Conservation Ecology of Cave Bats


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Conservation Ecology of Cave Bats

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Conservation Ecology of Cave Bats
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Bats in the Anthropocene: Conservation of Bats in a Changing World
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Furey, Neil M.
Racey, Paul A.
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Karst Cave ( local )
Cave Tourism ( local )
Incidental Disturbance ( local )
Karstic Cafe ( local )
Eidolon Helvum ( local )
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serial ( sobekcm )

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Caves and other subterranean sites such as mines are critical to the survival of hundreds of bat species worldwide, since they often provide shelter for most of a nation’s bat fauna. In the temperate zone, caves provide roosts for hibernation and for some species, breeding in summer, whereas in warmer regions, they support high species richness year round and enormous colonies that maintain substantial ecosystem services. Due to the solubility of the substrate, the highest densities of caves occur in karst landscapes. Given their importance for bats, relatively few studies have investigated factors involved in cave selection, although current evidence suggests that the density and size of caves are the best predictors of species diversity and population sizes. Thermal preferences have been established for some cave-dwelling species as well as their vulnerability to disturbance, particularly during hibernation and reproduction. Growth in limestone quarrying and cave tourism industries worldwide severely threatens cave-dwelling bats, in addition to loss of foraging habitat, hunting for bushmeat, incidental disturbance and disruptive guano harvesting. Apparent declines of cave bats in Europe and North America also pose serious concerns, as do global climate change predictions. The main conservation response to threats to cave bats in these continents has been gating, but this remains relatively untested as a means of protecting colonies in other regions. Research on sustainable harvesting of bats as bushmeat and their responses to different types of human disturbance at caves and loss of surrounding foraging habitats is required. More caves of outstanding importance for bats at national and international levels also require protection.

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ChristianC.Voigt· TiggaKingston Editors Bats in the Anthropocene: Conservation of Bats in a Changing World

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Bats in the Anthropocene: Conservation of Bats in a Changing World

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Christian C. Voigt · Tigga Kingston Editors Bats in the Anthropocene: Conservation of Bats in a Changing World

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Editors Christian C. Voigt Berlin Germany ISBN 978-3-319-25218-6þt ISBN 978-3-319-25220-9þ (eBook) DOI 10.1007/978-3-319-25220-9Library of Congress Control Number: 2015950865 Springer Cham Heidelberg New York Dordrecht London © The Editor(s) (if applicable) and The Author(s) 2016. The book is published with open access at SpringerLink.com. Open Access This book is distributed under the terms of the Creative Commons Attribution Noncommercial License, which permits any noncommercial use, distribution, and reproduction in any medium, provided the original author(s) and source are credited. All commercial rights are reserved by the Publisher, whether the whole or part of the material is concerned, specically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microlms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specic statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. The publisher, the authors and the editors are safe to assume that the advice and information in this book are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the editors give a warranty, express or implied, with respect to the material contained herein or for any errors or omissions that may have been made. Printed on acid-free paper Springer International Publishing AG Switzerland is part of Springer Science+Business Media (www.springer.com)Tigga Kingston Lubbock, TX USA

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For Thomas H. Kunz and Otto von Helversen for sharing with us their passion for bats. For Silke, Philippa and Florian (CCV) and for Danny (TK) for their inspiration and patience.

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viiContents1þt Bats in the Anthropoceneþ. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ....................................þt 1 Christian C. Voigt and Tigga Kingston Part Iþ Bats in Anthropogenically Changed Landscapes 2þt Urbanisation and Its Effects on Bats—A Global Meta-Analysisþ. . . . ....þt 13 Kirsten Jung and Caragh G. Threlfall 3þt Bats and Roadsþ. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ............................................þt 35 John Altringham and Gerald Kerth 4þt Responses of Tropical Bats to Habitat Fragmentation, Logging, and Deforestationþ. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ..........................................þt 63 Christoph F.J. Meyer, Matthew J. Struebig and Michael R. Willig 5þt Insectivorous Bats and Silviculture: Balancing Timber Production and Bat Conservationþ. . . . . . . . . . . . . . . . . . . . . . . . . . . . . .............................þt 105 Bradley Law, Kirsty J. Park and Michael J. Lacki 6þt Bats in the Anthropogenic Matrix: Challenges and Opportunities for the Conservation of Chiroptera and Their Ecosystem Services in Agricultural Landscapesþ. . . . . . . . ........þt 151 Kimberly Williams-Guillén, Elissa Olimpi, Bea Maas, Peter J. Taylor and Raphaël Arlettaz 7þt Dark Matters: The Effects of Articial Lighting on Batsþ. . . . . . . . . . ..........þt 187 E.G. Rowse, D. Lewanzik, E.L. Stone, S. Harris and G. Jones

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Contents viii 8þt Bats and Water: Anthropogenic Alterations Threaten Global Bat Populationsþ. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ............................................þt 215 Carmi Korine, Rick Adams, Danilo Russo, Marina Fisher-Phelps and David Jacobs Part IIþ Emerging Disesases 9þt White-Nose Syndrome in Batsþ. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ................................þt 245 Winifred F. Frick, Sébastien J. Puechmaille and Craig K.R. Willis 10þt Zoonotic Viruses and Conservation of Batsþ. . . . . . . . . . . . . . . . . . . . . .....................þt 263 Karin Schneeberger and Christian C. Voigt Part IIIþ Human-Bat Conicts 11þt Impacts of Wind Energy Development on Bats: A Global Perspectiveþ. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ........................................þt 295 Edward B. Arnett, Erin F. Baerwald, Fiona Mathews, Luisa Rodrigues, Armando Rodríguez-Durán, Jens Rydell, Rafael Villegas-Patraca and Christian C. Voigt 12þt Exploitation of Bats for Bushmeat and Medicineþ. . . . . . . . . . . . . . . . ................þt 325 Tammy Mildenstein, Iroro Tanshi and Paul A. Racey 13þt The Conict Between Pteropodid Bats and Fruit Growers: Species, Legislation and Mitigationþ. . . . . . . . . . . . . . . . . . . . . . . . . . . . ............................þt 377 Sheema Abdul Aziz, Kevin J. Olival, Sara Bumrungsri, Greg C. Richards and Paul A. Racey 14þt Bats and Buildings: The Conservation of Synanthropic Batsþ. . . . . . . .......þt 427 Christian C. Voigt, Kendra L. Phelps, Luis F. Aguirre, M. Corrie Schoeman, Juliet Vanitharani and Akbar Zubaid 15þt Conservation Ecology of Cave Batsþ. . . . . . . . . . . . . . . . . . . . . . . . . . . . ............................þt 463 Neil M. Furey and Paul A. Racey Part IVþ Conservation Approaches, Educational and Outreach Programs 16þt The Roles of Taxonomy and Systematics in Bat Conservationþ. . . . . . ......þt 503 Susan M. Tsang, Andrea L. Cirranello, Paul J.J. Bates and Nancy B. Simmons

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Contents ix17þt Networking Networks for Global Bat Conservationþ. . . . . . . . . . . . . . ..............þt 539 Tigga Kingston, Luis Aguirre, Kyle Armstrong, Rob Mies, Paul Racey, Bernal Rodríguez-Herrera and Dave Waldien 18þt Cute, Creepy, or Crispy—How Values, Attitudes, and Norms Shape Human Behavior Toward Batsþ. . . . . . . . . . . . . . . . ................þt 571 Tigga Kingston Indexþ. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .........................................................þt 597

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1Chapter 1Bats in the AnthropoceneChristian C. Voigt and Tigga Kingston© The Author(s) 2016 C.C. Voigt and T. Kingston (eds.), Bats in the Anthropocene: Conservation of Bats in a Changing World, DOIþ  10.1007/978-3-319-25220-9_1Abstractþ Humans have inadvertently changed global ecosystems and triggered the dawn of a new geological epoch, the Anthropocene. While some organisms can tolerate human activities and even ourish in anthropogenic habitats, the vast majority are experiencing dramatic population declines, pushing our planet into a sixth mass extinction. Bats are particularly susceptible to anthropogenic changes because of their low reproductive rate, longevity, and high metabolic rates. Fifteen percent of bat species are listed as threatened by the IUCN, i.e., they are considered Critically Endangered, Endangered or Vulnerable. About 18þ  % of species are Data Decient, highlighting the paucity of ecological studies that can support conservation status assessments. This book summarizes major topics related to the conservation of bats organized into sections that address: the response of bats to land use changes; how the emergence of viral and fungal diseases has changed bat populations; our perception of bats; and drivers of human–bat conicts and possible resolutions and mitigation. The book ends with approaches that might advance bat conservation through conservation networks and a better understanding of human behavior and behavioral change. C.C. Voigtþ  ()þ  Leibniz Institute for Zoo and Wildlife Research, Alfred-Kowalke-Str. 17, 10315 Berlin, Germany e-mail: voigt@izw-berlin.de C.C. Voigtþ  Institute for Biology, Freie Universität Berlin, Takustr. 6, 14195 Berlin, Germany T. Kingstonþ  ()þ  Department of Biological Sciences, Texas Tech University, Lubbock, TX, USA e-mail: tigga.kingston@ttu.edu

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2 C.C. Voigt and T. Kingston1.1þ The Emergence of a New Geological Epoch: The AnthropoceneThe world in which we live is fragile; a small layer of organismic activity covers the planet like a microbial lm on top of a large boulder. Nonetheless, humans treat the Earth as if anthropogenic impacts on this delicate biological layer may be absorbed by unfailing natural buffers. Yet, convergent and overwhelming evidence from all over the world underlines that mankind has already changed and continues changing the face of our planet. Among the many transformations humans imposed on our planet, some of the most severe appear to be (1) the addition of more than 550þ  billion metric tons of carbon to the atmosphere which are the main drivers of global climate change and ocean acidication (Gray 2007; Ciasi and Sabine 2013), (2) the alteration of the global nitrogen cycle by the use of articial fertilizers (Caneld etþ  al. 2010), (3) the routing of more than one third of global primary production to human consumption (Krausmann etþ  al. 2013), (4) the ongoing mass extinction of species (Barnosky etþ  al. 2011), and (5) the globalization of transport which has resulted in the spread of invasive species and pathogens (Lewis and Maslin 2015). It is now widely recognized that global ecosystem ser vices may be inadvertently suffering from human action, because human-induced environmental impacts are overriding natural process that have dominated our planet for millions of years (Steffen etþ  al. 2011). In the face of lasting human impacts on the Earth’s geological conditions and processes, many scientists, beginning with Paul Crutzen and Eugene Stoermer in 2000, now posit that our actions have brought us to the dawn of a new geological epoch—the Anthropocene. The pros and cons regarding this denition, which literally means “Human Epoch” and would succeed the Holocene, are still heavily debated (Monastersky 2015). Yet skeptics are declining in number, and much of the current debate focuses on the exact beginning of the Anthropocene, generally considered to be c. 1800. The Anthropocene working group of the Subcommission on Quaternary Stratigraphy reports to the International Commission on Stratigraphy with a proposal to formalize the Anthropocene in 2016. For the pur pose of this book, we do not refer to an exact starting point of the Anthropocene, but merely acknowledge the fact that humans have an impact on virtually all global ecosystems and that wildlife species such as bats (order Chiroptera) have adjusted to these changes, experienced substantial population declines, or gone extinct.1.2þ Bats in the Anthropocene: The Conservation of a Nocturnal TaxonBats (order Chiroptera) include more than 1300 extant species, forming the second largest mammalian order, and are unique among mammals in their evolution of powered ight. Although the common ancestor of living bats dates back to the K/T

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3 boundary (c. 70þ  mya), the most rapid radiation of any mammalian order resulted in all 18 extant families by the end of the Eocene c. 37þ  mya (Teeling etþ  al.2005). Moreover, although the majority of bat species are insectivorous, trophic diversity is extraordinary for a single order, with frugivores, nectarivores, piscivores, sanguinivores, and carnivores represented. Bats currently inhabit all continents except Antarctica, and in many parts of the world, especially the tropics, are the most species-rich mammalian group at a given locality, with alpha diversity reaching about 70 species in the Paleotropics (Kingston etþ  al. 2010) and over 100 in the Neotropics (Voss and Emmons 1996; Rex etþ  al. 2008). From any perspective, bats are an evolutionary and ecological success story. Nonetheless, bat populations are under severe threat in many regions of the world (Racey and Entwistle 2003). The last recorded case of a bat species driven to extinction is that of the Christmas Island pipistrelle, Pipistrellus murrayi (Lumsden and Schulz 2009; Lumsden 2009; Martin etþ  al. 2012), yet this species is most likely not the last one to vanish from our planet. The IUCN Bat Specialist Group is in the process of reassessing the Red List status of bat species, with the current assessments of 1150 species mostly completed in 2008, with 34 species assessed since. From these assessments, ve species were assessed as Extinct (giant vampire bat (Desmodus draculae), dusky ying fox (Pteropus brunneus), large Pelew ying fox (P. pilosus), dark ying fox (P. subniger), and Guam ying fox (P. tokudae)). The giant vampire bat is known only from the fossil and subfossil records, and the causes of its extinction are unknown. However, the four island Pteropus spp. are all victims of the Anthropocene, with hunting and habitat loss as the main drivers of extinction. Fifteen percent of bat species are listed in the threatened categories [Critically Endangered (CE), Endangered (EN), and Vulnerable (VU)] and 7þ  % are Near Threatened (Fig.þ  1.1). Around 18þ  % of species are Data Decient (DD), and there have been a wealth of new species discovered since the last assessment. The pattern of vulnerability is fairly consistent across families (Fig.þ  1.2), with the notable exception of the Pteropodidae with 36þ  % of species extinct or threatened, probably because of their size, their appeal as bushmeat and for traditional medicine, Fig.þ  1.1þ Red List status of the 1150 bat species assessed 2008–2014 (IUCN 2015). IUCN categories are EX Extinct, CR Critically Endangered, EN Endangered, VU Vulnerable, NT Near Threatened, DD Data Decient, LC Least Concern. Number of species and percentage of all species given as labels

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4 C.C. Voigt and T. Kingstonand because many form susceptible island populations. Even this depicts only part of the picture; populations are only considered stable in 21þ  % of all species and increasing in less than 1þ  %. Of the remaining species, populations are decreasing (23þ  %) or the trend is unknown (55þ  %). Moreover, of the 687 species assessed as Least Concern (LC), current specic threats were identied for about 27þ  % of species. Declining populations and identied threats suggest a bleak future, and it is probable that more species will satisfy the rigorous criteria of the threatened categories in the coming years. Globally, the major threats to bat species identied by IUCN assessments are land use change (logging, non-timber crops, livestock farming and ranching, wood and pulp plantations, and re), urbanization, hunting and persecution, quarrying and general human intrusions on bat habitats (Fig.þ  1.3). Bats are particularly susceptible to these human-induced perturbations of habitats because of their distinct life history. Bats are on the slow side of the slow-fast continuum of life histories (Barclay and Harder 2003). For example, they reproduce at a low rate (Barclay etþ  al. 2004) and are long-lived mammals (Munshi-South and Wilkinson 2010; Wilkinson and South 2002). Thus, bat populations recover slowly from increased mortality rates. Despite their low reproductive rate and longevity, bats have relatively high metabolic rates owing to their small size which leads to relatively high food requirements (Thomas and Speakman 2003). Lastly, bats are nocturnal animals with often cryptic habits. Even though they are present in many larger cities of the temperate zone, they often go unnoticed by their human neighbors. It is quite likely that perceptions of bats would be very different if Homo sapiens evolved as a nocturnal hominid. Or to put it in the words of Rich and Longcore: What if we woke up one morning and realize that we missed Fig.þ  1.2þ Red List status of bats by family. Abbreviations as for Fig.þ  1.1

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5 half of the story in our conservation efforts, namely the night part? (modied after Rich and Longcore 2004 , p. 1). This brings up an important question: Do noctur nal animals benet less from legal protection than diurnal animals? Are we more concerned about animals that we see and interact with during daytime? Do human societies perceive and evaluate, for example, fatalities of birds of prey at wind tur bines in a different way than bat fatalities when both ought to benet from the same level of protection? Do we consider recommendations to reduce light pol lution for the sake of nocturnal animals such as bats, or does the expansion of the human temporal niche into the night come at high costs for all nocturnal animals? In summary, we speculate that bats as nocturnal animals might be particularly exposed to human-induced ecological perturbations because we are driven by our visual system and therefore tend to neglect the dark side of conservation, i.e., the protection of nocturnal animals. 1.3 Why Care About Bat Conservation? The reasoning for the conservation of nature can be manifold, reaching from purely moral to monetary arguments and legal requirements. It may also vary according to the scale of the conservation approach, i.e., whether it is driven by Fig. 1.3 Frequency of threats listed in the IUCN assessments of bat species. a Distribution of major threats across assessments. Land use changes, urbanization. and hunting are aggregations of IUCN listed threats given in b – d . Frequency of threat and percentage contribution are given

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6 C.C. Voigt and T. Kingstonlocal, national, or international perspectives. Indeed, ethical considerations for the protection of species—although quite often neglected in modern civilization—should be the primary motivation; i.e., the obligation of humans to conserve nature for the simple reason of its existence and for the more selsh reason to make the diversity of biological life accessible and useable to following generations of humans. Lately, economic arguments for the conservation of nature are increasingly used, e.g., the importance of protecting water catchment areas to provide potable water or irrigation in agriculture. So-called ecosystem services of nature are highly valued in modern societies and therefore benet from increasing protection. Recent attempts to critically review the ecosystem services provided by bats have revealed that many species offer unique and large-scale monetary benets to agricultural industry (Kunz etþ  al. 2011; Ghanem and Voigt 2012; Maas etþ  al. 2015). For example, owers of the Durian tree are only effectively pollinated by the Dawn bat, Eonycteris spelaea, in Southeast Asia (Bumrungsri etþ  al. 2009). Durian is a highly valued fruit in Asia with Thailand producing a market value of durians of almost 600 million US$ annually (Ghanem and Voigt 2012). Other bats consume large amounts of pest insects, thereby offering services that could save millions of US$ for national industries (Boyles etþ  al. 2011; Wanger etþ  al. 2014). However, the monetary approach for protecting bat species is a doubleedged sword, since bat species without apparent use for human economy may not benet from protection compared to those that provide some ecosystem services. Moreover, arguments based on economic or utilitarian values of wildlife may appeal to self-interest motivations and suppress environmental concern (Kingston 2016). In this context, it is important to note that we have just started to under stand the ecological role bats ll in natural ecosystems. For example, bats have been recently documented as top-down regulators of insect populations in forest habitats of the tropics and temperate zone (Kalka etþ  al. 2008; Boehm etþ  al. 2011) and also in subtropical coffee and cacao plantations (Williams-Guillen etþ  al. 2008; Maas etþ  al. 2013). Finally, bats are protected by law in some countries. For example, they are covered by the Habitat Directive of the European Union and thus strictly protected in E.U. countries. Also, migratory bats benet from some level of protection because they are covered by the UN Convention for the Protection of Migratory Species. Countries that have signed this convention are obliged to support conservation actions that are benecial for migratory species. CITES (The Convention on International Trade in Endangered Species of Wild Fauna and Flora) protects threatened species through controls of international trade in specimens. The precarious conservation status of the ying foxes is apparent. Currently, Acerodon jubatus and ten Pteropus spp are on CITES Appendix I, with trade only permitted in exceptional circumstances, and the remaining Acerodon and Pteropus species on Appendix II, by which trade is controlled to avoid utilization incompatible with their survival.

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7 1.4þ About This BookThe idea to publish a book about bat conservation was stimulated by the “3rd International Berlin Bat Meeting: Bats in the Anthropocene” in 2013. The overall goal is to provide a summary of the major threats bats are facing in a rapidly changing world. The book is organized in four major sections: (1) bats in anthropogenically-shaped landscapes, (2) emerging diseases, (3) human–bat conicts, and (4) conservation approaches. The basic concept of chapters in all of these sections is to review the literature that is available in peer-reviewed journals. We are aware that many topics related to bat conservation have also been addressed in brochures or books published by non-governmental or governmental organizations. Sometimes these sources have been cited in the corresponding chapters, yet in most cases authors of this book have focused on the aforementioned sources of information. From our editorial perspective, the chapters cover the majority of relevant topics in bat conservation. However, we acknowledge that at least three topics are missing in this book. First, this book misses a chapter on “bats and global climate change,” because Jones and Rebelo (2013) published a recent review on this topic and the body of literature about this topic has not largely increased since then. Second, we did not commission a chapter on “Bats and chemical pollutants,” as current knowledge of heavy metals was recently synthesized by Zukal etþ  al. (2015) and information for other pollutants is sparse. That said, the subject is referenced in several chapters (Williams-Guillen etþ  al. 2015; Korine etþ  al. 2015; Voigt etþ  al. 2016). Third, we did not include a chapter on “island bats,” although many of them are endangered and some even are threatened by extinction, as Fleming and Racey (2010) provide a detailed overview of this topic in their recent book. Finally, authors integrate successful interventions into their accounts and make specic recommendations for future research, but additional evidencebased evaluations of the success of conservation interventions per se are found in Berthinussen etþ  al. (2014). The Anthropocene has gained momentum. It is a geological epoch that is not in equilibrium but is constantly changing by the action of mankind. For a handful of bat species anthropogenic changes may prove benecial, but for the vast majority our actions precipitate drastic population declines that must be slowed if we are to conserve the extraordinary diversity of this unique mammalian order. We hope that this book will stimulate new directions for research and support conservation interventions that will keep the night sky alive with bats in the Human Epoch.Acknowledgementsþ We would like to acknowledge the nancial support provided by the Leibniz Institute for Zoo and Wildlife Research in Berlin, Germany, by EUROBATS and a National Science Foundation grant to the Southeast Asian Bat Conservation Research Unit (NSF Grant No. 1051363) that enabled us to publish this book as an open-access electronic book. We thank Mark Brigham, Anne Brooke, Justin Boyles, Gabor Csorba, Brock Fenton, Jorge Galindo-González, Chris Hein, Carmi Korine, Allen Kurta, Pia Lentin, Herman Limpens, Lindy Lumsden, Jörg Müller, Alison Peel, Paul Racey, Hugo Rebelo, DeeAnn Reeder, Scott Reynolds, Danilo Russo, Armando Rodríguez-Durán, N. Singaravelan, Vikash Tatayah, Peter Taylor, and numerous anonymous reviewers for providing constructive comments on chapters of this book.

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8 C.C. Voigt and T. KingstonReferencesBarclay RMR, Harder LD (2003) Life histories of bats: life in the slow land. In: Kunz TH, Fenton MB (eds) Bat ecology. University of Chicago Press, Chicago and London Barclay RMR, Ulmer J, MacKenzie CJA, Thompson MS, Olson L, McCool J, Cropley E, Poll G (2004) Variation in the reproductive rate of bats. Can J Zool 82:688–693 Barnosky AD, Matzke N, Tomiya S, Wogan GOU, Swartz B, Quental TB, Marshall C, McGuire JL, Lindseay EL, Maguire KC, Mersey B, Ferrer EA (2011) Has the Earth’s sixth mass extinction already arrived. Nature 471:51–57 Berthinussen A, Richardson OC, Altringham JD (2014) Bat conservation: global evidence for the effects of interventions. Pelagic Publishing, Exeter Boehm SM, Wells K, Kalko EKV (2011) Top-down control of herbivory by birds and bats in the canopy of temperate Broad-Leaved Oaks (Quercus robur). PLoS One 6(4):e17857. doi:10.1371/journal.pone.0017857 Boyles JG, Cryan PM, McCracke GF, Kunz TH (2011) Economic importance of bats in agriculture. Science 332:41–42 Bumrungsri S, Sripaoraya E, Chongiri T, Sridith JK, Racey PA (2009) The pollination ecology of durian (Durio hibethinus, Bombacaceae) in southern Thailand. J Trop Ecol 25:85–92 Caneld DE, Glazer AN, Falkowski PG (2010) The evolution and future of Earth’s nitrogen cycle. Science 330:192–196 Ciasi P, Sabine C (2013) Chapter 6. In: Stocker TF, Qin D, Plattner G-K, Tignor MMB, Allen SK, Boschung H, Nauels A, Xia Y, Bex V, Midgley PM (eds) Climate change 2013: the physical science basis. Contribution of working group I to the fth assessment report of the inter governmental panel on climate change, pp 465–570 Fleming TH, Racey PA (2010) Island bats: evolution, ecology and conservation. University of Chicago Press, Chicago, p 592 Ghanem SH, Voigt CC (2012) Increasing awareness of ecosystem services provided by bats. Adv Study Behav 44:279–302 Gray V (2007) Climate change 2007: the physical science basis summary for policymakers. Energy Environ 18:433–440 IUCN (2015) The IUCN Red List of Threatened Species. Version 2015-3.http://www.iucnredlist. org . Downloaded on 9 September 2015 Jones G, Rebelo H (2013) Responses of bats to climate change: learning from the past and predicting the future. In: Adams RA, Pedersen SC (eds) Bat evolution, ecology, and conservation. Springer, New York, Berlin, pp 457–478 Kalka MB, Smith AR, Kalko EKV (2008) Bats limit arthropods and herbivory in a tropical for est. Science 320:71. doi:10.1126/science.1153352 Kingston T (2010) Research priorities for bat conservation in Southeast Asia: a consensusapproach. Biodivers Conserv 19:471–484 Kingston T (2016) Cute, Creepy, or Crispy How values, attitudes, and norms shape human behavior toward bats. In: Voigt CC, Kingston T (eds) Bats in the Anthropocene: conservation of bats in a changing world. Springer International AG, Cham, pp 571–588 Korine C, Adams R, Russo D, Fisher-Phelps M, Jacobs D (2015) Bats and water: anthropogenic alter ations threaten global bat populations. In: Voigt CC, Kingston T (eds) Bats in the Anthropocene: conservation of bats in a changing world. Springer International AG, Cham, pp 215–233 Krausmann F, Erb K-H, Gingrich S, Haberla H, Bondeau A, Gaube V, Lauk C, Plutzar C, Searchinger TD (2013) Global human appropriation of net primary production doubled in the 20th century. Proc Natl Acad Sci USA 110:10324–10329 Open Access This chapter is distributed under the terms of the Creative Commons Attribution Noncommercial License, which permits any noncommercial use, distribution, and reproduction in any medium, provided the original author(s) and source are credited.

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9 Kunz TH, Braun de Torrez E, Bauer DM, Lobova TA, Fleming TH (2011) Ecosystem services provided by bats. In: Ostfeld RA, Schlesinger WH (eds) The year in ecolgoy and conservation 2011: annals of the New York academy of sciences. Wiley, New York, USA, pp 1–38 Lewis SL, Maslin MA (2015) Dening the Anthropocene. Nature 519:171–180 Lumsden L (2009) The extinction of the Christmas Island Pipistrelle. Aus Bat Soc Newsl 33:21–25 Lumsden L, Schulz M (2009) Captive breeding and future in-situ management of the Christmas Island Pipistrelle Pipistrellus murrayi. A report to the Director of National Parks. Arthur Rylah Institute, Department of Sustainability and Environment, Heidelberg, Victoria Maas B, Clough Y, Tscharntke T (2013) Bats and birds increase crop yield in tropical agroforesty landscapes. Ecol Lett 16:14801 Maas B, Karp DS, Bumrungstri S, Darras K, Gonthier D, Huang JCC, Lindell CA, Maine JJ, Mestre L, Michel NL, Morrison EB, Perfecto I, Philpott SM, Sekergioglu CH, Silva RM, Taylor PJ, Tscharntke T, Van Bael SA, Whelan CH, Williams-Guillen K (2015) Bird and bat predation services in tropical forests and agroforestry landscapes. Biol Rev. doi:10.1111/brv.12211 Martin TG, Nally S, Burbidge AA, Arnall S, Garnett ST, Hayward MW, Lumsden L, Menhhorst P, McDonald-Madden E, Possingham HP (2012) Acting fast helps avoid extinction. Conserv Lett 5:274–280 Monastersky R (2015) Anthropocene: the human age—editorial. Nature 519:144–147 Munshi-South J, Wilkinson GS (2010) Bats and birds: exceptional longevity despite high metabolic rates. Ageing Res Rev 9:12–19 Racey PA, Entwistle AC (2003) Conservation ecology of bats. In: Kunz TH, Fenton MB (eds) Bat ecology. University of Chicago Press, Chicago, pp 680–743 Rex K, Kelm DH, Wiesner K, Kunz TH, Voigt CC (2008) Species richness and structure of three Neotropical bat assemblages. Biol J Linn Soc 94:617–629 Rich C, Longcore T (2004) Ecological consequences of articial night lighting. Island Press, Washington Steffen W, Grinevald J, Crutzen P, McNeill J (2011) The anthropocene: conceptual and historical perspectives. Philos Trans R Soc A 369:842–867 Teeling EC, Springer MS, Madsen O, Bates P, O’Brien SJ, Murphy WJ (2005) A molecular phylogeny for bats illuminates biogeography and the fossil record. Science 307:580–584 Thomas D, Speakman JR (2003) Physiological ecology and energetics of bats. In: Kunz TH, Fenton MB (eds) Bat ecology, University of Chicago Press. Chicago and London, pp 430–490 Voss RS, Emmons LH (1996) Mammalian diversity in Neotropical lowland rainforests: a preliminary assessment. Bull Am Mus Nat Hist 230:1–115 Voigt CC, Phelps KL, Aguirre L, Schoeman MC, Vanitharani J, Zubaid A (2016) Bats and buildings: the conservation of synanthropic bats. In: Voigt CC, Kingston T (eds) Bats in the Anthropocene: conservation of bats in a changing world. Springer International AG, Cham, pp 427–453 Wanger TC, Darras K, Bumrungsri S, Tscharntke T, Klein AM (2014). Bat pest control contributes to food security in Thailand. Biol Conserv 171:220–223 Wilkinson GS, South JM (2002) Life history, ecology and longevity in bats. Aging Cell 1:124–131 Williams-Guillen K, Perfecto I, Vandermeer J (2008) Bats limit insects in a neotropical agrofor estry system. Science 320:70. doi:10.1126/science.1152944 Williams Guillen etþ  al. (2016) Bats in the anthropogenic matrix: challenges and opportunities for the conservation of Chiroptera and their ecosystem services in agricultural landscapes. In: Voigt CC, Kingston T (eds) Bats in the Anthropocene: conservation of bats in a changing world. Springer International AG, Cham, pp 151–178 Zukal J, Pikula J, Bandouchova H (2015) Bats as bioindicators of heavy metal pollution: history and prospect. Mamm Biol-Z Saugertierkd 80:220–227

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Part IBats in Anthropogenically Changed Landscapes

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13Chapter 2Urbanisation and Its Effects on Bats—A Global Meta-AnalysisKirsten Jung and Caragh G. Threlfall© The Author(s) 2016 C.C. Voigt and T. Kingston (eds.), Bats in the Anthropocene: Conservation of Bats in a Changing World, DOIþ  10.1007/978-3-319-25220-9_2Abstractþ Urbanisation is viewed as the most ecologically damaging change to land use worldwide, posing signicant threats to global biodiversity. However, studies from around the world suggest that the impacts of urbanisation are not always negative and can differ between geographic regions and taxa. Bats are a highly diverse group of mammals that occur worldwide, and many species per sist in cities. In this chapter, we synthesise current knowledge of bats in urban environments. In addition, we use a meta-analysis approach to test if the general response of bats depends on the intensity of urbanisation. We further investigate if phylogenetic relatedness or functional ecology determines adaptability of species to urban landscapes and if determining factors for urban adaptability are consistent worldwide. Our meta-analysis revealed that, in general, habitat use of bats decreases in urban areas in comparison to natural areas. A high degree of urbanisation had a stronger negative effect on habitat use compared to an intermediate degree of urbanisation. Neither phylogenetic relatedness nor functional ecology alone explained species persistence in urban environments; however, our analysis did indicate differences in the response of bats to urban development at the family level. Bats in the families Rhinolophidae and Mormoopidae exhibited a negative association with urban development, while responses in all other families were highly heterogeneous. Furthermore, our analysis of insectivorous bats K. Jungþ  ()þ  Institute of Evolutionary Ecology and Conservation Genomics, University of Ulm, Ulm, Germany e-mail: kirsten.jung@uni-ulm.de C.G. Threlfallþ  ()þ  School of Ecosystem and Forest Sciences, University of Melbourne, Melbourne, Australia e-mail: caragh.threlfall@unimelb.edu.au

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14 K. Jung and C.G. Threlfallrevealed that the adaptability of individual families, e.g. Emballonuridae and Vespertilionidae, to urbanisation is not consistent worldwide. These results suggest that behavioural and/or morphological traits of individual species may better determine species’ adaptability to urban areas, rather than phylogenetic or functional classications, and that driving factors for species adaptability to urban areas might be regionally divergent. We thus argue that future research should focus on behavioural and morphological traits of bats, to assess if these determine urban adaptability in this species-rich group of mammals.2.1þ Introduction 2.1.1þ The Urban ContextUrbanisation results in extreme forms of land use alteration (Shochat etþ  al. 2006; Grimm etþ  al. 2008). In the last century, the human population has undergone a transition in which the majority of people now live in urban rather than rural areas (UNPD 2012). The rate of change at which urban areas are evolving due to natural population growth is dramatic, including signicant rural-to-urban migration and spatial expansion (Grimm etþ  al. 2008; Montgomery 2008; UN 2012; McDonnell and Hahs 2013). In the last 50þ  years, the global human population in urban areas increased from 2.53 to 6.97þ  billion people (UNPD 2012). Yet human pressure resulting from urbanisation is not uniformly distributed on the planet. While urbanisation in the developed countries is slowing down slightly, it is increasing rapidly in developing countries of Asia, Africa, Latin America and the Caribbean, many of which harbour hotspots of biodiversity (Myers etþ  al. 2000). In addition, over half of the urban population growth is projected to occur in smaller towns and cities (UN 2012). This implies that urbanisation is not a locally concentrated event, it is rather a fundamentally dispersed process and a happening worldwide (McDonald 2008). The ecological footprint of cities reaches far beyond their boundaries (McGranahan and Satterthwaite 2003; McDonald and Marcotullio 2013). Effects of cities operate from local (e.g. through urban sprawl) to global scales (e.g. through greenhouse gas emission) (McDonald etþ  al. 2008), and act both directly, through expansion of urban areas, and indirectly through growth in infrastructure and changes in consumption and pollution (McIntyre etþ  al. 2000; Pickett etþ  al. 2001). Apart from the obvious loss in natural area, expansion of cities also impacts the surrounding rural and natural habitats through increased fragmentation, and edge effects with increasing temperature and noise levels, which together introduce new anthropogenic stressors on fringe ecosystems (Grimm etþ  al. 2008) and nearby protected areas (McDonald etþ  al. 2008; McDonald and Marcotullio 2013). However, despite the radical land transformation and habitat loss incurred through urbanisation, many species (native and introduced) can still persist in urban environments and some even experience population increases (McKinney 2006). This

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15 2þ Urbanisation and Its Effects on Bats—A Global Meta-Analysissuggests that urban landscapes can actually provide suitable habitat for a variety of species, albeit an anthropogenically altered habitat. Nevertheless, our understanding of what constitutes a suitable habitat in urban areas and what determines a species’ adaptability to an urban environment is currently very limited. Generally, urban areas are characterised by high quantities of impervious sur faces (McKinney 2002). There are however many additional physical and chemical changes incurred via the process of urbanisation (McDonnell and Pickett 1990), such as increased pollution, eutrophication, increased waste generation, altered hydrology (Vitousek etþ  al. 1997; Grimm etþ  al. 2008), increased urban noise (e.g. Slabbekoorn and Ripmeester 2008) and articial light (Longcore and Rich 2004). Urban areas can provide a more thermally stable environment via the urban heat island effect (e.g. Zhao etþ  al. 2006); less radiation is reected during the day and more heat is trapped at night, which can increase minimum temperatures in cities (Grimm etþ  al. 2008). The changed climate prole of cities can benet some species by making the area more inhabitable year round. In addition, the planting of attractive introduced and native plant species throughout the suburbs and along city roads also changes the resources available to fauna, for example by providing nectar or fruits throughout the year. Altogether these changes can impact local species assemblages within cities and regional biodiversity beyond the municipal boundaries (Grimm etþ  al. 2008). Anthropogenic changes in urban ecosystems typically occur at rates drastically faster than long-lived organisms are capable of adapting to, and thus disrupt ecological processes that historically governed community structure (Duchamp and Swihart 2008). However, generalisations about the negative effects of urbanisation should not overlook biologically meaningful differences in how taxa respond to human land use (Dixon 2012). Some wildlife species are able to adjust to a life in urban areas. Among vertebrates, a range of birds are relatively abundant in urban environments and bird species richness may peak at intermediate levels of urbanisation because of increased heterogeneity of edge habitats (Blair 2001; McKinney 2002) and changes in resource availability due to provision of articial feeding stations (Sewell and Catterall 1998). In contrast, only a few mammals have been documented as successful species in urban areas (Macdonald and Newdick 1982; Septon etþ  al. 1995; Luniak 2004). For example, the grey-headed ying fox (Pteropus poliocephalus) has established a year-round camp in urban Melbourne, Australia, an area outside of its normal climatic range. Warmer temperatures from the urban heat effect, enhanced precipitation from local irrigation and year-round food resources appear to have facilitated the colony’s arrival and persistence (Parris and Hazell 2005). Many animals, however, disappear from cities because they depend on habitat features that no longer exist (Gilbert 1989; McKinney 2002; Luniak 2004; Haupt etþ  al. 2006; McDonnell and Hahs 2008). Declining species often suffer from increased habitat isolation, or face competition from invasive and more dominant species (McDonald and Marcotullio 2013). Some species in urban areas also suffer from additional stress (Isaksson 2010), increased infection and parasitism rates (Giraudeau etþ  al. 2014) and reductions in potential reproductive success (Chamberlain etþ  al. 2009). Urbanisation can also trigger a change

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16 K. Jung and C.G. Threlfallin behaviour (Ditchkoff etþ  al. 2006; Grimm etþ  al. 2008). For example, urban noise alters the pitch at which some birds call (Slabbekoorn and Peet 2003), and affects activity patterns of larger vertebrates (Ditchkoff etþ  al. 2006). Furthermore, increased articial lighting can potentially disturb the circadian rhythms of noctur nal animals and interfere with the navigation of migrating species (Longcore and Rich 2004; Hölker etþ  al. 2010; see Rowse etþ  al.,þ  Chap.þ  7 this volume).2.1.2þ Urban WildlifePersistence of wildlife in urban environments may be linked to opportunism and a high degree of ecological and behavioural plasticity (Luniak 2004). In contrast, species that decline in response to urbanisation are often habitat and resource specialists (McKinney and Lockwood 1999; Jokimäki etþ  al. 2011). Typically this results in altered assemblage structures in urban environments, often with a few highly abundant species, which account for a much higher proportion of the whole community in urban environments than in surrounding wild lands (Shochat etþ  al. 2006). In addition, many native species are replaced by non-native, weedy or pest species (McKinney 2002). The resulting mix of introduced and native species in urban areas can lead to novel species interactions and altered ecosystem functioning (Hobbs etþ  al. 2006). Often these non-native and introduced species are the same species across cities throughout the world. Thus, the ora and fauna of cities are becoming increasingly homogeneous (Hobbs etþ  al. 2006; Grimm etþ  al. 2008), however recent evidence suggests that many cities still retain several endemic species (Aronson etþ  al. 2014). Multi-scaled and multi-taxa investigations are required to provide detailed information about urban biodiversity (Clergeau etþ  al. 2006). To date, urban ecologists have focused on few taxa, examining the response of conspicuous species to an urbanisation gradient (McDonnell and Hahs 2008). Populationand assemblage-level responses to urbanisation have been examined most prolically for highly diverse and mobile bird taxa (McKinney 2002; McDonnell and Hahs 2008). Unfortunately, our understanding of how other wildlife, including bats, respond to the complex process of urbanisation is still limited (Barclay and Harder 2003). Research conducted to date provides a general indication that many bats may be declining due to urbanisation, however an understanding of the processes driving these patterns remains largely unknown.2.1.3þ Bats in Urban EnvironmentsBats likely form the most diverse group of mammals remaining in urban areas (van der Ree and McCarthy 2005; Jung and Kalko 2011). Of the studies conducted in urban landscapes to date, many show that overall bat activity and

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17 2þ Urbanisation and Its Effects on Bats—A Global Meta-Analysisspecies richness are greatest in more natural areas, and decreases with increasing urban inuence (Kurta and Teramino 1992; Walsh and Harris 1996; Gaisler etþ  al. 1998; Legakis etþ  al. 2000; Lesiñski etþ  al. 2000). However, certain bat species may better be able to adapt to urban landscapes (Avila-Flores and Fenton 2005; Duchamp and Swihart 2008). Coleman and Barclay (2011), however, cautioned that most researchers have worked in forested regions directing less attention to other biomes, including grasslands. They argue that because urban tree cover is fairly constant (<30þ  %) in all cities (McKinney 2002), urbanisation in tree-rich regions implies deforestation and thus reduced tree cover may cause the negative effect of urbanisation. In contrast, urban areas within grassland regions might enhance structural heterogeneity and thus benet species richness and relative abundance patterns (see Coleman and Barclay 2011 for more details). This is in accordance with the results of Gehrt and Chelsvig (2003, 2004) investigating the response of bats in and around the highly populated city of Chicago, USA. Here species diversity and occurrence were higher in habitat fragments within urban areas than in similar fragments in rural areas (Gehrt and Chelsvig 2003, 2004). However the large, forested parks in the region may offset the habitat loss caused by urbanisation, and hence they mitigate any negative impacts to bats at the regional scale. The majority of studies on bats in urban environments come from the temper ate regions of Europe and North America. Many studies focus on the response of bats to differently structured areas within the urban environment including historic and newly built city districts (Gaisler etþ  al. 1998; Legakis etþ  al. 2000; Guest etþ  al. 2002; Dixon 2012; Hale etþ  al. 2012; Pearce and Walters 2012), illuminated and non-illuminated areas (Bartonicka and Zukal 2003), industrial areas (Gaisler etþ  al. 1998) small and larger parklands (Kurta and Teramino 1992; Fabianek etþ  al. 2011; Park etþ  al. 2012) and areas that receive waste water (Kalcounis-Rueppell etþ  al. 2007). Most of these studies report relatively high bat activity and species richness in areas with remaining vegetation such as older residential areas, riverine habitats or parklands. Certain bat species appear to thrive in these urban environments, and success has been linked to species-specic traits (Duchamp and Swihart 2008). In particular, bat species with high wing loadings and aspect ratios, so presumed to forage in open areas (Norberg and Rayner 1987), which also roost primarily in human structures appeared to adjust to urban environments, provided that there is sufcient tree cover (Dixon 2012). Many of these studies imply that protecting and establishing tree networks may improve the resilience of some bat populations to urbanisation (Hale etþ  al. 2012). Populationand assemblage-level responses along gradients of urbanisation reveal that generally foraging activity of bats seems to be higher in rural and forested areas than in urban areas (Geggie and Fenton 1985; Kurta and Teramino 1992; Lesiñski etþ  al. 2000). However, it is important to note that some species might be highly exible in their habitat use. The European bat Eptesicus nilsonii, for example, spends a much higher proportion of its foraging time in urban areas after birth of the juveniles than before (Haupt etþ  al. 2006). This raises the importance of repeat observations during different seasons when investigating the response of bats to urbanisation.

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18 K. Jung and C.G. ThrelfallIn the Neotropics, most studies concerning bats and environmental disturbance have concentrated on fragmentation effects due to logging or agricultural land use (e.g. García-Morales etþ  al. 2013). Persistence of bats in fragmented landscapes has been associated with edge tolerance and mobility in phyllostomids (Meyer and Kalko 2008), and the predominant use of open space as foraging habitat for aerial insectivorous bats (Estrada-Villegas etþ  al. 2010). Of the few studies focusing on urban areas, most report an overall decrease in species richness and relative abundance of bats in urban areas (Avila-Flores and Fenton 2005; Siles etþ  al. 2005; Pacheco etþ  al. 2010; Jung and Kalko 2011) compared to forested areas. Predominantly, insectivorous bats seem to remain in large urban environments (Bredt and Uieda 1996; Filho (2011). Of these, it is typically members of the Molossidae, which are known to forage in the open spaces above the tree canopy that seem to tolerate and potentially prot from highly urbanised areas (AvilaFlores and Fenton 2005; Pacheco etþ  al. 2010; Jung and Kalko 2011). In addition, many buildings in cities provide potential roost sites that resemble natural crevices (Burnett etþ  al. 2001; Avila-Flores and Fenton 2005) and are known to be readily occupied by molossid bats (Kössl etþ  al. 1999; Scales and Wilkins 2007). In a smaller urban setting in Panama, where mature forest meets very restricted urban development, a high diversity of bats occurs within the town and bats frequently forage around street lights (Jung and Kalko 2010). Nevertheless, even in such a low impact urban setting some species of the bat assemblage such as Centronycteris centralis revealed high sensitivity and were never recorded within the town, albeit foraging frequently in the nearby mature forest (Jung and Kalko 2010). Recent investigations from large metropolitan urban centres in Australia show suburban areas can provide foraging habitat for bats (Rhodes and Catterall 2008; Threlfall etþ  al. 2012a), and support greater bat activity and diversity than more urban and even forested areas (Hourigan etþ  al. 2010; Basham etþ  al. 2011; Threlfall etþ  al. 2011, 2012b; Luck etþ  al. 2013). However, studies from regional urban centres in Australia suggest that any urban land cover, even if low-density residential, can decrease bat activity and species richness (Hourigan etþ  al. 2006; Gonsalves etþ  al. 2013; Luck etþ  al. 2013), and can deter some species of clutter-tolerant bats altogether (Gonsalves etþ  al. 2013; Luck etþ  al. 2013). Evidence also suggests that species adapted to open spaces and edges, such as those within the molossid family, do not display the same response to urbanisation in small regional versus large metropolitan urban centres, indicating subtle behavioural differences among species with similar ecomorphology (Luck etþ  al. 2013; McConville etþ  al. 2013a, b). The few studies that have investigated species-specic foraging and roosting requirements, suggest that although bats display high roost site delity within urban areas (Rhodes and Wardell-Johnson 2006; Rhodes etþ  al. 2006; Threlfall etþ  al. 2013a), species differ in their ability to forage successfully on aggregations of insects across the urban matrix, reecting variation in ight characteristics and sensitivity to articial night lighting (Hourigan etþ  al. 2006; Scanlon and Petit 2008; Threlfall etþ  al. 2013b). Asian bat assemblages comprise a variety of frugivore and insectivore bat species; however, there is limited information on urban impacts to bats in this region

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19 2þ Urbanisation and Its Effects on Bats—A Global Meta-Analysisof the world. Many roosting and foraging resources for frugivore species such as Cynopterus and Pteropus species are provided by exotic trees that grow easily in urban centres in Asia, for example Ficus, Livistona and Syzygium species, which have been studied in Hong Kong (Corlett 2005, 2006), India (Caughlin etþ  al. 2012) and Japan (Nakamoto etþ  al. 2007). Frugivore species in these systems provide critical seed dispersal services and can play a role in regeneration and pollination of some tree species (Mahmood-ul-Hassan etþ  al. 2010; Caughlin etþ  al. 2012). Radiotracking studies show that some bat species roost in forested areas (Nakamoto etþ  al. 2012) or in-built structures (Nadeem etþ  al. 2013), however many frugivore species appear to prot from the density of planted exotic vegetation and both frugivore and insectivore bats can benet from increased foraging resources in urban areas (Corlett 2005; Nakamoto etþ  al. 2007; Utthammachai etþ  al. 2008; Caughlin etþ  al. 2012; Nakamoto etþ  al. 2012). However, it appears that Asian bats, particularly large pteropodids, are also under threat from direct human impacts via hunting (Thomas etþ  al. 2013), in addition to human land use alteration, and hence, any impact of urbanisation may be confounded by direct human impacts. However, increasing land use change and growing urban populations have been stated as a likely cause of dramatic declines of many bat species (including pteropodids) in Singapore (Pottie etþ  al. 2005; Lane etþ  al. 2006), where it is suggested the reported declines may reect the declining status of bats in Southeast Asia more broadly (Lane etþ  al. 2006). The only study to our knowledge that has examined bat species distribution in relation to increasing urbanisation was conducted in Pakistan, where greater bat capture success was recorded in urban areas in comparison to suburban and rural areas (Nadeem etþ  al. 2013), and in line with other studies worldwide, the urban bat assemblage was dominated by a few common species. However, it is unclear whether these results were inuenced by trapping success, and as such, should be interpreted cautiously. The co-location of biodiversity and high human population densities raises the importance of conservation-related studies in urban areas where anthropogenic growth directly interacts with the highest levels of biodiversity (Rompré etþ  al. 2008). In these landscapes, it is especially important to identify the underlying mechanisms determining the potential of different species to adjust to urban environments. Currently, our general understanding of what inuences a species success and details of urban foraging and roosting habitat selection is incomplete. Yet, arguably the conservation of species such as bats in urban areas dependents upon this knowledge (Fenton 1997).2.2þ Evidence-Based Evaluation of the Effect of Urbanisation on Bats Worldwide Using a Meta-AnalysisWithin this book chapter, we were in particular interested in the general conclusions concerning the potential of bats to adjust to urban environments. We thus synthesised pre-existing data of published literature with a focus on bats in urban

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20 K. Jung and C.G. Threlfallversus natural environments in a worldwide meta-analysis. Meta-analysis has been previously used in ecology and conservation because results can lead to evidencebased environmental policies. Here, we investigated the general response of bats to urbanisation and tested whether this is consistent across cities differing in the intensity of urban development. In addition, we address the question of whether adaptability of species to urban landscapes correlates with phylogeny or rather functional ecology. Functional ecology of species can be linked to species traits, where traits refer to morphological, behavioural or physiological attributes of species (Violle etþ  al. 2007). Using such functional traits can improve understanding of and help predict how species respond to environmental change (Didham etþ  al. 1996; Flynn etþ  al. 2009), such as increasing urbanisation. A key challenge is to develop frameworks that can predict how the environment acts as a lter by advantaging or disadvantaging species with certain traits. Urbanisation has been demonstrated to select for, or against, species with specic response traits within ora and fauna communities, including remnant grasslands (Williams etþ  al. 2005), bat communities (Threlfall etþ  al. 2011) and bird communities (Evans etþ  al. 2011). To more fully understand and predict the impact of increasing urban land cover on urban bat communities, the identication and investigation of traits across a variety of studies in urban landscapes worldwide may prove useful. To do this, we investigated the response of bats to urbanisation using a functional ecology approach and fur ther investigated if these mechanisms are consistent worldwide and thus separately analysed the compiled literature for America (North and South America combined) versus Europe, Asia and Australia. Based on previous studies in urban and other human disturbed landscapes, we expected that predominant food item (fruits, nectar and insects), foraging mode (aerial, gleaning) and foraging space (narrow, edge and open, following Schnitzler and Kalko (2001)) may impact upon a species ability to adapt to urban environments, as suggested by (e.g. Avila-Flores and Fenton 2005; Jung and Kalko 2011; Threlfall etþ  al. 2011)2.2.1þ Data Acquisition and Meta-AnalysisWe used the Web of Knowledge (Thomson Reuter) to search for publications containing the following key words “bats” AND, “urban”, “urbanis(z)ation”, AND “gradient”, “community”, “assemblage”, “species composition”. This resulted in 99 studies reporting bat responses to urbanisation. In addition, we searched the reference list of these publications for further relevant articles. We compiled all studies focusing on bats in urban areas in our primary dataset. This selection also including different bat inventory methods such as acoustic monitoring, mist net and harp trap sampling as well as visual observations and roost surveys. In many of these articles however, quantitative data on bats were missing, sampling effort was not standardised, or studies did not reciprocally sample bats in urban versus natural areas. We excluded all of these studies from our nal dataset, as it

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21 2þ Urbanisation and Its Effects on Bats—A Global Meta-Analysiswas impossible to calculate a standard effect size of urbanisation. We thus only included studies into our nal meta-analysis that reported species-specic data on capture success, roosting individuals, occurrence counts or activity per sampling time in both urban and natural areas (Tableþ  2.1). In a few cases, we extracted data from graphs. We considered all of these measures as indicators of the relative intensity of habitat use and thus assumed comparability of these datasets and hence eligibility to be combined in a meta-analysis. Our nal data set for the metaanalysis consisted of 23 articles (Tableþ  2.1) and 96 bat species. Within this dataset we discriminated between studies with high (Nþ  þ  5) and intermediate intensity (Nþ  þ  5) of urbanisation following the individual authors’ statements in their articles (Tableþ  2.1). Our designation of ‘high’ and ‘intermediate’ was qualitative and based on descriptions of the urban study area from the original papers. For example, Avila-Flores and Fenton (2005) state that their study area of Mexico City is one of the “largest and most populated cities in the world”, hence we assigned this study a ‘high’ urban intensity. Gonsalves etþ  al. (2013) state that no quantication of urban intensity was made in their study, however they suggest that housing density in their study area was low and could be classied as suburban, hence we assigned this study an ‘intermediate’ urban intensity. This classication is by no means comprehensive, however we believe for comparative purposes these two classications give some indication and context of the intensity of urban development in the study area for each study used. Some articles (Nþ  þ  13) reported the response of bats to multiple intensities of urbanisation; here we extracted data on the highest, the lowest and the intermediate degrees of urbanisation. Data from urban parks, suburbia or small towns we considered as intermediate degrees of urbanisation. For each species reported in an article we compared the relative intensity of habitat use in urban (treatment group) versus natural areas (control group) and calculated the log odds ratio as a standardised effect size (Rosenberg etþ  al. 2000). A positive log odds ratioþ  >þ  0 indicated species that showed a higher intensity of habitat use in urban areas, while a negative log odds ratioþ  <þ  0 indicated higher intensity of habitat use in natural areas. For multiple reports on a species’ response to urbanisation in distinct articles we averaged the log odds ratios to avoid pseudoreplication. Species with incomplete identications were deleted from the dataset, except for Mormopterus species 2 (Australia) which has not yet been formally named (Adams etþ  al. 1988) and Eumops sp. (Panama) which most likely includes the two species Eumops glaucinus and Eumops auripendulus (Jung and Kalko 2011). For our analysis we thus considered each bat species (Nþ  þ  96) as a study case for our nal meta-analysis models. For all statistical analysis, we used the statistical software package R Version 2.1.4. (R Development Core Team 2011), package “metafor” (Viechtbauer 2013) (version 1.6-0). In a rst approach, we focused on the general response of bats to urbanisation and investigated if the overall response of bats depends on the degree of urbanisation. Hereby we distinguished between high and intermediate intensity of urbanisation (see above) and calculated log odds ratios for the respective contrast to natural areas. We then conducted a random effect model meta-analysis for the

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22 K. Jung and C.G. ThrelfallTableþ  2.1þ List of publications included in the meta-analyses Reference Country Urban intensity Study type N species urban N species suburban N species forest Survey method Considered habitat types Avila-Flores and Fenton (2005) Mexico High Urban gradient 2 3 4 Acoustic monitoring Residential areas, large parks, forest Basham etþ  al. (2010) Australia Intermediate Urban/forest NA 11 13 Acoustic monitoring Backyards, natural bushland Bihari (2004) Hungary High Urban/forest 1 NA 1 Roost survey Residential area, forested park Chirichella (2004) Italy High Urban gradient 1 1 1 Public survey Urban, suburban, forest Duchamp etþ  al. (2004) USA High Urban/forest 2 NA 2 Captures/ telemetry Urban, woods Fabianek etþ  al. (2011) Canada Intermediate Urban gradient NA 3 3 Acoustic monitoring Urban parksþ  <þ  100þ  ha, urban parksþ  >þ  100þ  ha Gaisler etþ  al. (1998) Czech Republic High Urban gradient 2 2 2 Acoustic monitoring Historical city centre, old suburbs, outskirts Gehrt and Chelsvig (2004) USA High Urban gradient 5 NA 4 Acoustic monitoring Urban index: 0 (urban) urban index: 5 (rural) Gonsalves etþ  al. (2013) Australia Intermediate Urban/forest NA 9 13 Acoustic monitoring Small urban, forest Hale etþ  al. (2013) United Kingdom High Urban gradient 2 2 2 Acoustic monitoring Dense urban, subur ban, rural Haupt etþ  al. (2006) Germany High Urban/forest 1 NA 1 Captures/ telemetry Urban areas, forest (before and after birth of juveniles Hourigan etþ  al. (2006) Australia Intermediate Urban gradient NA 8 8 Acoustic monitoring Intermediate suburbs, woodland Hourigan etþ  al. (2010) Australia High Urban gradient 13 13 13 Acoustic monitoring High and low-density residential, bush land (continued)

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23 2þ Urbanisation and Its Effects on Bats—A Global Meta-AnalysisGiven are the reference, the country where the study was carried out, the urban intensity (large or small) the study type (gradient, paired design), the number of bat species reported by each author for each habitat type, the survey methodology of the original data and the considered habitat types included in the meta-analysis. NA indicates that the corresponding study had no record of bats for the specied habitat type Tableþ  2.1þ (continued) Reference Country Urban intensity Study type N species urban N species suburban N species forest Survey method Considered habitat types Jung and Kalko (2010) Panama Intermediate Urban/forest NA 21 22 Acoustic monitoring Small town, forest Jung and Kalko (2011) Panama High Urban gradient 16 21 25 Acoustic monitoring Urban, small town, forest Kurta and Teramino (1992) USA High Urban/forest 4 NA 5 Captures Urban/rural Lesinski etþ  al. (2000) Poland High Urban gradient 3 3 3 Acoustic monitoring Central, suburban(III), suburban(V) Nadeem etþ  al. (2013) Pakistan High Urban gradient 4 4 3 Roost surveys Urban, suburban, rural dwelling Pottie etþ  al. (2005) Malaysia/ Singapore High Urban gradient 5 5 12 Roost sur veys/captures/ acoustic monitoring City/urban, suburban, primary forest Silva etþ  al. (2005) Brazil High Urban gradient 7 8 9 Acoustic monitoring Small farm, campus Funcesi, forest Threlfall etþ  al. (2011, 2012) Australia High Urban gradient 10 15 7 Acoustic monitoring Urban, suburban, forest Utthammachai etþ  al. (2008) Thailand High Urban gradient 1 1 1 Acoustic monitoring Forest patch, village/ others, urban Walters etþ  al. (2007) USA High Urban/forest 1 1 1 Captures Woodlots, lowdensity residential, commercial lands

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24 K. Jung and C.G. Threlfalleffect of high and intermediate urban development, respectively. Random effect models provide an unconditional inference of a larger set of studies from which only a few are included in the meta-analysis and assumed to be a random sample (Viechtbauer 2010). We compared both models based on the reported effect size and assessed the proportion of heterogeneity of bat responses between high and intermediate urban development (2 highly urban2 small urban/2 highly urban). In a second approach, we pooled data from high and intermediate urbanisation categories to investigate if the potential of bats to adjust to urban environments is determined by phylogeny or rather functional ecology using a mixed model metaanalysis. For this analysis we classied bats according to their taxonomic family and genus, their predominant food item (fruits, nectar and insects), foraging mode (aerial, gleaning) and foraging space (narrow, edge and open, following Schnitzler and Kalko (2001)) and included these classications as moderators in our mixed model meta-analysis. We further investigated in detail how each of the categorical moderators inuences effect size. Further, focusing on aerial insectivores, the majority of study cases in our dataset, we then investigated if moderators inuencing the adaptability to urban areas are consistent between North and South America versus Europe, Asia and Australia. P-levels for all models were assessed using a permutation test with 1000 randomizations. In none of our models did the funnel plot technique (Viechtbauer 2013) reveal any signicant publication bias or asymmetry in our dataset (function: regtest, package metaphor).2.2.2þ High Versus Lower Levels of Urbanisation Our random effect meta-analysis revealed that in general, urbanisation negatively affects bats, and areas with high (devianceþ  þ  453.14, z -valueþ  þ  3.9, pþ  <þ  0.001) and intermediate (devianceþ  þ  439.73; z -valueþ  þ  2.4, pþ  <þ  0.05) degrees of urban development reveal signicantly lower intensity of habitat use across species compared to natural areas (Fig.þ  2.1). A high degree of urbanisation had a stronger negative effect on the general intensity of habitat use (estimate: 1.47) than an intermediate degree of urban development (estimate: 0.79). However, in both high and intermediate urban development, we found signicant variation in the Effect size -2,0 -1,5 -1,0 -0,5 0,0 0,5 Urbanisation (high) Urbanisation (intemediate) -0.79 [-1.44, -0.13] -1.47 [-2.19, -0.73] Fig.þ  2.1þ Effect sizes of relative intensity of habitat use by bats in high and intermediate urban development, compared to natural areas. Solid symbols indicate the mean effect size (log odds ratio) and whiskers indicate the estimated standard error. Values of the estimated effect size, including the 95þ  % condence intervals are listed on the right side of the gure

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25 2þ Urbanisation and Its Effects on Bats—A Global Meta-Analysiseffect sizes (high urban development: Q(df 84)þ  þ  641.2, pþ  <þ  0.0001; intermediate urban development Q(df 85)þ  þ  989.9, pþ  <þ  0.0001), indicating a high variability in the response of bat species to urbanisation. This species-specic variability in the intermediate degree of urbanisation (2þ  þ  7.74) accounted for 21þ  % of the variability in the areas with high urban development (2þ  þ  9.80). This suggests that although intermediate urban development clearly has a negative inuence on bats it still permits the use of this habitat by more species showing fewer extremes in the species-specic response to urbanisation, compared to high urban development.2.2.3þ Phylogeny Versus Functional EcologyNeither phylogeny (QM(df3)þ  þ  11.57, pþ  >þ  0.05) nor functional ecology (QM(df3)þ  þ  12.18, pþ  >þ  0.05) explained the heterogeneity in bat response to urbanisation. However, a different pattern emerged when investigating the effect of single moderators in detail. Response to urbanisation differed between families (QM(df 10)þ  þ  32.4, pþ  þ  0.05) with bat species in the Rhinolophidae being negatively affected by urban development (pþ  <þ  0.01). In addition, bat species in the Mormoopidae tended to respond negatively towards urbanisation, as the 95þ  % condence interval did not overlap with zero. All other families revealed a high heterogeneity in the response to urbanisation. Effect size was neither genera— (QM(df46)þ  þ  81.4, pþ  >þ  0.05) nor species-specic (QM(df86)þ  þ  99.7, pþ  >þ  0.05). Effect size -4 -3 -2 -1 0123 Frugivore Insectivore Nectarivore Edge space Narrow space Open space Aerial Gleaning -0.99 [-3.44, 1.47] -1.16 [-1.83, -0.43] -0.16 [-4.44, 4.12] -0.96 [-1.81, -0.10] -2.55 [-4.18, -0.92] -0.72 [-1.84, 0.40] -1.05 [-1.73, -0.37] -1.62 [-3.44, 0.21] Fig.þ  2.2þ Effect of urbanisation (log odds ratio and the estimated standard error) on relative intensity of habitat use in relation to the predominant food item (a), foraging space (b), and foragþ­ ing mode (c). Solid symbols indicate the mean effect size (log odds ratio) and whiskers indicate the estimated standard error. Values of the estimated effect size, including the 95þ  % condence intervals are listed on the right side of the gure

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26 K. Jung and C.G. ThrelfallNone of the functional classications, food item, foraging mode and foraging space, revealed a signicant association with the persistence of bats in urban areas. However narrow space foragers (estimate 2.55þ  þ  0.83, pþ  þ  0.06) revealed a tendency to be associated with natural areas (Fig.þ  2.2). Europe, Asia, Australia Effect size -10 -8 -6 -4 -2 02 4 Emballonuridae Megadermatidae Miniopteridae Molossidae Nycteridae Rhinolophidae Vespertilionidae RE Model 1.50 [-1.53, 4.52] -3.22 [-9.50, 3.05] 1.42 [-6.78, 3.95] -0.71 [-3.13, 1.70] -2.20 [-8.56, 4.16] -6.59 [-9.84,-3.33] -0.00 [-1.17, 1.17] -0.64 [-1.68, 0.39] Northand South America Effect size -8 -6 -4 -2 02 4 Emballonuridae Molossidae Mormoopidae Noctiliondae Vespertilionidae RE Model -2.90 [-4.35, -1.44] -0.74 [-1.86, 0.38] -3.69 [-6.27, -1.11] 1.50 [-1.86, 4.86] -2.01 [-3.28, -0.75] -1.73 [-2.50, -0.96] (a) (b) Fig.þ  2.3þ Response of insectivorous bat families to urbanisation in a North and South America and b Europe, Asia and Australia. A negative effect size reects a higher association with natural areas, a positive effect size an association with urban areas. Depicted are the mean effect sizes (log odds ratio) and the estimated standard errors by family. Values of the estimated effect size, including the 95þ  % condence intervals are listed on the right side of the gure. The overall effect of urbanisation on insectivorous bats, based on the random effect model (RE Model), is given at the bottom of the respective gure

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27 2þ Urbanisation and Its Effects on Bats—A Global Meta-Analysis2.2.4þ Contrasting the Effects between North and South America and Europe, Asia and Australia Focusing on InsectivoresThe general response of insectivorous bats differed between the Americas and Europe, Asia and Australia. While insectivorous bats in the Americas revealed a signicant negative response to urbanisation (devianceþ  þ  171.18, z-valueþ  þ  4.4, pþ  <þ  0.001) the overall response of insectivorous bats to urbanisation in Europe, Asia and Australia was insignicant (devianceþ  þ  258.9, z-valueþ  þ  1.2, pþ  >þ  0.05, Fig.þ  2.3a, b). However, in both the Americas (QM(df5)þ  þ  35.1, pþ  <þ  0.05) and Europe, Asia and Australia (QM(df7)þ  þ  18.7, pþ  <þ  0.05) the response to urbanisation differed signicantly across families. Interestingly this family-level response was inconsistent between the Neoand Paleotropics. While Neotropical bats in the Emballonuridae showed a strong tendency to be associated with natural areas (estimate: 2.9þ  þ  0.7, pþ  þ  0.06), emballonurids in the Paleotropics (estimate: 1.5þ  þ  1.5, pþ  >þ  0.05) occurred frequently in urban areas. We found a similar trend in the globally distributed family of Vespertilionidae, which showed a higher association with natural areas in the Americas (estimate: 2.0þ  þ  0.6, pþ  >þ  0.05) but did not reveal any clear association in Europe, Asia and Australia (estimate: 0.0þ  þ  0.6, pþ  >þ  0.05) (Fig.þ  2.3a, b).2.3þ Adaptability of Species to Urban Areas: General Trends, Species-Specic Differences and Future ResearchUrban areas can provide suitable habitat for a variety of species, albeit an anthropogenically altered habitat (McKinney 2006). However, our general understanding of what inuences a species’ success in urban environments is limited. Arguably the conservation of species such as bats in urban areas is dependent upon this knowledge (Fenton 1997). Within this book chapter, we reviewed the existing literature on bats in urban areas. In addition, we combined published data in a metaanalysis to evaluate and derive general patterns in the response of bats to urban development. Our meta-analysis revealed that, in general, habitat use of bats decreases in urban areas. A high degree of urbanisation had a stronger negative effect on overall habitat use of bats compared to an intermediate degree of urban development. However, habitat use in intermediate urban development was much lower compared with natural areas. This is alarming, as it is generally thought that small towns and suburban landscapes could potentially provide suitable habitat for a wide range of species (McKinney 2006), including bats. The combination

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28 K. Jung and C.G. Threlfallof habitats with different complexity in smaller urban developments should lead to greater complementarity at a local scale and should favour species diversity and abundance. Some of the publications in our meta-analysis dataset indeed report a higher bat diversity, activity (Hourigan etþ  al. 2010; Threlfall etþ  al. 2011, 2012b) and feeding activity (Jung and Kalko 2011; Threlfall etþ  al. 2012a) at intermediate levels of disturbance compared to natural or urban habitats. Other studies reported that any urban land cover, even if low-density residential, can decrease bat activity and species richness (Hourigan etþ  al. 2006; Gonsalves etþ  al. 2013; Luck etþ  al. 2013), and even deter individual species (Jung and Kalko 2010; Gonsalves etþ  al. 2013; Luck etþ  al. 2013). Altogether, this strongly suggests regional differences in the intensity of urban development and points towards an interacting effect of the surrounding landscape (see Coleman and Barclay 2011). Results from recent urban bat studies suggest that bats of some families (e.g. molossids Jung and Kalko 2011) are better pre-adapted for life in an urban environment compared to others (e.g. rhinolophids Stone etþ  al. 2009; Threlfall etþ  al. 2011). Our analysis also indicated a family-specic effect of urbanisation and conrmed the negative response of Rhinolophidae to urban development across the Old World. However, the responses of Molossidae and Vespertilionidae, which are known to frequently roost in man-made structures in North and South America, did not reveal consistent associations with either urban or natural areas across continents. This might be due to the high morphological and behavioural heterogeneity within these families. We believe that the likely explanation for our results is that the response to urbanisation is dictated by the behavioural and morphological traits of species, regardless of geographic region or phylogeny. In particular, species foraging in open space seem to persist in urban areas, as due to their wing morphology (high aspect ratio and wing loading) they might be able to commute large distances between roosting sites and feeding areas (Jung and Kalko 2011). Thus traits predicting species mobility have been associated with urban tolerance (Jung and Kalko 2011; Threlfall etþ  al. 2012a), and the ability to forage around street lights (see Rowse etþ  al., Chap.þ  7 this volume). In addition, traits that allow for exible roost and foraging strategies confer an advantage for urban-tolerant species. Our current results support these ndings and thus suggest that adaptability of bats to urban environments (or disturbance in general) might be correlated with, and reected by, species behavioural exibility. Advancement of knowledge in this area will assist with conservation efforts of bat species globally, and potentially allow development of a predictive framework for assessing the impacts of urban development on bats. Open Access This chapter is distributed under the terms of the Creative Commons Attribution Noncommercial License, which permits any noncommercial use, distribution, and reproduction in any medium, provided the original author(s) and source are credited.

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31 2þ Urbanisation and Its Effects on Bats—A Global Meta-Analysis Kalcounis-Rueppell MC, Payne VH, Huff SR, Boyko AL (2007) Effects of wastewater treatment plant efuent on bat foraging ecology in an urban stream system. Biol Conserv 138(1–2):120–130 Kössl M, Mora E, Coro F, Vater M (1999) Two-toned echolocation calls from Molossus molossus in Cuba. J Mammal 80(3):929–932 Kurta A, Teramino JA (1992) Bat community structure in an urban park. Ecography 15:257–261 Lane DJW, Kingston T, Lee BPYH (2006) Dramatic decline in bat species richness in Singapore, with implications for Southeast Asia. Biol Conserv 131(4):584–593 Legakis A, Papdimitriou C, Gaetglich M, Lazaris D (2000) Survey of bats of the Athens metropolitan area. Myotis 38:41–46 Lesiñski G, Eb Fuszara, Kowalski M (2000) Foraging areas and relative density of bats (Chiroptera) in differently human transformed landscapes. Z Säugetierkunde 65:129–137 Longcore T, Rich C (2004) Ecological light pollution. Front Ecol Environ 2(4):191–198 Luck GW, Smallbone L, Threlfall C, Law B (2013) Patterns in bat functional guilds across multiple urban centres in south-eastern Australia. Landscape Ecol 28:455–469 Luniak M (2004) Synurbization—adaptation of animal wildlife to urban development. In: Shaw WW, Harris LK, VanDruff L (eds) Proceedings 4th international urban wildlife symposium. University of Arizona, Tucson, USA, pp 50–55 Macdonald DW, Newdick MT (1982) The distribution and ecology of foxes, Vulpes vulpes (L.) in urban areas. In: Bornkamm R, Lee JA, Seaward MRD (eds) Urban ecology. Oxford University Press, Oxford, pp 123–138 Mahmood-ul-Hassan M, Gulraiz TL, Rana SA, Javid A (2010) The diet of Indian ying-foxes (Pteropus giganteus) in urban habitats of Pakistan. Acta Chiropterologica 12(2):341–347 McConville A, Law B, Penman T, Mahony M (2013a) Contrasting habitat use of morphologically similar bat species with differing conservation status in south-eastern Australia. Austral Ecol 39(1):83–94 McConville A, Law BS, Mahony MJ (2013b) Are regional habitat models useful at a local-scale? A case study of threatened and common insectivorous bats in south-eastern Australia. PLoS ONE 8(8):1–10 McDonald RI (2008) Global urbanization: can ecologists identify a sustainable way forward? Front Ecol Environ 6(2):99–104 McDonald R, Marcotullio P (2013) Global effects of urbanization on ecosystem services. In: Elmqvist T, Fragkias M, Goodness J etþ  al (eds) Urbanization, biodiversity and ecosystem services: challenges and opportunities. Springer, Dordrecht McDonald RI, Kareiva P, Forman RTT (2008) The implications of current and future urbanization for global protected areas and biodiversity conservation. Biol Conserv 141(6):1695–1703 McDonnell MJ, Hahs AK (2008) The use of gradient analysis studies in advancing our under standing of the ecology of urbanizing landscapes: current status and future directions. Landscape Ecol 23(10):1143–1155 McDonnell M, Hahs A (2013) The future of urban biodiversity research: moving beyond the ‘low-hanging fruit’. Urban Ecosyst 16(3):397–409 McDonnell MJ, Pickett STA (1990) Ecosystem structure and function along urban-rural gradients: an unexploited opportunity for ecology. Ecology 71(4):1232–1237 McGranahan G, Satterthwaite D (2003) Urban centers: an assessment of sustainability. Annu Rev Environ Resour 28:243–274 McIntyre NE, Knowles-Yánez K, Hope D (2000) Urban ecology as an interdisciplinary eld: differences in the use of “urban” between the social and natural sciences. Urban Ecosyst 4(1):5–24 McKinney ML (2002) Urbanization, biodiversity, and conservation. Bioscience 52(10):883–890 McKinney ML (2006) Urbanization as a major cause of biotic homogenization. Biol Conserv 127(3):247–260

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32 K. Jung and C.G. Threlfall McKinney ML, Lockwood JL (1999) Biotic homogenization: a few winners replacing many losers in the next mass extinction. Trends Ecol Evol 14(11):450–453 Meyer CFJ, Kalko EKV (2008) Assemblage-level responses of phyllostomid bats to tropical for est fragmentation: land-bridge islands as a model system. J Biogeogr 35(9):1711–1726 Montgomery MR (2008) The urban transformation of the developing world. Science 319(5864):761–764 Myers N, Mittermeier RA, Mittermeier CG, da Fonseca GAB, Kent J (2000) Biodiversity hotspots for conservation priorities. Nature 403:853–858 Nadeem MS, Zafar S, Kayani AR, Mushtaq M, Beg MA, Nasir MF (2013) Distribution and roosting habitats of some Microchiropteran bats in Rawalpindi District, Pakistan. Pak J Zool 45(2):565–569 Nakamoto A, Kinjo K, Izawa M (2007) Food habits of Orii’s ying-fox, Pteropus dasymallus inopinatus, in relation to food availability in an urban area of Okinawa-jima Island, the Ryukyu Archipelago, Japan. Acta Chiropterologica 9(1):237–249 Nakamoto A, Kinjo K, Izawa M (2012) Ranging patterns and habitat use of a solitary ying fox (Pteropus dasymallus) on Okinawa-jima Island, Japan. Acta Chiropterologica 14(2):387–399 Norberg UM, Rayner JMV (1987) Ecological morphology and ight in bats (Mammalia; Chiroptera): wing adaptations, ight performance, foraging strategy and echolocation. Philos Trans R Soc Lond B Biol Sci 316(1179):335–427 Pacheco SM, Sodré M, Gama AR, Bredt A, Cavallini EM, Sanches Marques RV, Guimarães MM, Bianconi G (2010) Morcegos Urbanos: Status do Conhecimento e Plano de Ação para a Conservação no Brasil. Chiroptera Neotropical 16(1):629–647 Park KJ, Mochar F, Fuentes-Montemayor E (2012) Urban biodiversity: successes and challenges: bat activity in urban green space. Glasg Naturalist 25(4) Parris KM, Hazell DL (2005) Biotic effects of climate change in urban environments: the case of the grey-headed ying-fox (Pteropus poliocephalus) in Melbourne, Australia. Biol Conserv 124(2):267–276 Pearce H, Walters CL (2012) Do green roofs provide habitat for bats in urban areas? Acta Chiropterologica 14(2):469–478 Pickett STA, Cadenasso ML, Grove JM, Nilon CH, Pouyat RV, Zipperer WC, Costanza R (2001) Urban ecological systems: linking terrestrial ecological, physical, and socioeconomic components of metropolitan areas. Ann Rev Ecol Syst 32:127–157 Pottie SA, Lane DJW, Kingston T, Lee BPYH (2005) The microchiropteran bat fauna of Singapore. Acta Chiropterologica 7(2):237–247 Rhodes M, Catterall C (2008) Spatial foraging behaviour and use of an urban landscape by a fast-ying bat, the Molossid Tadarida australis. J Mammal 89(1):34–42 Rhodes M, Wardell-Johnson G (2006) Roost tree characteristics determine use by the whitestriped freetail bat (Tadarida australis, Chiroptera: Molossidae) in suburban subtropical Brisbane, Australia. Austral Ecol 31:228–239 Rhodes M, Wardell-Johnson GW, Rhodes MP, Raymond BEN (2006) Applying network analysis to the conservation of habitat trees in urban environments: a case study from Brisbane. Australia. Conservation Biology 20(3):861–870 Rompré G, Robinson WD, Desrochers A (2008) Causes of habitat loss in a neotropical landscape: the Panama Canal corridor. Landscape and Urban Planning 87(2):129–139 Rosenberg MS, Adams DC, Gurevitch J (2000) Metawin: manual of statistical software for metaanalysis. Sinauer, Sunderland Scales J, Wilkins KT (2007) Seasonality and delity in roost use of the mexican free tailed bat Tadarida brasiliensis in an urban setting. Western North American Naturalist 67(3):402–408 Scanlon AT, Petit S (2008) Effects of site, time, weather and light on urban bat activity and richness: considerations for survey effort. Wildlife Research 35(8):821–834 Schnitzler H-U, Kalko EKV (2001) Echolocation by insect-eating bats. Bioscience 51(7):557–569

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33 2þ Urbanisation and Its Effects on Bats—A Global Meta-Analysis Septon G, Marks JB, Ellestad T (1995) A preliminary assessment of peregrine Falcon Falco per egrinus recovery in the Midwestern North America. Acta Ornithologica 30:65–68 Sewell SR, Catterall CP (1998) Bushland modication and styles of urban development: their effects on birds in south-east Queensland. Wildlife Research 25(1):41–63 Shochat E, Warren PS, Faeth SH, McIntyre NE, Hope D (2006) From patterns to emerging processes in mechanistic urban ecology. Trends Ecol Evol 21(4):186–191 Siles L, Peñaranda D, Pérez-Zubieta JC, Barboza K (2005) Los murciélagos de la ciudad de Cochabamba. Revista Boliviana Ecologia 18:51–64 Slabbekoorn H, Peet M (2003) Birds sing at a higher pitch in urban noise. Nature 424(6946):267 Slabbekoorn H, Ripmeester EAP (2008) Birdsong and anthropogenic noise: implications and applications for conservation. Mol Ecol 17:72–83 Stone EL, Jones G, Harris S (2009) Street lighting disturbs commuting bats. Current Biology 19:1123–1127 Thomas NM, Duckworth J, Douangboubpha B, Williams M, Francis CM (2013) A checklist of bats (Mammalia: Chiroptera) from Lao PDR. Acta Chiropterologica 15(1):193–260 Threlfall CG, Law B, Penman T, Banks PB (2011) Ecological processes in urban landscapes: mechanisms inuencing the distribution and activity of insectivorous bats. Ecography 34(5):814–826 Threlfall CG, Law B, Banks PB (2012a) Inuence of landscape structure and human modications on insect biomass and bat foraging activity in an urban landscape. PLoS ONE 7(6):e38800 Threlfall CG, Law B, Banks PB (2012b) Sensitivity of insectivorous bats to urbanization: Implications for suburban conservation planning. Biol Conserv 146:41–52 Threlfall CG, Law B, Banks PB (2013a) Roost selection in suburban bushland by the urban sensitive bat Nyctophilus gouldi. J Mammal 94(2):307–319 Threlfall CG, Law B, Banks PB (2013b) The urban matrix and articial light restricts the nightly ranging behaviour of Gould’s long-eared bat (Nyctophilus gouldi). Austral Ecol 38(8):921–930 UN (2012) System Task Team on the post-2015 UN development agenda: sustainable urbanization. Thematic think piece UNPD (2012) World urbanization prospects: the 2011 revision. Highlights. United Nations Population Division, New York, USA Utthammachai K, Bumrungsri S, Chimchome V, Russ J, Mackie I (2008) The Habitat Use and Feeding Activity of Tadarida plicata in Thailand. Thai J For 27(2):21–27 van der Ree R, McCarthy MA (2005) Inferring persistence of indigenous mammals in response to urbanisation. Anim Conserv 8(3):309–319 Viechtbauer W (2010) Conducting meta-analysis in R with the metafor package. J Statistical Softw 36(3):1–48 Viechtbauer W (2013) The metafor package: a meta-analysis package for R. http://www. metafor-project.org Violle C, Navas M-L, Vile D, Kazakou E, Fortunel C, Hummel I, Garnier E (2007) Let the concept of trait be functional! Oikos 116(5):882–892 Vitousek PM, Mooney HA, Lubchenco J, Melillo JM (1997) Human domination of earth’s ecosystems. Science 277(5325):494–499 Walsh AL, Harris S (1996) Factors determining the abundance of Vespertilionid bats in Britain: geographical, land class and local habitat relationships. J Appl Ecol 33(3):519–529 Williams NSG, Morgan JW, McDonnell MJ, McCarthy MA (2005) Plant traits and local extinctions in natural grasslands along an urban–rural gradient. J Ecol 93(6):1203–1213 Zhao S, Da L, Tang Z, Fang H, Song K, Fang J (2006) Ecological consequences of rapid urban expansion: Shanghai, China. Front Ecol Environ 4(7):341–346

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35Chapter 3Bats and RoadsJohn Altringham and Gerald Kerth© The Author(s) 2016 C.C. Voigt and T. Kingston (eds.), Bats in the Anthropocene: Conservation of Bats in a Changing World, DOIþ  10.1007/978-3-319-25220-9_3Abstractþ The effects of roads on bats have been largely neglected until recently, despite growing evidence for profound effects on other wildlife. Roads destroy, fragment and degrade habitat, are sources of light, noise and chemical pollution and can kill directly through collision with trafc. The negative effects of roads on wildlife cannot be refuted but at the same time road building and upgrading are seen as important economic drivers. As a consequence, infrastructure projects and protection of bats are often in conict with each other. There is now growing evidence that fragmentation caused by roads reduces access to important habitat, leading to lower reproductive output in bats. This barrier effect is associated with reduced foraging activity and species diversity in proximity to motorways and other major roads. The effects of light and noise pollution may add to this effect in the immediate vicinity of roads and also make bats even more reluctant to approach and cross roads. Several studies show that vehicles kill a wide range of bat species and in some situations roadkill may be high enough to lead directly to population decline. Current mitigation efforts against these effects are often ineffective, or remain largely untested. The limited information available suggests that underpasses to take bats under roads may be the most effective means of increasing the safety and permeability of roads. However, underpass design needs further study and alternative methods need to be developed and assessed. J. Altringhamþ  ()þ  School of Biology, Faculty of Biological Sciences, University of Leeds, Leeds, UK e-mail: J.D.Altringham@leeds.ac.uk G. Kerthþ  Applied Zoology and Conservation, Zoological Institute and Museum, University of Greifswald, Greifswald, Germany e-mail: gerald.kerth@uni-greifswald.de

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36 J. Altringham and G. Kerth3.1þ IntroductionThe global road network gets longer, wider, faster and more complex as existing road systems are upgraded and new roads are built. Despite the widely acknowledged need to reduce our dependence on fossil fuel and growing concerns about the environmental impact of roads, improved communication by road, and even the act of road-building itself, are often seen as essential economic drivers. As road networks expand, trafc volumes increase and congestion remains a problem. A few statistics highlight the pervasive nature of our road networks: only 2þ  % of Germany is made up of landscape fragments greater than 100þ  km2 (Jaeger etþ  al. 2007) and only 17þ  % of the US landscape is more than 1þ  km from a road (Riiters and Wickham 2003). In 2012, the UK had 395,000þ  km of roads, of which over 50,000þ  km are major roads and 3700þ  km motorways (Defra 2013). Major roads account for only 13þ  % of all UK roads, but carry 65þ  % of the trafc. 50þ  % of all trafc is on motorways and other major roads in rural areas. Almost 20þ  % of major road length is dual carriageway. Over 3200þ  km have been added to the UK network in the last decade and many more have been upgraded. Roads have several negative impacts on animals. First, building roads and their ancillary structures destroys habitat directly. Secondly, the resulting road network fragments the landscape, potentially restricting animal movements, thereby blocking their access to the remaining habitat. Thirdly, roads are also sources of light, noise and chemical pollution, and so degrade the habitat around them. Moreover, the increased human access provided by roads usually accelerates urban, commer cial and agricultural development and increases human disturbance in many ways, e.g. through increased recreational pressure and the introduction of non-native predators and other invasive species. Finally, fast moving trafc kills animals directly. Broad reviews of the effects of roads on vertebrates include Bennett (1991), Forman and Alexander (1998), Trombulak and Frissell (2000), Cofn (2007), Fahrig and Rytwinski (2009), Laurance etþ  al. (2009), Benítez-L pez etþ  al. (2010), and Rytwinski and Fahrig (2012). Surprisingly, despite the many ways in which roads can impact on wildlife, it is only in the last 20þ  years that signicant attention has been given to what is now often referred to as ‘road ecology’ (Forman etþ  al. 2003). Little of this attention was directed at bats. Moreover, the few existing studies on the impact of roads on bats have all been carried out in North America and Europe. Globally many bat species are endangered (Racey and Entwistle 2003; Jones etþ  al. 2009), including regions with a dense infrastructure such as North America and Europe (Sa and Kerth 2004). As a consequence, in Europe, for example, bats are of high priority for conservation and all bat species have been strictly protected for two decades by European law (CMS 1994). Despite the importance of bats in conservation, rigorous, peer-reviewed studies on the impact of roads on bats have only begun to be published in the last few years. Only over the last decade it has been widely accepted that roads must have an effect on bats. As a result, mitigation against these effects is becoming increasingly integrated in the road building process and practical mitigation guidelines have been published in a number of countries (e.g. Highways Agency 2001, 2006; Limpens etþ  al. 2005). However, the

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37 3þ Bats and Roadsprecise nature and scale of the effects of roads on bats were mostly unknown, and as a consequence mitigation has often been poorly monitored and therefore rarely informed by sound evidence (Altringham 2008; O’Connor etþ  al. 2011). This review describes the ways in which roads do or may affect bats, discusses the available evidence in relation to each, and where appropriate suggests action for the future, in terms of both research and conservation action. Because work on the impacts of roads on bats is still scarce and biased towards the temperate zone, some work on other animals will be discussed, in particular birds, to help ll important gaps. Roads can affect bats in many ways, and because the mitigation solutions will to some extent be unique to each, the mechanisms will be discussed separately. However, there is considerable interaction between them and the impacts in many cases are cumulative, so some topics will appear under more than one heading. To our knowledge almost no studies have been published yet that investigated the effects of railways on bats (but see Vandevelde etþ  al. 2014). However, as linear development features, they have the potential to disrupt bats and will be discussed briey at the end of the review.3.1.1þ Bat Life HistoryIn order to assess the impact of roads on bats, an important consideration is of course the biology of the bats themselves. Bats are small mammals with the life history strategy of very much larger species (e.g. Barclay and Harder 2003; Altringham 2011). They have taken the low fecundity, long life option, often producing only a single pup each year, but frequently living for more than 10þ  years and not unusually 20 or more (e.g. Barclay and Harder 2003; Altringham 2011). Any external factors that reduce reproductive success, increase mortality, or both, can lead to severe population declines—and recovery will be slow (e.g. Sendor and Simon 2003; Papadatou etþ  al. 2011). Furthermore, bats typically have large summer home ranges compared to other similar sized mammals and many bats migrate over considerable distances between winter and summer roosts (Altringham 2011). Finally, bats are highly gregarious (Kerth 2008). As a result, negative impacts of roads on local bat colonies can affect large numbers of individuals simultaneously. Because of their particular life history, bats are susceptible to a wider range of environmental disturbances than many other small mammals.3.1.2þ Bat Conservation StatusA substantial number of the more than 1200 extant bat species are considered to be endangered (Racey and Entwistle 2003; Jones etþ  al. 2009). Reasons for the decline of bats include habitat loss, pollution, direct persecution and diseases (Jones etþ  al.

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38 J. Altringham and G. Kerth2009). Several of these threat factors are also relevant during the construction and maintenance of roads. In Europe, all bats are strictly protected, as all are listed in Annex 4 of the Habitats Directive, and several species have designated protected areas because they are also listed in the Annex 2 of the Habitats Directive (Council Directive 92/43/EEC). As a consequence, whenever bat populations are likely to be adversely affected by the construction of roads, environmental assessments are required and mitigation often becomes a necessity. Thus assessments of bats have been carried out during many recent infrastructure projects (e.g. Kerth and Melber 2009) and this process will continue to be important in the future.3.2þ The Effects of Roads on Bats—Habitat Destruction, Fragmentation, Degradation and Collision Mortality 3.2.1þ Loss of HabitatRoad development frequently involves the removal of trees and buildings that hold potential or actual bat roosts. The removal of trees, hedges, scrub, water bodies and unimproved (‘natural’) grassland also reduces available foraging habitat. The road surface alone destroys signicant areas of habitat: 7þ  ha for every 10þ  km of 7þ  m wide, two-lane road. Roadside hard shoulders, verges, junctions, service areas and other structures remove yet more potential habitat. As a result, road construction leads to the permanent loss of habitats for bats and thus is likely to reduce population sizes directly.3.2.2þ The Barrier EffectRoads are potential barriers to ight between roosts and foraging sites and between summer, mating and winter roosts. They could therefore reduce the available home range size and quality and may restrict migration, which could increase mortality and reduce reproductive potential. Roads may act as barriers because they interrupt existing linear ight lines, because some species are reluctant to cross open ground, because some species avoid lit areas (road and vehicle lights) and, at least initially, because they represent sudden changes in the bats’ familiar landscape. Roads may therefore fragment habitat, decreasing its accessible area and quality. Since habitat area and quality are major determinants of population size, then habitat fragmentation will lower the sustainable population size. Barriers such as roads may also limit the ow of individuals between populations with two major consequences. First, barriers may slow the recovery from local population declines since recruitment of individuals from neighbour ing populations (“rescue effect”) will be reduced and this will further increase

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39 3þ Bats and Roadsthe probability of local extinction. Secondly, barriers may also reduce gene ow between populations and increase inbreeding, reducing individual tness and increasing the risk of local extinction. Genetic isolation such as this can only occur with very low levels of dispersal. These factors may only be signicant for rare bat species that already have small and fragmented populations. Of course it may be that they are rare because of their susceptibility to these and other anthropogenic pressures. Genetic isolation as a direct result of roads has not been studied in bats. In several other mammal species an effect of roads on genetic population structure has been found (Frantz etþ  al. 2012). For example, Gerlach and Musolf (2000) have shown that populations of bank vole are genetically different either side of a fourlane highway. However, even in bat species such as Bechstein’s bat, Myotis bechsteinii, for which barrier effects of motorways haven been shown to occur in the summer habitat (Kerth and Melber 2009), local populations living in an area with several motorways show only weak genetic differentiation (Kerth etþ  al. 2002; Kerth and Petit 2005). In accordance with the ndings on Bechstein’s bats, population genetic studies on other temperate zone bats typically found no or very little evidence for genetic isolation on the regional scale (Moussy etþ  al. 2013), despite the dense road network in Europe and North America. This suggests that in the temperate zone roads probably have no signicant effect on gene ow in most bat species. For tropical bats much less data on population genetic structures are available but the situation may be different from the temperate zone. In general, mammal and bird species living in tropical rainforests are often particularly reluctant to cross open areas (Laurance etþ  al. 2009). Moreover, unlike most bats in Europe and North America, tropical bats often mate close to or at the breeding sites of the females. Both features make tropical bats likely to suffer more from fragmentation by roads by means of restricted gene ow than temperate zone bat species. Clearly, further studies are needed to test this. There is considerable evidence to suggest that roads act as barriers to bats dur ing foraging and movements between different day roosts (roost switching) in the summer habitat. Bats have been shown to make major detours to avoid roads or to nd appropriate crossing points (e.g. Kerth and Melber 2009). This behaviour could lead to longer journeys that consume time and energy or even deny bats access to parts of their habitat. In the study by Kerth and Melber (2009) of 32 radiotracked, female Bechstein’s bats, only three individuals, belonging to two different maternity colonies, crossed a four-lane motorway cutting through a German forest to forage (Fig.þ  3.1). All three bats used an underpass to cross the motor way. Other bats from four nearby colonies did not cross the motorway. Moreover, during roost switching none of the colonies crossed the motorway. In addition, foraging areas of females were smaller in those colonies whose home range was bounded by the motorway, relative to those bounded by more natural forest edges. Importantly, females in colonies bounded by the motorway had lower reproductive success than other females, persuasive evidence for the adverse effects on reproductive output. In the same study, six barbastelle bats, Barbastella barbastellus, belonging to one maternity colony, were also tracked and ve made several ights

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40 J. Altringham and G. Kerth over the road itself (Fig. 3.1 ). Moreover, the barbastelle bat colony used roosts on both sides of the motorway. These ndings highlight the fact that the effects of roads are species-specic, as will be discussed in more detail later. Berthinussen and Altringham ( 2012a ) observed only three bats ying over a six-lane motorway, all belonging to Nyctalus species, at heights above 20 m. Nyctalus species are Fig. 3.1 Home range use of two forest bat species living close to a motorway in Germany. The upper picture shows the polygons depicting the individual foraging areas of 32 BechsteinÂ’s bats belonging to six different colonies living in a German forest that is cut by a motorway. The lower picture shows the polygons depicting individual foraging areas of six barbastelle bats belonging to one colony living in the same forest as the BechsteinÂ’s bat colonies. From Kerth and Melber ( 2009 )

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41 3þ Bats and Roadsknown to y high and to forage in open spaces (e.g. Jones 1995), behaviour that is likely to make them less susceptible to the barrier effects of roads and to collision mortality. The absence of other species of bat ying over the road in this study suggests that the severance of linear elements by the road may have caused the abandonment of previous ight lines. Roads may be perceived as barriers by bats for several reasons: open spaces and articial light expose them to predation, and moving trafc and noise may be seen as threats. Small gaps (<5þ  m) in cover along ight routes can interrupt commuting bats (e.g. Bennett and Zurcher 2013), but many species will cross open spaces, even those adapted to forage in woodland (e.g. Kerth and Melber 2009; Abbott 2012; Abbott etþ  al. 2012a; Berthinussen and Altringham 2012b), although they will typically do so close to the ground (e.g. Russell etþ  al. 2009; Abbott 2012; Abbott etþ  al. 2012a; Berthinussen and Altringham 2012b). Abbott etþ  al. (2012a) observed low-ying species crossing at sites where mature hedgerows had been severed by the road, even when the gap was >50þ  m. However, Abbott (2012) found that the rate of bat crossing decreased with increasing distance between mature hedgerows on opposite sides of the road, suggesting a greater barrier effect. Russell etþ  al. (2009) reported that reduced cover at the roadside reduced the number of crossing bats. That some bats will cross roads is not an indicator that open roads are not a problem—the proportion of bats that do cross may be very small and they are at risk of collision with trafc. The presence of trafc does appear to have a direct effect on the likelihood of crossing, since Indiana bats, Myotis sodalis, reverse their ight paths and exhibit anti-predator avoidance behaviour in response to approaching vehicles (Zurcher etþ  al. 2010; Bennett and Zurcher 2013). No specic study has been made of crossing behaviour in relation to trafc volume and road width but anecdotal evidence suggests that it matters. For example, in the study of Kerth and Melber (2009) an individual Bechstein’s bat that ew over a two-lane road did only cross a four-lane highway through an underpass. Light and noise are discussed below. Evidence for a barrier effect is seen in other studies. Berthinussen and Altringham (2012a) found that total bat activity, the activity of the most abundant species (Pipistrellus pipistrellus) and the number of species, were all positively correlated with distance from a 40þ  year-old, six-lane, unlit motorway in rural north-west England (30–40,000þ  vehicles/day). Total activity increased more than threefold between 0 and 1600þ  m from the road. These effects were consistent over the two years of study and similar results were obtained on a rural motorway in south-west England (25–90,000þ  vehicles/day) (Berthinussen 2013). Unpublished work (A. Berthinussen and J.D. Altringham, in preparation) shows that this effect can extend to single carriageway (two-lane) roads. The most likely explanation for this spatially extensive reduction in bat activity is a long-term barrier effect, possibly in combination with increased mortality, driving colonies away from the road, and this is discussed further below.

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42 J. Altringham and G. Kerth 3.2.3 Roadkill Bats that attempt to cross roads risk collision, and hotspots for mortality have been found where yways cross roads and where there is favourable habitat for bats on both sides of a road (e.g. Lesi ski 2007 ; Russell et al. 2009 ; Medinas et al. 2013 ). Although agile and manoeuvrable in ight, most bat species y at low speeds (<20 km/h) and many y close to the ground (0–4 m: e.g. Russell et al. 2009 ; Berthinussen and Altringham 2012b ), particularly when crossing open spaces. In contrast to the majority of birds, most bats also spend most of the time they are out of the roost in ight. They make extensive use of linear landscape features, such as woodland edges and hedgerows along roads, for foraging and as navigational aids when commuting and several recent studies have shown how important these linear elements are to bats (e.g. Boughey et al. 2012 ; Frey-Ehrenbold et al. 2013 ; Bellamy et al. 2013 ). Flying close to such edges may also reduce predation risk. In combination, these behavioural traits make bats highly vulnerable to moving vehicles when either foraging along roads or when attempting to cross roads on commuting ights. Being small, bats can probably be pulled easily into the slip stream of passing vehicles. Russell et al. ( 2009 ) watched over 26,000 bat cross ings (primarily little brown bats, Myotis lucifugus ) on a highway in the USA. Bats approached the road using tree canopy cover and fewer bats were recorded cross ing where cover was absent. The lower the cover, the lower the bats crossed the road. Where bats were forced to cross an open eld on leaving the roost most did so at a height of less than 2 m. Berthinussen and Altringham ( 2012b ) recorded bats of four or more species crossing roads at mean heights well below 5 m (Fig. 3.2 ). Fig. 3.2 Boxplot of ight height above verge height of identied crossing bats. Median with upper and lower quartiles. Signicant differences shown for Myotis and Pipistrellus species ** P < 0.0005,*** P < 0.0001. Verges are elevated on either side of the road and are above road height, therefore negative values indicate bats ying across the road below the height of the verge. From Berthinussen and Altringham ( 2012b )

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43 3þ Bats and RoadsLesinski (2007) recorded bat casualties on an 8þ  km section of two-lane highway by weekly searches for carcasses over four summers. Casualties ranged from 0.3þ  bats/km/year in built-up areas to 6.8þ  bats/km/year where roads were bordered by trees. However, a study by Slater (2002) of the rate of removal of ‘carcasses’ (small pieces of chicken!) by scavengers on Welsh roads, suggests that a census of this kind may underestimate wildlife road kills as much as 12–16 fold, since dawn scavengers typically removed small carcasses within 30þ  min. More recently Santos etþ  al. (2011) have also shown that bat carcasses persist on roads in Portugal for a similarly brief period due to scavenging. Teixeria etþ  al. (2013) studied roads in Brazil and found that roadkill estimates increased 2–40 fold when scavenging and low detectability were accounted for. This wide variation was due to taxonomic differences and bats would be at the high end of this range. In addition, small bat carcasses are difcult to spot and many will be thrown clear of the road or carried some distance on the vehicle, suggesting that underestimates will be even greater. Arnett (2006) found that humans (in the absence of scavengers) were able to nd only 14 and 42þ  % of bat carcasses placed at two wind farm sites and Mathews etþ  al. (2013) reported that humans found only 20þ  % of bat carcasses at wind farms, relative to 73þ  % found by dogs. Road mortality studies will therefore inevitably under-estimate true mortality rates. A signicant proportion of European bat species, occupying a range of ecological niches, have been documented as roadkill (e.g. Billington 2001–2006; Lesiski 2007; Lesiski etþ  al. 2010). Woodland-adapted species should be most affected due to their characteristic low and slow ight, but this prediction was not supported by Lesiski etþ  al. (2010), as noctules (Nyctalus noctula) were killed in signicant numbers. Clearly other factors can play an important role locally. Forman etþ  al. (2003, pp 120–122) show that wildlife collisions increase as vehicle speed and trafc volume increase, and with proximity to wildlife habitat and wildlife movement corridors. There are no data on bats relating mortality to speed and trafc volume, but there is no reason to believe they will be different from that of other taxa. There are data from bats to show that roadkill is greater in good habitat and at natural crossing points (Lesiski etþ  al. 2010; Medinas etþ  al. 2013). The effects of trafc speed and volume, road width and height, habitat characteristics, and bat species on rates of roadkill should be explored in greater depth to help us understand how best to mitigate against the effects of roads. Collection of roadkill carcasses by Russell etþ  al. (2009) led to a conservative estimate of an annual mortality of 5þ  % of the bats in local roosts. Altringham (2008) arrived at a similar estimate, based on conservative calculations for a road in the UK crossed by lesser horseshoe bats from a large roost (data from Billington 2001–2006). Theoretical studies (e.g. Lande 1987; With and King 1999; Carr and Fahrig 2001) show that populations of animal species with low reproductive rates and high intrinsic mobility, such as bats, are more susceptible to decline and ultimately extinction by the additional mortality caused by roads.

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44 J. Altringham and G. Kerth3.2.4þ Habitat Degradation—Light, Noise and Chemical PollutionLight Several studies (e.g. Rydell 1992; Blake etþ  al. 1994; Stone etþ  al. 2009, 2012) have shown that road lighting deters many bat species, notably slow-ying, woodland-adapted species such as members of the genera Rhinolophus, Myotis and Plecotus, from approaching the road. Lighting will probably exacerbate the barrier effect of roads, since those species reluctant to cross open spaces are also those most likely to avoid light. Both high-pressure sodium and white LED light deter woodland-adapted species, even at low intensity (Stone etþ  al. 2009, 2012). Because light intensity drops rapidly away from the source and will often be blocked by vegetation, the effects of isolated sources are not likely to be far reaching in the landscape, but large arrays of high intensity lights will have a signicant effect close to roads. Light can also attract some bat species, in particular open air foragers such as Nyctalus and generalists like Pipistrellus (e.g. Rydell 1992; Blake etþ  al. 1994), since short wavelength light attracts insect prey, concentrating them around lights and increasing bat foraging efciency. This may be not be all good news, since bats exploiting insect swarms around lights may be at greater risk of collision with trafc. As discussed above, many woodland-adapted bats avoid all forms of visible light, so insects around lights are not available to them. Many insects may indeed be drawn out of woodland towards lights, reducing prey availability to woodland specialists. This could effectively enhance the edge effect around woodland. This has yet to be demonstrated but is worth investigation. The chapter by Rowse etþ  al. discusses the detrimental and benecial effects of articial lights on bats in detail. Noise Most insectivorous bats rely on hearing the returning echoes of their ultrasonic echolocation calls to orientate, detect prey and even communicate. Some species locate and capture prey by listening for sounds generated by their prey, such as wing movements or mating calls. Trafc noise may mask prey-generated sounds and the lower frequency components of echolocation calls. During indoor ight room experiments, simulated trafc noise reduced the feeding efciency of the greater mouse-eared bat, Myotis myotis, which typically hunts by listening for sounds made by its prey on the ground (Siemers and Schaub 2011). It is likely that habitats adjacent to noisy roads would therefore be unattractive as feeding areas for this and other species that glean their prey from the ground or vegetation by listening to rustling noises. Vehicle noise may also exacerbate the barrier effect: bats become less likely to y across a road as trafc noise increases (Bennett and Zurcher 2013). Currently, there are no published eld studies that have assessed the effect of trafc noise on bat diversity, abundance or breeding success. However, as described below, trafc noise, like light, is only likely to have a signicant effect over relatively short distances. Pollution Chemical pollution is another signicant factor potentially affecting bats close to roads: transport is the fastest growing source of greenhouse

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45 3þ Bats and Roadsgases. In the USA, over 50þ  % of domestic CO2 emissions come from cars, putting 1.7þ  billion tonnes into the atmosphere every year—a major contributor to climate change. In addition there are the local effects of other chemical pollutants. Automobile exhaust gases close to a road have been shown to be associated with a decline in arthropod diversity and abundance (Przybylski 1979). Motto etþ  al. (1970) and Muskett and Jones (1980) found signicant effects on invertebrates of lead and other metals from cars up to 30þ  m from roads.3.2.5þ Species-Specic EffectsBody size, wing form, echolocation call structure and feeding and roosting ecology all determine how bats y and use the landscape. Thus, it is not surprising that the effects of roads on bats are to a signicant extent species-specic. Larger, fastying species, adapted to foraging in the open, appear from most studies to be less affected by roads (e.g. Kerth and Melber 2009; Abbott etþ  al. 2012a; Berthinussen and Altringham 2012a), as they typically y high above the ground. Their greater ight efciency and speed relative to woodland-adapted species mean that even if they are forced to make long diversions to nd safe crossing points or to avoid roads altogether, the consequences are likely to be less important. Smaller, slower ying, woodland-adapted species are more manoeuvrable and typically capable of gleaning and hovering but this necessarily makes them less efcient yers (Altringham 2011). Woodland species are also more reluctant to y in the open and tend to commute along linear features in the landscape such as treelines, waterways, and woodland edges. These features provide protection from weather and predators, are sources of insect prey, and provide conspicuous acoustic and visual landmarks for orientation. Figureþ  3.3 shows schematically the main patterns of ight and habitat use by insectivorous bats. It is unfortunate that the species most likely to be affected by roads, the slow-ying, woodland-adapted bats, such as Rhinolophus and some Myotis species, are also those that have suffered most from human activity in Europe and North-America and are at highest risk of extinction there (Sa and Kerth 2004).3.2.6þ Road Class and SpeedThe greater width of motorways may make them more effective barriers (Berthinussen and Altringham 2012a) than most other roads. However, trafc density may be equally important (Russell etþ  al. 2009; Zurcher etþ  al. 2010; Bennett and Zurcher 2013) and many major non-motorway roads carry similar or greater trafc volumes, at comparable speed, to rural motorways. Even minor roads are avoided by many bat species. In a habitat suitability modelling (HSM) study in northern England based on extensive acoustic surveys,

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46 J. Altringham and G. Kerth Bellamy et al. ( 2013 ) found that only Nyctalus and Pipistrellus species showed a positive association with roads and then only when roads were at low densities and in close proximity to woodland. This association is likely due to the use by bats of hedgerows along roads that connect to woodland. Other species, particularly wood land specialists, such as Myotis and Plecotus species, avoided roads and all species avoided roads when they became dense around settlements. All road classes were combined in this study, but minor roads predominate in the region, so the effects of major roads were probably underestimated. Studies of birds support these conclu sions: Develey and Stouffer ( 2001 ) and Laurance et al. ( 2004 ) have shown that even narrow, unpaved forest roads can act as barriers to tropical forest birds. In the absence of further work on bats we can look at other animals. Forman et al. ( 2003 ) demonstrated that roads act as signicant barriers to a variety of mammals from voles to grizzly bears, that primary roads are signicantly more effective barriers than secondary roads, and the barrier effect increases with increasing trafc volume. The effects in some cases are severe. Gerlach and Musolf ( 2000 ) have shown that populations of bank vole are genetically distinct either side of a busy four-lane highway (50 m wide, 30,000 vehicles/day), but not either side of a two-lane country road (10 m, 5000 vehicles/day) or a railway. Highways can be major genetic barriers even to large and mobile animals such as coyotes and lynx (Riley et al. 2006 ) or red deer (Frantz et al. 2012 ). Fig. 3.3 Flight style and habitat use by insectivorous bats. Drawing by Tom McOwat

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47 3þ Bats and Roads3.2.7þ Cumulative Effects, Extinction Debt and the Importance of ScaleMost of the factors discussed above will be cumulative. The effects of each individually need not therefore be great for the combination to have a profound effect on a bat population. Furthermore, in many cases there will be a lag, known as the extinction debt, between cause and effect (e.g. Tilman etþ  al. 1994; Loehle and Li 1996). This is illustrated in Fig.þ  3.4. The effects of habitat loss and reduced habitat quality on the distribution of ying bats may be seen quickly, as bats alter their foraging and commuting behaviour to adapt as best they can to the altered landscape. Collision mortality, unless very high, may not have a signicant and detectable effect for several generations. The barrier effect may take several more generations to show itself, since it is likely to involve the decline and/or relocation of nursery and other roosts, but it too may be rapid, for example when bats are completely excluded from key foraging areas. Although no data exist for bats, a study of the effects of roads on wetland biodiversity (birds, mammals, reptiles, amphibian and plants) suggests that the full effects may not be seen for several decades (Findlay and Bourdages 2000). This has important implications for monitoring the effects of roads and assessing the effectiveness of mitigation, as discussed later. Fig.þ  3.4þ The multiple causes of bat population reduction by roads and the delayed response (extinction debt). Adapted from Forman etþ  al. (2003)

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48 J. Altringham and G. KerthBerthinussen and Altringham (2012a) found that the decline in diversity and abundance of bats extended to at least 1.6þ  km from a motorway. Which of the above mechanisms contribute to this extensive effect? Low activity and diversity close to the road may be due to most or all of the factors identied: habitat degradation resulting from light, noise and chemical pollution, a barrier effect, or increased mortality due to roadkill. Noise pollution can contribute only to shortrange effects, since noise levels in the study fell rapidly over the rst 200þ  m and were close to ambient thereafter. Lab studies on the gleaning greater mouse-eared bat Myotis myotis (Schaub etþ  al. 2008; Siemers and Schaub 2011) show that even species that hunt by listening for prey-generated noise are not likely to be affected by roads more than 60þ  m away. Light pollution was not considered by Berthinussen and Altringham, since the road sections studied were unlit. However, any effect of light pollution from road and vehicle lights is also likely to oper ate over relatively short distances, due to the inverse square relationship between distance and light intensity. In addition vegetation alongside of roads will further reduce the effect of light and noise pollution quickly. Road developments can disrupt local hydrology and polluted run-off may degrade wetland foraging habitats (Highways Agency 2001), but the scale of such effects will be very variable. As discussed above, chemical pollution is likely to be a factor only over relatively short distances unless dispersion is facilitated by drainage. The many processes that may be degrading roadside habitats need further study, but none of those discussed are likely to explain changes in bat activity over 1.6þ  km. Reduced activity over long distances can however be explained by the combination of a barrier effect and increased mortality due to roadkill. The home ranges of temperate insectivorous bat species typically extend 0.5–5þ  km from their roost (e.g. Bontadina etþ  al. 2002; Senior etþ  al. 2005; Davidson-Watts etþ  al. 2006; Smith and Racey 2008), and most species show high delity to roosts, foraging sites and commuting routes (e.g. Racey and Swift 1985; Entwistle etþ  al. 2000; Senior etþ  al. 2005; Kerth and van Schaik 2012; Melber etþ  al. 2013). A major road built close to a nursery roost has the potential to reduce the home range area of a colony through both destruction of habitat and the severance of commuting routes that reduces access to foraging areas. The bats have several options. One is to continue to use the roosts close to the road with a reduced foraging area, reduced resources and reduced reproductive potential (Kerth and Melber 2009). The colony is therefore likely to decline. Alternatively bats may cross the road to maintain their original home range area. Local habitat loss and degradation and increased roadkill will compromise the colony, which may therefore decline. Mortality from roadkill is likely to be high since most species cross at heights that put them in the paths of vehicles (e.g. Verboom and Spoelstra 1999; Gaisler etþ  al. 2009; Russell etþ  al. 2009; Berthinussen and Altringham 2012b). Bats may waste time and energy by commuting greater distances, either away from the road to nd new foraging sites, or to nd ‘safe’ crossing points along the road to commute to their original foraging sites. All of these outcomes will reduce the reproductive output of nursery colonies (e.g. Tuttle 1976; Kerth and Melber 2009). Alternatively the colonies may relocate away from the road, into habitat that is presumably already fully exploited by

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49 3þ Bats and Roadsother colonies. All ‘solutions’ will lead to a fall in bat density near to the road. The overall fall in habitat quality will most likely lead to reduced reproductive success and increased adult mortality and in long-lived bats these will have a profound effect on local colony size and overall population size (Sendor and Simon 2003; Papadatou etþ  al. 2011). Given the magnitude and spatial scale of the effects on bat activity and diver sity observed by Berthinussen and Altringham (2012a), it is likely that barrier and edge effects, together with increased roadkill are having a strong negative effect on the demographics and distribution of local bat populations in proximity to major roads. Similar effects have been found in other vertebrates. Reijnen and Foppen (1994) showed that a decreased density of willow warblers up to 200þ  m from a major highway was due to the negative inuence of the road on population sizes, with reduced breeding success and increased emigration of territorial males. Studies on breeding grassland birds revealed a decrease in density of seven out of 12þ  species, with disturbance distances up to 3500þ  m from the busiest roads (50,000 vehicles per day), with collision mortality being a major contributor (Reijnen etþ  al. 1996). A meta-analysis of 49 studies that between them investigated 234 bird and mammal species, found that bird population densities declined up to 1þ  km, and mammal population densities declined up to 5þ  km from roads (BenítezLópez etþ  al. 2010).3.2.8þ Secondary Effects—Inll and Increased Urban and Industrial DevelopmentBypasses are frequently built in the countryside to divert trafc around rather than through population centres, to reduce congestion and improve the environment for people in the town or village. In addition to the direct effects of the road itself, there are frequently other consequences. The typically narrow strip of land between the settlement and the new road may be too small to support viable bat populations. This land is also frequently taken over by residential and industrial/ commercial development and indeed this development is often part of the initial plan. This leads to further loss and degradation of habitat and a direct increase in trafc. Many of the secondary effects of roads are more severe in the tropics (Laurance etþ  al. 2009), where roads allow people easy access to the remaining undisturbed habitats, which as a consequence suffer further degradation and an increase in the hunting pressure for bush meat, including bats.3.3þ Can Roads Benet Bats?Although the balance of the impact of roads on bats is clearly strongly negative, there are potential benets.

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50 J. Altringham and G. KerthRoosts Some of the ancillary structures built with roads, in particular bridges (e.g. Keeley and Tuttle 1999), can provide roosts for bats. Road bridges over water or wooded valleys are the most likely to be used, those over busy roads much less so. Old stone road bridges over water are widely used by bats, most notably by Daubenton’s bat in Europe, but also other Myotis species and by Nyctalus species (e.g. Senior etþ  al. 2005; Celuch and Sevcik 2008; Angell etþ  al. 2013). In North America bridges are widely used by Brazilian free-tailed bats, Tadarida brasiliensis (e.g. Allen etþ  al. 2011) and some other species (e.g. Bennett etþ  al. 2008). Effective mitigation and compensation for the loss of roosting and foraging sites will make the environment close to a road more attractive to bats, but may do so at the expense of greater risk of collision with trafc. Light Articial light, particularly short-wavelength light such as mercury-vapour (not most LED lights) attract insects that are common prey to bats. Insect swarms around lights are exploited by open-air foraging bats such as Pipistrellus and Nyctalus (Rydell 1992; Blake etþ  al. 1994; Stone etþ  al. 2009, 2012). One consequence of this is that bats feeding around lights on busy roads may be at signicantly greater risk of mortality from collision with trafc. The balance between the positive and negative effects will be dependent on species, topography, the position of lights, etc. and further study would be useful. A very thorough discussion of the positive and negative effects of articial light can be found in the chapter by Rowse etþ  al. Flight corridors In rural environments roads are often bounded by hedgerows or treelines. The wide verges often associated with hedges in landscapes managed for wildlife can be among the most species-rich habitats in some agricultural areas. Minor roads in particular can therefore be both foraging sites and commuting routes, but even major roads are used by some species (e.g. Nyctalus leisleri, Waters etþ  al. 1999) where they are bounded by suitable habitat such as a woodland edge. Depending upon structure, this habitat could be used by a wide range of species. However, Bellamy etþ  al. (2013) found that even low road densities had a negative effect on most species of bats, most noticeably the woodland-adapted species Myotis and Plecotus. Only the distributions of common pipistrelles and noctules had a positive association with roads at low to moderate densities and only when in close proximity (<100þ  m) to woodland. A similar result was found for railway verges (Vandevelde etþ  al. 2014). As road density increased above moderate levels, the probability of presence of all species declined. The effects of roads of different classes have yet to be investigated in depth—the roads in this study were predominantly minor and rural.3.4þ Conservation in Principle: Avoidance, Mitigation, Compensation and EnhancementIn many countries, legislation has been passed stating that infrastructural development should be carried out in such a way as to minimise the impact of development on the environment, and on protected species such as bats in particular. In

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51 3þ Bats and Roadsprinciple, there should be no net loss to the environment. In the European Union this is formalised in the Habitats Directive (Council Directive 92/43/EEC). In practice, the system is usually awed, sometimes severely, due to a lack of knowledge, resources and commercial and political will. Poor goal-setting, planning and execution contribute to either failure, or the absence of any evidence for success, for all wildlife (Tischew etþ  al. 2010) and bats in particular (Altringham 2008; Berthinussen and Altringham 2012b; Stone etþ  al. 2013). As in many other areas of conservation a more scientically robust, evidence-based approach is urgently needed. European policy and practice also involve a hierarchal approach, starting with avoidance of environmental damage, moving to mitigation when damage is deemed to be unavoidable, then compensation when mitigation is not possible or only partial. Finally, there is an increasing expectation that replacing like with like is not enough, particularly given the uncertainty of success in mitigation and the continued loss of biodiversity. When habitat is lost or degraded, some level of habitat enhancement must accompany development so that in principle, the habitat is better than it was before development. The reality is less than perfect. The rst step in a conservation strategy to minimise the impact of a new road should be to select a route that avoids important bat habitat. To be effective this requires an understanding of the behaviour and ecology of the affected species and detailed knowledge of their distribution. Our knowledge in both areas is growing but far from complete. One approach that can deliver detailed, site-specic infor mation relatively quickly is GIS-based HSM, which can be based on existing data sets, such as those held by museums and record centres (e.g. Jaberg and Guisan 2001; Bellamy and Altringham 2015) or data collected specically for the pur pose, for example by acoustic survey (e.g. Bellamy etþ  al. 2013). This approach yields ne scale distribution maps of probability of occurrence for each species with an estimate of reliability, providing a useful practical tool. However, the route that best avoids bats may not meet human social and economic criteria, particularly if conservation is undervalued. The next step is therefore to build the road in such a way as to mitigate against its effects—that is remove or minimise the many detrimental effects described above. In principle, mitigation under European legislation (Habitats Directive, Council Directive 92/43/EEC) reduces ‘damage’ to a minimum that is consistent with maintaining bat populations in favourable conservation status. Where signicant loss cannot be avoided, it is expected that compensation will provide alternative roosting and foraging habitat to at least make good the loss. The expectation now is that there is in fact habitat enhancement, to allow for uncertainties in mitigation and to promote long-term habitat improvement. In practice, avoidance and mitigation are compromised by competing operational and nancial constraints. Furthermore, for practical and economic reasons, habitat restoration and creation are long-term processes and it may be many years before these sites are useful to bats, by which time a disturbed bat colony may have been lost. As we will show in the following section, the absence of adequate and well-planned survey and monitoring means that the consequences of roadbuilding and the effectiveness of current avoidance, mitigation, compensation and

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52 J. Altringham and G. Kerthenhancement practices are all largely unknown (Altringham 2008; O’Connor etþ  al. 2011). In some cases, they have even been shown to be ineffective (Berthinussen and Altringham 2012b).3.5þ Conservation in PracticeWe are not aware of any cases in which proposed roads have been rerouted to avoid key bat habitat. Almost all work in this area concerns attempts to remove or minimise the damaging effects of roads. This has usually involved building structures that aim to guide bats safely under or over roads to reduce both the bar rier effect and roadkill. The structures built may be multifunctional, for example underpasses for people and wildlife, and use by bats has often been an incidental and unanticipated use of structures built for other purposes, such a drainage culverts. Additional features include tree and hedge planting to guide bats towards crossing points, modied lighting schemes to achieve the same ends or deter bats from crossing at dangerous locations and a wide range of more general ‘enhancements’ to improve roosting or foraging opportunities.3.5.1þ Over-the-Road Methods: Gantries, Green Bridges, Hop-Overs and Adapted Road/Foot BridgesBat bridges or ‘bat gantries’ have been built on many UK and continental European roads in recent years. However, the most widely used design (Fig.þ  3.5) in the UK does not help bats to cross the road safely, even when on the line of pre-construction yways and after up to nine years in situ as shown in Fig.þ  3.6 (Berthinussen and Altringham 2012b). Other designs have yet to be tested effectively. Berthinussen and Altringham (2012b) found that only a very small propor tion of bats that approached gantries ‘used’ them (i.e. ew in close proximity to them) and for those that did, their ight paths were not raised above the trafc collision zone (Fig.þ  3.6). This failure of a widespread design highlights the need for effective monitoring and assessment to be an integral part of mitigation practice. Overpasses built to carry minor roads or footpaths appear to be largely ineffective (Bach etþ  al. 2004; Abbott etþ  al. 2012a) and certainly less effective than underpasses as crossing points (Bach etþ  al. 2004; Abbott etþ  al. 2012a). Most of the structures evaluated have been no more than footbridges and road bridges, with no adaptations to encourage bats, such as tree or shrub planting or careful design of lighting. To date studies have assessed only use, not effectiveness, in that the criterion for success in most studies has been use by an unspecied proportion of bats. A more useful approach would be to assess what proportion of bats crossing a road do so with the aid of crossing structure (Berthinussen and Altringham 2012b).

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53 3 Bats and Roads Land or green bridges have been designed and built specically for other wild life, and if planted with tall vegetation and linked to existing bat yways, they have obvious potential as bat crossing structures. As yet, few have been assessed, but bats have been shown to use one land bridge in Germany. Stephan and Bettendorf ( 2011 ) found that only a small proportion of woodland-adapted bats crossed a busy motorway using a new land bridge: most crossed the road itself at other locations. It will be interesting to see if bats adapt to it over time. Specic features of the design and connectivity to surrounding habitat of green bridges are Fig. 3.5 The most common bat gantry design in the UK—steel wires with plastic spheres at intervals that are intended to be acoustic guides for bats Fig. 3.6 Bat crossing activity at a ‘bat gantry’ that had been in place for nine years. Gaussian kernel and bandwidth of 1 m used ( n 1078). The gantry is located at distance 0 m on the x-axis, with distance from the gantry increasing to the left and right . The height of the gantry is marked by the square at 0 m, and the pre-construction commuting route is 10–15 m to the right . ‘Unsafe’ crossing heights are located below the dashed line , which is the maximum vehicle height in Europe. The dotted line marked verge shows the decrease in verge height above the road from left to right . From Berthinussen and Altringham ( 2012b )

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54 J. Altringham and G. Kerthprobably critical factors for bat use—as they will be for other structures. Further research is required before conclusions can be drawn, but several features are likely to be positively related to use: their strategic location on known ightlines, connectivity to treelines, mature vegetation on the bridge, and bridge width. ‘Hop-overs’ (Limpens etþ  al. 2005) have been put forward as a relatively low cost and unobtrusive way to encourage bats to cross roads at safe heights. These consist of close planting of trees up to the road edge on both sides of the road, with tall vegetation in the central reservation of wide roads. Branches should overhang the carriageway, ideally giving continuous canopy cover over the road. Safety concerns arising from overhanging branches may have led to reluctance to adopt hop-overs and even to remove trees from road margins. However, many roads have overhanging trees along their margins, so this is an illogical or at least inconsistent objection. The effectiveness of hop-overs has yet to be assessed. Russell etþ  al. (2009) observed that bat ights across a 20þ  m road gap were at greater heights where bats approached the road along ight routes with taller roadside vegetation and Berthinussen and Altringham (2012b) found a positive correlation between road-crossing height and the height of the roadside embankment.3.5.2þ Under-the-Road Methods: Underpasses, Culverts and Other ‘Tunnels’Many studies show that a wide range of bat species use underpasses to y beneath roads (e.g. Bach etþ  al. 2004; Kerth and Melber 2009; Boonman 2011; Abbott etþ  al. 2012a; Berthinussen and Altringham 2012b). However, most of these studies report only that a small number of bats of particular species were seen to y through an underpass. In some cases not reported here underpasses were monitored using automated bat detectors with no guarantee that detected bats actually ew through the underpass. For an underpass (or indeed any other mitigation structure) to be effective it must help to maintain bats in favourable conservation status. That is, it must protect the population, not a few individuals, by making a road permeable and safe to cross. Assessing abundance, let alone changes in abundance, is very difcult without considerable survey effort. It is also difcult to measure changes in the permeability of a road to bats without monitor ing a very large proportion of the bats in the vicinity of a newly built or upgraded road. Ideally, we would need data before the construction of the road and compare them with data after the road had been built. However, it is possible to deter mine whether the majority of bats at a location use an underpass (or bridge, gantry, etc.) to cross a road safely. Despite the existence of three underpasses within a 5þ  km stretch of motorway bisecting a forest, resident Bechstein’s bats rarely used them and lost access to important roosting and feeding habitat (Kerth and Melber 2009). Lesser horseshoe bats made frequent use of three underpasses along a 1þ  km stretch of motorway, but 30þ  % still crossed directly over the road at trafc height

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55 3þ Bats and Roads(Abbott etþ  al. 2012b). Some bats have been recorded making extensive detours to avoid crossing roads (e.g. Kerth and Melber 2009 and references cited in Bach etþ  al. 2004), but we do not know how prevalent this behaviour is: many bat species appear reluctant to deviate from their original ight paths after road sever ance (Kerth and Melber 2009; Abbott 2012; Berthinussen and Altringham 2012b). Where a road cuts through a dense network of ight routes it may not be straightforward providing a population with an adequate number of safe crossing points. Efforts to re-route bat ight paths, for example by planting new hedgerows linking old routes with new underpasses, should be undertaken well in advance of road clearance, and ideally tested for effectiveness before road opening. Bats were not diverted effectively to underpasses studied by Berthinussen and Altringham (2012b): the great majority of bats ew over the road, near to the original commuting routes. In the same study, one underpass on a known ightline was used by 96þ  % of the bats on the commuting route. Underpasses are more likely to be used if they are well connected to the landscape by treelines, hedges or watercourses (Boonman 2011; Abbott 2012), but there is scope for further study in this area. Where possible, they should be located on pre-construction ight routes and tall enough to allow bats to pass without changing ight height or direction (Berthinussen and Altringham 2012b). Even with these precautions, a high proportion of bats may ignore the underpass and y over the road above it, particularly if the underpass is too small. Underpass height, more than width, was the critical dimension determining the number of bats ying through underpasses in studies in Ireland (Abbott 2012; Abbott etþ  al. 2012a, b). Required heights of underpasses will generally be lower for woodland-adapted species (~3þ  m) compared to generalist edge-adapted species (~6þ  m), and open-air species are more likely to y high above roads. For small gleaning bat species, such as some Myotis species, which generally have small home ranges, it may be benecial to build a higher number of small underpasses (Fig.þ  3.7) along a road instead of a few large underpasses, which then would by located outside of the home range of most individuals. Mitigation practice would benet greatly from objective testing and reporting to determine if underpasses are actually providing safe passage for a high enough proportion of bats to protect a local population. Bats can potentially make use of underpasses that are used by people during the day but have little use at night, such as pedestrian underpasses, minor roads, railways and forestry or agricultural tracks. Use could be maximised by restricting lighting in and around these underpasses, placing them on tree and hedge lines, and making smaller wildlife underpasses or drainage culverts larger to accommodate woodland-adapted bat species. Provision of well-placed, numerous and spacious underpasses should be integral to the overall design of road mitigation, particularly near major roosts. Roads built on embankments are likely to be particularly dangerous to bats, particularly when they sever treelines, since bats appear to maintain ight height on leaving the treeline, bringing them into collision risk over raised road sections. These sites are ideal candidates for under passes, since they can be built relatively cheaply.

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56 J. Altringham and G. Kerth 3.5.3 Light Avoidance To reduce the potential for disturbance of roosts, ight routes and feeding sites lighting is often directed down toward the road surface, and light spill into the surroundings is minimised. However, since the most vulnerable bats, such as Rhinolophus species, y close to the ground, downward pointing lighting may still have a signicant impact on their behaviour. Restricting lighting in crossing structures such as pedestrian underpasses could increase their use by bats. In addi tion to choosing the intensity, wavelength and direction of lighting, it could also be controlled be timers and motion sensors. Lighting at river and stream crossings should always be avoided, as these are particularly important foraging areas and commuting routes for bats. Conversely, light may be used to purposely deect bats away from a dangerous ight route toward a safe crossing point. This has been done, but has not yet been tested for effectiveness and may exacerbate any barrier effect. This assessment is important not only to protect bats, but other wildlife too, since many species avoid light. 3.5.4 The Importance of Connectivity and the Maintenance of Existing Flightlines An important consideration that is frequently referred to is the need to maintain existing ightlines. There is evidence to support this and it is clearly a sensible precaution. As discussed above, Berthinussen and Altringham ( 2012b ) found that Fig. 3.7 A bat of the genus Myotis using a small underpass (about 2 m in diameter) to cross a motorway in Germany. Above the underpass, a wall was built to prevent bats from ying directly into the trafc. Similar walling/fencing has been used in the UK but has not yet been shown to be effective (e.g. Billington 2001 –2006)

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57 3þ Bats and Roadsan underpass on a pre-existing ightline was used by 96þ  % of the bats crossing the road, but attempts to deect bats to two other underpasses displaced from known routes were not successful. An extension of this is the general recommendation to maintain and enhance a ‘connected’ landscape, i.e. a landscape with a broad range and high density of inter connecting linear features such as hedgerows and treelines. This would not only increase the value of the landscape for foraging and commuting, but may give bats more exibility in how they adapt to a changing landscape and in particular the appearance of barriers in the form of roads. This makes intuitive sense, given the known behaviour of many bat species, and there is a growing body of evidence based on spatial analysis to support it (e.g. Boughey etþ  al. 2012; Bellamy etþ  al. 2013; FreyEhrenbold etþ  al. 2013; Bellamy and Altringham 2015). These studies highlight, using different approaches, the importance of these features to bats, and also reveal species differences: woodland-adapted species (e.g. Myotis, Plecotus, Rhinolophus) and small generalists (e.g. Pipistrellus) make more use of (and are more dependent upon) these features than larger open-air species (e.g. Nyctalus, Eptesicus).3.5.5þ Habitat Improvement and Effective Landscape-Scale PlanningSome general forms of mitigation not specically related to roads are also relevant, such as the planting of trees and the creation of ponds to replace lost habitat or enhance existing habitat as compensation for damage done by roads. Berthinussen and Altringham (2012a) have shown that the effects of major roads are less easily detected in high quality habitat. This is not a reason to build roads in high quality habitat, since a greater number of bats will still be affected than alongside a road through poor habitat, and the species affected may be more vulnerable. However, it is a reason to attempt to mitigate and compensate using habitat improvement, when a road is built in good habitat. Improvements must not increase roadkill or the costs may outweigh the benets, so habitat design will be an interesting challenge. Habitat improvement methods have not been tested effectively, so the scale of the benets is generally unknown. Habitat improvement and creation obviously have the potential to be benecial if done on an appropriate scale, but are unlikely to be effective in the short or even medium term, since new woodland and wetland take many years to become established. Over the time taken for habitat to mature, bat colonies may be lost, so long-term planning is needed. Considerable nancial incentives may be needed to persuade landowners to undertake habitat improvement. Woodland and wetland creation are more likely to be used for compensation and enhancement than direct mitigation. As discussed earlier, the Habitats Directive stipulates that in preparing development plans, the avoidance of damage is the preferred option. Mitigation and

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58 J. Altringham and G. Kerthcompensation should only be considered when alternative sites, routes or methods are unavailable and the avoidance of damage is not possible. There must also be over-riding social, economic or safety reasons for development. The planning of new road and rail routes now makes extensive use of GIS-based techniques to assist in the evaluation of the many factors involved. However, the environmental components of these analyses often rely on limited and biased data and do not take full advantage of the developing GIS and modelling techniques described earlier. GIS-based HSM is becoming widely used in ecology. HSM uses the detailed relationships between bat presence and habitat variables to build detailed and accurate distribution maps from relatively small datasets. Bellamy etþ  al. (2013) and Bellamy and Altringham (2015) have used HSM to produce high resolution, accurate predictive maps of the distribution of eight bat species in the Lake District National Park. Similar maps have been, and are being, prepared for other protected areas. These techniques determine the associations between bats and their habitat over multiple spatial scales to give greater accuracy and ecological insight. As our knowledge of bat distributions improves, we will be in a better position to identify those routes that will have minimum impact on bats, and better able to devise appropriate mitigation strategies.3.5.6þ RailThe effects of rail systems on both bats and other wildlife are even less well understood than those of roads. However, intuitively they have characteristics that may reduce their impact on wildlife. Rail systems are often (but not always) nar rower than roads, giving them a smaller footprint and potentially creating a lesseffective barrier to animal movement. Trains pass a given point on a network much less frequently than vehicles on roads, which are often continuous. On the busy East Coast line in northern England train noise was detectable for only 8þ  min/h and this noise decreased to background levels over very much shorter distances than road noise (Altringham 2012). It is nevertheless important that the effects of railways are assessed objectively, particularly in view of the proposed new HS2 line in England, on which trains will travel faster and more frequently. In a study on bat activity of railway verges, Vandevelde etþ  al. (2014) found that bat of the genus Myotis seem to avoid the vicinity of railways whereas species foraging in more open space such as pipistrelle and noctule bats use railway verges as foraging habitat.Open Access This chapter is distributed under the terms of the Creative Commons Attribution Noncommercial License, which permits any noncommercial use, distribution, and reproduction in any medium, provided the original author(s) and source are credited.

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63Chapter 4Responses of Tropical Bats to Habitat Fragmentation, Logging, and DeforestationChristoph F.J. Meyer, Matthew J. Struebig and Michael R. Willig© The Author(s) 2016 C.C. Voigt and T. Kingston (eds.), Bats in the Anthropocene: Conservation of Bats in a Changing World, DOIþ  10.1007/978-3-319-25220-9_4Abstractþ Land-use change is a key driver of the global biodiversity crisis and a particularly serious threat to tropical biodiversity. Throughout the tropics, the staggering pace of deforestation, logging, and conversion of forested habitat to other land uses has created highly fragmented landscapes that are increasingly dominated by human-modied habitats and degraded forests. In this chapter, we review the responses of tropical bats to a range of land-use change scenarios, focusing on the effects of habitat fragmentation, logging, and conversion of tropical forest to various forms of agricultural production. Recent landscape-scale studies have considerably advanced our understanding of how tropical bats respond to habitat fragmentation and disturbance at the population, ensemble, and assemblage level. This research emphasizes that responses of bats are often species and ensemble specic, sensitive to spatial scale, and strongly molded by the characteristics of the prevailing landscape matrix. Nonetheless, substantial knowledge gaps exist concerning other types of response by bats. Few studies have assessed responses at the genetic, behavioral, or physiological level, with regard to disease prevalence, or C.F.J. Meyerþ  ()þ  Centre for Ecology, Evolution, and Environmental Changes, Faculty of Sciences, University of Lisbon, Lisbon, Portugal e-mail: cmeyer@fc.ul.pt M.J. Struebigþ  Durrell Institute of Conservation and Ecology, School of Anthropology and Conservation, University of Kent, Canterbury, UK e-mail: m.j.struebig@kent.ac.uk M.R. Willigþ  Center for Environmental Sciences and Engineering and Department of Ecology and Evolutionary Biology, University of Connecticut, Storrs, CT, USA e-mail: michael.willig@uconn.edu

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64 C.F.J. Meyer et al.the extent to which human disturbance erodes the capacity of tropical bats to provide key ecosystem services. A strong geographic bias, with Asia and, most notably, Africa, being strongly understudied, precludes a comprehensive understanding of the effects of fragmentation and disturbance on tropical bats. We strongly encourage increased research in the Paleotropics and emphasize the need for long-term studies, approaches designed to integrate multiple scales, and answer ing questions that are key to conserving tropical bats in an era of environmental change and dominance of modied habitats (i.e., the Anthropocene).4.1þ Habitat Conversion: A Key Aspect of Global ChangeBats are valuable indicators of biodiversity and ecosystem health, and respond to a range of stressors related to environmental change (Jones etþ  al. 2009). Alteration in land use is one of the principal aspects of global environmental change and a key driver of biodiversity loss in terrestrial ecosystems. Indeed, biodiversity impacts of land-use change are generally considered to be more immediate than those from climate change (Sala etþ  al. 2000; Jetz etþ  al. 2007; Pereira etþ  al. 2010). However, the effects of land-use change on tropical species could exacerbate those of changing climate, leading to challenges for long-term conservation efforts (Struebig etþ  al. 2015), including those for bats. Over the last decades, human transformation of much of the Earth’s natural ecosystems has greatly accelerated, and the twenty-rst century will herald profound changes in land use, particularly in developing tropical countries (Lee and Jetz 2008). The most recent quantication of global forest change revealed an overall increasing trend in annual forest loss across the tropics between 2000 and 2012 (Hansen etþ  al. 2013), highlighting the continued prevalence of tropical deforestation. Drivers of tropical deforestation have shifted from being promoted mostly by government policies for rural development toward urban population growth and industrial-scale, export-oriented agricultural production (DeFries etþ  al. 2010). Fueled by unabated human population growth, global food demand is escalating, and the current trajectory of agricultural expansion will have serious negative long-term consequences for the preservation of the planet’s biodiversity (Tilman etþ  al. 2011; Laurance etþ  al. 2014). In tropical countries, conversion of natural habitats to agricultural and pastoral land is one of the greatest threats to biodiversity (Phalan etþ  al. 2013), as cropland expansion in recent decades has largely come at the expense of intact old-growth forest (Gibbs etþ  al. 2010). Rampant commercial logging is also a major force of tropical forest destruction and degradation, with around 20þ  % of such forests subjected to some level of timber harvesting (Asner etþ  al. 2009). Loss of habitat as a result of extensive land conversion and associated fragmentation are ubiquitous throughout the tropics. Resulting landscapes typically comprise a mosaic of human-modied habitats that include agroforests, agricultural land, and tree plantations, as well as remnants of old-growth, logged forest, and secondary forests regenerating from clearance or burning (Gardner etþ  al. 2009; Chazdon 2014).

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65 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …Indeed, except for large areas of tropical forest in Papua New Guinea and in the Amazon and Congo basins, such a description accurately characterizes most tropical landscapes (Melo etþ  al. 2013). Anthropogenic activities in many tropical countries have resulted in the creation of fragmented landscapes that are dominated by small (oftenþ  <þ  50þ  ha), isolated, and irregularly shaped forest patches. These patches are highly prone to edge effects (Broadbent etþ  al. 2008; Ribeiro etþ  al. 2009), dened as systematic changes in abiotic and biotic variables at the boundary between adjacent land-use types. Although deforestation and degradation of old-growth forests are the dominant forms of land-use alteration, forest regeneration and the expansion of secondary forests are the second most important type of land-use change occur ring across the tropics (Asner etþ  al. 2009; Dent and Wright 2009). These recovering forest habitats could potentially mitigate, or even reverse, current trends of forest loss and degradation as well as concomitant biodiversity loss (Wright and MullerLandau 2006; Dent and Wright 2009; Chazdon 2014). A pan-tropical meta-analysis of land-use change studies points to the irreplaceable value of old-growth forests, but also highlights the high species diversity found in regenerating logged forests compared to secondary forests (Gibson etþ  al. 2011). Although the long-term conservation value of regenerating forests has been questioned (Melo etþ  al. 2013), biodiversity representation clearly varies among logged and secondary habitats, and so not all recovering forests should be treated equally.4.2þ Tropical Bats in a Changing WorldBats exhibit the general mammalian pattern of greatest diversity in the tropics, from both a taxonomic and a functional perspective (Willig etþ  al. 2003). Bats also provide ecosystem services that are critically important in tropical ecosystems—as pollinators and seed dispersers for hundreds of plant species and as agents of suppression of arthropod herbivores and insect pest species (Muscarella and Fleming 2007; Kalka etþ  al. 2008; Williams-Guillén etþ  al. 2008; Kunz etþ  al. 2011; Maas etþ  al. 2013). Nonetheless, many tropical bat species face an uncertain future and show declining population trends due to many of the threats outlined previously (e.g., Kingston 2013). How do tropical bats fare in the Anthropocene, in which they are exposed to increasing levels of land-use change, potentially exacerbated by climate change (Struebig etþ  al. 2015), and the synergistic effects of both processes? Simple pan-þ­ tropical meta-analyses suggest that the impacts of land-use change on þ­ mammal diversity, particularly on bats, are somewhat less severe than for other animal groups (Gibson etþ  al. 2011). Nevertheless, such studies can potentially miss subtle, yet important, responses in assemblage structure. In this chapter, we summarize the accumulated knowledge on the responses of tropical bats to human-induced habitat fragmentation and forest disturbance. By providing a synthetic overview of the topic, we hope to shed light on the conservation value of anthropogenically modied habitats for bats across the major tropical regions and identify future research priorities.

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66 C.F.J. Meyer et al.4.3þ Review MethodologyWe followed a systematic review methodology (Pullin and Stewart 2006) to synthesize information about tropical bat responses to habitat fragmentation, logging, and deforestation. Studies were identied through a comprehensive search in the ISI Web of Science online database (accessed in September 2013), performing a topic search using the string “bat? AND *tropic* AND (fragment* OR logg* OR deforest* OR disturb*),” without restriction on publication year. The use of this combination of key words allowed for the identication of an inclusive set of studies on the effects of fragmentation, logging, and disturbance on tropical bats. The search identied 248 publications that were subsequently screened for suitability for the review based on the article’s title, abstract, and, when necessary, text. We excluded review articles and studies that were conducted in urban landscapes (see Chap. 2). As our purpose here was to review evidence for the effects of anthropogenic habitat modication on tropical bats, we also excluded studies that were conducted in naturally fragmented landscapes (e.g., forest islands embedded in savannah, oceanic islands). Our review thus focuses on a range of human-modied matrix types of varying structural complexity and contrast—from relatively low-contrast secondary forests, agroforests, and plantation forests, to high-contrast agricultural elds and water matrices resulting from dam construction. From the 248 studies, 93 met our criteria. In addition, we extended our search using the same key word combinations in Google Scholar through which we identied an additional eight relevant studies within the rst 100 records. Sixteen additional publications were found based on a search of our own literature databases, thus bringing the total number of studies considered in our synthesis to 117. Each article was characterized according to geographic region, taxonomic focus, response type, and disturbance type. Response types included (a) populationand assemblage-level responses, (b) genetic effects, (c) behavioral responses, (d) physiological responses, parasite and disease prevalence, and (e) effects on the provisioning of ecosystem services. Disturbance type included the following broad categories: (a) habitat fragmentation, (b) logging, (c) secondary forests and succession, (d) agroforestry systems, (e) tree plantations, and (f) agriculture.4.4þ Biases in Our Understanding of Responses of Tropical Bats to Habitat AlterationThe collated literature revealed substantial geographic and taxonomic biases in the current understanding of tropical bat responses to anthropogenic distur bance. Studies covered 34 distinct study landscapes in 21 countries. Despite a general increase in the number of studies over the last 20þ  years (Fig.þ  4.1), most research has been undertaken in the New World tropics (96 studies), with research in Southeast Asia and Australasia lagging far behind (19 studies) and studies in

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67 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …Africa being rare (2 studies; Fig.þ  4.2). Geographic variation in this research effort (Fig.þ  4.2) broadly parallels the pattern reported for multiple taxa across the tropics (Gibson etþ  al. 2011). A few notable differences include a disproportionately high number of bat studies in Mexico and low number of studies in Indonesia compared to other taxa. A large taxonomic bias therefore characterizes our understanding of disturbance effects on tropical bats as a consequence of the prevalence of studies in the Neotropics. With a few exceptions (Estrada etþ  al. 2004; Estrada Villegas Fig.þ  4.1þ Number of publications on the effects of fragmentation, logging, or disturbance on tropical bats based on a systematic search of the literature. There is a general increase in publications over the last 20þ  years (linear model t, Radj 2þ  þ  0.55, pþ  <þ  0.001). Data for 2013 represent an underestimate as the literature search did not include the entire year, and therefore, they were not considered in the model t 0 5 10 15 1995 2000 2005 2010 Publication year Number of studies Fig.þ  4.2þ Map illustrating the geographic distribution of research effort based on 117 studies of bats in anthropogenically modied landscapes. Sizes of orange circles represent the number of studies per site, where a site is dened as a particular study landscape. Colors of tropical countries represent the number of studies based on the pan-tropical analysis of the impact of distur bance and land conversion on birds, mammals, arthropods, and plants by Gibson etþ  al. (2011)

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68 C.F.J. Meyer et al.etþ  al. 2010; Williams-Guillén and Perfecto 2011), New World studies focused on the species-rich Phyllostomidae, in turn largely reecting the use of mist nets to capture bats. Phyllostomids are easily sampled with mist nets and dominate studies. In contrast, non-phyllostomids are underrepresented in samples based on mist netting. Although acoustic methods hold much promise for sampling non-phyllostomid and non-pteropodid bats, considerable difculties remain in the wider implementation of these techniques in tropical countries, including the lack of call libraries, taxonomic uncertainty, and practical challenges of tropical climates (Harrison etþ  al. 2012). As a result, acoustic sampling has not yet been employed intensively in landscape-scale studies of tropical bats (see also Cunto and Bernard 2012). Finally, a considerable bias exists with respect to studied aspects of fragmentation and disturbance. Comparatively few studies have targeted bat responses to logging or agroforestry (Fig.þ  4.3a). The vast majority of studies evaluated responses at the population or assemblage level. Far fewer have examined the consequences of anthropogenic disturbance for the provision of ecosystem services by bats. Genetic, physiological, and behavioral effects remain poorly explored, as do effects on disease dynamics associated with bat hosts (Fig.þ  4.3b).4.5þ Responses at the Population and Assemblage Level 4.5.1þ Habitat FragmentationHabitat fragmentation has become a major research theme in conservation biology, as reected in the burgeoning literature on the subject (Fahrig 2003; Ewers and Didham 2006a; Lindenmayer and Fischer 2006; Fischer and Lindenmayer 2007; Collinge 2009). Although the exact denition of “habitat fragmentation” 4Response type Region Neotropics Paleotropics Region Neotropics Paleotropics (b) Agriculture & residual tree cover Tree plantations Agroforestry Secondary forests & succession Logging Fragmentation Ecosystem service provisioning Behavior Physiology, parasite & disease prevalence Genetic effects Population/ assemblage level 01 02 03 04 05 0 02 00 60Number of studies Number of studiesType of habitat modification Region Neotropics Paleotropics Region Neotropics Paleotropics (a) Fig.þ  4.3þ Number of studies by region (Neotropics [nþ  þ  96 studies] vs. Paleotropics [nþ  þ  21 studies]) based on a type of disturbance or habitat modication and b type of response. Studies in many cases, especially for (a), matched more than one of the broad categories and were counted multiple times

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69 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …is contentious (Fahrig 2003; Ewers and Didham 2007; Fischer and Lindenmayer 2007), we follow a widely used denition—the landscape-scale process by which habitat loss results in the subdivision of continuous habitat into smaller patches that are isolated from each other by a matrix of modied habitat (Didham 2010). 4.5.1.1þ General Patterns Despite numerous and increasing attempts to detect consistent responses of tropical bats to habitat fragmentation, studies to date suggest relatively few generalizations. At the population level, many studies have documented that abundance responses to fragmentation are highly species and ensemble specic. For instance, in the Neotropics, abundances of gleaning animalivorous bats (Pons and Cosson 2002; Meyer etþ  al. 2008; Meyer and Kalko 2008a) and certain forest-dependent aerial insectivores (Estrada Villegas etþ  al. 2010) decline in response to fragmentation, whereas frugivorous and nectarivorous bats often increase (Sampaio etþ  al. 2003; Delaval and Charles-Dominique 2006; Meyer and Kalko 2008a). In the Paleotropics, insectivorous bat species that roost in tree cavities or foliage are more vulnerable to fragmentation than are cave-roosting species (Struebig etþ  al. 2008, 2009). At the assemblage level, studies that have compared fragmented and continuous forest in terms of species richness, diversity, and composition demonstrate inconsistent responses (Cosson etþ  al. 1999; Schulze etþ  al. 2000; Estrada and Coates-Estrada 2002; Faria 2006). Differences among sites with regard to fragmentation history and structural contrast between fragments and the surrounding matrix complicate the detection of general patterns. This may be a more important issue for the study of tropical bats compared to other taxonomic groups because of the wide range of dispersal abilities exhibited by chiropteran species. 4.5.1.2þ Area and Isolation Effects Early fragmentation studies generally emphasized the effects of area and isolation, reecting the pervasive inuence of island biogeographic theory (IBT, MacArthur and Wilson 1967) in ecology, while ignoring inuences of the surrounding landscape matrix. This same pattern is also apparent within the fragmentation literature on tropical bats. Studies have found evidence for effects of both fragment area (Cosson etþ  al. 1999; Struebig etþ  al. 2008, 2011) and isolation (Estrada etþ  al. 1993a; Meyer and Kalko 2008a, b) on populationand assemblage-level responses, whereas effects were weak or absent in others (Faria 2006; Pardini etþ  al. 2009). Moreover, bat ensembles and species often respond differentially to fragment area or isolation, with responses of some taxa being particularly strong (Struebig etþ  al. 2008; Estrada Villegas etþ  al. 2010). The relative importance of isolation versus area in shaping bat responses to fragmentation is governed by three main factors: the range of fragment sizes relative to isolation in the landscape, the history of landscape change (time since

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70 C.F.J. Meyer et al.isolation, rate of change), and, probably most importantly, the type and quality of matrix habitats in which fragments are embedded. For instance, the high explanatory power of area relative to isolation reported by Struebig etþ  al. (2008) likely reects the low structural contrast between fragments and matrix (mostly rubber and oil palm plantations), limited range of isolation distances compared to area in the study system, and a possible time lag in the realization of isolation effects due to landscape change being fairly recent. In contrast, isolation rather than island area best predicted bat species richness and composition on Neotropical land-bridge islands (Meyer and Kalko 2008a) where fragments were surrounded by water. The simplied dichotomous view of landscapes underlying IBT, albeit applicable in special cases (e.g., land-bridge islands), often fails to capture the inuence that other land-cover types in the surrounding matrix can have and so may not be broadly applicable to most anthropogenically modied landscapes (Kupfer etþ  al. 2006; Laurance 2008). After more than 40þ  years of research beyond the origins of IBT, it is now clear that for most animal taxa, including tropical bats, the majority of terrestrial habitat fragments are not islands in a homogeneous sea of inhospitable habitat. Indeed, island ecosystems support tropical bat biodiversity in fundamentally different ways compared to complex agricultural mosaic landscapes, the former adhering to IBT predictions of species loss, while countryside ecosystems are capable of maintaining high levels of species richness, evenness, and compositionally novel assemblages in human-made habitats (Mendenhall etþ  al. 2014). 4.5.1.3þ Responses to Landscape Structure Fragmentation studies have increasingly shifted their focus from being largely patch-centered toward taking a broader landscape-scale approach, thus acknowledging the overriding importance of the matrix and the existence of gradients of habitat conditions and quality as crucial determinants of species responses (Kupfer etþ  al. 2006; Driscoll etþ  al. 2013; Cisneros etþ  al. 2015). Such gradients are provided, for example, by mosaics of old-growth forest, successional habitat, and different forms of agriculture. This paradigm shift is to some degree reected within the more recent bat literature, as a growing number of studies have adopted matrix-inclusive approaches to studying fragmentation, although overall the number of studies is still small. In the broader literature, empirical evidence suggests widespread negative effects of habitat loss on many taxa (i.e., reduced abundance or density), whereas the effects of fragmentation per se are generally much weaker and may vary strongly in magnitude and direction of response (Fahrig 2003). In agreement with this, for est cover is a better predictor of bat assemblage characteristics (species richness or composition) than are measures of landscape conguration in Neotropical landbridge island systems (Meyer and Kalko 2008a; Henry etþ  al. 2010). On the other hand, consistent responses to landscape composition or conguration at the assemblage level were harder to identify in studies conducted in fragmented Neotropical

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71 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …rain forest landscapes in which the matrix was a mix of anthropogenic land uses (Gorresen and Willig 2004; Klingbeil and Willig 2009, 2010; Cisneros etþ  al. 2015). A difculty facing bat fragmentation studies is that responses tend to be highly species specic, which is often overlooked by diversity metrics applied at the assemblage level (Klingbeil and Willig 2009). This might be more important in low-contrast systems, in which the quality of matrix habitats likely mitigates some of the negative effects of fragmentation on biological communities. At the population level, available evidence suggests that tropical bats respond in complex ways to landscape composition (i.e., the amount of suitable habitat available across the patch types represented in the landscape) and conguration (Gorresen and Willig 2004; Henry etþ  al. 2007b; Klingbeil and Willig 2009, 2010). For instance, Klingbeil and Willig (2009, 2010) found that, apart from being scale dependent (see Sect.þ  4.5.1.4), abundance responses by phyllostomid bats to landscape structure in the Amazon were highly species and ensemble specic, and differed between seasons. In the dry season, abundances of frugivores responded primarily to changes in forest cover (i.e., landscape composition), whereas congurational metrics elicited the strongest response in the wet season. Gleaning animalivores showed the opposite pattern, responding to landscape conguration in the dry season and to landscape composition in the wet season. Such divergent responses suggest an important role of spatiotemporal variation in the abundance and diversity of food resources (Klingbeil and Willig 2010; Cisneros etþ  al. 2015). Together with seasonal differences in time and energy budgets linked to reproduction, these will affect species’ foraging and movement behavior, and could lead to seasonal shifts in diet composition (Durant etþ  al. 2013; Cisneros etþ  al. 2015). Such links remain little explored, yet future research in this regard may prove highly informative. 4.5.1.4þ Spatial and Temporal Scale Dependence in Responses to Fragmentation The scale at which bat species perceive their environment in fragmented landscapes is likely inuenced by spatiotemporal variation in the distribution of resources, as well as by species-specic differences in ecological traits such as diet, wing morphology, and movement behavior. For example, in a low-contrast fragmented system in Malaysia, the provision of large cave systems in the landscape provided clear population subsidies for cave-roosting bats, but also potentially masked the impact of forest fragmentation on this ensemble (Struebig etþ  al. 2009). Consequently, single-scale assessments may be inadequate for capturing the complex interactions between species’ ecology and landscape patterns (Gorresen and Willig 2004). While there is accumulating evidence of the diverse ways by which tropical bats respond to landscape structure, equally important is the increased recognition that the detection of such responses is also sensitive to the spatial scale at which the system is examined (Gorresen etþ  al. 2005).

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72 C.F.J. Meyer et al.Recent studies provide evidence for widespread scale dependence in associations between landscape metrics and bat responses at the assemblage, population, ensemble, and species levels (Gorresen and Willig 2004; Meyer and Kalko 2008a; Pinto and Keitt 2008; Klingbeil and Willig 2009, 2010; Henry etþ  al. 2010; Cisneros etþ  al. 2015). Pinto and Keitt (2008) quantied forest cover at a range of scales (buffers with radii from 50 to 2000þ  m) and found positive associations with bat abundance, whereby the scale that elicited the strongest response was species specic. Differential species responses to forest cover in this case were best explained by interspecic variation in diet, body size, and home range size. Similarly, multiple speciesand ensemble-specic abundance responses of phyllostomid bats to landscape characteristics at multiple focal scales (buffers with 1, 3, and 5 km radii) have been reported from moderately fragmented, lowland Amazonian forest (Klingbeil and Willig 2009) and highly fragmented Atlantic for est in Paraguay (Gorresen and Willig 2004). In both studies, species were demonstrated to interact with their environment simultaneously at a range of spatial scales. In the Amazon, a change in the focal scale of response occurred between dry and wet seasons, a nding which is likely linked to seasonal differences in food abundance and diversity as well as energetic constraints associated with reproduction (Klingbeil and Willig 2010; Cisneros etþ  al. 2015). Scale dependence in response patterns has also been observed in landscapes with an aquatic matrix (Meyer and Kalko 2008a; Henry etþ  al. 2010), suggesting that scale effects are ubiquitous and operate in fragmented landscapes across a broad range of matrix types. Overall, such ndings emphasize that multiscale approaches to determining the effects of landscape structure on tropical bats are essential. In agreement with recent ndings for tropical birds (Banks-Leite etþ  al. 2013), the available evidence suggests, however, that the extremely idiosyncratic responses of tropical bats to landscape structure make it difcult to identify any particular landscape predictor or spatial scale that performs best at predicting responses at the assemblage level. Despite the general importance of a landscape-level perspective in the study of habitat fragmentation, patch characteristics remain important for patch-dependent species (Driscoll etþ  al. 2013). However, fragmentation studies on tropical bats that have jointly assessed the relative contribution of patchand landscape-scale variables for explaining response patterns are scarce. Meyer and Kalko (2008a) found that the relative importance of localversus landscape-scale characteristics in explaining species richness and compositional patterns of phyllostomids on Panamanian land-bridge islands varied with spatial scale. At the patch scale, isolation distance from the mainland was the strongest predictor, whereas the proportion of forest cover in the surrounding landscape was the most prominent descriptor explaining variation in assemblage attributes at larger scales. Although the importance of spatial scale and spatial variation in matrix quality have received some attention in the bat fragmentation literature, we know little about how species responses to fragmentation vary over time or how they are mediated by changes to the matrix. Across many human-modied landscapes in the tropics, secondary forest regrowth may reclaim once deforested

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73 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …land, for instance in response to the abandonment of agriculturally unproductive areas (Bobrowiec and Gribel 2010; Chazdon 2014). Matrix recovery following disturbance can alter responses of fragment biota that may be driven by temporal changes in resource availability and of permeability of the matrix to disper sal (Bissonette and Storch 2007; Driscoll etþ  al. 2013). In this context, research at the Biological Dynamics of Forest Fragments Project in the Brazilian Amazon indicates strong divergence in phyllostomid bat assemblage structure, high levels of species turnover, and marked reorganization in the rank order of the most abundant species in response to changes in matrix quality and composition over 15þ  years (Meyer etþ  al., unpublished data). Time lags in the manifestation of species responses to fragmentation are ubiquitous and constitute an important temporal aspect to consider when studying fragmentation impacts (Ewers and Didham 2006a; Bissonette and Storch 2007), but so far have been rarely investigated in tropical bat studies. Notable exceptions are a series of studies conducted in the St. Eugène land-bridge island system in French Guiana, in which fragmentation effects prior to, and for several years after, fragmentation provided clear evidence for time lags in species loss (Cosson etþ  al. 1999; Pons and Cosson 2002; Henry etþ  al. 2010). These time lags occurred gradually over the course of ca. 10þ  years. Future assessments of tropical bat responses to fragmentation (and other types of anthropogenic disturbance) should therefore address not only the spatial but also the temporal dimension of human impacts. This is particularly notable as long-term studies in intact habitats reveal tropical bat assemblages to be highly dynamic in space and time (Pech-Canche etþ  al. 2011; Kingston 2013). 4.5.1.5þ Edge Effects Recent reviews concur that edge effects critically affect biodiversity in habitat fragments (Ewers and Didham 2006a; Fischer and Lindenmayer 2007; Laurance etþ  al. 2011). However, responses of tropical bats to habitat edges remain under studied, particularly in the Paleotropics. Current evidence from the Neotropics suggests that responses vary according to matrix contrast and land-use history, and are ensemble and species specic. Several studies have modeled bat responses in relation to the amount and complexity of edge habitat, revealing that some tropical bats are sensitive to habitat edges (Gorresen and Willig 2004; Meyer and Kalko 2008a; Klingbeil and Willig 2009, 2010; Henry etþ  al. 2010). While signicant associations between species richness or composition with edge density have been found in fragmented systems with a water matrix (Meyer and Kalko 2008a), studies conducted in a low-contrast landscape did not detect signicant edge-related responses at the assemblage level (Gorresen and Willig 2004; Klingbeil and Willig 2009, 2010). This again under lines the importance of matrix contrast in affecting species’ edge sensitivity and also shows that, at least in landscapes with low-contrast edges, composite community measures such as species richness may fail to capture edge responses that may

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74 C.F.J. Meyer et al.otherwise be evident at the species or ensemble level (Klingbeil and Willig 2009). At the population level, abundances of six frugivorous and gleaning animalivorous phyllostomid bat species in the Peruvian Amazon were positively related to edge density, whereby responses varied depending on spatial scale (Klingbeil and Willig 2009) and season (Klingbeil and Willig 2010). In contrast, in fragmented Atlantic forest, two frugivorous species exhibited negative responses to edge density (Gorresen and Willig 2004). The discrepancy in the direction of response may be explained by differences in the prevailing patterns of land conversion (smallvs. large-scale deforestation). A strong negative response of gleaning animalivores to edge cover was also found by Henry etþ  al. (2010) in a land-bridge island system in French Guiana. These studies indicate the sensitivity of phyllostomid bats to edges driven by changes in landscape conguration. However, quantifying the strength of edge effects requires explicit consideration of two distinct aspects: edge extent and edge magnitude. Edge extent is the distance over which a change in the response variable can be detected, and edge magnitude is the amplitude of the effect (Harper etþ  al. 2005; Ewers and Didham 2006b). The few studies that have examined the magnitude of edge effects on tropical bats by comparing interior sites of large, mature forest stands and forest edges reported declines in phyllostomid richness, in landscape matrices of high (water; Meyer and Kalko 2008a) and low structural contrast (secondary forest and shade cacao plantations; Faria 2006). The pattern of reduced species richness at edges in the low-contrast system was mainly attributable to the decline of gleaning animalivorous species (Faria 2006; Pardini etþ  al. 2009). Even though species composition did not signicantly change between forest edge and interior, Meyer and Kalko (2008a) found that gleaning animalivorous bats exhibited a strong negative numerical response toward edges. In fact, edge sensitivity was identied as the species trait that best explained species vulnerability to fragmentation (Meyer etþ  al. 2008). Similar to phyllostomids, aerial insectivorous bats in the same land-bridge island system had signicantly lower species richness at edges compared to interiors. The two functional groups of narrow-space foragers and open-space bats responded differently to forest edges. Open-space foragers had higher abundance counts at edges, whereas those of for est species were not signicantly altered (Estrada Villegas etþ  al. 2010). Comparing general bat activity, Estrada etþ  al. (2004) did not detect signicant differences between continuous forest interiors and forest–pasture edges. Only one study to date has tried to quantify the distance of edge inuence for tropical bats. Delaval and Charles-Dominique (2006) captured phyllostomid bats along 3-km transects perpendicular to the edges of a road traversing primary for est in French Guiana. Capture rates along the transects were more than seven times higher than those at a control site, 150þ  km inside the primary forest block. Moreover, along the transects abundances decreased with increasing distance from the road edge, a pattern attributable to the proliferation of opportunistic frugivores such as Carollia perspicillata and Artibeus jamaicensis that exploit abundant food resources provided by young regrowth along road margins. Species richness decreased signicantly with distance from the road edge, probably related to an

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75 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …inux of species from the open habitat into the edges. Species richness at edges was, however, not signicantly greater than that in the control site that harbored seven species not present at road edges or along transects. Differences in rank abundance patterns between transects and control site provided further evidence that even narrow road clearings can alter bat assemblage structure over distances of at least 3þ  km into forest interiors. Key research needs: Studies that try to disentangle the relative importance of habitat amount and habitat conguration in shaping species responses, in particular studies that identify portions of the gradient in habitat amount within which the effects of spatial arrangement become important, i.e., explicit tests of the “habitat threshold hypothesis” (Fahrig 2003). Research that addresses the relative tolerance of different species to changes in habitat conguration (see Villard and Metzger 2014). Studies that jointly assess the relative contribution of patchand landscape-scale variables to explaining response patterns. Long-term investigations that address the effects of matrix transformation on bat species responses over time. More studies that quantify edge effects in terms of both magnitude and extent. Further research investigating how consistently species respond to habitat edges across a broad range of edge types to identify ecological traits correlated with and potentially driving edge sensitivity (Ries and Sisk 2010). Studies that try to disentangle edge and area effects (Fletcher etþ  al. 2007; Banks-Leite etþ  al. 2010).4.5.2þ LoggingRain forests are selectively logged at 20 times the rate at which they are cleared (Asner etþ  al. 2009), and large expanses (403þ  millionþ  ha) are ofcially designated for timber extraction (Blaser etþ  al. 2011). Selective logging exposes vast areas to potentially detrimental edge effects (Broadbent etþ  al. 2008) and may often be the precursor to complete deforestation (Asner etþ  al. 2006). Yet, the impacts of selective logging on biodiversity depend critically on the harvest intensity (Asner etþ  al. 2013; Burivalova etþ  al. 2014) as well as the extraction techniques (Bicknell etþ  al. 2014). Selective harvesting methods range from large-scale conventional extraction that can cause substantial loss in canopy cover and associated mortality of non-harvested trees, to reduced-impact logging (RIL), in which collateral damage is reduced as a result of improved planning and control of harvesting activities (Putz etþ  al. 2008; Asner etþ  al. 2013). Recent meta-analyses indicate that selectively logged forests can retain a large proportion of the diversity of old-growth forest for a variety of taxa (Gibson etþ  al. 2011; Putz etþ  al. 2012) and the available evidence, though scant due to the

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76 C.F.J. Meyer et al.low number of studies (Fig.þ  4.3a), largely supports this notion for tropical bats (Bicknell etþ  al. 2014). At the assemblage level, selective logging appears to have little or no effect on bat species richness in the Neotropics (Ochoa 2000; Clarke etþ  al. 2005a, b; Castro-Arellano etþ  al. 2007). In contrast, compositional or structural differences between bat assemblages in logged and unlogged sites are more common, which suggests that if forests are unable to recover from logging distur bance, species losses may be detected in the long term (i.e., similar to time lags for fragmentation effects, see Sect.þ  4.5.1). Structural differences between bat assemblages in unlogged and logged forests are evident from changes in the propor tional abundance of species within ensembles (Clarke etþ  al. 2005a, b; Peters etþ  al. 2006) and shifts in species rank distributions and dominance (Castro-Arellano etþ  al. 2007). A consistent pattern emerging from Neotropical studies is that, similar to habitat fragmentation (see Sect.þ  4.5.1), selective logging appears to adversely affect the abundance of gleaning animalivorous phyllostomids, whereas frugivorous and nectarivorous species tend to increase in abundance (Ochoa 2000; Clarke etþ  al. 2005a, b; Peters etþ  al. 2006; Presley etþ  al. 2008). In a study in Trinidad, Clarke etþ  al. (2005a) found that the magnitude of change in species composition is linked to the intensity of timber harvesting. Comparing a continuous logging system with few harvest controls (open range [OR] system) to a polycyclic, selective system that incorporated stricter controls on felling (periodic block [PB] system), the study demonstrated that PB-managed sites resembled undisturbed primary forest much more closely in bat species composition and abundance than did OR forest. Despite structural changes associated with PB management, bat assemblages in such well-managed forest stands had great potential for recovery to near predisturbance levels (Clarke etþ  al. 2005b). The number of years post-logging was positively correlated with the number and abundance of species of gleaning animalivores but not frugivores, whereas the proportional abundance of the dominant frugivore decreased with forest recovery. Together, these ndings suggest that PB or similar low-intensity selective management systems may be compatible with the conservation of bat diversity. Unfortunately, similar studies that evaluate responses of tropical bats to different management systems or across a series of logged sites of different ages within the same general study landscape are lacking. Short-term population-level responses of phyllostomid bats to RIL in Amazonia were idiosyncratic (Castro-Arellano etþ  al. 2007) and RIL sites had reduced species richness, linked to the local absence of rare species from logged forest, whereas the populations of common species remained unaffected (Presley etþ  al. 2008). As argued by Presley etþ  al. (2008), landscape context may be important in mediating the effects of RIL on bats, and for this harvesting practice to be sustainable, it may be essential that RIL blocks be located in close proximity to undisturbed forest to facilitate rescue effects that can mitigate the negative impacts of RIL on rare species. Furthermore, due to the short post-harvest interval (<42þ  months) in both studies, the observed responses may be short term (Castro-Arellano etþ  al. 2007; Presley etþ  al. 2008), stressing the necessity for longer-term evaluations of logging impacts.

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77 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …In the only logging effect study on African bats, Monadjem etþ  al. (2010), using acoustic sampling, found no signicant differences in activity levels between primary and logged forests in Uganda for the insectivorous Neoromicia nana. Elsewhere in the Paleotropics, early studies reported higher species richness, diversity, and abundance in unlogged compared to selectively logged forest in Malaysia (Zubaid 1993) and profound changes in species composition due to logging in Sumatra (Danielsen and Heegaard 1995). However, in addition to having small sample sizes, these studies employed only mist nets, which are ineffective at capturing the numerous insectivorous species that dominate Paleotropical bat assemblages (Kingston 2013). Conclusions based on these studies alone should therefore be interpreted with caution. More recent studies in Southeast Asia have employed larger sampling effort and harp traps, which are adequate for sampling forest interior insectivores. In peninsular Malaysia, a comparison of forest reserves and adjacent logged-over forests >30þ  years post-extraction showed little overall difference in assemblage composition (Christine etþ  al. 2013). In nearly all site comparisons, species richness and abundances were higher in logged forest. However, certain treeor foliage-roosting species were only captured inside forest reserves, suggesting that forest reserves embedded in a matrix of production forest could play an important role as reservoirs to restock logged forest and to maintain populations of disturbance-sensitive species (Christine etþ  al. 2013). Logging effects may multiply spatially and temporally as a result of multiple harvesting cycles (Lindenmayer and Laurance 2012). However, only recently have researchers examined the impacts of multiple rounds of extraction. One such study examined bat assemblages on Borneo across a disturbance gradient ranging from old-growth to twice-logged to repeatedly logged forest (Struebig etþ  al. 2013). Logging had little effect on bat species richness, even in heavily degraded forest that had been logged multiple times, corroborating research on other taxa in the region (Edwards etþ  al. 2011). Changes in insectivorous bat assemblage structure and abundance between old-growth and repeatedly logged forest were nonetheless evident and degraded sites that were characterized by a low, open canopy har bored a depauperate bat fauna. Canopy height was an important determinant of assemblage change across the disturbance gradient, as was the availability of tree cavities for forest-roosting taxa. By quantifying microhabitat over the gradient, the study revealed that post-logging recovery of assemblages could be enhanced via restoration investments in canopy cover and tree cavity availability. Moreover, cave-dwelling hipposiderid and rhinolophid bats were less abundant in repeatedly logged sites, in line with ndings from a study in Vietnamese karst forests in which these taxa were also less abundant in logged than in primary forest (Furey etþ  al. 2010). A key theme emerging from the recent logging effect literature is the potential confounding issue of spatial pseudoreplication in study design, a problem whereby study sites in continuous forest stands are inappropriately treated as independent replicates (Ramage etþ  al. 2013). The most effective way to overcome these problems is to sample the same forest sites before and after logging. The only batlogging study to have implemented such a robust Before–After–Control–Impact

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78 C.F.J. Meyer et al.(BACI) design to date was undertaken in RIL forests in Guyana (Bicknell etþ  al. 2015). Differences in bat assemblage structure before and after logging were relatively weak and varied substantially across study sites. Although three species were classied as indicators of disturbed or undisturbed forest, there were no clear changes in bat assemblages at control sites, indicating that overall responses could not be reliably attributed to logging. In conclusion, given the paucity of studies available, it remains difcult to ascertain denitive responses of tropical bats to logging. The short-term effects appear to be relatively benign, especially in low-intensity extraction systems. Reported effects vary, largely owing to differences among studies with regard to the type of forest management system, and spatial and temporal variability in disturbance attributes, including time post-harvest. Key research needs: Studies comparing bat responses between different forest management systems and across a range of spatial and temporal scales. More studies implementing BACI designs, as exemplied by Bicknell etþ  al. (2015). Integration of logging disturbance into studies of forest fragmentation in order to distinguish true fragmentation responses from those of forest degradation.4.5.3þ Secondary Forests and SuccessionThe future of tropical biodiversity will critically depend on our ability to manage the large expanses of regenerating secondary forests (Chazdon etþ  al. 2009; Chazdon 2014) that account for approximately half of the remaining area of tropical moist forests (Asner etþ  al. 2009). Studies that have examined the conservation value of secondary forests for tropical bats are largely in line with assessments with regard to other tropical taxa (Barlow etþ  al. 2007; Gardner etþ  al. 2010) by suggesting that regenerating forests act as important repositories of bat biodiversity. Secondary forests are effective at conserving a subset of primary forest bat species richness (Louzada etþ  al. 2010), but usually host assemblages that differ in structure and composition from those in mature forest (Faria 2006; Barlow etþ  al. 2007). Secondary successional vegetation in Neotropical humid forests represents important habitat for many frugivorous and nectarivorous phyllostomids (e.g., Carollia spp., Sturnira spp., Glossophaga spp.). These taxa become numerically dominant in secondary forests representing early to intermediate stages (Brosset etþ  al. 1996; Castro-Luna etþ  al. 2007a, b; Willig etþ  al. 2007; de la PeñaCuéllar etþ  al. 2012; Vleut etþ  al. 2013). This pattern is likely attributable to an increase in the abundance, diversity, or quality of fruit and ower resources associated with early successional vegetation and emphasizes the fundamental impor tance of phyllostomid bats in the regeneration of tropical forests (Muscarella and Fleming 2007). In contrast, the abundance of frugivores was not elevated in earlier

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79 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …successional stages of tropical dry forest in Mexico (Avila-Cabadilla etþ  al. 2009). This likely reects distinct differences in the composition of early successional vegetation, and consequently resource scarcity, in tropical dry compared to wet forests. Pinto and Keitt (2008) found that the abundances of Sturnira spp. were positively associated with secondary forest cover, reecting the species’ preference for early successional vegetation. Conversely, Carollia spp. responded to forest cover that included both primary and secondary forests, implying that habitat connectivity may be more important than successional stage for populations in this genus. As with logged forests, these ndings suggest species-specic responses to secondary vegetation linked to interspecic differences in diet, home range size, and body size. Contrary to the exible responses observed for many frugivores and nectarivores, a large body of empirical evidence indicates that gleaning animalivorous phyllostomines are sensitive to forest degradation, as they are absent or occur at low abundance in secondary regrowth (Fenton etþ  al. 1992; Brosset etþ  al. 1996; Medellín etþ  al. 2000; Faria 2006; Castro-Luna etþ  al. 2007a, b; Mancina etþ  al. 2007; Willig etþ  al. 2007; Pardini etþ  al. 2009; Bobrowiec and Gribel 2010; de la PeñaCuéllar etþ  al. 2012; Vleut etþ  al. 2012, 2013). Some studies have detected a clear pattern of species richness increasing across successional gradients (Avila-Cabadilla etþ  al. 2009; de la Peña-Cuéllar etþ  al. 2012), but this pattern has not been evident in others (Castro-Luna etþ  al. 2007a; Mancina etþ  al. 2007). Nonetheless, for Neotropical wet and dry forests, oristically more diverse and structurally more complex habitats harbor greater taxonomic and functional richness than do early or intermediate stages of succession. Here, vegetation complexity appears to be an important factor shaping assemblage composition (Medellín etþ  al. 2000; Avila-Cabadilla etþ  al. 2009; Bobrowiec and Gribel 2010; Avila-Cabadilla etþ  al. 2012; de la Peña-Cuéllar etþ  al. 2012). Late successional forest stands often host many bat species not found in earlier stages, in particular rare taxa, and through succession, the number of species and ensembles increases for frugivorous, nectarivorous, and gleaning animalivorous taxa (AvilaCabadilla etþ  al. 2009, 2012; de la Peña-Cuéllar etþ  al. 2012). In tropical wet forest in Mexico, abundances of the most common bat species were associated positively or negatively with variation in canopy cover across successional stages, rather than with landscape attributes (Castro-Luna etþ  al. 2007a). In contrast, a study in Mexican tropical dry forest found evidence for an important role of local (vegetation complexity) and landscape attributes (area and cover of different vegetation types) as determinants of variation in abundance, which were ensemble specic and scale dependent (Avila-Cabadilla etþ  al. 2012). In Central Amazonia, gleaning animalivorous phyllostomid bats exhibited greater abundance and richness in Cecropia-dominated regrowth, whereas stenodermatine frugivores were more abundant in abandoned pastures and Vismia-dominated regrowth, demonstrating that different successional trajectories result from differences in land-use history (cutting versus cutting and burning) that lead to distinct differences in bat assemblage composition (Bobrowiec and Gribel 2010). Despite the recovery potential of Neotropical bat assemblages during succession, the conservation value of secondary forests for bats critically hinges

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80 C.F.J. Meyer et al.on landscape context and is maximized in mosaic landscapes in which patches of forest at different successional stages are located close to old-growth forest (Bobrowiec and Gribel 2010; Vleut etþ  al. 2012). Key research needs: Comprehensive assessments of the conservation value of secondary forests for bats in the Paleotropics, which are essentially lacking (but see Fukuda etþ  al. 2009). Studies addressing the recovery potential of Paleotropical bat assemblages dur ing secondary succession.4.5.4þ Agroforestry SystemsAs agriculture and associated biodiversity losses continue to rise across the tropics, agroforestry systems have been advocated as biodiversity-friendly alternatives, capable of conserving biodiversity while enhancing rural livelihoods (Perfecto and Vandermeer 2008; Clough etþ  al. 2011). Coffee (Coffea arabica, Coffea canephora) and cacao (Theobroma cacao) are the principal cash crops of many tropical countries (Donald 2004; Tscharntke etþ  al. 2011) and are the primary examples in the bat literature (but see bat inventories of Sumatran rubber agroforests in Prasetyo etþ  al. 2011). In traditional coffee and cacao agroforestry, these crops are commonly grown under a stratied canopy layer of a more or less diverse range of native shade tree species. Much of their potential for conservation derives from the fact that such traditional agroforestry systems resemble natural forest habitat in many structural aspects (Perfecto and Vandermeer 2008). Empirical studies that have assessed the value of agroforests for tropical bats to date come almost exclusively from the Neotropics (Fig.þ  4.3a). Pineda etþ  al. (2005) compared the bat fauna of Mexican cloud forest fragments and shade coffee plantations and found that both habitats had very similar species richness and composition, although there were changes in the species’ rank order between habitats. Large frugivorous phyllostomids (Artibeus spp.) reached higher abundance in shade coffee than in the natural habitat, possibly as a result of increased food availability due to the cultivation of important fruit tree species alongside coffee, a management strategy that also favored the abundance and richness of fruitand nectar-eating bats in coffee plantations elsewhere in Mexico (Castro-Luna and Galindo-González 2012a). Contrasting abundance responses for large Artibeus were found in another study in Mexico (Saldaña-Vázquez etþ  al. 2010). Here, shade coffee plantations and disturbed cloud forest fragments did not differ in abundance levels and also had similar availability of food plants. On the other hand, abundances of Sturnira spp. were higher in forest fragments, probably linked to a decline in food resources for these small frugivores in the coffee plantations. This reduction in resources resulted from the pruning of understory vegetation and was reinforced by the effects of a resource-poor pasture matrix surrounding the forest fragments.

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81 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …Williams-Guillén and Perfecto (2010, 2011) investigated how bat diver sity patterns in coffee agroforestry change with increasing management intensity. Phyllostomid bats maintained similar richness across management regimes, but showed signicant declines in abundance across the intensication gradient, from forest fragments through low-management shade polyculture and commer cial polyculture to high-management coffee monocultures (Williams-Guillén and Perfecto 2010). Compositional similarity differed signicantly between fragments and coffee plantations of all management intensities, and between high-shade polycultures and low-shade monocultures. The proportions of large frugivores increased with management intensity, in line with Pineda etþ  al.’s (2005) ndings. Conversely, those of nectarivorous and gleaning animalivorous bats decreased, the latter being absent from intensively managed coffee monocultures. Both for est fragments and the diverse and structurally complex shade polyculture systems may provide adequate roosting and food resources to sustain high levels of phyllostomid diversity. This contrasts strongly with the situation in low-shade monocultures, which offer reduced feeding and roosting opportunities, and may consequently serve more as commuting than foraging habitat. This was also suggested in a study on non-phyllostomid aerial insectivorous bats in the same landscape, which reported reduced foraging activity in the most intensively managed monocultures (Williams-Guillén and Perfecto 2011). Both of the functional groups of aerial insectivores, forest and open-space foragers, had similar species richness across habitat types. The two groups, however, showed opposite responses with respect to activity levels and compositional similarity. Forest-adapted species differed in ensemble composition across the management gradient and responded negatively to agricultural intensication in terms of activity. For open-space foragers, reductions in shade tree diversity and cover did not manifest in compositional changes, but were associated with increased levels of overall activity, albeit not feeding activity. Collectively, these studies demonstrate the high conservation value of structur ally diverse shade coffee for bats, but less so of intensively managed systems. The former constitutes a permeable high-quality matrix, while intensive coffee monocultures represent poor matrix habitat (Numa etþ  al. 2005). Landscape context, in particular the dominant matrix type, is an important modulator of how bat assemblages respond to agroforest management intensity. Forest fragments harbored signicantly greater phyllostomid richness than did management systems when the landscape matrix was dominated by sun coffee, whereas richness was similar among habitats in a shade coffee matrix (Numa etþ  al. 2005). For cacao, studies show results similar to those for coffee, supporting the notion that traditional, structurally complex shade cacao plantations sustain high levels of bat diversity. Insights come from a series of studies conducted in the Atlantic for est region of Una, Brazil. Cacao agroforests in this region provide foraging and roosting habitat for members of all feeding ensembles, including forest-þ­ dependent gleaning animalivorous species (Pardini etþ  al. 2009), primarily because of the structural complexity retained compared to intact forest (Faria etþ  al. 2006). In fact, bat assemblages in shade cacao showed greater richness, diversity, and abundance

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82 C.F.J. Meyer et al.than did those in nearby mature or secondary forest (Faria 2006; Faria and Baumgarten 2007; Pardini etþ  al. 2009). However, shade cacao plantations per se may not provide adequate habitat conditions for forest-dwelling bats, as the proximity of shade cacao to forest remnants was a key determinant of species persistence. Bat assemblages in plantations isolated by more than 1þ  km from forest were characterized by low richness and diversity, with clear shifts in species dominance, suggesting a crucial role of native forest remnants as population sources (Faria and Baumgarten 2007). Isolating distance to forest was also an important factor inuencing species richness and abundance in Mexican shade plantations (Estrada etþ  al. 1993a). These plantations maintained diverse and structurally similar bat assemblages to those in remnants of native forest (Medellín etþ  al. 2000; Estrada and Coates-Estrada 2001b). As for coffee (Numa etþ  al. 2005), landscapes dominated by cacao agroforests and comprising reduced native forest cover may harbor impover ished bat assemblages (Faria etþ  al. 2006; 2007), highlighting that landscape context generally plays a crucial role in determining bat species responses in tropical agroforestry landscapes, as it does for fragmented forest systems. In conclusion, both coffee and cacao, when grown under a traditional shade regime, comprise a high-quality matrix that offers suitable conditions for maintaining diverse phyllostomid assemblages. These agroecosystems, in turn, benet from pest control services provided by bats as has been shown for agroforests in the Neotropics (Williams-Guillén etþ  al. 2008) and Southeast Asia (Maas etþ  al. 2013) (see Chap. 6). Studies in cacao agroforestry at least in some cases entailed comparison between large tracts of mature forest and the agricultural system (Medellín etþ  al. 2000; Faria 2006), but these important baseline data are lacking for studies in coffee agroforests. Key research needs: Studies that assess response patterns for non-phyllostomid bats. Assessments of bat responses to cacao agroforestry intensication, especially in view of globally increasing levels of conversion of shade cacao systems into unshaded monocultures (Tscharntke etþ  al. 2011). Linkages between levels of bat biodiversity and crop yields.4.5.5þ Tree PlantationsGiven the extent to which forested land is being converted to tree plantations across much of the tropics (Gibbs etþ  al. 2010), there have been surprisingly few studies investigating the value of these habitats for bats. Three systems dominate tree plantation mosaics in the tropics: fast-growing timbers for the paper/pulp industry (e.g., Acacia, Eucalyptus), rubber (Hevea brasiliensis), and, increasingly, oil palm (Elaeis guineensis).

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83 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …In a multitaxon assessment in Brazil, Barlow etþ  al. (2007) found similar numbers of bat species in Eucalyptus plantations and secondary forests recovering from burning, but both habitats supported much lower richness than did unlogged forests. Bat assemblages in plantations were nested subsets of those in forests; approximately 11þ  % of all species were shared between plantations and primary forest, 4þ  % were shared with secondary forest, and 39þ  % found in all habitats (Louzada etþ  al. 2010). Nevertheless, three species (ca. 6þ  % of total) were captured exclusively in Eucalyptus plantations. A study in Brazilian Cerrado found lower species richness, diversity, and evenness of bat assemblages in Eucalyptus monocultures than in fragments of native Cerrado vegetation (Pina etþ  al. 2013). Gleaning animalivorous phyllostomid bats were not captured in plantation forests. An earlier comparative study in Sumatra documented a distinct shift in bat assemblage structure in rubber and oil palm plantations, which supported only 13–25þ  % of the bat species richness found in forest (Danielsen and Heegaard 1995). However, more recent surveys have revealed additional species utilizing rubber plantations, especially those grown as agroforests or close to forest areas (Prasetyo etþ  al. 2011). These studies point to an adverse response by bats to plantation development in both the New and Old World tropics. However, the extent to which these ndings reect true bat declines versus sampling bias (i.e., difculties in capturing bats in open plantation habitats) is open to question. Tree plantations present a much more open habitat compared to forests, but can provide canopy structure similar to that in forest. This may present difculties for capturing bats in these habitats, particularly in the Paleotropics, where much of the insectivorous bat fauna can only be captured in harp traps. Bat surveys in Sumatra and Borneo have resulted in extremely low capture rates for insectivorous species in oil palm plantations using mist nets and harp traps (Fukuda etþ  al. 2009; Syamsi 2013), a nding that could reect differ ential capture success in closed versus open habitats as well as true differences between habitats. Acoustic surveys could potentially contribute additional infor mation concerning bat activity and the structure of bat assemblages in these habitats. The rst insights from the Old World come from southern Thailand, where Phommexay etþ  al. (2011) sampled bats in forest and neighboring rubber plantations using bat detectors, mist nets, and harp traps. Although diversity and overall bat activity were much lower in plantations than in forests, differences between the two habitat types were not as severe as indicated by capture-based surveys. Acoustic sampling in plantations detected less than half the number of bat species found in forest and fewer bat passes. Although bat activity was clearly reduced in plantations, a substantial number of feeding buzzes were detected, suggesting that bats were still foraging in this modied habitat. Key research needs: Further studies, particularly those using acoustic methods, to accurately assess the conservation value of tree plantations for tropical bats.

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84 C.F.J. Meyer et al.4.5.6þ Agriculture and Residual Tree CoverAgricultural encroachment and cropland expansion are key threats to biodiver sity in tropical countries (Phalan etþ  al. 2013). The dominant crop will determine the permeability of the agricultural matrix, the likelihood of species persistence, and ultimately whether sustainable congurations in human-modied landscapes emerge in which biodiversity conservation and food production can be reconciled (Melo etþ  al. 2013). Apart from several studies in agroforestry systems (see Sect.þ  4.5.4) and oil palm plantations (see Sect.þ  4.5.5), little research has examined responses of tropical bats to forest conversion into other agricultural land uses, or the value of residual vegetation in agricultural matrices (Fig.þ  4.3a). By far, most of the available evidence comes from studies in Mexico and Central American tropical wet and dry forests. These studies generally suggest that human-modied landscapes comprising a heterogeneous mosaic of different landand tree-cover types can preserve species-rich bat assemblages (Estrada etþ  al. 1993a, b, 2004; Medellín etþ  al. 2000; Moreno and Halffter 2001; Estrada and Coates-Estrada 2002; Harvey etþ  al. 2006; Medina etþ  al. 2007; Barragan etþ  al. 2010; Mendenhall etþ  al. 2014). For instance, in a comparison of bat diversity in forest fragments, agricultural habitats, and live fences in Mexico, agricultural habitats contained 77þ  % of the species recorded, whereby species richness declined with increasing distance from forest fragments (Estrada etþ  al. 1993a). Certain frugivorous species (e.g., Carollia spp., Sturnira spp.) may become dominant in agricultural areas, whereas phyllostomine species are adversely affected by agriculture (Medellín etþ  al. 2000). A similar pattern was found by Willig etþ  al. (2007) in lowland Amazonian rain forest in Peru. Here, half of the frugivorous and nectarivorous species that responded consistently to habitat conversion reached highest abundances in agricultural areas, a response probably linked to the ample food resources provided by these habitats. Due to the presence of rare species not captured in forest, species richness in disturbed agricultural and early successional habitats was high compared to that in mature for est. However, the long-term persistence of most species likely still depends on the availability of forest (Willig etþ  al. 2007). Moreover, these ndings relate to smallscale habitat conversion and may not be generalizable to landscapes characterized by large-scale deforestation. Knowledge of the conservation value of agricultural habitats for bats in the Old World is scant (see Chap. 6). In a study in Fiji (Luskin 2010), foraging densities of the Pacic ying fox, Pteropus tonganus, an important seed disperser were four times higher in agricultural habitats than in remnants of dry forest, illustrating a strong preference for foraging on abundant food resources in farmland. Resource subsidies provided by farmland were responsible for sustaining high abundances of the species despite severe deforestation across the region. Roosting sites, however, were restricted to native forest fragments, highlighting their importance for population persistence. Agricultural habitats provided important resources for some species of pteropodid bats in Borneo, as evidenced by high capture rates

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85 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …in orchards relative to forest habitats (Fukuda etþ  al. 2009). Fukuda etþ  al. (2009) suggest that some pteropodids in Southeast Asian dipterocarp forests, which are characterized by a supra-annual owering and fruiting pattern, may augment food resources by feeding on cultivated plants during non-owering periods when food supply in the forest is scarce. However, other fruit bat species were restricted to forest, suggesting that the value of agricultural land is species specic. Sedlock etþ  al. (2008) reported that fewer species persist in mixed agricultural habitat than in tall secondary forest in the Philippines. Nevertheless, 19 of 26 species were present in agro-pastoral areas. Results from studies in the Paleotropics are thus largely congruent with those from the Neotropics in suggesting that agricultural habitats harbor considerable bat diversity and provide important foraging habitat for some fruit bat species. Linear landscape elements (corridors of residual vegetation such as live fences or strips of riparian forest) and scattered trees, commonly found in Neotropical countryside landscapes, may enhance functional connectivity (Villard and Metzger 2014), and studies indicate that bats extensively use them (Estrada and CoatesEstrada 2001a; Galindo-González and Sosa 2003; Estrada etþ  al. 2004; Harvey etþ  al. 2006; Medina etþ  al. 2007; Barragan etþ  al. 2010). For instance, in agricultural landscapes in Nicaragua, riparian forests and live fences harbor greater bat species richness and abundance than do secondary forest and pastures with low tree cover (Harvey etþ  al. 2006; Medina etþ  al. 2007). Riparian forests constitute favorable habitats for foraging and roosting, particularly in tropical dry for est ecosystems, where they often have higher tree diversity and food availability compared to other types of cover (Estrada and Coates-Estrada 2001a; Harvey etþ  al. 2006). Live fences and riparian corridors facilitate movement by bats across fragmented agricultural landscapes and may effectively reduce isolation between remnant forest patches, which, in turn, enhances species persistence at the landscape level. Similar to live fences, isolated pasture trees provide food and roosting opportunities for bats and act as important stepping stones for bat movement (Galindo-González and Sosa 2003), suggesting that they can render agro-pastoral landscapes more hospitable to bats and consequently deserve attention in conser vation strategies. In contrast, studies concur that pastures are low-quality habitat for bats, likely as a consequence of resource scarcity (food, roosts) and elevated predation pressure (Estrada etþ  al. 1993a, b, 2004; Harvey etþ  al. 2006; Griscom etþ  al. 2007; Medina etþ  al. 2007). Key research needs: In-depth studies in the Old World tropics that assess bat responses across a range of agricultural habitat types and landscape settings. Assessments of the value of residual tree cover in agricultural matrices for Paleotropical bats, particularly in Africa. Research addressing the effects of large-scale, commercial agriculture (e.g., cultivation of soybean, corn, sugarcane), which plays an increasingly signicant role in driving deforestation in some tropical regions such as the Amazon.

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86 C.F.J. Meyer et al.4.6þ Genetic ConsequencesTropical taxa are generally underrepresented in landscape genetic studies (Storfer etþ  al. 2010). Bats are no exception, as only few studies have assessed how they are affected by anthropogenic habitat loss and fragmentation at the genetic level (Fig.þ  4.3b). Meyer etþ  al. (2009) studied populations of two Neotropical bats in fragments that were isolated by a water matrix and detected signicant population differentiation that matched the species’ relative mobility. In contrast to the more mobile canopy frugivore, Uroderma bilobatum, population subdivision in the understory frugivore, C. perspicillata, showed a signicant effect of fragmentation and isolation by distance, as well as reduced genetic diversity on islands relative to mainland populations. Also employing mitochondrial DNA sequence data, Ripperger etþ  al. (2013) documented small-scale genetic differentiation for another small understory frugivore, Dermanura watsoni, in fragments embedded in a matrix dominated by agriculture. Landscape connectivity as measured by the amount of suitable habitat surrounding forest patches was most strongly correlated with genetic variation when quantied within small-scale (400þ  m) landscape buffers, likely reecting the reduced mobility of this species. Importantly, empirical levels of genetic diversity in fragments were best explained by past rather than present habitat conditions. Because anthropogenic habitat fragmentation is recent on evolutionary timescales, populations may not show immediate genetic responses to fragmentation, highlighting the importance of considering time lags in these scenarios. In a microsatellite study of three codistributed insectivorous bat species in for est fragments in peninsular Malaysia, Struebig etþ  al. (2011) observed area-related declines in genetic diversity in Kerivoula papillosa, the species that was most sensitive to fragmentation based on ecological characteristics (low vagility, low population density, tree-cavity-roosting habit). Based on the genetic-area relationship observed for K. papillosa, the authors estimated that preserving the genetic diversity of this species at levels similar to those of intact forest would require extensive areas (>10,000þ  ha), several times larger than necessary to maintain comparable levels of species richness. In view of the fact that most forest patches in heavily fragmented production landscapes across Southeast Asia are much smaller, it is evident that maintaining genetic diversity of the dozens of forest specialist species that exhibit trait combinations similar to those of K. papillosa constitutes a substantial conservation challenge (Struebig etþ  al. 2011). Roosting ecology and social organization may generally be important predictors of genetic structuring in insectivorous Old World bats. Rossiter etþ  al. (2012) found that less vagile, treeroosting species exhibit reduced gene ow, even across continuous intact rain for est, compared to more wide-ranging colonial cave-roosting species, indicating that the former should be disproportionately affected by landscape-scale habitat fragmentation. Only weak genetic population subdivision was demonstrated for Artibeus lituratus, an abundant, highly mobile, and generalist frugivore, in a study in

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87 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …fragmented Atlantic forest (McCulloch etþ  al. 2013). High levels of contemporary population connectivity in an abundant and widespread seed disperser like A. lituratus may buffer numerous plant species in Neotropical forests that rely on dispersal services of this bat species to counterbalance the negative impacts of deforestation. In summary, the available evidence suggests, both in the New and in the Old World tropics, and irrespective of fragment–matrix contrast, that some bat species may be vulnerable to genetic erosion as a result of small-scale habitat fragmentation. Further, studies indicate that susceptibility in this context is linked to individual species traits such as mobility or roosting habit. Key research needs: Increasing research on a broader range of species with different ecological and life-history traits, ideally using high-resolution genetic markers such as microsatellites or single-nucleotide polymorphisms (SNPs). Studies that quantify the extent to which frugivorous and nectarivorous bat species are capable of maintaining gene ow among plants in fragmented tropical landscapes.4.7þ Behavioral ResponsesIn addition to the direct effects on diversity and abundances, species’ responses to anthropogenic habitat modication and disturbance can manifest as behavioral changes, which may include disruptions to species’ dispersal, movement, activity patterns, and interspecic interactions (Fischer and Lindenmayer 2007). Few studies so far have addressed these issues for tropical bats (Fig.þ  3.3b). Although a number of studies have reported movement distances and space use for a variety of tropical bat species (not reviewed here), few have explicitly addressed these phenomena in anthropogenically modied landscapes. Mark–recapture and radiotracking studies in the Neotropics suggest that in areas where landscape connectivity is relatively high, bats may regularly traverse open areas between forest fragments or between fragments and continuous forest. Evidence for interhabitat movements comes from landscapes with agricultural matrices (Estrada etþ  al. 1993a; Estrada and Coates-Estrada 2002; Bianconi etþ  al. 2006; Medina etþ  al. 2007; Mendes etþ  al. 2009; Trevelin etþ  al. 2013) or from those with a more inhospitable aquatic matrix (Albrecht etþ  al. 2007; Meyer and Kalko 2008a). Recapture data from a study in a fragmented landscape in Malaysia also indicate long-distance between-habitat movements for some cave-roosting species (Struebig etþ  al. 2008). Whether a species is able to move over fragmented landscapes may be linked to the species’ foraging ecology (Albrecht etþ  al. 2007; Henry etþ  al. 2007b). Overall, these studies were fundamental in determining the general capacity of tropical bats to move across human-modied habitats. However, they provide mostly circumstantial evidence and cannot establish whether

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88 C.F.J. Meyer et al.anthropogenic disturbance elicits direct behavioral responses in bats that manifest as changes in movement distances or patterns of space use. Better insights into how habitat modication inuences movement behavior can be gained through detailed radiotracking or long-term banding studies that compare movement patterns for species with different autecologies. Such studies, although difcult and costly to implement, would ideally compare continuous forest with fragments or other disturbed habitats. Studies that have assessed behavioral changes to habitat modication in terms of effects on temporal activity patterns have followed such a rigorous approach. Disturbance-related changes in resource abundance, diversity, or predictability can be assumed to potentially alter temporal activity of species that exploit those resources (Presley etþ  al. 2009b). Presley etþ  al. (2009a) found no interspecic differences in activity patterns of eight abundant frugivorous bats in primary lowland Amazonian rain forest. However, for ve species, activity patterns differed between primary or secondary forest and agricultural habitats, whereby bats in larger agricultural areas exhibited reduced crepuscular activity compared to those in undisturbed forest. Elsewhere in Amazonia, Castro-Arellano etþ  al. (2009) detected no differences in activity levels for nectarivores and gleaning animalivores in response to RIL. Conversely, understory frugivores (Carollia spp.) decreased activity at dusk. Another study found reduced activity by some frugivores in small forest clearings created by tree removal, although the overall effects of RIL on activity patterns of frugivores were negligible (Presley etþ  al. 2009b). In all cases, the curtailment of activity in open areas at twilight or during periods of high lunar illumination was best explained by increased predation risk (SaldañaVázquez and Munguía-Rosas 2013). Habitat modication and disturbance may consequently inuence energy budgets of bats as they have less time available for foraging, with possible negative repercussions for their ability to meet daily energy requirements. Human disturbance may also affect roosting behavior and roost site selection. In fragmented rain forest in Mexico, Evelyn and Stiles (2003) found that both sexes of cavity-roosting Sturnira lilium selected large-diameter trees in mature for est stands, as did females of the foliage-roosting Artibeus intermedius, whereas males of the latter species roosted in secondary forest. These ndings under score that preferences in terms of roosting and foraging habitat are not necessar ily correlated and point to the importance of preserving mature forest patches in human-dominated landscapes for meeting the roosting requirements of tree-cavityroosting species. Key research needs: More studies, particularly in the Paleotropics, that assess the extent to which human-driven habitat change affects bat behavior in terms of roosting and for aging ecology. Research that addresses how such behavioral changes translate into tness consequences (e.g., in terms of survival, reproductive success, physiology) that may affect long-term population persistence.

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89 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …4.8þ Effects on Selected Species InteractionsIn recent years, bats have moved to the forefront of public attention, mostly as a result of accumulating evidence that they comprise important reservoir hosts for numerous zoonotic viruses (e.g., lyssaviruses, SARS, Ebola) that may pose a serious health risk to humans (Calisher etþ  al. 2006; Hayman etþ  al. 2013, Chap. 10). Recent studies have highlighted the urgency of gaining a better understanding of how habitat loss, land-use change and disturbance and an associated increase in bat–human interactions may, for instance, accelerate viral spillover (Peel etþ  al. 2013). However, few studies to date have explored to what extent these stressors inuence patterns of parasite and disease prevalence and transmission, as well as physiological stress responses in bats (Fig.þ  4.3b). Cottontail etþ  al. (2009) found that trypanosome prevalence in A. jamaicensis was signicantly higher in fragmented sites than in continuous forest, linked to a loss of bat species richness and fragmentation-related changes in vegetation cover that may favor disease transmission. The negative relationship between trypanosome prevalence and bat species richness reects the “dilution effect,” i.e., a situation in which high host species richness reduces parasite transmission if vectors feed on multiple host species that vary in their ability to contract, amplify, or transmit the pathogen (Ostfeld and Keesing 2012). In contrast, prevalence of hemoparasitic nematodes (Litomosoides spp.) showed no signicant difference among habitats, probably as a result of greater host specicity (Cottontail etþ  al. 2009). In another study, fragmentation affected the physiological condition of A. obscurus, as evidenced by elevated hematocrit levels in forest fragments ver sus continuous forest, even though similar abundances in both habitats indicated a high degree of fragmentation tolerance. The opposite pattern was documented for A. jamaicensis, suggesting that abundance may in many instances be misleading as a metric of fragmentation sensitivity (Henry etþ  al. 2007a). Pilosof etþ  al. (2012) found a signicant effect of anthropogenic disturbance on the abundance of ectoparasitic bat ies in three of four widespread Neotropical host bat species, whereby the direction of response differed among species. Species-specic roosting habits likely play a key role in mediating the effects of disturbance on parasite transmission. A study in Mexico found signicantly lower prevalence of antirabic antibodies in non-hematophagous bats in disturbed agricultural areas (22.7þ  %) compared to relatively undisturbed dry forest sites (51.9þ  %), a pattern which may arise because of more frequent interspecies encounters in the undisturbed habitat (Salas-Rojas etþ  al. 2004). The important role of animalivorous, frugivorous, and nectarivorous bats in arthropod suppression, seed dispersal, and pollination in tropical ecosystems is widely acknowledged (Kunz etþ  al. 2011). The degree to which such interactions are susceptible to habitat modication and disturbance is generally better under stood for seed dispersal than for pollination or arthropod suppression. Mostly using fecal analysis or seed traps, numerous studies in various human-modied landscapes across the Neotropics have documented the quantity and diversity of

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90 C.F.J. Meyer et al.seeds carried by bats (mostly Carollia spp., Sturnira spp., Artibeus spp.) into a diverse range of disturbed habitats including pastures, agricultural areas, coffee plantations, and secondary forests (Medellín and Gaona 1999; Galindo-González etþ  al. 2000; García etþ  al. 2000; Aguiar and Marinho-Filho 2007; Hanson etþ  al. 2007; Wieland etþ  al. 2011; Castro-Luna and Galindo-González 2012b; GarcíaEstrada etþ  al. 2012; García-Morales etþ  al. 2012; Gorchov etþ  al. 2013). Voigt etþ  al. (2012) showed that bats of the genus Carollia were likely to carry seeds from midsuccessional forest into adjacent primary forest, suggesting that directionality of seed transfer between disturbed areas and undisturbed forest may change with forest recovery. Isolated g trees in abandoned pastures are attractive for many frugivores and may function as regeneration nuclei that effectively facilitate forest recovery (Guevara etþ  al. 2004). Overall, these studies provide little evidence for major disruptions of seed dispersal mutualisms in response to habitat fragmentation and disturbance, although minor effects were detectable. For instance, small frugivorous bats disperse fewer large seeds in small, disturbed compared to large, undisturbed forest patches (Melo etþ  al. 2009), suggesting a negative impact of disturbance on the dispersal of larger-seeded trees. Although Old World fruit bats in some areas may disperse seeds of early successional species (Hamann and Curio 1999), seed input into deforested or degraded areas tends to be low in humanmodied landscapes in the Paleotropics (Duncan and Chapman 1999; Ingle 2003). Pteropodids generally play a much less signicant role as dispersers of early successional plants compared to phyllostomids, but are important dispersers of late successional canopy trees (Muscarella and Fleming 2007). How habitat modication affects seed dispersal of large-seeded canopy trees by pteropodid fruit bats in Paleotropical forests requires further detailed study. Research in fragmented Central American dry forest ecosystems found a decline in ower visitation rates, number of pollen grains deposited, and fruit set of certain bombacaceous tree species, suggesting that habitat disruption can impair the pollination services of nectarivorous phyllostomids, with negative consequences for plant reproductive success (Stoner etþ  al. 2002; Quesada etþ  al. 2003). However, effects were dependent on plant species (Quesada etþ  al. 2004), making general predictions regarding the effects of habitat modication on the disruption of bat pollination difcult. Through its inuence on bat foraging behavior, habitat disturbance may also limit pollen exchange between trees, leading to higher progeny relatedness in isolated trees relative to those in undisturbed forest (Quesada etþ  al. 2013). In a fragmented landscape in tropical Australia, common blossom bats (Syconycteris australis) were high-quality pollinators of the rain forest tree Syzygium cormiorum, as inferred based on pollen loads, visitation rates, and movement patterns (Law and Lean 1999). Nectarivorous bats often attain higher abundance in response to anthropogenic disturbance (see Sect.þ  4.5), suggesting that provisioning of pollination services may potentially be resistant and resilient to environmental perturbation.

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91 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …Key research needs: Detailed studies that address the causal links between human-induced landscape change and bat physiological and immune responses, as well as disease susceptibility. Studies, particularly in the Paleotropics, that document the full dispersal cycle—from seed deposition through germination, seedling establishment, and recruitment—and how it is affected by habitat alteration. Further studies across a range of pollinator and plant species, as well as fragmented landscapes with different degrees of connectivity, to directly relate behavior and movement of pollinators with reproductive success and gene ow of trees. Studies that address the extent to which arthropod suppression services are affected by more intensive forms of habitat alteration and disturbance such as those associated with secondary forests, tree plantations, or cropland (see Wanger etþ  al. 2014).4.9þ General Conclusions and Future Research DirectionsAs a consequence of a rapid increase in the annual number of publications over the past quarter century, ecological understanding has broadened and deepened concerning the inuence of land conversion and habitat fragmentation on tropical bats at the level of populations, ensembles, and assemblages. Nonetheless, large geographic and taxonomic biases characterize current understanding. Although many studies document that human-induced changes in land use alter bat species abundances and taxonomic dimension of biodiversity, surprisingly few studies have explored how these changes manifest with regard to genetic, behavioral, physiological, or disease-related phenomena. Similarly, little is known about the way in which land-use change affects functional or phylogenetic dimensions of biodiversity (but see Cisneros etþ  al. 2015). Studies generally are not conducted in a spatially explicit manner (Fig.þ  4.4a), so multiscale (e.g., alpha, beta, and gamma diversities) or cross-scale interactions cannot be explored fully, and conclusions must be tempered in the absence of a more integrated understanding of the role of unmodied habitat in rescuing local populations from extinction. Key insights from landscape-scale studies comprise the speciesand ensemble-specic nature of responses, as well as their dependence on spatial scale. The most fundamental developments include the recognition that habitat fragmentation is a complex process involving the nature of patches (i.e., landscape composition and conguration), as well as the nature of the matrix that arises as a consequence of direct, human modications of the landscape (Fig.þ  4.4b). Finally, the consequences of changes in the bat fauna from habitat conversion and fragmentation have not been

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92 C.F.J. Meyer et al. (a) (b)

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93 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging …quantied with regard to the maintenance of vital ecosystem processes or services. Clearly, we are still far from a comprehensive understanding of how tropical bats respond to habitat modication. To advance ecological understanding, we have highlighted a number of more specic research needs across all themes in this chapter. We further stress the following key research directions as particularly worthy of pursuit, many of which have been summarized in different context for mammals in general (e.g., Willig 2001). 1.þt Geographic and taxonomic biases toward the Neotropics and a focus on just one bat family, Phyllostomidae, need to be overcome. Although research efforts in Southeast Asia are gaining momentum (Kingston 2013), Africa deserves greatly intensied research activities. As technological advances now make acoustic sampling of aerial insectivorous bats increasingly timeand cost-þ­ efcient, this ensemble should regularly be targeted in ecological research, including environmental impact assessments. 2.þt Research should be broadened to encompass the full spectrum of possible responses at the level of populations, ensembles, assemblages, and metacommunities. Novel mechanistic insights could be gained by studies that assess behavioral responses to particular types of habitat conversion or habitat fragmentation. Similarly, studies are needed to investigate physiological and immune responses, as well as disease susceptibility across a broad range of host and vector species. A better understanding of the genetic effects on bats from habitat modication requires integrated research on a suite of different species that explore the link between patterns of genetic variation and species’ ecological and life-history traits. In general, the way in which species traits and Fig.þ  4.4þ Two conceptual models that indicate the pathways whereby land-use changes affect bats in ways that a are not spatiotemporally explicit or b are spatiotemporally explicit. In both scenarios, effects of land-use change are mediated by alterations in the vegetation, but the under lying mechanisms differ (contrast the purple boxes with the blue boxes). Nonetheless, populations and assemblages of bats respond via similar mechanisms associated with feeding, roosting, and movement opportunities (green boxes). Generally, studies that explore the effects of habitat conversion (e.g., effects of logging or agriculture) on bats are not spatially explicit. Land-use change is reected in habitat conversion that directly alters the composition and structure of the vegetation, with effects on the abundance and distribution of food resources or roosts, and the existence of “yways” whereby bats navigate through the forest. In concert, these three characteristics affect the population dynamics of different bat species and the interaction likelihoods among species (e.g., bat species, other animal species, and disease-causing microorganisms). As a consequence, changes in bat species abundance distributions (e.g., richness, evenness, dominance, diversity, rarity) emerge with cascading effects on the vegetation as a consequence of altered seed dispersal, pollination, or regulation of insect herbivores. Generally, studies of habitat fragmentation are spatially explicit and explore how land-use change affects a focal habitat type (e.g., forest) by creating a network of patches embedded in a matrix of human-modied habitats. Such studies have the potential to explore how patch characteristics (e.g., landscape composition and conguration of forest patches) as well as matrix characteristics (e.g., structural or compositional attributes of the converted land) interact to affect the bat fauna. See text for additional detailsW

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94 C.F.J. Meyer et al.environmental factors interact to shape species responses to landscape change is unclear, as trait-based approaches have been rare (but see Farneda etþ  al. 2015). Understanding how functional and phylogenetic biodiversity changes during habitat conversion and secondary succession is investigated rarely and remains poorly understood. Much also remains to be learned about how habitat disruption and modication affect the provisioning of critical ecosystem ser vices, especially ower pollination and arthropod suppression. 3.þt Multiscale studies provide a more comprehensive understanding of pattern–process relationships in heterogeneous human-modied landscapes than do single-scale assessments. Future research should address bat responses to landscape change with respect to both spatial and temporal dimensions. Considerable progress in the eld could be made by directing greater research effort and resources toward long-term studies that are capable of unveiling novel insights, which are hard or impossible to obtain from short-term, crosssectional studies (cf. Lindenmayer etþ  al. 2011). Studies currently underway at the Biological Dynamics of Forest Fragments Project (BDFFP) in Brazil (Meyer etþ  al, unpublished data) or at the Stability of Altered Forest Ecosystems (SAFE) Project in Borneo (e.g., Struebig etþ  al. 2013) provide examples of rst efforts in this direction. The need for broader geographic coverage notwithstanding, directing more research to well-studied systems or long-term study sites, allows the responses of bats to land-use change to be compared to those of other taxa (e.g., Barlow etþ  al. 2007; Bicknell etþ  al. 2015; Ewers etþ  al. 2015). 4.þt We stress the importance of robust study designs for assessing faunal responses to habitat alteration. Studies should have adequate replication (cf. Ramage etþ  al. 2013) and involve controls or reference sites. Lack of controls is an important shortcoming of many of the reviewed studies, which often focused on comparisons of different types of disturbed habitats. This clearly limits their ability to ascribe observed effects to disturbance. We echo Kingston’s (2013) call for studies to collect predisturbance, baseline information whenever possible, given that tropical bat assemblages exhibit considerable spatiotemporal variability even in unmodied habitats. In this context, Before–After–Control– Impact designs (e.g., Bicknell etþ  al. 2015), in which sites affected by human disturbance are compared with undisturbed reference sites, both before and after impact, enhance inferential strength (Smith 2013), and add scientic rigor to future assessments of the effects of habitat modication on tropical bats. Finally, an improved ecological understanding of bat responses to land-use change will be of little use to society unless it can be translated into improved management practices that ensure their long-term conservation and provision of critical ecosystem services. Across all themes in this chapter, we urge bat researchers to apply more of their science to policy and management questions. Examples of such applications include the effectiveness of specic management practices (e.g., farming intensity, cutting cycles) and mitigation measures (e.g., riparian conservation set-asides, articial roosts).

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95 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging … Acknowledgmentsþ We would like to thank the editors for the invitation to contribute to this volume and Brock Fenton and Jorge Galindo-González for helpful comments on the manuscript. Open Access This chapter is distributed under the terms of the Creative Commons Attribution Noncommercial License, which permits any noncommercial use, distribution, and reproduction in any medium, provided the original author(s) and source are credited.References*Studies considered in our systematic review Aguiar LMS, Marinho-Filho J (2007) Bat frugivory in a remnant of Southeastern Brazilian Atlantic forest. Acta Chiropt 9:251–260* Albrecht L, Meyer CFJ, Kalko EKV (2007) Differential mobility in two small phyllostomid bats, Artibeus watsoni and Micronycteris microtis, in a fragmented Neotropical landscape. Acta Theriol 52:141–149* Asner GP, Broadbent EN, Oliveira PJC etþ  al (2006) Condition and fate of logged forests in the Brazilian Amazon. Proc Natl Acad Sci 103:12947–12950 Asner GP, Keller M, Lentini M etþ  al (2013) Selective logging and its relation to deforestation. In: Keller M, Bustamante M, Gash J, Silva Dias P (eds) Amazonia and Global Change. American Geophysical Union, Washington, D.C., pp 25–42 Asner GP, Rudel TK, Aide TM etþ  al (2009) A contemporary assessment of change in humid tropical forests. Conserv Biol 23:1386–1395 Avila-Cabadilla LD, Sanchez-Azofeifa GA, Stoner KE etþ  al (2012) Local and landscape factors determining occurrence of phyllostomid bats in tropical secondary forests. PLoS ONE 7:e35228* Avila-Cabadilla LD, Stoner KE, Henry M etþ  al (2009) Composition, structure and diversity of phyllostomid bat assemblages in different successional stages of a tropical dry forest. For Ecol Manage 258:986–996* Banks-Leite C, Ewers RM, Metzger J-P (2010) Edge effects as the principal cause of area effects on birds in fragmented secondary forest. Oikos 119:918–926 Banks-Leite C, Ewers RM, Metzger J-P (2013) The confounded effects of habitat disturbance at the local, patch and landscape scale on understorey birds of the Atlantic Forest: Implications for the development of landscape-based indicators. Ecol Indic 31:82–88 Barlow J, Gardner TA, Araujo IS etþ  al (2007) Quantifying the biodiversity value of tropical primary, secondary and plantation forest. Proc Natl Acad Sci USA 104:18555–18560* Barragan F, Lorenzo C, Moron A etþ  al (2010) Bat and rodent diversity in a fragmented landscape on the Isthmus of Tehuantepec, Oaxaca, Mexico. Trop Conserv Sci 3:1–16* Bianconi GV, Mikich SB, Pedro WA (2006) Movements of bats (Mammalia, Chiroptera) in Atlantic Forest remnants in southern Brazil. Rev Bras Zool 23:1199–1206* Bicknell JE, Struebig MJ, Edwards DP etþ  al (2014) Improved timber harvest techniques maintain biodiversity in tropical forests. Curr Biol 24:R1119–R1120 Bicknell JE, Struebig MJ, Davies ZG (2015) Reconciling timber extraction with biodiversity conservation in tropical forests using Reduced-Impact Logging. J Appl Ecol 52:379–388 Bissonette JA, Storch I (2007) Temporal dimensions of landscape ecology: wildlife responses to variable resources. Springer, New York Blaser J, Sarre A, Poore D etþ  al (2011) Status of tropical forest management. ITTO technical series no. 38. International Tropical Timber Organization, Yokohama, Japan Bobrowiec PED, Gribel R (2010) Effects of different secondary vegetation types on bat community composition in Central Amazonia, Brasil. Anim Conserv 13:204–216* Broadbent EN, Asner GP, Keller M etþ  al (2008) Forest fragmentation and edge effects from deforestation and selective logging in the Brazilian Amazon. Biol Conserv 141:1745–1757

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101 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging … Perfecto I, Vandermeer J (2008) Biodiversity conservation in tropical agroecosystems. Ann NY Acad Sci 1134:173–200 Peters SL, Malcolm JR, Zimmerman BL (2006) Effects of selective logging on bat communities in the Southeastern Amazon. Conserv Biol 20:1410–1421* Phalan B, Bertzky M, Butchart SHM etþ  al (2013) Crop expansion and conservation priorities in tropical countries. PLoS ONE 8:e51759 Phommexay P, Satasook C, Bates P etþ  al (2011) The impact of rubber plantations on the diver sity and activity of understorey insectivorous bats in southern Thailand. Biodivers Conserv 20:1441–1456* Pilosof S, Dick CW, Korine C etþ  al (2012) Effects of anthropogenic disturbance and climate on patterns of bat y parasitism. PLoS ONE 7:e41487* Pina SMS, Meyer CFJ, Zortéa M (2013) A comparison of habitat use by phyllostomid bats (Chiroptera: Phyllostomidae) in natural forest fragments and Eucalyptus plantations in the Brazilian Cerrado. Chiroptera Neotrop 19:14–30* Pineda E, Moreno C, Escobar F etþ  al (2005) Frog, bat, and dung beetle diversity in the cloud for est and coffee agroecosystems of Veracruz, Mexico. Conserv Biol 19:400–410* Pinto N, Keitt TH (2008) Scale-dependent responses to forest cover displayed by frugivore bats. Oikos 117:1725–1731* Pons J-M, Cosson J-F (2002) Use of forest fragments by animalivorous bats in French Guiana. Rev Ecol-Terre Vie 57:117–130* Prasetyo PN, Noerfahmy S, dan Tata HL (2011) Jenis-jenis Kelelawar Agroforest Sumatera (Bat species in Sumatran agroforest). World Agroforestry Centre—ICRAF, SEA Regional Ofce, Bogor, Indonesia* Presley SJ, Willig MR, Wunderle Jr. JM etþ  al (2008) Effects of reduced-impact logging and forest physiognomy on bat populations of lowland Amazonian forest. J Appl Ecol 45:14–25* Presley SJ, Willig MR, Castro-Arellano I etþ  al (2009a) Effects of habitat conversion on temporal activity patterns of phyllostomid bats in lowland Amazonian rain forest. J Mammal 90:210–221* Presley SJ, Willig MR, Saldanha LN etþ  al (2009b) Reduced-impact logging has little effect on temporal activity of frugivorous bats (Chiroptera) in lowland Amazonia. Biotropica 41:369–378* Pullin AS, Stewart GB (2006) Guidelines for systematic review in conservation and environmental management. Conserv Biol 20:1647–1656 Putz FE, Sist P, Fredericksen T etþ  al (2008) Reduced-impact logging: Challenges and opportunities. For Ecol Manage 256:1427–1433 Putz FE, Zuidema PA, Synnott T etþ  al (2012) Sustaining conservation values in selectively logged tropical forests: the attained and the attainable. Conserv Lett 5:296–303 Quesada M, Stoner KE, Rosas-Guerrero V etþ  al (2003) Effects of habitat disruption on the activity of nectarivorous bats (Chiroptera: Phyllostomidae) in a dry tropical forest: implications for the reproductive success of the neotropical tree Ceiba grandiora. Oecologia 135:400–406* Quesada M, Stoner KE, Lobo JA etþ  al (2004) Effects of forest fragmentation on pollinator activity and consequences for plant reproductive success and mating patterns in bat-pollinated bombacaceous trees. Biotropica 36:131–138* Quesada M, Herrerias-Diego Y, Lobo JA etþ  al (2013) Long-term effects of habitat fragmentation on mating patterns and gene ow of a tropical dry forest tree, Ceiba aesculifolia (Malvaceae: Bombacoideae). Am J Bot 100:1095–1101* Ramage BS, Sheil D, Salim HMW etþ  al (2013) Pseudoreplication in tropical forests and the resulting effects on biodiversity conservation. Conserv Biol 27:364–372 Ribeiro MC, Metzger JP, Martensen AC etþ  al (2009) The Brazilian Atlantic Forest: how much is left, and how is the remaining forest distributed? Implications for conservation. Biol Conserv 142:1141–1153 Ries L, Sisk TD (2010) What is an edge species? The implications of sensitivity to habitat edges. Oikos 119:1636–1642

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102 C.F.J. Meyer et al. Ripperger SP, Tschapka M, Kalko EKV etþ  al (2013) Life in a mosaic landscape: anthropogenic habitat fragmentation affects genetic population structure in a frugivorous bat species. Conserv Genet 14:925–934* Rossiter SJ, Zubaid A, Mohd-Adnan A etþ  al (2012) Social organization and genetic structure: insights from codistributed bat populations. Mol Ecol 21:647–661* Sala OE, Chapin FS, Armesto JJ etþ  al (2000) Global biodiversity scenarios for the year 2100. Science 287:1770–1774 Salas-Rojas M, Sánchez-Hernández C, Romero-Almaraz MdL etþ  al (2004) Prevalence of rabies and LPM paramyxovirus antibody in non-hematophagous bats captured in the Central Pacic coast of Mexico. Trans R Soc Trop Med Hyg 98:577–584* Saldaña-Vázquez R, Sosa V, Hernández-Montero J etþ  al (2010) Abundance responses of frugivorous bats (Stenodermatinae) to coffee cultivation and selective logging practices in mountainous central Veracruz, Mexico. Biodivers Conserv 19:2111–2124* Saldaña-Vázquez RA, Munguía-Rosas MA (2013) Lunar phobia in bats and its ecological cor relates: a meta-analysis. Mamm Biol 78:216–219 Sampaio EM, Kalko EKV, Bernard E etþ  al (2003) A biodiversity assessment of bats (Chiroptera) in a tropical lowland rainforest of Central Amazonia, including methodological and conser vation considerations. Stud Neotrop Fauna Environ 38:17–31* Schulze MD, Seavy NE, Whitacre DF (2000) A comparison of the phyllostomid bat assemblages in undisturbed neotropial forest and in forest fragments of a slash-and-burn farming mosaic in Petén, Guatemala. Biotropica 32:174–184* Sedlock JL, Weyandt SE, Cororan L etþ  al (2008) Bat diversity in tropical forest and agro-pastoral habitats within a protected area in the Philippines. Acta Chiropt 10:349–358* Smith EP (2013) BACI design. Encyclopedia of environmetrics. Wiley, London Stoner KE, Quesada M, Rosas-Guerrero V etþ  al (2002) Effects of forest fragmentation on the Colima long-nosed bat (Musonycteris harrisoni) foraging in tropical dry forest of Jalisco, Mexico. Biotropica 34:462–467* Storfer A, Murphy MA, Spear SF etþ  al (2010) Landscape genetics: where are we now? Mol Ecol 19:3496–3514 Struebig MJ, Kingston T, Zubaid A etþ  al (2008) Conservation value of forest fragments to Palaeotropical bats. Biol Conserv 141:2112–2126* Struebig MJ, Kingston T, Zubaid A etþ  al (2009) Conservation importance of limestone karst outcrops for Palaeotropical bats in a fragmented landscape. Biol Conserv 142:2089–2096* Struebig MJ, Kingston T, Petit EJ etþ  al (2011) Parallel declines in species and genetic diversity in tropical forest fragments. Ecol Lett 14:582–590* Struebig MJ, Turner A, Giles E etþ  al (2013) Quantifying the biodiversity value of repeatedly logged rainforests: gradient and comparative approaches from Borneo. In: Guy W, Eoin JOG (eds) Advances in ecological research, vol 48. Academic Press, New York, pp 183–224* Struebig M, Wilting A, Gaveau DLA etþ  al (2015) Targeted conservation safeguards a biodiversity hotspot from climate and land-cover change. Curr Biol 25:372–378 Syamsi F (2013) Chiroptera community in oil palm plantation. J Indonesian Nat Hist 1:49* Tilman D, Balzer C, Hill J etþ  al (2011) Global food demand and the sustainable intensication of agriculture. Proc Natl Acad Sci 108:20260–20264 Trevelin LC, Silveira Mc, Port-Carvalho M etþ  al (2013) Use of space by frugivorous bats (Chiroptera: Phyllostomidae) in a restored Atlantic forest fragment in Brazil. Ecol Manage 291:136–143* Tscharntke T, Clough Y, Bhagwat SA etþ  al (2011) Multifunctional shade-tree management in tropical agroforestry landscapes—a review. J Appl Ecol 48:619–629 Villard M-A, Metzger JP (2014) Beyond the fragmentation debate: a conceptual model to predict when habitat conguration really matters. J Appl Ecol 51:309–318 Vleut I, Levy-Tacher SI, Galindo-González J etþ  al (2012) Tropical rain-forest matrix quality affects bat assemblage structure in secondary forest patches. J Mammal 93:1469–1479*

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103 4þ Responses of Tropical Bats to Habitat Fragmentation, Logging … Vleut I, Levy-Tacher SI, de Boer WF etþ  al (2013) Tropical secondary forest management inuences frugivorous bat composition, abundance and fruit consumption in Chiapas, Mexico. PLoS ONE 8:e77584* Voigt CC, Voigt-Heucke SL, Kretzschmar AS (2012) Isotopic evidence for seed transfer from successional areas into forests by short-tailed fruit bats (Carollia spp.; Phyllostomidae). J Trop Ecol 28:181–186* Wanger TC, Darras K, Bumrungsri S etþ  al (2014) Bat pest control contributes to food security in Thailand. Biol Conserv 171:220–223 Wieland LM, Mesquita RCG, Bobrowiec PED etþ  al (2011) Seed rain and advance regeneration in secondary succession in the Brazilian Amazon. Trop Conserv Sci 4:300–316* Williams-Guillén K, Perfecto I (2010) Effects of agricultural intensication on the assemblage of leaf-nosed bats (Phyllostomidae) in a coffee landscape in Chiapas, Mexico. Biotropica 42:605–613* Williams-Guillén K, Perfecto I (2011) Ensemble composition and activity levels of insectivorous bats in response to management intensication in coffee agroforestry systems. PLoS ONE 6:e16502* Williams-Guillén K, Perfecto I, Vandermeer J (2008) Bats limit insects in a Neotropical agrofor estry system. Science 320:70* Willig MR (2001) Exploring biodiversity in time and space: Protable directions for mammalogy in the 21st century. Mastozool Neotrop 8:107–109 Wright SJ, Muller-Landau H (2006) The future of tropical forest species. Biotropica 38:287–301 Willig MR, Patterson BD, Stevens RD (2003) Patterns of range size, richness, and body size in the Chiroptera. In: Kunz TH, Fenton MB (eds) Bat Ecology. University of Chicago Press, Chicago, pp 580–621 Willig MR, Presley SJ, Bloch CP etþ  al. (2007) Phyllostomid bats of lowland Amazonia: Effects of habitat alteration on abundance. Biotropica 39:737–746* Zubaid A (1993) A comparison of the bat fauna between a primary and fragmented secondary forest in peninsular Malaysia. Mammalia 57:201–206*

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105Chapter 5Insectivorous Bats and Silviculture: Balancing Timber Production and Bat ConservationBradley Law, Kirsty J. Park and Michael J. Lacki© The Author(s) 2016 C.C. Voigt and T. Kingston (eds.), Bats in the Anthropocene: Conservation of Bats in a Changing World, DOIþ  10.1007/978-3-319-25220-9_5Abstractþ Forests are one of the most important habitats for insectivorous bats as they offer the potential for both roosting and foraging. We reviewed silvicultural literature from North America, Australia, and Europe and found that diverse research approaches have revealed commonalities in bat responses to forest silviculture. Almost all silvicultural treatments evaluated were compatible with some use by forest bats, though different bat ensembles respond in different ways. Ensemble ecomorphology was a consistent predictor of how bats respond to vegetative clutter and its dynamic changes as forests regenerate and develop a dense structure following harvesting. Sustaining high levels of bat diversity in timber production forests requires a mix of silvicultural treatments and exclusion areas staggered across the landscape, regardless of forest type or geographic region. Use of edge habitats, exclusion areas/set-asides, and riparian corridors for roosting and foraging by bats were consistent themes in the literature reviewed, and these habitat elements need to be considered in forest planning. Densities of hollow or dead trees sufcient to support large populations of roosting bats are unknown and remain a major knowledge gap, but will likely be species contingent. New paradigm shifts in forest management away from the use of even-aged systems to multi-spatial scale retention of mature forest including trees with cavities should be benecial to bats, which are inuenced by landscape-scale management. Such B. Lawþ  ()þ  Forest Science, NSW Primary Industries, Sydney, Australia e-mail: brad.law@dpi.nsw.gov.au K.J. Parkþ  Biological & Environmental Sciences, University of Stirling, Scotland, UK M.J. Lackiþ  Department of Forestry, University of Kentucky, Kentucky, USA

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106 B. Law et al.an approach is already in use in some regions, though there is a limited guidance on what constitutes a reasonable landscape threshold for retention. The effectiveness of such an approach will require long-term monitoring and research, especially with population studies which are currently lacking.5.1þ IntroductionForests are one of the most important habitats for bats as they offer the potential for both roosting and foraging, and most species are reliant on forests for at least some part of their life cycle. Humans are also heavily reliant on the resources produced by forests, in particular timber. Consequently, forests are highly managed and modied in many areas. Understanding the effect that human manipulation of forested landscapes has on the resources required by bats is therefore of great importance to their conservation. The use of silvicultural techniques to manipulate tree stands for timber production or biodiversity conservation goals presents several challenges. Forest bats are mobile and, as with forest birds, can use a large three-dimensional space to meet their life requisites (Kroll etþ  al. 2012). Therefore, stand-level considerations alone are insufcient in sustaining habitat conditions for many forest bats as landscapelevel needs are of equal or greater concern (Duchamp etþ  al. 2007). Secondly, for est bats require roosting sites, high-quality foraging habitats, drinking sites, and features that provide connectivity among landscape elements. Providing all of these habitat requirements for an entire assemblage of bats simultaneously on a managed forested landscape is a difcult challenge, necessitating hierarchical approaches that assess spatial juxtaposition of habitat elements on the landscape and that implement silvicultural systems using multiple treatments applied both within and among stands. Silvicultural practices vary greatly around the world. For example, in the northern hemisphere, clear felling typically results in cleared areas of 40–180þ  ha sur rounded by relatively even-aged forests (Thomas 1988; Grindal and Brigham 1999; Swystun etþ  al. 2001). In parts of Europe and North America, however, patch sizes are considerably smaller and some countries have abandoned clear felling altogether, favouring a more selective logging approach. Similarly, in parts of Australia, where broad scale clear-fall techniques are not utilised, selective logging results in a multi-aged forest (Nicholson 1999). A key feature of insectivorous bats is their sophisticated sensory system, which enables them to navigate and forage in the dark. The foraging efciency of echolocating bats is constrained by variations in vegetation because the echoes returning from prey need to be distinguished from background echoes returning from vegetation. These ‘clutter’ echoes can mask the echoes of prey making foraging inefcient in situations where vegetation is dense (Schnitzler etþ  al.

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107 5þ Insectivorous Bats and Silviculture …2003). Forest bat species differ in echolocation signal design and wing morphology and this inuences their ight behaviour and their tolerance to clutter, allowing classication into three broad foraging ensembles: (1) closed-space species are slow ying and highly manoeuvrable bats that can forage close to vegetation; (2) edge-space species exploit edge habitat and other linear features; and (3) open-space foragers have lower manoeuvrability and y faster above the for est canopy or within large gaps in the forest. Changes to forest structure that inuence the degree of clutter can, therefore, alter the availability of foraging habitat for each ensemble. Our aim in this chapter was to explore how insectivorous bats respond to differ ent silvicultural approaches used in forests around the world, incorporating studies within natural, or semi-natural, forests to intensive management within plantation forestry. We focus on three broad areas: North America, Australasia (including New Zealand), and Europe and refer the reader to Meyer etþ  al. (2016) (Chap. 3) for tropical forests. While the majority of studies included in this review are published in scientic journals, we also include information from the grey literature (e.g. reports, conference proceedings, and unpublished theses) and some unpublished data where appropriate. We look to highlight both commonalities and differences in the various approaches to the issue in different regions. We suggest that ecomorphology is one of the keys to understanding how bats use their environment and we use ecomorphological traits as a framework for predicting how the three broad functional ensembles of bats respond to forest logging (Hanspach etþ  al. 2012; Luck etþ  al. 2013). Conceptual models have been proposed previously for the relationship between the abundance of bats and key ecological resources manipulated by forest management (Fig.þ  5.1; Hayes and Loeb 2007). These posit the inuence of thresholds for certain variables such as water availability, where fur ther increases do not result in increased bat abundance. We assess the extent to which these models t current data and extend them to (1) consider the time since logging as a response variable and (2) include an ecomorphological framework for the response of bats. We emphasise the importance of a long-term perspective when assessing bat responses in forests given that forests are longlived ecosystems that undergo dynamic changes after disturbance. Finally, we consider the merits of multi-spatial scale management for bats and recommend future areas of research to advance the effective management of this diverse and functionally important group. There is some specialised terminology within this chapter that may be unfamiliar to those new to silvicultural literature, so we have provided a glossary at the end of the chapter with denitions. While the term woodland is often used to describe vegetation communities comprising trees but with a more open and lower canopy cover than forests, this denition varies by country. Here, we use the term forest to encompass the various denitions of woodland.

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108 B. Law et al. 5.2 Major Forest Areas 5.2.1 North America Management of forests in North America is undergoing a renaissance, of sorts, as threats associated with habitat loss and fragmentation, climate change, increased re frequency, and introduction of forest insect pests are leading to paradigm shifts in how forests should be managed to sustain biodiversity, increase carbon seques tration, and maintain the capacity for resource extraction (Boerner et al. 2008 ; Parks and Bernier 2010 ; Moore et al. 2012 ). Historically, even-aged management was practiced across the continent with clearcuts, shelterwood cuts, seed-tree cuts, and deferment cuts all used in management of forests regardless of region or forest type. These practices have reached their zenith in south-eastern pine plantations where production forestry has led to short rotation harvests of monotypic stands Fig. 5.1 Conceptual models illustrating hypothesised relationships among the abundance of bats and ecological resources within forests (Hayes and Loeb 2007 )

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109 5þ Insectivorous Bats and Silviculture …of loblolly pine, Pinus taeda (Wear and Greis 2013). More recently, þ­ silvicultural approaches have focused on mimicking natural disturbance events or þ­ ecologically based forestry (Mitchell etþ  al. 2002; Long 2009), resulting in þ­ application of þ­ uneven-aged or multi-aged silvicultural systems (O’Hara 2002, 2009), and prescribed res (Boerner etþ  al. 2008), in both pine and hardwood forests. North America is >24þ  millionþ  km2 in total land surface and lies entirely within the northern hemisphere. The continent supports a rich diversity of plant species across eight major forest types (Young and Geise 2003, Fig.þ  5.2) with each type encompassing from 1 to 8 subtypes (SAF 2010). Latitude plays a prominent role in the distribution of forest types across the continent, with a north-to-south pattern of northern coniferous, northern hardwood, central broad-leaved, oak–pine, bottomland hardwood, and tropical forests (Young and Geise 2003). Two other forest Fig.þ  5.2þ a Standing dead ponderosa pine (Pinus ponderosa) used as a roost tree by long-legged myotis (Myotis volans) in Oregon, b forested landscape treated using clearcut logging in Idaho with natural regeneration present, c stand of dead trees in California typical of habitats used by barkand cavity-roosting bats in western coniferous forests, and d bottomland hardwood forest in Western Kentucky, with hollow roost tree of Ranesque’s big-eared bat (Corynorhinus ranesquii) in the centre. Photograph credits M. Baker (a), M. Lacki (b, c), and J. Johnson (d)

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110 B. Law et al.types, Pacic coast and Rocky Mountain, are distributed largely in a north–south direction paralleling several mountain ranges and, thus, cross a greater expanse of latitudes. The northern coniferous and boreal forest, dominated by spruce, r, and larch, covers the largest extent of North American land surface of any forest type, followed by Rocky Mountain and central broad-leaved forests. Rocky Mountain forests are dominated by pines across much of their range, with central broadleaved forests supporting oaks, hickories, maple, and beech. Bottomland hardwoods, comprising gums, bald cypress, oaks, and willows, represent the smallest land area of any major forest type in North America (1.25þ  millionþ  ha remaining; Mississippi Museum of Natural Science 2005). Globally, North America has experienced some of the greatest forest losses with a 5.1þ  % decline in forest land cover from 2000 to 2005 (Hansen etþ  al. 2010). Declines in forest cover have been greatest in the south-eastern USA, where 3.5þ  millionþ  ha have been lost from 1992 to 2001 (World Resources Institute 2014). Recent shifts in the region-wide approach to management of south-eastern bottomland hardwood forests, however, have brought about a reversal in the trend of loss of these forests (USDA Forest Service 2009; Miller etþ  al. 2011).5.2.2þ EuropeEurope consists of 50 countries and is just over 10þ  millionþ  km2 in land area. Forests cover approximately 45þ  % of the land area, most of which is found within the Russian Federation which comprises 40þ  % of the land area of Europe (FAO 2012). Europe’s native forest is very diverse with 13 broad categories encompassing 74 types (EEA 2006). Boreal forest consisting primarily of spruce or pine species dominates in northerly latitudes that comprise Scandinavia (Fig.þ  5.3). This is replaced by hemiboreal forest and nemoral coniferous and mixed broadleaved/coniferous forest in southern Sweden and much of eastern central Europe, with alpine coniferous forest along the mountain ranges. Moving west, mesophytic deciduous and beech forest dominates, but there is increasing amounts of plantation forest. In the southern parts of Europe coniferous (pines, rs, junipers, cypress, cedar), broadleaved (oak, chestnut) and evergreen broadleaved forests are the main wooded habitats. Parts of Europe have undergone extensive defor estation and cover has been fragmented and depleted for several centuries. While 26þ  % of Europe’s forest area is classied as primary, this falls to <3þ  % excluding the Russian Federation, and approximately 52þ  % of all forests in Europe are now designated primarily for production (FAO 2012). In Europe, as in North America and Australia, there is growing interest in silvicultural practices that mimic natural forest ecosystem processes with the aim of developing mixed, structurally diverse stands (Lähde etþ  al. 1999). This is a result of a move away from treating forests, particularly plantations, solely as a resource for timber, and an increased emphasis on sustainable management for multiple objectives including biodiversity conservation and recreation (Mason and Quine 1995). In practice, this has meant a

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111 5þ Insectivorous Bats and Silviculture …reduction in clear felling, although this varies greatly between countries. For example, it has been largely phased out in Switzerland and Slovenia, but is still the primary form of logging in the UK (Fries etþ  al. 1997; Mason etþ  al. 1999), but recent modications include retaining stands with longer rotations where possible (Mason and Quine 1995), reducing the removal of deadwood (Humphrey and Bailey 2012), and techniques geared to mimic natural disturbance such as prescribed burning.5.2.3þ AustraliaIt is estimated that forests covered about a third of the Australian continent at the time of European settlement in 1788, but by the mid-2000s this had been reduced Fig.þ  5.3þ a New Forest, United Kingdom: wood pasture, a historical European land management system providing shelter and forage for grazing animals as well as timber products, b doubleleadered Corsican pines (Pinus nigra ssp. laricio) are used as roost sites by Natterer’s bats (M. nattereri) in Tentsmuir forest in Scotland, UK; c wooded landscape, including olive groves, used extensively in southern Italy by Rhinolophus euryale; d typical Bechstein’s bat (Myotis bechsteinii) foraging habitat in England, UK: a mixture of oak (Quercus robur) and hazel (Corylus avellana) woodland. Photograph credits J Sjolund, G Mortimer (b), D Russo (c), F Greenaway (d)

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112 B. Law et al.to about 19þ  % cover (Montreal Process Implementation Group for Australia 2008). Five million hectares of forest are classied as old growth (22þ  %) and over 70þ  % of these occur in conservation reserves. Timber harvesting on public land is now restricted to 9.4þ  millionþ  ha, or about 25þ  % of the areas potentially suitable for timber production, and much of this has been previously logged. Eucalypts dominate the forests of Australia, and they are highly diverse comprising 500–600 species (Fig.þ  5.4, Florence 1996). Eucalypt forests range from those with a high diversity of eucalypt species to those dominated by one or a few species, the latter most often occurring in the tall wet forests of temperate southern Australia, including Tasmania (Florence 1996). These different eucalypt species and forest communities grow on different soils, under varying climates and natural disturbance regimes that in turn inuence the variety of silvicultural practices applied. Fire is Fig.þ  5.4þ Eucalypt forests of Australia: a narrow vehicle tracks through regrowth wet sclerophyll forest are used extensively by bats; b recently thinned regrowth forest potentially increases ight space and foraging opportunities for bats; c senescing crown of a Blackbutt Eucalyptus pilularus supports multiple hollow branches where bats, including maternity colonies, selectively roost; d an old-growth, spotted gum forest, Corymbia maculata, supports high densities of hollows and an open zone above a dense understorey/shrub layer, providing a variety of niches for foraging and roosting bats. Photographs B. Law

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113 5þ Insectivorous Bats and Silviculture …also a driving force behind the distribution and composition of eucalypt forests, and it occurs as massive wildres that sweep across the landscape and less intensive prescribed burns that aim to reduce fuel loads and minimise damaging wildres. To some extent, silvicultural practices aim to mimic these disturbance events and maximise regeneration after harvest. Silviculture of Australian eucalypts is thus highly variable, although the techniques applied largely resemble those used elsewhere around the globe. For example, silviculture varies from clearcut practices in the tall wet eucalypt forests of temperate southern Australia (Tasmania and Victoria) to group selection and single tree selection in warm temperate and subtropical areas to the north. Clearcuts aim to mimic broadly the massive stand replacement events created by wildres, which are an irregular feature of tall eucalypt forests in Australia. However, one important difference between clearcuts and wildres is that wildres leave legacies in the form of dead trees with hollows that can remain standing for decades. Regrowth after harvesting may take many decades to self-thin sufciently for the forest to begin to resemble the openness of mature or unlogged forest (Florence 1996). Selective logging can occur at a range of intensities that are almost a continuum from very low levels of tree removal targeting specic size/species of trees with ~10þ  % of tree basal area removed to almost a seed-tree retention silviculture with >60þ  % of stand basal area removed. In selectively harvested forests, nominal ‘rotations’ are about 60–80þ  years though these develop from repeated logging visits to the same coupes every 10–30þ  years to produce a dynamic of multi-aged mosaics of even-aged regeneration cohorts (Curtin etþ  al. 1991). Selective logging is most commonly applied to forests comprising mixed eucalypt species and uneven ages. Rainforest has a restricted occurrence in Australia, and logging of this forest type is no longer permitted.5.3þ Complexity of Bat Habitat Needs 5.3.1þ Mature, Large Diameter TreesOlder age classes of trees, especially old-growth forests, have historically been viewed as important habitats for bats (Altringham 1996; Fisher and Wilkinson 2005; Hayes and Loeb 2007) and are likely to contain a greater diversity and abundance of insect prey (e.g. Fuentes-Montemayor etþ  al. 2012; Lintott etþ  al. 2014). Early studies demonstrated variation in bat activity across stands of different age classes, with the levels of bat activity higher in older, mature stands than young stands (Thomas 1988; Erickson and West 1996; Crampton and Barclay 1998; Law and Chidel 2002). Older forests possess canopies that are more fully developed than regenerating or early-seral forests, with complex crown architecture (Wunder and Carey 1996). Old-growth forests are also likely to contain a larger number of microhabitats which are associated with higher bat species richness and higher levels of activity in common and Nathusius pipistrelles, Pipistrellus pipistrellus

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114 B. Law et al.and P. nathusii, in oak, Quercus spp., forests in southern France managed for coppice (Regnery etþ  al. 2013a). In a parallel study, time since cutting was the best predictor of the number of tree microhabitats which were 13 times more abundant in stands >90þ  years post-cutting, than those <30þ  years in age (Regnery etþ  al. 2013b). Considerable research has been undertaken on roost selection since pioneering radio-tracking studies in Australia (Lunney etþ  al. 1988; Taylor and Savva 1988). A consistent trend throughout the world is that most bats prefer to roost in larger diameter trees (>30þ  cm, Russo etþ  al. 2004; ~80þ  cm, Baker and Lacki 2006; see also Kalcounis-Rüppell etþ  al. 2005), often in older forest stands or mature forests (Lunney etþ  al. 1988; Taylor and Savva 1988; Brigham etþ  al. 1997; Betts 1998; Crampton and Barclay 1998; Sedgeley and O’Donnell 1999; Law and Anderson 2000; Lumsden etþ  al. 2002; Mazurek and Zielinski 2004; Russo etþ  al. 2004, 2010; Ruczyn´ski etþ  al. 2010). Such trees have a greater likelihood of supporting larger populations of roosting bats and persist for longer than smaller diameter dead trees (Lacki etþ  al. 2012); thus, their identication and provision in residual patches during timber harvesting is important. Where mature forest is absent across large areas at least some species nd roosts in scattered hollow trees in regrowth for est where habitat trees were not specically retained, indicating that bats typically roost in the largest available trees. One Australian study found that the 4-g eastern forest bat, Vespadelus pumilus, which ranges over relatively small areas, maintains similar sizes of maternity colonies in the scarce roosts remaining within regrowth forest compared to maternity colonies in old-growth forest (Law and Anderson 2000). Russo etþ  al. (2010) found evidence of roost selection exibility in barbastelle bats, Barbastella barbastellus; dead and dying trees, a favoured roost site for this species, were six times more common in unmanaged than managed European beech, Fagus sylvatica, forests in central Italy. Bats, however, were able to roost within managed forest, albeit in smaller numbers by exploiting roost sites in live trees and rock crevices. Few studies have investigated roost selection in younger forest where roosts are scarce, so generalisations are difcult (although see section on Plantations below).5.3.2þ Deadwood Availability and Hollow Tree DensityUntil the late twentieth century, in many parts of Europe and North America, deadwood in managed forests was removed due to concerns over forest health. While this is still common practice in some areas, the key role played by dead and decaying wood in the functioning and productivity of forest ecosystems, and its importance for biodiversity, has gained increasing recognition over the past 20þ  years (Humphrey 2005). In Australia, deadwood removal has been conned to plantations, though recognition of the importance of specically retaining old trees with hollows in managed forests originated in the 1980s. A preference for roosts in dead and dying trees has been noted for Barbastella and Nyctalus species in Europe (Russo etþ  al. 2004; Ruczyn´ski and Bogdanowicz 2008; Hillen etþ  al. 2010),

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115 5þ Insectivorous Bats and Silviculture …and high densities of dead trees appear to be strongly correlated with the presence of roosts of bark and cavity-roosting bats in forested ecosystems across North America (Mattson etþ  al. 1996; Sasse and Pekins 1996; Rabe etþ  al. 1998; Waldien etþ  al. 2000; Cryan etþ  al. 2001; Bernardos etþ  al. 2004; Broders and Forbes 2004; Miles etþ  al. 2006; Perry and Thill 2007b; Arnett and Hayes 2009). The importance of high roost density has also been reported in Australia. In dry Jarrah forest of Western Australia, both Gould’s long-eared bat, Nyctophilus gouldi, and the southern forest bat, Vespadelus regulus, preferred roosting in older forest that contained a much higher density of trees with hollows (16–32þ  treesþ  ha 1) than shelterwood creation and gap release sites (8–12þ  treesþ  ha 1) (Webala etþ  al. 2010). These mature forest hollow tree densities are comparable to average densities of live and dead hollow trees in roost areas used by Gould’s wattled bat, Chalinolobus gouldii, (17þ  ha 1) and the lesser long-eared bat, N. geoffroyi, (18þ  ha 1) in a fragmented landscape in south-eastern Australia (Lumsden etþ  al. 2002). Greater densities of hollow trees likely facilitate roost switching in bark and cavityroosting bats or ssion–fusion behaviours (Kerth and König 1999; Willis and Brigham 2004). These behaviours lead to complex patterns of use and movement among available roost trees by colonies of forest bats. The variation in numbers of roosts between core and peripheral areas of roost networks is further inuenced by the density and spatial distribution of available roost trees, as demonstrated for Ranesque’s big-eared bat, Corynorhinus ranesquii, in south-eastern bottomland hardwood forests of North America (Johnson etþ  al. 2012b). Roost networks of northern long-eared bat, Myotis septentrionalis, in actively managed forests were scale-free and connected to a single central-node roost tree (Johnson etþ  al. 2012a). A similar pattern was observed for the open-space foraging white-striped free-tail bat, Tadarida australis, in south-east Queensland (Rhodes etþ  al. 2006). Given these patterns, we postulate that implementation of silvicultural systems, which promote retention of higher densities of dead and old living trees across forested ecosystems, should benet barkand cavity-roosting bats and facilitate ‘natural patterns’ in colony behaviours, social interactions, and the use of roost networks.5.3.3þ Understory VegetationThe extent and composition of understory vegetation in forests strongly inuences insect prey availability, the ability of bats to access the forest interior, and the microclimates available and is also likely to affect risk of predation. The degree to which understory cover affects the use of forests by bats depends greatly on their wing morphology and foraging behaviour, with some bats benetting from a more open forest with little in the way of cover, while other species rely heavily on a well-developed dense understory (e.g. Hill and Greenaway 2008; Müller etþ  al. 2012). Vegetation structure revealed by LiDAR in Germany indicated that while high levels of understory cover were preferred by edge-space and gleaning

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116 B. Law et al.species, open-space foragers were more associated with relatively open for est stands (Jung etþ  al. 2012). Foraging intensity also varies with canopy height, with the activity of open-space foragers highest above the canopy (Kalcounis etþ  al. 1999; Müller etþ  al. 2013), although few studies have surveyed bats at those heights. Similarly, in forest fragments in Scotland (UK), high activity levels of edge-space species, e.g. Pipistrellus spp., are related to low tree densities and an open understory, while closed-space gleaning species, e.g. Natterer’s bat, Myotis nattereri, showed the opposite trend. These studies are supported by numerous species-specic studies. For example, roosts of Bechstein’s bat, Myotis bechsteinii, and the barbastelle bat, B. barbastellus, are strongly associated with areas of thick understory (Greenaway and Hill 2004), and core foraging areas for brown longeared bat, Plecotus auritus, a closed-space species, were associated with more cover and a well-developed understory layer more than peripheral areas (Murphy etþ  al. 2012). An Australian study of vertical stratication (excluding above the canopy) in spotted gum forest also found the understorey to support the greatest insect abundance, although bat activity was up to 11 times greater in the canopy where there was less clutter and presumably insects were more accessible (Adams etþ  al. 2009). There was no evidence that any one ensemble or ensemble species foraged exclusively at a particular height, although the open-space ensemble was most activity in the canopy.5.3.4þ Slope and AspectSlope and aspect inuence roost selection in forest bats by creating variation in the amount and extent of solar heating at roosting sites due to differences in shading effects and the length of the day that roosts are in direct sunlight. Studies have demonstrated the importance of both slope position and reproductive stage in roost selection. For example, long-legged myotis, Myotis volans, in the northwestern USA switch between riparian bottoms and upper-slope positions during pregnancy, but select roosts in upper-slope positions during lactation, where they would be exposed to greater solar radiation (Baker and Lacki 2006). Studies of bats in south-eastern forests of North America have also observed preferences for roosting in upper-slope positions by foliage-roosting eastern red bat, Lasiurus borealis, and barkand cavity-roosting bats (Myotis and Eptesicus) (Hutchinson and Lacki 2000; Lacki and Schwierjohann 2001; Perry etþ  al. 2008), suggesting that higher slopes are important for roost selection in some forest bat species in both eastern and western parts of North America and should be accounted for in forest planning. Use of lower slope positions and riparian corridors for roosting is common in several bats in eastern and south-eastern forests, however, including barkand cavity-roosting (Watrous etþ  al. 2006; Perry and Thill 2008; Fleming etþ  al. 2013) and foliage-roosting species (Perry etþ  al. 2007a; Hein etþ  al. 2008b; O’Keefe etþ  al. 2009). Roosting on lower slopes was also found in a subtropical Australian forest, where lactating eastern forest bats, V. pumilus, roost in hollow

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117 5þ Insectivorous Bats and Silviculture …trees in riparian zones during early summer, but shift up-slope during autumn when bats begin to mate (Law and Anderson 2000). Riparian zones may provide more buffered conditions for maternity roosts in warm, subtropical locations. In the northern hemisphere, selection of south-eastern-facing (Willis and Brigham 2005), south-facing (Klug etþ  al. 2012), and eastern-facing (Perry and Thill 2007a) sides of tree canopies by hoary bats, Lasiurus cinereus, is associated with positive energy savings and is hypothesised to facilitate rapid growth of young (Klug etþ  al. 2012). Eastern red bat, L. borealis, another foliage-roosting þ­ species, was observed using the south aspect of tree canopies that were also located in south-facing slope positions (Mormann and Robbins 2007). Collectively, these behaviours suggest consideration be given to creating and maintaining edge habitats for foliage-roosting bats at the landscape scale, especially along south-facing slopes in the northern hemisphere in areas with sufcient topographic relief.5.3.5þ Forest EdgeLoss and fragmentation of forest habitat are accompanied by an increase in the ratio of forest edge to interior forest, and the response of bats to this can vary among species. Roosting ecology and edge-afnity have been identied as good predictors of the sensitivity of individual bat species to habitat fragmentation; ‘forest interior’ species (often tree-roosting bats) are negatively affected by fragmentation, as opposed to species which show afnity for forest edges (Meyer etþ  al. 2008). Edge habitats can inuence roosting behaviour in barkand cavity-roosting Myotis species differently. Indiana bat, M. sodalis, and northern long-eared bat, M. septentrionalis, two species with overlapping distributions in North America and similar preferences for roosting in dead trees (Foster and Kurta 1999; Lacki etþ  al. 2009), choose roosts differently in the same forested landscapes. M. sodalis prefers roosts in edge habitats with low vegetative clutter and higher solar exposure of roost trees and M. septentrionalis selects roosts in shaded environments within intact forests (Carter and Feldhamer 2005). Russo etþ  al. (2007) found that barbastelle bat, B. barbastellus, emerged later from tree roosts in more open forests, probably as a result of increased predation risks, and suggested that it was important to ensure canopy heterogeneity to provide a range of roosting conditions. Edge effects also inuence foraging behaviour in forest bats although results from studies comparing bat activity at the edge compared to forest interior show contrasting results; all ve species spanning the open/edge-space/closed-space spectra that were assessed in forests in Canada showed higher activity at the forest edge than in the interior (Jantzen and Fenton 2013). Bat activity was also high along coupe edges 5–8þ  years after clear fell in Tasmania (Law and Law 2011), partly because bats avoided the large harvested gaps in these coupes. In contrast, of three species surveyed within forest fragments on farmland in the UK, one edge-space species showed similar levels of activity at edge verses interior while the other two species (one edge-space and one closed-space) showed higher levels of activity within the

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118 B. Law et al.forest interior (Fuentes-Montemayor etþ  al. 2013). In Australia, harvested forests are often interspersed with old logging trails and re trails, providing ‘edge habitats’ that facilitate the use of dense forest regenerating after harvest by bats that possess a range of traits (Crome and Richards 1988; Law and Chidel 2002; Webala etþ  al. 2011). Activity on trails in regrowth forest is as high as it is in mature forest. Most importantly, foraging activity is typically much higher on forest trails than within the forest remote from trails or along narrow riparian zones (Law and Chidel 2002; Lloyd etþ  al. 2006; Webala etþ  al. 2011). Use of trails as linear edges in regenerating forest has also been reported in North America (Menzel etþ  al. 2002). These obser vations highlight the importance of edge habitats to many bat species within each ensemble, in all the regions covered in this chapter.5.4þ Bat Responses to Silvicultural TreatmentsSilviculture involves a diverse range of techniques to manipulate growth conditions, extract resources, and facilitate regeneration within forests. These inuence the composition and density of tree species present, the extent and composition of the understorey vegetation and ultimately the resources available for bats. Here, we focus on the techniques for which there is at least some information on the response of bats to (1) different logging strategies, (2) thinning regimes, and (3) the use of harvest exclusion areas. We also examine the use of timber plantations by bats which, in some regions, is the focus of silvicultural activities. There is very little information on the effects of other techniques such as coppice and the use of chemical applications (e.g. herbicides to clear vegetation), and we highlight important knowledge gaps in the concluding section.5.4.1þ LoggingHistorically, the strategy for logging in forest managed for timber extraction was to remove all trees within an area (clearcuts) as this is considered the most economically protable method. In production State Forests in Australia, selective harvesting was most common before World War II, but it was subsequently recognised that this adversely affected the regeneration and growth of many of the fastest growing, commercial species, which subsequently led to increased intensity of harvests. Recent concern over the environmental (including biodiversity loss and soil erosion) and visual impacts, however, has led to increased use of more selective forms of logging including variable retention and group selection techniques, which are reviewed here. A review of published data sets on response of forest bats to silvicultural logging indicates that there are major gaps in our understanding of relationships of bats with timber harvesting practices (Tableþ  5.1). In particular, there is a notable

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119 5þ Insectivorous Bats and Silviculture … Tableþ  5.1þ Summary of bat response in activity and roost selection to silvicultural treatments referred to in this review for North America and Australasia Treatment(s) Treatment conditions Forest type Bat species Bat response Source Bat activity Even-aged treatments North America Clearcut 30þ  ha Pacic coast M. lucifugus None Lunde and Harestad 1986 Clearcut Not dened Northern hardwood L. borealis Decrease Hart etþ  al. 1993 L. cinereus Increase Myotis sp. Decrease Clearcut 2–3þ  years old Pacic coast E. fuscus Increase Erickson and West 1996 L. noctivagans Increase C. townsendii Increase Clearcut Not dened Northern coniferous Multiple Mixed Grindal 1996 Clearcut Along streams Pacic coast Myotis sp. Decrease Hayes and Adam 1996 Clearcut 5–17þ  years old Pacic coast Multiple Decrease Parker etþ  al. 1996 Clearcut and residual patches Varied patch isolation Northern coniferous Multiple Mixed Swystun etþ  al. 2001 Clearcut and residual patches 8–10þ  ha; 1.5þ  years old; 0.2–0.46þ  ha Northern hardwood M. lucifugus Increase Hogberg etþ  al. 2002 M. septentrionalis Increase L. noctivagans None Clearcut 10þ  ha Northern coniferous L. noctivagans Increase Patriquin and Barclay 2003 M. lucifugus Increase M. septentrionalis Decrease Clearcut; deferment harvest 5þ  years old; 6–10þ  m2/ha residual Northern hardwood L. cinereus Increase Owen etþ  al. 2004 L. noctivagans Increase Myotis sp. None Shelterwood harvest 10þ  ha; 30–50þ  % decline in volume Central broad-leaved L. borealis Increase Titchenell etþ  al. 2011 L. noctivagans Increase E. fuscus Increase (continued)

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120 B. Law et al. Tableþ  5.1þ (continued) Treatment(s) Treatment conditions Forest type Bat species Bat response Source Seed-tree and shelterwood harvest 7.7þ  m2/ha residual; 18þ  m2/ha residual Northern hardwood Multiple Increase Dodd etþ  al. 2012 Australasia Clearcut; post-wildre 0–250þ  years old Tall mountain ash eucalypt Total activity Increase with age Brown etþ  al. 1997 Clearcut and Variable retention 10–27þ  ha; 8þ  years old; 0.5–1þ  ha retention Tall wet eucalypt forest Multiple Mixed Law and Law 2011 Plantations Noncommercial mixed; <10 and 20–25þ  years old Eucalypts Multiple Positive, older plantations Law and Chidel 2006 Plantations Low rainfall monoculture; <11þ  years old Eucalypts Multiple Neutral Law etþ  al. 2011 Uneven-aged treatments North America Group selection cuts 0.1–0.8þ  ha; 9þ  years old Northern hardwood Multiple Increase Krusic etþ  al. 1996 Group selection cuts 60þ  % decline in volume Northern coniferous Multiple Increase Perdue and Steventon 1996 Small cutblocks 0.5–1.5þ  ha Northern coniferous Multiple Increase Grindal and Brigham 1998 Group selection cuts 0.02–0.5þ  ha gaps Southern oak–pine Multiple Increase Menzel etþ  al. 2002 Canopy gaps 16–33.5þ  m wide Northern hardwood E. fuscus Increase Ford etþ  al. 2005 L. cinereus Increase Myotis sp. Decrease Australasia Selective 18þ  % basal removal 1–6þ  years old Tropical rainforest Multiple Mixed Crome and Richards 1988 Selective 3 age classes Wet sclerophyll eucalypt Multiple Mixed de oliveira etþ  al. 1999 Alternate coupe 15þ  ha coupes, 22þ  years old Dry sclerophyll eucalypt Multiple Mixed Law and Chidel 2001 (continued)

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121 5þ Insectivorous Bats and Silviculture … Tableþ  5.1þ (continued) Treatment(s) Treatment conditions Forest type Bat species Bat response Source Group selection/ plantation/old growth 13–97þ  ha catchments; 16þ  years old; tracks versus interior Wet sclerophyll eucalypt Multiple Mixed Law and Chidel 2002 Group selection cuts 3 age classes; riparian buffers 10–50þ  m Wet and dry sclerophyll eucalypt Multiple Mixed Lloyd etþ  al. 2006 Group selection cuts Old vs young regrowth; tracks vs interior; vertical stratication Spotted gum eucalypt Multiple Mixed Adams etþ  al. 2009 Gaps and shelterwood 3 age classes; gaps <10þ  ha; tracks vs interior Dry sclerophyll eucalypt Multiple Mixed Webala etþ  al. 2011 Variable retention 10–100þ  % retention; 100þ  ha blocks Tableland eucalypt Guilds Mixed Law unpubl. data Intermediate treatments North America Thinning 10–13þ  years old Pacic coast Multiple Increase Erickson and West 1996 Thinning 10þ  ha; 55þ  % decline in density Pacic coast Multiple Increase Humes etþ  al. 1999 Thinning 25þ  % decline in density; Northern coniferous Multiple None Patriquin and Barclay 2003 Thinning 45þ  % decline in density Northern pine plantation Multiple None Tibbels and Kurta 2003 Thinning 18þ  m2/ha residual Southern oak–pine E. fuscus Increase Loeb and Waldrop 2008 L. borealis Increase P. subavus None Salvage logging Control, moderate, and heavily logged sitesþ  þ  4 replicates (12–16þ  ha); 1þ  year post-re Douglas, white and ponderous r Multiple Positive Hayes 2009 (continued)

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122 B. Law et al. Tableþ  5.1þ (continued) Treatment(s) Treatment conditions Forest type Bat species Bat response Source Australasia Thinning 4–9þ  years old Spotted gum eucalypt Multiple None Adams and Law (2011) Europe Salvage logging 4 stand types varying by logging & structureþ  þ  8 replicates (5þ  þ  ha each) Norway spruce, beech and silver r Multiple Varied between for aging guilds Mehr etþ  al. 2012 Roost selection Even-aged treatments North America Clearcut 7–18þ  ha Northern coniferous M. evotis Positive, tree stumps Vonhof and Barclay 1997 Cutblocks with residual patches Not dened Northern coniferous Myotis sp. Positive, edges Grindal 1999 Australasia Clearcut 11þ  years old Dry sclerophyll eucalypt Multiple Positive, mature forest and diameter Taylor and Savva 1988 Plantation/ regrowth versus old growth 30þ  years old Wet sclerophyll forest V. pumilus Positive, gullies and diameter Law and Anderson 2000 Plantation Exotic; mosaic age classes Pinus radiata C. tuberculatus Positive, old age classes and near water Borkin and Parsons 2011b Uneven-aged treatments North America Group selection and thinning 13.8þ  m2/ha residual Southern oak–pine M. septentrionalis Positive Perry and Thill 2007b Group selection and thinning 13.8þ  m2/ha residual Southern oak–pine 5 of 6 species Positive Perry etþ  al. 2008 Australasia Alternate coupe 10–20þ  ha; 2–3þ  years old Dry sclerophyll eucalypt N. gouldi Positive, gullies and diameter Lunney etþ  al. 1988 (continued)

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123 5þ Insectivorous Bats and Silviculture …lack of long-term, longitudinal studies that track changes in bat assemblages and their forest habitat over time. Studies on bats in even-aged systems have largely focused on responses to clearcuts with limited exploration of two-age systems such as seed tree, shelterwood, or deferment harvests (Owen etþ  al. 2004; Titchenell etþ  al. 2011). Clearcut harvests have been used with less frequency, especially on public lands, for some time now (USDA and USDI 1994), though they still per sist in cool temperate forests, such as those of Tasmania (Law 1996), and some European countries. Patterns in bat responses to clearcuts are still helpful, however, in understanding the potential effects on bats of future directions in forest management based on even-aged systems. Bat responses to uneven-aged systems, such as small cutblocks, patch cuts, or group selection harvests, have received greater attention and have been evaluated across multiple bat species and forest types, so inferences can be drawn on the efcacy of these silvicultural systems for bats. In North America, more studies have evaluated bat response to thinning than any other silvicultural treatment, with thinning often applied in combination with other treatments on the same landscape (Erickson and West 1996; Patriquin and Barclay 2003; Loeb and Waldrop 2008; Perry etþ  al. 2008). Studies of treatment combinations are important as future directions in the management of forests in North America are emphasising multi-treatment prescriptions (Aubry etþ  al. 2009; Harrod etþ  al. 2009; Hessburg etþ  al. 2010), to increase structural habitat complexity, both vertically and horizontally, while reducing the impact of insect infestations and the threats of wildre and global climate change (Boerner etþ  al. 2008; Parks and Bernier 2010; Duerr and Mistretta 2013). Some forest management strategies specically target bats, though often bats are catered for under broad forest prescriptions that aim to accommodate the needs of a range of forest-dependent species in an area (Law 2004). There is a surprising lack of European studies on the effects of any logging strategy on bats and the only study found for this review which directly related Tableþ  5.1þ (continued) Treatment(s) Treatment conditions Forest type Bat species Bat response Source Gaps and shelterwood gaps <10þ  ha; buffers; 20–30þ  years old Dry sclerophyll eucalypt V. regulus Positive, mature forest and diameter Webala etþ  al. 2011 N. gouldi Positive, retained trees & diameter Intermediate treatments North America Thinning 150– 309þ  trees/ ha southern pine plantation L. borealis Positive Elmore etþ  al. 2004 Thinning 13.8þ  m2/ha residual Southern oak–pine L. borealis Positive Perry etþ  al. 2007a L. cinereus Positive

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124 B. Law et al.to this issue was one on the effects of salvage logging (see Sect.þ  4.1.4). This is especially concerning given the strict protection afforded to all bat species and particularly bat roosts under the EU Habitat Directive; this prohibits deliberate disturbance of all bats during any stage of their life cycle as well as the destruction of breeding sites or resting areas. As such, the timing of forest harvesting needs to consider whether bats may be roosting in targeted areas (e.g. Forestry Commission 2005). There are no such restrictions in Australia; though for New Zealand pine plantations, Borkin etþ  al. (2011) recommends that harvests should be planned when bats are not heavily pregnant nor have non-volant dependents. In eastern North America, logging is currently restricted from 15 October through 31 March across the distribution of the endangered Indiana bat, Myotis sodalis, as this bat uses live and dead trees as maternity sites during the growing season (USFWS 2009). Restrictions are further constrained to a start date of 15 November within 16þ  km of known hibernacula of the species (USFWS 2009). The implications of white-nose syndrome and the extensive mortalities of cave-hibernating bats in North America (USFWS 2012) are likely to add species of forest bats to the threatened and endangered species list in the USA, leading to further restrictions on logging. Missing in all of the dialogue, however, is any direct link of impact, or mortality of bats, during logging operations and studies of these potential impacts are needed (but see Borkin etþ  al. 2011). 5.4.1.1þ Clearcut and Deferment Harvests Response of forest bats to clearcut harvests has been mixed across forest types and species of bats (Tableþ  5.1). For example, three studies each in different locations within the Pacic coast forest type found no response to clearcuts by little brown bats, Myotis lucifugus, in British Columbia (Lunde and Harestad 1986), a decrease in overall bat activity over clearcuts in south-eastern Alaskan rainforests (Parker etþ  al. 1996), and an increase in activity of big brown bats, Eptesicus fuscus, silver-haired bats, Lasionycteris noctivagans, and Townsend’s big-eared bats, Corynorhinus townsendii, in clearcuts in western Washington (Erickson and West 1996). Patterns in bat activity recorded in and around clearcut harvests are inuenced by three factors: the number of years post-harvest when data were collected, the size and shape of cutblocks studied, and the assemblage of bat species present in the area. When reported, the age of clearcut stands in North America evaluated post-harvest ranged from 1.5 to 17þ  years. This range in age is wide and likely spans considerable variation in above-ground habitat structure due to differences in the amounts of regeneration present; thus, a varied response by bats across studies and geographic locations should be expected. In montane eucalypt forests of south-eastern Australia, bat activity peaked in 165-year-old wildre regrowth rather than in younger regrowth from clear-felling operations (Brown etþ  al. 1997). Unfortunately, the size and shape of clearcuts studied are rarely reported so an evaluation of the effects of cut size and shape on bat activity cannot be made.

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125 5þ Insectivorous Bats and Silviculture …Focusing on traits is likely to provide more insights into the response of bats to the large gaps created by clearcut harvests. In North America, two trends are evident. First, the creation of less obstructed ight space over clearcut stands generally leads to increased levels of activity of edge/open-space bats that possess moderate to high aspect ratios and often higher wing loadings (Lacki etþ  al. 2007). This mix of bats includes the foliage-roosting Lasiurus species, along with others (Lasionycteris and Eptesicus) (Tableþ  5.1). The length of years post-harvest at which this increase in bat activity is sustained is less clear and likely is affected by tree species composition and the speed at which regeneration proceeds in harvested stands at a particular geographic location. Second, the response to clearcut harvests between Myotis species varies both within and among species (Patriquin and Barclay 2003), with some increase in activity associated with linear edge habitats at the periphery of cuts but reduced activity in the centre of harvested stands, except where residual patches are left behind (Hogberg etþ  al. 2002). As our ability to distinguish among Myotis species increases with technological advances in acoustic detectors and software packages (Britzke etþ  al. 2011), resolution among the full suite of Myotis bats in North America should become possible allowing for a more in-depth and complete evaluation of bat response to edge effects in actively managed forests. Data on bat responses to even-aged systems other than clearcuts are sorely lacking. A study of bat activity in deferment harvests found high levels of activity of silver-haired bats, L. noctivagans, in stands with 6–10þ  m2/ha of basal area remaining (Owen etþ  al. 2004), and the only study examining bat activity in shelterwood harvests (30 to 50þ  % reductions in basal area) observed higher levels of activity in three species of bats that have wing morphologies and echolocation call structures possessed by edge/open-space bats (Titchenell etþ  al. 2011). Patterns of habitat use by radio-tagged northern long-eared bats, M. septentrionalis, a closedspace bat, showed this species spent limited time in deferment harvest stands, especially harvested sites with more open canopies and less cluttered foraging space (Owen etþ  al. 2003). For roosting bats, gap release and shelterwood systems retain tall and large diameter hollow-bearing trees within stands possessing less clutter than sur rounding forest regenerating after harvest and these offer potential roosts for bats. However, in Western Australia, southern forest bat, V. regulus, avoided locating roosts in shelterwood treatments when older forest was available nearby (Webala etþ  al. 2010). In general, remnant trees in these silvicultural treatments, including retained ‘habitat trees’, were not preferred as roost sites by V. regulus, though a second species (N. gouldi) frequently used such trees. One possible reason for avoiding using ‘habitat trees’ as roosts was the relatively low density of hollow roosts (see 3.2 Deadwood availability and hollow tree density). 5.4.1.2þ Variable Retention Harvests Variable retention has recently been proposed as an alternative to standard clearcuts, whereby old-growth elements are retained within the clearcut coupe

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126 B. Law et al.(Baker and Read 2011). Variable retention increases the availability of edges, for example, around retained patches (aggregates) of undisturbed forest within the clear-fell coupe and along coupe boundaries as well as increasing the area of open space. Openand edge-space ensembles would be expected to benet from this treatment. The 200-ha Silvicultural Systems Trial, in Tasmania, provides one of the main experimental sites in Australia for investigating responses to variable retentions. Bat activity was similar in control coupes of 45to 60-m-tall old-growth Messmate Stringybark, Eucalyptus obliqua, forest, compared to variable retention coupes 5–8þ  years after logging (Law and Law 2011). Activity was lower above the dense young regeneration of clear-fell-burn-sow (no retention) coupes and marginally lower for dispersed tree retention coupes. This suggests that the retention of old-growth elements as aggregates or patches moderates the unsuitable young regrowth zone for total bat activity, while retention of dispersed individual trees is less effective. Surprisingly, bat activity was low at the retained aggregates themselves, both in their centre and along the edge, and it is not known to what extent bats roost in these locations. Overall the results are consistent with conceptual models (Fig.þ  5.1), whereby activity is predicted to be higher in areas of medium clutter levels and where hollow abundance is high. Individual bat taxa responded to treatments consistent with predictions from ecomorphology. Closed-space bats were less active in clearcuts than unharvested forest, large edge-space bats were more active in clearcuts (especially along edges), and smaller edge-space bats were less inuenced by patch type and location within coupes; consistent with other studies of forest clearcuts from North America (Grindal and Brigham 1999; Menzel etþ  al. 2002; Patriquin and Barclay 2003). The age of regenerating forest is likely to be an important inuence on how bats respond to variable retention. An unreplicated, operational scale (100-ha for est blocks) experiment established in 1984 in the temperate forests of southern New South Wales (Waratah Creek) (Kavanagh and Webb 1998) was sampled acoustically for bats after 18þ  years of regrowth. Treatments retained different amounts of tree canopy within four different forest blocks comprising 100þ  % (control), 50þ  % (0.5þ  ha patches in a chessboard pattern), 25, or 10þ  % tree canopy retention. Control sites supported 2–4 times more activity than logged sites, with 10þ  % retention supporting the lowest activity level with just 50 bat passes per night of sampling (Fig.þ  5.5; B. Law, unpubl. data). Thus, bat activity remained low even 18þ  years after logging and the amount of canopy retained within a block had little impact on activity, except for the block with the most intensive logging which supported the lowest activity level. As expected, the activity of closed-space bats was similar, though low, between the control and treatments, after 18þ  years. Activity of edge-space bats was three times lower in logged stands, suggesting a loss of edges and spaces between trees, especially in the treatment where logging was most intense. Logging treatments had little effect on openspace bats that forage above the canopy, except that activity was lower where logging intensity was greatest.

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127 5þ Insectivorous Bats and Silviculture …5.4.1.3þ Group Selection Harvests Changes in ensemble activity with group selection harvest are likely to depend on gap size, with an increase in edge-space activity if gaps are small and openspace activity if gaps are large. Immediately after harvest, closed-space bats are expected to decline, but we predict subsequent recovery if the retention of roost trees is catered for. All studies examining bat responses in North America to group selection harvests, canopy gaps, or small cutblocks consistently reported increases in activity of bats, primarily open/edge-space species, with the opening up of forest canopies, regardless of forest type or assemblage of bats present (Tableþ  5.1). The one exception was a decline in activity of Myotis bats in canopy gaps in forests of the central Appalachian Mountains, with this drop off in use inversely correlated with increasing diameter of canopy gaps (Ford etþ  al. 2005). In this study, the maximum gap diameter examined was 33.5þ  m in width, with the decline in activity with increasing gap size largely attributable to response of closed-space Myotis species. Studies in oak–pine forests in Arkansas have demonstrated the use of dead and live trees along gap edges for roosting by several bat species (Perry and Thill 2007b; Perry etþ  al. 2008), demonstrating the importance of maintaining canopy gaps in managed forests. The almost univer sal response by bats of increased activity with canopy gap formation means this silvicultural treatment holds much promise for management of foraging habitat Fig.þ  5.5þ Total bat activity (762 passes, 10 taxa) recorded 18þ  years after logging in an unreplicated, variable intensity logging experiment in New South Wales, Australia. Data are mean number of passes per night for two Anabat detectors deployed per forest block (~100þ  ha) over two entire nights of recording and exclude activity on trails (B. Law, unpubl. data). Different bat ensembles are open-space, edge-space, closed-space, and unknown

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128 B. Law et al.of bats in the short-term. Use of gaps by forest bats following a decade or more of successional change is likely to be different, however, with overall declines in activity plausible as open/edge-space species disappear or decline in abundance with increasing gap clutter. Such temporal changes need to be identied along with the optimal gap size(s) and the density of gaps required by different species of bats to permit commercially viable, sustained yield harvests while fostering high levels of bat activity and provision of roosting habitat in managed forests. In contrast to many North American studies that have been undertaken in gaps soon after harvesting, in Australia, most bat research has focused on the use of older regrowth regenerating from group selection harvest, particularly character ising bat species by their traits in relation to the use of these dense stands. There is a general pattern of forest clutter increasing over time after group selection harvest so that old regrowth (>30þ  years) has signicant higher clutter levels than young or older forest, which constrains use by bats to closed-space species with a low wing aspect ratio (Law and Chidel 2002; Webala etþ  al. 2011). Less manoeuvrable edge-space species with a high wing aspect ratio tend to be scarce in regrowth forest (except on yways provided by tracks and creeks), although their activity is greater in the subcanopy and canopy than understorey (Adams etþ  al. 2009). Vegetation is more cluttered in regrowth at these upper heights (closer stems and less vertical space in the subcanopy), and this leads to less bat activity in such situations (Adams etþ  al. 2009). It is not known whether open-space and low-frequency edge-space species are active above the canopy of these young forests, although this was conrmed by Müller etþ  al. (2013) for mature forests in Europe. 5.4.1.4þ Salvage Logging Salvage logging involves the removal of dead wood after a natural disturbance (e.g. windthrow, forest res, and insect outbreaks) and has been employed even in protected forests, provoking some controversy. To our knowledge, no research has examined the implications for roost availability of this practice, although removal of standing dead wood will inevitably reduce the abundance and diversity of roosts and would have a considerable impact when carried-out over large scales (Lindenmayer and Noss 2006). We found two studies which investigated changes in bat activity following salvage operations. In Germany, closed-space species reduced their activity in both types of forest clearing (bark beetle and logging), while the activity of open-space species slightly increased, and edge-adapted species showed a mixed response (Mehr etþ  al. 2012). These results are similar to a study in Oregon where the highest bat activity was in the more intensely logged sites (Hayes 2009).

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129 5þ Insectivorous Bats and Silviculture …5.4.2þ Recovery Times After Timber HarvestLong-term studies are largely missing from assessments of the response of bats to silvicultural methods. A typical approach uses chrono-sequences or snapshots of comparisons between different silvicultural methods or logging histories and makes the assumption that the matching of treatments is equal and evenly distributed across the same environmental niche and landscape context. Most impor tantly, a one-year snapshot may not be representative of temporal variation and dynamism over a longer period (Recher etþ  al. 1983; Maron etþ  al. 2005); thus, conservation plans developed from snap-shots can have limitations. Long-term studies are ideal for tracking changes to vegetation structure as forests regenerate after harvesting and how different ensembles of bats respond to these dynamics. One longitudinal study in Australian eucalypt forests, initiated in 1998, has been investigating alternate-coupe-integrated harvesting for woodchips and sawlogs, and although currently unpublished, a summary is presented here (B. Law and M. Chidel, unpubl. data). Alternate-coupe harvesting divides management units (e.g. 200-ha areas) into small (~15þ  ha) coupes that are alternately harvested in a chessboard fashion, every 20þ  years. In 1998, bat activity was recorded after 22þ  years of regrowth from the rst cycle of logged coupes (Law and Chidel 2001). Bat activity in the cluttered regrowth was about half that of adjacent, more open unlogged coupes. This effect was most notable for less manoeuvrable, openand edge-space vespertilionids that were more active in unlogged coupes. The site was then sampled at intervals over 13þ  years following the second round of alternate-coupe logging (B. Law and M. Chidel, unpubl. data; Fig.þ  5.6). During this period, total bat activity remained low in old regrowth coupes (22þ  years old in 1998). Activity in unlogged controls remained similar to the initial samples taken prior to second round harvesting. Within the recently logged coupes, activity peaked soon after logging in the large gaps, but it quickly declined and remained at low levels (similar to that found in old regrowth coupes) once young regenerating eucalypts established within eight years of logging. In terms of clutter and total bat activity, these results are only partly consistent with the conceptual models of Hayes and Loeb (2007). The model predicts low bat activity when clutter is very low, yet this was not the case in this study, possibly because gaps were patchy within the 15-ha coupes due to the requirement for retention of 5 habitat trees per ha plus equivalent numbers of recruits, indicating that gap size or scale is likely to be an important issue inuencing activity. High activity at inter mediate clutter levels (unlogged coupes) and low activity at high clutter levels (old regrowth coupes) are consistent with the model. The response of individual species and ensembles are yet to be analysed for this study. The lack of recovery after 36þ  years in old regrowth coupes is consistent with a number of other studies where low activity persisted for more than 30þ  years after disturbance (Brown etþ  al. 1997; Adams etþ  al. 2009; Webala etþ  al. 2011), but differs from selective harvesting of wet sclerophyll forest in subtropical Queensland where recovery of bat activity was apparent in a site logged 33þ  years previously

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130 B. Law et al.(de Oliveira etþ  al. 1999). It is important to note that none of these studies þ­ consider activity levels on tracks, riparian zones, or other areas of retention that potentially could ameliorate the effects of clutter from dense regrowth and loss of tree hollows.5.4.3þ Thinning Young ForestsThe goal of thinning is to improve the quality and growth of the remaining trees (especially diameter) by reducing the density of trees in a stand. Reducing tree density will decrease canopy cover, at least initially, with increased light levels reaching the forest oor and thus inuencing understory cover. Adams and Law (2011) reviewed the literature on thinning and bats and proposed hypotheses for testing that included: (1) activity of edgeand open-space species will increase from pretreatment levels where thinning reduces stem separation to 7þ  m (~200þ  stemsþ  perþ  ha) but will remain at low levels where average stem separation is Ol d regrowth Recent ly logged Unlogged 1998 2002 2007 2012 Y ear 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 log_activity Fig.þ  5.6þ Changes in total bat activity over 14þ  years in an alternate-coupe logging system in southern Australia (B. Law and M. Chidel, unpubl. data). The dashed vertical line indicates second round logging of the alternate unlogged coupes in 1999, which took place 23þ  years after the rst round of logging of adjacent coupes in 1976. All but two unlogged coupes were harvested in 1999 and are thereafter referred to as recently logged coupes. Bat activity is a log transformation of the number of passes per night (95þ  % condence limits) after adjusting with mean nightly temperature as a covariate

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131 5þ Insectivorous Bats and Silviculture …less than 3þ  m (~1100 stems per ha); (2) highly cluttered forests will have low bat activity away from yways, regardless of the number of potential roosting sites and the abundance of insects, while bat activity in open forests will be highest where roost availability and insect abundance are high. Consistent with the hypotheses, bat responses to silvicultural thinning have been examined across several forest types in North America with increases in bat activity associated with thinning in Pacic coast (Erickson and West 1996; Humes etþ  al. 1999) and southern oak–pine (Loeb and Waldrop 2008) forests, but not in northern red pine, Pinus resinosa, plantations (Tibbels and Kurta 2003) or northern coniferous forests (Patriquin and Barclay 2003). An explanation for these differences is not readily clear, as the extent of thinning is not always reported in metrics that can be compared among study sites, and the suite of bat species present varies among locations. Further, data for bat activity within the Myotis genus could not be resolved to the species level with technologies used, preventing an evaluation of responses by ensemble. Patterns in roost selection of Lasiurus species in southern oak–pine forests indicate that thinned stands are frequently selected by these bats for roosting. Thus, as with clearcut harvests and larger-sized canopy gaps, stands thinned to basal areas <14þ  m2/ha appear to be well suited to less manoeuvrable edge-space Lasiurus species by providing suitable roosting and foraging habitats (Perry and Thill 2007a; Perry etþ  al. 2007a, 2008). The response of bats to forest thinning has received little attention in Australia. A preliminary study found high variability in activity for all bats and ensembles between thinned and unthinned eucalypt stands and among vegetation layers within the forest (Adams and Law 2011). Unexpectedly, thinned regrowth had a higher percentage cover for the shrub layer, and the vertical gap between canopy and understory trees was halved, which represented an increase in clutter in the zone where bats frequently y and this could have undermined any benet of wider stem spacings. However, the variability in bat activity within the thinned/ control treatments was too high to unequivocally state that thinning had no effect. While thinning is a commonly employed silvicultural technique across Europe, there has been no study of its effects on bat activity, occurrence, or species richness. There are, however, a few studies which have looked at effects of tree density on bats, thereby providing indirect evidence on likely effects of thinning. For example, in one study, where tree density varied between 180 and 2500 stems per ha in mixed deciduous/coniferous fragments within agricultural landscapes in Scotland (UK), activity of the soprano pipistrelle, Pipistrellus pygmaeus (an edgespace forager), decreased with increased tree density. In contrast, the abundance and activity of Myotis spp., and the abundance of Diptera, both increased with tree density (Fuentes-Montemayor etþ  al. 2013). This mirrors ndings by Müller etþ  al. (2012) where the activity of closed-space foragers and prey abundance increased at higher vegetation densities, while the activity of open-space foragers, and to a lesser extent, edge-space foragers declined.

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132 B. Law et al.5.4.4þ Harvest Exclusion AreasGiven the low levels of bat activity observed in young regenerating forest after logging, mitigations are needed to ameliorate the effect of high clutter levels and lower numbers of tree hollows. Edge habitat, such as tracks and clearcut boundaries, is extensively used by a range of bat species (Sect.þ  3.5). In Australia, har vest exclusion areas that support naturally open, undisturbed forest constitute a much greater proportion of the forest landscape compared to forest tracks and are therefore expected to be more important at ameliorating logging impacts on bats given that they also provide roosts in the hollows of old trees. Provided attention is paid to the size and location of harvest exclusion areas these can play a vital role in landscape connectivity, acting as corridors across forested landscapes, per mitting bats to reach otherwise isolated blocks of preferred habitat within landscapes where fragmentation has altered the matrix and created an abundance of suboptimal habitat blocks. As the extent of habitat fragmentation increases, so does the importance of corridors on the landscape (Duchamp etþ  al. 2007). Indiana bats, M. sodalis, preferred to y along wooded corridors and avoided open elds in Michigan, even though commuting distances increased by more than 50þ  % (Murray and Kurta 2004), with similar results for Pipistrellus spp. in the UK (Downs and Racey 2006). Activity of bats in heavily fragmented, pine plantations in South Carolina demonstrated more use by bats of edges along corridors than habitats within the corridor interior or nearby stands of timber (Hein etþ  al. 2009a), with bat activity directly correlated with the height of the corridor overstorey. Riparian corridors in timber production forests are often excluded from har vesting in order to ameliorate impacts of harvesting on water quality as well as providing unharvested productive habitat for biodiversity. Riparian corridors are important areas of bat foraging activity (Hayes and Adam 1996; Zimmerman and Glanz 2000; Brigham 2007), with male and female bats segregating themselves along corridor reaches in upland landscapes, with males more abundant at higher elevations (Grindal etþ  al. 1999; Senior etþ  al. 2005). Activity of bats along ripar ian corridors appears to be scale-dependent, with vegetation architecture, i.e. shrub and tree cover, inuencing the use of foraging space by bats at the local, or nest spatial, scale more than landscape habitat measures or abundance of insect prey (Ober and Hayes 2008). Abundance of Lepidoptera was high in riparian corridors in Arkansas prompting the authors to hypothesise that Ozark big-eared bat, Corynorhinus townsendii ingens, a moth strategist (Dodd and Lacki 2007), feeds extensively in and around riparian corridors in the Ozark Mountains (Dodd etþ  al. 2008). Use of best management practices along streamside management zones for sustaining healthy, riparian ecosystems is a well-established forest management practice in many regions of North America (Stringer and Perkins 2001; Lee etþ  al. 2004). Regardless, data on how these practices inuence habitat use by forest bats in riparian areas remain limited, with experimental studies sorely needed on the effects of habitat quality within corridors (stand age and composition) and corridor dimensions (size and width) on roosting and foraging ecology of bats. One study

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133 5þ Insectivorous Bats and Silviculture …in Australia demonstrated that bat activity, foraging rates, and species richness in riparian corridors within selectively harvested eucalypt forest was maintained at levels similar to riparian areas in mature forest (Lloyd etþ  al. 2006). Higher activity was recorded on larger rather than smaller order streams, a pattern also not affected by harvesting history. Such results highlight the benets of buffers, with riparian areas effectively providing habitat for foraging and commuting bats in selectively logged forests where clutter levels are likely to be high. Mitigating the loss of roosting habitat in hollow-bearing trees is arguably even more important than maintaining suitable foraging habitat. Forested corridors are critical habitat elements for North American foliage-roosting bats by providing both roosting and foraging opportunities. Male Seminole bats, Lasiurus seminolus, in south-eastern loblolly pine, P. taeda, plantations chose roost trees in forested corridors within harvest exclusion zones over 60þ  % of the time, even though corridors represented only 11þ  % of the landscape area (Hein etþ  al. 2008a). Corridors were 100 to 200þ  m in width and comprised largely of older-aged forests in riparian and upland slope positions. Use of forested corridors for roosting has been observed in other foliage-roosting species in south-eastern forests, with tricoloured bats, Perimyotis subavus, selecting riparian corridors (O’Keefe etþ  al. 2009), male evening bats, Nycticeius humeralis, choosing upland corridors of mature forest (Hein etþ  al. 2009b), and eastern red bats, L. borealis, roosting in the vicinity of gated roads (O’Keefe etþ  al. 2009). Greenbelts in riparian corridors, or unharvested inclusions of mature mixed-pine hardwoods 50þ  years in age, were important roosting habitats for these same species in southern oak–pine forests of Arkansas (Perry etþ  al. 2007b; Perry and Thill 2008). Harvest exclusion areas, especially those surrounding streams, are commonly used as roosting habitat by many tree hollow roosting Australian bats such as Gould’s long-eared bat, N. gouldi, eastern forest bat, V. pumilus, and southern forest bat, V. regulus (Lunney etþ  al. 1988; Law and Anderson 2000; Webala etþ  al. 2010). A range of factors will inuence the pattern of roosting close to creek-lines, but a large pool of older and mature trees in a variety of decay classes is likely to be important. Riparian areas often support a different vegetation type, with rainforest being particularly common in Australia. The specialist golden-tipped bat, Kerivoula papuensis, preferentially roosts in the suspended nests of small birds within riparian rainforest and such areas are excluded from harvesting (Schulz 2000; Law and Chidel 2004). Jarrah forest in Western Australia offers one example of providing pools of mature trees using zoning. Since 2004, Fauna Habitat Zones (i.e. areas of mature forest >200þ  ha set 2–4þ  km apart within areas available for logging) have been retained for species, including bats, that rely on blocks of forest supporting mature forest attributes or characteristics (Webala etþ  al. 2010). In some forest blocks, approximately 54þ  % of the total area (11,740þ  ha) is currently reserved from logging as conservation reserves, informal reserves (riparian buffers, diverse ecotype zones, road reserves), old-growth forest, and fauna habitat zones. Of these, about 39þ  % are permanently reserved, including riparian buffers, from logging in the

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134 B. Law et al.future. Testing the effectiveness of this level of retention remains a priority for forest bat research. Collectively, these ndings indicate that forested corridors are important habitat elements for roosting bats in forests across the globe.5.4.5þ PlantationsThere is no internationally agreed denition of forest plantation and many very old forests we may think of as natural have been planted. However, for the purposes of this review, the term plantation is used to mean forests planted primarily for timber extraction using intensive management techniques. Timber plantations are per haps the most extreme form of silviculture as they require replanting of typically exotic trees, with site and soil preparation required over large scales. Seedlings are planted at high densities to maximise growth and form of trees, and this has the consequence of producing high levels of clutter as the trees grow. All the silvicultural practices outlined in this section are also applicable to plantation forests. The response of bats has been documented in eucalypt plantations in Australia and pine plantations in New Zealand. As expected, bat activity in young plantations of eucalypts (<10þ  years) is typically low and considerably less than that found in nearby forest, and, somewhat surprisingly, activity is similar to levels over adjacent cleared farms (Law and Chidel 2006; Law etþ  al. 2011). Bat activity is higher in older eucalypt plantations (~25þ  years), especially where drought and lack of maintenance leads to tree mortality and the creation of gaps (Law and Chidel 2006). Closed-space species (Nyctophilus) show some association with plantations as do open-space species (Mormopterus ridei), which presumably use the space above plantations together with adjacent open paddocks. Radio-tracked bats avoid roosting in young eucalypt plantations where tree hollows are absent, even though decorticating bark is present (Law etþ  al. 2011). Despite limitations in habitat quality, plantation forests provide large areas of additional habitat for threatened long-tailed bats, Chalinolobus tuberculatus, in New Zealand (Borkin and Parsons 2011a). Borkin and Parsons (2011b) found these bats roosting in crevices, ssures, and small hollows in the oldest stands of Monterey pine, Pinus radiata, plantations (25–30þ  years), with females choosing to roost within 150þ  m of waterways. In these plantations, bats selected home ranges with higher proportions of relatively old stands than available (Borkin and Parsons 2011a). Males selected edges with open unplanted areas within their home ranges, which females avoided, instead selecting older stands for foraging. Borkin etþ  al. (2011) also documented the response to the clear-fell harvest of a pine plantation and found a pattern of declining numbers of roosts used, as well as smaller roosting areas and colony sizes. Over 3þ  years, 21þ  % of known roosts were lost with 15þ  % due to forestry operations and 6þ  % due to natural tree fall. To mitigate har vest operations, it was suggested that some suitable foraging and roosting areas should be retained within bat home ranges. Borkin etþ  al. (2011) further suggested that priority management for this declining New Zealand bat should focus on

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135 5þ Insectivorous Bats and Silviculture …plantation areas closest to water and harvests should be planned when bats are not heavily pregnant nor have non-volant dependents. Pine plantations in the south-eastern USA are actively managed landscapes with extensive amounts of fragmentation and edge development. Nevertheless, these landscapes often support a diverse bat assemblage, in part due to enhanced foraging conditions along edge interfaces and to suitable foraging and roosting habitats along forested-riparian corridors (Miller 2003; Elmore etþ  al. 2004; Hein etþ  al. 2008b, 2009a). Experimental studies have demonstrated that activity of bats is affected by edge habitats, with highest levels of activity occurring along the edge interface regardless of echolocation call structure or wing morphology (Jantzen and Fenton 2013). Tree canopies also serve as edge interfaces in forested environments, with more manoeuvrable, high-frequency bats foraging along canopies and edges more often than less manoeuvrable, low-frequency bats (Pettit and Wilkins 2012). Relationships of age, formation, and structural characteristics of edge habitats with activity of foraging bats are complex, with newly formed, high-contrast edges supporting higher bat activity and stronger depth of edge inuence, than older more developed, cantilevered edges which possess less contrast between adjacent habitats (Jantzen and Fenton 2013). Regardless, data indicate that managed forests with an abundance of edge habitat, typical of plantation forests in south-eastern North America, can support a diverse assemblage of forest bat species. Spruce, pine, and r species account for the largest share of the forest plantation area in Europe, with Eucalyptus species introduced from Australia common in the south. While eucalypt plantations appear to be avoided by some bats (Di Salvo etþ  al. 2009), positive selection was found for the Mediterranean horseshoe bat, Rhinolophus euryale, in the Basque country (Aihartza etþ  al. 2003). In Spain, R. euryale and Mehely’s horseshoe bat, R. mehelyi, both closed-space foragers, were radio-tracked foraging in eucalypt plantations and dehesa (managed oak savanna) in proportion to, or greater than, their availability (Russo etþ  al. 2005a, b ). Numerous acoustic and radio-tracking studies have documented avoidance of bats from non-native coniferous plantations in Europe (e.g. Entwhistle etþ  al. 1996; Walsh and Harris 1996). Perhaps as a consequence of this, the effects of plantation forestry practices on bat populations in Europe have been largely ignored, and surprisingly little is known about the use of timber plantations by bats. However, several long-running articial ‘bat box’ schemes operated by the UK’s Forestry Commission have indicated that some plantations contain large roosting bat populations (Park etþ  al. 1998). Radio-tracking of Natterer’s bat, Myotis nattereri, a species previously associated primarily with deciduous forests has uncovered the extensive use of areas used for commercial forestry, both for roosting and foraging (Mortimer 2006). This study conducted in a plantation in Scotland found that M. nattereri preferentially foraged within areas of Corsican pine, Pinus nigra var. maritima, and roosted in cavities formed from live double-leadered Corsican pine (Mortimer 2006). Given life-history parameters of the bats studied (survival, population densities) were similar or higher than those described within deciduous for ests, and that double-leadered trees are usually targeted for removal by þ­ foresters as uneconomic, such ndings illustrate the importance of studies in plantation forests.

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136 B. Law et al.A high percentage of open ground in some planted forests can benet species that specialise on the predation of ground dwelling prey. Greater mouse-eared bat, Myotis myotis, for example, while often associated with deciduous forests, was found preferentially foraging in mature spruce monocultures with a high percentage of open ground in Germany, and intensively managed orchards and lowland forests with no undergrowth in Switzerland (Arlettaz 1999; Zahn etþ  al. 2004). These studies collectively suggest that it is the forest structure that may be more important than tree species composition in many cases. Therefore, it seems clear that timber plantations have the potential to be of value to bats, but we lack an understanding of how populations of different species are affected by current silviculture practices.5.4.6þ PreyThe response of bat prey is also a critical issue when evaluating silvicultural treatments. Lepidoptera (moths–a fundamentally important prey group of bats) in temperate zone forests of North America differ little in species richness between stands regenerating after harvest and stands that remain unharvested (Burford etþ  al. 1999; Summerville and Crist 2002; Dodd etþ  al. 2008). Group selection logging of Australian eucalypt forests has found greater insect biomass in old regrowth Jarrah forest (>30þ  years since logging) than younger forest treatments (Webala etþ  al. 2011) and a similar trend was found in spotted gum forests in eastern Australia (Adams etþ  al. 2009). An additive effect of insect abundance and an index of vegetation openness in the spotted gum forests inuenced bat activity, especially edge-space species with medium to high echolocation frequency. High values of insects and openness correlated with high levels of bat activity (Adams etþ  al. 2009). Thus, dense clutter appears to constrain activity of some species even where insect abundance is high. This varies between bat ensembles, however, with closed-space foragers able to take advantage of the higher insect densities often associated with clutter, particularly Diptera, an important taxa for many bats (Müller etþ  al. 2012; Fuentes-Montemayor etþ  al. 2013; see also Sect.þ  4.3). While the prey base of bats can probably be sustained with application of many silvicultural systems, clearcut stands regenerating as monocultures support reduced levels of moth diversity, indicating that plant species richness is important for providing adequate populations of lepidopteran prey for insectivorous bats in managed for ests (Summerville and Crist 2002; Dodd etþ  al. 2012).5.5þ Multi-spatial Scale Forest ManagementIntegrating silvicultural systems into managed forested landscapes in ways that promote habitat for forest bats must account for the fact that bats are highly mobile and exhibit considerable variability in the use of habitats both spatially

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137 5þ Insectivorous Bats and Silviculture …and temporally (Duchamp etþ  al. 2007). Given that resource requirements differ among species and also sex, age, and reproductive classes within species (Perry etþ  al. 2007a; Perry and Thill 2007b; Henderson etþ  al. 2008), designing a forestedlandscape matrix with a mosaic of resources that addresses the needs of all bat species in the region will likely require the application of a mix of silvicultural methods, each implemented with different objectives in mind (Guldin etþ  al. 2007). These would include the retention of mature forest habitat at the landscape and stand scale in the form of large reserves, narrow and large strips, streamside reserves, aggregates, and clumps (Gustafsson etþ  al. 2012). Lindenmayer and Franklin (2002) proposed a strategic landscape-scale approach with þ­ conservation measures applied at multiple spatial scales for forests. The four main strategies identied for conservation at multiple spatial scales include: (1) establishment of large ecological reserves, (2) application of landscape-level measures in offreserve areas, (3) application of stand-level measures in off-reserve areas, and (4) monitoring and adaptive management. There are limited data on bats for setting overall retention thresholds at the landscape scale. Gustafsson etþ  al. (2012) suggested a strict minimum of 5–10þ  % retention of old-growth forest to achieve a positive ecological response for biodiversity. However, considerably higher levels are often recommended. For example, in Tasmania, 30þ  % is retained in some state forests (Gustafsson etþ  al. 2012; see also Biaowieza Forest in Europe ~20þ  %, Ruczyn´ski etþ  al. 2010). This retention should be spread across the landscape to facilitate dispersal. A key question is whether there are thresholds for the retention of mature forest that can optimise the trade-off between biodiversity conservation and production. A recent study on Tasmanian bats, using both radio-tracking and acoustic detectors, assessed the response of bats to multi-spatial scale forest management (Cawthen etþ  al. 2013). At broader scales, maternal bat colonies selected roosts in landscapes with the highest availability of hollow-bearing trees. At more nescales, however, maternal colonies did not exhibit strong selection for roost trees in patches with the highest availability of hollow-bearing trees. Instead, other attributes such as hollow type were important. For overall bat activity, the extent to which bats used different types of retained forest patches varied with the composition of the surrounding landscape. Large strips and small patches of wooded habitat were used by bats to a greater extent in landscapes with less mature forest in the surrounding area (<1þ  km radius). For small patches, this corresponded to landscapes with <22þ  % mature forest in the surrounding 1þ  km. No thresholds in bat activity were identied for large patches (370þ  ha) or small corridors (3þ  ha). Overall, these results indicate that in the landscapes sampled, activity is low in small retained patches where mature forest is readily available nearby, though these habitat elements do provide roosts and connectivity (and probably foraging habitat) where mature forest is rare or has been lost. Thus, the type, amount, and spatial arrangement of mature forest existing in the landscape need to be considered when retaining forest habitat at ner-spatial scales. Clearly, the extent to which forest bats respond to changes at the landscape scale remains only partially understood. Studies of bat activity at stand and

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138 B. Law et al.landscape scales in both north-western and north-eastern forests of North America demonstrated that patterns in habitat use of bats were largely determined by habitat characteristics at the local or stand level and not at landscape scales (Erickson and West 2003; Ford etþ  al. 2006), suggesting that silvicultural systems that create a mosaic of treatments across forested landscapes with local differences in habitat structure will support a higher overall diversity of bat species (Wigley etþ  al. 2007). This approach has been recommended in published studies (Loeb and Waldrop 2008; Perry etþ  al. 2008); however, other sources report both stand and landscape metrics in North America and Europe to be important in selection of activity areas of bats (Loeb and O’Keefe 2006; Yates and Muzika 2006; FuentesMontemayor etþ  al. 2013), with tri-coloured bats, P. subavus, and eastern red bats, L. borealis, most affected by local stand structure, northern long-eared bats, M. septentrionalis, negatively affected by forest edge, and Indiana bats, M. sodalis, positively affected by dead tree density and non-forested land cover. Other studies corroborate that selection of roosting sites in both barkand cavity-roosting and foliage-roosting bat species is strongly inuenced by landscape-scale metrics in both eastern and western forests of North America (Limpert etþ  al. 2007; Perry etþ  al. 2008; Arnett and Hayes 2009; Lacki etþ  al. 2010).5.6þ Summary and Future PossibilitiesThis review of the effects of silvicultural systems on forest bats demonstrated that almost all treatments evaluated were compatible with some use by forest bats, depending on the suite of species considered: closed-space species feed in intact forests, but respond to creation of small canopy gaps and less to reduced tree densities and open-edge interfaces; edge-space species exploit edge habitat along tracks, coupe edges, and other linear features such as creeks, but fare poorly within dense regrowth that often dominates soon after harvest; and, open-space foragers benet temporarily from silvicultural treatments that signicantly reduces cluttered air space and provides edge interfaces for roosting. These patterns were largely consistent across three different continents. To sustain high levels of bat diversity in managed forests at the landscape scale, a balance of needs for these three groupings of bats is desirable and will likely require a mix of silvicultural treatments and exclusion areas staggered across the landscape, regardless of forest type or geographic region. Use of edge habitats, exclusion areas/set-asides, and riparian corridors for roosting and foraging by bats was a consistent theme in the literature reviewed, and these habitat elements need to be considered in forest planning. These landscape features accompany forest fragmentation, however, and it remains unclear to what extent increasing loss of the unharvested forest matrix will lead to declines in population numbers of for est bats. Unfortunately, data on densities of occupied roosts and, thus, potential for landscape-scale population estimates of bats are few (Clement and Castleberry 2013; Fleming etþ  al. 2013). Regardless, population studies could integrate the

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139 5þ Insectivorous Bats and Silviculture …potential benets of multiple prescriptions at a scale over which bats themselves sample the landscape. Population studies are likely to provide the ultimate test of the effectiveness of a silvicultural regime, especially when such studies take a long-term perspective. Long-term studies on forest bats are notably lacking in the published literature. Application of silvicultural treatments in regenerating forests to reduce tree densities and open gaps in the forest canopy shows promise for creating forested landscapes that support diverse and sustainable populations of bats. Forests with reduced tree density and vegetative clutter permit higher levels of light penetration, with this increased exposure hypothesised to enhance the suitability of live and dead trees for roosting by barkand cavity-roosting bats in temperate climates (Boyles and Aubrey 2006). Further, LiDAR studies demonstrate that reduced clutter in the midand understory layers of forests is correlated with higher levels of activity by low-frequency (34þ  kHz) open-space bats (Britzke etþ  al. 2011; Dodd etþ  al. 2013). However, closed-space bat species that glean insects from vegetation and manoeuvre well within clutter benet from a relatively dense understorey and higher tree densities, which can act as sources of insect prey (FuentesMontemayor etþ  al. 2013). Therefore, management that encourages habitat heter ogeneity to full the requirements of different species is needed. Bat activity is also vertically stratied, but there is a paucity of information on the effects of high canopy forest structure on bat activity (Adams etþ  al. 2009; Müller etþ  al. 2013), and research to address this gap would be valuable. The quality and density of old trees in exclusion areas must not be overlooked. Roost abundance stands out as a key variable in our conceptual model (Fig.þ  5.1). The posited relationship is for increasing bat populations with increasing numbers of roosts, though with a threshold at the upper end of roost abundance rather than at low roost abundance. Densities of hollow trees sufcient to support populations of roosting bats are unknown and remain a major knowledge gap (Law 1996), but will likely be species contingent and based on roost switching behaviours and social dynamics within colonies (Johnson etþ  al. 2013) and the density of other hollow-dependent fauna. Even small colonies of bats can require a large þ­ number of roosts over the active season. For example, Russo etþ  al. (2005a, b) estimated that over a period of a month a colony of 12 female barbastelle bats, B. þ­ barbastellus, would require approximately 18 different trees for roosting. Although the retention and sustained recruitment of large mature trees at various stages of decay is essential in harvested forests for the future long-term maintenance of bat roosts and other hollow-dependent fauna, this might best be achieved through regular harvest exclusion areas (unharvested buffers, old-growth forest, etc.) that can maintain high local densities of potential roosts. There remains little guidance on how much undisturbed forest should be retained at a landscape scale. Paradigm shifts in forest management away from even-aged to retention systems (Puettmann etþ  al. 2009) are already in place in Pacic coast forests of North America and Australian eucalypt forests and are being encouraged for use in management of forests globally (Gustafsson etþ  al. 2012; Lindenmayer etþ  al. 2012). These systems allow for maintenance in post-harvest forests of tree species compositions,

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140 B. Law et al.canopy structures, and ecosystem functions typical of preharvest conditions. We conclude from our review that the use of multi-scale retention systems may be a compatible approach for sustaining habitats of bats in forests. These silvicultural systems are designed to provide spatial variation in retained tree densities and distribution of residual patches of uncut forest, both of which lead to habitat complexity within stands and across landscapes. These systems intentionally mimic natural disturbance regimes and have broad biodiversity benets across multiple taxa (Long 2009). Retention of old forest patches is likely to be most important where harvest intensity is high, such as in clearcut or heavy selection practices, or where retention of critical habitat components is low. Stand-level (site-scale) retention should be greater where old-growth forest in the surrounding landscape is scarce and where logging practices are more intense. The effectiveness of this multi-scale approach will require testing through monitoring and research tailored for different environments, multiple taxa and silvicultural practices. Monitoring the effectiveness of these strategies is an essential part of adaptive management and a fundamental part of ecological sustainable forestry and the ‘social license to operate’ that is increasingly required by forest certication schemes (Lindenmayer and Franklin 2002).Acknowledgementsþ Thanks to L. Cawthen, L. Lumsden, T. Kingston, J. Müller, D. Russo, and J. Williams for helpful comments on a draft.þ GlossaryClearcut/Clear-fell Harvestþ Also referred to as uniform selection and heavy group selection, it removes all trees from a large management area and allows natural regeneration to take place, resulting in even-aged regrowth with high stem density. The aim is to mimic natural stand replacing events such as wildre or large storms Coupe/Cutblocksþ A dened area of forest, which may vary in size, in which har vesting takes place usually over one year Deferment Harvestsþ Sometimes also referred to as a shelterwood or clearcut with reserves. A deferment harvest retains a limited number of canopy trees (reserve trees) while allowing regeneration in the understory. These two tree levels are then allowed to develop together until the end of the next rotation, whereupon other trees are retained for canopy cover Forest Zoningþ Where management for multiple objectives in a forest incorporates broad exclusion areas such that logging is excluded from patches of forest deemed to be environmentally sensitive or where patches of forest are specied to allow different silvicultural practices (Florence 1996) Gap Releaseþ Creation of canopy gaps typically <0.1þ  ha to allow the growth of younger, often suppressed trees Green Treeþ The retention of live trees on an otherwise harvested area as part of a variable retention harvest

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141 5þ Insectivorous Bats and Silviculture …Group Selection Harvestþ Removes all trees from small patches, with the aim of using disturbance to stimulate regeneration of new trees, but simultaneously maintaining a well-connected mosaic of patches of varying size, containing varying numbers of residual mature trees LiDARþ A remote sensing technology that measures distance by illuminating a tar get with a laser and analyses the reected light Patch Cutsþ An area of felling smaller than a clearcut but removing a larger number of trees than a group selection harvest Prescriptionsþ Targeted retention that aims to mitigate the effects of logging on environmental features. Hollow tree retention and riparian exclusion zones are two common prescriptions, but can also include exclusion zones surrounding signicant bat roosts SeedTree Harvestþ The retention of a few residual trees in a harvested area to provide seeds for the forest to regenerate Self-thinningþ Density-dependent mortality within an even-aged stand of trees as they grow in size, leading to reduced tree density Shelterwood Harvestþ See deferment harvest. Shelterwood Systemsþ Removal of canopy trees in a series of selective harvests leaving sufcient trees for regeneration and shelter. New seedlings are left to establish before mature trees are removed Silvicultureþ The art and science of manipulating a stand of trees by controlling the supplies of water, nutrients, and solar radiation by altering forest structure, towards a desired future condition (Guldin etþ  al. 2007), typically for timber production but also for biodiversity conservation goals Single Tree Selectionþ Removes a scattering of high value individual trees from management areas, with repeat cuts taking place at regular intervals over time. However, intensity can vary. Cumulative effects can result in reduced hollow tree density unless there is a specic retention of old trees Standþ A group of forest trees sufciently uniform in species composition or age to be considered a management unit Thinningþ Felling to decrease tree stem density within young regrowth forests to reduce competition for resources among trees and promote the growth of the stand (Florence 1996) Variable Retention Harvestsþ Creation of multi-aged stands in clearcut zones by retaining clumps, patches, or aggregates of old trees within the clearcutOpen Accessþ This chapter is distributed under the terms of the Creative Commons Attribution Noncommercial License, which permits any noncommercial use, distribution, and reproduction in any medium, provided the original author(s) and source are credited.

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150 B. Law et al. Swystun MB, Psyllakis JM, Brigham RM (2001) The inuence of residual tree patch isolation on habitat use by bats in central British Columbia. Acta Chiropter 3:197–201 Taylor RJ, Savva NM (1988) Use of roost sites by four species of bats in state forest in southeastern Tasmania. Aust Wildl Res 15:637–645 Thomas DW (1988) The distribution of bats in different ages of Douglas-r forests. J Wildl Manage 52:619–626 Tibbels AE, Kurta A (2003) Bat activity is low in thinned and unthinned stands of red pine. Can J For Res 33:2436–2442 Titchenell MA, Williams RA, Gehrt SD (2011) Bat responses to shelterwood harvests and forest structure in oak-hickory forests. For Ecol Manage 262:980–988 U.S. Department of Agriculture Forest Service (2009) U.S. forest resource facts and historic trends: forest facts 1952–2007 US metric revised rev072411. FS-801þ  M. Available at: http://a.fs.fed.us U.S. Department of Agriculture and U.S. Department of the Interior (1994) Record of decision for amendments to Forest Service and Bureau of Land Management planning documents within the range of the northern spotted owl. U.S. Forest Service, Portland, Oregon U.S. Fish and Wildlife Service (2009) Range-wide Indiana bat protection and enhancement plan guidelines. Available at: http://www.fws.gov/frankfort/pdf/inbatpepguidelines U.S. Fish and Wildlife Service (2012) North American bat death toll exceeds 5.5 million from white-nose syndrome. News Release, U.S. Fish and Wildlife Service, Arlington, Virginia Vonhof MJ, Barclay RMR (1997) Use of tree stumps as roosts by the western long-eared bat. J Wildl Manage 61:674–684 Waldien DL, Hayes JP, Arnett EB (2000) Day-roosts of female long-eared myotis in western Oregon. J Wildl Manage 64:785–796 Walsh AL, Harris S (1996) Factors determining the abundance of vespertilionid bats in Britain: geographical, land class and local habitat relationships. J Appl Ecol 33:519–529 Watrous KS, Donovan TM, Mickey RM etþ  al (2006) Predicting minimum habitat characteristics for the Indiana bat in the Champlain Valley. J Wildl Manage 70:1228–1237 Wear DN, Greis JG (2013) The southern forests futures project. USDA Forest Service, Southern Research Station, Gen. Tech. Rep. SRS-178, Asheville, North Carolina Webala PW, Craig MD, Law BS etþ  al (2010) Roost site selection by southern forest bat Vespadelus regulus and Gould’s long-eared bat Nyctophilus gouldi in logged jarrah forests, south-western Australia. For Ecol Manage 260:1780–1790 Webala PW, Craig MD, Law BS etþ  al (2011) Bat habitat use in logged jarrah eucalypt forests of south-western Australia. J Appl Ecol 48:398–406 Wigley TB, Miller DA, Yarrow GK (2007) Planning for bats on forest industry lands in North America. In: Lacki MJ, Hayes JP, Kurta A (eds) Bats in forests: conservation and management. Johns Hopkins University Press, Baltimore, pp 293–318 Willis CKR, Brigham RM (2004) Roost switching, roost sharing and social cohesion: forestdwelling big brown bats, Eptesicus fuscus, conform to the ssion-fusion model. Anim Behav 68:495–505 Willis CKR, Brigham RM (2005) Physiological and ecological aspects of roost selection by reproductive female hoary bats (Lasiurus cinereus). J Mammal 86:85–94 World Resources Institute (2014) Southern forests for the future. Available at: http://www.seesou thernforests.org Wunder L, Carey AB (1996) Use of the forest canopy by bats. Northwest Sci 70:79–85 Yates MD, Muzika RM (2006) Effect of forest structure and fragmentation on site occupancy of bat species in Missouri Ozark forests. J Wildl Manage 70:1238–1248 Young RA, Giese RL (eds) (2003) Introduction to forest ecosystem science and management, 3rd edn. John Wiley and Sons, New York Zahn A, Haselbach H, Güttinger R (2004) Foraging activity of central European Myotis myotis in a landscape dominated by spruce monocultures. Mamm Biol 70:265–270 Zimmerman GS, Glanz WE (2000) Habitat use by bats in eastern Maine. J Wildl Manage 64:1032–1040

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151Chapter 6Bats in the Anthropogenic Matrix: Challenges and Opportunities for the Conservation of Chiroptera and Their Ecosystem Services in Agricultural LandscapesKimberly Williams-Guillén, Elissa Olimpi, Bea Maas, Peter J. Taylor and Raphaël Arlettaz© The Author(s) 2016 C.C. Voigt and T. Kingston (eds.), Bats in the Anthropocene: Conservation of Bats in a Changing World, DOIþ  10.1007/978-3-319-25220-9_6Abstractþ Intensication in land-use and farming practices has had largely negative effects on bats, leading to population declines and concomitant losses of ecosystem services. Current trends in land-use change suggest that agricultural areas will further expand, while production systems may either experience further intensication (particularly in developing nations) or become more environmentally friendly (especially in Europe). In this chapter, we review the existing literature K. Williams-Guillénþ  ()þ  Paso Pacíco and School of Natural Resources and Environment, Ventura, CA, USA e-mail: kim@pasopacico.org K. Williams-Guillénþ  University of Michigan, Ann Arbor, MI, USA E. Olimpiþ  Environmental Studies Department, University of California, Santa Cruz, CA, USA e-mail: eolimpi@gmail.com B. Maasþ  Agroecology Group, University of Gottingen, Gottingen, Germany e-mail: beamaas@gmx.at P.J. Taylorþ  South African Research Chair in Biodiversity Value and Change and Centre for Invasion Biology, University of Venda, Thohoyandou, South Africa e-mail: peter.taylor.univen@gmail.com R. Arlettazþ  Division of Conservation Biology, Institute of Ecology and Evolution, University of Bern, Bern, Switzerland e-mail: raphael.arlettaz@iee.unibe.ch R. Arlettazþ  Swiss Ornithological Institute, 6204 Sempach, Switzerland

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152 K. Williams-Guillén et al.on how agricultural management affects the bat assemblages and the behavior of individual bat species, as well as the literature on provision of ecosystem services by bats (pest insect suppression and pollination) in agricultural systems. Bats show highly variable responses to habitat conversion, with no signicant change in species richness or measures of activity or abundance. In contrast, intensication within agricultural systems (i.e., increased agrochemical inputs, reduction of natural structuring elements such as hedges, woods, and marshes) had more consistently negative effects on abundance and species richness. Agroforestry systems appear to mitigate negative consequences of habitat conversion and intensication, often having higher abundances and activity levels than natural areas. Across biomes, bats play key roles in limiting populations of arthropods by consuming various agricultural pests. In tropical areas, bats are key pollinators of several commercial fruit species. However, these substantial benets may go unrecognized by farmers, who sometimes associate bats with ecosystem disservices such as crop raiding. Given the importance of bats for global food production, future agricultural management should focus on “wildlife-friendly” farming practices that allow more bats to exploit and þ­ persist in the anthropogenic matrix so as to enhance provision of ecosystem ser vices. Pressing research topics include (1) a better understanding of how local-level þ­ versus landscape-level management practices interact to structure bat assemblages, (2) the effects of new pesticide classes and GM crops on bat populations, and (3) how increased documentation and valuation of the ecosystem services provided by bats could improve attitudes of producers toward their conservation.6.1þ IntroductionAgricultural areas cover approximately 40þ  % of our planet’s terrestrial ecosystems (FAOSTAT 2011), with the 5þ  billion ha of land under farming and grazing now surpassing the extent of the world’s forested areas (Robertson and Swinton 2005; Power 2010). Agricultural areas are expected to continue to expand with increasing human population growth and resultant resource use: Lowand middle-income countries will experience a 100þ  % increase in demand for agricultural products by 2050 (Defries etþ  al. 2010; FAO 2011). In the face of increasing pressure on natural resources, the conservation of remaining natural areas is critical for the survival of multitudes of species. However, the ubiquity of agriculture means that farmland cannot be ignored in the context of landscape-level approaches to biodiversity conservation (Vandermeer and Perfecto 2007; Loos etþ  al. 2014). A growing body of research demonstrates that not only do some agricultural systems harbor high levels of biodiversity and provide a variety of ecosystem ser vices (Tilman 1999; Foley etþ  al. 2005; Tscharntke etþ  al. 2005), but also that char acteristics of these agricultural systems may have profound effects upon remaining natural areas (Perfecto and Vandermeer 2010). Agricultural matrices can vary drastically in their quality and permeability, impacting dispersal rates, and hence,

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153 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities …long-term population stability of organisms found in less disturbed areas (Ricketts 2001; Laurance 2008; Perfecto and Vandermeer 2010; Tscharntke etþ  al. 2012). On a local scale, different agricultural management approaches often coexist. Some rely on varying chemical inputs (pesticides, fertilizer), or novel plant types (e.g., genetically modied crops incorporating genes for characteristics such as insecticide functions), resulting in environmental contamination, pollution, and dissemination of toxins that could negatively impact biodiversity across multiple spatial scales (Nelson etþ  al. 2009; Power 2010). As a consequence, agricultural management has effects not only on biodiversity, but also on human health and economies. In the tropics, the expansion of export-oriented agriculture results from population growth and shifts in consumption patterns of developing nations, and is car ried out mostly to the detriment of old growth forests and extensively managed grasslands such as pastures (Defries etþ  al. 2010; Lambin and Meyfroidt 2011). As a consequence, croplands are still expanding dramatically, and agricultural practices are likely to further intensify in the near future (more chemical and mechanical inputs, reliance on genetically modied plants with novel manufactured traits). Short-term increases in yield will come at the cost of reduced structural and taxonomic diversity within agricultural systems (Loos etþ  al. 2014) and concomitant loss of crucial ecosystem services. An additional factor affecting agriculture in the Anthropocene is climate change and the need to adapt cultures to novel environmental conditions: Many areas may become unsuitable for cultivation of their current dominant crops, while extreme weather events may result in reduced yields. Resulting declines in calorie availability, particularly in the developing world (Nelson etþ  al. 2009), will increase the need for agricultural practices that meet both productivity and sustainability goals (Tilman etþ  al. 2002; McShane etþ  al. 2011; Tscharntke etþ  al. 2012). These trends portend major shifts in land-use patterns (Lambin and Meyfroidt 2011) and hence biodiversity, with agricultural intensication, forest and tree roost loss anticipated to have particularly negative effects on bat species richness, abundance, and functional diversity (Fischer etþ  al. 2009, 2010; Jones etþ  al. 2009). These emerging trends pose major threats to farmland bat assemblages and populations (Jones etþ  al. 2009; Kunz etþ  al. 2011) and could negatively impact human populations by altering the ecosystem services that bats provide. Thus, there is a critical need to assess how agricultural management affects bat populations, and how affected bat populations will in turn affect agricultural production. In this chapter, we review the effects of agricultural land use and management on bat assemblages and the behavior and ecology of individual bat species at eld, farm, and landscape scales (Vickery and Arlettaz 2012). We also review the developing literature on ecosystem services—and disservices—provided by bats in agricultural areas. Finally, we synthesize this information to suggest key management recommendations necessary to maintain bat populations in agricultural landscapes and highlight critical knowledge gaps that must be resolved in order to conserve bat diversity and ecosystem functions in a planet increasingly dominated by food production.

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154 K. Williams-Guillén et al.6.2þ MethodsWe used the Web of Knowledge, Google Scholar, and PubMed search engines to locate publications with the keywords “bats” AND “agriculture,” “agroforestry,” “farm,” and “farmland.” Given the potential importance of bats in provisioning ecosystem services in agricultural areas, we also searched for “bats” AND “ecosystem services,” “pollination,” “pest consumption,” “pest control,” and “pest limitation.” The majority of sources stemmed from peer-reviewed publications, although we also included Master’s and Ph.D. theses and published reports if results from the study in question were not available as journal articles. We also inspected the bibliographies of relevant publications. Each co-author focused on a specic geographic area (RA, assisted by Olivier Roth: Europe; BM: Australia and tropical Asia; EO: temperate North America; PT: sub-Saharan Africa; KWG: tropical Americas). Our searches were limited to publications with English language text or summaries. We focused on agriculture and animal husbandry for the production of calories for human or animal consumption, excluding forestry systems dedicated to timber or ber production (see Law etþ  al., Chap.þ  4), studies in which fallows or abandoned elds were the only agricultural systems investigated, as well as investigations that focused on fragmentation without explicit consideration of the effect of agricultural matrix (see Meyer etþ  al., Chap. 3). We divided results from the literature search into two broad categories of investigations: (1) How agricultural practices affect bat assemblages, ecology, behavior, and/or physiology; and (2) how bats affect agriculture through the provision of ecosystem services such as pollination and pest suppression. Within the rst category, most studies addressed effects of land conversion and agricultural management on bat assemblage structure, abundance, activity levels, and behavior. We further subdivided results to consider habitat conversion to agriculture and agricultural intensication. We dene agricultural intensication as consisting of at least one of the following: decreased structural complexity of native vegetation (natural and seminatural elements structuring the landscapes such as woodland patches and hedges), increased application of agrochemicals (pesticides, fertilizer), increased crop plant density, increased mechanization, or increased reliance on GM plants. We reviewed results from searches to locate studies which contrasted aspects of bat assemblage structure, abundance, activity, ranging behavior, or diet in either natural and agricultural habitat, or different agricultural systems of contrasting management. To better quantify the responses of bats to habitat conversion and agricultural intensication across multiple disparate studies, we conducted a meta-analysis. We emphasize that this meta-analysis is based on correlational studies, rather than from controlled experiments; because assignment of treatment locations is not randomized in the majority of these studies, confounding factors could result in spurious effect sizes (Egger etþ  al. 1998). We thus view our meta-analysis as a tool for exploring trends across a diverse suite of studies, with limited conclusive power.

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155 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities …We reviewed studies for the inclusion of mean values for at least one response variable in both natural and agricultural areas, or two or more agricultural areas of differing intensication; 32 studies using mist netting, harp trapping, acoustic monitoring, or a combination of these methods included appropriate data. We classify the response variable metrics into two separate categories for analyses, measures of species richness and measures of relative activity or abundance (i.e., pass rates from acoustic monitoring or capture rates from mist netting). We also consider habitat conversion and intensication responses separately. For each pairwise comparison (natural–agricultural, or agricultural–agricultural), we calculated the effect size as the log odds ratio of the mean value from the lower intensity system divided by the mean from the higher intensity system. Thus, a positive effect size indicates higher species richness or activity/abundance in natural versus agricultural areas or lower intensity versus higher intensity agriculture. We followed García-Morales etþ  al. (2013) and considered mean effect sizes with 95þ  % condence intervals that did not include 0 as indicative of a signicant effect. In the case of studies comparing multiple natural or agricultural habitats or presenting means for multiple species or species groups (i.e., producing multiple pairwise comparisons for any given combination of metric and response type), we averaged the odds ratio to avoid pseudo replication. Due to the diverse nature of the studies and a lack of clarity about numbers of replicates in some studies, we did not weight studies by sample size or replicates. For our analysis, we thus considered each study as an equally weighted case for the nal model. We conducted analyses in R Version 3.0.2 (R Development Core Team 2013) using the packages lme4 and lmerTest. This diverse set of studies includes different methods (e.g., acoustic monitoring versus mist netting) from different regions with ecologically and taxonomically characteristic bat assemblages. To account for some of this variation, we included study method and continent as random effects. Fixed factors included latitudinal zone (temperate, subtropical, and tropical) and whether or not the high-intensity system comprised an agroforestry system (including monocultural orchards). We also located several studies on ecotoxicology and demography, focusing on the effects of pesticide and GMOs use on bats. A complete review of the effects of pesticides on bats is beyond the scope of this chapter, particularly since bats and contaminants have received recent reviews (O’Shea and Johnston 2009; Bayat etþ  al. 2014). We therefore focus on studies that explicitly link bat agrochemical exposure to changes in bat populations. Similarly, although fertilizers comprise a large portion of the chemical inputs to agriculture, their impacts on bats are indirect. In considering the benets of bats for agricultural production (i.e., crop yield), we focus on the provision of two ecosystem services: agricultural pest limitation by insectivorous bats and pollination by tropical bats. We did not consider their role as seed dispersers since human management of farmland vegetation limits the effect and value of bat seed dispersal. Similarly, although bat pollination is key for the unmanaged reproduction of several economically important crops, such as

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156 K. Williams-Guillén et al.bananas and agaves (Kunz etþ  al. 2011), we did not consider these particular crops because they are mostly propagated vegetatively in such plantations. We instead focus on crops that are almost exclusively reliant on bat pollination under standard cultivation practices. Multiple investigations have characterized the diets of insectivorous bats at the order level, claiming potential consumption of pest insects. To more condently assess consumption of insects damaging crops, we focused on studies in which known (species level identity) or probable (family level identity) agricultural pests were identied from feces of bats foraging in farms or areas dominated by agriculture. We exclude dietary studies that have sampled exclusively from natural habitats or do not describe the agricultural systems within which bats may have been foraging. We also briey contrast these with ecosystem disservices of bats in agricultural areas. Bats are associated with costs to agriculturalists, particularly in the subtropics and tropics where frugivorous bats raid crops and sanguivorous bats attack domestic livestock. As with other sections, we focus on direct impacts on productive systems and do not consider the impacts of bat transmission of disease except where it directly impacts agriculture. The majority of the nearly 140 investigations reviewed in this chapter have been conducted in temperate North America and Europe (Fig.þ  6.1). The bulk of studies documenting how habitat conversion or agricultural intensication affects bats has been conducted in Europe and the Neotropics (Fig.þ  6.1, Tableþ  6.1). Within temperate zones, studies have focused mainly on annual cultivars and pasture, while research in tropical areas is dominated by studies on agroforestry systems, particularly coffee and cacao. Results on ecotoxicology of farmland bats come primarily from North America. Studies demonstrating the consumption of agricultural pests also derive primarily from North America, whereas studies of other ecosystem services provided by bats are limited to the tropics. Fig.þ  6.1þ Locations of studies on effects of habitat conversion or agricultural intensication (red diamonds) on bats, pesticide contamination (pink triangles) on bats, and ecosystem services (green squares) provided by bats in agriculture

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157 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities …Tableþ  6.1þ Studies investigating the effects of agriculture on bat assemblage structure, ecology, or behavior across six continents Source Biome/life zone Agricultural system Bat taxa assessed Conversion response Intensication response North America Braun de Torrez (2014)aTemperate woodland savannah (mesquite-juniper) Native and commercial pecan groves General bat assemblage species richness, abundance, activity, habitat use species richness, abundance, activity, by species Farrow and Broders (2011) Boreal forest, temper ate broadleaf forest Mixed agricultural landscape Perimyotis subavus activity levels Gehrt and Chelsvig (2003)aTemperate prairie, woodlands, and wetlands “Intensive” agricultural landscape General bat assemblage activity, habitat selection Henderson and Broders (2008) Boreal forest, temper ate broadleaf forest Mixed agricultural landscape Myotis septentrionalis mobility Rambaldini and Brigham (2011)aMontane forest Vineyards Antrozous pallidus activity Tuttle etþ  al. (2006) Arid desert Rangeland with troughs Myotis spp., Antrozous pallidus drinking efciency Europe Arlettaz and Perrin (1995), Arlettaz (1996,1999), Arlettaz etþ  al. (1997), 2001) Temperate agricultural landscape within European Alps Mixed farmland Myotis myotis, Myotis blythii foraging habitat selection habitat use, dietary diversity Bontadina etþ  al. (2002) Temperate agricultural landscape Mixed farmland Rhinolophus hipposideros foraging habitat selection Boughey etþ  al. (2011) Temperate agricultural landscape Conventional farmland Pipistrellus spp., Nyctalus noctula, Eptesicus serotinus activity Davy etþ  al. (2007)aMediterranean landscape Olive groves General bat assemblage activity activity (continued)

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158 K. Williams-Guillén et al.Tableþ  6.1þ (continued) Source Biome/life zone Agricultural system Bat taxa assessed Conversion response Intensication response Dietz etþ  al. (2013) Temperate agricultural landscape Traditional farmlands (orchards, meadows, pastures) Rhinolophus fer rumequinum, Myotis emarginatus habitat selection habitat selection Downs and Racey (2006) Temperate agricultural landscape Mixed farmland with woodlands and hedges Pipistrellus spp., M. daubentonii activity activity Drescher (2004) Temperate agricultural landscape Apple orchards, vineyards, pastures Myotis myotis foraging activity foraging activity Ekman and de Jong (1996)aTemperate agricultural landscape Habitat islands within crop elds Myotis brandti, Eptesicus nilssoni, Plecotus auritus, Pipistrellus pipistrellus activity, species occurrence Flaquer etþ  al. (2008) Mediterranean landscape Olive groves, fruit orchards, rice Myotis emarginatus foraging activity (olive groves) foraging activity (orchards, rice) Frey-Ehrenbold etþ  al. (2013)aTemperate agricultural landscape Seminatural to intensively managed farmland General bat assemblage activity, species richness Fuentes-Montemayor etþ  al. (2011) Temperate agricultural landscape Conventional farmland and agri-environment scheme farmlands Pipistrellus spp. activity Fuller etþ  al. (2005)aTemperate agricultural/ woodland landscape Organic and nonorganic cereals General bat assemblage species density, activity Jones and Morton (1992) Temperate agricultural/ woodland landscape Hay/silage, grazing Rhinolophus ferrumequinum activity de Jong (1995) Temperate agricultural landscape Agriculture-dominated landscape Myotis spp., Pipistrellus pipistrellus, Plecotus auritus activity Lesin´ski etþ  al. (2013) Temperate agricultural and woodlands Annual crops (organic and conventional) Eptesicus serotinus activity (continued)

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159 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities …Tableþ  6.1þ (continued) Source Biome/life zone Agricultural system Bat taxa assessed Conversion response Intensication response Lisón and Calvo (2011)aSemiarid Mediterranean landscape Rainfed olive/almond, irrigated peach/citrus General bat assemblage activity Lisón and Calvo (2013) Semiarid Mediterranean landscape Rain fed crops/xerophytic vegetation Pipistrellus spp. activity (varies by species) Lundy and Montgomery (2010) Temperate agricultural landscape Improved and unimproved pasture General bat assemblage foraging activity Obrist etþ  al. (2011)aTemperate agricultural landscape Managed and abandoned chestnut orchards General bat assemblage foraging activity Pocock and Jennings (2007)aTemperate agricultural landscape Organic/conventional farmland; hay/silage elds General bat assemblage activity (loss of linear features), activity (agrochemical use, silage) Rainho (2007) Semiarid Mediterranean landscape Cereal crops and olive groves General bat assemblage activity activity Russ and Montgomery (2002) Temperate agricultural landscape Mixed farmland with woodlands, tree lines General bat assemblage activity activity Russo and Jones (2003)aMediterranean landscape Traditional farmland habitats, chestnut woodland General bat assemblage activity Russo etþ  al. (2002) Mediterranean landscape Olive groves, traditional farmlands Rhinolophus euryale home range composition, foraging time Stahlschmidt etþ  al. (2012), Stahlschmidt and Brühl (2012)aTemperate agricultural/ woodland landscape Apple orchard General bat assemblage activity (Pipistrelle, Eptesicus), activity (Myotis) (continued)

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160 K. Williams-Guillén et al.Tableþ  6.1þ (continued) Source Biome/life zone Agricultural system Bat taxa assessed Conversion response Intensication response Verboom and Huitema (1997) Temperate agricultural landscape Linear features in traditional farmland landscapes Pipistrellus pipistrellus, Eptesicus serotinus activity Walsh and Harris (1996a, b) Temperate agricultural landscape Mixed farmland General bat assemblage activity activity Wickramasinghe etþ  al. (2003, 2004) Temperate agricultural landscape Organic and conventional farmland General bat assemblage activity Australia Fischer etþ  al. (2010) Temperate woodland/ agricultural landscape Low tree density livestock pasture General bat assemblage activity Hanspach etþ  al. (2012) Temperate woodland/ agricultural landscape Pasture with varying levels of tree cover General bat assemblage activity, species richness (peaks at inter mediate tree cover) activity, species richness Lentini etþ  al. (2012)aTemperate woodland/ agricultural landscape Cereal, canola, and pasture-dominated landscape with and without linear features Mollossidae; Vespertilionidae; Emballonuridae activity, species richness, feeding Lumsden etþ  al. (2002) Temperate woodland/ agricultural landscape Fragmented humandominated landscape Nyctophilus geoffroyi, Chalinolobus gouldii roost locations roost locations Lumsden and Bennett (2005)aTemperate woodland/ agricultural landscape Pasture with varying densities of trees General bat assemblage abundance, activity abundance, activity Neotropics Avila-Cabadilla etþ  al. (2009) Tropical dry forest Pasture Phyllostomids abundance, species richness Castro-Luna and Galindo-González (2012)aTropical montane rainforest Diverse and simplied shade coffee, pasture Frugivorous phyllostomids abundance, species richness (continued)

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161 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities …Tableþ  6.1þ (continued) Source Biome/life zone Agricultural system Bat taxa assessed Conversion response Intensication response Estrada etþ  al. (1993) Lowland tropical rainforest Shaded (coffee, cacao, mixed) and unshaded (citrus, allspice) plantations, pastures Phyllostomids, nonphyllostomids sampled with mist nets species richness species richness Estrada and CoatesEstrada (2001) Lowland tropical rainforest Shaded (coffee, cacao, mixed) and unshaded (citrus, allspice) plantation General bat assemblage abundance, expected species richness abundance, expected species richness Estrada and CoatesEstrada (2002)aLowland tropical rainforest Coffee, cacao, citrus, banana, pasture General bat assemblage abundance, species richness Estrada etþ  al. (2004)aLowland tropical rainforest Fencerows, citrus, pasture Non-phyllostomids activity activity Faria (2006)aBrazilian Atlantic forest Shade cacao in a forest dominant landscape General bat assemblage abundance, species richness Faria etþ  al. (2006)aBrazilian Atlantic forest Shade cacao in a cacao dominant matrix General bat assemblage species richness Faria and Baumgarten (2007)aBrazilian Atlantic forest Shade cacao in two contrasting landscapes General bat assemblage abundance, species richness García Estrada etþ  al. (2006, 2012) Montane tropical rainforest Shade coffee Phyllostomids abundance, dietary diversity, species richness abundance, dietary diversity, species richness Harvey etþ  al. (2006)aTropical dry forest High and low tree cover pasture General bat assemblage abundance, species richness (frugivores, nectarivores) abundance, species richness (frugivores), abundance (nectarivores) Harvey and González Villalobos (2007)aTropical humid forest, premontane wet forest Cacao agroforest, banana agroforest, plantain monoculture General bat assemblage abundance, species richness abundance, species richness (continued)

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162 K. Williams-Guillén et al.Tableþ  6.1þ (continued) Source Biome/life zone Agricultural system Bat taxa assessed Conversion response Intensication response Medellin etþ  al. (2000) Montane tropical rainforest Cacao, oldeld, and corn General bat assemblage abundance, species richness abundance, species richness Medina etþ  al. (2007)aTropical moist forest High and low tree cover pasture General bat assemblage abundance, species richness abundance, species richness Numa etþ  al. (2005)aTropical montane rainforest Sun and shade coffee in contrasting landscapes Phyllostomids abundance, estimated species richness estimated species richness (within landscape), estimated species richness (between landscapes) Pineda etþ  al. (2005)aTropical montane cloud forest Shade coffee General bat assemblage abundance, species richness Saldaña Vázquez etþ  al. (2013) Tropical montane rainforest Shade coffee Sturnira ludovici abundance, females Sosa etþ  al. (2008) Tropical montane rainforest Shade coffee General bat assemblage species richness abundance Williams-Guillén and Perfecto (2010)aTropical montane rainforest Shade coffee Phyllostomids abundance, species richness abundance, species richness Williams-Guillén and Perfecto (2011)aTropical montane rainforest Shade coffee Non-Phyllostomids activity (cluttered space foragers), abundance (open space foragers), species richness activity (cluttered space foragers), abundance (open space foragers), species richness Vargas Espinoza etþ  al. (2008) Premontane tropical rainforest Citrus orchards General bat assemblage abundance, species richness Africa Noer etþ  al. (2012) Subtropical Savanna Sugarcane Chaerephon pumilus, Mops condylurus foraging time (continued)

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163 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities …Tableþ  6.1þ (continued) Source Biome/life zone Agricultural system Bat taxa assessed Conversion response Intensication response Randrianandrianina etþ  al. (2006)aTropical rainforest/ mixed use landscape Smallholder gardens General bat assemblage species richness, activity, foraging Sirami etþ  al. (2013)aSubtropical grassland, fynbos Intensive wheat, vineyards, orchards General bat assemblage abundance, species richness abundance, species richness Taylor etþ  al. (2013b)aSubtropical Savanna Macadamia General bat assemblage activity (molossids), activity (vespertilionids) Asia Fukuda etþ  al. (2009)aFragmented tropical rainforest Orchards, palm oil Hipposideridae, Vespertilionidae, Pteropodidae abundance, species richness abundance, species richness Furey etþ  al. (2010)aLimestone karst, tropical rainforest Landscapes dominated by rice paddies and degraded forest Insectivorous bat assemblage abundance, species richness Graf (2010)aMontane tropical rainforest Forest and shade cacao General bat assemblage abundance, species richness species richness Mildenstein etþ  al. (2005) Tropical moist forest Fruit orchards and hardwood tree plantations Pteropodidae bat assemblage habitat selection Sedlock etþ  al. (2008) Montane and premontane rainforest Mosaic of pasture, root crops, orchards, and fallows Insectivorous bat assemblage species accumulation Van Weerd and Snelder (2008)aTropical moist forest Village homegarden polycultures, shrublands used for grazing General bat assemblage abundance, species richness abundance, species richnessa Studies included in meta-analysis “Conversion response” indicates effects of agriculture versus non-anthropogenic habitat; “intensication response” indicates effects of agricultural intensica tion (i.e., increased amount of agrochemicals, decreased structural complexity, infrastructure construction). “ ” indicates a negative effect on specied response variable; “ ” variable effects depending on species, ensemble, or habitat contrast; “ ” no marked effects observed; “ ” positive effect on response variable

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164 K. Williams-Guillén et al.6.3þ Effects of Agricultural Intensity on Bat Assemblage Structure, Behavior, and EcologyWe found 70 studies addressing the effects of habitat conversion or management on the assemblage structure, behavior, or ecology of bats. Fifty-two studies assessed bats in both natural and agricultural areas. Twenty-two studies (42þ  %) demonstrated negative effects of habitat conversion, twelve (23þ  %) showed variable responses (e.g., only some species or ensembles declined, different agricultural systems were associated with different effects), twelve (23þ  %) showed increased richness, activity, or abundance in agricultural areas, and six (12þ  %) showed little or no difference between agricultural and natural areas. Forty-ve studies addressed some aspect of agricultural intensication, with 38 of these (84þ  %) documenting a negative effect of intensication on bats, four showing variable or neutral (9þ  %) responses, while three studies (7þ  %) documented increases in bat richness, abundance, or activity in more intensive systems. Response variables differ in response to habitat conversion and agricultural intensication (Fig.þ  6.2, Tableþ  6.2), with measures of species richness showing no signicant change between treatments. In contrast, measures of relative activity and abundance show stronger responses (Fig.þ  6.2). Agroforestry systems are more structurally similar to the original non-anthropogenic land uses, making them less intensive than annual crops dominated by one plant species or pasture systems lacking structural complexity. This relationship presumably explains why agricultural systems that incorporate trees and other large woody perennials on farms and throughout the agricultural landscape have little effect on bat activity and abundance (Fig.þ  6.2). Agroforestry systems appear to mitigate negative effects on bat assemblages in cases of both habitat conversion and agricultural intensication (Tableþ  6.2). Several studies have considered the effects of agricultural management at landscape scales versus focusing exclusively on farm-level management practices (Estrada etþ  al. 1993; Ekman and de Jong 1996; Verboom and Huitema 1997; Numa etþ  al. 2005; Faria etþ  al. 2006, 2007; Faria and Baumgarten 2007; FuentesMontemayor etþ  al. 2011; Boughey etþ  al. 2011; Maas etþ  al. 2013). Within agricultural areas, bat activity increases with proximity to natural areas (Estrada etþ  al. 1993; Verboom and Huitema 1997; Boughey etþ  al. 2011) and in less fragmented landscapes (Fuentes-Montemayor etþ  al. 2011; Frey-Ehrenbold etþ  al. 2013) or in landscapes with more natural elements such as hedgerows and woodlots (Verboom and Huitema 1997). Agricultural areas also serve as matrix habitat connecting fragmented nonanthropogenic habitats. Although one study has suggested that landscapes dominated by crops and open elds have a stronger negative inuence on bats than water (Ekman and de Jong 1996), a recent analysis of bat responses to isolation on islands versus in forest fragments embedded in agricultural matrix suggests that the anthropogenic matrix is more permeable than water matrix (Mendenhall etþ  al.

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165 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities …2014). Thus, agricultural intensication at the landscape level should make the matrix less permeable due to the reduction of natural resources and structural elements such as trees, affecting not only the persistence of bats in fragmented landscapes, but also the degree to which bat assemblages show a negative response to agriculture. A few investigations have conrmed such interactions between farmand landscape-level intensication: Intensication in cacao matrices in Brazil (Faria etþ  al. 2006, 2007; Faria and Baumgarten 2007) and coffee matrices in Colombia (Numa etþ  al. 2005) resulted in reductions in the species richness and abundance of bats in diverse shade agroforests relative to forest fragments. In Europe, effects of landscape management on bat assemblage structure and ecology in temperate landscapes dedicated to the production of annual crops remain largely unexplored compared to the extensive information available at the eld and farm scales. Relative Abundance/Activity Species Richness Fig.þ  6.2þ Mean effect size (log odds ratio, circles) 95þ  % CI of relative abundance and activity (left) and species richness (right) of habitat conversion versus agricultural intensication (top row ), and of contrasts (both habitat conversion and agricultural intensication) with and without agroforestry systems (bottom row). Positive effect sizes indicate reductions in relative abundance and activity or species richness in response to habitat conversion and intensication

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166 K. Williams-Guillén et al.6.4þ Pesticide Impacts on Bat PopulationsAgricultural intensication may remove potential habitat for bats and their prey; the effects of increased agrochemical inputs, such as increased exposure and changes in prey availability, may put resident bats under further pressure. Tableþ  6.2þ Effects of latitudinal zone and agroforestry systems on effect size (log odds ratio) for two response variable types under habitat conversion and agricultural intensication Parentheses indicate random effects, and bold text indicates best tting model based on AIC value Response variable Land change type Model AIC 2P Abundance/ activity Habitat conversion Effect sizeþ  ~þ  (Method)þ  þ  (Continent) 60.7 Effect sizeþ  ~þ  Agroforestryþ  þ  (Met hod)þ  þ  (Continent) 49.7 13.00 <0.001 Effect sizeþ  ~þ  Latitudeþ  þ  (Method)þ  þ  (Continent) 62.0 0.00 1.000 Effect sizeþ  ~þ  Agroforestryþ  þ  Latitu deþ  þ  (Method)þ  þ  (Continent) 51.9 12.15 <0.001 Intensication Effect sizeþ  ~þ  (Method)þ  þ  (Continent) 52.4 Effect sizeþ  ~þ  Agroforestryþ  þ  (Met hod)þ  þ  (Continent) 49.2 5.22 0.022 Effect sizeþ  ~þ  Latitudeþ  þ  (Method)þ  þ  (Continent) 53.6 0.00 1.000 Effect sizeþ  ~þ  Agroforestryþ  þ  Latitu deþ  þ  (Method)þ  þ  (Continent) 50.6 4.923 0.026 Species richness Habitat conversion Effect sizeþ  ~þ  (Method)þ  þ  (Continent) 20.7 Effect sizeþ  ~þ  Agroforestryþ  þ  (Meth od)þ  þ  (Continent) 21.7 0.99 0.319 Effect sizeþ  ~þ  Latitudeþ  þ  (Method)þ  þ  (Continent) 24.0 0.00 1.000 Effect sizeþ  ~þ  Agroforestryþ  þ  Latitu deþ  þ  (Method)þ  þ  (Continent) 24.1 1.82 0.178 Intensication Effect sizeþ  ~þ  (Method)þ  þ  (Continent) 22.9 Effect sizeþ  ~þ  Agroforestryþ  þ  (Meth od)þ  þ  (Continent) 24.4 0.54 0.460 Effect sizeþ  ~þ  Latitudeþ  þ  (Method)þ  þ  (Continent) 26.3 0.06 0.806 Effect sizeþ  ~þ  Agroforestryþ  þ  Latitu deþ  þ  (Method)þ  þ  (Continent) 27.0 1.34 0.248

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167 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities …Bats may directly consume pesticides by feeding on fruits, owers, and arthropods exposed to chemical application. Even bats foraging outside of agricultural areas can be exposed to pesticides via biomagnication as residues are incorporated into the tissues of organisms at higher trophic levels (Bayat etþ  al. 2014). Investigations of exposure of bats to pesticides and its effects on physiology and mortality rst appeared in the 1970s, amid a wave of growing concern regarding the effects of organochlorine pesticides (e.g., DDT, DDE, dieldrin, lindane, endosulfan, aldrin) on ecosystems and observations of declining bat populations at high-prole sites such as the Carlsbad Caverns in New Mexico, USA (Clark 1988, 2001). In some cases, DDT and other organochlorines were even applied directly to bat roosts in efforts to exterminate “vermin” (Kunz etþ  al. 1977), and declines in high-prole bat colonies were linked to organochlorine use (Clark etþ  al. 1978; Clark 2001). Even sublethal exposure to pesticides can have negative consequences for bats, resulting in increased metabolic rates (Swanepoel etþ  al. 1998), and ingestion of pesticide residues on arthropods may poses a potential reproductive risk to certain bat species (Stahlschmidt and Brühl 2012). Organochlorine residues have been documented in bats in a wide variety of both agricultural and non-agricultural landscapes, although several studies have found increased contaminant loads in bats sampled near agricultural areas (Clark and Prouty 1976; White and Krynitsky 1986) or near sites of pesticide manufacture (O’Shea etþ  al. 2001). In some cases, temporal changes in levels of different contaminants reect shifts in local agricultural practice as farmers adopt new pesticide regimes (Miura etþ  al. 1978; Clark etþ  al. 1980). Organochlorines are notorious for their persistence in ecosystems, and a variety of studies demonstrate that bats continue to harbor these contaminants in their tissues 20–30þ  years after the use of these pesticides was banned in sampling areas (Clawson and Clark 1989; Guillén etþ  al. 1994; Schmidt etþ  al. 2000; Sasse 2005). In some cases, persistence may reect the continued use of these pesticides in lower income nations, as may be the case for the migratory Tadarida brasiliensis (Thies and Thies 1997; Bennett and Thies 2007). Investigations in India (Senthilkumar etþ  al. 2001) and Benin (Stechert etþ  al. 2014) have detected levels or metabolites of organochlorines in bat samples indicative of continued recent use in these regions, especially to ght against malaria. Furthermore, pesticide standards vary between different countries, application often appears to occur non-selectively, and farmers with limited training (especially in developing countries, where agricultural expansion is greatest) are likely to be unaware of the multitude of negative nontargeted environmental impacts affecting human health and biodiversity (Tilman etþ  al. 2001; Yadav 2010). Despite the clear negative impacts of organochlorines on bats, the effects of agrochemical classes such as pyrethroids and neonicotinoids remain largely unknown (O’Shea and Johnston 2009; Quarles 2013; Bayat etþ  al. 2014), although

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168 K. Williams-Guillén et al.recent research demonstrates a negative impact on birds (Hallmann etþ  al. 2014). In North America, pesticide contamination has been implicated in bat mortality associated with the fungal pathogen causing white-nose syndrome (WNS), since pesticide load can lead to immunosuppression and endocrine disruption that could make bats more vulnerable to infection (Kannan etþ  al. 2010). “Back of the envelope” calculations suggest declines in bat populations attributed to WNS could translate into an additional 1320þ  metric tons of insects escaping predation each year (Quarles 2013). The trickle-down impacts on agricultural production could be substantial, although quantitative evidence is lacking. The effects of GM crops incorporating insecticidal traits have been investigated largely in the context of the provisioning of predation services (Federico etþ  al. 2008; Lopez-Hoffman etþ  al. 2014; see next section); however, declines in pest numbers associated with the use of these crops could result in population declines of insectivorous bats (LopezHoffman etþ  al. 2014).6.5þ Ecosystem Services Provided by Bats in Agricultural Systems 6.5.1þ Insectivorous Bats and Pest LimitationOf the potential ecosystem services provided by bats, their role in consuming insect pests has received the most attention within agricultural systems. Insectivorous bats have a global distribution and have long been identied as key suppressors of arthropod pests in agricultural systems (Kunz etþ  al. 2011). However, surprisingly little evidence exists quantifying the impact of their predation on arthropod populations, plant damage, or its economic value (Boyles etþ  al. 2013; Maas etþ  al. 2013). Several studies have characterized diets of insectivorous bats (reviewed by Kunz etþ  al. 2011), and the recent development of DNA-based methods for dietary analysis provides an unprecedented amount of detail on the composition of bat diets and allows for the identication of individual pest species. Although few studies have documented direct impacts of bat predation on agricultural pests, an increasing body of evidence documents pest consumption, impacts on arthropods, and estimates of direct economic impacts. We review 15 studies documenting the consumption of known or probable crop pests by insectivorous bats (Tableþ  6.3). The diets of temperate North American insectivores have received particular attention. Many bat species consume lepidopterans, and studies in North America demonstrate bat predation on devastating pests such as corn earworm (Helicoverpa zea) and fall armyworm (Spodoptera frugiperda) moths (Lee and McCracken 2005; McCracken etþ  al. 2012). Bat species across the world feed on folivorous beetles from a variety of damaging families

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169 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities … Tableþ  6.3þ Dietary investigations of insectivorous bat in agricultural areas documenting consumption of pest insect families or species Study region Source Bat species Crop Pest insects consumed Africa (South Africa) Taylor etþ  al. (2012, 2013a) Various species Macadamia nuts Hemiptera: Nezara viridula Africa (Swaziland) Bohmann etþ  al. (2011) Chaerephon pumilus, Mops condylurus Sugarcane Hemiptera: Aphidadae, Lygaeidae, Pentatomidae Lepidoptera: Eldana saccharina, Mythimna phaea Asia (Thailand) Leelapaibul etþ  al. (2005) Chaerephon plicatus Rice Hemiptera: Sogatella sp. Europe (Switzerland) Arlettaz and Perrin (1995, 1997, 2001) Myotis myotis, M. blythii Agricultural landscape with orchards, pasture Coleoptera: Melolontha sp. Latin America (Mexico) WilliamsGuillén (unpublished data) Various species Shade coffee Coleoptera: Hypothenemus hampeii, Rhabdopterus jansoni Orthoptera: Idiarthron subquadratum North America (Canada) Clare etþ  al. (2011) Myotis lucifugus Agricultural landscape Coleoptera: Phyllophaga spp., Amphimallon majale, Phyllobius oblongus; Curculionidae, Chrysomelidae Diptera: Delia antiqua Hemiptera: Aphididae Lepidoptera: Korscheltellus lupulina North America (Canada) Rambaldini and Brigham (2011) Antrozous pallidus Grapes Coleoptera: Curculionidae, Tenebrionidae Orthoptera: Acrididae North America (USA) Braun de Torrez (2014) Various species Pecan Lepidoptera: Acrobasis nuxvorella (continued)

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170 K. Williams-Guillén et al.and species, particularly weevils, leaf beetles, and scarab beetles. Bats may also be underappreciated predators of hemipteran pests, with many studies demonstrating consumption of leafhoppers, froghoppers, spittle bugs, and stink bugs. We emphasize that direct consumption alone is not sufcient to prove that bats are limiting insect pests: Damaging insects may comprise a small proportion of the diet, and nearly every study summarized in Tableþ  6.3 also demonstrated consumption of the predatory arthropods that comprise part of the assemblage of natural enemies. Such intraguild predation could counteract the pest-limiting effects of bat insectivory (Brashares etþ  al. 2010), although herbivores generally comprise the majority of diet by volume in investigations using fecal pellet dissections (Kunz etþ  al. 2011). That the relative abundance, diets, and movements of bats may track populations of agricultural pests (Lee and McCracken 2005; McCracken etþ  al. 2012; Taylor etþ  al. 2013b) suggests that many species are indeed preying heavily on herbivorous insects. This has been assessed in mouse-eared bats, Myotis spp., that track cyclic, massive local aggregations of cockchafers known since centuries for the damages they cause to fruit trees in Central Europe (Arlettaz 1996; Arlettaz etþ  al. 2001). During lactation, small bat species consume 75þ  % to over 100þ  % of their body weight each night (Kurta etþ  al. 1989; Kunz etþ  al. 1995, 2011), and a single maternity colony of 1 million Brazilian free-tailed bats is capable of consuming over 8 tons of insects per night (Kunz etþ  al. 2011). These numbers suggest the staggering potential for bat predation to limit pest insect Tableþ  6.3þ (continued) Study region Source Bat species Crop Pest insects consumed North America (USA) Lee and McCracken (2005) Tadarida brasiliensis Landscape with corn and cotton Coleoptera: Scarabaeidae Hemiptera: Cercopidae, Delphacidae, Pentatomidae Lepidoptera: Spodoptera frugiperda, Helicoverpa zea North America (USA) McCracken etþ  al. (2012) Tadarida brasiliensis Corn, cotton Lepidoptera: Helicoverpa zea North America (USA) Storm and Whitaker (2008) Eptesicus fuscus Agricultural landscape Coleoptera: Curculionidae Hemiptera: Cicadelidae North America (USA) Whitaker (1995) Eptesicus fuscus Agricultural landscape Coleoptera: Curculionidae, Scarabaeidae Hemiptera: Cicadellidae, Pentatomidae

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171 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities …populations and provide a valuable ecosystem service for agricultural production. Until recently, surprisingly little work had quantied the impact of bat predation on insect biomass (Maas etþ  al. 2015). Exclosure studies have long been a mainstay for studying the impacts of bird predation; however, it was widely assumed that such methods would not be suitable to measure the impact of bat insectivory, due to the misconception that all insect eating bats take highly mobile, ying prey. However, bats capable of gleaning insect prey from substrates exist throughout the world, and their impacts could be monitored via exclosure studies and disentangled from those of birds. This approach has been used fruitfully in the past ve years, demonstrating signicant increases in arthropod density when bats are absent, in agroecosystems (Williams-Guillén etþ  al. 2008; Maas etþ  al. 2013), reforestation (Morrison and Lindell 2012), and natural forests (Kalka etþ  al. 2008). In Mexican polycultural shade coffee, arthropod densities on coffee plants during the rainy season nearly doubled in the absence of bats, with marked increases in densities of hoppers, katydids, cockroaches, and beetles (Williams-Guillén etþ  al. 2008). However, no effects on plant damage were observed in that study, perhaps as a result of the short duration of the study or release of spiders and other arthropod predators. In Indonesian shade cacao, excluding bats resulted in a 29þ  % increase in arthropod numbers (Maas etþ  al. 2013). Although herbivory did not differ signicantly between cacao plantations with different levels of shade or proximities to primary habitats within the landscape, exclosure of bats resulted in a signicant decrease in yields, with the effects of bird and bat predation together valued at an astonishing US $730 per ha and year (bat predation was valued at US $520 per ha and year). However, the effects of bat predation on crop pests are not universal: An exclosure study in Costa Rican coffee found that excluding bats alone had virtually no effect on the density or damage caused to beans by the devastating coffee berry borer (Karp etþ  al. 2013). Exclosure studies are not suitable to measure the impact of high-ying insectivores, such as molossids. However, careful extrapolations taking into account bat feeding rates, population sizes, pest reproduction, and survivorship, and the costs of inputs allow for estimation of the economic impact of predation for other bats, particularly molossids forming large colonies. Cleveland etþ  al. (2006) estimate that Mexican free-tailed bats (T. brasiliensis) feeding on the cotton bollworm moth in Texas provide pest limitation services worth roughly US $183 per ha and year to cotton growers. Extending these estimates to agricultural areas throughout the USA suggests that bat predation could have a value of nearly US $23þ  billion annually (Boyles etþ  al. 2011). These benets hold for both conventional and transgenic cotton (Federico etþ  al. 2008), although the introduction of Bt cotton (a genetically modied organism whose tissues produce an insecticide derived from the bacterium Bacillus thuringiensis), coupled with reduced area in cotton cultivation, has led to a decline in the overall value of this pest limitation service (Lopez-Hoffman etþ  al. 2014).

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172 K. Williams-Guillén et al.Valuation of bat-mediated pest suppression is limited for staple crops and for sites outside the southern USA. In northern Mexico, the impact of T. brasiliensis predation on avoided agricultural costs across a variety of staple and commodity crops was estimated at a far more modest $19 per ha and year (Gándara Fierro etþ  al. 2006). In Thailand, the value of wrinkle-lipped bat (Tadarida plicata) predation on a major rice pest, the white-back planthopper (Sogatella furcifera), was estimated to have a monetary value of $1.2 million annually (Wanger etþ  al. 2014). This estimate results in a seemingly paltry $0.13 per ha and year value considered against Thailand’s 8.7þ  million ha (Redfern etþ  al. 2012) of rice paddies, but in this case an economic approach obscures the true value of the service: This single bat species prevents the loss of nearly 2900þ  metric tons of rice per year, enough to feed Thailand’s entire population of 66.8þ  million people for a week. Such investigations underscore the potentially grave consequences for human food security should global bat populations continue declining (Kunz etþ  al. 2011).6.5.2þ Nectarivorous Bats and Pollination ServicesPollination services to crops by bats are poorly documented. Bats are key pollinators of wild Agave and Musa spp. (Kunz etþ  al. 2011). Although these plants are propagated vegetatively under cultivation, bat pollination plays a critical role in sustaining genetic diversity in the wild relatives of these domestic species, a key aspect of maintaining future food security (Hopkins and Maxted 2011). Within the Americas, several bat pollinated cacti are commercially important fruit species (Kunz etþ  al. 2011). Several species of the hemiepiphytic cactus Hylocereus (pitahaya, dragonfruit) endemic to the Neotropics are now cultivated worldwide. In Mexico, visitation of Hylocereus undatus fruits by bats resulted in signicantly higher fruit set than did visitation by diurnal pollinators (Valiente-Banuet etþ  al. 2007). Although H. undatus is self-compatible, other species such as H. costaricensis (an important fruit crop in southern Mesoamerica) apparently rely on pollination by bats and sphingid moths (Weiss etþ  al. 1994; Le Bellec etþ  al. 2006). Nectarivorous bats, particularly the cave nectar bat (Eonycteris spelaea) feed on the owers of tree beans or petai (Parkia spp.) (Bumrungsri etþ  al. 2008a, b, 2013) and durian (Durio zibethinus) (Bumrungsri etþ  al. 2008b), pollinating these plants in the process. The economic value of this pollination has been estimated at over US $13þ  million annually in three provinces of Thailand (Petchmunee 2008).

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173 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities …6.6þ The Issue of Ecosystem Disservices of Bats to Agricultural ProductionUnfortunately, while the ecosystem services provided by bats are largely invisible, their disservices are obvious. In the Paleotropics, crop raiding by frugivorous pteropodids can cause substantial losses of commercial fruits (see Aziz etþ  al., Chap.þ  12). For example, in Indian vineyards, Cynopterus sphinx damages up to 90þ  % of the crop along peripheries of plantations and may cause revenue losses of up to US $590 per ha and year (Srinivasulu and Srinivasulu 2002). In the Neotropics, sanguivorous vampire bats can cause substantial economic damage: Estimates for 1968 placed losses at $47.5þ  million USD for over 512,000 rabiesrelated cattle deaths in Latin America (Arellano-Sota 1988). Harassment by vampire bats can put cattle off their feed, resulting in annual weight losses estimated at roughly 40þ  kg/head and milk production loss of 261þ  L/head (Schmidt and Badger 1979). These estimates fail to take into account the effects of vampire bats on the medium and small domestic animals (e.g., chickens, pigs, goats) that provide critical sources of animal protein for millions of smallholder farmers across the region. Not surprisingly, farmers with rst-hand experiences of economic losses engendered by bats are more likely to have negative attitudes or report a willingness to destroy bat roosts (Reid 2013). Failure to explicitly address the negative impacts of some bat species likely reduces the efcacy of conservation messages; meanwhile, practical measures to reduce these disservices could benet multiple bat species by reducing indiscriminate persecution. Different functional groups provide most of the ecosystem services (insectivores, nectarivores) and disservices (frugivores, sanguivores). However, local farmers may not distinguish between these groups. For example, farmers and agricultural technicians in Latin America often attempt to cull vampire bat populations by destroying bat roosts; unfortunately, the widespread belief that all bats are “vampiros” frequently results in the destruction of colonies of benecial bat species (Mayen 2003; Aguiar etþ  al. 2010). If local people perceive the ecosystem services of one bat group as offsetting the damages of another, then an ecosystem service approach could provide a framework for bat conservation more broadly. Unfortunately, the extent to which knowledge of ecosystem services changes attitudes toward bats in developing countries remains unknown.6.7þ DiscussionOur review suggests that in all biogeographic regions investigated, at least some bat species persist in and exploit agricultural areas. In many agricultural systems (e.g., tropical agroforestry or historical landscapes of Europe), bat assemblages

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174 K. Williams-Guillén et al.maintain richness and may even exceed abundances observed in unmanaged areas. Nevertheless, agricultural intensication has a generally negative effect on bats and thus presumably on the ecosystem services they provide. Our analysis did not address differences between bat taxa in their sensitivity to habitat change and intensication. However, evidence from speciose assemblages suggests that forest-adapted insectivorous species are particularly sensitive to habitat conversion (Medellin etþ  al. 2000; Faria and Baumgarten 2007; Williams-Guillén and Perfecto 2010), implying that in some regions, this valuable ecosystem service could be particularly vulnerable to loss in the face of habitat loss. Although few investigations have considered the scale of intensication, limited information suggests that less managed systems embedded in regions dominated by intensive agriculture may show depauperate bat faunas (Numa etþ  al. 2005; Faria etþ  al. 2007). Declines in bat populations in agricultural regions are concerning not only from the point of view of biodiversity conservation but also regarding human well-being and food security, especially in many tropical areas where smallholder farming systems are dominant. Ongoing losses of these generalist vertebrate predators could have major impacts on insect pest limitation for a wide variety of staple and commodity crops. However, the smallholder farmers in developing nations who most depend on the ecosystem services provided by bats (due to limited access to manufactured inputs or cultivation of bat pollinated crops) may have highly negative attitudes toward these mammals as a result of visible damages caused to crops and livestock (López del Toro etþ  al. 2009; Reid 2013), whereas benecial impacts on crop yield productivity and the value of biodiversity (i.e., increased ecosystem resilience) are often unknown or unappreciated (Williams-Guillén, unpublished data). These results suggest a pressing need to reassess common approaches to conservation and agricultural management in the Anthropocene.6.7.1þ Sparing, Sharing, and the Devaluation of Manufactured CapitalGiven the anticipated need to nearly double global food production in the twenty-rst century, a vigorous debate has emerged with respect to the most viable path to increase production without degrading ecosystem services or reducing biodiversity: land sparing, which posits that increased intensication and yields will reduce pressure to convert non-agricultural lands, versus land sharing, in which agricultural areas are less intensively farmed in order to increase associated biodiversity and habitat permeability (Fischer etþ  al. 2008). Given the vagility and critical role of bats in agricultural production, land sharing approaches might be preferable with respect to the provision of batdependent ecosystem services. Many sensitive bat ensembles and species (e.g.,

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175 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities …many forest-adapted and insectivorous species, e.g., from Phyllostomidae or Vespertilionidae) will require well-structured farmland, i.e., cultivated landscapes including patches of natural and seminatural features for their longterm existence. However, not only do many bat species thrive in diverse agricultural landscapes, but also their loss could affect the provision of pest suppression and pollination services and result in reduced crop productivity. Given the many disadvantages of chemical control of pests, managing agricultural landscapes to maximize the abundance and diversity of bats and other natural enemies must form a key aspect of sustainable agricultural production. However, the design and management of such systems to maximize bat diver sity, activity, and ecosystem services is largely unknown, although European conservationists are at the forefront with their strategies to promote biodiver sity-friendly farming. Chemical and mechanical inputs are not the only tools of agricultural intensication. Within recent decades, genetic modication of crops (e.g., Bt corn and cotton) has become increasingly prevalent (James 2011). In the short term, adoption of such varieties does reduce the need to rely on bats and other predators for pest limitation (Lopez-Hoffman etþ  al. 2014), resulting in a “devaluation” of the natural capital provided by bats, and undermines arguments for bat conservation that are based exclusively on provision of ecosystem services. However, as is the case with pesticides, insects are rapidly evolving resistance to Bt crops across the world, resulting in a rapid devaluation of manufactured capital (Lopez-Hoffman etþ  al. 2014). While the value of bats’ natural capital may uctuate, it likely devalues far less slowly: Bats and insects are engaged in an evolutionary arms race dating back millions of years (Conner and Corcoran 2012). Without bats to buffer the inevitable loss of efcacy of chemical inputs and GM crops, the technological advances that make agricultural intensication possible leave production vulner able to potentially catastrophic failures to limit pest damage.6.8þ Research Priorities 6.8.1þ Filling in Biogeographical Knowledge GapsAlthough the effects of habitat conversion and management have been well investigated in Europe and the Neotropics, the extent to which these processes may differ in other regions of the world remains unknown. We highlight a particular lack of knowledge from Africa and Asia; we did not nd any studies from East Asia, although we suspect information exists in the Chinese language literature. Understanding the types and magnitudes of ecosystem services provided by bats in a variety of agricultural systems and regions is particularly important.

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176 K. Williams-Guillén et al.6.8.2þ Linking Farm Management, Ecosystem Services, and Landscape-Level ProcessesThe effects of farm-level management on biodiversity and ecosystem services cannot be adequately considered without taking account of landscape-level processes (Tscharntke etþ  al. 2005; Vickery and Arlettaz 2012). Nevertheless, the extent to which localand landscape-level management interact to shape pest suppression or pollination services is largely uninvestigated. The effect of bats in limiting arthropod pests in agricultural areas is still poorly documented. However, the limited data that exist can demonstrate a vexing degree of divergence in results. For example, bats in Mexican shade coffee have substantial effects on herbivorous insects (Williams-Guillén etþ  al. 2008), while bats in Costa Rican shade coffee had no signicant effect on herbivores (Karp etþ  al. 2013). In Indonesian cacao agroforestry systems, insectivorous bats strongly contribute to the suppression of many differ ent pest insect groups and crop yield productivity across gradients of local shadetree management and forest proximity within the agricultural landscape (Maas etþ  al. 2013). In general, the study sites differ in landscape structure and land use, local farm history and management, habitat dynamics and conversion, intensity of farming practices, and vertebrate insectivore assemblage structure. Elucidating the factors of bat ecosystem service provision is key to managing agricultural areas to sustain bat populations and enhance food production (Maas etþ  al. 2015).6.8.3þ Pest Suppression in the Face of Climate Change, Pesticides, and GM CropsNot only will warming climates lead to shifts in the areas suitable for agricultural production, but it will also likely lead to range expansions of tropical pests, increases in pest numbers and damage, with a parallel risk of a drop in the efcacy of pest suppression by natural enemies that might be negatively affected by climate change (Thomson etþ  al. 2010; Bebber etþ  al. 2013). Such changes will make the ecosystem services provided by generalist predators like insectivorous bats more valuable than ever before. However, if agricultural adaptation to climate change relies on landscape-level intensication as a strategy, bats are likely to decline further, reducing their provision of pest suppression services. Despite the myriad negative effects of pesticides (i.e., affecting livelihoods, food security, environment, and health; reviewed by Yadav 2010), farmers across the world might turn to agrochemicals as a rst response to increases in pest damage (Wilson and Tisdell 2001), with the Old World’s rapid development of more environmentally friendly farming practices appearing as an exception in this general move. As reviewed in this chapter, older pesticide classes such as organochlorines

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177 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities …have particularly detrimental effects on bat populations. However, the degree to which newer pesticide classes affect bats is largely unknown. The neonicotinoids, once touted for their low toxicity, have now been linked to major declines in bees (Van der Sluijs etþ  al. 2013) and more recently in several species of passerines as a result of insect resource depletion (Hallmann etþ  al. 2014). The extent to which use of next-generation pesticides and GM crops is driving and interacting with bat declines and resultant increases in pest damage is a critical research area.6.8.4þ Quantifying Impact and Value Across Crops and BiomesAdditional valuation of bats’ ecosystem services could provide both guidance for bat management priorities in agricultural areas and compelling rationales for conservation. However, valuation efforts have focused almost exclusively on commodity crops quantied along the single dimension of monetary value. Most of the world’s smallholder farmers focus on staple crop cultivation and may not have the means to substitute the manufactured capital of pesticides and GM crops for bat predation. As Wanger etþ  al. (2014) demonstrate, valuation based on dollars of damage prevented misses many of the criteria most important to subsistence farmers seeking food security. There is an urgent need to better understand the importance of bat ecosystem services across a variety of crop types, regions, and management approaches. Research also highlights the importance of better quantifying the uctuations in bat service provision across years and seasons, in relation to population uctuations, reproductive phenology, and agricultural management (Lopez-Hoffman etþ  al. 2014; Wanger etþ  al. 2014; Maas etþ  al. 2015). This level of local, nuanced knowledge is key to managing pest suppression services in such a way that they are actively used as alternatives to agrochemical inputs and GM crops, and to contribute to more biodiversity-friendly and sustainable land-use practices (Tilman etþ  al. 2002; Maas etþ  al. 2015).6.8.5þ Changing Attitudes and Behaviors Toward Bats in the Developing WorldAlthough the conservation of tropical biodiversity is highly benecial to global society (Rands etþ  al. 2010), ultimately it is the attitudes and beliefs of farmers and other rural populations that will determine its fate (Brechin etþ  al. 2002; Tscharntke etþ  al. 2012). Throughout the world, bats are subject to misconceptions and poor public perceptions (see Kingston and Barlow, this volume Chap.þ  17). However, exposure to environmental education can signicantly

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178 K. Williams-Guillén et al.decrease negative attitudes toward bats (López del Toro etþ  al. 2009; Prokop etþ  al. 2009; Reid 2013). These results suggest that reducing bat disservices, conducting environmental education, and building local valuation of benecial bats could work in concert to improve conservation outcomes. As much as there is a critical need to manage agricultural landscapes to conserve bats, there is a parallel need to understand the local drivers of attitudes toward bats and to develop culturally appropriate, evidence-based interventions that encourage farmers to sustainably manage bat populations and other biodiversity associated with ecosystem services and ecosystem resilience.Acknowledgmentsþ We would like to thank the researchers who made their results available to us, particularly those who shared work in preparation or press. We are very grateful to Olivier Roth who carried out a thorough literature search on European farmland bats and Jean-Yves Humbert for feedback on the meta-analysis approach. Justin Boyle, Pia Lentin, Tigga Kingston, and an anonymous reviewer provided detailed feedback that greatly improved this manuscript. Finally, we thank the editors of the current volume for invitation to review this topic and for their patience. Open Access This chapter is distributed under the terms of the Creative Commons Attribution Noncommercial License, which permits any noncommercial use, distribution, and reproduction in any medium, provided the original author(s) and source are credited.ReferencesAguiar LMS, Brito D, Machado RB (2010) Do current vampire bat (Desmodus rotundus) population control practices pose a threat to Dekeysers nectar bats (Lonchophylla dekeyseri) longterm persistence in the Cerrado? Acta Chiropterol 12:275–282 Arellano-Sota C (1988) Vampire bat-transmitted rabies in cattle. Rev Infect Dis 10:707–709 Arlettaz R (1996) Feeding behaviour and foraging strategy of free-living mouse-eared bats, Myotis myotis and Myotis blythii. Anim Behav 51:1–11 Arlettaz R (1999) Habitat selection as a major resource partitioning mechanism between the two sympatric sibling bat species Myotis myotis and Myotis blythii. J Anim Ecol 68:460–471 Arlettaz R, Perrin N (1995) The trophic niches of sympatric sibling Myotis myotis and Myotis blythii: do mouse-eared bats select prey? Symp Zool Soc Lond vol 67, London: The Society 1960–1999, pp 361–376 Arlettaz R, Perrin N, Hausser J (1997) Trophic resource partitioning and competition between the two sibling bat species. J Anim Ecol 66:897–911 Arlettaz R, Christe P, Lugon A etþ  al (2001) Food availability dictates the timing of parturition in insectivorous mouse-eared bats. Oikos 95:105–111 Avila-Cabadilla LD, Stoner KE, Henry M, Alvarez-Añorve MY (2009) Composition, structure and diversity of phyllostomid bat assemblages in different successional stages of a tropical dry forest. For Ecol Manage 1–11 Aziz SA, Olival KJ, Bumrungsri S, Richards GC, Racey PA (2016) The conict between pteropodid bats and fruit growers: species, legislation and mitigation. In: Voight CC, Kingston T (eds) Bats in the Anthropocene: conservation of bats in a changing world. Springer International AG, Cham, pp. 377–420 Bayat S, Geiser F, Kristiansen P, Wilson SC (2014) Organic contaminants in bats: trends and new issues. Environ Int 63:40–52

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185 6þ Bats in the Anthropogenic Matrix: Challenges and Opportunities … Swanepoel RE, Racey PA, Shore RF, Speakman JR (1998) Energetic effects of sublethal exposure to lindane on pipistrelle bats (Pipistrellus pipistrellus). Environ Pollut 104:169–177 Taylor PJ, Nemudivhiso P, Mphethe V etþ  al (2012) Bats as biocontrol agents in macadamia orchards, Levubu Valley, Limpopo Province: results of a pilot project and future prospects. South Afr Macademia Grow Assoc Yearb 20:41–51 Taylor PJ, Bohmann K, Steyn JN etþ  al (2013a) Bats eat pest green vegetable stink bugs (Nezara viridula): diet analyses of seven insectivorous species of bats roosting and foraging in macadamia orchards. South Afr Macademia Grow Assoc Yearb 21:37–43 Taylor PJ, Monadjem A, Steyn JN (2013b) Seasonal patterns of habitat use by insectivorous bats in a subtropical African agro-ecosystem dominated by macadamia orchards. Afr J Ecol 51:552–561 Thies ML, Thies KM (1997) Organochlorine residues in bats from Eckert James River Cave, Texas. Bull Environ Contam Toxicol 58:673–680 Thomson LJ, Macfadyen S, Hoffmann AA (2010) Predicting the effects of climate change on natural enemies of agricultural pests. Biol Control 52:296–306 Tilman D (1999) Global environmental impacts of agricultural expansion: the need for sustainable and efcient practices. Proc Natl Acad Sci USA 96:5995–6000 Tilman D, Fargione J, Wolff B etþ  al (2001) Forecasting agriculturally driven global environmental change. Science 292:281–284 Tilman D, Cassman KG, Matson PA etþ  al (2002) Agricultural sustainability and intensive production practices. Nature 418:671–677 Tscharntke T, Klein AM, Kruess A etþ  al (2005) Landscape perspectives on agricultural intensication and biodiversity-ecosystem service management. Ecol Lett 8:857–874 Tscharntke T, Clough Y, Wanger TC etþ  al (2012) Global food security, biodiversity conservation and the future of agricultural intensication. Biol Conserv 151:53–59 Tuttle SR, Chambers CL, Theimer TC (2006) Potential effects of livestock water-trough modications on bats in northern Arizona. Wildl Soc B 34:602–608 Valiente-Banuet A, Santos Gally R, Arizmendi MC, Casas A (2007) Pollination biology of the hemiepiphytic cactus Hylocereus undatus in the Tehuacán Valley, Mexico. J Arid Environ 68:1–8 Van der Sluijs JP, Simon-Delso N, Goulson D etþ  al (2013) Neonicotinoids, bee disorders and the sustainability of pollinator services. Curr Opin Environ Sustain 5:293–305 Van Weerd M, Snelder DJ (2008) Human-altered tree-based habitats and their value in conserving bird and bat diversity in northeast Luzon, Philippines. In: Snelder DJ, Lasco RD (eds) Smallholder tree growing for rural development and environmental services. Springer, Dordrecht, pp 347–377 Vandermeer J, Perfecto I (2007) The agricultural matrix and a future paradigm for conservation. Conserv Biol 21:274–277 Vargas Espinoza A, Aguirre LF, Galarza MI, Gareca E (2008) Ensamble de murciélagos en sitios con diferente grado de perturbación en un bosque montano del Parque Nacional Carrasco, Bolivia. Mastozool Neotrop 15:297–308 Verboom B, Huitema H (1997) The importance of linear landscape elements for the pipistrelle Pipistrellus pipistrellus and the serotine bat Eptesicus serotinus. Landsc Ecol 12:117–125 Vickery J, Arlettaz R (2012) The importance of habitat heterogeneity at multiple scales for birds in European agricultural landscapes. In: FR J (ed) Birds and habitat: relationships in changing landscapes. Cambridge University Press, Cambridge, pp 177–204 Walsh AL, Harris S (1996a) Factors determining the abundance of vespertilionid bats in Britain: geographical, land class and local habitat relationships. J Appl Ecol 33:519–529 Walsh AL, Harris S (1996b) Foraging habitat preferences of vespertilionid bats in Britain. J Appl Ecol 33:508–518 Wanger TC, Darras K, Bumrungsri S etþ  al (2014) Bat pest control contributes to food security in Thailand. Biol Conserv 171:220–223 Weiss J, Nerd A, Mizrahi Y (1994) Flowering behavior and pollination requirements in climbing cacti with fruit crop potential. HortScience 29:1487–14982

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187Chapter 7Dark Matters: The Effects of Articial Lighting on BatsE.G. Rowse, D. Lewanzik, E.L. Stone, S. Harris and G. Jones© The Author(s) 2016 C.C. Voigt and T. Kingston (eds.), Bats in the Anthropocene: Conservation of Bats in a Changing World, DOIþ  10.1007/978-3-319-25220-9_7Abstractþ While articial lighting is a major component of global change, its biological impacts have only recently been recognised. Articial lighting attracts and repels animals in taxon-specic ways and affects physiological processes. Being nocturnal, bats are likely to be strongly affected by articial lighting. Moreover, many species of bats are insectivorous, and insects are also strongly inuenced by lighting. Lighting technologies are changing rapidly, with the use of light-emitting diode (LED) lamps increasing. Impacts on bats and their prey depend on the light spectra produced by street lights; ultraviolet (UV) wavelengths attract more insects and consequently insectivorous bats. Bat responses to lighting are species-specic and reect differences in ight morphology and performance; fast-ying aerial hawking species frequently feed around street lights, whereas relatively slowying bats that forage in more conned spaces are often light-averse. Both highpressure sodium and LED lights reduce commuting activity by clutter-tolerant bats of the genera Myotis and Rhinolophus, and these bats still avoided LED lights when dimmed. Light-induced reductions in the activity of frugivorous bats may affect ecosystem services by reducing dispersal of the seeds of pioneer plants and hence reforestation. Rapid changes in street lighting offer the potential to explore E.G. Rowseþ  · E.L. Stoneþ  · S. Harrisþ  · G. Jonesþ  ()þ  School of Life Sciences, University of Bristol, Bristol, UK e-mail: Gareth.Jones@bristol.ac.uk D. Lewanzikþ  Department of Evolutionary Ecology, Leibniz Institute for Zoo and Wildlife Research, Berlin, Germany D. Lewanzikþ  Animal Behaviour, Freie Universität Berlin, Berlin, Germany E.G. Rowse and D. Lewanzik: Equal contributors.

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188 E.G. Rowse et al.mitigation methods such as part-night lighting (PNL), dimming, directed lighting, and motion-sensitive lighting that may have benecial consequences for lightaverse bat species.7.1þ IntroductionAnthropogenic change is altering ecosystems at unprecedented rates and humans now dominate most ecosystems (Vitousek etþ  al. 1997; McDonald 2008). Urbanisation in particular has major impacts on bat activity and abundance (Jung and Threlfall 2016), and one aspect of global change that occurs predominately, but not exclusively, in urban areas is increased articial light at night. Almost a fth of the global land area was affected by light pollution in 2001 (Cinzano etþ  al. 2001). Although night-time brightness generally increased in Europe between 1995 and 2010, regional patterns are complex, with some localised declines (Bennie etþ  al. 2014). However, the biological impacts of light pollution have only recently been recognised (Longcore and Rich 2004). Being nocturnal, bats are likely to be affected by light pollution. In this chapter, we review the types of articial light that bats experience, describe how light pollution has become more widespread in recent years, show how technological changes may lead to signicant reductions in light pollution and describe some of the physiological consequences of light pollution that may be relevant to bats. We then discuss how articial lighting affects the insect prey of bats, and why some bats may benet from the growth in articial lighting, whereas others are affected detrimentally. After highlighting some aspects of bat vision, we describe the shift from observational to experimental studies of how bats respond to lighting. Finally, we identify some of the major knowledge gaps and suggest priorities for future research on the effects of articial lighting on bats.7.2þ Types of Articial LightThe electromagnetic spectrum encompasses radiation with wavelengths ranging from less than a nanometre (gamma rays) to a kilometre (radio waves) (Campbell 2011). While humans perceive wavelengths between 400 and 700þ  nm as ‘visible light’ (Purves and Lotto 2003), birds, sh and invertebrates can detect light in the ultraviolet (UV) range (10–400þ  nm). Recent work suggests that UV sensitivity may be widespread among mammals (Douglas and Jeffery 2014), and snakes and beetles can detect spectral emissions in the infrared range (700–1000þ  nm) (Schmitz and Bleckmann 1998; Land and Nilsson 2012). Articial lighting has inltrated all aspects of human life both indoors and outside (Gaston etþ  al. 2012). Here, we focus on street lighting because of its univer sal use and potential for ecological impacts (Gaston etþ  al. 2012). Different types of street light have distinct spectral signatures (Fig.þ  7.1); their primary emissions

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189 7þ Dark Matters: The Effects of Articial Lighting on Bats Fig.þ  7.1þ The spectral content of different light types varies considerably. The spectral composition of common lighting technologies is shown. From Gaston etþ  al. (2013)

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190 E.G. Rowse et al.depend on the type of reactive material or coating in the lamps (Buchanan 2006). Incandescent lamps, developed by Thomas Edison in 1880, mainly emit long wavelengths with a maximum intensity between 900 and 1050þ  nm (Elvidge etþ  al. 2010). Despite improvements such as the quartz halogen lamp, which uses an inert gas to preserve the tungsten lament, incandescent lamps are still relatively inefcient because their emissions are predominantly near the infrared spectrum and so largely invisible to humans (Elvidge etþ  al. 2010). Gas discharge lamps, developed by the mid-twentieth century, produce light by passing electric arcs through gas-lled bulbs (Elvidge etþ  al. 2010). These are further classied as low-pressure discharge and high-intensity discharge (HID) lamps (Elvidge etþ  al. 2010). Low-pressure discharge lamps include the compact uorescent lamp (CFL) and low-pressure sodium (LPS) lamps. Fluorescent lamps produce distinct emission peaks, which combine to emit a ‘white’ light (Royal Commission on Environmental Pollution 2009; Elvidge etþ  al. 2010), whereas LPS lamps have a narrow spectral signature, emitting monochromatic orange light with a peak intensity of 589þ  nm (Fig.þ  7.1) (Rydell 2006; Elvidge etþ  al. 2010). HID lamps include high-pressure mercury vapour (HPMV) lamps, which produce a bluish-white light, and high-pressure sodium (HPS) and metal halide lamps that have broader spectral emissions (Fig.þ  7.1) (Davies etþ  al. 2013). Emissions from HPMV lamps extend into the UV range (Rydell 2006; Elvidge etþ  al. 2010), whereas HPS lamps emit yellow-orange light and metal halide lamps ‘white’ light (Royal Commission on Environmental Pollution 2009; Davies etþ  al. 2013; Gaston etþ  al. 2013). The colour rendering index (CRI) compares how accurately a light source replicates the full range of colours of an object viewed in natural light on a scale of 0–100, where 100 is equivalent to natural light (Schubert and Kim 2005; Elvidge etþ  al. 2010; Davies etþ  al. 2013). HPS lamps typically have a CRI between 7 and 32, whereas metal halide lamps have a CRI ranging from 64 to 100, reecting their ability to render colour more suited for human vision (Elvidge etþ  al. 2010; Gaston etþ  al. 2012). Gas discharge lamps replaced incandescent lamps because of their energy efciency and improved longevity (Schubert and Kim 2005), and LPS (44þ  %) and HPS (41þ  %) lamps came to dominate street lighting in the UK (Royal Commission on Environmental Pollution 2009) and elsewhere. The luminous efcacy (LE) (amount of light produced per watt of electricity) of gas discharge lamps is ve times higher than incandescent lamps (Schubert and Kim 2005; Elvidge etþ  al. 2010). However, with pressure to reduce energy use and CO2 emissions, the lighting industry is now turning to light-emitting diodes (LEDs) (Elvidge etþ  al. 2010; Gaston etþ  al. 2012). LEDs have broad spectral signatures, typically 400–700þ  nm, with very few emissions in the UV range (Elvidge etþ  al. 2010). This is achieved mainly through the use of cerium-doped yttrium aluminium garnet (YAG:Ce) phosphors with a gallium nitride (GaN) which converts monochromatic blue to ‘white’ light. However, more recently LEDs are able to produce light by combining multiple monochromatic sources (red, green and blue), which allows for greater control over spectral emissions (Narendran etþ  al. 2004; Gaston etþ  al. 2012, 2013; Davies etþ  al. 2013). LED lamps have comparable CRI scores to metal

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191 7þ Dark Matters: The Effects of Articial Lighting on Batshalide lamps (65–100) (Elvidge etþ  al. 2010) but benet from lower running costs (Gaston etþ  al. 2012); low energy consumption (Elvidge etþ  al. 2010); controllability of spectral, temporal and intensity of emissions; reduced CO2 emissions (Hölker etþ  al. 2010a); and smart lighting capabilities that enable dimming in response to weather, trafc and lunar conditions (Bennie etþ  al. 2014).7.3þ The Growth of Light PollutionLight pollution is dened as the changing of natural light levels in nocturnal landscapes (nightscapes) through articial lighting sources (Falchi etþ  al. 2011; Kyba and Hölker 2013). Here, we focus on ecological light pollution, i.e. the direct ecological effects of light as opposed to astronomical light pollution, which describes the light that disrupts viewing of stars and other celestial matter (Longcore and Rich 2004). Ecological light pollution can be caused by glare (extreme contrasts between bright and dark areas), over-illumination, light clutter (unnecessary numbers of light sources), light trespass (unwanted light) and skyglow, where articial light is directed towards the sky, scattered by atmospheric molecules and reected back to earth (Royal Commission on Environmental Pollution 2009; Gaston etþ  al. 2012; Kyba and Hölker 2013). Articial lighting has increased as a result of urbanisation, population growth, economic development and advances in lighting technologies and provides numer ous economic, commercial, recreational and security benets (Riegel 1973; Hölker etþ  al. 2010a; Davies etþ  al. 2012). However, light pollution is now of global concern: the accelerated use of electric lighting, growing at 6þ  % per year, has escalated light pollution to threat status (Hölker etþ  al. 2010a, b). Satellite images suggest that 19þ  % of the global land surface surpassed the threshold for acceptable lighting levels (Cinzano etþ  al. 2001). However, satellites are unable to capture all illumination from light sources (Bennie etþ  al. 2014). While light pollution is currently more apparent in developed nations (Fig.þ  7.2), projected increases in industrial and urban growth suggest that light pollution will become more spatially heterogeneous both locally and regionally (Cinzano etþ  al. 2001; Gaston etþ  al. 2012; Hölker etþ  al. 2010b; Bennie etþ  al. 2014). In the UK, street lighting consumes approximately 114þ  Twh of energy annually (International Energy Agency 2006) and is growing at 3þ  % per annum (Royal Commission on Environmental Pollution 2009). The number of lighting installations is increasing (Gaston etþ  al. 2012), and the change in emissions due to increased use of broad spectrum technologies is also likely to affect light pollution as these sources emit higher levels of blue light. This scatters more into the atmosphere than green or red light, ultimately making a bigger contribution to skyglow (Benenson etþ  al. 2002; Falchi etþ  al. 2011; Kyba and Hölker 2013). The growth in light pollution will be further exacerbated because, as LEDs become cheaper, non-essential uses, such as advertising and architectural lighting, may increase (Schubert and Kim 2005).

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192 E.G. Rowse et al.7.4þ Projected Changes in TechnologyInternational lighting policies are prioritising energy-efcient technologies to reduce costs and CO2 emissions. The European Ecodesign Directive, for instance, encour ages moves from energy-intensive technologies such as incandescent, LPS and HPMV lamps (Hölker etþ  al. 2010a) to ‘whiter’ lighting with higher colour rendering capabilities (Gaston etþ  al. 2012). This may reduce CO2 emissions in the EU by as much as 42þ  Mt per year. A number of pilot studies in cities around the world (including Adelaide, Hong Kong, London, Mumbai, New York, Sydney and Toronto) have compared LED lamps against existing lighting technologies. After a three-year trial, the City of Sydney Council agreed to switch to LEDs on 6500 outdoor lights due to their reduced energy consumption, cost-effectiveness and improved illuminance (The Climate Group 2014). Future research will focus on increasing the efciencies of LEDs: the LE of a LED is 60–90þ  lm/W, compared to 80–120þ  lm/W for HPS lamps (California Lighting Technology Center 2010). More effective ways of producing light are also being investigated, such as combining multiple monochromatic sources as opposed to using phosphors: this will increase control over spectral emissions (Schubert and Kim 2005; Gaston etþ  al. 2012).7.5þ The Biological Effects of Light PollutionThe number of studies revealing negative consequences of articial night lighting on a multitude of both diurnal and nocturnal vertebrates and invertebrates is increasing rapidly (reviewed in Rich and Longcore 2006). Most negative effects Fig.þ  7.2þ Articial lighting is currently most widespread in the developed world. Global use of lighting at night in 2000. From NASA Earth Observatory/NOAA NGDC (2012)

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193 7þ Dark Matters: The Effects of Articial Lighting on Batsare due to the disruption of natural circadian and circannual cycles, which in turn can affect a whole range of species interactions, physiological processes and behaviours.7.5.1þ Impacts of Light Pollution on Intraand Inter-specic CompetitionLight-induced changes in circadian activity patterns can alter competition both within species (e.g. for mates) and between species (e.g. interference and exploitation competition). These are best documented for birds. For instance, early singing may be a signal of male quality in songbirds and increases the rate of extra-pair copulations, which are usually higher in older males. In territories affected by articial light, males of several songbird species start singing earlier at dawn and thereby gain access to about twice as many extra-pair mates (Kempenaers etþ  al. 2010; Nordt and Klenke 2013; Dominoni etþ  al. 2014). The effect of articial light on paternity gain is even stronger in yearlings than in adults, and so street lights might result in maladaptive mate choice of females by articially increasing the extra-pair success of yearlings (Kempenaers etþ  al. 2010). Whether similar maladaptive effects occur with nocturnal species is less clear. Articial light can affect niche partitioning by extending the activity of diur nal species, bringing them into inter-specic competition with nocturnal species (Longcore and Rich 2004; Rich and Longcore 2006). The scissor-tailed ycatcher Tyrannus forcatus, for example, will catch insects at street lights until at least 3þ  h after sunset (Frey 1993); this may increase exploitation and interference competition with insectivorous bats. Light pollution may also cause inter-specic competition between bats, with light-sensitive bat species excluded from illuminated resources exploited by light-tolerant species (Arlettaz etþ  al. 2000).7.5.2þ Effects of Articial Light on Physiological HomeostasisLight-induced changes in circadian rhythms may induce physiological aberrations. For instance, exposure of captive mice to light at night disrupts metabolic signals, leading to increased body mass and decreased glucose tolerance (Fonken etþ  al. 2010). Dim night-time light can also impair learning and memory, affect stress hormone levels, compromise immune function and cause depressive-like behaviour in rodents (Bedrosian etþ  al. 2011, 2013; Fonken etþ  al. 2012). In humans, depression, obesity and cancer risk relate to light pollution and associated disruptions of the circadian system (Fonken and Nelson 2011; Kronfeld-Schor and Einat 2012; Haim and Portnov 2013).

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194 E.G. Rowse et al.Light pollution can also result in a decoupling of seasonal behaviours and physiological adaptations from the optimal time of year. So, for instance, reproduction might be desynchronised from peak food availability; even very low light levels at night advance avian reproduction (Dominoni etþ  al. 2013) so that birds breed earlier close to street lights than in darker territories (Kempenaers etþ  al. 2010). Light-induced decoupling can even reverse an animal’s seasonal phenotype, so that it exhibits a long-day phenotype in winter and vice versa. In sheep, 1þ  h of light during the dark phase is enough to mimic a long-day during short-day conditions (Chemineau etþ  al. 1992). Also in primates, articial light at night can induce a long-day phenotype; these animals had higher core body temperatures, showed less locomotor activity during the nocturnal activity period and had fainter torpor bouts compared with short-day photoperiod acclimated animals (Le Tallec etþ  al. 2013). Voles that experienced light interference at night showed reduced winter acclimatisation of their thermoregulatory system to such a degree that they reduced heat production and died under winter eld conditions (Haim etþ  al. 2004, 2005). Thus, light pollution may have deleterious impacts on survival when animals expend too much energy during winter (Haim etþ  al. 2004): this may be relevant for hibernating bats.7.5.3þ Interference of Light Pollution with Nocturnal NavigationA well-documented effect of light pollution not mediated through circadian rhythms is the impact on movement decisions of visually orienting animals. Nesting attempts of female sea turtles are disrupted by articial light, and light attracts or confuses the hatchlings, rendering them more vulnerable to predation, exhaustion and dehydration (Salmon 2006; Perry etþ  al. 2008; Berry etþ  al. 2013). Birds migrating at night often approach bright lights instead of following their normal migration route, possibly because the light interferes with their magnetic compass (Poot etþ  al. 2008). Birds may also be trapped within the sphere of light, milling around illuminated objects until they die through collisions or exhaustion (Gauthreaux and Belser 2006; Montevecchi 2006; Spoelstra and Visser 2014). This may have relevance to bats, which also use magnetic compasses for navigation (Holland etþ  al. 2006). Similarly many insects, particularly moths (Lepidoptera), use articial lights rather than the moon for orientation and die of exhaustion when circling a lamp or following a collision with the hot cover. Articial light also provokes a ‘dazzling effect’: many insects become immobilised when approaching a lamp and rest on the ground or in vegetation, becoming easy prey (Eisenbeis 2006). Light pollution may even be a driver of an insect biodiversity crisis (Conrad etþ  al. 2006). The ‘vacuum cleaner’ effect, i.e. the long-distance attraction of light-susceptible species to lamps, removes large numbers of insects from the ecosystem, even

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195 7þ Dark Matters: The Effects of Articial Lighting on Batsresulting in local extinctions. This ight-to-light behaviour strongly depends on spectral output of the lighting: white HPMV lamps have a high UV proportion of their spectrum, and so four times as many moths are captured at HPMV lights compared to yellow/orange HPS lights (Eisenbeis 2006). Warm-white and cool-white LED lights induce less ight-to-light behaviour than HPS lights (Huemer etþ  al. 2010; Eisenbeis and Eick 2011), and the virtually monochromatic deep-orange LPS lights are least attractive to insects (Rydell 1992; Blake etþ  al. 1994; Eisenbeis 2006; Frank 2006). Several spiders, amphibians, reptiles, birds and bats focus their foraging on insects accumulated at street lights (Rich and Longcore 2006). For bats, this can also be advantageous because articial light disrupts the evasive behaviour of most nocturnal Lepidoptera, rendering them more vulnerable to bat attacks (Svensson and Rydell 1998; Acharya and Fenton 1999).7.6þ Bat VisionVision is important in the lives of many bats; see reviews in Suthers (1970), Altringham and Fenton (2003) and Eklöf (2003). A number of species rely on vision to a large extent (Altringham 2011). Since vision is important to both bats and their predators, we briey summarise some key recent ndings relevant to bats’ perception of articial lighting. Most pteropodids do not echolocate and use vision to locate fruit and owers. Some echolocating bats use vision to complement auditory information when hunting (Eklöf and Jones 2003) and, if vision and echolocation provide conicting information, visual information is used in preference (Orbach and Fenton 2010). Vision can also be more effective than echolocation over long distances (Boonman etþ  al. 2013), and the California leaf-nosed bat Macrotus californicus relies more on vision when hunting prey under low levels of illumination equivalent to a moonlit night (Bell 1985). Recent research on bat vision has focussed on the molecular evolution of lightsensitive pigments (Jones etþ  al. 2013). As for most nocturnal mammals, bat retinas are dominated by rods: they are highly sensitive under low light and confer monochromatic vision. The opsin DNA sequences of rhodopsin (the opsin in rods) were intact in 15 bat species (Zhao etþ  al. 2009a) and wavelengths of maximum absorbance were 497–501þ  nm. Colour vision in mammals results in part from opsins in the cones that are sensitive to short and medium wavelengths. Zhao etþ  al. (2009b) sequenced a short-wavelength sensitive opsin gene (Sws1) that is most sensitive to blue-violet wavelengths, and a medium-to-long-wavelength sensitive opsin gene (M/lws) in a range of bat species; maximum absorbance of red light wavelengths by the M/lws opsin was at 545–553þ  nm. Although many bats resemble diurnal mammals in having the potential for dichromatic vision, with both genes being intact, Sws-1 was pseudogenised in all the rhinolophid and hipposiderid bats studied and in some pteropodids, especially cave-roosting taxa. Immunohistochemistry suggests that

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196 E.G. Rowse et al.the primary visual cortex may not respond to stimulation by UV light in these taxa (Xuan etþ  al. 2012a), and behavioural responses to UV were also lacking (Xuan etþ  al. 2012b). The lesser Asiatic yellow bat Scotophilus kuhlii and Leschenault’s rousette Rousettus leschenaultii showed behavioural (Xuan etþ  al. 2012b) and immunohistochemical responses in the primary visual cortex (Xuan etþ  al. 2012a) to UV light at 365þ  nm. Two phyllostomid species (Pallas’s long-tongued bat Glossophaga soricina and Seba’s short-tailed bat Carollia perspicillata) possess signicant cone populations and express opsins that are sensitive to short and long wavelengths. The short-wavelength opsin is sensitive to UV and may be advantageous for the detections of UV-reecting owers (Winter etþ  al. 2003; Müller etþ  al. 2009). Other bat species with intact Sws1 genes may be UV sensitive, as ancestral reconstructions suggest UV sensitivity, with maximal sensitivity close to 360þ  nm (Zhao etþ  al. 2009b). Whether differences in UV sensitivity among bat taxa affect how species with intact and pseudogenised Sws1 genes respond to different types of lighting remains unknown. Nevertheless the ndings are of interest given that the wavelengths of maximum absorbance in bat opsins lie close to some of the peak emissions of wavelengths in a range of light types (Davies etþ  al. 2013). Moreover emerging LED lighting technologies do not emit UV wavelengths, whereas older technologies, especially HPMV lamps, emit wavelengths that extend into the UV range and so HPMV lights may have been particularly conspicuous to horseshoe bats.7.7þ Observational Studies on Bats at Street LightsBats have been observed foraging around lights ever since articial lighting became pervasive (Shields and Bildstein 1979; Belwood and Fullard 1984; Barak and Yom-Tov 1989; Acharya and Fenton 1999). Articial light attracts many positively phototactic insects (Rydell 1992; Eisenbeis 2006), and most insectivorous bats are probably opportunistic feeders. Thus, they quickly identify and exploit insect accumulations such as swarming termites (Gould 1978) and insect clusters at articial lights (Fenton and Morris 1976; Bell 1980; de Jong and Ahlén 1991). So some insectivorous bats probably prot from street lights because resource predictability and high insect densities increase foraging efciency (Rydell 1992, 2006). For instance, 18 of 25 Neotropical insectivorous bat species which could be detected by acoustic monitoring were observed foraging around street lights in a small settlement. While more species were recorded in mature forest, total bat activity was lowest in forest but highest around street lights (Jung and Kalko 2010). Bats prey on relatively large insects at street lights, mostly moths (Fenton and Morris 1976; Belwood and Fullard 1984; Acharya and Fenton 1992; Acharya 1995; Hickey etþ  al. 1996; Acharya and Fenton 1999; Jacobs 1999; Pavey 1999; Fullard 2001). While moths are the most numerous insects around articial lights (Huemer etþ  al. 2010; Eisenbeis and Eick 2011), their contribution to a bat’s diet can be much higher than expected from their relative abundance at street lights

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197 7 Dark Matters: The Effects of Articial Lighting on Bats (Belwood and Fullard 1984 ). This implies that bats focus on larger moths rather than smaller prey at street lights. Although moths were only captured in 36 % of attacks, northern bats Eptesicus nilssonii probably gain more than twice as much energy when feeding on moths at street lights than smaller dipterans in woodlands (Rydell 1992 ). Aggregations of large insects around lamps enable bats to reduce foraging time and hence energy costs while maximising energy returns (Acharya and Fenton 1999 ; Jung and Kalko 2010 ). Big brown bats Eptesicus fuscus , for instance, spend less than half as much time outside the roost where in habitats where they forage at street lights than where they do not use lamps for hunting (Geggie and Fenton 1985 ). Hence, foraging at lights might be benecial when a high foraging ef ciency compensates for the potentially higher predation risk. Bat activity and foraging efciency at street lights are mainly determined by the number and size of prey insects available, both of which are strongly affected by the spectral characteristics of the light (Blake et al. 1994 ). Thus, the type of light indirectly inuences bat activity. The lightÂ’s attractiveness for insects increases with its UV spectral content. Aerial-hunting long-legged myotis Myotis volans and California myotis M. californicus consistently preyed on insects clustered in the cone of experimental black (UV) lights in North America (Bell 1980 ). While black light is not used for street lighting, similar results are seen with street lights that produce UV emissions. Thus, bat density can be an order of magnitude higher in towns illuminated by HPMV compared with those illuminated by HPS lights and road sections illuminated by HPMV rather than deep-orange LPS lights (Rydell 1992 ). In Britain, mean bat activity, likely to be mainly common pipistrelles Pipistrellus pipistrellus , is usually equal to or lower along roads lit by LPS lights than in dark sections, whereas bat activity is higher under HPMV than LPS lights or sections with no light (Fig. 7.3 ; Blake et al. 1994 ). Fig. 7.3 Bat activity varies according to the type of articial lighting. Activity of pipistrelle Pipistrellu s spp. bats (mean and SD) along a 28 km stretch of road near Aberdeen, Scotland. a rural sections of the road without streetlamps, b village sections with sodium ( orange ) lamps and c a village with high-pressure mercury vapour lamps. From Rydell and Racey ( 1995 )

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198 E.G. Rowse et al.7.8þ Experimental Studies on Bats at Street LightsDrawing conclusions from observational studies can be difcult, especially since confounding factors other than the presence of street lights can affect bat activity. Experimental eld studies have demonstrated species-specic impacts of street lighting. Two 70þ  W HPS (DW Windsor Ltd, UK) lights, spaced and orientated to replicate street lights, were installed along preferred commuting routes of lesser horseshoe bats Rhinolophus hipposideros. The commuting activity of R. hipposideros (Fig.þ  7.4) and Myotis spp. was signicantly reduced, and the onset of commuting delayed, on lit nights (Stone etþ  al. 2009; Stone 2011). The following year the experiment was repeated on the same routes using white LED lights (Monaro LED, DW Windsor Ltd), at low (3.6þ  lux), medium (6.6þ  lux) and high (49.8þ  lux) light intensities. Activity of both R. hipposideros and Myotis spp. was signicantly reduced during all lit treatments, and for R. hipposideros, the effect size at 49.8þ  lux was the same as that under HPS illumination. So both HPS and LED light distur bance caused spatial avoidance of preferred commuting routes by R. hipposideros and Myotis spp. (Stone etþ  al. 2009), with no evidence of short-term habituation. Further work is needed to test for long-term habituation. In contrast, there was no signicant change in bat activity under HPS and LED light treatments for P. pipistrellus, and for bats in the genera Eptesicus and Nyctalus (Fig.þ  7.5). R. hipposideros and many other slow-ying species rely on linear habitat features for shelter from wind, rain and predators; acoustic orientation; and foraging 0 20 40 60 80 100 120 Control Noise Lit 1 Lit 2 Lit 3 Lit 4N oise Treatment M ean (a nt ilogge d) ba t p assesp = <0 .0 3 p = <0 .0 2 p = <0 .0 09 Fig.þ  7.4þ Light-averse bat species show reduced activity along commuting routes subjected to high-pressure sodium (HPS) lighting. Activity of lesser horseshoe bats Rhinolophus hipposideros (mean passes and SE) in relation to lighting treatment. Signicant within-subject differences with p values are shown. Treatments were control nights (no lighting treatment or generator), noise controls (HPS light units installed but switched off, generator running at night), 4 nights where lighting was switched on and powered by the generator (Lit 1 to Lit 4) and a nal noise control. From Stone etþ  al. (2009)

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199 7 Dark Matters: The Effects of Articial Lighting on Bats (Verboom and Spoelstra 1999 ; Verboom et al. 1999 ). Using suboptimal routes with reduced cover to avoid articial lighting may increase vulnerability to aerial pred ators and energetic costs due to increased exposure to wind and rain. So bats may have to travel further to reach foraging areas, reducing foraging time and increas ing energetic losses, with consequential negative effects on reproduction rates and tness. For example, juvenile growth rates were suppressed in the grey bat Fig. 7.5 Bats respond in different ways to LED lighting. Although the light-averse Rhinolophus hipposideros showed higher activity under more dimmed treatments compared with less dimmed ones, activity was still less than under unlit conditions. Myotis spp. showed negligible activity under all dimmed treatments. Geometric mean and condence limits for bat passes along treat ment hedges subjected to LED illumination at different light intensities are illustrated. Treatments were control nights (no lighting treatment or generator), noise controls (LED light units installed but switched off, generator running at night), 3 nights where illumination levels were modied (low light mean 3.6 lux; medium light mean 6.6 lux; and high light mean 49.8 lux), and a nal noise control. Bat passes were monitored on Anabat bat detectors and are shown for a Rhi nolophus hipposideros , b Myotis spp., c common pipistrelle Pipistrellus pipistrellus , d soprano pipistrelle Pipistrellus pygmaeus and e Nyctalus/Eptesicus . From Stone et al. ( 2012 )

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200 E.G. Rowse et al.Myotis grisescens with increased travel distance to foraging grounds (Tuttle 1976). Compensating for energetic losses by increasing foraging time may not be possible if, for instance, emergence and/or commuting is delayed by light pollution (Stone etþ  al. 2009). Such delays also increase the risk that bats will miss the dusk peak in insect abundance, reducing the quality of foraging time. Delayed emer gence could therefore affect the tness of both individuals and the roost as whole. Light disturbance along the commuting routes may isolate bats from their for aging grounds if the energetic costs of using alternative routes exceed the benets. The commuting costs for P. pipistrellus become prohibitive when foraging areas are more than 5þ  km from the roost (Speakman 1991). Since bats select roosts based on the quality of surrounding habitat features, including linear connectivity (Jenkins etþ  al. 1998; Oakeley and Jones 1998), maintaining optimal commuting routes is paramount. Whether tness, or likely proxies of tness, is affected by lighting needs further evaluation.7.9þ Winners and Losers: Light-Tolerant and Light-Averse BatsBats show variable responses to light pollution. Insectivorous bats that hunt in open spaces above the canopy (open-space foragers) or along vegetation edges such as forest edges, tree lines or hedgerows (edge foragers) are the species most tolerant of articial lighting. They have evolved traits advantageous for foraging in sparsely structured habitats (Norberg and Rayner 1987; Neuweiler 1989) and so are preadapted to foraging in urban habitats (Rydell 2006; Jung and Kalko 2010; Jung and Threlfall 2016). Open-space foragers, such as the noctule Nyctalus noctula, typically have long narrow wings with a high aspect ratio, often combined with a high wing loading (weight/wing area). They have to y fast to remain airborne and so use high-intensity, low-frequency narrowband echolocation calls that facilitate long-range detection of insects (Norberg and Rayner 1987; Rydell 2006; Kalko etþ  al. 2008). When foraging at street lights, open-space foragers typically y above the lamps, diving into the light cone to catch insects (Jung and Kalko 2010). Edge foragers generally use echolocation calls with a conspicuous narrowband component, but usually also include a frequency-modulated ‘broadband’ component during the search phase, which is advantageous for ranging when ying close to obstacles. They comprise relatively fast-ying species with above-average aspect ratio and wing loading (e.g. P. pipistrellus), and species with an average aspect ratio and wing loading (e.g. E. nilssonii). Edge foragers tend to be more manoeuvrable than open-space foragers (Norberg and Rayner 1987; Kalko etþ  al. 2008), and some can even conduct circuits inside the light cone when hunting insects at street lights (Jung and Kalko 2010). Though most edge foragers y with agility and speed (Norberg and Rayner 1987), they differ in their degree of synanthropism. While Kuhl’s pipistelle

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201 7þ Dark Matters: The Effects of Articial Lighting on BatsPipistrellus kuhlii is recorded almost exclusively at street lights in southern Switzerland, P. pipistrellus forage to a similar extent both at lights and at least 100þ  m from lights (Haffner and Stutz 1985). Even within a species, foraging activity at lamps can be highly variable depending on the quantity of insects available: Geggie and Fenton (1985) never observed E. fuscus foraging around street lights in an urban environment, whereas in rural habitats feeding activity was greater at lights than in areas without lights. In spring and autumn, when articial lights attract numerous insects in Sweden, E. nilssonii activity is about 20-fold higher in towns with street lighting than in non-illuminated towns, forest and farmland (de Jong and Ahlén 1991; Rydell 1991), with the bats ying back and forth above the street lights, regularly diving to within 1þ  m of the ground to catch insects. Although fast-ying species adapted to forage in open areas, particularly bats of the genera Eptesicus, Nyctalus and Pipistrellus, may benet from the increased foraging opportunities provided at lamps that attract high densities of insects, Stone etþ  al. (2009, 2012) found no signicant increases in bat activity for these ‘light-tolerant’ species during lit treatments. This could be due to two factors. First, HPS lights are less attractive to insects than white lights because their spectral content has less UV (Blake etþ  al. 1994); for example, HPS street lights attracted fewer insects than white lights in Germany (Eisenbeis and Eick 2011). Second, the experimental nature of the study may have affected the results, since bats may need time to nd and recognise newly installed lights as an attractive for aging source. Though a relatively high proportion of aerial insectivorous bats may forage in suburban habitats, bat activity and the number of bat species decrease signicantly towards highly urbanised areas. This is probably because both roosts and appropriate insect habitats are lacking, and those insects which are present might not aggregate at street lamps because the pervasive articial lighting in city centres causes a dilution effect, rendering the lights less attractive for bats (Gaisler etþ  al. 1998; Avila-Flores and Fenton 2005; Frank 2006; Rydell 2006; Jung and Kalko 2011; Jung and Threlfall 2016). In Panama, 18 of 25 insectivorous bat species frequently foraged around street lamps in a settlement bordering mature forest; the reduced vegetation cover in town constrained strictly forest-dwelling species from hunting at lamps (Jung and Kalko 2010). Yet, even some closely related and ecologically similar species may differ in their tolerance of urban habitats, and their potential to adapt to anthropologically altered habitats is best viewed from a species-specic perspective. As compared to open-space foragers, bats at the other end of the wing shape spectrum, such as many horseshoe bats (Rhinolophidae) with their low aspect ratio wings and a low wing loading, rarely forage near articial lights (Rydell 2006; Stone etþ  al. 2009, 2012). They are mostly forest-dwelling and their short broad wings facilitate the high manoeuvrability needed for hawking insects in a cluttered environment (Norberg and Rayner 1987). However, their morphology only allows slow ight speeds, which might render them more vulnerable to predators when ying in a sphere of light away from protective vegetation cover (Jones and Rydell 1994; Rydell etþ  al. 1996). Most forest-dwelling bat species emerge from

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202 E.G. Rowse et al.their roosts relatively late in the evening, presumably to minimise predation risk from diurnal birds of prey (Jones and Rydell 1994) and so may be ‘hard-wired’ to be light-averse. Furthermore, slow-hawking bats use echolocation calls that are adapted for short-range prey detection among clutter (Norberg and Rayner 1987), and so these may not be suitable for orientation in semi-open habitats where most street lights are positioned. Myotis spp. in Canada and Sweden and brown long-eared bats Plecotus auritus in Sweden were only recorded away from street lights (Furlonger etþ  al. 1987; Rydell 1992). In Australia, the chocolate wattled bat Chalinolobus morio avoided parks when lights were switched on (Scanlon and Petit 2008). Despite having street-lit areas in their home range, they were never utilised by greater horseshoe bats Rhinolophus ferrumequinum (Jones and Morton 1992; Jones etþ  al. 1995). Articial light reduced the foraging activity of pond bats Myotis dasycneme over rivers in the Netherlands (Kuijper etþ  al. 2008), and commuting activity of R. hipposideros and Myotis spp. was reduced under LED and HPS street lights (Stone etþ  al. 2009, 2012). It is likely that the Myotis spp. in Stone etþ  al.’s studies were Natterer’s bats Myotis nattereri (Stone 2011). M. nattereri emerges from roosts relatively late (Jones and Rydell 1994), at median light levels (3.5þ  lux, Swift 1997), lower than those recorded for R. hipposideros (Stone etþ  al. 2009). M. nattereri and R. hipposideros use different echolocation strategies (Parsons and Jones 2000) but have similar ight and foraging patterns. M. nattereri has broad wings, prefers foraging in woodlands and is slow-ying and manoeuvrable, often foraging close to vegetation to glean prey (Arlettaz 1996; Swift 1997). This suggests that light-dependent predation risk limits the ability of these bats to take advantage of illuminated areas. Nevertheless, one large-eared horseshoe bat Rhinolophus philippinensis was repeatedly observed traversing 200þ  m of open grassland to for age extensively around articial lights in Australia. The same lights were also used by eastern horseshoe bats Rhinolophus megaphyllus (Pavey 1999). Extinction risk is highest in bat species with low aspect ratios (Jones etþ  al. 2003; Sa and Kerth 2004), which are the species that show aversion to articial lighting. Thus, species that may suffer most from light pollution are likely to be already threatened taxa.7.10þ Effects of Light Pollution on Ecosystem Services Provided by BatsThe impacts of lighting go far beyond changing the physiology, behaviour and/ or distribution of individual species. Since congeners interact with each other as well as their prey and predators, light pollution is likely to have far-reaching consequences for the entire biome and the ecosystem services that bats provide. Insectivorous bats, for instance, signicantly reduce the number of insects that cause damage to ora and fauna (Ghanem and Voigt 2012). The value of

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203 7 Dark Matters: The Effects of Articial Lighting on Bats insectivorous bats to the US agricultural industry by reducing insect populations was estimated to be $23 billion/year (Boyles et al. 2011 ). Most studies to date have been on temperate-zone insectivorous bats. However, many tropical bats feed on nectar and fruits, thereby pollinating owers and dis persing seeds of several hundred species of plants (Ghanem and Voigt 2012 ). Consequently, frugivorous bats are key for succession and maintaining plant diver sity, especially in fragmented Neotropical landscapes (Medellin and Gaona 1999 ; Muscarella and Fleming 2007 ). However, very little is known about the impact of light pollution on this feeding guild. Southern long-nosed bats Leptonycteris yerbabuenae , a nectarand fruit-eating species, used areas of relatively low light intensity when commuting (Lowery et al. 2009 ) and Oprea et al. ( 2009 ) rarely captured frugivorous bats along roads, although some were present in municipal parks. However, neither study could disentangle the inuence of lighting from other factors related to urbanisation, such as altered vegetation cover or increased noise levels. Lewanzik and Voigt ( 2014 ) provided the rst experimental evidence for light avoidance by frugivorous bats. They found that SowellÂ’s short-tailed bat Carollia sowelli , a specialist on fruits of the genus Piper , harvested only about half as many fruits in a ight cage compartment lit by a sodium vapour street light than in a dark compartment, and free-ranging bats neglected ripe fruits that were experimentally illuminated (Fig. 7.6 ). Lewanzik and Voigt ( 2014 ) concluded that articial light might reduce nocturnal dispersal of pioneer plant seeds. Since Fig. 7.6 Articial lighting reduces and delays feeding behaviour on pepper plants by a frugivorous bat. a Percentage of harvested infructescences of Piper sancti felices among 14 marked plants harvested by SowellÂ’s short-tailed bats Carollia sowelli in nonilluminated conditions ( black ) and under conditions where plants were illuminated by a street lamp ( grey ) in the eld, b time after sunset when infructescences were harvested. From Lewanzik and Voigt ( 2014 )

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204 E.G. Rowse et al.bat-mediated seed intake is particularly important during the early stages of succession (Medellin and Gaona 1999; Muscarella and Fleming 2007), light pollution might slow down the reforestation of cleared rainforests (Lewanzik and Voigt 2014).7.11þ Knowledge Gaps, Future Challenges and Mitigation Strategies 7.11.1þ Knowledge GapsLight pollution has only recently been acknowledged as a threat to biodiversity (Hölker etþ  al. 2010b), and there are still many unknowns about the interactions between bat species and articial lighting sources (Hölker etþ  al. 2010a). Most studies have focused on specic ecological behaviours such as foraging (Rydell 1992; Blake etþ  al. 1994), predator–prey interactions, particularly with moths (Rydell etþ  al. 1995; Svensson and Rydell 1998), commuting routes (Stone etþ  al. 2009, 2012) and roost emergence (Downs etþ  al. 2003). No long-term studies have been carried out to determine whether any of these behavioural changes have tness consequences (Beier 2006; Stone etþ  al. 2012). The only indication of potential population-level responses has been shown in Hungary on Myotis species, where juveniles roosting in illuminated buildings had a lower body mass than their counterparts in unlit roosts (Boldogh etþ  al. 2007). However, this study did not establish whether a lower body mass in these juveniles reduced their survival rate after hibernation. It is particularly important to understand higher level responses for bat species because they have low fecundity rates, usually only producing one pup per year (Dietz etþ  al. 2009), and so populations are sensitive to sudden changes (Stone etþ  al. 2012). Further studies are needed to address the impact of articial lighting at the community level (Davies etþ  al. 2012). The current literature highlights that articial lighting causes species-specic responses (Rydell 1992; Stone etþ  al. 2009, 2012; Jung and Kalko 2010), which could cause light-tolerant species to exclude light-averse species (Polak etþ  al. 2011; Stone etþ  al. 2012). Such competitive interactions have been proposed as the driving force behind changes in bat populations in Switzerland, where decreases in photosensitive R. hipposideros have been linked to increases in light-tolerant P. pipistrellus (Stutz and Haffner 1984; Arlettaz etþ  al. 2000). It is believed that by avoiding street lights, R. hipposideros are foregoing protable prey sources exploited by P. pipistrellus (Arlettaz etþ  al. 1999, 2000). So far research has focussed largely on insectivorous bats in temperate zones. Further research in tropical ecosystems is needed. For example, the forested areas of South-east Asia contain a high diversity and abundance of horseshoe bat species that are likely to be negatively affected by light pollution, and the impact of light

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205 7þ Dark Matters: The Effects of Articial Lighting on Batspollution on pollination and seed dispersal in the tropics and subtropics needs fur ther investigation. Research on the impacts of different light spectra in emerging technologies on bat activity and reproduction will be valuable; this is currently being investigated in the Netherlands as part of a large-scale investigation exposing a wide range of taxa to white, red and green LED lighting (see http://www.lichtopnatuur.org). With the current plans to switch to broader spectrum lighting sources, it is impor tant to understand more about the spectral sensitivities of bats (Davies etþ  al. 2012, 2013), especially given the recent ndings on opsin genes highlighted above. Determining if there are spectral and intensity thresholds for different species would aid mitigation strategies and improve conservation initiatives (Stone etþ  al. 2012; Gaston etþ  al. 2013).7.11.2þ Mitigation StrategiesThe most effective approach to reduce the detrimental effects of articial lighting is to limit the growth of lighting by restricting unnecessary installations or removing them from areas already saturated with articial lighting sources. This has the greatest potential to reduce light pollution and minimise ecological effects (Gaston etþ  al. 2012). Turning off lights in areas commonly used by light-averse bats to for age, commute or roost during key times such as reproduction (Jones 2000) may be effective. Bats are faithful to maternity roosts due to the specic conditions they provide, and so conserving them is important for maintaining bat populations (Lewis 1995; Mann etþ  al. 2002). However, some photosensitive bats may be disrupted even if areas were only lit for a short period of time (Boldogh etþ  al. 2007), and switching off lighting may be challenged if it is perceived to jeopardise public safety (Lyytimäki and Rinne 2013). Reducing the duration of illumination through part-night lighting (PNL) schemes could also help limit the adverse effects of light on nocturnal animals (Gaston etþ  al. 2012). This has already been adopted by a number of local authorities in the UK, which switch off lights in specied areas between midnight and 05.30 to reduce CO2 emissions and save money (Lockwood 2011). Since April 2009, lights along sections of motorways have also been switched off between these hours (Royal Commission on Environmental Pollution 2009). While this may help to reduce light pollution, it is unlikely to have signicant ecological benets since the lights remain switched on in the early part of the night, when bats and other nocturnal species undertake key activities such as foraging and commuting (Gaston etþ  al. 2012). Intelligent lighting schemes, such as the use of motion sensors, have already been implemented in Portugal and may have more ecological benets. The lights remain switched off unless needed and so still provide all the perceived public safety benets (Royal Commission on Environmental Pollution 2009). However, these uctuations in lighting levels may also be damaging to bats (Longcore and Rich 2004).

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206 E.G. Rowse et al.It is also important to reduce the trespass of articial lighting to minimise the impact on bats. Newer technologies such as LEDs produce more directional light (Gaston etþ  al. 2012), preventing the horizontal or upward emissions which contribute most to light pollution (Falchi etþ  al. 2011). Effective luminaire design, installation of shielding xtures and correct column height can also help focus light and avoid wasteful emissions (Royal Commission on Environmental Pollution 2009). In Lombardia, Italy, for example, 75þ  % of light pollution was due to poorly designed luminaires; the other 25þ  % was unavoidable reection from road surfaces (Falchi 2011). Vegetation canopies such as hedgerows can also help decrease light trespass, which is crucial for many bat species that use linear features as commuting routes (Rydell 1992; Fure 2006). Diminishing trespass could create dark refuges, providing corridors for bats to forage in fragmented habitats (Longcore and Rich 2004; Stone etþ  al. 2012; Gaston etþ  al. 2012). Light intensity has a signicant effect on bat activity (Stone etþ  al. 2012) and delays roost emergence (Downs etþ  al. 2003). If bats delay foraging, they risk missing the peak abundance in insects that occurs shortly after dusk, so may not meet their energy requirements, which in turn could reduce tness (Jones and Rydell 1994; Stone etþ  al. 2012). In addition to implementing PNL, many local authorities are also dimming lights in specied areas (Gaston etþ  al. 2012). This relies on local authorities already having lights such as LEDs that have the necessary centralised management system (International Energy Agency 2006). These schemes are more environmentally friendly and cost-effective (Gaston etþ  al. 2012). However, dimming lights may not be benecial to all bat species; Daubenton’s bats Myotis daubentonii, for instance, only emerge from their roosts at very low light levels (less than 1þ  lux) (Fure 2006) and R. hipposideros and Myotis spp. avoid commuting routes illuminated to 3.6þ  lux (Stone etþ  al. 2012). Since illumination levels of street lights are usually between 10 and 60þ  lux (Gaston etþ  al. 2012), it may not be feasible to dim lighting to such low intensities without compromising public per ceptions of safety (Stone etþ  al. 2012; Lyytimäki and Rinne 2013).7.11.3þ Future ChallengesWith a number of changes to street lighting planned in the coming years, including dimming, PNL and modications to luminaire design to reduce light pollution, energy expenditure and greenhouse gas emissions, nightscapes could increase in heterogeneity, making it even more challenging to understand the impacts of articial lighting on biodiversity (Gaston etþ  al. 2012). This is further complicated because current metrics for measuring emissions from light sources omit key biological information (Longcore and Rich 2004; Gaston etþ  al. 2012). Illumination is measured in lux, which is dened as the brightness of a light according to human spectral sensitivities; spectral sensitivities of other taxa are often very different from ours (Peitsch etþ  al. 1992; Briscoe and Chittka 2001). In bats, for example, many species can detect wavelengths in

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207 7þ Dark Matters: The Effects of Articial Lighting on Batsthe UV range (Winter etþ  al. 2003; Wang etþ  al. 2004; Müller etþ  al. 2009). So HPS and LPS lamps could have the same intensity of light, e.g. 50þ  lux, but HPS lamps emit UV wavelengths, whereas LPS lamps do not, thereby affecting both bats and their insect prey in different ways (Longcore and Rich 2004). Since lux is commonly used as a metric by lighting engineers, designers and environmental regulators, migrating from this measure may thwart interdisciplinary communication (Longcore and Rich 2004). Another challenge is to nd more effective ways of quantifying the impact of articial lighting on bat species. Current methods use acoustic survey methods to quantify bat activity; this underestimates the activity of bats that use low-intensity echolocation calls (O’Farrell and Gannon 1999). Crucially, we also need to deter mine whether articial lighting has tness consequences (Stone etþ  al. 2012). A decrease in bat activity may have no relevance for tness if, for example, the bats are able to utilise equally suitable alternative sites nearby. A transdisciplinary approach needs to be adopted to minimise the impact of light on biodiversity, reduce CO2 emissions, increase energy efciency and reduce costs (Hölker etþ  al. 2010a; Gaston etþ  al. 2012). Scientists, policymakers and engineers need to work together to implement successful strategies (Stone etþ  al. 2012). Moreover, it is vital to nd ways to broaden awareness of light pollution and its ecological impacts. Since the public plays an integral part in agreeing mitigation schemes such as dimming lights, their support is pivotal in moving forward (Hölker etþ  al. 2010a).Acknowledgementsþ EGR, ELS, SH and GJ thank NERC for support. DL was supported by the Federal Ministry for Education and Research (BMBF) as part of the network project ‘Loss of the Night’. Open Access This chapter is distributed under the terms of the Creative Commons Attribution Noncommercial License, which permits any noncommercial use, distribution, and reproduction in any medium, provided the original author(s) and source are credited.ReferencesAcharya L (1995) Sex-biased predation on moths by insectivorous bats. Anim Behav 49:1461–1468 Acharya L, Fenton MB (1992) Echolocation behaviour of vespertilionid bats (Lasiurus cinereus and Lasiurus borealis) attacking airborne targets including arctiid moths. Can J Zool 70:1292–1298 Acharya L, Fenton MB (1999) Bat attacks and moth defensive behaviour around street lights. Can J Zool 77:27–33 Altringham JD (2011) Bats: from evolution to conservation, 2nd edn. Oxford University Press, Oxford Altringham JD, Fenton MB (2003) Sensory ecology and communication in the Chiroptera. In: Kunz TH, Fenton MB (eds) Bat ecology. University of Chicago Press, Chicago, pp 90–127

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213 7þ Dark Matters: The Effects of Articial Lighting on Bats Speakman JR (1991) Why do insectivorous bats in Britain not y in daylight more frequently? Funct Ecol 5:518–524 Spoelstra K, Visser ME (2014) The impact of articial light on avian ecology. In: Gill D, Brumm H (eds) Avian urban ecology: behavioural and physiological adaptations. Oxford University Press, Oxford, pp 21–28 Stone EL (2011) Bats and development: with a particular focus on the impacts of articial lighting. Dissertation, University of Bristol, Bristol Stone EL, Jones G, Harris S (2009) Street lighting disturbs commuting bats. Curr Biol 19:1123–1127 Stone EL, Jones G, Harris S (2012) Conserving energy at a cost to biodiversity? Impacts of LED lighting on bats. Glob Change Biol 18:2458–2465 Stutz H-PB, Haffner M (1984) Summer colonies of Vespertilio murinus Linnaeus, 1758 (Mammalia: Chiroptera) in Switzerland. Myotis 22:109–112 Suthers RA (1970) Vision, olfaction and taste. In: Wimsatt WA (ed) Biology of bats, vol II. Academic Press, New York, pp 265–281 Svensson AM, Rydell J (1998) Mercury vapour lamps interfere with the bat defence of tympanate moths (Operophtera spp.; Geometridae). Anim Behav 55:223–226 Swift SM (1997) Roosting and foraging behaviour of Natterer’s bats (Myotis nattereri) close to the northern border of their distribution. J Zool 242:375–384 The Climate Group (2014) Sydney LED trial: nal report. http://www.theclimategroup.org/_ assets/les/Sydney.pdf. Accessed 1 June 2014 Tuttle MD (1976) Population ecology of the gray bat (Myotis grisescens): factors inuencing growth and survival of newly volant young. Ecology 57:587–595 Verboom B, Spoelstra K (1999) Effects of food abundance and wind on the use of tree lines by an insectivorous bat, Pipistrellus pipistrellus. Can J Zool 77:1393–1404 Verboom B, Boonman AM, Limpens HJGA (1999) Acoustic perception of landscape elements by the pond bat (Myotis dasycneme). J Zool 248:59–66 Vitousek PM, Mooney HA, Lubchenco J etþ  al (1997) Human domination of earth’s ecosystems. Science 277:494–499 Wang D, Oakley T, Mower J etþ  al (2004) Molecular evolution of bat color vision genes. Mol Biol Evol 21:295–302 Winter Y, López J, von Helversen O (2003) Ultraviolet vision in a bat. Nature 425:612–614 Xuan F, Hu K, Zhu T etþ  al (2012a) Immunohisochemical evidence of cone-based ultraviolet vision in divergent bat species and implications for its evolution. Comp Biochem Physiol B 161:398–403 Xuan F, Hu K, Zhu T etþ  al (2012b) Behavioral evidence for cone-based ultraviolet vision in divergent bat species and implications for its evolution. Zoologia 29:109–114 Zhao H, Ru B, Teeling EC etþ  al (2009a) Rhodopsin molecular evolution in mammals inhabiting low light environments. PLoS ONE 4:e8326 Zhao H, Rossiter SJ, Teeling EC etþ  al (2009b) The evolution of color vision in nocturnal mammals. Proc Natl Acad Sci USA 106:8980–8985

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215Chapter 8Bats and Water: Anthropogenic Alterations Threaten Global Bat PopulationsCarmi Korine, Rick Adams, Danilo Russo, Marina Fisher-Phelps and David Jacobs© The Author(s) 2016 C.C. Voigt and T. Kingston (eds.), Bats in the Anthropocene: Conservation of Bats in a Changing World, DOIþ  10.1007/978-3-319-25220-9_8Abstractþ Natural bodies of open water in desert landscapes, such as springs and ephemeral pools, and the plant-life they support, are important resources for the survival of animals in hyper arid, arid and semi-arid (dryland) environments. Human-made articial water sources, i.e. waste-water treatment ponds, catchments and reservoirs, have become equally important for wildlife in those areas. Bodies of open water are used by bats either for drinking and/or as sites over which to forage for aquatic emergent insects. Due to the scarcity of available water for replenishing water losses during roosting and ight, open bodies of water of many shapes and sizes may well be a key resource inuencing the survival, C. Korineþ  ()þ  Mitrani Department of Desert Ecology, Jacob Blaustein Institute for Desert Research, BenGurion University of the Negev, Sede Boqer Campus, Midreshet Ben-Gurion, 84990 Beersheba, Israel e-mail: ckorine@bgu.ac.il R. Adamsþ  School of Biological Sciences, University of Northern Colorado, Greeley, CO 80639, USA e-mail: battings@yahoo.com D. Russoþ  Wildlife Research Unit, Laboratorio di Ecologia Applicata, Dipartimento di Agraria, Università degli Studi di Napoli Federico II, Via Università, 100, 80055 Portici, Napoli, Italy e-mail: danrusso@unina.it M. Fisher-Phelpsþ  Department of Biological Sciences, Texas Tech University, Lubbock, TX 79409, USA e-mail: m.sher-phelps@ttu.edu D. Jacobsþ  Department of Biological Sciences, University of Cape Town, 7701 Rondebosch, Republic of South Africa e-mail: David.Jacobs@uct.ac.za

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216 C. Korine et al.activity, resource use and the distribution of insectivorous bats. In this chapter, we review the current knowledge of bats living in semiand arid regions around the world and discuss the factors that inuence their richness, behaviour and activity around bodies of water. We further present how increased anthropogenic changes in hydrology and water availability may inuence the distribution of species of bats in desert environments and offer directions for future research on basic and applied aspects on bats and the water they use in these environments.8.1þ General IntroductionDryland environments which include hyper-arid, arid and semi-arid regions can be highly complex and diverse, despite being occasionally perceived as simple ecosystems supporting low species diversity (Ayal etþ  al. 2005). Aridity is described by ratio of precipitation to potential evapotranspiration ratio (P/ETP) (UNESCO 1979, Fig.þ  8.1) and dryland environments are ecosystems in which typically food availability is low, precipitation is limited and unpredictable, ambient temperature is high, humidity is low, and drinking water is scarce (Noy-Meir 1973). Consequently, there are large variations in primary production by plants that can strongly affect overall species diversity and interactions (Evenari etþ  al. 1971). Furthermore, the distribution, abundance and persistence of several desert-dwelling mammal species is affected by water availability, especially during dry summer months, when the challenges of minimizing energy use and water losses is greatest (Calder 1984; Morton etþ  al. 1995; Lovegrove 2000; Marom etþ  al. 2006). In desert environments, bats are an important component of the mammalian fauna. Carpenter (1969) asserted that, based on the number of species and abundance, bats are one of the most successful desert mammals, although they are outnumbered by rodents in the driest parts of the Sahara and the Namib Desert (Findley 1993). In the deserts of Israel, insectivorous bats are the most diverse Fig.þ  8.1þ The arid lands of the world (U.S. Geological Survey, science information services)

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217 8þ Bats and Water: Anthropogenic Alterations …group of mammals (Mendelssohn and Yom-Tov 1999), with 12 species recorded in the Negev Desert (Korine and Pinshow 2004) and 17 species in the Dead Sea area (Yom-Tov 1993). Benda etþ  al. (2008) recorded 14 species of insectivorous bats in Sinai, highlighting the diversity of these mammals in desert environments. The dryland regions of South America are the most species-rich habitats of the region and have the highest number of endemic species, even when compared to the tropical lowland Amazon forest (Mares 1992; Ojeda and Tabeni 2009; Sandoval and Barquez 2013). In the Yungas dry forest of Argentina, 55þ  % of the bat species may be endemics (Sandoval etþ  al. 2010). However, this area is severely under-protected and very little research has been conducted on the bat fauna (Mares 1992; Sandoval and Barquez 2013) In Mongolia, more than half of the bat species only occur in arid and semi-arid regions (Nyambayar etþ  al. 2010). Most bats, and in particular desert-dwelling bats, use open water sources for drinking water and/or as a foraging site (Vaughan etþ  al. 1996; Grindal etþ  al. 1999; Ciechanowski 2002; Campbell 2009, Fig.þ  8.2) with various studies reporting high levels of bat activity over open bodies of water (Rydell etþ  al. 1994; Walsh etþ  al. 1995; Young and Ford 2000; Mickeviciene and Mickevicius 2001; Ciechanowski 2002; Russo and Jones 2003; Korine and Pinshow 2004; Williams and Dickman 2004; Anderson etþ  al. 2006; Davie etþ  al. 2012; Monamy etþ  al. 2013), making even small springs, ephemeral pools and waterholes key foraging areas for insectivorous bats worldwide (Racey 1998). Water availability was even proposed as a mechanism for elevational patterns of species richness of bats in arid mountains (McCain 2007). In this chapter, we review our current knowledge of bats and water across regionally different semi-arid and dryland environments, and the factors that may Fig.þ  8.2þ A drinking event of the lesser horseshoe bat (Rhinolophus hipposideros) from a spring in the Dead Sea, Israel. Photo by Jens Rydell

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218 C. Korine et al.inuence their richness, behavior and activity around bodies of water. We discuss how anthropogenic development may inuence water availability and thus the distribution of species of bats in desert environments. Dryland environments are also predicted to be particularly sensitive to climate change, and we will discuss patterns by which climate disruption may further reduce water availability in arid regions. Finally, we offer directions for future research on basic and applied aspects on bats and the water they use in these environments.8.2þ Ecology of Bats and Water in Drylands Environments 8.2.1þ Water Sources Used by BatsPermanent and ephemeral pools are the central characteristic of many watersheds in dry, arid and semi-arid regions. Temporary pools have largely been ignored in management programs due to their relatively small size and apparent lack of benet for human use (Schwartz and Jenkins 2000). However, during spring and early summer, temporary pools may serve as important foraging grounds for aquatic and terrestrial species, some of which are regionally or locally rare and/or endemic (Nicolet etþ  al. 2004). Temporary pools in the Negev Desert had equivalent levels of species richness of bats and activity to permanent pools (Razgour etþ  al. 2010) and the activity of bats was reduced signicantly when bodies of open water were dried (Korine and Pinshow 2004), highlighting the importance of pools of all shapes and sizes to desert wildlife. In the arid regions of Mongolia, even suboptimal water sources such as small human-dug wells and salty lakes are used by bats and are an important resource for their continued survival (Nyambayar etþ  al. 2010). Conservation efforts should therefore focus on those sources offering only temporary water availability because although they support similar bat species richness and activity levels as permanent pools, they are less likely to be protected due to their ephemeral nature. That said, the importance of permanent pools can be underestimated if landscape availability of water is not considered through time. Geluso and Geluso (2012) analyzed 34þ  years of data in relation to capture rates gathered at a single drinking site, which was sampled once yearly, in the San Mateo Mountains of New Mexico. They found that in non-drought years capture success was significantly lower because bats were more dispersed across the landscape. However, in drought years, capture rates at the only available water source skyrocketed, thereby indicating the importance of open-water to local species of bats. Data gathered on foraging patterns of bats in Utah indicated a strong afnity by Myotis bats for riparian and edge habitats as compared to other surrounding areas (Rogers etþ  al. 2006). Similarly, Grindal etþ  al. (1999) showed that bat activity levels were signicantly greater in riparian versus upland areas in British Columbia and capture rates were higher for females than for males indicating that female bats may be more dependent on water-driven attributes of a particular area. Williams

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219 8þ Bats and Water: Anthropogenic Alterations …etþ  al. (2006) sampled across 22.5þ  km of the Muddy River oodplain in the Mojave Desert in Nevada, which was highly disturbed by long-standing ood control, livestock grazing, and the invasion of non-native plant species, and found that the riparian woodland habitat, which represents less than 1þ  % of the area, accounted for greater than 50þ  % of all bat activity. Areas of historically less disturbed mesquite bosque habitat maintained higher bat activity than more disturbed areas. Fortunately, restoration of habitats can increase local species richness. In Arizona, red bats (Lasiurus blossevillii), which had not been reported before, were captured along riparian-restoration areas of the lower Colorado River. The Arizona myotis (Myotis occultus), presumed extirpated, was also captured after restoration (Calvert 2012). In Africa, there is evidence that bat activity is higher around bodies of water than in adjacent areas. For example, in two regions in southern Africa, bat abundance was higher in riverine habitat than in adjacent, dryer savannah (Rautenbach etþ  al. 1996; Monadjem and Reside 2008). Differences in species richness and diversity between riverine and savannah habitats were not the same in the two regions. In the Kruger National Park, there was no difference in bat species richness or evenness between riverine habitat and savannah (Rautenbach etþ  al. 1996). In contrast, at another site in Swaziland, the riverine habitat had higher species richness and diversity (Monadjem and Reside 2008). In both regions, the two assemblages differed in the relative densities of the various species, with the savannah assemblages forming a subset of the riverine assemblages (Rautenbach etþ  al. 1996; Monadjem and Reside 2008). This reinforces the notion that bat assemblages in less mesic regions are extensions of bat assemblages in more mesic regions, but that not all species are inclined to make use of less mesic habitats when conditions are favorable. Some of them, particularly fruit eating bats (e.g. Epomophorus crypturus; Thomas and Fenton 1978) may be restricted to riverine habitats (Monadjem and Reside 2008). Australian studies also indicate high levels of bat activity around bodies of water (Lumsden and Bennett 1995; Williams and Dickman 2004; Grifths etþ  al. 2014a). Young and Ford (2000) found that species richness of bats, abundance, and capture success in the semi-arid Idalia National Park was greatest in areas adjacent to water, with 97þ  % of captures occurring at sites with water. Bats in Uluru National Park and the north-eastern edge of the Simpson Desert predominantly use oasis habitats that have permanent or temporary water sources even in years with higher than average annual rainfall (Coles 1993; Williams and Dickman 2004). Multiple species of Australian insectivorous bats have even been recorded ying, foraging, and perhaps drinking over hypersaline environments (Laegdsgaard etþ  al. 2004; Gonsalves etþ  al. 2012; Grifths etþ  al. 2014a, b). Pteropus species in New Guinea have been recorded drinking seawater (Iudica and Bonaccorso 2003) but the prevalence of bats drinking hypersaline water in arid environments is not understood, despite natural hypersaline water bodies being common in arid and semi-arid areas in Western Australia (Halse etþ  al. 2003; Timms 2005). In the arid regions of Mongolia, bats are mostly frequently found in association with water (Dolch etþ  al. 2007; Davie etþ  al. 2012).

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220 C. Korine et al.8.2.2þ Bodies of Water as a Drinking SourceWater sources that are used by bats are likely to be pools in streams, lakes, ponds, slow-owing streams and rivers and articial bodies of water with similar proper ties such as farm and urban dams (Jackrel and Matlack 2010; Sirami etþ  al. 2013), canals (e.g. Lisón and Calvo 2011), cattle troughs, swimming pools and settling ponds at waste water treatment facilities (Vaughan etþ  al. 1996; Abbott etþ  al. 2009; Naidoo etþ  al. 2013, 2014) and mines having natural seepage (Donato etþ  al. 2007; Grifths etþ  al. 2014a). Both the size and accessibility of the water source inuence whether a bat can drink from it. Bats drink water by swooping over a water source while lapping at the surface (Harvey etþ  al. 1999). Because bats drink on the wing, small and more maneuverable bats are able to drink from smaller pools, whereas less maneuverable bats need a large surface area of water to skim (Tuttle etþ  al. 2006). In the Negev Desert, Razgour etþ  al. (2010) found that both within and between pools, species richness of bats and activity signicantly increased with pond size. Furthermore, manipulations that decreased pond size led to a signicant reduction in species richness and activity and affected the bat assemblage composition. The size and situation of articial water sources similarly affect their use by bats. In the arid Texas Panhandle, USA, bats preferentially drank water from larger livestock tanks that were full and had only light vegetation around. They tended to avoid smaller, half-full tanks with denser vegetation around them (Jackrel and Matlack 2010). Although there are many anecdotal observations (Nickerson and O’Keefe 2013) of bats drinking from swimming pools there have been no formal studies of this. Despite the central nature of drinking and water availability for bats, there are a surprisingly small number of studies addressing this topic in Europe, even though many species do drink at open water sources regularly to rehydrate (e.g. Russo etþ  al. 2012). Some appear more sensitive than others to water deprivation because of their stricter dependence on water habitats. For instance, in water-denial experiments Daubenton’s bat, Myotis daubentonii, a species selectively dwelling in riparian habitat and above bodies of open water, has been found to undergo a greater body mass loss and to show signs of dehydration earlier than the brown long-eared bat, Plecotus auritus, a forest bat (Webb etþ  al. 1995). Drinking sites are also of chief importance for European bats outside the semiarid Mediterranean region. In the Bavarian Forest, Germany, oligotrophic, acidic ponds are used by over a dozen species of bats for drinking (Seibold etþ  al. 2013). Likewise, in the Italian Apennines, water cattle troughs built for traditional livestock breeding are frequently used to drink by over a dozen species of bats. Such small (often less than 15þ  þ  1.5þ  m) pools of water are locally of extreme importance (Russo etþ  al. 2010, 2012) for several threatened species (Fig.þ  8.3). These pools also concentrate insects, so bats occasionally forage there, but their importance for drinking is overwhelming (Russo etþ  al. 2012). The disappearance of traditional livestock breeding due to rural depopulation in many Apennine areas has led to the abandonment of the cattle troughs, implying an unstudied yet potentially high cost for bat

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221 8þ Bats and Water: Anthropogenic Alterations …populations (Fig.þ  8.3). In Italian forests, bats also drink from the small ephemeral pools which form following heavy rain and only last few days or weeks (D. Russo, pers. obs.). Eavesdropping on other drinking bats is likely to play an important role in locating such sites and this behaviour is typical of species with manoeuvrable ight such as the barbastelle bat, Barbastella barbastellus, and the greater horseshoe bat Rhinolophus ferrumequinum.8.2.3þ Bodies of Water as a Foraging HabitatThe tendency for higher insect abundance near water sources attracts bats to use water sources as foraging habitats. Furthermore, calm surface water provides a less cluttered acoustic signal return from the echolocation pulses (Mackey and Barclay 1989; Siemers etþ  al. 2001), and there is some evidence, at least for echolocating bats, that activity over calm pools of water is higher than that over fastowing rifes (von Frenckell and Barclay 1987). Bat activity in a transect from dry woodland savannah to riverine habitat in southern Africa was correlated with insect abundance—both bat activity and insect abundance were higher in riverine habitat (Rautenbach etþ  al. 1996) suggesting that bats were attracted to this habitat because of the feeding opportunities it provided. Drought is known to reduce the abundance of insects in temperate zones (Frampton etþ  al. 2000) and thus affect reproduction in insectivorous bats (Rhodes 2007). An eight year study by Bogan and Lytle (2011) on aquatic insects living in two study pools of a formerly perennial desert stream in the Whetstone Mountains of Arizona, USA, showed that complete water loss followed by intermittent ow caused a catastrophic regime shift in community structure that did not recover to the pre-drying conguration even after four years. Ledger etþ  al. (2011) found signicant reduction in and suppression of secondary productivity by drought that could have severe constraining effects on terrestrial vertebrate predator populations, and Love etþ  al. (2008) found similar effects in Arkansas, USA. Furthermore, Fig.þ  8.3þ Cattle troughs used by drinking bats in the Italian Apennines. Photo by Luca Cistrone

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222 C. Korine et al.desert bats in Arizona responded to articial-light-induced food patches (Fenton and Morris 1975) and one would presume this would be similar when small pools of water create swarms of high insect density. All of these data together suggest that small water sources with intermittent ow are vitally important as foraging sites to at least some insectivorous desert bat species. In Europe, three species of bats are aquatic habitat specialists: Daubenton’s bat, M. daubentonii, the long-ngered bat, Myotis capaccinii, and the pond bat, Myotis dasycneme. Besides taking insects in ight by aerial hawking, they typically for age very close to the water surface, from which prey is gaffed with their large feet or the inter-femoral membrane and transferred to the mouth while on the wing (Kalko and Schnitzler 1989; Siemers etþ  al. 2001). Chironomidae and Trichoptera are frequent prey items of these bats (e.g. Biscardi etþ  al. 2007; Krüger etþ  al. 2012). M. capaccinii may seize adult chironomids from the water surface as they emerge from pupal casings. Trawling bats mainly forage over calm water whose surface is free from ripples (Rydell etþ  al. 1999) as echoes from clutter interfere with prey detection (Siemers and Schnitzler 2004). On windy nights, M. capaccinii and M. daubentonii are less active (Russo and Jones 2003), presumably because wind reduces prey density and generates ripples on the water surface affecting target detection. In such circumstances, bats forage at sheltered sites where water is calmer (Lewis and Stephenson 1966; Lewis 1969). Several other species of bats frequent riparian habitats to forage and/or drink, especially the soprano pipistrelle, Pipistrellus pygmaeus (e.g. Nicholls and Racey 2006), Nathusius’ pipistrelle, Pipistrellus nathusii (Flaquer etþ  al. 2009), and other Pipistrellus spp. (Scott etþ  al. 2010), Schreiber’s bat Miniopterus schreiber sii (Serra-Cobo etþ  al. 2000) and noctules, Nyctalus spp. (Rachwald 1992; Racey 1998; Vaughan etþ  al. 1997). The stricter reliance on riparian habitats is one of the main ecological factors distinguishing P. pygmaeus from its sibling P. pipistrellus (but see Warren etþ  al. 2000) and allowing interspecic niche partitioning and thus coexistence (Oakeley and Jones 1998; Nicholls and Racey 2006; DavidsonWatts etþ  al. 2006; Sattler etþ  al. 2007). However, local factors such as elevation or landscape composition may inuence differences across species. At larger scales, the presence of main rivers and wetland areas are important as migratory paths and offer important stopover sites to migrating bats across Europe (Flaquer etþ  al. 2009). Rivers and riparian vegetation also constitute important linear landscape elements used for navigation by several European bats (Serra-Cobo etþ  al. 2000; Russo etþ  al. 2002). As might be expected given the above, the quality of foraging areas lacking water is inuenced by their distance to water. In Portugal, proximity to a drinking water source increased foraging habitat quality for Mehely’s horseshoe bat Rhinolophus mehelyi and M. schreibersii (Rainho and Palmeirim 2011). Similarly, a radio-tracking study of R. mehelyi in Spain showed that although this species hunted predominately in forest, the foraging areas were always within 500þ  m of a water source (Salsamendi etþ  al. 2012), possibly to allow for easy rehydration between foraging bouts or perhaps to take advantage of water-emergent forest insects. In historic landscape parks of England (Glendell and Vaughan 2002) as

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223 8þ Bats and Water: Anthropogenic Alterations …well as in German forests (Kusch and Idelberger 2005) the relative area of available water surface is an effective proxy for levels of bat activity. Australian bats have also been documented preferentially foraging around water sources. When compared to other habitat types in the Simpson Desert, more feedings buzzes were recorded around permanent and temporary water sources (Williams and Dickman 2004). Bats will also forage over hypersaline water bodies but more feeding buzzes are recorded over freshwater sites (Grifths etþ  al. 2014b). There is also evidence (e.g. Aldridge and Rautenbach 1987; Schoeman and Jacobs 2003, 2011; Naidoo etþ  al. 2011, 2013) that insects associated with freshwater habitats (e.g. Plecoptera, Ephemeroptera and Trichoptera) occur in the diet of southern African bats.8.2.4þ Water, Roosts and ReproductionThe propensity for female bats to choose roost sites that are relatively high in ambient temperature is thought to help them save metabolic energy by allowing for continued gestation of the young during torpor (Speakman etþ  al. 1991; Adams and Thibault 2006; Daniel etþ  al. 2010). The cost of such a choice in roost sites in arid regions, however, is the propensity for high-levels of evaporative water loss during the diurnal roosting cycle (Webb 1995) and this is further exacerbated when females are lactating (Kurta etþ  al. 1990). The only quantitative eld study to assess the need for drinking water by lactating female bats in drylands used PIT-tagged lactating and non-reproductive females from a maternity colony of fringed myotis (Myotis thysanodes) in Colorado, USA. Adams and Hayes (2008) found that lactating females visited to drink an average of seven times more per night than did non-breeding adult females. In addition, lactating females visited to drink consistently night after night regardless of daily relative humidity and temperatures, whereas non-reproductive females visited more when temperatures were high and relative humidity low (Adams and Hayes 2008). In addition, Adams (2010) synthesized 13þ  years of capture data from the same eld sites in Colorado, USA and found that summer mean precipitation had the highest correlation with reproductive frequency followed closely by mean stream discharge rates. Of these two, the latter showed the most abrupt effect on bat reproduction. When stream discharge rates were lower than 7þ  m/s, the frequency of reproductively active females captured plummeted, in some years by as much as 50þ  %. When female reproductive condition was plotted against mean stream discharge, the frequency of lactating females tracked the amount of available water, whereas the frequency of pregnant females was not correlated. This suggests that during drought years pregnant females may give birth, but do not have access to enough drinking water to support lactation. O’Shea etþ  al. (2010) using mark/recapture of big brown bats, Eptesicus fuscus, at maternity colonies in Ft. Collins, Colorado, USA found that rst year survival was lowest in bats born dur ing a drought year, although other factors were also at play.

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224 C. Korine et al.Several species of bats have been found to roost close to bodies of water to minimize the energy expenditure required to reach important drinking or foraging sites (Racey 1998; Korine etþ  al. 2013). The need to drink directly after emerging from the roosts may be the main factor determining the proximity of roosts to water, especially for maternity colonies (Racey 1998). M. daubentonii, whose foraging strictly depends on water habitat, often uses bridges over rivers, as well as buildings or cavity-bearing trees in the immediate surroundings of riparian biotopes (Racey 1998; Parsons and Jones 2003; Luan and Radil 2010; Encarnação 2012). Several other species, such as Natterer’s bat (Myotis nattereri), pipistrelles (Pipistrellus spp.) and brown long-eared bat, also tend to roost in landscapes comprising bodies of water that provide drinking and foraging opportunities (Racey 1998; Entwistle etþ  al. 1997; Oakeley and Jones 1998). Floodplain forests of central Europe host important reproductive colonies of tree-roosting noctule bat Nyctalus noctula (Görföl etþ  al. 2009). Myotis macropus, an Australian species, has a variable roosting behaviour but the primary force behind roost selection is proximity to waterways (Campbell 2009).8.3þ Threats to Water Sources Used by BatsIn drylands, where water resources are scarce, any loss of or degradation to open water source, such as a reduction in water quality, may create cascading affects that will be harmful to the wildlife that depends on it. When bats drink from a polluted source they ingest toxins directly and during foraging they indirectly ingest toxins that may have bio-accumulated within their insect prey. For example, if insect larvae feed on microorganisms in polluted water, they concentrate the pollutants in their bodies and when they metamorphose into adults these are consumed by bats. The effect of environmental chemical containments on bats was reviewed in 2001; most studies have occurred in Europe (~50þ  %) and North America (~34þ  %) mostly pertaining to organochlorine insecticides (58þ  %), metals (30þ  %), and polychlorinated biphenyls or PCBs (13þ  %) (Clark and Shore 2001). There are hardly any reports on the effect of polluted water on bat activity and richness in the drylands of North Africa, the Middle East and South America. Levels of bat activity in the Negev Desert were very high over wastewater treatment ponds (Korine and Pinshow 2004), however species richness was low and the majority of the activity was attributed to Kuhl’s pipistrelle (Pipistrellus kuhlii). Pilosof etþ  al. (2013) showed that sewage pollution in the Negev desert affected the immune response of Kuhl’s pipistrelle and Naidoo etþ  al. (2014) reported on DNA damage to bats that forage at wastewater treatments work.

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225 8þ Bats and Water: Anthropogenic Alterations …8.3.1þ Loss of Sources of WaterAn estimated two-thirds of Earth’s freshwater owing to oceans is obstructed by anthropogenic development (Nilsson and Berggren 2000), with approximately 75,000 dams in the USA alone and the majority of natural wetlands having been destroyed as well. Although not the scope of this chapter, it is important to mention that for bats, wetlands provide critical foraging habitat (Johnson etþ  al. 2008; Rainey etþ  al. 2006) with absolute area and connectivity of wetlands being impor tant components for foraging (Lookingbill etþ  al. 2010). Indeed, a recent report on total wetland loss in the USA from 2004–2009, showed a 25þ  % reduction from the previous reporting period. In addition, a total of 95,000 acres of saltwater wetlands and 265,720 acres of freshwater wetlands were lost (Dahl and Stedman 2013). The situation is exacerbated in the western USA, where livestock grazing has damaged at least 80þ  % of stream and riparian ecosystems (Belsky and Matzke 1999). The consequences for bats are illustrated by observed declines in bat activity as related to ow-reduction and drying along the San Pedro River in Arizona. Moreover, these declines corresponded to declines in insect availability at perennial sites and both bat activity and insect activity declined to imperceptible levels in areas where the river dried up (Hagen and Sabo 2012). European rivers, lakes and wetlands are among the most seriously altered ecosystems. Human impact has caused a major structural or chemical degradation of such ecosystems with fatal repercussions for their associated biota (e.g. Abel 1996). Alteration of European rivers has often led to the loss of channel features, oodplain connectivity and structure of bank vegetation. A threatened vespertilionid, M. capaccinii, selects foraging sites where water is less polluted and ripar ian vegetation better preserved. Along with the loss or disturbance of suitable cave roosts (Papadatou etþ  al. 2008), riparian habitat alteration poses the main threat to this bat (Biscardi etþ  al. 2007). Australian rivers have the highest variation in ow and ooding in the world (Williams 1981; Puckridge etþ  al. 1988). Anthropogenic activities such as extraction and diversion of water have had adverse impacts on rivers in the arid-zone of Australia (Walker 1985; Kingsford and Thomas 1995). High natural variation in water availability coupled with anthropogenic activities and climate change has the potential to catastrophically affect arid-species that depend on water availability (Roshier etþ  al. 2001; McKenzie etþ  al. 2007; Saunders etþ  al. 2013). A major concern associated with natural rivers and lakes in urban areas is that they may be polluted by runoff from roads or other sources. When bats drink from these sources, they ingest these pollutants directly or indirectly by feeding on aquatic-emergent insects. Sources of pollution of farm and golf course dams include feces from livestock and wild animal, nitrate and phosphate in fertilizers, metals, pathogens, sediments and pesticides. Unfortunately, little research has been done on the use of polluted urban water sources by bats and the probable health impacts on bats. The little evidence that does exist suggests that at least

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226 C. Korine et al.some species of bats may not avoid polluted bodies of water in arid areas (Pilosof etþ  al. 2013; Korine etþ  al. 2015). In Durban, South Africa bat abundance and species richness were higher over a polluted than over an unpolluted river and bat feeding activity (measured by feeding buzzes in the echolocation sequences) was also higher at the polluted river. There was, however, no difference in insect diver sity between the two rivers (Naidoo etþ  al. 2011) and, with the exception of a single species, Rufous mouse-eared bat, Myotis bocagii, proportions of prey items in the diets of bats did not correspond to their proportion in the insect fauna. M. bocagii fed predominantly on Diptera and this was also the most abundant insect in the insect light traps (Naidoo etþ  al. 2011).8.3.2þ MiningMining is a major anthropogenic source of environmental destruction and contamination globally. Toxins associated with extensive mining operations, in par ticular, gold mining is well documented. Cyanide used to extract gold from ore is commonly stored in open ponds, some of which are 200 acres in size. The actual numbers of bats, and other wildlife killed by drinking at these ponds is poorly understood and very difcult to track as many affected individuals either become submerged, or die from drinking contaminated water after leaving the site. Between 1980 and 1989, 34þ  % of all known mammals killed at cyanide ponds used for mining gold in California, Nevada, and Arizona were bats (Clark and Hothem 1991). Other heavy metals used in mining operations such as arsenic, cadmium, chromium, copper, lead, mercury, methyl mercury, nickel, and zinc have been found in bat carcasses. In Arizona, USA where at least 20þ  % of bat populations are in decline (King etþ  al. 2001), Mexican free-tailed bats (Tadarida brasiliensis) living 8þ  km from a major copper smelting mine had accumulated signicant levels of atmospheric mercury in their tissues (Petit 2007). In another study in Arizona, pallid bats (Antrozous pallidus), western pipistrelles (Parastrellus hesperus), and T. brasiliensis had elevated mercury levels in their liver and muscles that they most likely acquired via drinking from contaminated free-water sources (Reidinger 1972; see also Syaripuddin etþ  al. 2014). Besides contaminated ponds, natural water ows through thousands of abandoned mines in the western USA (used by bats for hibernaculum and maternity roosts) may be highly contaminated with heavy metals. For example, at Sheep Tank Mine overlooking the Colorado River in Arizona, barium, manganese and zinc were detected in soil samples at concentrations 10 times normal levels and E. fuscus captured at the site had higher concentrations of these elements than those collected from three other sites (King etþ  al. 2001). Other species included in the study had high arsenic levels as well as other contaminants (copper, lead, barium, manganese, and zinc) (King etþ  al. 2001). Bats and other terrestrial vertebrates can also be exposed to high levels of contaminants by ingesting aquatic emergent

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227 8þ Bats and Water: Anthropogenic Alterations …insects living in toxic streams and High levels of bioaccumulated cadmium and zinc are known to occur as far as 381þ  km downstream from the pollution source, whereas lead was found to be transferred from sediments to chironomids (midges) only as far as 40þ  km downstream (Cain etþ  al. 1992). Thus, large stretches of streams and rivers far from the point source of contamination pose threats to bats and other aquatic and terrestrial wildlife. Bats are also known to y and possibly forage/drink over gold mines in Australia (Donato and Smith 2007; Smith etþ  al. 2008). High bat activity was recorded over gold mine water bodies containing cyanide (Grifths etþ  al. 2014a). Grifths etþ  al. (2014b) suggested that elevated salt levels in water bodies at gold mines may decrease bat activity, foraging, and drinking. Bats, including the Vulnerable (IUCN 2014) ghost bat, Macroderma gigas, have also been recorded around an Australian copper mine in the Great Sandy Desert, although the mine’s effects on individuals or the population is unknown (Read 1998). Africa is rich in mineral resources and this makes mining activities relatively common so likely a serious threat to water quality and therefore to bats. A matter of grave concern is that no research has been done in Africa in this regard. This situation prevails despite evidence that mining activities do pollute surface water in Africa (Olade 1987; Naicker etþ  al. 2003).8.3.3þ AgricultureOrganochlorine pollution of streams and rivers, and other sources, is of major concern for bats (see Bayat etþ  al. 2014 for review). Experimental testing of organochlorine insecticides such as DDT on two species widely distributed throughout the USA, found that Myotis lucifugus was approximately twice more sensitive than were E. fuscus. Furthermore, juvenile E. fuscus were 1.5 times more sensitive than adults (Clark etþ  al. 1978). In addition, tests showed that individuals of T. brasiliensis poisoned with DDT survived for some time but later died of DDT poisoning mobilized from fat during active ight after being starved (Clark etþ  al. 1975). Laboratory studies also show that presence of organochlorine in tissues can accelerate the catabolism of fat, causing DDE-dosed bats (M. lucifugus) to lose weight faster than control bats (Clark and Stafford 1981). Although banned in the USA in 1972, signicant levels of DDT and DDE have been documented in tissues collected from bats foraging and drinking at the Rocky Mountain Arsenal Superfund Site (O’Shea etþ  al. 2001). High DDT concentrations are also found in M. lucifugus tissues in the Eastern United States (Kannan etþ  al. 2010). Furthermore, post-ban persistence of DDT in USA bats has been veried by sampling guano at roost sites (Clark etþ  al. 1982; Reidinger and Cockrum 1978; Bennett and Thies 2007). DDT has also been found in bat tissues in Australia despite being banned since 1987 (Mispagel etþ  al. 2004; Allinson etþ  al. 2006). DDT for agricultural use was essential banned worldwide in 2001, but recent work from Africa showed that DDT is probably still being used and accumulating in the tissues of multiple species of bats (Stechert etþ  al. 2014).

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228 C. Korine et al.The two most common agricultural pollutants are nitrogen and phosphorus and sources of these pollutants include inorganic and organic fertilizers, leguminous crops, septic tanks, farm and municipal waste water treatment facilities, and, in the case of phosphorous, run-off from groundwater discharge and atmospheric deposition. An excess of these nutrients is the leading cause of aquatic eutrophication (Shabalala etþ  al. 2013). Inorganic pollutants such as metals from agricultural and industrial run-off can also accumulate in these sites as well as in the tissues of insects using these bodies of water. Bats feedings on such insects are thus at risk of ingesting high levels of toxic metals such cadmium, chromium and nickel (see Naidoo etþ  al. 2013).8.3.4þ Waste WaterEuropean bats foraging in aquatic habitats are known to be largely exposed to toxic heavy metals which bioaccumulate in their insect food (Pikula etþ  al. 2010). Organic pollution of rivers is also known to affect bat foraging, but its effects are variable. A British study compared the differences in bat activity found respectively upstream and downstream from sewage outputs and showed that downstream activity of pipistrelle bats decreased whereas that of M. daubentonii increased relative to upstream sites (Vaughan etþ  al. 1996). The latter species is thought to benet from the higher downstream abundance of pollution-tolerant prey such as chironomids. However, an Irish study obtained opposite results, with P. pygmaeus being more common downstream of sewage efuent discharges than M. daubentonii (Abbott etþ  al. 2009). Park and Cristinacce (2006) compared the effects of two types of sewage treatment works for foraging bats: those with percolating lter beds, often hosting many insects potentially important for bats, and the “activated sludge” system—gradually replacing the former—in which sewage and bacteria-laden sludge are mixed and agitated so that they prove inhospitable for the invertebrate fauna. The study showed that both insect biomass and bat activity were higher at percolating lter beds and that bat activity there was comparable to that recorded at nearby natural foraging habitats. However, bats may run serious risks when foraging at such sewage treatment works: endocrine disrupting chemicals, which may alter the endocrine functions in exposed animals, have been found to concentrate in bat insect prey at percolating lter beds, with potentially harmful effects on foraging bats (Park etþ  al. 2009). There has been very little research in Africa on the concentration of pollutants in tissues of bats and no work on the long and short term effects of these pollutants on the health of bats. There is some evidence of the presence of the toxic metals cadmium, chromium and nickel in tissues of African bats foraging at sites downstream of waste water treatment plants (Naidoo etþ  al. 2013). Furthermore, bats for aging over waste water treatment facilities display increased haematocrit and DNA damage and decreased antioxidant capacity in muscle tissue compared to bats that forage over unpolluted sites. Although these effects were not lethal they may result

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229 8þ Bats and Water: Anthropogenic Alterations …in long-term negative effects on the health of bats (Naidoo etþ  al. 2014). These metals were probably ingested by bats via their insect prey. There is evidence that aerial insects developing in sewage sludge and waste water at sewage treatment plants can accumulate pollutants that could disrupt endocrine functioning (Park etþ  al. 2009). However, a similar study on the activity of the insectivorous bat, the banana bat, Neoromicia nana, at three urban rivers systems above and downstream of where sewage efuent enters these rivers revealed that the relative abundance and feeding activity of N. nana were higher at polluted sites downstream of where sewage entered the system than at the unpolluted sites upstream (Naidoo etþ  al. 2013). In this case the bats may have been attracted by the higher abundance of dipterans over the polluted sites. Diptera were the dominant prey items in both the insect fauna at the polluted sites and in the diets of the bats (Naidoo etþ  al. 2013). This also appeared to be the case for M. bocagii which also fed predominantly and opportunistically on Diptera (Naidoo etþ  al. 2011). The response by bats to rivers affected by waste water treatment efuent may vary both between and within species. In North America (Kalcounis-Rueppell etþ  al. 2007) and England (Vaughan etþ  al. 1996), some species were more active upstream from where waste water efuent entered the rivers while others were more active downstream. It appears that these differences arise from the differ ential effects of euthrophication on insect prey as well as on the responses of bats. Some species take advantage of eutrophication that causes an increase in the abundance of their preferred prey, and other species which apparently do not feed on insects that are affected by eutrophication, prefer to forage in less polluted habitats. Furthermore, these differences may also result from differences in the foraging behavior of the same species at different sites. For example, N. nana fed opportunistically on the small abundant dipterans at wastewater polluted sites, but at unpolluted river sites fed selectively on insects from other orders (Naidoo etþ  al. 2013). Another major anthropogenic compound found in open bodies of water in the USA is polychlorinated biphenyl or PCB, a common industrial waste product that was banned by the United States in 1979 and the United Nations in 2001. PCB poisoning in pregnant M. lucifugus led to stillborn young (Clark and Krynitsky 1978). Aquatic-emergent insects are key exporters of contaminants to terrestrial ecosystems (Menzie 1980; Runck 2007) and data show signicant lateral transfers of PCBs to terrestrial riparian predators such as spiders, reptiles and amphibians (Walters etþ  al. 2008). High concentrations of PCB’s have been found in fat tissues of M. lucifugus in New York and Kentucky (Kannan etþ  al. 2010). Along the fresh water tidal river, the Biesbosch, in the Netherlands, direct transfer from river sediments to chironomids to pond bats occurred in concentrations known to cause negative reproductive effects in mink (Reinhold etþ  al. 1999). Frick etþ  al. (2007) investigated the effects of an accidental chemical spill (metam sodium) on Yuma myotis (Myotis yumanensis) in California and found reduced female juvenile sur vival, but not adult female survival. The spill-affected population declined signicantly during the rst years of the study. Although the population increased in year four, this also coincided with an end to an extensive regional drought. Controlled

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230 C. Korine et al.experimental exposure to Lindane (an organochlorine used in wood preservatives) at sublethal levels in P. pipstrellus increased 24þ  h metabolic rates of a 7.3þ  g individual by 15þ  % and in a 6.3þ  g individual by 23þ  %, thereby posing a signicant threat to survivorship of free-living individuals (Swanepoel etþ  al. 1999) and showing that sub-lethal exposure can affect energetic balance.8.4þ Mitigation and RestorationBoth the availability and distribution of water in drylands have been drastically altered by natural processes such as decline in annual precipitation, and by anthropogenic developments such as irrigation for agriculture, over exploitation of groundwater and human-induced climate changes.8.4.1þ Restoration of Water Sources and Related HabitatsMost wetlands have been altered globally due to anthropogenic disruption, pollution, and outright destruction. In some, but too few, places, humans have begun to restore some of those wetlands. For example, in the USA, the Sierra Nevada Conservancy is working in cooperation with State Parks, the Department of Toxic Substances, California State University Chico and others, to identify mercury sources and potential remediation strategies for an abandoned hydraulic mine discharging sediment and heavy metals into the Yuba River and removing mercury from dredged sediment that have accumulated in the Combie Reservoir. In California, restoration of the Cosumnes River oodplain re-established bat activity that broadly corresponded with ooding and an increase in aquatic emer gent insects (Rainey etþ  al. 2006). Furthermore restoration of riparian habitat, frequently damaged by cattle as well as other anthropogenic uses, and wetlands commonly destroyed by human development, is essential and is occurring in some areas, but well below necessary levels for bat conservation (Goodwin etþ  al. 1997). Despite some of the negative effects highlighted in the previous section concerning waste water efuent, wastewater reclamation is an important process especially in areas where water is scare (Anderson etþ  al. 2001). Wastewater can be used to construct articial wetlands that provide habitat for wildlife if the water is properly treated (Greenway and Simpson 1996; Fujioka etþ  al. 1999; Greenway and Woolley 1999; Greenway 2005). Some studies have found that increased nutrient loads, such as those caused by wastewater efuent may have a positive effect on insect and bat abundance both in US and European streams (Kokurewicz 1995; Vaughan etþ  al. 1996; Abbott etþ  al. 2009). One US study found that bat activity and foraging levels were the same up-stream and down-stream of wastewater discharge but community structure was altered, with the riparian-specialist Perimyotis subavus being more abundant (Kalcounis-Rueppell etþ  al. 2007).

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231 8þ Bats and Water: Anthropogenic Alterations …8.4.2þ Articial Water SourcesOne way to overcome the diminishing of natural water sources in many drylands is the development of articial catchments which are widely used for wildlife management (Krausman etþ  al. 2006). There has long been controversy regarding the effects of catchments on local wildlife, in which critics argue that these developments do not yield expected benets to game species and may have opposing impacts such as predation (O’Brien etþ  al. 2006). Small articial ponds may be of utmost importance for wildlife (Russo etþ  al. 2012). The large-scale expansion of intensive agriculture in semiarid Mediterranean climates has often been sustained by hydraulic engineering works, to cope with the scarcity of natural irrigation water. In southeastern Spain, Lisón and Calvo (2011) studied the effects on bats of a water transfer channel and a related network of irrigation ponds in a mixed landscape of traditional and intensive agricultural landscape. In general, articial bodies of water had a positive effect on bat activity, but this mainly regarded common, generalist species (P. pipistrellus and P. pygmaeus) most likely because of the absence of foraging habitats suitable for more specialized species (those bearing a higher conservation value) such as riparian vegetation. In Catalonia, rice paddies sustain high bat activity, providing large amounts of insect prey. However, roost availability was the main limiting factor and installing bat boxes represents a valuable strategy to increase bat populations (Flaquer etþ  al. 2006). In the arid Ikh Nart Nature Reserve in Mongolia, signicantly more bats were caught at natural springs relative to human-made wells and no bats were captured at sites without water (Davie etþ  al. 2012). This suggests that at least for this area, replacing lost natural water sources with articial ones may not be as effective for preserving bat populations as conserving natural water sources. Paradoxically, the creation of large water reservoirs may prove harmful to the entire bat community. Rebelo and Rainho (2009) looked at the effects on bats of the largest reservoir in Europe, created by construction in 2001 of the Alqueva dam, in Alentejo, Southern Portugal. The project led to the deforestation and submersion of an area of ca. 250þ  km2. Consequently, bat populations were affected by the sudden disappearance of ca. 200þ  km of riparian habitat, together with largescale roost loss and the replacement of important habitat with a vast homogeneous one which was not used by foraging bats. Noticeably, bat activity showed a strong decline in the submerged areas but increased in the surrounding unaffected habitat. The expansion of Mediterranean species into surrounding arid wildlife communities may have a negative impact on local populations such as competition for the use of pools for drinking and foraging. Nine of the 12 Negev species of bats (Korine and Pinshow 2004) are associated with arid areas, and the Kuhl’s pipistrelle, the European free-tailed bat (Tadrida teniotis), and the rare lesser horseshoe bat (Rhinolophus hipposideros)—are Mediterranean species that have expanded their distribution into the Negev in the twentieth century (Yom-Tov and Mendelssohn 1988). The most common bat in some desert habitats and in

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232 C. Korine et al.particular at articial water sites in the Negev is Kuhl’s pipistrelle (Korine and Pinshow 2004). The expanded distribution is probably linked to human settlements and in particular to articial bodies of water since non-desert species of bats must drink on a daily basis and drink more frequently compared with desert-dwelling bats (Razgour 2010). Kuhl’s pipistrelle competes for the use of pools for drinking and foraging, resulting in temporal and spatial partitioning between local desert bat species (Razgour etþ  al. 2011). The documented competition between Kuhl’s pipistrelle and desert-dwelling bat species (Polak etþ  al. 2011; Razgour etþ  al. 2011), combined with the increasing development of bodies of open water in the Negev and other drylands, may lead to further resource competition resulting in loss to the region’s biodiversity. Korine etþ  al. (2015) have shown that species richness and activity of desert dwelling bats did not differ between articial and natural bodies of water in the Negev desert, however several species of bats drank or foraged only at natural bodies of water.8.5þ Conclusion and Future DirectionsHuman population growth, land use change and habitat loss have led to massive habitat alterations and destruction, particularly of water sources in arid regions. The availability of water (temporary/permanent) appears to have a strong positive inuence on species of bats richness and activity. This suggests that large temporary pools are important for the conservation of bats in arid environments. A reduction in the availability of temporary pools, due to intensication of arid conditions, is expected to predominantly affect species of bats that forage over water, and will most likely increase interspecic competition for foraging space above the pools. These problems are likely be exacerbated in species of bats that are able to extend into arid areas because of their association with humans. Studies on the distribution of bats in drylands on a large scale should be the focus of future research to understand how climate change and introduction of articial bodies of water effect species distribution, activity and richness. Studies are strongly needed in arid regions to understand the best and most efcient way to provide safe articial water sources for bats that can mitigate increased incidences of drought due to climate change and, in some cases, the total loss of available water, especially in the more temperate arid regions with shorter growing seasons. For example, placement of articial water sources near maternity roosts is instrumental in arid temperate areas with shorter growing seasons (Adams 2010). However, the introduction of articial bodies of water may promote invasion by non-native species and range expansion of others, leading to resource competition. In regions of Europe likely to become water-stressed because of human induced climate changes, bats may be affected as they may lack the physiological means to cope with water limitation (Sherwin etþ  al. 2013). Africa, as well as other arid areas such as the Negev and the Mongolian deserts, has a high diversity of bats but compared to other areas of the world its bat fauna

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233 8þ Bats and Water: Anthropogenic Alterations …has been little studied. Fundamental research is most needed throughout Africa and other arid zones on how often bats need to drink and whether this varies across species, geographically and seasonally. Comparative studies on bats with distributions restricted to arid regions and species that have populations in mesic and arid regions would be particularly informative in this regard. For example, the diversity of renal capacities and habitat use amongst African species of bats of the same family (Happold and Happold 1988), and the emergence of robust family level phylogenies (e.g. Stoffberg etþ  al. 2010) provide an excellent opportunity to study the evolution of renal form and function in African bats in an ecological context. Special focus should be placed on research determining the extent to which African bats are reliant on articial water sources. Such research should tar get arid zone species of bats, especially those species that live in close association with humans because these are the species likely to be impacted by insufcient or polluted water sources. Research is also needed on whether all water sources are used for both drinking and foraging and how bats respond to decreases in water quality as a result of pollutants. Do certain species of bats avoid drinking from low quality bodies of water as shown by Korine etþ  al. (2015)? Would bats still use polluted bodies of water for feeding but not for drinking? If so, how do they detect low quality water, do they do so before they are adversely affected by it and do they have alternative water sources? How are desert-dwelling bats affected by pollutants in water or by waterborne toxins and pollutants in the insect fauna, and are such bats able to deal with such pollutants physiologically? Although least is known about bats and water in sub-Sahara Africa, studies thus far in other regions of the world are in their infancy in terms of understanding the long-term effects of decreased water availability on bat and other wild populations. Due to human destruction of wetlands and riparian habitats as well as unsustainable human population growth that more and more is utilizing greater amounts of fresh water, availability of fresh water to sustain wildlife populations are reaching critically low levels, especially in areas suffering from extended droughts due to human-induced climate disruption. Because water is a key ingredient of all life, focus on this topic needs to increase and because bats act as ‘canaries in a global coal mine,’ studies concerning bats and water are key to better management of water resources in natural and articial areas.Acknowledgmentsþ We would like to acknowledge four anonymous reviewers and Dr. Tigga Kingston for their helpful comments on the chapter. RAA thanks the University of Northern Colorado, the Department of Open Space and Boulder Mountain Parks, Boulder County Department of Parks and Open Space for providing funding for research. This is paper number 871 of the Mitrani Department of Desert Ecology. Open Access This chapter is distributed under the terms of the Creative Commons Attribution Noncommercial License, which permits any noncommercial use, distribution, and reproduction in any medium, provided the original author(s) and source are credited.

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241 8þ Bats and Water: Anthropogenic Alterations … Stechert C, Kolb M, Bahadir M, Djossa BA, Fahr J (2014) Insecticide residues in bats along a land use-gradient dominated by cotton cultivation in northern Benin, West Africa. Environ Sci Pollut R 21:8812–8821 Stoffberg SM, Jacobs DS, Mackie IJ, Matthee CA (2010) Molecular phylogenetics and historical biogeography of Rhinolophus bats. Mol Phylogenet Evol 54:1–9 Swanepoel RE, Racey PA, Shore RF etþ  al (1999) Energetic effects of sublethal exposure to lindane on pipistrelle bats (Pipistrellus pipistrellus). Environ Poll 104:169–177 Syaripuddin K, Kumar A, Sing K W, Halim MRA, Nursyereen MN, Wilson JJ (2014) Mercury accumulation in bats near hydroelectric reservoirs in Peninsular Malaysia. Ecotoxicology 23:1164–1171 Thomas DW, Fenton MB (1978) Notes on the dry season roosting and foraging behaviour of Epomophorous gambianus and Roussettus aegyptiacus (Chiroptera: Pteropodidae). J Zool 186:403–406 Timms BV (2005) Salt lakes in Australia: present problems and prognosis for the future. Hydrobiologia 552:1–15 Tuttle SR, Chambers CL, Theimer TC (2006) Potential effects of livestock water-troughs modications on bats in Northern Arizona. Wildlife Soc Bull 34:602–608 UNESCO (1979) Map of the World Distribution of Arid Regions. UNESCO, Paris Vaughan N, Jones G, Harris S (1996) Effects of sewage efuent on the activity of bats (Chiroptera: Vespertilionidae) foraging along rivers. Biol Cons 78:337–343 Vaughan N, Jones G, Harris S (1997) Habitat use by bats (Chiroptera) assessed by means of a broad-band acoustic method. J Appl Ecol 34:716–730 von Frenckell B, Barclay RMR (1987) Bat activity over calm and turbulent water. Can J Zool 65:219–222 Walker KF (1985) A review of the ecological effects of river regulation in Australia. In: Davies BR, Walmsley RD (eds) Perspectives in Southern Hemisphere Limnology. Springer Netherlands, pp. 111–129 Walsh AL, Harris S, Hutson AM (1995) Abundance and habitat selection of foraging vespertilionid bats in Britain: a landscape-scale approach. Symp Zoo Soc Lond 67:325–344 Walters DM, Fritz KM, Otter RR (2008) The dark side of subsidies: adult stream insects export organic contaminants to riparian predators. Ecol Appl 18:1835–1841 Warren RD, Water DA, Altringham JD etþ  al (2000) The distribution of Daubenton’s bats (Myotis daubentonii) and pipistrelle bats (Pipistrellu spipistrellus) (Vespertilionidae) in relation to small-scale variation in riverine habitat. Biol Cons 92:85–91 Webb PI (1995) The comparative ecophysiology of water balance in micro-chiropteran bats. In: Racey PA Swift SM (eds) Ecology, evolution and behaviour of bats. Oxford University Press, Oxford, pp 203–218 Webb PI, Speakman JR, Racey PA (1995) Evaporative water loss in two sympatric species of vespertilionid bat, Plecotus auritus and Myotis daubentoni: relation to foraging mode and implications for roost site selection. J Zool 235:269–278 Williams AJ, Dickman CR (2004) The ecology of insectivorous bats in the Simpson Desert central Australia: habitat use. Aust Mammal 26:205–214 Williams WD (1981) Running water ecology in Australia. In: Lock MA, Dudley D (eds) Perspectives in Running Water Ecology. Springer US, pp. 367–392 Williams JA, O’Farrell MJ, Riddke BR (2006) Habitat use by bats in a riparian corridor of the Mojave desert in southern Nevada. J Mammal 87:1145–1153 Yom-Tov Y (1993) Character displacement among the insectivorous bats of the Dead Sea area. J Zool 230:347–356 Yom-Tov Y, Mendelssohn H (1988) Changes in the distribution and abundance of vertebrates in Israel during the 20th century. In: Tchernov E, Yom-Tov Y (eds) The zoogeography of Israel. Dr. W. Junk, Dordrecht, pp 515–547 Young RA, Ford GI (2000) Bat fauna of a semi-arid environment in central western Queensland, Australia. Wildlife Res 27:203–215

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Part IIEmerging Disesases

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245Chapter 9White-Nose Syndrome in BatsWinifred F. Frick, Sébastien J. Puechmaille and Craig K.R. Willis© The Author(s) 2016 C.C. Voigt and T. Kingston (eds.), Bats in the Anthropocene: Conservation of Bats in a Changing World, DOIþ  10.1007/978-3-319-25220-9_9Abstractþ White-nose syndrome (WNS) is an infectious disease of hibernating bats that has killed millions of bats since it rst emerged in eastern North America in 2006. The disease is caused by a pathogenic fungus, Pseudogymnoascus (for merly Geomyces) destructans that was likely introduced to North America by human trade or travel, demonstrating the serious problem of global movement of pathogens by humans in the Anthropocene. Here, we present a synthesis of the current state of knowledge on WNS, including disease mechanisms, disease ecology, global distribution and conservation and management efforts. There has been rapid research response to WNS and much about the disease is now well under stood. However, critical gaps in our knowledge remain, including ways to limit spread, or effective treatment options to reduce disease mortality. There are several hibernating bat species in North America that are threatened with extinction from WNS. Protecting those species has become a race against time to nd and implement creative solutions to combat the devastating impacts of this disease. W.F. Frickþ  ()þ  Ecology and Evolutionary Biology, University of California Santa Cruz, Santa Cruz, CA, USA e-mail: wfrick@ucsc.edu S.J. Puechmailleþ  Zoology Institute, Ernst-Moritz-Arndt University, Greifswald, Germany S.J. Puechmailleþ  School of Biological and Environmental Sciences, University College Dublin, Dublin, Ireland C.K.R. Willisþ  Department of Biology, University of Winnipeg, Winnipeg, MB, Canada

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246 W.F. Frick et al.9.1þ IntroductionIn late winter of 2007, biologists at the New York State Department of Environmental Conservation encountered a macabre scene during their annual winter surveys of hibernating bats in caves and mines in northern New York State: heaps of dead bats piled on cave oors (Fig.þ  9.1) (Veilleux 2008). Bats were also seen ying out in the middle of winter onto the snowy landscape and the number of citizen reports of dead bats found in backyards was much higher than normal. A white fuzzy growth was observed on muzzles and wings of the few remaining live bats, which led to the name white-nose syndrome (WNS) (Veilleux 2008; Reeder and Turner 2008; Turner and Reeder 2009). WNS is now recognized as one of the most devastating wildlife epidemics in recorded history and has caused the death of millions of bats in eastern North America. The research and management response to WNS has been rapid and we know much more about WNS than when those rst dead bats were observed in New York, although there is still a great deal about this wildlife disease that is yet to be resolved. The rst evidence of WNS in North America is dated to a photograph taken by a caver at Howe’s Cave in 2006 (Turner and Reeder 2009). Howe’s Cave is a popular tourist attraction that receives hundreds of thousands of visitors each year, many of whom visit from other parts of the world. The white fuzzy growth visible on bats is caused by a pathogenic fungus, which was described as Geomyces destructans (Gargas etþ  al. 2009; Blehert etþ  al. 2009), but was recently re-named Pseudogymnoascus destructans after closer evaluation of its taxonomic allies (Minnis and Lindner 2013). The fungus infects the skin tissues, including the wings and tail membranes, and causes bats to arouse too frequently from torpor Fig.þ  9.1þ Bats that died from WNS during winter at Aeolus Cave in Vermont, USA. Photo by Al Hicks

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247 9þ White-Nose Syndrome in Batsduring hibernation (Lorch etþ  al. 2011; Warnecke etþ  al. 2012) (Fig.þ  9.2). Bats die before spring brings warmer weather and insects for food. WNS has spread rapidly and by 2014 was found in 25 U.S. states and 5 Canadian Provinces (Fig.þ  9.3). A conrmed case of WNS is dened by the presence of cupping erosions on the skin caused by infection by P. destructans, which is determined by histopathological examination (Meteyer etþ  al. 2009). There are currently seven hibernating species in North America that have been conrmed with infections characteristic of WNS, including Myotis lucifugus, Myotis septentrionalis, Myotis sodalis, Myotis leibii, Myotis grisescens, Eptesicus fuscus and Perimyotis subavus. There are several additional species for which P. destructans has been detected on skin tissues using swab sampling and quantitative PCR methods (Muller etþ  al. 2013), but that have not been conrmed with characteristic skin lesions that dene the disease. Two of the species conrmed with WNS (M. sodalis, M. grisescens) were already listed as federally endangered under the US Endangered Species Act before WNS emerged and several other species have been predicted to go globally or regionally extinct due to mortality from WNS (Frick etþ  al. 2010; Langwig etþ  al. 2012; Thogmartin etþ  al. 2013). The US Fish and Wildlife Service listed M. septentrionalis as federally threatened in 2015 due to the risk of extinction Fig.þ  9.2þ A hibernating little brown myotis (Myotis lucifugus) with typical WNS infection visible on skin tissues. Photo by Ryan von Linden

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248 W.F. Frick et al.from WNS-associated mortality. In addition, a status review of M. lucifugus is being conducted to determine whether listing as federally endangered is warranted of this once common species (Frick etþ  al. 2010). In Canada, three species, M. lucifugus, M. septentrionalis and P. subavus were listed as endangered in 2015. The rapid spread and extensive mortality associated with WNS raise serious concerns about population viability for species that are being impacted by this disease. In this chapter, we review what is currently known about WNS, focusing on mechanisms of disease, disease ecology, global distribution patterns and conservation and management. We rst explain why WNS belongs in a volume addressing bats in the Anthropocene. We review what is known about disease mechanisms, including what we currently understand about the physiology of the disease and immune response in bats. We then review what is currently known about disease ecology of WNS, including the population impacts to species, and then highlight Fig.þ  9.3þ Map of current distribution and past spread of WNS across North America. Conrmed WNS cases are those where disease has been conrmed by histological examination of tissues. Suspect cases are those that are either a molecular detection of Pseudogymnoascus destructans by quantitative PCR (Muller etþ  al. 2013) or by visual signs and/or aberrant behaviour consistent with WNS disease at a site. Updated versions of this map are made publically available at whitenosesyndrome.org

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249 9þ White-Nose Syndrome in Batsunanswered questions about transmission dynamics. We discuss global distributions patterns, focusing on what is known about WNS in Europe. We conclude by discussing current conservation and management strategies. Wildlife disease is increasingly recognized as a major conservation threat (Daszak 2000). Global movements of humans increase the probability and rate at which we introduce pathogens into naïve ecosystems (Cunningham etþ  al. 2003). This human-mediated spread of pathogens has been dubbed “pathogen pollution” to highlight the role of human trade and travel in the spread of wildlife pathogens (Cunningham etþ  al. 2003). The fungus P. destructans was presumably introduced to North America from Europe by people, most likely from someone who had visited caves in Europe and subsequently visited Howe’s Cave with contaminated boots or gear (Puechmaille etþ  al. 2011c; Leopardi etþ  al. 2015). No bats are known to migrate between the Americas and other continents, implicating human trade or travel in the trans-Atlantic arrival of the fungus (Wibbelt etþ  al. 2010). Ironically, bats are often seen as reservoirs of diseases with consequences to human health (e.g. rabies, SARS, etc.). In the case of WNS, humans were most likely the unwitting transcontinental carrier of a pathogen that has killed millions of bats and now threatens species with extinction. The emergence of WNS has dramatically changed conservation planning and population monitoring of temperate bats in North America (Foley etþ  al. 2011). On the positive side, this crisis prompted collaborative research efforts among bat conservationists in North America and in Europe. Although mortality from WNS is currently restricted to North America, the pathogen is a potential threat to hiber nating bat populations in other parts of the globe and is a global concern for bats in the Anthropocene (Puechmaille etþ  al. 2011c).9.2þ Disease MechanismsChallenge or inoculation studies (e.g. Lorch etþ  al. 2011; Warnecke etþ  al. 2012; Wilcox etþ  al. 2014) and comparative studies of bats from affected versus unaffected hibernacula (Moore etþ  al. 2011; Storm and Boyles 2011; Reeder etþ  al. 2012; Brownlee-Bouboulis and Reeder 2013) have led to progress in our understanding of mechanisms underlying WNS. The wings of bats are physiological active tissue involved in gas exchange and uid balance. In general, results of physiological studies are converging on a consensus that cutaneous infection of the wings accounts for the physiological and behavioural effects of WNS (Cryan etþ  al. 2010). Lorch etþ  al. (2011) experimentally inoculated the wings of healthy M. lucifugus with P. destructans for comparison to sham-inoculated controls. They housed bats in temperatureand humidity-controlled incubators that maintained environmental conditions approaching natural hibernacula [82þ  % relative humidity (RH) at 6.5þ  °C]. This experiment resolved a critical question by demonstrating that experimental infection with P. destructans caused the dening characteristics

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250 W.F. Frick et al.of WNS (e.g. cupping erosions in the epidermis associated with fungal growth, Meteyer etþ  al. 2009). They also found that P. destructans spread from infected to un-infected bats housed in the same cages but did not spread between cages in the same incubator conrming contact but not airborne transmission of the causal pathogen under laboratory conditions. Lorch etþ  al. (2011) did not detect differences in survival between infected and un-infected bats possibly because the experimental duration was shorter than a typical hibernation season and/ or because humidity in this experiment was lower than that of hibernacula used by M. lucifugus in the wild, potentially inuencing hibernation patterns of both control and infected bats. Warnecke etþ  al. (2012) repeated aspects of Lorch etþ  al.’s (2011) experiment but increased ambient humidity to >97þ  % RH at 7þ  °C and ran the experiments for 120þ  days (vs. 102þ  days in Lorch etþ  al. 2011). In Warnecke etþ  al.’s (2012) experiment, all sham-inoculated bats survived four months of hiber nation, while infected bats exhibited a signicant increase in the frequency of periodic arousals, reduced fat reserves and reduced survival, thus conrming that infection with P. destructans alone causes the pathology that denes WNS, altered torpor behaviour and mortality. A eld study comparing arousal frequency of bats in affected versus unaffected caves (Reeder etþ  al. 2012) also reported a difference in arousal frequency similar to that observed by Warnecke etþ  al. (2012). Together these ndings suggest a strong role for increased arousal frequency and altered energy balance in WNS pathophysiology. Comparisons of control and infected bats have also provided insight into immune responses (or lack of responses) of bats during and after hibernation. Hibernators generally exhibit down-regulated immune function during winter and bat species affected by WNS appear to be no exception (Meteyer etþ  al. 2009, 2012; Moore etþ  al. 2011). During hibernation, there is little evidence of initiation of an inammatory response or recruitment of immune cells in bats infected by P. destructans based on histopathology (Meteyer etþ  al. 2009, 2012). Despite the absence of an inammatory response, however, variation in other aspects of cellular immunity may have a role to play. Moore etþ  al. (2013) found differences in immunological responses of M. lucifugus in affected versus unaffected hibernacula, specically higher leukocyte counts, reduced antioxidant activity and lower levels of interleukin-4 (an important precursor for differentiation of T-cells) in bats from WNS-affected caves. Although comparisons between populations of bats in different hibernacula are challenging to interpret because of the potential for underlying differences between bats independent of infection, these ndings suggest that even the hardest-hit bat species attempt some, albeit weak, immune response to P. destructans infection. This also raises the possibility that some bats may be better equipped to resist infection than others (Puechmaille etþ  al. 2011c) with the potential for directional selection on immune function if these differences are heritable and provide a survival advantage. Immune responses of bats to WNS could be as much a disadvantage as an advantage. Meteyer etþ  al. (2012) recently reported the disheartening paradox that some sur vivors of WNS exhibit characteristic signs of immune reconstitution inammatory syndrome (IRIS). When infected bats emerge from hibernation and their immune

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251 9þ White-Nose Syndrome in Batsfunction resumes, they exhibit a massive neutrophilic inammatory response to the fungal infection. This response appears to dramatically increase tissue damage and may reect an over-reaction to infection because euthermic body temperatures in spring would likely be sufcient to combat the fungal infection (Chaturvedi etþ  al. 2010; Puechmaille etþ  al. 2011b; Verant etþ  al. 2012). The response is likely energetically expensive and the resulting wing damage could compromise ight ability and, therefore, spring energy balance by increasing healing and immunity costs, while reducing potential foraging efciency at a time when energy balance is critical to support reproduction. Further studies of the role of IRIS in the ecology of WNS are essential for understanding the potential for populations to recover from WNS. A down-regulated immune response in hibernating bats generally, combined with increased arousal frequency (Boyles and Willis 2010; Reeder etþ  al. 2012; Warnecke etþ  al. 2012) and possibly increased metabolic rate and body temperature during torpor following infection (Storm and Boyles 2011; Verant etþ  al. 2014), appears to result in premature fat depletion and starvation. However, why fungal infection would increase arousal frequency is still not fully understood. Cryan etþ  al. (2010) proposed the hypothesis that fungal damage to the wings of bats could lead to increased evaporative water loss (EWL) across damaged epidermis. Rates of EWL during torpor are a strong predictor of arousal frequency in hiber nators (Ben-Hamo etþ  al. 2013; Thomas and Cloutier 1992; Thomas and Geiser 1997) so an increase in EWL or uid loss due to skin damage from infection by P. destructans could lead to the observed effects on arousals. Willis etþ  al. (2011) used data on water loss and arousal frequency in healthy bats, combined with an individual-based model quantifying survival of hypothetical populations of bats, to demonstrate that even a small increase in EWL resulting from infection could cause the same patterns of arousal and mortality observed for infected bats, thus highlighting the plausibility of the dehydration hypothesis. Two independent datasets from both captive and free-ranging bats also support a role for dehydration and uid loss in WNS pathophysiology (Cryan etþ  al. 2013; Warnecke etþ  al. 2013). In addition to high hematocrit levels consistent with dramatic uid loss, Cryan etþ  al. (2013) and Warnecke etþ  al. (2013) both found evidence of electrolyte depletion (with no evidence of renal pathology), consistent with hypotonic dehydration due to uid loss across damaged wings. Presumably infected bats lose uid containing both water and electrolytes across injured wing tissue but can only replenish or partially replenish water stores by drinking, because electrolytes are not available in hibernacula. Warnecke etþ  al. (2013) also found preliminary evidence of a respiratory response to metabolic acidosis in infected bats which they hypothesized reect reduced perfusion of infected tissues, localized anaerobic metabolism and acidosis, and increased respiratory rate to increase CO2 excretion and counter acidosis. In addition to increased arousal frequency, these physiological responses also predict increased metabolic costs and elevated body temperature during torpor. To date, measurements of torpid body temperature with enough precision to test this hypothesis are unavailable but these would be valuable, especially alongside measurements of metabolism during torpor and arousal in infected versus un-infected bats.

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252 W.F. Frick et al.Other physiological mechanisms could also be at play. Willis and Wilcox (2014) reviewed three (of many potential) hormone systems that could be inuenced by WNS, both within individuals and via selection on traits which could favour survival. For example, the lipostat hormone leptin is strongly associated with winter energy balance and pre-hibernation fattening. Bats must enter a state of leptin resistance during fall to accumulate adequate fat stores to survive the winter. If, as the evidence suggests, WNS represents a challenge for hibernation endurance, bats with the greatest leptin resistance (and therefore potential fat stores) in autumn may be best equipped to survive increased arousals associated with WNS (Willis and Wilcox 2014). Interactions between WNS and other hor mone systems important for seasonal energetics, body temperature regulation and energy and uid balance (e.g. glucocorticoids, melatonin, thyroid hormone, vasopressin, androgens) could also play important roles in disease dynamics and evolution of remnant populations and are worth further study. In addition to physiological research, recent studies have also examined behavioural mechanisms associated with WNS that could reect either adaptive responses to disease or maladaptive pathological responses. Langwig etþ  al. (2012) reported that a much greater proportion of the M. lucifugus surveyed in WNS-affected caves after the emergence of the disease were hibernating solitarily (i.e. without clustering) compared to bats surveyed before WNS. This could reect a behavioural change by individuals following infection or selection by WNS for bats which tend to roost individually (Langwig etþ  al. 2012). Wilcox etþ  al. (2014) reported behavioural observations of bats inoculated with P. destructans and found evidence supporting the former hypothesis. Infected bats gradually reduced their clustering behaviour as hibernation progressed. Wilcox etþ  al. (2014) also observed a reduction in behavioural activity during arousals, in general, for affected bats. Taken together, reduced clustering and reduced activity by infected bats could reect general patterns known as “sickness behaviour”, a coordinated response to infection character ized in part by lethargy presumably to save energy for immune responses (Adelman and Martin 2009). These behaviours could also reduce the potential for transmission among individuals in a social group within a hibernaculum. Even bats that have already been infected with P. destructans could benet by reduced subsequent exposure to other infected individuals because new contacts could lead to additional areas of infection in the wings, exacerbating disease severity. On the other hand, reduced clustering behaviour could increase energy expenditure and EWL leading to negative consequences for survival. More work is needed to understand the sur vival consequences of a range of physiological and behavioural responses to WNS.9.3þ Disease Ecology of WNSOne of the dening characteristics of WNS is that it is a multi-host disease, meaning that P. destructans infects multiple bat species. Although all hibernating bat species in northeastern North America can be infected with P. destructans and

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253 9þ White-Nose Syndrome in Batsdevelop the cupping erosions in their skin tissues that characterize the disease, population impacts from WNS vary widely among species (Langwig etþ  al. 2012; Turner etþ  al. 2011). Prior to the emergence of WNS in North America, all six hibernating bat species that occur in the northeastern United States had positive population growth trends (Frick etþ  al. 2010; Langwig etþ  al. 2012). With the emer gence of WNS, four of these six species suffered severe population declines (M. septentrionalis, M. lucifugus, M. sodalis and P. subavus) (Langwig etþ  al. 2012). Two species (M . leibii and E . fuscus) have experienced less severe impacts from disease (Langwig etþ  al. 2012). In addition, species of the genus Corynorhinus do not appear to get sick and die from WNS, despite occurring in WNS-affected caves in states in the mid-Atlantic region, such as West Virginia and Virginia. Why some species suffer higher mortality than others is an important area of current research, but there are no clear-cut answers yet. Langwig etþ  al. (2012) showed that differences in roosting microclimates (temperature and RH) were correlated with differential impacts among sites for some species. For example, sites with warmer roosting temperatures had the highest declines for M . lucifugus and sites with highest RH had the highest declines for M . sodalis, suggesting that roosting microclimates could play an important role in WNS impacts (Langwig etþ  al. 2012). Differences in environmental conditions as well as exposure, transmission, susceptibility, torpor physiology and immune response among species could contribute to observed differences in mortality. Future research focusing on differences in these factors among species will be critical for identifying the risks to particular species. Understanding whether transmission is dependent on the density of hibernating populations is key to determining whether WNS will cause bats to go extinct or whether bat populations will stabilize at low numbers. For diseases where transmission is density-dependent, the probability of extinction is much lower because transmission rates decline as populations become smaller (De Castro and Bolker 2004). Langwig etþ  al. (2012) showed that for bats that hibernate in dense clusters (e.g. M. lucifugus and M. sodalis), there was no evidence for density-dependent declines, meaning that declines from WNS were equally severe in populations that ranged from 100 to 100,000 bats. In contrast, there was evidence that declines were smaller in smaller populations for species that roost solitarily (e.g. P. subavus and M. septentrionalis). Although the declines were density-dependent in M. septentrionalis, declines were not predicted to stabilize before populations went extinct in this species, suggesting that this species is at serious risk of extinction from WNS. Determining whether a pathogen can persist in an environmental reservoir is also important for understanding disease transmission dynamics and extinction risk from disease (De Castro and Bolker 2004). Pathogens that can persist in an environmental reservoir are more likely to drive species extinct because hosts can get infected from the environment even if only a few individuals remain. Studies have shown that P. destructans is found in sediments and environmental substrates in hibernacula (Puechmaille etþ  al. 2011a; Lindner etþ  al. 2011; Lorch etþ  al. 2013a, b). Lorch etþ  al. (2013b) demonstrated that viable P. destructans can be cultured

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254 W.F. Frick et al.from samples taken during late summer when bats have been absent for several months, suggesting that P. destructans persists in the environment between hiber nation seasons. An unpublished experiment conducted by Al Hicks at the New York Department of Environmental Conservation demonstrated that naïve bats that had never been exposed to P. destructans could contract disease and die from WNS when placed in an infected hibernaculum with no access to other infected bats (Hicks, pers. comm.). The evidence to date suggests that hibernacula are environmental reservoirs for P. destructans, which has potentially dire consequences if the environment proves a major source of transmission. WNS is a seasonal disease and recent work by Langwig etþ  al. (2015) describes how the seasonal patterns of transmission of P. destructans are driven by hiber nation. Bats begin to become infected in the fall when they return to hibernacula during fall swarm and transmission spikes in early winter once bats begin hiber nating. Infection intensity increases during hibernation and peaks in late winter at which time most bats have become infected. These seasonal patterns are similar to temporal prevalence of visual signs of P. destructans growth on bats at sites in Europe as described by Puechmaille etþ  al. (2011a), where a peak of infection was also observed in late hibernation when most individuals present were infected. In Langwig etþ  al.’s study, most bats cleared infection during summer and prevalence of infection fell to zero by late summer at maternity roosts. The seasonal timing of infection suggests that mortality occurs at a time of maximal impact for populations (before the birth pulse). However, a peak in transmission after bats begin hibernating in early winter may reduce the rate of spread among hibernacula since bats presumably move among sites less frequently once they start hibernating compared to during the fall swarm period.9.4þ Status of P. Destructans/WNS in EuropeIn contrast to the severe impacts WNS has on North American bat species, P. destructans is commonly found on bats in Europe but is not associated with mass mortality (Wibbelt etþ  al. 2010; Puechmaille etþ  al. 2011a). Europe is a putative source of the pathogen and the pathogen likely arrived in North America by some means of human trade or travel. Ongoing studies on global distribution of P. destructans (S.J. Puechmaille and J.R. Hoyt, unpublished data), including surveys in temperate Asia, may reveal important insights about the global distribution of the pathogen. Pseudogymnoascus destructans was rst reported in Europe by Puechmaille etþ  al. (2010) who sampled a hibernating Myotis myotis from southwestern France showing the typical powdery white fungal growth on its nose. Since then, the fungus has been morphologically and genetically conrmed in 14 countries in Europe (France, Portugal, Belgium, The United Kingdom, The Netherlands, Germany, Switzerland, Austria, Slovakia, Poland, Hungary, Ukraine and Estonia) and

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255 9þ White-Nose Syndrome in Batsconvincing photographic evidence further supports its presence in an additional four countries (Luxembourg, Denmark, Romania and Turkey [the European part]) (Martínková etþ  al. 2010; Puechmaille etþ  al. 2010, 2011a; Kubátová etþ  al. 2011; Simonovicová etþ  al. 2011; Mestdagh etþ  al. 2012; Wibbelt etþ  al. 2010, 2013; Burger etþ  al. 2013; Paiva-Cardoso etþ  al. 2014; Sachanowicz etþ  al. 2014). At the continental scale, most European reports are from northeastern France through Belgium, the Netherlands, Germany and the Czech Republic, but it remains unclear whether this pattern of higher prevalence of the fungus is real or reects sampling bias (Puechmaille etþ  al. 2011a). Studies conducted in Italy, Slovenia and Sweden, where P. destructans was not detected (Voyron etþ  al. 2010; Nilsson 2012; Mulec etþ  al. 2013), support the hypothesis that P. destructans occurrence and/or prevalence varies between different geographic regions in Europe (Puechmaille etþ  al. 2011a). Puechmaille etþ  al. (2011a) demonstrated that the prevalence of visible signs of P. destructans on bat wings and nose drastically varied through the hibernation period with the rst cases appearing around mid-January. The number of cases increased to reach a peak in March and declined as bats emerged from hibernation. This pattern further complicates comparisons of prevalence of visual signs of fungal growth on bats between sites, regions or years unless surveys are car ried out at the same time. Work done in the Czech Republic and Slovakia detected differences in prevalence of bats suspected to carry P. destructans (based on visual observations) between sub-mountain humid to mesic regions (higher prevalence) and mountainous and limestone regions (lower prevalence) (Martínková etþ  al. 2010), supporting the idea that P. destructans is not equally abundant across Europe. Nevertheless, the differences in sampling strategy (spatio-temporal), sampling intensity (number of sites, number of samples), nature of the samples collected (e.g. swab from the bat vs. environment vs. guano) and analysis techniques (e.g. culture, PCR detection) between different European studies make quantication of these neand large-scale patterns challenging (Puechmaille etþ  al. 2011a). All conrmed cases of P. destructans infection come from fungal material collected on bats with the exception of a case from Estonia where the fungus has been isolated and cultured from the walls of the hibernation site, representing the rst published isolation of viable spores from the environment in Europe or North America (Puechmaille etþ  al. 2011a). In terms of species, available data suggest that M. myotis is the most commonly infected species (ca. 66þ  % of cases) with P. destructans in Europe (Martínková etþ  al. 2010; Puechmaille etþ  al. 2011a). The fungus is known to also infect another nine species of European Myotis (ranked by decreasing order of prevalence): M. dasycneme, M. mystacinus, M. blythii, M. daubentonii, M. brandtii, M. emarginatus, M. nattereri, M. bechsteinii and M. escalerai/sp. A. The list of species with P. destructans infection is likely to increase as sampling intensity increases as illustrated by the recent Zukal etþ  al. (2014) study which reported infection of a few individuals from three more species of the family Vespertilionidae, Eptesicus nilssonii, Plecotus auritus and Barbastella Barbastellus, as well as on a single individual of Rhinolophus hipposideros, of the family Rhinolophidae.

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256 W.F. Frick et al.Owing to the protection of bats across Europe and the absence of mass mortality, only three studies with limited to moderate numbers of samples have investigated the pathology of P. destructans during the hibernation period (Pikula etþ  al. 2012; Wibbelt etþ  al. 2013; Bandouchova etþ  al. 2015). In Europe, P. destructans invasion of the wing membrane is generally restricted to the epidermis and adnexae without deep invasion into the underlying connective tissue but with occasional formation of neutrophilic pustules, contrasting with the common and extensive invasion of dermal connective tissue in bats from North America (Pikula etþ  al. 2012; Wibbelt etþ  al. 2013; Zukal etþ  al. 2014; Bandouchova etþ  al. 2015). Based on investigation of two euthanized individuals, P. destructans invasion in the skin of the muzzle seems to be more pronounced than invasion of the wing membrane (Pikula etþ  al. 2012; Wibbelt etþ  al. 2013). As damage to the skin of the muzzle may not be as physiologically important for homeostasis as damage to the wing membranes (Cryan etþ  al. 2010; Reeder etþ  al. 2012; Warnecke etþ  al. 2013), we suggest that it may be important to differentiate the pathology of P. destructans on the wing and on the muzzle. If dehydration and uid loss play an important role in WNS pathophysiology, quantifying wing damage consistently (e.g. following Reeder etþ  al. 2012 or an alternative scoring system) alongside physiological measures of disease severity will be critical for a better understanding of the disease, its progression and species-specic attributes, compared to the commonly reported dichotomous presence/absence of the disease. The term WNS was originally used to describe the symptoms associated with bats in the eld before the disease was fully characterized as a cutaneous infection of skin tissues by the pathogenic fungus, P. destructans (Blehert etþ  al. 2009; Meteyer etþ  al. 2009). As such, the name ‘WNS’ has changed from referring to a set of symptoms, including visible fungal growth on skin surfaces, depletion of fat reserves, altered torpor patterns and aberrant winter behaviour (Blehert etþ  al. 2009) to referring to the presence of disease as dened by the presence of cutaneous infection characterized by cupping erosions (Meteyer etþ  al. 2009). This has led to confusion and some debate about whether the term WNS should be used to describe infections occurring in Europe, which are pathologically similar to those in North America but which do not include mass mortality or aberrant winter behaviour (Puechmaille etþ  al. 2011a). Despite its original denition as a syndrome (Veilleux 2008; Reeder and Turner 2008; Turner and Reeder 2009), the term WNS is now routinely used to refer the cutaneous infection caused by P. destructans, which have been documented in Europe (Pikula etþ  al. 2012; Wibbelt etþ  al. 2013; Zukal etþ  al. 2014). Some have advocated a name change to clarify a difference between a ‘syndrome’ and a ‘disease’ caused by fungal infection (Chaturvedi and Chaturvedi 2011). Inconsistency in the literature could lead to confusion but recent use of the term white-nose disease (WND; Paiva-Cardoso etþ  al. 2014) could clar ify the situation by providing terminology reminiscent enough of WNS to avoid confusion but technically consistent with the denition of a disease. Recent work comparing colony sizes of hibernating vespertilionid bats in North America before and after the emergence of WNS, to current colony sizes in Europe, reveals an intriguing pattern. Before WNS emerged in North America,

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257 9þ White-Nose Syndrome in Batscolony sizes of hibernating bats were, on average, about 10-fold larger than those of similar species in Europe (Frick etþ  al. 2015). However, after the emer gence of WNS, colony sizes in eastern North America are no longer statistically different from those in Europe (Frick etþ  al. 2015), raising the following question: Were hibernating bat colonies in Europe once much larger prior to the emergence of WNS there? If WNS is indeed acting as a hidden force on bat populations in Europe, then small winter colony sizes in eastern North America may become the norm for species in North America that manage to persist. However, Frick etþ  al. (2015) also show that 69þ  % of winter colonies of M. septentrionalis were entirely eliminated within 7þ  years of WNS detection, suggesting that this species is rapidly disappearing from the landscape. The predicted extinction of M. septentrionalis from WNS begs the question whether past extinctions of bat species may have also occurred in Europe.9.5þ Conservation and ManagementConservation and management strategies for WNS in North America have focused primarily on preventing spread of the pathogen to new areas through decontamination protocols as well as cave closures to limit the potential for human-mediated spread. Decontamination of gear used in hibernacula by both recreational cavers and bat researchers is an important management strategy to reduce the risk of spread of P. destructans by humans. P. destructans spores have been found on eld gear after use in infected sites and therefore utmost precaution is needed to reduce the chance that researchers and cavers spread P. destructans to new areas. Cave closures have been controversial and have been met with some resistance by some members of the caving community. Some cave closures have subsequently been relaxed in parts of the western United States where P. destructans has not yet spread. Determining whether cave closures are effective can be challenging given that the absence of spread in areas is hard to measure. Bats are capable of spreading the fungus, but the primary focus of closing caves and advocating decontamination was to slow spread by people, especially to distant locations. Finding a treatment for infected bats has proved elusive and difcult. Several studies have examined the efcacy of treating bats with anti-fungal chemicals, such as terbinane, but none have shown any promise. There has also been inter est in alternative forms of treatment, including use of naturally occurring bacteria (Fritze etþ  al. 2012; Hoyt etþ  al. 2015) or volatile compounds (Cornelison etþ  al. 2014). Recent work by Cornelison etþ  al. (2014) showed that a volatile organic compounds (VOCs) inhibited growth of P. destructans in vitro. Similarly, a recent study by Hoyt etþ  al. (2015) showed that Pseudomonas bacteria that naturally occur on hibernating bats inhibit growth of P. destructans in vitro. Other strains of Pseudomonas found in Europe have shown similar results (Fritze etþ  al. 2012). Research on these biological control treatment options is still in early stages and although early lab results have shown promise, experimental and eld trials will

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258 W.F. Frick et al.need to be conducted before the efcacy of these approaches is fully evaluated. The WNS research and management community is developing standards and protocols for evaluating the safety and efcacy of biological treatment options. Other ideas for active management have included building articial hibernacula that can be cleaned and decontaminated each summer between hibernating seasons. An experimental articial hibernacula was built in Tennessee and existing military bunkers have been used as articial hibernaculum in the northeastern US. The goal of these structures is to provide a place for bats to hibernate that does not serve as an environmental source of transmission when bats re-enter the hiber naculum in fall. To date there have been no studies to determine whether bats will use these articial hibernacula naturally and whether survival will be improved in these sites. Given what we know about the potential role that electrolyte depletion plays in the physiology of the disease, some researchers have also explored the potential for electrolyte therapy for hibernating bats by providing access to electrolyte supplements during hibernation. Experimental trials to test this are underway. Finally, bats are very difcult to breed in captivity and, while the prospect of captive breeding and management of bats has been explored, it remains doubtful whether this approach could be useful as a management tool for bat species affected by WNS. However, if breeding programmes could be developed, they could provide a supply of animals for laboratory studies to reduce potential impacts of research on wild populations.9.6þ ConclusionsAlthough we have learned a great deal about WNS in the past seven years, there are still many unanswered questions about disease mechanisms, ecology, transmission dynamics, long-term impacts, global distribution patterns and potential treatment options that will be important for managing WNS and its impacts on bats. The US Fish and Wildlife Service has been pivotal in terms of coordinating meetings for information exchange among researchers and state biologists as well as directly funding much of the research on WNS in both the US and Canada. Research priorities for management and conservation of species have focused on topics such as establishing that P. destructans was the causative agent of infection, trying to identify potential treatment of infection, the physiology of infection and mechanisms of mortality, characterizing the environmental reservoir and under standing transmission and immunological response. For many of us, working on WNS is a grim business. There is nothing quite like the experience of going underground and entering a chamber that was formally home to thousands of bats and seeing empty walls and a few straggling survivors covered in white fungus. However, the sense of commitment within the WNS community and the dedication of researchers and managers to try and nd new ways to understand and solve this crisis provide a certain hope. We have yet to nd

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259 9þ White-Nose Syndrome in Batsa way to stop bats dying from WNS, but we are trying hard to do so. Whether we are able to prevent species extinctions may rely, in part, on the creativity to nd solutions before it is too late and the willingness of agency biologists to implement creative solutions without clear assurances of outcomes.Open Access This chapter is distributed under the terms of the Creative Commons Attribution Noncommercial License, which permits any noncommercial use, distribution, and reproduction in any medium, provided the original author(s) and source are credited.ReferencesAdelman JS, Martin LB (2009) Vertebrate sickness behaviors: adaptive and integrated neuroendocrine immune responses. Integr Comp Biol 49:202–214 Bandouchova T, Bartonicka T, Berkova H, Brichta J, Cerny J, Kovacova V, Kolarik M, Köllner B, Kulich P, Martínkova N, Rehak Z, Turner GG, Zukal J, Pikula J (2015) Pseudogymnoascus destructans: evidence of virulent skin invasion for bats under natural conditions, Europe. Transbound Emerg Dis 62:1–5 Ben-Hamo M, Muñoz-Garcia A, Williams JB, Korine C, Pinshow B (2013) Waking to drink: rates of evaporative water loss determine arousal frequency in hibernating bats. J Exp Biol 216:573–577 Blehert DS, Hicks AC, Behr M, Meteyer CU, Berlowski-Zier BM, Buckles EL, Coleman JTH, Darling SR, Gargas A, Niver R, Okoniewski JC, Rudd RJ, Stone WB (2009) Bat white-nose syndrome: an emerging fungal pathogen? Science 323:227 Boyles JG, Willis C (2010) Could localized warm areas inside cold caves reduce mortality of hibernating bats affected by white-nose syndrome? Front Ecol Environ 8:92–98 Brownlee-Bouboulis SA, Reeder DM (2013) White-nose syndrome-affected little brown myotis (Myotis lucifugus) increase grooming and other active behaviors during arousals from hiber nation. J Wildl Dis 49:850–859 Burger K, Gebhardt S, Wolfhahrt G, Wibbelt G, Reiter G (2013) First conrmed records of Geomyces destructans (Blehert and Gargas 2009) in Austria. Ber Naturwiss-Med Ver Innsbruck 98:127–135 Chaturvedi V, Chaturvedi S (2011) What is in a name? A proposal to use geomycosis instead of white nose syndrome (WNS) to describe bat infection caused by Geomyces destructans. Mycopathologia 171:231–233 Chaturvedi V, Springer DJ, Behr MJ, Ramani R, Li X, Peck MK, Ren P, Bopp DK, Wood B, Samsonoff WA, Butchkoski CM, Hicks AC, Stone WB, Rudd RJ, Chaturvedi S (2010) Morphological and molecular characterizations of Psychrophilic fungus Geomyces destructans from New York bats with white nose syndrome (WNS). PLoS ONE 5:e10783 Cornelison CT, Gabriel KT, Barlament C, Crow SA (2014) Inhibition of Pseudogymnoascus destructans growth from conidia and mycelial extension by bacterially produced volatile organic compounds. Mycopathologia 177:1–10 Cryan PM, Meteyer CU, Blehert DS, Lorch JM, Reeder DM, Turner GG, Webb J, Behr M, Verant M, Russell RE, Castle KT (2013) Electrolyte depletion in white-nose syndrome bats. J Wildl Dis 49:398–402 Cryan PM, Meteyer CU, Boyles JG, Blehert DS (2010) Wing pathology of white-nose syndrome in bats suggests life-threatening disruption of physiology. BMC Biol 8:135 Cunningham AA, Daszak P, Rodriguez JP (2003) Pathogen pollution: dening a parasitological threat to biodiversity conservation. J Parasitol 89:S78–S83 Daszak P (2000) Emerging infectious diseases of wildlife—threats to biodiversity and human health. Science 287:443–449

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260 W.F. Frick et al. De Castro F, Bolker B (2004) Mechanisms of disease-induced extinction. Ecol Lett 8:117–126 Foley J, Clifford D, Castle K, Cryan PM, Ostfeld RS (2011) Investigating and managing the rapid emergence of white nose syndrome, a novel, fatal, infectious disease of hibernating bats. Conserv Biol 25:223–231 Frick WF, Puechmaille SJ, Hoyt JR, Nickel BA, Langwig KE, Foster JT, Barlow KE, Bartonika T, Feller D, Haarsma AJ, Herzog C, Horáek I, van der Kooij J, Mulkens B, Petrov B, Reynolds R, Rodrigues L, Stihler CW, Turner GG, Kilpatrick AM (2015) Disease alters macroecological patterns of North American bats. Glob Ecol Biogeogr 24:741–749 Frick WF, Pollock JF, Hicks AC, Langwig KE, Reynolds DS, Turner GG, Butchkoski CM, Kunz TH (2010) An emerging disease causes regional population collapse of a common North American bat species. Science 329:679–682 Fritze M, Huong Pham TL, Irmtraut I (2012) Effekt des bodenbakteriums Pseudomonas veroniilike PAZ1 auf das wachstum des white-nose erregers Geomyces destructans in antagonistentests. Nyctalus 17:104–107 Gargas A, Trest MT, Christensen M, Volk TJ, Blehert DS (2009) Geomyces desctructans sp. nov. associated with bat white-nose syndrome. Mycotaxon 108:147–154 Hoyt JR, Cheng TL, Langwig KE, Hee MM, Frick WF, Kilpatrick AM (2015) Bacteria isolated from bats inhibit the growth of Pseudogymnoascus destructans, the causative agent of whitenose syndrome. PLoS ONE 10:e0121329 Kubátová A, Koukol O, Nováková A (2011) Geomyces destructans, phenotypic features of some Czech isolates. Czech Mycol 63:65–75 Langwig KE, Frick WF, Reynolds R, Parise K, Drees KP, Hoyt JR, Cheng TL, Kunz TH, Kilpatrick AM (2015) Host and pathogen ecology drive the seasonal dynamics of a fungal disease, white-nose syndrome. Proc Roy Soc Lond B 282:20142335 Langwig KE, Frick WF, Bried JT, Hicks AC, Kunz TH, Kilpatrick AM (2012) Sociality, density-dependence and microclimates determine the persistence of populations suffering from a novel fungal disease, white-nose syndrome. Ecol Lett 15:1050–1057 Leopardi S, Blake D, Puechmaille SJ (2015) White-Nose Syndrome fungus introduced from Europe to North America. Curr. Biol 25:R217–219 Lindner DL, Gargas A, Lorch JM, Banik MT, Glaeser J, Kunz TH, Blehert DS (2011) DNAbased detection of the fungal pathogen Geomyces destructans in soils from bat hibernacula. Mycologia 103:241–246 Lorch JM, Meteyer CU, Behr MJ, Boyles JG, Cryan PM, Hicks AC, Ballmann AE, Coleman JTH, Redell DN, Reeder DM, Blehert DS (2011) Experimental infection of bats with Geomyces destructans causes white-nose syndrome. Nature 480:376–378 Lorch JM, Lindner DL, Gargas A, Muller LK, Minnis AM, Blehert DS (2013a) A culture-based survey of fungi in soil from bat hibernacula in the eastern United States and its implications for detection of Geomyces destructans, the causal agent of bat white-nose syndrome. Mycologia 105:237–252 Lorch JM, Muller LK, Russell RE, O’Connor M, Lindner DL, Blehert DS (2013b) Distribution and environmental persistence of the causative agent of white-nose syndrome, Geomyces destructans, in bat hibernacula of the eastern United States. Appl Environ Microbiol 79:1293–1301 Martínková N, Bakor P, Bartonika T, Blaková P, Cervený J, Falteisek L, Gaisler J, Hanzal V, Horáek D, Hubálek Z, Jahelková H, Kolaík M, Korytár L, Kubátová A, Lehotská B, Lehotský R, Luan RK, Májek O, Matj J, Rehák Z, Šafá J, Tájek P, Tkadlec E, Uhrin M, Wagner J, Weinfurtová D, Zima J, Zukal J, Horáek I (2010) Increasing incidence of Geomyces destructans fungus in bats from the Czech Republic and Slovakia. PLoS ONE 5:e13853 Mestdagh X, Baltus L, Hoffman L, Titeux N (2012) Découverte de chauves-souris au nez blanc au Luxembourg. Bull Soc Nat Luxemb 113:141–149 Meteyer CU, Barber D, Mandl JN (2012) Pathology in euthermic bats with white nose syndrome suggests a natural manifestation of immune reconstitution inammatory syndrome. Virulence 3:583–588

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261 9þ White-Nose Syndrome in Bats Meteyer CU, Buckles EL, Blehert DS, Hicks AC, Green DE, Shearn-Bochsler V, Thomas NJ, Gargas A, Behr MJ (2009) Histopathologic criteria to conrm white-nose syndrome in bats. J Vet Diagn Invest 21:411–414 Minnis AM, Lindner DL (2013) Phylogenetic evaluation of Geomyces and allies reveals no close relatives of Pseudogymnoascus destructans, comb. nov., in bat hibernacula of eastern North America. Fungal Biol 117:638–649 Moore MS, Reichard JD, Murtha TD, Zahedi B, Fallier RM, Kunz TH (2011) Specic alterations in complement protein activity of little brown myotis (Myotis lucifugus) hibernating in whitenose syndrome affected sites. PLoS ONE 6:e27430 Moore MS, Reichard JD, Murtha TD, Nabhan ML, Pian RE, Ferreira JS, Kunz TH (2013) Hibernating little brown myotis (Myotis lucifugus) show variable immunological responses to white-nose syndrome. PLoS ONE 8:e58976 Mulec J, Covington E, Walochnik J (2013) Is bat guano a reservoir of Geomyces destructans? Open J Vet Med 03:161–167 Muller LK, Lorch JM, Lindner DL, O’Connor M, Gargas A, Blehert DS (2013) Bat white-nose syndrome: a real-time TaqMan polymerase chain reaction test targeting the intergenic spacer region of Geomyces destructans. Mycologia 105:253–259 Nilsson S (2012) Surveillance of Geomyces destructans in Swedish bats and bat hibernacula. SLU, Department of Biomedical Sciences and Veterinary Public Health, Upsalla, Sweden Paiva-Cardoso MdN, Morinha F, Barros P, Vale-Gonçalves H, Coelho AC, Fernandes L, Travassos P, Faria AS, Bastos E, Santos M, Cabral JA (2014) First isolation of Pseudogymnoascus destructans in bats from Portugal. Eur J Wildl Res 60:645–649 Pikula J, Bandouchova H, Novotny L, Meteyer CU, Zukal J, Irwin NR, Zima J, Martínková N (2012) Histopathology conrms white-nose syndrome in bats in Europe. J Wildl Dis 48:207–211 Puechmaille SJ, Verdeyroux P, Fuller H, Gouilh MA, Bekaert M, Teeling EC (2010) White-nose syndrome fungus (Geomyces destructans) in Bat, France. Emerg Infect Dis 16:290–293 Puechmaille SJ, Wibbelt G, Korn V, Fuller H, Forget F, Mühldorfer KM, Kurth A, Bogdanowicz W, Borel C, Bosch T, Cherezy T, Drebet M, Görföl T, Haarsma AJ, Herhaus F, Hallart G, Hammer M, Jungmann C, Le Bris Y, Lutsar L, Masing M, Mulkens B, Passior K, Starrach M, Wojtaszewski A, Zöphel U, Teeling EC (2011a) Pan-European distribution of white-nose syndrome fungus (Geomyces destructans) not associated with mass mortality. PLoS ONE 6:e19167 Puechmaille SJ, Fuller H, Teeling EC (2011b) Effect of sample preservation methods on the viability of Geomyces destructans, the fungus associated with white-nose syndrome in bats. Acta Chiropterol 13:217–221 Puechmaille SJ, Frick WF, Kunz TH, Racey PA, Voigt CC, Wibbelt G, Teeling EC (2011c) White-nose syndrome: is this emerging disease a threat to European bats? Trends Ecol Evol 26:570–576 Reeder DM, Turner GG (2008) Working together to combat white nose syndrome: a report of a meeting on 9–11 June 2008, in Albany, New-York. Bat Res News 49:75–78 Reeder D, Frank CL, Turner GG, Meteyer CU (2012) Frequent arousal from hibernation linked to severity of infection and mortality in bats with white-nose syndrome. PLoS ONE 7:e38920 Sachanowicz K, Stpie A, Ciechanowski M (2014) Prevalence and phenology of white-nose syndrome fungus Pseudogymnoascus destructans in bats from Poland. Central Eur J Biol 9:437–443 Simonovicová A, Pangallo D, Chovanová K, Lehotská B (2011) Geomyces destructans associated with bat disease WNS detected in Slovakia. Biologia 66:562–564 Storm JJ, Boyles JG (2011) Body temperature and body mass of hibernating little brown bats Myotis lucifugus in hibernacula affected by white-nose syndrome. Acta Theriol 56:123–127 Thogmartin WE, Sanders-Reed CA, Szymanski JA, McKann PC, Pruitt L, King RA, Runge MC, Russell RE (2013) White-nose syndrome is likely to extirpate the endangered Indiana bat over large parts of its range. Biol Conserv 160:162–172

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262 W.F. Frick et al. Thomas DW, Cloutier D (1992) Evaporative water loss by hibernating little brown bats, Myotis lucifugus. Physiol Zool 65:443–456 Thomas DW, Geiser F (1997) Periodic arousals in hibernating mammals: is evaporative water loss involved? Funct Ecol 11:585–591 Turner GG, Reeder DM (2009) Update of White Nose Syndrome in bats, September 2009. Bat Res News 50:47–53 Turner GG, Reeder DM, Coleman JTH (2011) A ve-year assessment of mortality and geographic spread of white-nose syndrome in North American bats, with a look to the future. Bat Res News 52:13–27 Veilleux JP (2008) Current status of white-nose syndrome in the Northeastern United States. Bat Res News 49:15–17 Verant ML, Boyles JG, Waldrep W, Wibbelt G, Blehert DS (2012) Temperature-dependent growth of Geomyces destructans, the fungus that causes bat white-nose syndrome. PLoS ONE 7:e46280 Verant ML, Meteyer CU, Speakman JR, Cryan PM, Lorch JM, Blehert DS (2014) White-nose syndrome initiates a cascade of physiologic disturbances in the hibernating bat host. BMC Physiol 14(1):10 Voyron S, Lazzari A, Riccucci M, Calvini M, Varese GC (2010) First mycological investigations on Italian bats. Hystrix Italian J Mammal 22:189–197 Warnecke L, Turner JM, Bollinger TK, Lorch JM, Misra V, Cryan PM, Wibbelt G, Blehert DS, Willis CKR (2012) Inoculation of bats with European Geomyces destructans supports the novel pathogen hypothesis for the origin of white-nose syndrome. Proc Natl Acad Sci 109:6999–7003 Warnecke L, Turner JM, Bollinger TK, Misra V, Cryan PM, Blehert DS, Wibbelt G, Willis CKR (2013) Pathophysiology of white-nose syndrome in bats: a mechanistic model linking wing damage to mortality. Biol Lett 9:20130177 Wibbelt G, Kurth A, Hellmann D, Weishaar M, Barlow A, Veith M, Prüger J, Görföl T, Grosche L, Bontadina F, Zöphel U, Hans-Peter S, Cryan PM, Blehert DS (2010) White-nose syndrome fungus (Geomyces destructans) in bats Europe. Emerg Infect Dis 16:1237 Wibbelt G, Puechmaille SJ, Ohlendorf B, Mühldorfer K, Bosch T, Görföl T, Passior K, Kurth A, Lacremans D, Forget F (2013) Skin lesions in European hibernating bats associated with Geomyces destructans, the etiologic agent of white-nose syndrome. PLoS ONE 8:e74105 Wilcox A, Warnecke L, Turner JM, McGuire LP, Jameson JW, Misra V, Bollinger TC, Willis CKR (2014) Behaviour of hibernating little brown bats experimentally inoculated with the pathogen that causes white-nose syndrome. Anim Behav 88:157–164 Willis CKR, Menzies AK, Boyles JG, Wojciechowski MS (2011) Evaporative water loss is a plausible explanation for mortality of bats from white-nose syndrome. Integr Comp Biol 51:364–373 Willis C, Wilcox A (2014) Hormones and hibernation: possible links between hormone systems, winter energy balance and white-nose syndrome in bats. Horm Behav 66:66–73 Zukal J, Bandouchova H, Bartonika T, Berkova H, Brack V, Brichta J, Dolinay M, Jaron KS, Kovacova V, Kovarik M, Martínková N, Ondracek K, Rehák Z, Turner GG, Pikula J (2014) White-nose syndrome fungus: a generalist pathogen of hibernating bats. PLoS ONE 9:e97224

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263Chapter 10Zoonotic Viruses and Conservation of BatsKarin Schneeberger and Christian C. Voigt© The Author(s) 2016 C.C. Voigt and T. Kingston (eds.), Bats in the Anthropocene: Conservation of Bats in a Changing World, DOIþ  10.1007/978-3-319-25220-9_10Abstractþ Many of the recently emerging highly virulent zoonotic diseases have a likely bat origin, for example Hendra, Nipah, Ebola and diseases caused by coronaviruses. Presumably because of their long history of coevolution, most of these viruses remain subclinical in bats, but have the potential to cause severe illnesses in domestic and wildlife animals and also humans. Spillovers from bats to humans either happen directly (via contact with infected bats) or indirectly (via intermediate hosts such as domestic or wildlife animals, by consuming food items contaminated by saliva, faeces or urine of bats, or via other environmental sources). Increasing numbers of breakouts of zoonotic viral diseases among humans and livestock have mainly been accounted to human encroachment into natural habitat, as well as agricultural intensication, deforestation and bushmeat consumption. Persecution of bats, including the destruction of their roosts and culling of whole colonies, has led not only to declines of protected bat species, but also to an increase in virus prevalence in some of these populations. Educational efforts are needed in order to prevent future spillovers of bat-borne viruses to humans and livestock, and to further protect bats from unnecessary and counterproductive culling. K. Schneebergerþ  ()þ  · C.C. Voigtþ  Department of Evolutionary Ecology, Leibniz Institute for Zoo and Wildlife Research, Berlin, Germany e-mail: schneeberger@izw-berlin.de

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264 K. Schneeberger and C.C. Voigt10.1þ IntroductionOver the past decades, the emergence of zoonotic viruses (those that are naturally transmitted between vertebrate animals and humans) from bats has been the subject of increasing attention from both scientists and the general public (e.g. Quammen 2013). During outbreaks of diseases in humans and livestock, bats are now often the primary focus of searches for a reservoir host (Chua etþ  al. 2002a; Leroy etþ  al. 2005; Li etþ  al. 2005; Halpin etþ  al. 2007; Towner etþ  al. 2007; Lau etþ  al. 2010; Wibbelt etþ  al. 2010; Memish etþ  al. 2013). Identication of bats as natural hosts for emerging viruses has important implications for bat conservation. We review the current state of research of four important families of emerging zoonotic viruses for which bats are natural reservoir hosts and discuss direct and indirect conservation implications.10.2þ Emerging Viral Diseases: Why Bats?Although bats have been identied as carriers of many highly virulent human pathogens (Chen etþ  al. 2014), evidence of pathogen-related clinical signs or disease in bats is scarce, particularly for intracellular pathogens such as viruses (Brook and Dobson 2015). Post-infection survival is supported by the frequent identication of antibodies to known viruses in apparently healthy bats and longterm survival of these bats (e.g. Hayman etþ  al. 2010). Additionally, viruses isolated or genetically detected from bat populations are highly diverse and often ancestral to related viruses in human and other mammalian species (e.g. Towner etþ  al. 2009; Drexler etþ  al. 2012; Baker etþ  al. 2013a; Tong etþ  al. 2013; Vidgen etþ  al. 2015). Together, these ndings suggest a long history of coevolution between many bat-virus relationships identied to date. Recent progress in the eld of bat immunology and genomics has identied key differences in bat immunity and physiology that evolved concomitantly with the evolution of ight, resulting not only in apparently increased immunotolerance of intracellular pathogens, but also in increased longevity and decreased tumour production (Baker etþ  al. 2013b; Zhang etþ  al. 2013; Brook and Dobson 2015). Immunotolerance and incomplete clearance of viral infections are also likely to favour the establishment of persistent infections (Virgin etþ  al. 2009), as proposed for a number of bat-borne viruses (Plowright etþ  al. 2015). Various ecological and life-history factors play a key role in the susceptibility of individuals and populations to pathogens (Allen etþ  al. 2009; Turmelle etþ  al. 2010; Schneeberger etþ  al. 2013), and notable differences exist between bats and terrestrial mammals such as rodents (Luis etþ  al. 2013). For example, the often high-population densities and the usually gregarious roosting behaviour of bats increase the likelihood of both intraand interspecies transmission of viruses (Luis etþ  al. 2013; Streicker etþ  al. 2010). Large-scale movements of bats due to

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265 10þ Zoonotic Viruses and Conservation of Batstheir ability for powered ight are also likely to facilitate viral transmission within and among species, including the exchange of novel viruses and virus variants across biomes or even continents (Calisher etþ  al. 2006; Epstein etþ  al. 2009; Peel etþ  al. 2013). The extreme relative longevity of bats compared to other mammals of similar size (Wilkinson and South 2002) and the potential for persistent and/or subclinical viral infections could further increase transmission potential (Calisher etþ  al. 2006). Reduction of body temperature associated with hibernation of temper ate zone bats lowers both viral activity and the metabolism of hosts, leading to increased incubation periods and therefore reduced likelihood of epizootic fadeout (of rabies, for example; George etþ  al. 2011). Bats are ancient mammals in evolutionary terms, and virus utilisation of highly conserved cellular receptors could facilitate transmission to other mammals (Calisher etþ  al. 2006), for example, as has been suggested for henipaviruses (Negrete etþ  al. 2005). Lastly, it was recently speculated that, similar to the febrile response of other mammals, the relatively high body temperature (about 38–41þ  °C) and metabolism of bats during ight may select for viruses tolerant to such conditions, meaning the normal febrile defence mechanism of other mammals is ineffective (“Flight as fever hypothesis”, O’Shea etþ  al. 2014), making bat-borne viruses potentially more virulent and lethal for other, non-ying mammals.10.3þ Zoonotic Viruses of Bats and Their SpillOver 10.3.1þ RhabdovirusesRabies virus (RABV) is the longest and best-known member of the genus Lyssavirus (family Rhabdoviridae) and still one of the most signicant zoonoses known from bats (recent reviews include: Banyard etþ  al. 2011; Banyard etþ  al. 2014 and Kuzmin 2014). The genus is rapidly expanding, with 14 of the currently recognised species (plus another known from genetic material only), and all but two (Mokola and Ikoma viruses) having been isolated from bats (Tableþ  10.1). Lyssaviruses spill over directly from bats to domestic animals, other wildlife and humans, or indirectly to humans via these other species. All lyssaviruses are potentially neurotropic, meaning that the virus infects nerve cells and replicates in the brain, resulting in clinical signs consistent with classical rabies (Schnell etþ  al. 2009). Although isolated from a variety of tissues and body uids in the late stages of infection, the predominant route of transmission is via saliva (mostly via biting; Kuzmin 2014). Lyssaviruses can be divided into two distinct “phylogroups” (Badrane etþ  al. 2001, Tableþ  10.1), reecting biological and genetic differences, and they are distributed globally in bats. Classical rabies virus occurs in bats across North, Central and South America (Messenger etþ  al. 2003; Banyard etþ  al. 2011) and was rst associated with vampire bats following an outbreak in cattle in South America in 1911 (Carini 1911). It is reported most frequently in the common

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266 K. Schneeberger and C.C. Voigt Tableþ  10.1þ Known lyssaviruses and their association with different bat species (adapted from Banyard etþ  al. 2014) Geographical distribution Lyssavirus species Phylogroup Bat species most commonly associated with lyssavirus infection Common name Known human cases The Americas Rabies virus (RABV) I Eptesicus fuscus Big brown bat Yes Tadarida brasiliensis Mexican/ Brazilian free-tail bat Lasionycteris noctivagens Silver-haired bat Perimyotis subavus Tri-coloured bat Desmodus rotundus Vampire bat Eurasia European bat lyssavirus type 1 (EBLV-1) I Eptesicus serotinus Serotine bat Yes European bat lyssavirus type 2 (EBLV-2) I Myotis daubentonii Daubenton’s bat Yes Bokeloh bat lyssavirus (BBLV) I Myotis nattereri Natterer’s bat No Aravan virus (ARAV) I Myotis blythi Lesser mouse-eared bat No Irkut virus (IRKV) I Murina leucogaster Greater tube-nosed bat Yes Khujand virus (KHUV) I Myotis mystacinus Whiskered bat No West Caucasian bat virus (WCBV) NAaMiniopterus schreibersii Common bent-wing bat No Lleida bat lyssavirus (LLEBV) NAaMiniopterus schreibersii Common bent-wing bat No Africa Duvenhage virus (DUVV) I Miniopterus sp? Undened Yes Nycteris thebaica Egyptian slit-faced bat Lagos bat virus (LBV) II Eidolon helvum Straw-coloured fruit bat No Rousettus aegyptiacus Egyptian fruit bat Epomorphorus wahlbergi Wahlberg’s epauletted fruit bat (continued)

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267 10þ Zoonotic Viruses and Conservation of Batsvampire bat (Desmodus rotundus; Kuzmin etþ  al. 2011a), which has a wide distribution across Mexico, Central America, and South America. Bites from this species appear to be responsible for the majority of human and domestic animal rabies infections of bat origin in South and Central America, with increased prey availability via expansion of livestock into new areas across the region hypothesised to be contributing to increasing incidences (Schneider etþ  al. 2009; Ruiz and Chávez 2010). In Canada and the USA, 51 cases of human rabies transmitted by non-haematophagous bats were recognised or inferred between 1951 and 2006 (mostly silver-haired bats (Lasionycteris noctivagans), eastern pipistrelle bats ( Perimyotis subavus) and Brazilian/Mexican free-tailed bats (Tadarida brasiliensis)) (Constantine and Blehert 2009; Banyard etþ  al. 2011). However, across the Americas, only 15þ  % of human rabies cases between 1993 and 2002 were reported as resulting from encounters with bats (Belotto etþ  al. 2005). Reported antibody prevalences against RABV in D. rotundus include 3–28þ  % in Peru (Streicker etþ  al. 2012) and 12þ  % in Brazil (Almeida etþ  al. 2011). Depending on the year, location and species, prevalence in other bats varies from relatively low 2þ  % in T. brasiliensis in New Mexico (Steece and Altenbach 1989) and 2.5þ  % in the little brown bats (Myotis lucifugus) in New York (Trimarchi and Debbie 1977), to 58þ  % in Seba’s short-tailed bat (Carollia perspicillata) in Peru (Salmón-Mulanovich etþ  al. 2009) and 67þ  % in T. brasiliensis in Texas (Baer and Smith 1991). As with other lyssaviruses discussed below, the potential for high antibody prevalences in bat populations and infrequent reports of mortality suggest that many individuals exposed to the virus survive, contrary to the over whelmingly lethal nature of lyssavirus infections in other mammalian species (reviewed in Banyard etþ  al. 2011). The mechanisms for this remain unclear.aLyssaviral phylogenies infer WCBV, LLEBV and IKOW that are more genetically distinct from other species, and they have not yet been assigned a phylogroup (Kuzmin 2014)bBarrett (2004) Tableþ  10.1þ (continued) Geographical distribution Lyssavirus species Phylogroup Bat species most commonly associated with lyssavirus infection Common name Known human cases Mokola virus (MOKV) II not detected Yes Shimoni bat virus (SHIBV) II Hipposideros commersoni Commerson’s leaf-nosed bat No Ikoma virus (IKOV) NAanot detected No Australasia Australian bat lyssavirus (ABLV) I Pteropus scapulatusbLittle red ying fox Yes Saccolaimus aviventris Yellow-bellied sheath-tailed bat Pteropus alecto Black ying fox

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268 K. Schneeberger and C.C. VoigtSeven bat lyssaviruses have been isolated in Eurasia (Tableþ  10.1). European bat lyssavirus type 1 and type 2 (EBLV-1 and EBLV-2; Bourhy etþ  al. 1992) are the most widely recognised and studied. Five fatal cases of human infections with EBLV have so far been reported, three from EBLV-1 (Roine etþ  al. 1988; Selimov etþ  al. 1989; Botvinkin etþ  al. 2005) and two from EBLV-2 (Lumio etþ  al. 1986; Fooks etþ  al. 2003; Nathwani etþ  al. 2003). Spillover of EBLV-1 into other mammals has also been observed, but rarely, with examples including zoo bats (Rønsholt etþ  al. 1998), sheep (Tjørnehøj etþ  al. 2006), domestic cats (Dacheux etþ  al. 2009) and a stone marten (Müller etþ  al. 2004). While EBLV-1 and EBLV-2 have been detected in a range of bat species (reviewed in Schatz etþ  al. 2013), they are most frequently associated with serotine bats (Eptesicus serotinus) and Daubenton’s bat (Myotis daubentonii), respectively. The dynamics of EBLV infections in their natural hosts is poorly understood, but banding and recapture data and the frequent capture of apparently healthy bats with antibodies against EBLV suggest that many bats survive infection (Serra-Cobo etþ  al. 2002; Amengual etþ  al. 2007; Schatz etþ  al. 2013). In cases where bats develop clinical symptoms of EBLV infection, the affected individuals are often unable to y, are generally weak and show abnormal behaviour, including attempts to bite (Banyard etþ  al. 2011). Experimental studies suggest that variable development of clinical signs may be related to inoculation route and dose (reviewed in Banyard etþ  al. 2011). Comparatively, little is known about the remaining Eurasian bat lyssaviruses, which have each been isolated from bats only once: West Caucasian bat virus (WCBV, Botvinkin etþ  al. 2003), Bokeloh bat lyssavirus (BBLV, Freuling etþ  al. 2011), Aravan virus (ARAV, Kuzmin etþ  al. 1991), Irkut virus (IRKV, Botvinkin etþ  al. 2003) and Khujand virus (KHUV, Kuzmin etþ  al. 2001), or is only known from partial genetic sequence data (Lleida virus, Ceballos etþ  al. 2013, Tableþ  10.1). Of these, only IRKV has been detected in other mammals (a human who developed rabies after a bat bite, Leonova etþ  al. 2009). WCBV appears to have a large geographical range. It was isolated from Miniopterus schreibersii in Russia, but cross-reactive antibodies have also been detected in Miniopterus bats in Kenya (Kuzmin etþ  al. 2008a). The relatively wide distribution and migratory behaviour of Miniopterus spp. may facilitate cross-continental transmission of this virus. Alternatively, given the close relationship between WCBV and Ikoma virus (IKOV), which was recently isolated in neighbouring Tanzania, the serological ndings from Kenya could in fact indicate exposure to IKOV or another related lyssavirus rather than WCBV (Marston etþ  al. 2012; Horton etþ  al. 2014). Similarly, serological surveys have detected antibodies against ARAV virus and KHUV virus in Indian ying foxes (Pteropus giganteus) from Bangladesh (Kuzmin etþ  al. 2006), and ARAV, KHUV, IRKV or Australian bat lyssavirus in Lyle’s ying foxes (P. lylei) and dawn bats (Eonycteris spelaea) from Thailand (Lumlertdacha etþ  al. 2005). Yet, given the limited lyssavirus surveillance in bats performed to date in this region and that individuals in these studies tested positive to multiple viruses, these results likely represent cross-reactivity of serological assays to unknown lyssaviruses.

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269 10þ Zoonotic Viruses and Conservation of BatsAfrica also hosts signicant lyssavirus diversity, with ve species identied, though only three of these isolated from bats to date (Tableþ  10.1). Duvenhage virus (DUVV, Meredith etþ  al. 1971) is the only phylogroup I lyssavirus in Africa and is more closely related to RABV, ABLV and the majority of the European species than other known African lyssaviruses. Since it was rst isolated from a human in 1970, two more fatal human infections of DUVV have been reported, one in South Africa in 2006 (Paweska etþ  al. 2006) and one from the Netherlands in 2007 after obtaining the infection in Kenya (van Thiel etþ  al. 2008). DUVV has been isolated from bats twice, once from a presumed M. schreibersii bat in South Africa and once from an Egyptian slit-faced bat (Nycteris thebaica) in Zimbabwe (Schneider etþ  al. 1985; Foggin 1988; Paweska etþ  al. 2006). No further information is so far available on this apparently rare African lyssavirus. In contrast, Lagos bat virus (LBV) is the most widely detected lyssavirus in Africa (Banyard etþ  al. 2011). In 1956, this virus was rst isolated from a strawcoloured fruit bat (Eidolon helvum; Boulger and Portereld 1958). Since then, the virus has been isolated and neutralising antibodies detected in a variety of fruit bat species, one insectivorous bat species, domestic cats, domestic dogs and a water mongoose, but not in humans (reviewed in Banyard etþ  al. 2011). E. helvum and Rousettus aegyptiacus are likely primary reservoir hosts for LBV, with seroprevalences ranging from 6 to 80þ  % and 29 to 46þ  %, respectively, depending on the region (Hayman etþ  al. 2008, 2012; Kuzmin etþ  al. 2008b; Dzikwi etþ  al. 2010; Peel etþ  al. 2013). LBV has been isolated from healthy, rabid and dead bats (reviewed in Banyard etþ  al. 2011), but longitudinal studies in Ghana (Hayman etþ  al. 2012) and surveys across continental Africa (Peel etþ  al. 2010, 2013) suggest widespread exposure, no difference in survival between seropositive and seronegative E. helvum, and viral persistence in very small, isolated island populations. Early infection experiments with LBV suggested that LBV and other phylogroup II viruses were less pathogenic than other lyssaviruses (Boulger and Portereld 1958; Badrane etþ  al. 2001). However, recent experimental infections indicated the potential for comparable mortality between LBV and RABV and indicated that signicant differences might instead exist between different LBV isolates (Kuzmin etþ  al. 2010; Markotter etþ  al. 2009). Of the other African lyssaviruses, only Shimoni bat virus (SHIBV) has been detected in bats (Commerson’s leaf-nosed bat (Hipposideros commersoni) in Kenya; Kuzmin etþ  al. 2010) and only Mokola virus (MOKV) has been detected in humans (on two occasions in Nigeria, Familusi and Moore 1972; Familusi etþ  al. 1972). MOKV has also been isolated from cats and small wild mammals, however, the natural reservoir host is unknown (Nel 2001). Ikoma virus was isolated from a rabid African civet (Civettictis civetta), but it is believed that the civet was a spillover host and the true reservoir host is yet to be identied (Horton etþ  al. 2014). The only lyssavirus detected in Australia to date—Australian bat lyssavirus (ABLV)—has two known lineages, one circulating in ying foxes and one in an insectivorous bat (Fraser etþ  al. 1996; Gould etþ  al. 2002; Warrilow 2005). In 1996, shortly after ABLV was rst isolated from a black ying fox (P. alecto) that was

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270 K. Schneeberger and C.C. Voigtunable to y (Fraser etþ  al. 1996), a 39-year-old woman died of clinical rabies after being bitten by a yellow-bellied sheath-tail bat (Saccolaimus aviventris; Gould etþ  al. 2002). Two subsequent human cases have been identied, a woman who died in 1998, 27þ  months after being bitten by a ying fox (Hanna etþ  al. 2000), and a child who died in 2014 after being scratched by a ying fox (Francis etþ  al. 2014). Experimental infection of grey-headed ying foxes (P. poliocephalus) with ABLV resulted in clinical signs of weakness, trembling and limb paralysis in three out of ten individuals (McColl etþ  al. 2002). As with other bat lyssaviruses, a small proportion of ABLV-positive bats succumb to encephalitis-like symptoms (Hooper etþ  al. 1997), yet serological tests show a high prevalence of antibodies in populations of surviving bats (McColl etþ  al. 2000).10.3.2þ ParamyxovirusesThe most notable viruses from the Paramyxoviridae family in bats are those of the genus Henipavirus, which are the subject of many reviews (e.g. Halpin and Rota 2015; Smith and Wang 2013, Luby and Gurley 2012; Clayton etþ  al. 2013; Middleton and Weingartl 2012; Field and Kung 2011). The rst recognised henipavirus, Hendra virus (HeV), was rst detected during an outbreak of infectious respiratory disease in horses and then humans in Hendra, Australia, in 1994 (Murray etþ  al. 1995). Ultimately, 13 of 20 infected horses died or were euthanised, and of two humans working closely with horses who became infected, one died from acute pneumonia (Murray etþ  al. 1995; Plowright etþ  al. 2015). This spillover was preceded a month earlier by another involving two horses and one human over 800þ  km away in Mackay, but which went unrecognised until 1995 (Rogers etþ  al. 1996; O’Sullivan etþ  al. 1997). An initial serological survey of 46 wildlife species (excluding bats) failed to identify a reservoir host; however, serological evidence of HeV infection was later identied in all four species of ying foxes native to Australia (Young etþ  al. 1996). Virus isolation (Halpin etþ  al. 2000) and experimental studies (Halpin etþ  al. 2011) have conrmed pteropodid bats as reser voir hosts for henipaviruses (with a lack of clinical signs), with evidence that black (P. alecto) and spectacled ying foxes (P. conspicillatus) are the main reservoir species for HeV (Smith etþ  al. 2014; Goldspink etþ  al. 2015). Because HeV is frequently detected in the urine of wild ying foxes (Smith etþ  al. 2014), the predominant transmission route to horses is likely via material recently contaminated with bat urine (e.g. pastures) or via direct transmission (Martin etþ  al. 2015). Recognised spillover events from bats to horses occurred sporadically from 1994 to 2004 and annually since 2006, with ve spillover events resulting in ongoing transmission to humans in close contact with horses (a total of seven human cases and four deaths; Field etþ  al. 2010). Spillover events are spatiotemporally clustered, occurring year-round in the northern tropics, but seasonally clustered in winter with a peak in July in subtropical regions (Plowright etþ  al. 2015).

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271 10þ Zoonotic Viruses and Conservation of BatsThe relative importance of various hypothesised drivers of HeV dynamics in bats and subsequent spillover to horses is still unclear (Plowright etþ  al. 2015). Nipah virus (NiV), the second henipavirus to be recognised, was rst isolated in 1999 from pigs and encephalitic pig workers in Malaysia (Center of Disease Control and Prevention 1999). NiV spillover has not been observed since this time in Malaysia; however, annual seasonal outbreaks with high case fatality (average 73þ  %) have occurred in people in Bangladesh since 2001 (Hsu etþ  al. 2004; Luby etþ  al. 2009; Luby and Gurley 2012), with occasional spillover also occurring in neighbouring India (Chadha etþ  al. 2006; Harit etþ  al. 2006). Due to the close relatedness of HeV and NiV, fruit bats were targeted, and serological evidence quickly identied them to be the natural reservoir of NiV (Enserink 2000; Yob etþ  al. 2001). This was subsequently supported by isolation of NiV from the urine of P. hypomelanus (Chua etþ  al. 2002a), P. vampyrus (Rahman etþ  al. 2010) and P. lylei (Reynes etþ  al. 2005), and seroconversion in the absence of clinical signs following experimental infections in P. vampyrus (Halpin etþ  al. 2011). Antibodies against NiV and NiV-related viruses have now been detected in a variety of bat species (including non-pteropid bats) across a wide geographical area (summarised in Breed etþ  al. 2013). NiV transmission to humans appears to occur via a wider variety of routes compared with HeV. Infection of domestic animal intermediate hosts (via consumption of salivaor urinecontaminated partially eaten fruits or raw date palm sap) has been implicated as a source of human infections in both Malaysia and Bangladesh (Chua etþ  al. 2002b; Chowdhury etþ  al. 2014). In Malaysia, human infections resulted from direct contacts with infected pigs (Chua etþ  al. 1999; Paton etþ  al. 1999; Parashar etþ  al. 2000), whereas in Bangladesh, transmission to humans regularly occurs via consumption of contaminated date palm sap (Luby etþ  al. 2006; Rahman etþ  al. 2012) or directly from human to human (e.g. via nursing sick individuals or preparation for burial; Hughes etþ  al. 2009). The risk of direct human infection with NiV from bats is considered to be lower than horizontal transmission once the virus enters the human population (Gurley etþ  al. 2007; Luby etþ  al. 2009; Chong etþ  al. 2003). A third henipavirus, Cedar Virus (Marsh etþ  al. 2012), has been isolated from urine collected under a mixed P. alecto/P. scapulatus roost in Australia. In contrast to HeV and NiV, however, it appears to be of low pathogenicity and failed to induce clinical signs in experimentally infected laboratory animal species (Marsh etþ  al. 2012). Serological evidence from South-East Asia and Australasia (Breed etþ  al. 2013) and the wide diversity of paramyxovirus sequences detected in Australia (Vidgen etþ  al. 2015) suggest more henipaviruses are yet to be found. Additionally, although henipaviruses were long thought to be restricted to Asia and Australia, antibodies cross-reactive to HeV and NiV were detected in Madagascar in 2007, suggesting a potentially wider geographical distribution of henipa-related paramyxoviruses (Iehlé etþ  al. 2007). This was supported by serological ndings and molecular detection of henipaor henipa-like viruses in mainland Africa and its offshore islands (Hayman etþ  al. 2008, 2012; Peel etþ  al. 2010, 2013; Drexler etþ  al. 2012). Indeed, a recent serological study indicates that these

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272 K. Schneeberger and C.C. Voigtviruses are also occasionally transmitted to humans in Africa (Pernet etþ  al. 2014), though no African henipavirus has been isolated to date. Viruses from the paramyxovirus genus Rubulavirus (a genus which includes the human mumps virus) have also been frequently detected in bats (Barr etþ  al. 2015). Menangle virus was isolated from pigs following the birth of unusually high numbers of stillborn and deformed piglets in Australia (Philbey etþ  al. 1998). Two piggery personnel had neutralising antibodies against Menangle virus after having recovered from an unexplained febrile illness (Philbey etþ  al. 1998). Flying fox colonies roosting in close proximity to the piggeries were a suspected source of infection for pigs, with subsequent transmission to humans (Philbey etþ  al. 1998). This was supported by serological evidence from P. poliocephalus, P. alecto and P. conspicillatus, and recent virus isolation from P. alecto (Barr etþ  al. 2012). Other isolated bat rubulaviruses with unknown or limited understanding of their zoonotic potential include Tioman virus from Malaysia (Chua etþ  al. 2001), Tuhokovirus 1, 2 and 3 from China (Lau etþ  al. 2010), Achimota virus 1 and 2 from Ghana (Baker etþ  al. 2013c) and Hervey, Grove, Teviot and Yeppoon paramyxoviruses from Australia (Barr etþ  al. 2015). Neutralising antibodies to Tioman virus and Achimota viruses have been detected in humans, suggesting previous exposure and infection with the virus (Yaiw etþ  al. 2007; Baker etþ  al. 2013c). Pigs experimentally infected with Tioman virus produced neutralising antibodies and excreted virus in saliva, but were either asymptomatic or developed only a fever (Yaiw etþ  al. 2008). Undetected infection in pigs could therefore facilitate transmission to humans. Finally, viral fragments related to rubulaviruses and the proposed genus Jeilongvirus have also been detected outside the range of fruit bats, in European insectivorous bat species (Kurth etþ  al. 2012). However, nothing is yet known about the relevance of these viruses as potentially zoonotic threats to humans.10.3.3þ CoronavirusesBat coronaviruses were rst identied from species of the genus Miniopterus (Poon etþ  al. 2005), however, with unknown zoonotic potential. The most prominent coronavirus, the one causing severe acute respiratory syndrome (SARS), was followed by a pandemic spread in humans after the rst outbreak in China in 2002 (Rota etþ  al. 2003). Soon after the outbreak, the virus was detected in masked palm civet (Paguma larvata) and raccoon dogs (Nyctereutes procyonoides) in a market in Guangdong Province, where SARS was rst reported (Guan etþ  al. 2003). A survey of common wildlife species in the area identied bats to be the natural reservoir of SARS coronavirus, with viruses from bats showing greater genetic diversity than the ones isolated from other species, including humans (Li etþ  al. 2005). Bats can regularly be found in markets in China, which makes direct transmission of the virus from bats to humans likely (Li etþ  al. 2005). The followed pandemic spread with 8096 conrmed cases of which 774 were fatal can be

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273 10þ Zoonotic Viruses and Conservation of Batsaccounted to rapid interindividual transmission of the virus once it entered the human population (World Health Organization 2003). Outside Asia, SARS-like coronaviruses have been detected in the lesser horseshoe bat (Rhinolophus hipposideros) from Europe (Rihtari etþ  al. 2010), in Chaerephon sp. from Kenya (Tong etþ  al. 2009) and in Hipposideros commersoni from Nigeria (Quan etþ  al. 2010). Antibodies against SARS coronavirus are present in various African bat species (Müller etþ  al. 2007). As with many newly detected viruses, their potential threat as a zoonotic disease is yet unclear. Since the outbreak of SARS in Asia has been traced to bats as natural hosts of the virus, the same was suspected to be the case for Middle East respiratory syndrome (MERS), an infection that has been occasionally spreading among humans of the Arabian peninsula since 2012 (Zaki etþ  al. 2012). Most human infections have been traced down to close contacts with dromedary camels (Camelus dromedarius), which carry a virus with a similar genome organisation as human MERS (Hemida etþ  al. 2014). There is at least one report of direct transmission of the virus from camels to humans via contact with infected animals (Memish etþ  al. 2014). However, a small fragment of a coronavirus PCRed from an Egyptian tomb bat (Taphozous perforatus) showed 100þ  % nucleotide identity to virus from the human index case-patient of MERS, suggesting that this species may be one of the putative natural reservoirs of the virus (Memish etþ  al. 2013). Bat-derived MERS virus has been shown to be able to use human receptors and thus could potentially infect human cells (Yang etþ  al. 2014). However, given the generally low prevalence of MERS virus in bat populations, a direct spillover from bats to humans is unlikely, and transmission probably happens mainly via camels as intermediate hosts (Memish etþ  al. 2013). In fact, no other bat has yet been found to carry MERS virus since the one reported by Memish and colleagues in 2013. The intensied search for viruses in bats worldwide has led to the detection of coronaviruses other than SARS and MERS, whose potential to be or become zoonotic has yet to be investigated (Woo etþ  al. 2006; Tang etþ  al. 2006; Dominguez etþ  al. 2007; Carrington etþ  al. 2008; Brandão etþ  al. 2008; Misra etþ  al. 2009; Pfefferle etþ  al. 2009; Donaldson etþ  al. 2010; Watanabe etþ  al. 2010; Drexler etþ  al. 2010; Falcón etþ  al. 2011; Annan etþ  al. 2013; Ge etþ  al. 2013; Anthony etþ  al. 2013; Ithete etþ  al. 2013). No clinical symptoms associated with infections with SARS-like and other coronaviruses have yet been described for bats.10.3.4þ FilovirusesEbola virus is the most prominent lovirus, causing severe haemorrhagic fever in humans with high mortality and fast spreading among African populations. The recent outbreak in 2013 in west Africa has resulted in the most severe epidemy of Ebola so far, with more than 11,000 lethal cases (as by September 2015; according to World Health Organization;http://apps.who.int/ebola/ebola-situation-reports).

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274 K. Schneeberger and C.C. VoigtAll Ebola outbreaks recorded until 2004 in Gabon and the Republic of the Congo have been linked to handling of gorilla, chimpanzee or duiker carcasses, species that can carry the Ebola virus (Leroy etþ  al. 2004; Pigott etþ  al. 2014). It has thus became apparent that spillover from animals to humans occurs through hunting, butchering and consumption of bushmeat (Gonzalez etþ  al. 2005; Li and Chen 2014; Chap. 12), followed by fast human-to-human transmission (World Health Organization 2014). An outbreak of Ebola in Congo in 2007 that resulted in 260 infected humans of whom 186 died has been traced to a potential direct transmission from a dead fruit bat that the rst human victim bought from hunters to eat (Leroy etþ  al. 2009). Antibodies against Ebola virus have since been detected in a total of 14 bat species, with seroprevalences of up to 44þ  % depending on species and location (Olival and Hayman 2014). Experimental infection of several bat species with Ebola led to high replication of the virus, but to no apparent signs of illness, suggesting that Ebola infections are subclinical in these species (Swanepoel etþ  al. 1996). One Eidolon helvum has survived for at least 13þ  months after being tested seropositive for Ebola virus and Lagos bat virus, indicating longterm survival of an individual bat following exposure to these viruses (Hayman etþ  al. 2010). The recent outbreak of Ebola in Guinea and neighbouring countries in 2013—countries that are at signicant distance to the previous outbreaks in central Africa—has caused speculations about a possible transmission of the virus by migrating fruit bats (Bausch and Schwarz 2014; Vogel 2014). However, as the strain of the west African Ebola virus is a genetic outlier within the known Ebola viruses, it has been argued that the west African variant may have emerged from local wildlife populations rather than from migrating individuals (Gatherer 2014). Furthermore, although speculated (Saéz etþ  al. 2015), it is yet not clear whether the spillover of Ebola virus in west Africa originated from bats. Marburg virus is the only lovirus that has so far been directly isolated from bats (Towner etþ  al. 2009; Amman etþ  al. 2012; Pourrut etþ  al. 2005). The rst outbreak of the virus was caused by a spillover from laboratory monkeys to humans in Marburg, Germany, in 1967 (Jacob and Solcher 1968). In 2007, mine workers in a cave in Uganda were diagnosed with Marburg haemorrhagic fever that potentially resulted from a spillover of the virus from a colony of Rousettus aegyptiacus, where 5.1þ  % of tested individuals carried the virus (Towner etþ  al. 2009). The high divergence of the genome sequence of Marburg in this population suggests a long-term association of the virus with the host, leading to the assumption that bats are the natural reservoir (Towner etþ  al. 2009). However, given that no other bat species has yet tested positive for the virus (Towner etþ  al. 2007), and seroprevalence being generally low in R. aegyptiacus (Pourrut etþ  al. 2009), spillovers from bats to humans may be rare events. The Reston Ebolavirus has rst been detected in 1989 in crab-eating macaques (Macaca fascicularis) imported from the Philippines to be used for animal testing in laboratories in Reston, USA (Jahrling etþ  al. 1990). During a second outbreak in 1990, animal handlers developed antibodies but did not get sick (Center for Disease Control and Prevention 1990). In 2008, Reston Ebolavirus was isolated from pigs in the Philippines (Marsh etþ  al. 2011), and soon after, some

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275 10þ Zoonotic Viruses and Conservation of Batssampled R. amplexicaudatus had antibodies against the virus, while 16 other bat species tested negative against Reston Ebolavirus (Taniguchi etþ  al. 2011). Screening for antibodies of the Ebola virus and Reston Ebolavirus in bats in Bangladesh has found seropositive R. leschenaultii, suggesting that these loviruses or related strains are distributed at a much larger geographic range than previously assumed (Olival etþ  al. 2013).10.4þ Main Conservation Issues Related to Bat Viruses 10.4.1þ Direct Effect: Viruses Killing BatsFrom all the viruses described above, only a few seem to affect bats. Although experimental infection with RABV leads to mortalities between 40 and 90þ  % depending on the bat species (Sétien etþ  al. 1998; Jackson etþ  al. 2008; Turmelle etþ  al. 2010), there are no observed mass mortalities in natural populations (Pawan 1959). The only virus that may be largely lethal for bats is the Lloviu virus, which is closely related to Ebola and Marburg virus, but not yet of zoonotic relevance. It was detected during investigations of a massive die-off of Miniopterus schreibersii in a cave in Spain (Negredo etþ  al. 2011). However, a causal connection between the detected virus and death of the bats has not yet been conrmed, and other bat species roosting in the same caves appeared to remain unaffected (Roué and Nemoz 2004). The lack of reports of viruses that are detrimental for bat health should not imply that viruses in general are not of importance for the conservation of bat populations. Similar to white-nose syndrome causing mass mortalities in North American bats (Frick etþ  al. 2010), newly emerging viruses may put local populations at threat. This may be especially the case if pathogens cross geographical borders and infect naïve bat populations. Pseudogymnoascus destructans—the causative fungus responsible for white-nose syndrome—likely originated from Europe, where it seemingly causes no bat fatalities, in contrast to North America (Puechmaille etþ  al. 2010; Frick etþ  al. 2010; Frick etþ  al. 2015, Chap. 9).10.4.2þ Indirect Effects: Biased Public PerceptionGenerally, the public perception of bats as aesthetically less appealing mammals as well as folklores that often associate bats with negative stigma makes batrelated conservation efforts time-consuming and demanding (Fenton 1997; Allen 2004; Knight 2008). The recent outbreaks of viral zoonotic diseases with the identication of bats as putative natural hosts have further complicated bat conservation efforts (Li etþ  al. 2005; Knight 2008). Following numerous and often lurid reports of fatal zoonotic diseases by the media, public perception of bats is mostly

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276 K. Schneeberger and C.C. Voigtskewed by fear and lack of information (Kingston 2016, Chap. 18). Therefore, it is important to highlight the context of bat-associated infections in order to provide more evidence-based information about the emergence and transmission of batrelated zoonotic diseases, which may lead to a more balanced reputation of bats. Depending on educational, cultural, legal and medial background of the targeted audience, specic aspects need to be taken into account. In Europe and North America, rabies is, so far, the only viral disease that is associated with bats. The fact that lyssaviruses are occasionally found in temper ate zone bats sometimes nds its way to the media, not always in favour of bats. Biased newspaper articles or press campaigns may result in the public misconception that bats are aggressive animals or that their mere presence can lead to human infections with these viruses. Although there are anecdotal reports of unprovoked attacks of bats on humans and dogs (Baer and Smith 1991), bats, as is the case of most mammals, usually only bite when handled or provoked. Furthermore, once bitten or scratched by a bat, immediate post-exposure vaccination can prevent a person from contracting rabies (see Sect. 10.5.2). In the case of the 37-year-old woman who died from a bat lyssavirus infection in Kenya, staff members of the health facility which the woman visited after being scratched by a bat were unaware of the possibility of rabies transmission (van Thiel etþ  al. 2009). Likewise, two persons in Europe who worked regularly with bats and died from rabies after being bitten and scratched by bats received neither prenor post-exposure treatment (Roine etþ  al. 1988; Nathwani etþ  al. 2003). These two cases triggered a Europe-wide serological screening effort involving more than 11,000 bats, with seroprevalences varying depending on the species and location (Racey etþ  al. 2012). EBLV-1 was most commonly detected in the serotine bat (Eptesicus serotinus), while EBLV-2 was very uncommon in all bat species. As a result, the public has been persuaded not to handle bats or to do so only with gloves and, in the case of bat workers, to receive preand/or post-exposure immunisation. Two fatal cases in which persons contracted rabies in Australia (Samaratunga etþ  al. 1998; Hanna etþ  al. 2000) triggered a similar campaign on this continent (Speare etþ  al. 1997, but see Francis etþ  al. 2014). Efcient education of medical professionals worldwide seems to be pivotal for implementing the correct treatment after scratches or bites from bats. In addition, vaccination should be mandatory for those who are frequently exposed to bats (Rupprecht and Gibbons 2004). Studies on animal models have shown that rabies vaccine also provides protection against other, although not all, lyssaviruses’ variants (Brookes etþ  al. 2005; Hanlon etþ  al. 2005). However, there is no known case of a person developing bat-associated rabies despite having been vaccinated, neither prenor post-exposure. Thus, getting infected by some sort of bat-related virus is unlikely in Europe and North America and decreases virtually to zero if people who experienced bat bites and scratches are treated appropriately. There is no case known for paramyxoviruses having spilled over to humans by direct contact with bats. An extensive serological survey among people frequently handling bats in Australia revealed no antibodies against Hendra virus (Arklay etþ  al. 1996). The virus apparently needs horses as amplier hosts, from where the virus can further be transmitted to persons in close contact with infected

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277 10þ Zoonotic Viruses and Conservation of Batsindividuals. Nevertheless, the outbreak of Hendra increased the unpopularity of ying foxes in Australia, making conservation of the four native species challenging (Thiriet 2011). Unlike Hendra, Nipah virus has likely been acquired by humans via consumption of contaminated date palm sap (Luby etþ  al. 2006; Rahman etþ  al. 2012), followed by person-to-person transmission (Gurley etþ  al. 2007). Although diseases associated with Hendra virus and Nipah virus have high mortality rates, the risk of infection for humans seems to be low (Chong etþ  al. 2003), and countermeasures may be taken in order to prevent future spillover events (see Sect. 10.5.2). MERS, just as Hendra virus, apparently needs livestock as an amplier host. In contrast to dromedaries (Hemida etþ  al. 2014), seroprevalence of MERS seems to be low in bats (Memish etþ  al. 2013), making direct transmission from bats to humans unlikely. As long as details on MERS infections in dromedaries and how to mitigate them are missing, it is hard to give recommendations to people who might be at risk. In contrast to MERS, the spillover of SARS into the human populations most likely happened via the wildlife market, either directly from a bat, or from other wildlife species. Likewise, the hunting, butchering and consumption of chimpanzees, gorillas and bats seem to have been sources of Ebola spillovers from wildlife to humans. The education of local communities needs to carefully balance information about the potential risk of acquiring infectious diseases by consuming bushmeat, without implying that bats need to be eradicated in order to prevent spillovers. The recent outbreak of Ebola resulting in several thousand human victims, and with bats frequently being reported as the likely source of origin, has undoubtly led to severe loss of reputation of bats on this continent, which makes the conservation of threatened populations and species even more challenging, not only in Africa, but also worldwide.10.4.3þ Indirect Effect—CullingThe direct persecution of bats often seems to be the most effective way to deal with bat-borne diseases to members of the public. Killing of bats has long been acceptable, even if they are protected (Chap. 14). Even though culling may be ofcially banned and thus not supported by authorities or governmental programs, large-scale killing of bats or the destruction of roost trees may still be commonly practiced in areas where zoonotic diseases are spreading. In Australia, for example, ying foxes are frequently harassed and killed, both legally (under permits issued by state wildlife management agencies) and illegally. This happened most prominently during periods when Hendra virus emerged in Australian ying fox populations (Roberts etþ  al. 2012). Half of the ying fox species native to Australia have declined about 30þ  % in population size during the last decade, and killing of bats usually does not lead to legal measures (Booth 2005). Furthermore, large-scale culling leads to a change of movement behaviour of bats, with new, susceptible individuals being recruited from nearby colonies

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278 K. Schneeberger and C.C. Voigt(Field 2009). Instead of reducing the viral prevalence, this may therefore lead to the exact opposite (see below). In the attempt to reduce rabies incidences, vampire bats are regularly culled in many parts of Latin America (Streicker etþ  al. 2012). In Brazil, for example, governmental programs are in action that involve targeted campaigns against vampire bats. During these measures, vampire bats are captured and poisoned or coated with anticoagulant and released, so that allogrooming kills their conspecics (Medellin 2003). Furthermore, bat roosts are destroyed using re and explosives (Mayen 2003), which also leads to dramatic declines of non-target bats (Furey and Racey 2015, Chap. 15). Besides the questionable methods involved, instead of reducing viral abundance in the population, culling of wildlife can lead to an increase in viral spreading. New hosts are recruited and the dispersal probability of infected individuals increases, which results in transmission of the disease to naïve hosts (Donnelly etþ  al. 2005; Choisy and Rohani 2006; Streicker etþ  al. 2012). This was the case for vampire bats in Peru, where culling failed to reduce seroprevalence of rabies in bat populations, but rather had the opposite effect (Streicker etþ  al. 2012). Therefore, persecution of bats as potential carriers of zoonotic diseases has been denounced as useless and even counterproductive by both conser vationists and experts on disease transmission (Hutson and Mickleburgh 2001; Knight 2008).10.4.4þ Indirect Effect—Killing of Bats for Virus SurveysIn the scope of recently emerging zoonotic diseases, the search for new batborne viruses has become a well-funded eld in the scientic community. While research is important to advance our understanding about the emergence of diseases and to possibly prevent further spillover events, the methods involved in these surveys are sometimes questionable from the perspective of bat conservation (Racey 2015). Some of the investigated bat species are listed as near threatened or vulnerable by the International Union for the Conservation of Nature (IUCN), with decreasing population sizes even in many species of least concern. While most surveillance studies that involve species of conservation concern use nonlethal methods such as antibody screening in blood (Hayman etþ  al. 2008; Young etþ  al. 1996; Lumlertdacha etþ  al. 2005; Wacharapluesadee etþ  al. 2005; Reynes etþ  al. 2005), others have involved the killing of a considerable number of bats of various conservation status (e.g. in Yob etþ  al. 2001; Kuzmin etþ  al. 2008b, 2010, 2011b; Dzikwi etþ  al. 2010 and Sasaki etþ  al. 2012). In order to limit such detrimental sur veys, the Food and Agriculture Organization of the United Nations (2011) has published a guideline for investigating the role of bats in emerging zoonotic diseases, including non-invasive protocols, which not only reduce the impact on bat populations, but also minimise the transmission risk of viral diseases. Such protocols have now been widely adopted, as for example by Ecohealth Alliance and other international research groups and networks.

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279 10þ Zoonotic Viruses and Conservation of Bats10.5þ Counter Measures in Favour of Bat Conservation 10.5.1þ Preventing the Emergence of New Viral DiseasesIn general, preventing the emergence of infectious diseases in wildlife populations is extremely challenging and usually underfunded, with only few practical suggestions being discussed (Daszak etþ  al. 2000). For example, it is important that translocations of animals across geographical borders need to follow strict guidelines in order to prevent the introduction of exotic pathogens in novel areas (e.g. Woodroofe 1999). Furthermore, an integration of knowledge about disease dynamics, as well as ecological and immunological aspects of the host, may contribute to a better understanding of emerging infectious diseases in wildlife species such as bats (Daszak etþ  al. 2000).10.5.2þ Educational EffortsAs many bat-borne viral diseases have high lethality rates for humans, preventing spillover events are of central importance. In particular, spillover by direct contact to bats, such as via bites or bat consumption, may bear severe risks to humans that could be minimised by educational programs (Kingston 2016, Chap. 18). Reducing the risk of outbreaks of zoonotic viruses may also lead to more positive attitudes towards bats, which may further be increased by highlighting their ecological importance as pollinators, seed dispersers and pest control for agriculture (Ghanem and Voigt 2012). Moreover, conservation measures that promote the preservation of bat habitats serve a dual role as they can decrease the contact zone between bats and humans, thus reducing the risk of spillover. As aforementioned vaccination against rabies and other lyssaviruses should be mandatory for persons working with bats and recommended for other people at risk. A signicant problem is that both preand post-exposure treatments are expensive and thus may not be readily available in developing countries, such as in Central and South America. Here, building houses in a bat-proof manner in order to avoid vampire bites during sleep and decreasing the risk of direct contact with other bats has so far been the best solution (Greenhall 1964; Voigt etþ  al. 2016, Chap. 14). A different issue is the transmission of Nipah viruses via consuming raw date palm sap contaminated by urine, faeces or saliva of bats (Luby etþ  al. 2006; Rahman etþ  al. 2012). Here, cooking the sap at temperatures above the level that viruses tolerate is an effective measure to prevent spillover (Hughes etþ  al. 2009). Additionally, preventing bats from accessing date palms and thus contaminating the sap has been proved to be both efcient and relatively cheap (Nahar etþ  al. 2010, 2013). The traditional “bamboo skirt” method for example uses inexpensive, recyclable bamboo to cover the part of the date palm where the sap is collected, preventing bats and

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280 K. Schneeberger and C.C. Voigtother vertebrates from getting access. Furthermore, in contrast to bird nets, this measure is non-lethal to the bats and therefore of high conservation value to local populations. However, such protective measures are reported to be rarely used in Bangladesh (Nahar etþ  al. 2010, 2013). This could potentially be changed by encouraging local farmers to use this method, emphasising its inexpensiveness and efciency while highlighting the reduced risk of acquiring Nipah virus disease. One of the key issues both for conservation and public health is the direct transmission of SARS and Ebola via wildlife markets. In South-East Asia, ying foxes are hunted regularly for the purpose of food (Mickleburgh etþ  al. 2002; Mildenstein etþ  al. 2016, Chap. 12), sometimes even authorised by the local Wildlife Department such as in Malaysia (Breed etþ  al. 2006). Likewise, fruit bats are consumed regularly throughout Africa (Mickleburgh etþ  al. 2009; Mildenstein etþ  al. 2016, Chap. 12). Since bats are suggested as potential reservoir for the recent outbreak of Ebola, Guinea banned bats for sale from markets (Gatherer 2014). Educational efforts to reduce the threat both to public health by zoonotic diseases and to the conservation of local bat populations are challenging, as they are usually impeded by the lack of understanding of entrenched cultural behaviours and social components (Pooley etþ  al. 2015; Kingston 2016, Chap. 18). In Ghana, for example, where the consumption of bats is part of the local culture and traditions, a survey revealed that knowledge about the ecological and economical value of bats would not make people refrain from killing and eating bats (Kamins etþ  al. Fig.þ  10.1þ Intact trees with colonies of Eidolon helvum (left) in Yaoundé, Cameroon, as compared to former roosting trees that have been cut (right) after bats were suspected to be the source of the recent Ebola outbreak in western Africa (photograph credits: Simon Ghanem)

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281 10þ Zoonotic Viruses and Conservation of Bats2014). Usually, the direct economic benet from selling hunted bats is more valuable to an individual person than the indirect, not always obvious economic value of bats, for example, for agriculture. However, about half of the hunters stated they would stop hunting bats if they could make them sick (Kamins etþ  al. 2014). This highlights the potential effectiveness of public education, but careful consideration is needed to avoid demonising bats in the process (Pooley etþ  al. 2015). The recent Ebola epidemic in western Africa for example has led to an increase in the persecution of bats, with roosts being destroyed and colonies being killed by communities (Fig.þ  10.1). Although preventing bats from being consumed may have higher priorities due to public health reasons, the culling of whole colonies as a likely result may be much more of a threat for the conservation of bats than the bushmeat trade (Pooley etþ  al. 2015).10.5.3þ Environmental ConservationCombining knowledge about the ecology of the host species as well as the disease dynamics of the virus may be crucial for establishing efcient disease prevention programs (e.g. Plowright etþ  al. 2015). Here, it needs to be noted that the emer gence of zoonotic diseases from bats also seems to be a consequence of anthropogenic alteration of natural environments (e.g. Daszak etþ  al. 2001). For example, in Central and South America, the conversion of forested habitats into pastures shifted the dominant food source for vampire bats from native vertebrates to livestock. This has increased rabies transmission from vampire bats to livestock and domestic animals in many parts of Latin America (Schneider etþ  al. 2009). Where bat habitats have been converted largely into agricultural farmland, the remaining bat populations are forced to concentrate in patches that provide them with resources they need. Flying foxes, for example, are highly sensitive to landscape modications, as they require large forested areas for foraging. Where natural habitats are scarce, ying foxes may use fruiting or owering trees in agricultural, suburban and urban areas, which increases the contact zone and spillover risk between bats and livestock or humans (Daszak etþ  al. 2006; Plowright etþ  al. 2015). Indeed, contact between bats and naïve hosts as a consequence of human landscape modication and encroachment likely sparked the transmission of Hendra viruses to horses (Epstein etþ  al. 2006) and Nipah virus to pigs (Chua etþ  al. 1999; Field etþ  al. 2001).10.5.4þ Conservation of Bat Populations and Population DynamicsRemoving individuals or colonies from regional populations, either by unsustainable hunting or culling, can cause an increase in relative local resource availability, creating regional gradients along which bats from other populations may move,

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282 K. Schneeberger and C.C. Voigtwhich may lead to an increase of virus movement (Field 2009). In Australia, for example, roosts that became empty after culling, disturbing or relocating colonies of ying foxes are usually reoccupied by immigrating individuals (Roberts etþ  al. 2012). Anthropogenic transformation of bat habitats in Australia has also been shown to lead to decreased migration in Pteropus bats, which can itself lead to a decline in population immunity (Plowright etþ  al. 2011). This could give rise to more viral shedding after local viral reintroduction, a mechanism that may be facilitated by urban habituation of fruit bat and the resulting increased contact with human and domestic animal populations (Epstein etþ  al. 2006; Plowright etþ  al. 2011). In Australia, all recently emerged bat-associated viruses—Hendra, Menangle and Australian bat lyssavirus—are hypothesised to be associated with habitat loss due to deforestation and agricultural intensication (Jones etþ  al. 2013). Therefore, protection of remaining natural habitats of bats along with farm management aiming at decreasing the contact zone between bats and livestock as well as education plans increasing awareness of environmental issues and safety may play a crucial role in the avoidance of future spillovers of bat-borne diseases to livestock and human populations, and promote further protection of local bat populations.10.6þ ConclusionBats harbour viruses that may become zoonotic. Circumstances facilitating spillover include direct contact with bats (bites, scratches, consumption of bats), contact with material contaminated by bat saliva, faeces or urine and amplication via intermediate hosts such as domestic animals or other wildlife species. Conservational actions are not only important to prevent spillovers, but also because emerging zoonotic viruses often lead to persecution of bats. In order to reduce the transmission risk of viruses from bats to human and livestock and to protect bat species at threat, educational efforts are needed. However, entrenched cultural and social components often act as barriers to efcient changes on how people think about and respond to bats. Whenever possible, educational efforts should be done in an informative, non-lurid way, presenting the facts rather than provoking additional fears to the already bad reputation of bats. Wherever possible, solutions should be found to enable the existence of bats in anthropogenic landscape, including the development of more affordable and readily available vaccinations (e.g. against rabies), and the reduction of potential contact between bats and humans and livestock. This however also includes that the natural habitats of bats need to be better protected to provide bat populations with sufcient space and to prevent range expansion into urban and suburban areas, where contact with humans and livestock may increase the risk of spillover events. Bat-borne viruses should be considered during bat conservation efforts, and it should be equally noticed that appropriate conservation measures may even reduce the risk of viral spillover from bat populations into human populations.

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283 10þ Zoonotic Viruses and Conservation of Bats Acknowledgmentsþ We thank Alison Peel, Paul Racey and an anonymous reviewer for constructive comments that helped to improve the manuscript. Open Access This chapter is distributed under the terms of the Creative Commons Attribution Noncommercial License, which permits any noncommercial use, distribution, and reproduction in any medium, provided the original author(s) and source are credited.ReferencesAllen GM (2004) Bats: biology, behavior, and folklore. Dover Publications Allen LC, Turmelle AS, Mendonça MT, Navara KJ, Kunz TH, McCracken GF (2009) Roosting ecology and variation in adaptive and innate immune system function in the Brazilian freetailed bat (Tadarida brasiliensis). J Comp Physiol B 179(3):315–323 Almeida MFD, Martorelli LFA, Sodré MM etþ  al (2011) Rabies diagnosis and serology in bats from the State of São Paulo, Brazil. Rev Soc Bras Med Trop 44(2):140–145 Amengual B, Bourhy H, López-Roig M, Serra-Cobo J (2007) Temporal dynamics of European bat Lyssavirus type 1 and survival of Myotis myotis bats in natural colonies. PLoS ONE 2(6):e566 Amman BR, Carroll SA, Reed ZD etþ  al (2012) Seasonal pulses of Marburg virus circulation in juvenile Rousettus aegyptiacus bats coincide with periods of increased risk of human infection. PLoS Pathog 8:e1002877 Annan A, Baldwin HJ, Corman VM etþ  al (2013) Human betacoronavirus 2c EMC/2012–related viruses in bats, Ghana and Europe. Emerg Infect Dis 19(3):456 Anthony S, Ojeda-Flores R, Rico-Chávez O etþ  al (2013) Coronaviruses in bats from Mexico. J Gen Virol 94(Pt 5):1028–1038 Arklay A, Selvey L, Taylor R, Gerrard J (1996) Screening of bat carers for antibodies to equine morbillivirus. Commun Dis Intell 20:477 Badrane H, Bahloul C, Perrin P, Tordo N (2001) Evidence of two Lyssavirus phylogroups with distinct pathogenicity and immunogenicity. J Virol 75(7):3268–3276 Baer GM, Smith J (1991) Rabies in nonhematophagous bats. Nat Hist Rabies 2:105–120 Baker KS, Leggett RM, Bexeld NH etþ  al (2013a) Metagenomic study of the viruses of African straw-coloured fruit bats: detection of a chiropteran poxvirus and isolation of a novel adenovirus. Virology 441(2):95–106 Baker ML, Schountz T, Wang LF (2013b) Antiviral immune responses of bats: a review. Zoonoses Public Health 60(1):104–116 Baker KS, Todd S, Marsh GA etþ  al (2013c) Novel, potentially zoonotic paramyxoviruses from the African straw-colored fruit bat Eidolon helvum. J Virol 87(3):1348–1358 Banyard AC, Hayman D, Johnson N, McElhinney L, Fooks AR (2011) Bats and lyssaviruses. Adv Virus Res 79:239–289 Banyard AC, Evans JS, Luo TR, Fooks AR (2014) Lyssaviruses and bats: emergence and zoonotic threat. Viruses 6(8):2974–2990 Barr JA, Smith C, Marsh GA, Field H, Wang LF (2012) Evidence of bat origin for Menangle virus, a zoonotic paramyxovirus rst isolated from diseased. J Gen Virol 93:2590–2594 Barr J, Smith C, Smith I etþ  al (2015) Isolation of multiple novel paramyxoviruses from pteropid bat urine. J Gen Virol 96:24–29 Barrett JL (2004) Australian bat lyssavirus. Ph.D. thesis, University of Queensland, Brisbane Bausch DG, Schwarz L (2014) Outbreak of Ebola virus disease in Guinea: where ecology meets economy. PLoS Neglected Trop Dis 8(7):e3056 Belotto A, Leanes L, Schneider M, Tamayo H, Correa E (2005) Overview of rabies in the Americas. Virus Res 111(1):5–12

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290 K. Schneeberger and C.C. Voigt Peel AJ, Sargan DR, Baker KS etþ  al (2013) Continent-wide panmixia of an African fruit bat facilitates transmission of potentially zoonotic viruses. Nat Commun 4:2770 Pernet O, Schneider BS, Beaty SM etþ  al (2014) Evidence for henipavirus spillover into human populations in Africa. Nat Commun 5:5342 Pfefferle S, Oppong S, Drexler JF etþ  al (2009) Distant relatives of severe acute respiratory syndrome coronavirus and close relatives of human coronavirus 229E in bats, Ghana. Emerg Infect Dis 15:1377–1384 Philbey AW, Kirkland PD, Ross AD etþ  al (1998) An apparently new virus (family Paramyxoviridae) infectious for pigs, humans, and fruit bats. Emerg Infect Dis 4:269–271 Pigott DM, Golding N, Mylne A etþ  al (2014) Mapping the zoonotic niche of Ebola virus disease in Africa. Elife 3:e04395 Plowright RK, Foley P, Field HE etþ  al (2011) Urban habituation, ecological connectivity and epidemic dampening: the emergence of Hendra virus from ying foxes (Pteropus spp.). Proc R Soc B 278:3703–3712 Plowright RK, Eby P, Hudson PJ etþ  al (2015) Ecological dynamics of emerging bat virus spillover. Proc R Soc B 282(1798):20142124 Pooley S, Fa JE, Nasi R (2015) No conservation silver lining to Ebola. Conserv Biol 29(3):965–967 Poon L, Chu D, Chan K etþ  al (2005) Identication of a novel coronavirus in bats. J Virol 79:2001–2009 Pourrut X, Kumulungui B, Wittmann T etþ  al (2005) The natural history of Ebola virus in Africa. Microbes Infect/Inst Pasteur 7:1005–1014 Pourrut X, Souris M, Towner JS etþ  al (2009) Large serological survey showing cocirculation of Ebola and Marburg viruses in Gabonese bat populations, and a high seroprevalence of both viruses in Rousettus aegyptiacus. BMC Infect Dis 9:159 Puechmaille SJ, Verdeyroux P, Fuller H, Gouilh MA, Bekaert M, Teeling EC (2010) White-nose syndrome fungus (Geomyces destructans) in bat, France. Emerg Infect Dis 16(2):290 Quammen D (2013) Spillover: animal infections and the next human pandemic. WW Norton & Company Quan P-L, Firth C, Street C etþ  al (2010) Identication of a severe acute respiratory syndrome coronavirus-like virus in a leaf-nosed bat in Nigeria. MBio 1:e00208–e00210 Racey PA, Hutson AM, Lina PHC (2012) Bat rabies, public health and european bat conservation. Zoonoses Public Health (Special Issue—Bats). 60:58–68 Racey PA (2015) The uniqueness of bats. In: Lin-Fa W, Cowled C (eds) Bats and Viruses: A New Frontier of Emerging Infectious Diseases. Wiley Blackwell. p. 1–22 Rahman SA, Hassan SS, Olival KJ etþ  al (2010) Characterization of Nipah virus from naturally infected Pteropus vampyrus bats, Malaysia. Emerg Infect Dis 16(12):1990 Rahman MA, Hossain MJ, Sultana S etþ  al (2012) Date palm sap linked to Nipah virus outbreak in Bangladesh, 2008. Vector-Borne Zoonotic Dis 12:65–72 Reynes J-M, Counor D, Ong S etþ  al (2005) Nipah virus in Lyle’s ying foxes, Cambodia. Emerg Infect Dis 11:1042–1047 Rihtaric D, Hostnik P, Steyer A, Grom J, Toplak I (2010) Identication of SARS-like coronaviruses in horseshoe bats (Rhinolophus hipposideros) in Slovenia. Arch Virol 155:507–514 Roberts BJ, Catterall CP, Eby P, Kanowski J (2012) Long-distance and frequent movements of the ying-fox Pteropus poliocephalus: implications for management. PLoS ONE 7:e42532 Rogers RJ, Douglas IC, Baldock FC etþ  al (1996) Investigation of a second focus of equine mor billivirus infection in coastal Queensland. Aust Vet J 74(3):243–244 Roine R, Hillbom M, Valle M etþ  al (1988) Fatal encephalitis caused by a bat-borne rabies-related virus. Clinical ndings. Brain: a journal of neurology 111:1505–1516 Rønsholt L, Sørensen K, Bruschke C etþ  al (1998) Clinically silent rabies infection in (zoo) bats. Vet Rec 142:519–520 Rota PA, Oberste MS, Monroe SS etþ  al (2003) Characterization of a novel coronavirus associated with severe acute respiratory syndrome. Science 300:1394–1399 Roué SY, Nemoz M (2004) Unusual mortality in Schreiber’s long ngered bat, at several nurser ies. Wild Dis Assoc Newsl 14:4

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291 10þ Zoonotic Viruses and Conservation of Bats Ruiz M, Chávez CB (2010) Rabies in Latin America. Neurol Res 32:272–277 Rupprecht CE, Gibbons RV (2004) Prophylaxis against rabies. New Engl J Med 351:2626–2635 Saéz AM, Weiss S, Nowak K, Lapeyre V, Zimmermann F etþ  al (2015) Investigating the zoonotic origin of the West African Ebola epidemic. EMBO Mol Med 7(1):17–23 Salmón-Mulanovich G, Vásquez A, Albújar C, Guevara C (2009) Human rabies and rabies in vampire and nonvampire bat species, Southeastern Peru, 2007. Emerg Inf Dis 15:1308–1310 Samaratunga H, Searle J, Hudson N (1998) Non-rabies Lyssavirus human encephalitis from fruit bats: Australian bat Lyssavirus (pteropid Lyssavirus) infection. Neuropathol Appl Neurobiol 24:331–335 Sasaki M, Setiyono A, Handharyani E etþ  al (2012) Molecular detection of a novel paramyxovirus in fruit bats from Indonesia. Virol J 9:240 Schatz J, Fooks AR, McElhinney L etþ  al (2013) Bat rabies surveillance in Europe. Zoonoses Public Health 60(1):22–34 Schneeberger K, Czirják GÁ, Voigt CC (2013) Measures of the constitutive immune system are linked to diet and roosting habits of neotropical bats. PLoS ONE 8:e54023 Schneider L, Barnard B, Schneider H etþ  al (1985) Application of monoclonal antibodies for epidemiological investigations and oral vaccination studies. In: Rabies in the tropics. Springer, Verlag Schneider MC, Romijn PC, Uieda W etþ  al (2009) Rabies transmitted by vampire bats to humans: an emerging zoonotic disease in Latin America? Rev panam de salud pública 25:260–269 Schnell MJ, McGettigan JP, Wirblich C, Papaneri A (2009) The cell biology of rabies virus: using stealth to reach the brain. Nat Rev Microbiol 8:51–61 Selimov MA, Tatarov AG, Botvinkin AD etþ  al (1989) Rabies-related Yuli virus; identication with a panel of monoclonal antibodies. Acta Virol 33(6):542–546 Serra-Cobo J, Amengual B, Abellán C, Bourhy H (2002) European bat lyssavirus infection in Spanish bat populations. Emerg Inf Dis 8:413–420 Sétien A, Brochier B, Tordo N etþ  al (1998) Experimental rabies infection and oral vaccination in vampire bats (Desmodus rotundus). Vaccine 16:1122–1126 Smith I, Wang LF (2013) Bats and their virome: an important source of emerging viruses capable of infecting humans. Curr Opin Virol 3(1):84–91 Smith C, Skelly C, Kung N, Roberts B, Field H (2014) Flying-fox species density-a spatial risk factor for Hendra virus infection in horses in Eastern Australia. PLoS ONE 9(6):e99965 Speare R, Skerratt L, Foster R etþ  al (1997) Australian bat lyssavirus infection in three fruit bats from north Queensland. Commun Dis Intell 21:117–119 Steece R, Altenbach JS (1989) Prevalence of rabies specic antibodies in the Mexican free-tailed bat (Tadarida brasiliensis mexicana) at Lava Cave, New Mexico. J Wildl Dis 25:490–496 Streicker DG, Turmelle AS, Vonhof MJ etþ  al (2010) Host phylogeny constrains cross-species emergence and establishment of rabies virus in bats. Science 329(5992):676–679 Streicker DG, Recuenco S, Valderrama W etþ  al (2012) Ecological and anthropogenic drivers of rabies exposure in vampire bats: implications for transmission and control. Proc R Soc B 279:3384–3392 Swanepoel R, Leman PA, Burt FJ etþ  al (1996) Experimental inoculation of plants and animals with Ebola virus. Emerg Inf Dis 2:321–325 Tang X, Zhang J, Zhang S etþ  al (2006) Prevalence and genetic diversity of coronaviruses in bats from China. J Virol 80:7481–7490 Taniguchi S, Watanabe S, Masangkay JS etþ  al (2011) Reston Ebolavirus antibodies in bats, the Philippines. Emerg Infect Dis 17:1559–1560 Thiriet D (2011) Conservation shouldn’t be a popularity contest. Conversation:1–4 Tjørnehøj K, Fooks AR, Agerholm JS, Rønsholt L (2006) Natural and experimental infection of sheep with European bat lyssavirus type-1 of Danish bat origin. J Comp Pathol 134(2):190–201 Tong S, Conrardy C, Ruone S etþ  al (2009) Detection of novel SARS-like and other coronaviruses in bats from Kenya. Emerg Inf Dis 15:482–485

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292 K. Schneeberger and C.C. Voigt Tong S, Zhu X, Li Y etþ  al (2013) New world bats harbor diverse inuenza A viruses. PLoS Pathog 9(10):e1003657 Towner JS, Pourrut X, Albariño CG etþ  al (2007) Marburg virus infection detected in a common African bat. PLoS ONE 2:e764 Towner JS, Amman BR, Sealy TK etþ  al (2009) Isolation of genetically diverse Marburg viruses from Egyptian fruit bats. PLoS Pathog 5:e1000536 Trimarchi C, Debbie J (1977) Naturally occurring rabies virus and neutralizing antibody in two species of insectivorous bats of New York State. J Wildl Dis 13:366–369 Turmelle A, Jackson F, Green D, McCracken G, Rupprecht C (2010) Host immunity to repeated rabies virus infection in big brown bats. J Gen Virol 91:2360–2366 van Thiel P, Van den Hoek J, Eftimov F etþ  al (2008) Fatal case of human rabies (Duvenhage virus) from a bat in Kenya: The Netherlands, December 2007. Euro Surveill 13:118 van Thiel P-PA, de Bie RM, Eftimov F etþ  al (2009) Fatal human rabies due to Duvenhage virus from a bat in Kenya: failure of treatment with coma-induction, ketamine, and antiviral drugs. PLoS Negl Trop Dis 3:e428 Vidgen ME, de Jong C, Rose K etþ  al (2015) Novel paramyxoviruses in Australian ying-fox populations support host-virus coevolution. J Gen Virol 96(7):1619–1625 Virgin HW, Wherry EJ, Ahmed R (2009) Redening chronic viral infection. Cell 138(1):30–50 Vogel G (2014) Are bats spreading ebola across sub-saharan Africa? Science 344:140 Voigt, CC, Phelps KL, Aguirre L, Schoeman MC, Vanitharani J, Zubaid A (2016) Bats and buildings: the conservation of synanthropic bats. In: Voigt, CC, Kingston, T (eds) Bats in the Anthropocene: conservation of bats in a changing world. Springer International AG, Cham, pp. 427–453 Wacharapluesadee S, Lumlertdacha B, Boongird K etþ  al (2005) Bat Nipah virus, Thailand. Emerg Inf Dis 11:1949–1951 Warrilow D (2005) Australian bat lyssavirus: a recently discovered new rhabdovirus. In: Fu ZF (ed) The world of Rhabdoviruses. Springer, Berlin Watanabe S, Masangkay JS, Nagata NK etþ  al (2010) Bat coronaviruses and experimental infection of bats, the Philippines. Emerg Inf Dis 16:1217–1223 Wilkinson GS, South JM (2002) Life history, ecology and longevity in bats. Aging Cell 1(2):124–131 Woodroffe R (1999) Managing disease threats to wild mammals. Anim Conserv 2(03):185–193 World Health Organisation (2003) Summary of probable SARS cases with onset of illness from 1 November 2002 to 31 July 2003 (http://www.who.int/csr/sars/country/table2004_04_21) World Health Organization (2014) Ebola virus disease, Fact Sheet. (http://www.who.int/mediace ntre/factsheets/fs103/en) Wibbelt G, Moore MS, Schountz T, Voigt CC (2010) Emerging diseases in Chiroptera: why bats? Biol Lett 6:438–440 Woo PC, Lau SK, Li KS etþ  al (2006) Molecular diversity of coronaviruses in bats. Virology 351:180–187 Yaiw KC, Crameri G, Wang L etþ  al (2007) Serological evidence of possible human infection with Tioman virus, a newly described paramyxovirus of bat origin. J Infect Dis 196:884–886 Yaiw KC, Bingham J, Crameri G etþ  al (2008) Tioman virus, a paramyxovirus of bat origin, causes mild disease in pigs and has a predilection for lymphoid tissues. J Virol 82:565–568 Yang Y, Du L, Liu C etþ  al (2014) Receptor usage and cell entry of bat coronavirus HKU4 provide insight into bat-to-human transmission of MERS coronavirus. Proc Nat Acad Sci USA 111:12516–12521 Yob JM, Field H, Rashdi AM etþ  al (2001) Nipah virus infection in bats (order Chiroptera) in peninsular Malaysia. Emerg Inf Dis 7:439–441 Young PL, Halpin K, Selleck PW etþ  al (1996) Serologic evidence for the presence in Pteropus bats of a paramyxovirus related to equine morbillivirus. Emerg Inf Dis 2:239–240 Zaki AM, Van Boheemen S, Bestebroer TM etþ  al (2012) Isolation of a novel coronavirus from a man with pneumonia in Saudi Arabia. New Engl J Med 367:1814–1820 Zhang G, Cowled C, Shi Z etþ  al (2013) Comparative analysis of bat genomes provides insight into the evolution of ight and immunity. Science 339(6118):456–460

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Part IIIHuman-Bat Conicts

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295Chapter 11Impacts of Wind Energy Development on Bats: A Global PerspectiveEdward B. Arnett, Erin F. Baerwald, Fiona Mathews, Luisa Rodrigues, Armando Rodríguez-Durán, Jens Rydell, Rafael Villegas-Patraca and Christian C. Voigt© The Author(s) 2016 C.C. Voigt and T. Kingston (eds.), Bats in the Anthropocene: Conservation of Bats in a Changing World, DOIþ  10.1007/978-3-319-25220-9_11Abstractþ Wind energy continues to be one of the fastest growing renewable energy sources under development, and while representing a clean energy source, it is not environmentally neutral. Large numbers of bats are being killed at utilityscale wind energy facilities worldwide, raising concern about cumulative impacts of wind energy development on bat populations. We discuss our current state of knowledge on patterns of bat fatalities at wind facilities, estimates of fatalities, mitigation efforts, and policy and conservation implications. Given the magnitude and extent of fatalities of bats worldwide, the conservation implications of under standing and mitigating bat fatalities at wind energy facilities are critically impor tant and should be proactive and based on science rather than being reactive and arbitrary. E.B. Arnettþ  ()þ  Department of Natural Resource Management, Texas Tech University, Lubbock, TX, USA e-mail: earnett@trcp.org E.F. Baerwaldþ  Department of Biological Sciences, University of Calgary, Calgary, Canada F. Mathewsþ  College of Life and Environmental Sciences, University of Exeter, Exeter, UK L. Rodriguesþ  Instituto da Conservação da Natureza e das Florestas, Lisbon, Portugal A. Rodríguez-Duránþ  Universidad Interamericana, Bayamón, Puerto Rico J. Rydellþ  Biology Department, Lund University, 223 62 Lund, Sweden

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296 E.B. Arnett et al.11.1þ IntroductionDeveloping renewable energy alternatives has become a global priority, owing to long-term environmental impacts from the use of fossil fuels, coupled with a changing climate (Schlesinger and Mitchell 1987; McLeish 2002; Inkley etþ  al. 2004) and because of growing concerns about negative effects from the use of nuclear power (Voigt etþ  al. 2015a). Wind power is one of the fastest growing renewable energy sources worldwide (Fig.þ  11.1), in part due to recent costcompetitiveness with conventional energy sources, technological advances, and tax incentives (Bernstein etþ  al. 2006). Although presently wind power contributes only about 4þ  % of the global electricity demand, some countries provide greater than 20þ  % of their demand from wind (e.g., Denmark [34þ  %] and Spain and Portugal [21þ  %]; World Wind Energy Association, www.wwindea.org). By the end of 2013, the Global Wind Energy Council reported that 318,105þ  MW of wind power capacity was installed worldwide (http://www.gwec.net/wp-content/uploads/2014/04/5_17-1_global-installed-wind-power-capacity_regionaldistribution.jpg). The World Wind Energy Association (http://www.wwindea.org) projects that by 2020, more than 700,000þ  MW could be installed globally. Wind energy development is not environmentally neutral, and impacts to wildlife and their habitats have been documented and are of increasing concern. Wind energy development affects wildlife through direct mortality and indirectly through impacts on habitat structure and function (Arnett etþ  al. 2007; Arnett 2012; NRC 2007; Strickland etþ  al. 2011). Bats are killed by blunt force trauma or barotrauma and may also suffer from inner ear damage and other injuries not readily noticed by examining carcasses in the eld (Baerwald etþ  al. 2008; Grodsky etþ  al. 2011; Rollins etþ  al. 2012; Fig.þ  11.2). Kunz etþ  al (2007a) proposed several hypotheses that may explain why bats are killed and some of these ideas have subsequently been discussed by others (e.g., Cryan and Barclay 2009; Rydell etþ  al 2010a). Collisions at turbines do not appear to be chance events, and bats probably are attracted to turbines either directly, as turbines may resemble roosts (Cryan 2008), or indirectly, because turbines attract insects on which the bats feed (Rydell etþ  al. 2010b). Horn etþ  al. (2008) and Cryan etþ  al. (2014) provide video evidence of possible attraction of bats to wind turbines. Regardless of causal mechanisms, bat fatalities raise serious concerns about population-level impacts because bats are long-lived and have exceptionally low reproductive rates, and their population growth is relatively slow, which R. Villegas-Patracaþ  Unidad de Servicios Profesionales Altamente Especializados, Instituto de Ecología, Coatepec, Veracruz, Mexico C.C. Voigtþ  Department of Evolutionary Ecology, Leibniz Institute for Zoo and Wildlife Research, Berlin, Germany

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297 11 Impacts of Wind Energy Development on Bats … limits their ability to recover from declines and maintain sustainable populations (Barclay and Harder 2003 ). Additionally, other sources of mortality cumulatively threaten many populations. For example, white-nosed syndrome causes devastat ing declines in bat populations in the USA and Canada (e.g., Frick et al. 2010 ), and national programs for improving insulation of buildings, particularly in Northern Europe, cause losses of roosting opportunities for bats such as the com mon pipistrelle ( Pipistrellus pipistrellus ; Voigt et al. 2016 ). Thus, high wind tur bine mortality poses a serious threat to bats unless solutions are developed and Fig. 11.1 Annual installed global wind energy capacity (MW) from 1996–2013 (modied from the Global Wind Energy Council, http://www.gwec.net/global-gures/graphs/ ) Fig. 11.2 Blunt force trauma ( a ) and barotrauma ( b , c ) in three noctule bats ( Nyctalus noctula ) killed at wind turbine in Germany. a Ventral view of an open fracture of the left humerus at the height of the elbow joint. b Ventral view of the opened abdominal cavity with blood effusion in the thoracic cavity visible behind the diaphragm (hemothorax). c Ventral view of opened car cass without bone fractures, but severe bleeding in the abdominal cavity (hemoabdomen) (picture courtesy: Gudrun Wibbelt, IZW)

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298 E.B. Arnett et al.implemented (Arnett and Baerwald 2013). In this chapter, we build on previous reviews of existing information (e.g., Arnett etþ  al. 2008; Rydell etþ  al. 2010a; Arnett and Baerwald 2013; EUROBATS 2014), synthesize information on bat fatalities at wind energy facilities worldwide, discuss unifying themes and policy and conser vation implications, and offer insights for future directions of research and mitigation of bat fatalities at wind facilities.11.2þ Composition and Estimates of Bat FatalitiesWe present information on estimates of bat fatalities as reported in published literature or publically available reports, but caution that studies had varying levels of effort, used different estimators (e.g., Huso 2011; Korner-Nievergelt etþ  al. 2013) and different methods to quantify bias (Arnett etþ  al. 2008; Strickland etþ  al. 2011), thus biasing estimates. Also, most estimators fail to adequately account for unsearched area near turbines (Huso and Dalthorp 2013), which further biases estimates. Some studies report fatalities/turbine and others fatalities/MW of installed capacity. As such, data presented here offer a general and relative sense of fatalities within and among continents and do not represent quantitative comparisons.11.2.1þ North AmericaFrom 2000 to 2011 in the USA and Canada, annual bat fatality rates were highest at facilities located in the Northeastern Deciduous Forest (6.1–10.5þ  bats/MW; Fig.þ  11.3) and Midwestern Deciduous Forest-Agricultural (4.9–11.0þ  bats/MW) regions dened by Arnett and Baerwald (2013: 438). Average fatality rate in the Fig.þ  11.3þ Wind energy facilities on forested ridges in the eastern USA have consistently documented high fatality rates of bats (photograph by E.B. Arnett)

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299 11þ Impacts of Wind Energy Development on Bats …Great Plains region was moderately high (6 bats/MW, 95þ  % CI: 4.0–8.1 bats/MW), while the Great Basin/Southwest Desert region (1.0–1.8þ  bats/MW) consistently reports the least variable and lowest fatality rates for bats (Arnett etþ  al. 2008; Arnett and Baerwald 2013; Johnson 2005). Wind energy facilities in this region occur in habitats generally offering few roosting resources, possibly (but untested) poor foraging opportunities, and may not be in migratory pathways, thus render ing these sites less risky to bats (Arnett and Baerwald 2013). However, facilities in other regions report high fatality rates of bats where there are large expanses of prairie and agricultural lands with few roosting resources, foraging opportunities, and likely migratory routes (e.g., Summer view Alberta, Canada, 8–14.6 bats/MW; Baerwald etþ  al. 2008). Thus, current patterns in the Great Basin/Southwest region reported by Arnett and Baerwald (2013) may simply reect biased reporting and an absence of evidence as opposed to evidence of absence (Huso and Dalthorp 2013). Twenty-one of the 47 species of bats known to occur in the USA and Canada have been reported killed at wind energy facilities, and fatalities are skewed toward migratory species often referred to as “tree bats” that include hoary bats (Lasiurus cinereus; 38þ  %), eastern red bats (Lasiurus borealis; 22þ  %), and silverhaired bats (Lasionycteris noctivagans; 18.4þ  %) that comprise a total of 78.4þ  % of the recovered bat turbine fatalities in the USA and Canada (Arnett and Baerwald 2013). However, other species also are affected, sometimes seriously. Fatalities of the cave-living Brazilian free-tailed bats (Tadarida brasiliensis) are quite frequent in the southern USA during the maternity period in summer (Miller 2008; Piorkowski and O’Connell 2010). In the USA, two species listed as threatened or endangered under the Endangered Species Act also have been killed by turbines, the Indiana bat (Myotis sodalis) and Hawaiian hoary bat (Lasiurus cinereus semotus; Arnett and Baerwald 2013). In the Oaxacan Isthmus region of Mexico, 32 of the 42 species of bats known to occupy this region (García-Grajales and Silva 2012; Briones-Salas etþ  al.þ  2013) were found killed (Villegas-Patraca etþ  al. 2012). These bats belonged to ve different families (Mormoopidae, Molossidae, Vespertilionidae, Phyllostomidae, and Emballonuridae), although 52þ  % of the fatalities belonged to just two species, Davy’s naked-backed bat (Pteronotus davyi; 40.2þ  %) and the ghost-faced bat (Mormoops megalophylla; 11.9þ  %), both of the family Mormoopidae. These two species are particularly abundant in the area studied and form colonies with thousands of individuals in caves (García-Grajales and Silva 2012). Both are aer ial-hawking and relatively fast-ying bats (Bateman and Vaughan 1974; Adams 1989). Also, unlike those species killed most frequently in Holarctic regions of North America, these species do not tend to roost in trees. Ninety-seven percent of bat fatalities found at wind turbines are resident species. This differs consider ably from the USA, Canada, and parts of northern Europe, suggesting that wind turbines are equally dangerous to resident cave bats assumed to be non-migratory as to migratory tree-roosting species. The common theme is rather that the most frequently killed species are adapted to ight and echolocation in the open air (e.g., bats that have a relatively high wing loading).

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300 E.B. Arnett et al.11.2.2þ EuropeRydell etþ  al. (2010a) synthesized data from 41 sites in 5 countries in northwestern Europe and found that the Black Forest region in Germany (nþ  þ  10) had the highest annual fatality rates, averaging 10.5þ  bats killed/MW. Some regions in Germany had relatively low estimated annual fatality rates, averaging around 1.1–1.2þ  bats/ MW (Rydell etþ  al. 2010a), yet some of these studies did not control for carcass removal and searcher efciency. The single comprehensive study that covered most parts of Germany did take the aforementioned eld biases into account when estimating annual fatality rates of 10–12þ  bats per wind turbines, translating to 6–8þ  bats per MW produced (Korner-Nievergelt etþ  al. 2013). Studies from mostly agricultural areas of Austria (nþ  þ  3), Switzerland (nþ  þ  3), and England (nþ  þ  1) yielded mean annual fatalities rates of 2.5, 5.3, and 0.6þ  bats killed/MW, respectively (Rydell etþ  al. 2010a). In France, some particularly dangerous sites are located near water along the river Rhone in the east (Dubourg-Savage etþ  al. 2011) and on the Atlantic coast in the west (Rydell etþ  al. 2010a). In Spain, bat fatalities from 56 wind facilities ranged from 0.00 to 0.80 bats/MW per year (Camina 2012), but most studies did not correct for scavenger removal and searcher biases and therefore underestimate fatalities. In Portugal, annual fatality rates at 28 facilities ranged from 0.07 to 11.0/MW (L. Rodrigues, Instituto da Conservação da Natureza e das Florestas, unpublished data). Generally, data from Europe are inconsistently collected, rendering comparisons and generalizations across countries difcult. Nevertheless, it is clear that bats are frequently killed at wind tur bines throughout the continent, with some facilities experiencing considerably higher fatality rates relative to others. Members of EUROBATS recently synthesized data from several countries and reported 6429 documented bat kills of 27 species collected at wind facilities in Europe from 2003 to 2014 (EUROBATS 2014), but some studies used to derive estimates of fatality rates did not incorporate eld bias or area corrections. The species of bats found most frequently at wind facilities across northern Europe were the common pipistrelle, common noctule (Nyctalus noctula), Nathusius’ pipistrelle (Pipistrellus nathusii), and Leisler’s bat (Nyctalus leisleri). In Germany, nearly 70þ  % of recorded deaths represent the latter three species and the particolored bat (Vespertilio murinus), all of which are long-distance migrants (Hutterer etþ  al. 2005). Owing to its central geographical location on the European continent, Germany appears to provide ecological stepping stones for many longdistance bat migrants from northeastern populations (Steffens etþ  al. 2004; Voigt etþ  al. 2012). However, resident species or short-distance migrants, including common pipistrelle and northern bats (Eptesicus nilssonii), also are frequently killed in northern Europe (Rydell etþ  al. 2010a). The majority (>90þ  %) of bats killed at wind turbines in southern Europe belong to the various pipistrelle and noctule species: common pipistrelle, Nathusius’ pipistrelle, soprano pipistrelle (Pipistrellus

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301 11þ Impacts of Wind Energy Development on Bats …pygmaeus), Kuhl’s pipistrelle (Pipistrellus kuhlii), and Savi’s pipistrelles (Hypsugo savii) and the common noctule, giant noctule (Nyctalus lasiopterus) and Leisler’s bat (Nyctalus leisleri). Some of these are long-distance migrants (e.g., Nathusius’ pipistrelle and common noctule) that often roost in tree holes, while others are resident and usually house-living species that do not migrate long distances (e.g., Kuhl’s pipistrelle and Savi’s pipistrelle). Rare species, such as the barbastelle (Barbastella barbastellus) and the Myotis and Plecotus spp., also are killed occasionally, but in smaller numbers. Thus, bats killed at wind turbines in southern Europe generally belong to the same genera as those in northern Europe (Pipistrellus and Nyctalus spp.), but include several non-migratory species such as Kuhl’s and Savi’s pipistrelles.11.2.3þ AfricaLittle work has been done on wind energy facilities in Africa, and prior to 2012, no studies had been published from the continent. During a pilot study at a single turbine located in the Eastern Cape of South Africa, Doty and Martin (2012) found 18 carcasses of 2 species of bats—the Cape serotine (Neoromicia capensis) and Egyptian free-tailed bat (Tadarida aegyptiaca). No estimates of fatality rates were provided, likely because of small sample size of recovered carcasses and no bat carcasses were used during eld bias trials. In the Western Cape of South Africa, Aronson etþ  al. (2013) reported only one carcass of a Cape serotine. These studies conrm at least some species of bats are vulnerable to wind turbine mortality in South Africa, which could have implications for ecosystem function and conservation of bats in this region.11.2.4þ New Zealand and AustraliaIn Australia, Hall and Richards (1972) were the rst to report bat fatalities at a wind facility in the world, and 22 white-striped free-tailed bats (Tadarida australis) were found over a 4-year period. Little work had been done in the region since this pioneering discovery, until Hull and Cawthen (2012) surveyed two wind facilities in Tasmania, where they recorded 54 bat fatalities from two species, Gould’s wattled bats (Chalinolobus gouldii) and an unknown Vespadelus sp. More recently, Bennett (2012) found white-striped free-tailed bats at two turbines located in Victoria. While no estimates of fatality rates were provided for these studies, they indicate that some species of bats are at risk of wind turbine mortality in this part of the world.

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302 E.B. Arnett et al.11.2.5þ South America, Central America, and the CaribbeanFew studies have been done in Latin American regions on bat fatalities caused by wind turbines. Puerto Rico hosts 13 species of bats of ve families. Five of these 13 species belong to the family Phyllostomidae, which feed on fruits and nectar and forage in the understory and canopy (Gannon etþ  al. 2005). It was originally speculated that these species would be at low risk for mortality caused by wind turbines based on their life histories and foraging patterns. Species in the family Molossidae also occur in Puerto Rico, and conversely, these species have been considered to be at higher risk to turbine collisions because they y high in open spaces. Species from both families of bats have been detected during preconstruction surveys in areas where wind facilities were proposed. Twenty months of ongoing post-construction surveys in Puerto Rico revealed 30 carcasses from 11 of the 13 species, for a corrected mortality rate of about 10 bats/turbine /year (Rodríguez-Durán, Universidad Interamericana, unpublished data). Aside from the expected mortality of species in the family Molossidae, it was surprising that fruit and nectar feeding species of phyllostomids were followed in number of fatalities given their ight and foraging patterns. One important hazard for bats in this region relates to their use of hot caves as roosts (Rodríguez-Durán 2009; Ladle etþ  al.þ  2012). Although little studied, these systems may be ubiquitous throughout parts of México, Panamá, Colombia, Venezuela, Brazil, and the Greater Antilles. Phyllostomids and mormoopids (family Mormoopidae) form large aggregations in hot caves and commute to foraging areas ying long distances at high altitude. This reliance on hot caves may place them at risk from wind facilities located near their feeding sites or along their commuting routes.11.2.6þ AsiaOn the island of Taiwan off the Chinese mainland, wind facilities have been established along the western coastline, predominantly in former mangrove wetlands. Bat fatalities have been recorded at three of these facilities (C.H. Chou, Endemic Species Research Institute, unpublished data). Carcass searches and acoustic monitoring indicated regular feeding activity of bats near turbines in summer, and 51 dead bats were found. However, the study is ongoing and no eld bias correction experiments have been conducted yet, so corrected fatality estimates are not available. The Japanese pipistrelle (Pipistrellus abramus), which is a non-migratory open-air foraging bat, was killed most frequently (nþ  þ  39). Six other species have also been found killed, although in smaller numbers (1–4 individuals for each species), namely Horikawa’s brown bat (Eptesicus serotinus horikawai), common house bat (Scotophilus kuhlii), Chinese noctule (Nyctalus plancyi velutinus), Taiwanese golden bat (Myotis formosus avus), a recently described mouse-eared bat (Myotis secundus), and Japanese long-ngered bat (Miniopterus fuliginosus).

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303 11þ Impacts of Wind Energy Development on Bats …Three other species have been observed foraging around the turbines, but have not yet been found during carcass searches. These species are the yellow-necked sprite (Arielulus torquatus), Taiwanese tube-nosed bat (Murina puta), and East Asian free-tailed bat (Tadarida insignis). Several of these species (e.g., yellownecked sprite, Taiwanese golden bat, Taiwanese tube-nosed bat, Chinese noctule, Horikawa’s brown bat, and M. secundus) are all island endemics, some of which occur in sparse and probably small and vulnerable populations. Nevertheless, the pattern conforms to that of most regions around the world, since the mortality predominantly (but not exclusively) affects species that feed in the open air (C.H. Chou, Endemic Species Research Institute unpublished data).11.2.7þ ConclusionsBats are killed at wind turbines worldwide, and those fatalities are not restricted to migratory species at high latitudes, as previously suggested (e.g., Kunz etþ  al. 2007a; Arnett etþ  al. 2008). Hence, the bias toward tree-roosting migrants observed in North America and to some extent also in northern Europe is not consistent elsewhere. An emerging hypothesis is that bats that regularly move and feed in less cluttered and more open air-space are most vulnerable to collisions with wind turbines, regardless of continent, habitat, migratory patterns, and roost prefer ences. The species most often killed at wind turbines throughout Europe belong to aerial-hawking and relatively fast-ying, open-air species, and this is consistent with the pattern found in North America and Mexico. However, other species, including gleaning insectivores and even fruit feeders, also are killed occasionally. The vulnerability of tropical bat faunas is a potentially serious problem that must be addressed immediately and preferably before extensive wind facilities are planned and constructed. While fatalities of endangered species like the Indiana bat are important from a legal perspective, they currently appear to be biologically irrelevant in comparison with those for hoary and eastern red bats, for example. However, fatalities of listed species worldwide may become increasingly important as wind energy development expands. The paucity of studies in most regions of the world is alarming, particularly in Mexico, Central and South America, the Caribbean, Africa, New Zealand, and Australia. Notably, we could not nd information on bat fatalities at wind facilities from mainland Asia, but the data from Taiwan indicate that the bat fauna of eastern Asia may be highly vulnerable at wind turbines. Turbine fatalities may be a serious threat to bats in, for example, China where wind energy development is substantial (Global Wind Energy Council, http://www.gwec.net/global-gures/graphs/#). This situation is further complicated by the fact that in most countries information gathered is sequestered either by wind energy companies or government agencies and not made readily available. The importance of having access to this information cannot be overstated for all regions of the world.

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304 E.B. Arnett et al.11.3þ Patterns of Bat Fatality 11.3.1þ Temporal PatternsIn the temperate Northern Hemisphere, most bat fatalities occur during late summer and early autumn. In the USA, fatalities peak in mid-July through early September in most parts of the country (Johnson 2005; Arnett etþ  al. 2008; Baerwald and Barclay 2011; Arnett and Baerwald 2013). Studies from Europe demonstrate a similar pattern (e.g., in Germany, where most (about 90þ  %) bat fatalities at wind turbines occur between mid-July and the end of September; Brinkmann etþ  al. 2011; Lehnert etþ  al. 2014). Some studies from northern Europe and North America demonstrate smaller peaks of fatalities during spring (Arnett etþ  al. 2008; Rydell etþ  al. 2010a). In Greece and on the Iberian Peninsula of Spain and Portugal, the pattern is similar, with most (>90þ  %) fatalities in late summer (Georgiakakis etþ  al. 2012; Camina 2012; Amorim etþ  al. 2012), but in some places, particularly at high elevation sites, fatalities occur from May to October and without any obvious concentration in the late summer period (DubourgSavage etþ  al. 2011; Camina 2012). Such consistent temporal patterns of fatality are helpful when predicting high-risk periods and applying some mitigation measures such as raising turbine cut-in speed (Arnett etþ  al. 2011, Baerwald etþ  al. 2009). Hull and Cawthen (2012) noted that fatalities predominantly occurred in autumn in Tasmania, where the climate is temperate. However, in the tropical Isthmus of Tehuantepec in Mexico, while 46þ  % of bat fatalities were found in the summer rainy season, no clear pattern in bat deaths associated with any season emerged. In summary, while there are clear temporal patterns and a distinct late summer fatality peak in high-latitude temperate regions (north Europe and North America), the pattern becomes less obvious in warmer climates at lower temperate latitudes (south Europe) and temporal patterns may dissipate entirely in tropical regions (e.g., southern Mexico).11.3.2þ Spatial PatternsArnett and Baerwald (2013) noted that the spatial context of bat kills, both among turbines within a facility and among different facilities, could be useful for developing mitigation strategies. They hypothesized that if, for example, kills were concentrated at specic turbines, then curtailment, removal, or relocating that turbine may reduce bat deaths. However, if fatalities are broadly distributed, then facilitywide mitigation strategies would be necessary (Arnett etþ  al. 2008). Thus far, studies worldwide have failed to detect specic turbines responsible for most fatalities at any given facility. Other patterns at scales beyond individual turbines have been reported that may assist with assessing risk. Baerwald and Barclay (2011) found no differences in

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305 11þ Impacts of Wind Energy Development on Bats …fatalities on the east vs. west side of a facility in southern Alberta, but the fatality rate was higher at the north end. Baerwald and Barclay (2011) hypothesized that because fall migrations are from north to south, higher fatality rates could be expected at the more northerly turbines rst encountered by migrating bats. At a landscape scale, Baerwald and Barclay (2009) found both higher activity and fatality rates of bats at wind facilities near the foothills of the Rocky Mountains as compared to eastward prairie grasslands. They speculated that turbine proximity to stopover and roost sites in foothills habitat signicantly increased fatality rates assuming that geographical landmarks are used for navigating migration routes and that bats judge nightly travel distances between suitable diurnal roosting sites.11.3.3þ Habitat RelationshipsRelationships between bat fatalities and habitat or topographic characteristics may be useful for developing mitigation strategies (e.g., to avoid placing turbines near places where many bats move or forage, such as near open water sources, wetlands, or known roosts; Arnett etþ  al. 2008; Arnett and Baerwald 2013; Rydell etþ  al. 2010a). Johnson etþ  al. (2004) did not nd a signicant relationship between the number of bat fatalities and any of the 10 cover types within 100þ  m of turbines at facilities in Minnesota or any relationship between fatalities and distance to near est wetland or woodlot. In assessing the type of vegetation present in areas where the fatalities were found in wind facilities in the Isthmus of Tehuantepec, 79.6þ  % occurred in agricultural areas. In Oklahoma in 2004, Piorkowski and O’Connell (2010) found that turbines in eroded ravine topography accounted for higher fatality rates than those in areas of low topographic relief and reported some evidence that turbines in mixed cedar/pasture habitats killed more bats than those in cropland and prairie habitats. However, these patterns were not repeated in 2005 or for both years of the study when combined, and Piorkowski and O’Connell (2010) speculated that bats may have exhibited different habitat use patterns in differ ent years or they did not measure factors better explaining annual differences they observed. Interestingly, Grodsky (2010) found that bat fatalities were actually lower near the Horicon Marsh in Wisconsin. Hull and Cawthen (2012) found no relationships between bat fatalities and proximity of turbines to the coast or vegetation. Hence, correlating high-risk locations with particular habitat types or topographic patterns has proven difcult and inconsistent. Analyses of fatalities reported from Spain and Portugal, where most wind facilities are located on top of hills and mountains, suggest that the most signicant environmental predictor of fatality rate is proximity to steep slopes with bare rock and no vegetation. Bare rock is warmed by the sun and radiates heat during the night, which likely facilities insect activity over the rocks (Ancilotto etþ  al. 2014), possibly explaining higher fatality at sites near steep, rocky slopes. Alternatively, rocks on tops of hills and mountains might provide suitable roosts.

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306 E.B. Arnett et al.Piorkowski and O’Connell (2010) documented the rst evidence of fatality of Mexican free-tailed bats at a North American wind facility that could be attributed to the site’s proximity (~15þ  km) to a large maternity colony. In Wisconsin, Grodsky (2010) found no relationship between distances of turbines from a large hibernaculum (Neda Mine), but in this case, hibernating bats did not belong to the species most vulnerable to wind turbine mortality (see above). Georgiakakis etþ  al. (2012) reported that the most frequently killed species at wind facilities in Greece exhibited different spatial patterns of fatality, speculating that this resulted from some turbines being located closer to roosts and/or commuting cor ridors. It may not be enough to consider the proximity of a facility to a maternity or hibernation site, but rather where it is located relative to feeding grounds or movement corridors (Arnett and Baerwald 2013). We are not aware of other studies demonstrating similar relationships or patterns with large maternity or winter roosts.11.3.4þ Climate and Weather VariablesArnett (2005) was rst to employ daily carcass searches and relate them to weather variables, discovering that most bats were killed on low-wind nights when power production appeared insubstantial. Based on this approach, Arnett etþ  al. (2008) estimated that 82–85þ  % of bat fatalities at two facilities in the eastern USA occurred on nights with median nightly wind speeds of <6þ  m/s. Since this pivotal discovery, studies worldwide document that most bat fatalities occur during low-wind periods. In the USA, for example, Jain etþ  al. (2011) found that maximum wind speeds when bat collisions likely occurred ranged from 2.4 to 5.3þ  m/s. Korner-Nievergelt etþ  al. (2013) found that maximum collision rates of bats occurred at wind speeds between 3.5 and 5.7þ  m/s. Several other studies from Europe demonstrate a similar pattern (e.g., Amorim etþ  al. 2012). Indeed, this consistency suggesting bat fatality is highest during lower wind speeds greatly assists predicting high-risk periods during which to apply operational mitigation. Fatalities appear to increase as ambient temperature rises, a relationship observed in North America (e.g., Grodsky 2010; Young etþ  al. 2011) and Europe (e.g., in Portugal; Amorim etþ  al. 2012). Amorim etþ  al. (2012) also found that bat fatalities increased with decreasing relative humidity. The effect of high temperature on fatality rate seems to apply both on the broader regional and climatic scales and according to daily changes in the weather (Dubourg-Savage etþ  al. 2011 and unpublished data). Hence, at least in southern Europe, high fatality rates at wind turbines are most likely in warm and dry geographic areas (Mediterranean and low elevation) and also in warm weather (most common in late summer). In the end, this suggests that fatalities may be correlated with periods of high insect activity, which generally is most likely to occur under warm and dry conditions (Heinrich 1993).

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307 11þ Impacts of Wind Energy Development on Bats …Bat fatalities also have been correlated with other climatic factors that could assist with predicting high-risk periods. Baerwald and Barclay (2011) reported that species–specic fatalities were affected by greater moon illumination. They also observed that falling barometric pressure and the number of deaths were cor related and that whereas fatalities of silver-haired bats increased with increased activity of this species, moon illumination, and south-easterly winds, hoary bat mortality increased most signicantly with falling barometric pressure. Interestingly, neither hoary bat activity nor fatality was inuenced by any measured variables other than falling barometric pressure (Baerwald and Barclay 2011). Again, this could result from decreasing barometric pressure that triggers insect ight activity and therefore may motivate foraging efforts among bats by indicating a potential increase in food availability (Wellington 2011).11.4þ Offshore Wind FacilitiesPotential impacts of offshore wind-energy development on bats are poorly under stood, although observations in Europe and anecdotal accounts of bats occurring offshore suggest that impacts may occur. Bats are known to regularly migrate across the Baltic and North Seas and visit offshore facilities (Hutterer etþ  al. 2005; Boshamer and Bekker 2008; Ahlén etþ  al. 2009; Poerink etþ  al. 2013; Rydell etþ  al. 2014). Ahlén etþ  al. (2009) recorded 11 species of bats ying and feeding over the sea up to 14þ  km from the shore. In spring and late summer, migrating bats are found along coastlines of the Baltic Sea and southeastern North Sea in northern Europe, including all offshore islands where observations have been made (Rydell etþ  al. 2014). This suggests bats, including Nathusius’ pipistrelles, soprano pipistrelles, and common noctules, migrate on a broad front across the Baltic Sea and along its coasts, using small islands for stopovers. Researchers in North America also have reported activity of bats in both near and offshore habitats, suggesting impacts are highly probable at facilities located in such places. Cryan and Brown (2007) discovered longitudinal movement by hoary bats from inland summer ranges to coastal regions during autumn and winter and suggested that coastal regions with non-freezing temperatures may be important wintering areas for hoary bats. Off the coast of Maryland, Johnson etþ  al. (2011) recorded ve species of bats, including eastern red bats, big brown bats (Eptesicus fuscus), hoary bats, tri-colored bats (Perimyotis subavus), and silver-haired bats, on a barrier island and concluded these species used this island during migration, which could have implications for wind energy development near and offshore. It seems likely that near and offshore wind facilities also will kill bats, but it is difcult or impossible to nd bat fatalities at sea and no attempts to assess offshore turbine bat fatality have been made to date. Arnett and Baerwald (2013) suggested that impacts of the rst several offshore wind-energy facilities proposed and built in North America, including those on inland waters such as the Great Lakes, be evaluated extensively both for fatalities and displacement effects.

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308 E.B. Arnett et al.They also suggested that a method for predicting fatalities at existing and planned wind facilities offshore will be required to understand impacts and develop mitigation strategies, because nding and retrieving dead birds and bats from water bodies will be a considerable challenge (Arnett etþ  al. 2007; Arnett 2012).11.5þ Estimating RiskKunz etþ  al. (2007b) found a positive correlation between post-construction bat activity and fatality from carcass searches conducted simultaneously. However, Kunz etþ  al. (2007b) warned of several limitations of their analysis and noted that it was unclear whether pre-construction call rates could predict risk and level of post-construction fatality rates. When comparing 5 sites with fatality and activity data, and tall turbines (towers 65þ  m), Baerwald and Barclay (2009) found a signicant positive relationship between post-construction activity and fatality at 5 wind facilities in Alberta. Amorim etþ  al. (2012) and Korner-Nievergelt etþ  al. (2013) also found increasing number of bat fatalities with increasing acoustic bat activity at facilities in Portugal and Germany, respectively. These studies correlating post-construction bat activity with fatality suggest that it may be possible to use indices of pre-construction bat activity to predict future fatality and, thus, risk and need for mitigation. However, while numerous studies have documented pre-construction activity of bats with hopes of inferring risk of collision mortality, these studies have yet to link with post-construction fatality data gathered from carcass searches. Hein etþ  al. (2013) were the rst to correlate pre-construction acoustic activity with post-construction fatalities from 12 paired study sites in the USA and found that no statistically signicant relationship existed between bat fatalities/MW and bat passes/detector night and only a small portion of the variation in fatalities was explained by activity. Thus, Hein etþ  al. (2013) concluded that prediction of risk prior to construction of a wind facility is highly variable and imprecise and acoustic data may not necessarily predict bat fatality in any reliable way. One explanation as to why correlations between pre-construction measurements of bat activity with similar measurements made post-construction or fatality estimates are weak could be that bats are attracted to the turbines once they are built and sites are used differently by at least some species (open-air bats) afterward (Horn etþ  al. 2008; Kunz etþ  al. 2007b; Arnett etþ  al. 2008; Cryan etþ  al. 2014). Theoretical estimations of exposure risk of bats to collisions with turbines based on models may also improve our understanding of factors inuencing fatality and the context of fatalities. Species distribution models developed in Italy suggest that 41þ  % of the region offers suitable foraging habitat for 2 species of bats vulnerable to wind turbines, Leisler’s bat and the common pipistrelle, and these same areas encompass over 50þ  % of existing or planned wind farms (Roscioni etþ  al. 2013). The authors believe fatality risk for these species is increased by the common proximity to forest edges, but this contradicts other ndings from southern Europe, suggesting the opposite relationship (Dubourg-Savage etþ  al. 2011).

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309 11þ Impacts of Wind Energy Development on Bats …Roscioni etþ  al. (2014) further investigated habitat connectivity as a surrogate for assessing risks of wind facilities to bat migration and commuting in Italy. Using species distribution models, they found that most corridors used by bats were concentrated in an area where existing (54þ  %) and planned (72þ  %) wind facilities would interfere with important corridors connecting the western and the eastern parts of the region. In Portugal, mortality risk models indicated wind farms located in humid areas with mild temperatures and within 600þ  m of steep slopes had higher probabilities of mortality (Santos etþ  al. 2013). They also demonstrated that high mortality risk areas overlapped greatly with the potential distribution of Leisler’s bat in Portugal, suggesting that populations of this species may be at high risk to turbine fatalities (Santos etþ  al. 2013). They also found that a large extent of the area predicted to be high risk for mortality overlapped with sites highly suitable for wind farm construction.11.6þ Cumulative ImpactsEstimates of fatalities, and thus any estimate of cumulative fatalities, are conditioned by eld methodology for each study (e.g., search interval) and how each study did or did not account for sources of eld sampling bias when calculating fatality rate estimates. Arnett and Baerwald (2013) synthesized information from 122 post-construction fatality studies (2000–2011) from 73 regional facilities in the USA and Canada and developed a regional weighted mean estimate of cumulative bat fatalities for the USA and Canada. Assuming fatality rates were (1) representative of all regional sites and (2) consistent from year to year without behavioral modication or mitigation, Arnett and Baerwald (2013) estimated cumulative bat fatalities in the USA and Canada ranged from 0.8 to 1.7 million over a 12-year period from 2000 to 2011. This estimate was projected to increase by 0.2–0.4 million bats in 2012 based on the assumptions and installed wind power capacity. Smallwood (2013) estimated 888,000 bats killed/year at wind facilities in the USA, while Hayes (2013) concluded that over 600,000 bats may have been killed by wind turbines in 2012 alone. However, neither of these estimates used all data available at the time they were published, nor did they weight their estimates by regionally collected data and installed wind energy capacity as Arnett and Baerwald (2013) did; the latter approach likely provides a more conservative and accurate estimate based on the studies and installed capacity from each region. When controlling for eld biases, an estimated 10–12 bats are killed annually at each wind turbine in Germany, if no mitigation measures have been implemented (Brinkmann etþ  al. 2011). Assuming these numbers are representative of all types of wind turbines for all of Germany, it has been suggested that more than 200,000 bats were killed at onshore wind turbines in Germany, assuming no behavioral modication or mitigation measures were practiced (Voigt etþ  al. 2015a). Over the past ten years of wind energy development, it is estimated that

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310 E.B. Arnett et al.more than two million bats may have been killed by wind turbines in Germany, based on the reported large-scale development of wind turbines in that country (Berkhout etþ  al. 2013; Voigt etþ  al. 2015a). Importantly, the context of wind turbine fatalities remains poorly understood, in part because little population data exist for most species of bats (O’Shea etþ  al. 2003) and this hinders understanding population-level impacts, as well as effectiveness of mitigation measures. Population estimates for most species of bats around the world are lacking, and some bat populations are suspected or known to be in decline (e.g., Frick etþ  al. 2010; Hutson etþ  al. 2001; Ingersoll etþ  al. 2013). Other populations, such as hibernating species in Europe, appear to be increasing (9 of 16 species examined by Van der Meij etþ  al. (2014) increased at their hibernation sites from 1993 to 2011), but these species are not largely affected by wind turbines. In addition to natural and other forms of anthropogenic-induced mortality, wind turbine mortality further compounds population declines for many species of bats and warrants mitigation.11.7þ Mitigating Bat MortalityAs reported previously, most bat fatalities occur during relatively low-wind condiþ­tions over a relatively short period of time in late summer (Arnett etþ  al. 2008) and operational adjustments under these conditions and during this time could reduce impacts on bats (Arnett 2005; Arnett etþ  al. 2008; Kunz etþ  al. 2007a). Behr and von Helversen (2006) were the rst to examine operational mitigation in Germany, documenting around 50þ  % fewer bats killed at turbines having their cut-in speed (wind speed at which turbines begin producing electricity into the power grid) raised above the set manufacture’s cut-in speed of 4.0þ  m/s. In the synthesis of operational mitigation studies in the USA and Canada, Arnett etþ  al. (2013a) reported that most studies documented at least a 50þ  % reduction in bat fatalities when turbine cutin speed was increased by 1.5þ  m/s above the manufacturer’s cut-in speed, with up to a 93þ  % reduction in bat fatalities in one study (Arnett etþ  al. 2011). Baerwald etþ  al. (2009) demonstrated benecial reductions (~60þ  %) with a low-speed idling approach. Young etþ  al. (2011) discovered that feathering turbine blades (pitched 90° and parallel to the wind) at or below the manufacturer’s cut-in speed resulted in up to 72þ  % fewer bats killed when turbines produced no electricity into the power grid. Arnett etþ  al. (2013a) noted that studies failing to demonstrate statistically signicant effects could be explained by lack of treatments being implemented during the study (i.e., winds were either too low or high to enable comparison of treatments). In Portugal, a mitigation study found that estimated mortality at turbine with raised cut-in speed was 0.3 bats/turbine compared to 1.6 bats/turbine at tur bines operating normally, which resulted in a 78.5þ  % reduction in bat fatalities assuming all turbines at the facilities had raised cut-in speed (LEA 2010). More recently, situation-dependent operation protocols, so-called algorithms, were developed for the operation of wind turbines. These algorithms consider a

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311 11þ Impacts of Wind Energy Development on Bats …number of parameters such as ambient temperature, wind speed, season, and time of day as well as recorded bat activities for dening a set of operation rules for wind turbines (Korner-Nievergelt etþ  al. 2013). However, these algorithms have been formulated for a single type of turbine and for a limited number of sites. Thus, the suggested algorithms may be unsuitable for other places with varied geographical and topographic characteristics, bat communities, and turbine types (Voigt etþ  al. 2015a). Few studies have disclosed actual power loss and economic costs of operational mitigation, but those that have suggest that <1þ  % of total annual output would be lost if operational mitigation was employed during high-risk periods for bat fatalities. While costs of lost power due to mitigation can be factored into the economics, nancing, and power purchase agreements of new projects, altering turbine operations even on a limited-term basis potentially poses difculties on existing projects. Although curtailment is relatively straightforward to implement on large modern turbines, for older models and for small to medium energy-generating tur bines, there often is no way to remotely control or change cut-in speed; some tur bines would require a technician to physically change turbine operating systems (which is not feasible). However, raising cut-in speed or altering blade angles to reduce rotor speed (termed “low-speed idling” by Baerwald etþ  al. 2009) where blades are near motionless in low wind speeds remain the only proven solutions to mitigating bat kills at wind facilities. The fact that it may be difcult to apply these mitigation techniques to some old turbines should not compromise its use on contemporary turbines. Other approaches to mitigating bat fatalities have been suggested, including projecting electromagnetic signals from small, portable radar units (Nicholls and Racey 2009) and ultrasonic broadcasts (Arnett etþ  al. 2013b). However, the for mer approach has not been tested at large, utility-scale facilities, and none are yet being implemented broadly at wind energy facilities. Future studies of any mitigation approach must demonstrate greater or equal effectiveness to operational adjustments and also be cost-competitive with different operational strategies for mitigation.11.8þ Conservation PolicyIn this section, we discuss a few selected issues regarding policy and regulation of wind facilities as they relate to wildlife impacts and successful integration of science, policy, and management to improve siting that minimizes risk to wildlife, including bats. This discussion is by no means exhaustive or comprehensive, but rather offers examples of policy issues from different regions of the world. In the USA, the federal government’s role in regulating wind power development is limited to projects occurring on federal lands, impacting federal trust species, or projects that have some form of federal involvement (e.g., interconnect with a federal transmission line) or require federal permits. The primary federal

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312 E.B. Arnett et al.regulatory framework for protecting wildlife from impacts from wind power includes three laws—the US Migratory Bird Treaty Act, the Bald and Golden Eagle Protection Act, and the Endangered Species Act (ESA; GAO 2005; NRC 2007). Because wind-energy development has primarily occurred on non-federal land, regulating such facilities is largely the responsibility of state and local governments (GAO 2005). The primary permitting jurisdiction for wind-energy facilities in many instances is a local planning commission, zoning board, city council, or county board of supervisors or commissioners, and typically, these local jurisdictional entities regulate wind projects under zoning ordinances and building codes (GAO 2005), often without the basic knowledge needed to make informed decisions. Additionally, each state may enforce its laws regarding wind energy and wildlife impacts or establish cooperative efforts to address impacts. The US Fish and Wildlife Service has voluntary guidelines designed to help wind energy project developers avoid and minimize impacts of land-based wind projects on wildlife and their habitats (U.S. Fish and Wildlife Service 2012). In the USA, most species of bats, including migratory tree-roosting species killed most frequently by tur bines, are not protected under federal, state, or provincial laws (Arnett 2012; Cryan 2011). Documented presence or fatality of species listed as threatened or endangered under the ESA (e.g., Indiana bat) does not necessarily mandate monitoring or mitigation as one might expect; rather, all efforts are voluntary even in cases involving a listed species, although threat of prosecution under the ESA increases when operators fail to collaborate or develop a conservation and mitigation plan. Until recently, an Environmental Impact Assessment (EIA) in Canada was required under the Canadian Environmental Assessment Act (CEAA) when a Federal Authority initiated a wind project, granted any form of nancial assistance or land for the project, and/or performed a regulatory duty in relation to the project, such as issuing a permit or license. Given that the Canadian Federal Government provided nancial incentives for wind energy from 2002 to 2011, EIAs of wind energy projects were generally mandatory. However, wind energy projects no longer require federal environmental assessments (Canadian Environmental Assessment Act 2012), but projects may still require an environmental assessment if requested by the province or territory. Bats fall under the jurisdiction of the individual provinces (ten) and territories (three). As such, there are no Canada-wide bat and wind-energy policies or regulations; each province or territory sets their own policy and/or regulation regarding bats and wind energy projects (e.g., Ontario Ministry of Natural Resources 2011; Government of Alberta 2013). The Ontario Ministry of Natural Resources set a mortality threshold of 10þ  bats/turbine/year, which if exceeded triggers operational mitigation across the wind facility from 15 July to 30 September for the duration of the project (Ontario Ministry of Natural Resources 2011). This mortality threshold was based on fatality rates of bats at wind energy projects in Ontario and across North America. Ontario’s guidelines do not explicitly consider cumulative effects (i.e., operational mitigation is only triggered by project-specic fatality rates). In Alberta, the integration of data, including acoustic data, collected during both the preand postconstruction monitoring, helps guide the mitigation framework (Government of

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313 11þ Impacts of Wind Energy Development on Bats …Alberta 2013). For example, if less than ve migratory-bat passes/detector night are recorded during pre-construction acoustic monitoring, then the project is considered to be a potentially acceptable risk, but if greater than ten migratory-bat passes/detector night were recorded, the project is considered to have a potentially high risk of bat fatalities and will likely require operational mitigation (Government of Alberta 2013). Unlike Ontario’s guidelines, Alberta’s mitigation framework explicitly considers cumulative effects (i.e., the proximity and risk at wind energy projects in the area are considered when determining the need for operational mitigation). Given the wide-ranging movement patterns of migratory tree bats and the tendency for wind energy projects to be clustered, from a conser vation perspective, a policy which considers cumulative effects is superior to one that does not (Arnett etþ  al 2013c). Development of wind facilities in Mexico is regulated by laws and norms that have been enacted to achieve sustainable development. The General Law of Ecological Balance and Environmental Protection (GLEBEP) and its regulations are the main legal instruments the Mexican government has to protect ecosystems (www.semarnat.gob.mx). According to the GLEBEP, for a wind facility to be built, it is necessary to develop an EIA to determine the environmental feasibility of the project. The environmental authority has developed methodological guidelines for productive sectors to perform studies that meet the minimum information necessary for evaluation. In this case, the wind facilities are included in the energy sector, which includes, among others, hydroelectric, thermal, combined cycle plants, transmission lines, dams, and electrical substations. The rst wind facility EIA (2000–2004) was completed with evaluations similar to those used for any other infrastructure (e.g., hydroelectric, thermoelectric, etc.) and was therefore not focused on impacts associated with wind facilities; collisions of birds and bats are not considered in the EIA. Thus, mitigation to reduce these impacts was not required by any regulatory authority at the rst wind facilities in Mexico. Recently, the Mexican government has begun considering negative impacts on birds and bats and has incorporated measures including an annual monitoring program in these taxa during the entire cycle of wind energy projects. While there is no regulatory framework specically for protection or conservation of bats in Mexico, there is an ofcial standard that includes listings of ora and fauna found in risk categories similar to the Red List of the International Union for Conservation of Nature. The NOM-059-SEMARNAT-2010 (SEMARNAT 2010) includes the three categories of risk in order from most to least critical: in danger of extinction, threatened, and under special protection. The inclusion of species in each of these three categories is in accordance with technical and scientic criteria (SEMARNAT 2010). The NOM-059-SEMARNAT-2010 includes 38 species of bats, 19 of which warranted special protection. To date, however, only four species included in the NOM-059 have been found killed at wind facilities, and none of the species killed most frequently by turbines in Mexico are included in the NOM59 given their abundance and wide distribution (Ceballos and Arroyo-Cabrales 2012; Ceballos etþ  al. 2005). Apart from these legal instruments, there are no other legal mechanisms in Mexico to protect Mexican bats per se.

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314 E.B. Arnett et al.In the European Union, all species of bats, regardless of numbers, are strictly protected by law and it is illegal to deliberately kill or harm bats irrespective of any population effects. “Deliberately” means in this case that the actor is aware that activity may have an effect but still carries out the activity. The EIA Directive 85/337/EEC (amended to Council Directive 97/11/EC in 2011) legally requires an assessment to be carried out for all wind facilities with 5 or more turbines, or which are over 5þ  MW capacity. In addition, member states must restore or maintain their bat populations in favorable conservation status (Council Directive 92/43/EEC). All members of the EU have translated these directives into their domestic legislation which in theory should provide high levels of protection and a consistent way of handling the issue, based on the EUROBATS guidelines (Rodrigues etþ  al. 2015). However, while bat issues are taken very seriously in some countries, this is not true in others. In the UK and Republic of Ireland, it is an offense to deliberately or recklessly kill or injure a bat or to deliberately disturb bats in a way that would signicantly affect their local distribution or abundance, and detailed guidance is in place about the requirements for EIA. The EUROBATS guidance (Rodrigues etþ  al. 2015) proposes that turbines should not be placed closer than 200þ  m to woodland, whereas the current recommendations for the UK are that the blade tips should be at least 50þ  m from woodland or hedgerows. It is argued that a smaller buffer size is acceptable because the activity of bats found in the UK tends to decline rapidly with increasing distance from linear landscape features and woodlands (Natural England 2014). It is ofcially acknowledged that risk assessments for bats in the UK and Republic of Ireland are hampered by a lack of evidence in crucial areas (Bat Conservation Ireland 2012; Natural England 2014). Not only are collision rates unknown, but population estimates, and therefore inferences about the impact of turbine collisions on population viability, are uncertain. Standardized post-construction monitoring, including acoustic and carcass surveys, is recommended for sites identied as “high risk” (Bat Conservation Ireland 2012; Natural England 2014). In reality, little post-construction monitoring occurs. This is at least partly because responsibility for requiring and enforcing survey conditions lies with local planning authorities, which are reluctant to impose conditions which may be open to legal challenge. A particular difculty is that while generic guidance on survey designs is available, there is no standardized methodology and so it is extremely difcult to judge whether a particular level of bat activity would place a site as being in a “high risk” category: This point has already been raised at a legal review. Further, there has been no assessment of the relationship between pre-construction and post-construction acoustic surveys (or collision risk). It is therefore unclear how data collected pre-construction can be used to predict post-construction risk, particularly given evidence from the USA demonstrating a poor relationship between pre-construction activity surveys and bat fatalities (Hein etþ  al. 2013). A nal difculty for Local Authorities is that the legal basis of bat protection relates to the conservation status of local populations (except in Scotland, where recklessly killing a bat is also an offense). Given that local population sizes are very poorly characterized, it is unclear how mitigation (such as

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315 11þ Impacts of Wind Energy Development on Bats …raising turbine cut-in speed) could be enforced on the basis of reduced bat activity or bat fatalities. In Sweden, there were no national guidelines until 2011 (Rydell etþ  al. 2012), and more than ten years after the rst turbines were built. Hence, many wind tur bines in Sweden were constructed in poor locations with respect to bats, such as along the coast, and with operation permitted without any mitigation measures. After 2011, however, the Swedish Environmental Protection Agency concentrated resources to achieve national acceptance for the guidelines among decision makers, the industry and NGOs, and to implement them on all new project. In 2014, all proposed wind facilities are subject to a pre-construction survey and an evaluation regarding the risk to bats. In the Netherlands, all bats are species of annex II and/ or IV of the European Habitats Directive EIA, and risk assessment and evaluation under Flora and Fauna law (research on what species are present) and possibly Nature Conservation law (when EHD annex II species are present) are obligatory. When risk species are present and fatalities cannot be excluded, a permit for construction is needed under FF law and NC law, and information on fatalities needs to be established using protocols and a curtailment may be required (Boonman et al. 2013; Limpens 2013). Mitigation studies have shown that bat fatalities can be reduced substantially (e.g., Baerwald etþ  al. 2009; Arnett etþ  al. 2011). Although curtailing turbines holds great promise, the problem is that developing thresholds—those values that trigger some action—to mitigate bat kills is difcult, especially when supporting data are limited or imperfect (Arnett etþ  al. 2013c). In Germany, recent models accounting for multiple environmental variables that predict and reduce collision rate and further minimize loss of energy production offer promise for mitigating bat fatalities (Korner-Nievergelt etþ  al. 2013). However, these algorithms and most other operation protocols still tolerate an arbitrary number of bat fatalities (currently two bat fatalities per wind turbine per year where these algorithms are used in Germany) (Voigt etþ  al. 2015a); any such fatality trigger in Europe is seemingly in opposition to current law, given that European legislation does not allow deliber ate killing of any bat, regardless of population effects. Additionally, with increasing numbers of wind turbines, xed annual “per capita” (i.e., per wind turbine) mortality rates may not be acceptable in light of limited bat population sizes, and the acceptance of a reduced fatality rate may not necessarily be consistent with national and EU legislation (Voigt etþ  al 2015a). Regulatory authorities in the US state of Pennsylvania and the Canadian province of Ontario set thresholds for initiating curtailment based on the annual mean number of bats killed per turbine (28 and 10 bat fatalities/turbine, respectively Arnett etþ  al. 2013c). Arnett etþ  al. (2013c) argued that this approach sets a dangerous precedent and has several aws, none the least of which is the assumption that bat populations are currently stable and remain so. This approach also ignores expanding development of wind turbines that will likely yield increasing bat fatalities per population or region. Policy and management efforts to mitigate bat fatalities and conserve bat populations affected by wind turbines should be proactive and based on science rather than being reactive and arbitrary.

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316 E.B. Arnett et al.11.9þ Future DirectionsPopulation data are generally lacking, and this not only impedes our understanding of actual impacts of wind turbines, but also impedes knowing the effectiveness of mitigation efforts. For example, we do not know whether raising turbine cut-in speed that might result in 50þ  % fewer bat fatalities will mitigate population-level impacts or simply delay inevitable losses (Arnett and Baerwald 2013). The lack of population data also makes it difcult to set triggers for mitigation (i.e., number of bats killed per turbine or MW that requires mitigation; but see Arnett etþ  al. 2013c). However, population data are not likely to be available for most bat species in the near future, and thus, wind operators should practice the precautionary principle and implement operational mitigation at sites where bat fatalities are high, or are predicted to be high, even in the absence of population data. Several knowledge gaps remain that must be lled in the immediate future. Most notably, many regions of the world lack any publicly available monitoring information on the impacts of wind energy facilities on bats and other wildlife (e.g., China). We strongly encourage wind energy developers and governments to end this trend and gather needed data to inform siting and operating wind facilities around the world. In addition to population studies and basic monitoring data described above, some key research priorities germane to all regions of the world include: 1.þt Evaluating effectiveness of pre-construction bat activity surveys in predicting future fatalities at wind facilities. 2.þt Determining whether approaches such as temperature proles in relation to weather types and seasons (and different regions) or habitat suitability modeling for bats can effectively predict high-risk sites and be used by planning authorities and industry to help situate wind turbines in areas where the potential for conict with bats is minimized. 3.þt Evaluate methods for assessing the risk and minimizing and mitigating impacts posed by offshore wind turbines to bats (including approaches such as radar and collision sensors). 4.þt Investigate the extent of migratory activity worldwide, particularly offshore, by international collaboration using a range of techniques [stable isotopes (Voigt etþ  al. 2012; Lehnert etþ  al. 2014; Baerwald etþ  al. 2014), population genetics, potentially GPS tracking, etc.]. Although difcult, it may be possible to gather and pool acoustic data of activity patterns established through automated real-time recorders in numerous wind energy facilities to evaluate spatial and temporal patterns. 5.þt Future operational mitigation experiments should be designed to determine which factors (e.g., habitat, insect occurrence, temperature, wind, humidity, moon illumination) or combination of factors (Weller and Baldwin 2012) will best improve predictability of bat fatalities, while minimizing economic costs. Alternative mitigation approaches to operational adjustments should be proven equally or more effective at reducing bat fatalities at operating wind facilities before being accepted as viable mitigation approaches.

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317 11þ Impacts of Wind Energy Development on Bats …6.þt Detailed meta-analyses of existing data on cumulative fatality impacts and factors inuencing fatalities are needed. It should be noted that data disclosure from many sites by some companies hinders such analyses. We strongly encourage developers to follow guidelines (e.g., Kunz etþ  al. 2007b; Rodrigues etþ  al. 2015; Strickland etþ  al. 2011; Bat Conservation Ireland 2012) consistently when implementing preand post-construction monitoring. Data should be placed into the public domain or preferably published in refereed journals. There are a number of policies, regulatory, and communication challenges we face in protecting bats while developing wind energy responsibly across the globe (Arnett 2012). Unless there is a federal, state, or provincial nexus, most research, siting, and mitigation efforts by wind energy developers and operating companies will be voluntary, likely without regard for cumulative effects. Sites that do trigger a regulatory nexus will usually be driven by endangered species issues (e.g., Indiana bats in the USA). It is apparent most local jurisdictional entities, regardless of country, lack experience in wildlife science, and unless they coordinate with their wildlife or natural resource agency specialists, concerns about bat fatalities may never be addressed in decision making for wind energy development. Another key issue is consistent application of regulations. The authors have encountered many situations where different individuals had varied interpretations of the same law or guidance policy, and this creates untenable situations, consider able uncertainty, consternation, and lack of trust among stakeholders that seems completely unnecessary and easily remedied (Arnett 2012). We encourage continuing cooperation among all stakeholders, gathering needed information, avoiding construction in high-risk sites, considering cumulative effects, and implementing mitigation where needed even when no regulatory process is triggered (Arnett 2012; Arnett and Baerwald 2013). Wind energy developers should, however, be treated fairly and consistently to ensure proactive measures are implemented. Arnett (2012) noted that when some companies choose to cooperate, while others may not, unnecessary angst is generated and deters resolving wildlife impacts and other issues. Decision making must be based on the best available science. Also, consistent policy, accountability, effective siting and mitigation strategies, and a “level-playing eld” for the industry (i.e., consistent requirements and incentives for all companies) are fundamental if we are to successfully develop wind energy that protects bats and other wildlife.Acknowledgementsþ We appreciate the invitation from C. Voigt and T. Kingston to prepare this chapter. Portuguese data were collected by AgriPro Ambiente, Bio3, Biota, Colmus, Ecomind, Ecosativa, Ecosfera, EDP, ENEOP2, EolFlor, LEA, Naturibérica, NOCTULA, Plecotus, PROCESL, Proco Ambiente, ProSistemas, STRIX, and Tecneira. We acknowledge C.-H. Chou at the Endemic Species Research Institute in Taiwan for letting us cite his important but yet unpublished data. This chapter was greatly enhanced by reviews from C. Hein, T. Kingston, H. Limpens, and 2 anonymous reviewers. We wish to dedicate this chapter to our friend Dr. Thomas Kunz for all of his wisdom, knowledge, and efforts in regard to wind energy impacts on bats.

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318 E.B. Arnett et al.ReferencesAdams JK (1989) Pteronotus davyii. Mamm Species 346:1–5 Ahlén I, Baagøe HJ, Bach L (2009) Behavior of Scandinavian bats and foraging at sea. J Mammal 90:1318–1323 Amorim F, Rebelo H, Rodrigues L (2012) Factors and mortality at a wind farm in the Mediterranean region. Acta Chirop 14:439–457 Ancilotto L, Rydell J, Nardone V, Russo D (2014) Coastal cliffs on islands as foraging habitat for bats. Acta Chirop 16:103–108 Arnett EB (2005) Relationships between bats in Pennsylvania and West Virginia: an assessment of bat fatality search protocols, patterns of fatality, and behavioral interactions with wind turbines. Bat conservation international, Austin, Texas, USA http://www.batsandwind.org/p df/postconpatbatfatal.pdf. Accessed 22 Dec 2014 Arnett EB (2012) Impacts of wind energy development on wildlife: challenges and opportunities for integrating science, management. In: Sands JP, DeMaso SJ, Brennan LA, Schnupp MJ (eds) Wildlife science: with management. Taylor and Francis, New York, pp 213–237 Arnett EB, Baerwald EF (2013) Impacts of wind energy development on bats: implications for conservation. In: Adams RA, Peterson SC (eds) Bat evolution, ecology, and conservation. Springer, New York, pp 435–456 Arnett EB, Inkley DB, Larkin RP, Manes S, Manville AM, Mason JR, Morrison ML, Strickland MD, Thresher R (2007) Impacts of wind energy facilities on wildlife and wildlife habitat. Wildl Soc Tech Rev 07-2 (The Wildlife Society, Bethesda, Maryland) Arnett EB, Brown K, Erickson WP, Fiedler J, Henry TH, Johnson GD, Kerns J, Kolford RR, Nicholson CP, O’Connell T, Piorkowski M, Tankersley R Jr (2008) Patterns of fatality of bats at wind energy facilities in North America. J Wildl Manage 72:61–78 Arnett EB, Huso MMP, Schirmacher MR, Hayes JP (2011) Changing wind turbine reduces at wind facilities. Front Ecol Environ 9:209–214. doi:10.1890/100103 Arnett EB, Johnson GD, Erickson WP, Hein CD (2013a) A synthesis of studies to reduce at wind energy facilities in North America. Bat conservation international, Austin, Texas http://www.batsandwind.org/pdf/Operational%20Mitigation%20Synthesis%20FINAL%20 REPORT%20UPDATED.pdf. Accessed 22 Dec 2014 Arnett EB, Hein CD, Schirmacher MR, Huso MMP, Szewczak JM (2013b) Evaluating the effectiveness of an ultrasonic acoustic deterrent for reducing. PLoS ONE 8(6):e65794. doi:10.1371/journal.pone.0065794 Arnett EB, Barclay RMR, Hein CD (2013c) Thresholds for bats killed. Front Ecol Environ 11:171. doi:10.1890/1540-9295-11.4.171 AronsonJB, Thomas AJ, Jordaan SL (2013) Bat fatality at a wind energy facility in the western Cape. Afr Bat Conserv News 31:9–12 (ISSN 1812-1268) Baerwald EF, Barclay RMR (2009) Geographic variation in activity and fatality bats at wind energy facilities. J Mammal 90:1341–1349 Baerwald EF, Barclay RMR (2011) Patterns of activity and fatality bats at a wind energy facility in Alberta. J Wildl Manage 75:1103–1114 Baerwald EF, D’Amours GH, Klug BJ, Barclay RM (2008) Barotrauma is a signicant cause of bat fatalities at wind turbines. Curr Biol 18(16):R695–R696 Baerwald EF, Edworthy J, Holder M, Barclay RMR (2009) A large-scale mitigtion experiment to reduce at wind energy facilities. J Wildl Manage 73:1077–1081 Baerwald EF, Patterson WP, Barclay RMR (2014) Origins patterns of bats killed in southern Alberta: evidence from stable isotopes. Ecosphere 5(article 118):1–17 Open Access This chapter is distributed under the terms of the Creative Commons Attribution Noncommercial License, which permits any noncommercial use, distribution, and reproduction in any medium, provided the original author(s) and source are credited.

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319 11þ Impacts of Wind Energy Development on Bats … Barclay RMR, Harder LM (2003) Life histories of bats: life in the slow lane. In: Kunz TH, Fenton MB (eds) Bat ecology. University of Chicago Press, Chicago, pp 209–259 Bat Conservation Ireland (2012) Wind turbine/wind development bat survey guidelines— version 2.8. Bat conservation Ireland http://www.batconservationireland.org/pubs/ reports/BCIreland%20Wind%20Farm%20Turbine%20Survey%20Guidelines%20 Version%202%208.pdf. Accessed 20 January 2015 Bateman GC, Vaughan TA (1974) Nightly activities of mormoopid bats. J Mammal 55:45–65 Behr O, von helversen O (2006) Gutachten zur Beein -trächtigung im freien Luftraum jagender und ziehender Fle -dermäuse durch bestehende Windkraftanlagen. Wirkungs -kontrolle zum Windpark “Rosskopf” (Freiburg i Br.) im Jahr 2005. Report to Regiowind GmbH & Co., Freiburg Bennett E (2012) Hepburn wind farm bird and bat mortality survey interim report 11th July 2011–9th Jan 2012. Elmoby ecology http://hepburnwind.com.au/downloads/Hepburn%20 Wind%20interim%20report%20Bird%20&%20Bat%20Monitoring.pdf. Accessed 25 Aug 2014 Berkhout V, Faulstich S, Görg P, Kühn P, Linke K, etþ  al. (2013) Wind energy 2012. FraunhoferInstitut für Windenergie und Energiesystemtechnik. IWES, Kassel Bernstein MA, Grifn J, Lempert R (2006) Impacts on energy expenditures of use. Technical report prepared for the energy future coalition, RAND Corporation, Santa Monica, CL Boonman M, Limpens HJGA, La Haye MJJ, van der Valk M, Hartman JC (2013) Protocollen leermuisonderzoek bij windturbines. Rapport 2013.28, Zoogdiervereniging & Bureau Waardenburg. 29þ  pp 1 bijlage Boshamer JPC, Bekker JP (2008) Nathusius’ pipistrelles (Pipistrellus nathusii) and other species of bats platforms in the Dutch sector of the North Sea. Lutra 51:17–36 Brinkmann R, Behr O, Niermann I, Reich M (2011) Entwicklung von Methoden zur Untersuchung und Reduktion des Kollisionsrisikos von Fledermausen an onshore-Windenergieanlagen. Schriftenreihe Institut fur Umweltplanung. Cuvillier Verlag Gottingen, p 457 Briones-Salas M, Peralta-Pérez My, García-Luis M (2013) Acoustic characterization of new species of bats for the state of Oaxaca. Theyra 4:15–32 Camina A (2012) At wind farms in—lessons to be learned. Acta Chirop 14:205–212 Canadian Environmental Assessment Act (2012) An ACT respecting the environmental assessment of certain activities and the prevention of signicant adverse environmental effects. Assented to 29 June 2012, in force 06 July 2012. (Online). Government : Justice Laws http://laws-lois.justice.gc.ca/eng/acts/C-15.21/page-1.html. Accessed 6 Jan 2015 Ceballos G, Arroyo-Cabrales J (2012) Lista actualizada de los mamíferos de México 2012. Rev Mex Mastozoología 1:27–80 Ceballos G, Oliva G, coords. (2005) Los mamíferos silvestres de México. Comisión Nacional para el Conocimiento y Uso de la Biodiversidad y Fondo de Cultura Económica, México, DF Cryan PM (2008) Mating behavior as a possible cause of bat fatalities at wind turbines. J Wildl Manage 72:845–849 Cryan PM (2011) Wind turbines as landscape impediments to connectivity of bats. Environ Law 41:355–370 Cryan PM, Barclay RMR (2009) Causes of bat fatalities at wind turbines: hypotheses and predictions. J Mammal 90:1330–1340 Cryan PM, Brown AC (2007) Migration of bats past a remote island offers clues towards the problem of bat fatalities at wind turbines. Biol Conserv 139(1–2):1–11 Doty AC, Martin AP (2012) Assessment of bat and avian mortality at a pilot wind turbine at Coega, Port Elizabeth, Eastern Cape, South Africa. New Zeal J Zool. doi:10.1080/0301422 3.2012.741068 Dubourg-Savage M-J, Rodrigues L, Santos H, Georgiakakis P, Papadotou E, Bach L, Rydell J (2011) Patterns of comparing north and south. Poster and abstract. In: Conference on wind energy and environmental impacts, Trondheim, Norway, May 2011

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320 E.B. Arnett et al. EUROBATS (2014) Report of the intersessional working group on wind turbines and bat populations. http://www.eurobats.org/sites/default/les/documents/pdf/Advisory_Committee/Doc. AC_.20.5.ReportIWGWindTurbines_0.pdf. Accessed 1 Aug 2015 Frick WF, Pollock JF, Hicks AC, Langwig KE, Reynolds DS, Turner GG, Butchkoski CM, Kunz TH (2010) An emerging disease causes regional population collapse of a common North American bat species. Science 329:679–682 Gannon MR, Kurta A, Rodríguez-Durán A, Willig MR (2005) Bats of Puerto Rico: an island focus and a Caribbean perspective. Texas Tech University Press, Lubbock García-Grajales J, Silva AB (2012) Revisión al conocimiento de los murciélagos del estado de Oaxaca. Therya 3:277–293 Georgiakakis P, Kret E, Cárcamo B, Doutau B, Kafkaletou-Diez A, Vasilakis D, Papadatou E (2012) Bat fatalities at wind farms in north-eastern Greece. Acta Chirop 14:459–468 Cryan PM, Gorresen, PM, Hein CD, Schirmacher MR, Diehl RH, Huso MM, Hayman DTS, Fricker PD, Bonaccorso FJ, Johnson DH, Heist K, Dalton DC (2014) Behavior of bats at wind turbines. Proc Nat Acad Sci. doi:10.1073/pnas.1406672111 Government Accountability Ofce (GAO) (2005) Wind Power: impacts on wildlife and government responsibilities for regulating development and protecting wildlife. Report to Congressional Requesters, GAO-05-906, U.S. Government Accountability Ofce, Washington Government of Alberta (2013) Bat mitigation framework for wind power development. In: Environment and sustainable resource development. Alberta, Canada Grodsky SM (2010) Aspects of bird and bat mortality at a wind energy facility in southeastern Wisconsin: impacts, relationships, and cause of death. Thesis, University of Wisconsin, Madison Grodsky SM, Behr MJ, Gendler A, Drake D, Dieterle BD, Rudd RJ, Walrath NL (2011) Investigating the causes of death for wind turbine-associated bat fatalities. J Mammal 92:917–925 Hall LS, Richards GC (1972) Notes on Tadarida australis (Chiroptera:Molossidae). Aust Mammal 1:46–47 Hayes MA (2013) Bats killed in large numbers at United States wind energy facilities. Bioscience 63:975–979 Hein CD, Gruver J, Arnett EB (2013) Relating pre-construction bat activity and post-construction bat fatality to predict risk at wind energy facilities: a synthesis. Bat conservation inter national, Austin, Texas http://www.batsandwind.org/pdf/Pre-%20Post-construction%20 Synthesis_FINAL%20REPORT.pdf. Accessed 23 Dec 2014 Heinrich B (1993) The hot-blooded insects. Springer, Berlin Horn J, Arnett EB, Kunz TH (2008) Behavioral responses of bats to operating wind turbines. J Wildl Manage 72:123–132 Hull CL, Cawthen L (2012) Bat fatalities at two wind farms in Tasmania, Australia: bat characteristics, and spatial and temporal patterns. New Zeal J Zool. doi:10.1080/03014223.2012.7 31006 Huso MMP (2011) An estimator of wildlife fatality from observed carcasses. Environmetrics 22:318–329. doi:10.1002/env.1052 Huso MMP, Dalthorp D (2013) Accounting for unsearched areas in estimating wind turbinecaused fatality. J Wildl Manage. doi:10.1002/jwmg.663 Hutson AM, Mickleburgh SP, Racey PA (2001) Global status and conservation action plan: microchiropteran bats. IUCN/SSC Chiroptera Specialist Group, Cambridge Hutterer R, Ivanova T, Meyer-Cords C, Rodrigues L (2005) Bat migrations in Europe: a review of banding data and literature. Naturschutz und Biologische Vielfalt, Bonn 28:1–176 Ingersoll TE, Sewall BJ, Amelon SK (2013) Improved analysis of long-term monitoring data demonstrates marked regional declines of bat populations in the Eastern United States. PLoS ONE 8(6):e65907. doi:10.1371/journal.pone.0065907

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321 11þ Impacts of Wind Energy Development on Bats … Inkley DB, Anderson MG, Blaustein AR, Burkett VR, Felzer B, Grifth B, Price J, Root TL (2004) Global climate change and wildlife in North America. Wildl Soc Tech Rev 04-1 (The Wildlife Society, Bethesda, Maryland) Jain AA, Koford RR, Hancock AW, Zenner GG (2011) Bat mortality and activity at a northern Iowa wind resource area. Am Mid Nat 165:185–200 Johnson GD (2005) A review of bat mortality at wind-energy developments in the United States. Bat Res News 46:45–49 Johnson GD, Perlik MK, Erickson WP, Strickland MD (2004) Bat activity, composition, and collision mortality at a large wind plant in Minnesota. Wildl Soc Bull 32:1278–1288 Johnson JB, Gates JE, Zegre NP (2011) Monitoring seasonal bat activi ty on a coastal barrier island in Maryland, USA. Environ Monit Assess 173:685–699 Korner-Nievergelt F, Brinkmann R, Niermann I, Behr O (2013) Estimating bat and bird mortality occurring at wind energy turbines from covariates and carcass searches using mixture models. PLoS ONE 8(7):e67997. doi:10.1371/journal.pone.0067997 Kunz TH, Arnett EB, Erickson WP, Hoar AR, Johnson GD, Larkin RP, Strickland MD, Thresher RW, Tuttle MD (2007a) Ecological impacts of wind energy development on bats: questions, research needs, and hypotheses. Front Ecol Environ 5:315–324 Kunz TH, Arnett EB, Cooper BM, Erickson WP, Larkin RP, Mabee T, Morrison ML, Strickland MD, Szewczak JM (2007b) Methods and metrics for studying impacts of wind energy development on nocturnal birds and bats. J Wildl Manage 71:2449–2486 Ladle RJ, Firmino JVL, Malhado ACM, Rodríguez-Durán A (2012) Unexplored diversity and conservation potential of Neotropical hot caves. Conserv Biol 26:978–982 LEA (2010) Monitorização dos efeitos da Medida de Minimização de Mortalidade do Parque Eólico do Outeiro Relatório nal. Laboratório de Ecologia Aplicada da Universidade de Trás-os-Montes e Alto Douro. Vila Real, 78þ  pp Lehnert LS, Kramer-Schadt S, Schonborn S, Lindecke O, Niermann I, Voigt CC (2014) Wind farm facilities in Germany kill noctule bats from near and far. PLoS ONE 9(8):e103106. doi:10.1371/journal.pone.0103106 Limpens HJGA, Boonman M, Korner-Nievergelt F, Jansen EA, van der Valk M, La Haye MJJ, Dirksen S, Vreugdenhil SJ (2013) Wind turbines and bats in the Netherlands—measuring and predicting.þ  Report 2013.12, Zoogdiervereniging & Bureau Waardenburg. 77þ  pp 2 appendices McLeish T (2002) Wind power. Nat N Eng 11:60–65 Miller A (2008) Patterns of avian and bat mortality at a utility-scalewind farm on the southern high plains. Thesis, Texas Tech University, Lubbock National Research Council [NRC] (2007) Ecological impacts of wind-energy projects.þ  National Academies Press, Washington Natural England (2014) Natural England technical information note TIN051. Bats and onshore wind turbines, 3rd edn, March 2014 http://publications.naturalengland.org.uk/search?qBat sandonshorewindturbines&num100. Accessed 4 Jan 2015 Nicholls B, Racey PA (2009) The aversive effect of electromagnetic radiation on foraging bats— a possible means of discouraging bats from approaching wind turbines. PLoS ONE 7:e6246. doi:10.1371/journal.pone.0006246 O’Shea TJ, Bogan MA, Ellison LE (2003) Monitoring trends in bat populations of the United States and Territories: status of the science and recommendations for the future. Wild Soc Bull 31:16–29 Ontario Ministry of Natural Resources (2011) Bats and bat habitats: guidelines for wind power projects. Ontario, Canada Piorkowski MD, O’Connell TJ (2010) Spatial pattern of summer bat mortality from collisions with wind turbines in mixed-grass prairie. Am Mid Nat 164:260–269 Poerink BJ, Lagerveld S, Verdaat H (2013) Bat activity in the Dutch offshore wind farm OWEZ and PAWP. IMARES Report no. C026/13 / tFC report no. 20120402. The Fieldwork Company, Wageningen, Holland http://edepot.wur.nl/248979. Accessed 29 Nov 2014

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322 E.B. Arnett et al. Rodrigues L, Bach L, Dubourg-Savage MJ, Karapandza B, Kovac D, Kervyn T, Dekker J, Kepel A, Bach P, Collins J, Harbusch C, Park K, Micevski B, Minderman J (2015) Guidelines for consideration of bats in wind farm projects – revision 2014. EUROBATS Publication Series no. 6 (English version). UNEP/EUROBATS Secretariat, Bonn, Germany, UNEP/EUROBATS Secretariat. http://www.eurobats.org/sites/default/les/documents/publications/publication_series/pubseries_no6_english.pdf. Accessed 1 Aug 2015 Rodríguez-Durán A (2009) Bat assemblages in the West Indies: the role of caves. In: Fleming TH, Racey P (eds) Island bats: evolution, ecology, and conservation. University of Chicago Press, Chicago, pp 265–280 Rollins KE, Meyerholz DK, Johnson GD, Capparella AP, Loew SS (2012) A forensic investigation into the etiology of bat mortality at a wind farm: barotrauma or traumatic injury? Vet Pathol 49:362–371 Roscioni F, Russo D, Di Febbraro M, Frate L, Carranza ML, Loy A (2013) Regional-scale modeling of the cumulative impact of wind farms on bats. Biodivers Conserv. doi:10.1007/ s10531-013-0515-3 Roscioni F, Rebelo H, Russo D, Carranza ML Di Febbraro M, Loy A (2014) A modeling approach to infer the effects of wind farms on landscape connectivity for bats. Landscape Ecol. doi 10.1007/s10980-014-0030-2 Rydell J, Bach L, Dubourg-Savage M, Green M, Rodrigues L, Hedenstrom A (2010a) Bat mor tality at wind turbines in northwestern Europe. Acta Chirop 12:261–274 Rydell J, Bach L, Dubourg-Savage MJ, Green M, Rodrigues L, Hedenström A (2010b) Mortality of bats at wind turbines links to nocturnal insect migration? Euro J Wildl Res 56:823–827 Rydell J, Engström H, Hedenström A, Larsen JK, Pettersson J, Green M (2012) The effect of wind power on birds and bats: a synthesis report. Report 6511, Swedish Environmental Protection Agency, Stockholm, Sweden Rydell J, Bach L, Bach P, Guia Diaz L, Furmankiewicz J, Hagner-Wahlsten N, Kyheröinen EM, Lilley T, Masing M, Meyer MM, Petersons G, Suba J, Vasko V, Vintulis V, Hedenström A (2014) Phenology of migratory bat activity across the Baltic Sea and the south-eastern North Sea. Acta Chirop 16:139–147 Santos H, Rodrigues L, Jones G, Rebelo H (2013) Using species distribution modelling to predict bat fatality risk at wind farms. Biol Conserv 157:178–186 Schlesinger ME, Mitchell JFB (1987) Climate model simulations of the equilibrium climate response to increased carbon dioxide. Rev Geophys 25:760–798 Smallwood KS (2013) Comparing bird and bat fatality-rate estimates among North American wind-energy projects. Wildl Soc Bull 37:19–33 Steffens R, Zophel U, Brockmann D (2004) 40th anniversary bat marking Centre Dresden— evaluation of methods and overview of results. Sachsisches Landesamt für Umwelt und Geologie, p 132 Strickland MD, Arnett EB, Erickson WP, Johnson DH, Johnson GD, Morrison ML, Shaffer JA, Warren-Hicks W (2011) Comprehensive guide to studying wind energy/wildlife interactions. National Wind Coordinating Collaborative, Washington http://www.batcon.org/pdfs/ wind/National%20Wind%20Coordinating%20Collaborative%202011_Comprehensive%20 Guide%20to%20Studying%20Wind%20Energy%20and%20Wildlife%20Interactions.pdf. Accessed 4 Jan 2015 U.S. Fish and Wildlife Service (2012) Land-based wind energy guidelines. U.S. sh and wildlife service, Arlington, VA, USA http://www.fws.gov/windenergy/docs/WEG_nal.pdf. Accessed 1 February 2015 Van der Meij T, Van Strien AJ, Haysom KA, Dekker J, Russ J, Biala K, Bihar Z, etþ  al. (2014) Return of the bats? A prototype indicator of trends in European bat populations in under ground hibernacula. Mammal Biol. doi:10.1016/j.mambio.2014.09.004 Villegas-Patraca R, Macías-Sánchez S, MacGregor-Fors I, Muñoz-Robles C (2012) Scavenger removal: bird and bat carcass persistence in a tropical wind farm. Acta Oecol 43:121–125 Voigt CC, Popa-Lisseanu AG, Niermann I, Kramer-Schadt S (2012) The catchment area of wind farms for European bats: a plea for international regulations. Biol Conserv 153:80–86

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323 11þ Impacts of Wind Energy Development on Bats … Voigt CC, Lehnert LS, Petersons G, Adorf F, Bach L (2015a) Wildlife and renewable energy: German politics cross migratory bats. Eur J Wildl Res (in press) Voigt CC, Aguirre L, Phelps K, Schoeman C, Vanitharani J, Zubaid A (2016) Bats and buildings: the conservation of synanthropic bats. In: Voight CC, Kingston T (eds) Bats in the anthropocene: conservation of bats in a changing world. Springer International AG, Cham, pp. 427–453 Weller TJ, Baldwin JA (2012) Using echolocation monitoring to model bat occupancy and inform mitigations at wind energy facilities. J Wildl Manage 76:619–631 Wellington WG (2011) The effects of variations in atmospheric pressure upon insects. Can J Res 1946, 24d(2):51–70 Young DP, Bay K, Nomani S, Tidhar WL (2011) Nedpower mount storm wind energy facility post-construction avian and bat monitoring: July–October 2010. Western EcoSystems Technology, Inc, Cheyenne, Wyoming http://www.batsandwind.org/pdf/WV%20-%20 Young%20et%20al.%202011%20-%20Mount%20Storm%20Fall%202010%20Report(%20 2-10-11).pdf. Accessed 4 Jan 2015

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325Chapter 12Exploitation of Bats for Bushmeat and MedicineTammy Mildenstein, Iroro Tanshi and Paul A. Racey© The Author(s) 2016 C.C. Voigt and T. Kingston (eds.), Bats in the Anthropocene: Conservation of Bats in a Changing World, DOIþ  10.1007/978-3-319-25220-9_12Abstractþ Bat hunting for consumption as bushmeat and medicine is widespread and affects at least 167 species of bats (or c. 13þ  % of the world’s bat species), in Africa, Asia, across the islands of Oceania, and to a lesser extent in Central and South America. Hunting is particularly prevalent among the large-bodied fruit bats of the Old World tropics, where half (50þ  %, 92/183) the extant species in the family Pteropodidae are hunted. Pteropodids that are hunted are six times more likely to be Red Listed as threatened: 66þ  % of species in IUCN threatened categories (CR, EN, VU, NT), compared to 11þ  % of species in the ‘Least Concern’ (LC) category. However, there still appears to be an information gap at the international level. One third of the hunted species on the Red List are not considered threatened by that hunting, and nearly a quarter of the bat species included in this review are not listed as hunted in IUCN Red List species accounts. This review has resulted in a comprehensive list of hunted bats that doubles the number of species known from either the IUCN Red List species accounts or a questionnaire circulated in 2004. More research is needed on the impacts of unregulated hunting, as well as on the sustainability of regulated hunting programs. In the absence of population size and growth data, legislators and managers should be precautionary in their attitude towards T. Mildensteinþ  ()þ  Department of Biology, Cornell College, Mt. Vernon, IA, USA e-mail: TMildenstein@cornellcollege.edu I. Tanshiþ  Department of Biological Sciences, Texas Tech University, Lubbock, TX, USA I. Tanshiþ  Department of Animal and Environmental Biology, University of Benin, Benin City, Nigeria P.A. Raceyþ  Centre for Ecology and Conservation, College of Life and Environmental Sciences, University of Exeter, Penryn Campus, Penryn TR10 9FE, UK

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326 T. Mildenstein et al.hunting. Roost site protection should be a priority as it is both logistically simpler than patrolling bat foraging grounds and reduces the comparatively larger scale mortality and stress that hunting at the roost can cause. Education and awareness campaigns within local communities should demonstrate how bats are a limited resource and emphasize characteristics (nocturnal, slow reproducing and colonial) that make them particularly vulnerable to hunting pressure.12.1þ IntroductionMost of the chapters in this book (Voigt and Kingston 2016) consider negative consequences for bats from indirect effects of anthropogenic perturbations. In contrast, this chapter explores the direct exploitation of bats by humans for bushmeat and medicine. A global review of bats as bushmeat was published in 2009 providing an over view of bat hunting based on published literature and a questionnaire widely distributed among bat biologists in 2004 (Mickleburgh etþ  al. 2009). Here, we summarize what is currently known about the exploitation of bats for consumption and medicinal use, synthesizing the 2009 review with what has been published since and unpublished information the authors have gathered from colleagues. The result is a comprehensive list of hunted bats species that contains nearly twice as many species as known from either the IUCN Red List species accounts or the 2009 review. It is unclear whether the increased concern about hunting is the result of greater actual hunting pressure, or just represents our increased understanding of hunting impacts. What is clear is that reviews now explicitly attribute species declines and extinction risk to hunting pressure. Most of the chapter discusses conservation needs in light of what is known about bat hunting. We summarize regional patterns in bat hunting and protection efforts and highlight areas of conservation concern. We provide details of current research aimed at learning more about hunting impacts, and we give examples of management activities to strengthen protection efforts against population-level effects of hunting. We end with recommendations for research to better under stand the effects of hunting on bat populations as well as strategies for effectively managing hunting to support bat conservation. All common and scientic species names follow Simmons (2005).12.2þ Background 12.2.1þ Overview of Bat HuntingBat hunting is widespread and affects at least 167 species of bats (or c. 13þ  % of the world’s 1331 bat species, Bat Conservation International 2015), occurring in Africa, Asia, across the islands of Oceania, and in some parts of Central and

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327 South America (compiled from IUCN 2014; Mickleburgh etþ  al. 2009, and personal communications by the authors; Appendix). Hunting is particularly prevalent on the large-bodied fruit bats (family Pteropodidae) in the Old World tropics, where half (50þ  %, 92/183) of all extant species experience hunting pressure (Mickleburgh etþ  al. 2009; IUCN 2014). A much smaller proportion of insectivorous (<8þ  %, 75/962 species) are hunted, particularly members of the Emballonuridae, Hipposideridae and Molossidae in Asia and Southeast Asia, Vespertilionidae in North Africa and West and Central Asia, and Phyllostomidae in Brazil, Peru, and Bolivia as well as Rhinolophidae in sub-Saharan Africa, (Mickleburgh etþ  al. 2009; Lizarro etþ  al. 2010; IUCN 2014) (Appendix). Bats are hunted for a variety of reasons, from their perceived medicinal proper ties e.g. Nicobar ying fox, Pteropus faunulus, Kingston etþ  al. (2008); ‘small bats’ in Nepal (Tuladhar-Douglas 2008); fat from pteropodid species in Pakistan (Roberts 1997) to their use in ornate decoration such as the teeth of the Makira ying fox ( Pteropus cognatus) used for necklaces (James etþ  al. 2008). Bats are also hunted for sport by urban residents seeking country pursuits (e.g. large fruit bat hunting at Subic Bay, Philippines, S. Stier, pers. comm.) and tourists seeking exotic eating experiences (e.g. Pacic ying fox, Pteropus tonganus, hunting is offered as a recreation option at hotels in Vanuatu; A. Brooke pers. comm. in Hamilton and Helgen 2008). However, the most widespread reason for bat hunting, by far, is for consumption; all 167 species that are hunted are, at least in part, wanted for their meat as a source of protein. Bat meat ranges in value from a highly sought-after delicacy served at special ceremonies and traditional celebrations (e.g. Pteropus mariannus in the Mariana Islands) to “nger food” consumed in social drinking settings (e.g. many bat species in Southeast Asia, Mildenstein 2012; and in West Africa, M. Abedi-Lartey pers. comm.). Elsewhere, it provides an alternative source of protein for local people for whom meat is an expensive commodity (Jenkins and Racey 2008) and in extreme cases, bats are consumed as starvation food (Goodman 2006). The intensity and frequency of bat harvesting varies from year round to periodical depending on the seasonality of the species, hunters’ lifestyles, and/or local legislation. On the Islands of São Tomé and Príncipe, bats are hunted opportunistically for food all year round (Carvalho etþ  al. 2014). In Southeast Asia, regular har vest of bats occurs in Indonesia and the Philippines (T. Mildenstein, unpublished data). A migratory species, the African straw-colored fruit bat (Eidolon helvum) is hunted whenever it is present in Accra (Fig.þ  12.1) and Kumasi, Ghana, between November and March (Kamins etþ  al. 2011). Reduced hunting intensity in March is likely due to the northward migration of forest resident bats and/or the shift in hunters’ occupation to farming. Similarly, in Madagascar, local legislation species two hunting seasons—one for fruit bats, and another for Commerson’s leafnosed bat (Hipposideros commersoni) (Jenkins and Racey 2008), though actual hunting intensity is driven more by local availability (e.g. the lychee season for fruit bats). In some localities in east and western Nigeria, year round harvest of the Egyptian rousette (Rousettus aegyptiacus) is known (Fig.þ  12.1) (I. Tanshi pers. obs.), and E. helvum was documented as hunted during peak population periods in the southwest (Funmilayo 1978; Halstead 1977).

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328 T. Mildenstein et al.Bushmeat is preferred to domestic livestock in many places because of the taste and perceived higher nutritional alue (Mbete etþ  al. 2011, T. Mildenstein unpublished data). In locations where domestic meats and sh are generally preferred, such as Madagascar (Randrianandrianina etþ  al. 2010), bushmeat becomes more important in periods of food shortage (Jenkins and Racey 2008). Similarly, on the island of Yap (Micronesia), hunting is socio-economically based, and bats are less desirable than seafood. Only people of lower social ranks with no access to the coast hunt fruit bats (Falanruw 1988). Consumption of bushmeat varies indirectly with the availability of other protein sources (e.g. in west Africa: Brashares etþ  al. (2004)). In areas where bats are eaten, they are rarely the only available source of protein. The exception to this is in times of food insecurity, when people turn to bats as a food source, especially following natural disasters (e.g. typhoons: Aldabra ying fox, Pteropus aldabrensis, Mickleburgh etþ  al. 2008a; Vanuatu ying fox, Pteropus anetianus, Helgen and Hamilton 2008a; Ontong Java ying fox, Pteropus howensis, Helgen and Allison 2008; Rodrigues ying fox, Pteropus rodricensis, Mickleburgh etþ  al. 2008b; Samoan Flying Fox, Pteropus samoensis and P. tonganus, Brooke 2001, and P. mariannus, Esselstyn etþ  al. 2006, Fig.þ  12.1þ Collection and sales of bats in Africa a R. aegyptiacus collected by a hunter with sticks from a limestone cave in Etapkini near Calabar, Nigeria (credit I. Tanshi), b Fruit bat kebab on sale in Kumasi, Ghana (credit M. Abedi-Lartey), c E. helvum and H. monstrosus on sale in a small market by the River Congo in Kisangani, DRC (credit Guy-C. Gembu)

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329 USFWS 2009) and during civil unrest (e.g., Bougainville monkey-faced fruit bat, Pteralopex anceps antrata, S. Hamilton, pers. comm.). Similarly, species found in low-lying areas (e.g. P. aldabrensis and P. howensis) may become increasingly important food to local communities as rising sea-levels destroy other food sources (Mickleburgh etþ  al. 2008a; Helgen and Allison 2008). Twenty years ago marked the end of a long period of international trade in the Pacic with many pteropodids being imported into Guam and the Northern Mariana Islands. Once local bat populations were depleted, bats were imported from other island groups and mainland Southeast Asia (e.g. Wiles and Payne 1986; Wiles 1992; Stinson etþ  al. 1992). Protracted international effort eventually led in 1987–1989 to the addition of pteropodid species to the Appendices of the Convention on International Trade of Endangered Species (CITES), which has stopped legal trade of bats between nation states, although a black market still occurs (e.g. into Europe, Samuel 2013). Currently, hunting of bats for trade tends to be locally-based, and not inter national, but varies widely in intensity. An extensive commercial chain of bat trade exists outside markets in Ghana (Kamins etþ  al. 2011). Other high levels of trade, include that of the large ying fox (Pteropus vampyrus) in Kalimantan, Indonesia (Harrison etþ  al. 2011) and of the Malagasy ying fox (Pteropus rufus) in Madagascar (Jenkins etþ  al. 2007; Oleksy etþ  al. 2015b). More commonly, bats are traded locally and on a lesser scale, with relatively few individuals sold in markets (e.g. P. vampyrus in the Philippines, Sheffers etþ  al. 2012; and in Southeast Asia, Mickleburgh etþ  al. 2009). Prices per bat range from <1 USD in Southeast Asia (Indonesia: Heinrichs 2004; the Philippines: T. Mildenstein unpublished data) to more than 130 USD when acquired through black market trading (e.g. P. mariannus on Guam and the Northern Mariana Islands, USFWS 2009).12.2.2þ Hunting Overview by Region12.2.2.1þ Africa In total, 55 species of bats are hunted in Africa, including mainly abundant largebodied fruit bats (Mickleburgh etþ  al. 2009) such as E. helvum, Franquet’s epauletted fruit bat (Epomops franqueti), Gambian epauletted fruit bat (Epomophorus gambianus), hammer-headed fruit bat (Hypsignathus monstrosus), R. aegyptiacus and medium-sized species like Angolan soft-furred fruit bat (Myoncyteris angolensis) (formerly Lissonycteris), Peter’s lesser epauletted fruit bat (Micropteropus pusillus) and to a lesser degree insectivorous bats such as the large slit-faced bat (Nycteris grandis), Maclaud’s horseshoe bat (Rhinolophus maclaudi), Ruwenzori horseshoe bat (Rhinolophus ruwenzori) and Hipposideros species. Although insectivorous bats are considered to be less palatable in many regions and may appear to be under low hunting pressure, (Kamins etþ  al. 2011; Dougnon etþ  al. 2012) this is not necessarily the case. Goodman (2006) showed that in addition to fruit bats,

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330 T. Mildenstein et al.mainly the Malagasy straw-colored fruit bat (Eidolon dupreanum), P. rufus, and the Malagasy rousette (Rousettus madagascariensis), insectivorous H. commer soni is frequently hunted, especially during periods of food shortages. In addition, while fruit bats are probably the most commonly hunted group, 64þ  % of the 55 bat species hunted in Africa are animalivores (Appendix). Hunting bats for food is common in West and Central African states where it can be a major threat to their populations (Funmilayo 1978; Mickleburgh etþ  al. 2009; Kamins etþ  al. 2011). Frequent bat hunting is recorded from Benin Republic, Ghana, Guinea, Liberia and Nigeria (Funmilayo 1978; Anstey 1991; Kamins etþ  al. 2011; Dougnon etþ  al. 2012) (Fig.þ  12.1), as well as in Cameroon, Congo Republic, Democratic Republic of Congo (DRC), Equatorial Guinea and Gabon. High levels of hunting have also been reported in the past from islands off Africa—the Comoros, Madagascar, Mauritius and Rodrigues and São Tomé and Príncipe as well as Pemba Island, Tanzania (Jenkins and Racey 2008; Carvalho etþ  al. 2014), although conservation efforts have reduced this pressure in some of these islands (Trewhella etþ  al. 2005). While occasional bat hunting occurs in Mali and Zambia, there is almost no hunting in East Africa, except eastern Uganda, and bat hunting is rare in South Africa (Mickleburgh etþ  al. 2009). Bats are also persecuted because of negative perceptions in Ethiopia (Mickleburgh etþ  al. 2009) but that is not the focus of this chapter. Although, Halstead (1977) reported the potential for sustainable harvesting of bats at the Ile Ife campus, over-exploitation of E. helvum in southwestern Nigeria was also evident (Funmilayo 1978). People in rural areas in southern Nigeria admit to eating bats occasionally, whereas in parts of eastern Nigeria, R. aegyptiacus is hunted intensively (Fig.þ  12.1). Over 3000 individuals of this species have been collected in one night from a cave in Buanchor village by several hunters who hunt more than once a month (I. Tanshi, unpublished). North Africa and West and Central Asia. Bat hunting is less prevalent in North Africa and West and Central Asia. Of the 98 bat species that occur in this region, ve are known to be hunted and these are for medicinal purposes, of the 98 bat species that occur in this region, ve (all Vespertilionidae) are known to be hunted: long-ngered Myotis (Myotis capaccinii), Geoffroy’s myotis ( Myotis emarginatus), whiskered myotis (Myotis mystacinus), Natterer’s myotis ( Myotis nattereri), Maghrebian myotis (Myotis punicus) (Tableþ  12.1, Appendix). 12.2.2.2þ Asia In Asia, hunting is known to affect 64 species, which represents the largest absolute number of hunted bat species in a region. Southeast Asia. The hunting pressure on bats is greatest in Southeast Asia, where 56, or 17þ  % of the region’s bat species are hunted (Tableþ  12.1, Appendix). Bat hunting is widespread in 10 out of the 11 countries (Brunei, Cambodia, Indonesia, East Timor, Laos, Malaysia, Myanmar, Philippines, Thailand, and Vietnam). Only in Singapore are bats not thought to be hunted heavily (Mildenstein 2012; IUCN 2014).

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331 High levels of hunting occur in Indonesia, where there is a long history of bat consumption (Fujita 1988) and large numbers of individuals are still sold in mar kets (e.g. P. vampyrus, Harrison etþ  al. 2011; Sulawesi fruit bat, Acerodon celebensis, gray ying fox, Pteropus griseus, black ying fox, Pteropus alecto, Heinrichs 2004). Hunting pressure is also high in the Philippines, with a third (24/75) of its species known to be hunted. Although Philippine bats are protected from hunting by the Philippine Wildlife Act and the Philippine Cave Management Act, these laws are not well enforced, and hunting for personal consumption and local trade is widespread. In Malaysia, hunting of some species is regulated, which may curb some of the hunting pressure but has not reduced hunting rates to sustainable levels (Epstein etþ  al. 2009). The laws and levels of enforcement are different for the different regions of Malaysia. All bats are legally protected in Sarawak, but this is not the case in Sabah and peninsular Malaysia. Illegal hunting still occurs in orchards and by sport hunters in Sarawak at places where enforcement is lacking. Legal protection for Old World frugivorous bats is reviewed by Abdul-Aziz etþ  al. (2015). In Buddhist countries (Cambodia, Myanmar, Thailand and Vietnam), most roost sites of large fruit bats are found in the gardens around temples and monasteries because of the protection the monks provide (e.g. Ravon etþ  al. 2014; T. Mildenstein unpublished data). Whether this degree of protection is sufcient to maintain stable populations of these species has yet to be investigated (Tableþ  12.1, Appendix). South Asia. In Bangladesh, large fruit bats are hunted for food by members of tribal groups (Mickleburgh etþ  al. 2009). In India and Pakistan, bats are classied as vermin and are persecuted, although they are consumed infrequently, and more often killed for medicinal purposes (Noureen 2014). The exception is the Indian ying fox (Pteropus giganteus), which is eaten by indigenous forest-dwelling people (Mickleburgh etþ  al. 2009). On the Andaman and Nicobar islands, black-eared Tableþ  12.1þ Proportion of bats hunted by region (Calculated by total number of extant bats species hunted divided by the total number of bat species in the region) Taxon Region Total# On Red List Not on list Total hunted %hunted Chiroptera 1146 97 70 167 14.6 Caribbean islands 106 0 0 0 0.0 East Asia 130 3 4 7 5.4 Europe 42 0 0 0 0.0 Meso America 177 0 0 0 0.0 North Africa 41 3 1 4 9.8 North America 49 0 0 0 0.0 North Asia 43 0 0 0 0.0 Oceania 173 25 15 40 23.1 South America 249 0 8 8 3.2 South and Southeast Asia 365 43 20 63 17.3 SE 333 36 20 56 16.8 South 114 8 5 13 11.4 Sub-Saharan Africa 249 25 26 51 20.5 West and Central Asia 94 1 0 1 11

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332 T. Mildenstein et al.ying fox (Pteropus melanotus) and P. faunulus are hunted and eaten on special occasions (Mickleburgh etþ  al. 2009) (Tableþ  12.1, Appendix). North Asia. Bats are not specically protected in China and many species are eaten, especially in southern China, where bats are found regularly in markets (Mickleburgh etþ  al. 2009) (Tableþ  12.1, Appendix). Requests from international agencies following the SARS outbreak, (which resulted in several hundred human deaths) that wildlife legislation be introduced in China prohibiting inter alia hunting and sale of bats have been ignored. 12.2.2.3þ Pacic (Oceania) Bats are often the only native mammals on remote Pacic Islands, and there is a long history of bat species being hunted in many of these areas. Bats are eaten on American Samoa, the Cook Islands and Niue, the Federated States of Micronesia, Fiji, Guam and the Northern Mariana Islands, New Caledonia, Palau, the Solomon Islands, and Vanuatu (Chambers and Esrom 1991; Mickleburgh etþ  al. 2009). In total, 40 bat species are affected, 23þ  % of Oceania’s bats, making this the region with the highest proportion of hunted bat species on the planet. The value of bat meat is highly variable in Oceania. It is a sought-after delicacy on Guam and the Mariana Islands, where the bats are strictly protected by the United States’ Endangered Species Act (USFWS 2009). In contrast, in the nearby Federated States of Micronesia, the same bat species are rarely eaten (Mickleburgh etþ  al. 2009). In American Samoa, (another United States territory), bats were consumed regularly in the past (Brooke 2001) but are now highly protected. Bat meat is also a delicacy in the Cook Islands, Niue, and Raratonga (Brooke and Tshapka 2002) and is a popular food on Fiji, New Caledonia, and Vanuatu (Mickleburgh etþ  al. 2009). 12.2.2.4þ South America Bat hunting is much less common in South America, occurring in highly localized areas and affecting eight species in the families Phyllostomidae (7 spp.) and Vespertilionidae (1 sp.) (Tableþ  12.1, Appendix).12.3þ Why Bat Hunting is a Conservation Problem 12.3.1þ Negative Impacts on Bat Populations and EcosystemsBats are particularly vulnerable to the effects of hunting for a number of reasons. They are long-lived for their body size (ve species live >30þ  years, Racey 2015) and reproduce slowly, with generally one young per year. They have a slow rate of

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333 fetal growth and long gestation periods (Racey and Entwistle 2000). Females and young bats are thus sensitive to hunting disturbance during a large portion of the year. Bats are nocturnal, making them susceptible to hunting at their roost sites by day when humans can easily nd them. This is especially a concern for the highly sought-after fruit bats in the Old World, which tend to roost conspicuously, aggregating in large numbers in the forest canopy (e.g. Mildenstein etþ  al. 2008). Whether roosting colonies are in caves, cliffs, or trees, hunting at the roost site is likely to affect the entire colony. Hunting disturbance at the roost site causes injury to many bats from the spread of shot gun pellets, large-scale infant mortality when pups fall from eeing mothers (Mildenstein and Stier unpublished data; R. Ulloa, pers. comm.), and higher stress levels as resting bats are startled and forced to ee from hunters (Van der Aa etþ  al. 2006). In their survey of P. rufus in Madagascar, Mackinnon etþ  al. (2003) recorded a high incidence of abandonment of historical roosts, which they attributed, at least in part, to high hunting pressure. Finally, bat colonies are characterized by high roost site delity (e.g. Banack 1996; Brooke etþ  al. 2000; Gumal 2004; Stier and Mildenstein 2005). Hence, bats may be reluctant to leave when hunting starts and may nd it difcult to nd alter native roost sites after eeing hunters. Because bats are likely to eventually return to the preferred roost site, they are predictable prey for hunters. The overall effect of hunting at roost sites is reduction of bat population densities to a fraction of local carrying capacity (e.g. Mildenstein 2012). These population-level impacts may also have negative ecological consequences. Some bat species play prominent roles in insect population control, pollination, seed germination and dispersal, and in many areas, bat species are essential to forest regeneration (e.g. large fruit bats are primary seed dispersers for hemiepiphytic gs, Shanahan 2001; Oleksy etþ  al. 2015a). On isolated islands, where there is little ecological redundancy, bats are often recognized as keystone species due to their unique roles in seed dispersal (Shilton and Whittaker 2009). Mortality due to hunting may, therefore, have cascading effects on ecological communities (e.g. Mildenstein 2012) and ecosystem function (e.g. McConkey and Drake 2006).12.3.2þ Negative Impacts on HumansThe negative impacts of bat hunting extend beyond natural ecosystems to human communities. Bats in their natural ecological roles perform valuable ecosystem services benecial to humans (e.g. insect suppression: Cleveland etþ  al. 2006, pollination: Bumrungsri etþ  al. 2008b, 2009, seed dispersal maintaining local water sheds: Banack 1998; Stier and Mildenstein 2005), all of which are reduced when bats are hunted. Bat colonies have also proved valuable as eco-tourism attractions supporting local economies (e.g., in Costa Rica, Lao People’s Democratic Republic, Madagascar, the Philippines, and North America; examples in Pennisi etþ  al. 2004). However, hunted bats that are wary of human presence often do not

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334 T. Mildenstein et al.maintain colonies in locations easily viewed by people. Hence, reductions in bat populations as a result of hunting could have expensive ramications on local communities’ water supplies, agriculture, and eco-tourism industries. Finally, the hunting of bats may also expose human communities to potentially zoonotic pathogens (Leroy etþ  al. 2005). In the past decade, considerable attention has been paid to bats as natural reservoirs of emerging infectious diseases (Calisher etþ  al. 2006). Studies that link infectious disease outbreaks to bats demonstrate the spillover potential through contact with bats or exposure to faeces and urine in bat habitats (reviewed by Plowright etþ  al. 2015). Most notable are the Ebola virus outbreaks, which have attracted international attention. Leroy etþ  al. (2009) suggest that the 2007 emergence of Ebola virus in the Occidental Kasai province of DRC could be attributable to the consumption of freshly killed bats. The authors trace the virus spread from a rst patient with bat bushmeat contact to an outbreak of the disease in 260 persons resulting in 186 deaths in 2007. The re-emergence of the disease in 2014 may also have arisen from contact with bats (Saéz etþ  al. 2015) and has proven far more deadly.12.4þ Overhunting as a Growing Concern for ConservationHuman communities have long exploited bat populations for consumption. Current hunting pressure, however, is likely to be much greater than historical pressure with increases in human population density, greater accessibility to natural areas, technological advances in bat capture methods and transport options, and relaxed adherence to cultural taboos (Brooke and Tschapka 2002; Millner-Gulland and Bennett 2003). Hence, bat hunting is likely to be unsustainable (Bradshaw etþ  al. 2009), especially when coupled with other anthropogenic stressors (such as those described throughout this book). Overhunting (commonly also “unregulated” hunting, although not all unregulated hunting is unsustainable, nor regulated hunting sustainable) is a globallyrecognized threat to many wild species of animals (Robinson and Bennett 2000; Milner-Gulland and Bennett 2003). For bats, overhunting has been a conservation concern for over three decades (Lemke 1986; Mickleburgh e