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Dupont, Jennifer Maria.
Ecological dynamics of livebottom ledges and artificial reefs on the inner central West Florida Shelf
h [electronic resource] /
by Jennifer Maria Dupont.
[Tampa, Fla] :
b University of South Florida,
Title from PDF of title page.
Document formatted into pages; contains 164 pages.
Dissertation (Ph.D.)--University of South Florida, 2009.
Includes bibliographical references.
Text (Electronic dissertation) in PDF format.
ABSTRACT: The West Florida Shelf (WFS) is one of the largest and most diversely-used continental shelf/slope systems in the world. The presence of paleoshorelines and scarped hardbottom outcrops (up to 4 m in relief) along the inner shelf (10-30 m depth) provide important habitat for a variety of infaunal, epifaunal, and fish assemblages that contribute to the productivity of the region. This dissertation will present a comprehensive overview of the geological, physical, and chemical settings of the inner West Florida Shelf, with particular focus on biological and ecological community dynamics of epibenthic macroinvertebrates, algae, and fish assemblages. Baseline and comparative data sets are presented in the form of historic and modern species lists, with focus on seasonal and intra-annual variations. Quantitative effects of disturbances (e.g., hurricanes, thermal stresses, and red tides) and subsequent recovery rates are discussed as they periodically perturb inner-shelf systems and can have significant effects on community structure. Benefits of and recommendations for using artificial reefs as restoration tools along the inner shelf, as mitigation for future disturbances, are presented.
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Advisor: Pamela Hallock Muller, Ph.D.
Gulf of Mexico
x Marine Science
t USF Electronic Theses and Dissertations.
Ecological Dynamics of Livebottom Ledges and Artificial Reefs on the Inner Central West Florida Shelf by Jennifer Maria Dupont A dissertation submitted in partial fulfillment of the requirements for the degree of Doctor of Philosophy Department of Marine Science College of Marine Science University of South Florida Major Professor: Pamela Hallock Muller, Ph.D. Walter Jaap, M.S. Gabriel Vargo, Ph.D. David Mann, Ph.D. John Ogden, Ph.D. Date of Approval: January 15, 2009 Keywords: Gulf of Mexico, restoration, di sturbance ecology, red tide, epibenthos Copyright 2009, Jennifer Maria Dupont
Dedication This is dedicated to my parents, Lars and Marianne Dupont. Thank you for your love, support, and inspiration. I am the luckiest da ughter ever and I hope th at I have made you proud.
Acknowledgements I thank my advisor, Pamela Hallock Muller and Â“co-adviso rÂ”, Walter Jaap, for their input, encouragement, and support throughout my time at USF. I also thank the rest of my committee, Gabriel Va rgo, John Ogden, and David Mann, as well as my defense chairperson, John Lawrence, for their assistance and expertise. E. Watkins, A. Taylor, G. Matzke and Gulfstream Natural Gas Syst ems provided boat support, photographic and survey data, and personnel to support the artificial reef wor k. The monthly samplings at the natural livebottom ledges would not have been possible without assistance from C. Coy, M. Terrell, and volunteer divers at th e Florida Aquarium. The HAB group at the Fish and Wildlife Research Institute, particularly R. Pigg, conducted water sample phytoplankton counts and provided environmenta l data from the 2005 red-tide event. I thank D. Palandro (FWRI) for assistance with maps and figures and N. Ogden for algal identification. Volunteer divers that assisted with data co llection include students from the SCUBAnauts, International; W. Dent; B. Meister; M. Peters; M. Stephens; and J. Galkiewicz. Funding was provided by the USF College of Marine Science Endowed Fellowships including the Gulf Oceanogra phic Charitable Trust Fellowship and the William and Elsie Knight Fellowship, and by the Florida Sea Grant/Royal Caribbean Scholarship.
i Table of Contents List of Tables iv List of Figures v Abstract x 1. Overview of the Inner Central West Florida Shelf 1 1.1. Introduction 1 1.2. Geological Setting 4 1.3. Physical Oceanography 8 1.4. Chemical Oceanography 10 1.5. Biological Oceanography 11 1.5.1. Faunal Zones 11 1.5.2. The 2005 Red Tide 16 1.6. Overview of Dissertation 20 2. Central West Florida Shelf Natural Ledge Dynamics 22 2.1 Abstract 22 2.2 Introduction 23 2.3 Historic Data 26 2.4 Background and Rationale for Current Study 30 2.5 Methods 31 2.5.1. Site Characteristics 31 2.5.2. Benthic Community Data 32 2.5.3. Fish Assemblage Data 35 2.5.4. Abiotic Data 37 2.6 Results 37 2.6.1. Historic Data 37 2.6.2. Benthic Community Data 48 2.6.3. Fish Assemblage Data 57 2.6.4. Abiotic Data 60 2.7 Discussion 63 2.7.1. Benthic Communities 62 2.7.2. Juvenile Coral Recruitment 66 2.7.3. Fish Assemblages 70 2.7.4. Marginal West Florida Shelf Assemblages and Disturbance 73
ii 3. Artificial Reefs as Restoration Rools: A Case Study on the West Florida Shelf 81 3.1 Abstract 81 3.2 Introduction 82 3.3 Artificial Reef Project Background and Objective 84 3.4 Artificial Reef Site Description 84 3.5 Environmental Conditions 86 3.6 Artificial Reef Size and Design 87 3.7 Monitoring Methods 89 3.8 Results and Performance Evaluation 91 3.9 Ecological and Legal Framework 97 3.10 Conclusion 98 4. Ecological Impacts of the 2005 Red Ti de on Artificial Reef Epibenthic Macroinvertebrate and Fish Communities in the Eastern Gulf of Mexico 100 4.1 Abstract 100 4.2 Introduction 101 4.3. Methods 103 4.3.1. Study Area Characteristics 103 4.3.2. Benthic Communities 105 4.3.3. Fish Communities 109 4.4 Results 110 4.4.1 Study Area Characteristics 110 4.4.2. Benthic Communities 112 4.4.3. Fish Communities 115 4.5 Discussion 123 4.5.1. Benthic Communities 123 4.5.2. Fish Communities 126 4.5.3. Red Tides as a Community Structuring Force 131 5. Enhancement of Natural Ledge Substr ate Via Deployment of Artificial Reefs Along the West Florida Shelf 133 5.1 Introduction 133 5.2 Defining Current Artifici al Reef Data issues 134 5.3 Problem #1: Rationale for In cluding Benthic Communities in Production Calculations 135 5.3.1. Artificial Reef Contributions 135 5.3.2. Substrate, Benthic Fauna, and Increased Food Availability 136 5.3.3. Provision of Shelter 140 5.3.4. Recruitment Habitat 141 5.4 Problem #2: Rationale for Dism issing Â“Lack of Definitive DataÂ” Argument Against Artificial Reef Use Along the West Florida Shelf 143 5.5. Conclusion 147 References 149
iii About the Author End Page
iv List of Tables Table 2.1. Coordinates of sites sampled 1965-1967 (Station B), 2005-2007 (Artificial Reefs), and 2 006-2007 (FWRI1 and MT). 26 Table 2.2. Coral species r ecorded at shallow har dbottom sites on the inner west Florida shelf. Data are compiled from three data sets and species presence is denoted by the corresponding number: (1) 1965-1967 (Hourglass Program), (2) 2006-2007 (Natural Reef Ledges) and (3) 2005-2007 (Artif icial Reefs). Reproductive modes are also listed. N/K=Not known. 39 Table 2.3. Echinoderm species recorded on shallow hardbottom sites on the inner WFS. Data are compiled fr om three data sets and species presence is denoted by the corresponding number: (1) 1965-1967 (Hourglass Program), (2) 2006-2007 (N atural Reef Ledges) and (3) 2005-2007 (Artificial Reefs). 40 Table 2.4. Benthic algae species record ed on shallow hardbottom sites on the inner WFS. Data are compiled fr om three data sets and species presence is denoted by the corresponding number: (1) 1965-1967 (Hourglass Program), (2) 2006-2007 (N atural Reef Ledges) and (3) 2005-2007 (Artificial Reefs). 41 Table 2.5. Fish species recorded on sha llow hardbottom sites on the inner WFS. Data are compiled from thre e data sets and species presence is denoted by the corresponding number: (1) 1965-1967 (Hourglass Program), (2) 2006-2007 (Natural Reef Ledges) and (3) 2005-2007 (Artificial Reefs). 46 Table 2.6. ANOVA and TukeyÂ’s posthoc comparison results for the 4 categories that displayed significan t seasonal differences. Results from the TukeyÂ’s test first list the categoryÂ’s determining season followed by the season(s) that it differs significantly from. 56 Table 2.7. ANOSIM analys is and SIMPER pair-wise comparisons of fish assemblages during fall, winter, spring, and summer samplings. Only those seasons that differed significantly from one another (ANOSIM R>0.3; p<0.05) are shown. The SIMPER results list the
v 4 top species contributing to th e dissimilarity between the two seasons. Their relative abundance ch ange (+/-) is also listed. 60 Table 2.8. Natural disturbances acti ng on WFS livebottom areas. The Â“level influencedÂ” column specifies whethe r benthic (B) or fish (F) levels are most influenced by the resp ective disturbance process. 79 Table 3.1. Abiotic and habitat characterization data summary presented as mean (n=10) values ( S.E). Da ta were recorded during each of the 5 sampling times. 87 Table 4.1 Abiotic and habitat characterization data summary presented as mean values ( S.E). Data were recorded during each of the five fish censuses. 111 Table 4.2. Results from the SIMPER test to determine discriminating benthic categories for pairs of sampling times that differed significantly (ANOSIM R 0.5, p 0.5%). Discriminating categories satisfy the conditions of contributing signif icantly and consistently to the average dissimilarity. The average percent cover change of each category is shown in the last column. 115 Table 4.3. List of the 71 fish species observed during census activites from March 2005 to March 2007 at GNGS ar tificial reef sites including limestone boulder (LB), reef module (RM), and reference sites. Commercial importance and primary habitat are noted. 116 Table 4.4. Matrix of significant (S) and non-significant (N) temporal fishassemblage trends at LB, RM, a nd R sites, respectively. Summer 2005 (S05) sampling occurred prior to the peak of the red-tide event. Summer 2007 (S07) represents the final sampling in the focused two-year time series. 120 Table 4.5. Results from SIMPER analyses performed on significantly different assemblages (determined by ANOSIM in Table 4.4 above) to determine discriminating species and their average change (+/) in abundance between time periods. AP = Archosargus probatocephalus, SS = Serranus subligarius, SV = Stegastes variabilis, MP = Mycteroperc a phenax, LM = Lachnolaimus maximus, CS = Chasmodes saburrae, HB = Halichoeres bivittatus, CC = Caranx crysos, DF = Diplectrum formosum, HP = Haemulon plumierii, CX = Calamus sp., CB = Calamus bajonado, LR=Lagodon rhomboides. 121
vi List of Figures Figure 1.1. Offshore (40-50 km) site s surveyed extending from Clearwater Beach south to the mouth of Ta mpa Bay. FWRI1, MT, and Station B are natural hardbottom ledges an d the GNGS sites are a set of artificial reefs deployed in 2001. 5 Figure 1.2. Faunal zone delineations of the WFS (not to scale) as proposed by Lyons and Collard (1974). Chap ters 2-5 will concentrate areas along the inner shelf zone (10-30 m). 13 Figure 2.1. Map (Google Earth) of the study sites located along the shallow inner west Florida shelf. 28 Figure 2.2. Examples of hardbottom ledge communities along the shallow inner west Florida shelf. 29 Figure 2.3. Boxplot of pe rcent coral cover at FWRI 1 and MT from February 2006 to December 2007. The bars represent the interquartile ranges, sample means are designated by a diamond, and medians by a horizontal line. 50 Figure 2.4 Boxplot of macroalgal cove r at FWRI1 and MT from February 2006 to December 2007. 51 Figure 2.5. Boxplot of bare substrate c over at FWRI1 and MT from February 2006 to December 2007. 51 Figure 2.6. Regression analysis of macroalgae vs. bare substrate. 52 Figure 2.7. Boxplot of percentage of tran sect photos in which juvenile corals were identified at FWRI1 and MT from February 2006 to December 2007. 53 Figure 2.8. Boxplot of Porifera cover at FWRI1 and MT from February 2006 to December 2007. 54 Figure 2.9. Boxplot of other living fa una cover at FWRI1 and MT from February 2006 to December 2007. 54
vii Figure 2.10. Boxplot of dead coral with algae cover at FWRI1 and MT from February 2006 to December 2007. 55 Figure 2.11. Boxplot of bleached coral c over at FWRI1 and MT from February 2006 to December 2007. 55 Figure 2.12. Principal Components Analys is of Log (x+1) transformed data from seven major categories plus ju venile corals. Eigen values for the first two components are 2.7 and 1.2, respectively, and cumulatively explain approximate ly 50% of the variance. 57 Figure 2.13. Non-metric multidimensional scaling ordination (2-D) of fish assemblage samples over the 22-month study period. Separation of seasonal groups [e.g., July-Sep data grouped towards upper right is confirmed via a 3-D analysis wh ich yields a lowe r stress (0.14)]. The 3-D graph is not shown due to poor visual representation. A priori groupings were analy zed based upon those grouping. 59 Figure 2.14. Ten-day average bottom temp erature data at FWRI1 from March to December, 2007. 61 Figure 2.15. Daily fluctuations in bot tom temperature during March 2007 at FWRI1. Data were collected every 10 minutes and daily averages are plotted. 62 Figure 2.16. Photograph depicting cond itions at FWRI1 in February 2006. Anomalously high percent cover of the Â“MacroalgaeÂ” category was attributed to this unidentified algal growth. The growth had disappeared by the April 2006 sampli ng. The coral in the picture is Solenastrea hyades. 64 Figure 3.1. Artificial reef module de signed by H. Hudson (U.S. patent #5215406) and constructed of limestone in a concrete matrix (each module occupies 8.7 m3). The cavity passes through the entire length of the module. Three groups of 17 modules were placed at each of the three Reef Module (RM) artificial reef habitats. 85 Figure 3.2. Species richness trends over the five samplings: Summer 2005 (S05), Winter 2005 (W05), Summer 2006 (S06), Winter 2006 (W06), and Summer 2007 (S07). Th e data are part of a larger study that assessed the effects of a Karenia brevis (red tide) bloom that passed through the area imme diately after the Winter 2005 sampling, extirpating a majority of the benthos and a ltering the fish assemblage. Recovery trajectories of species richness at the three
viii habitat types (RM Â– Reef Modules, LB Â– Limestone Boulders, and Ref Â– Reference) are very similar. 92 Figure 3.3. Sampling distribution plots (i ncluding median, interquartile range, upper and lower limits, and outliers) of the five commercially important fish species whose a bundances were significantly higher (ANOVA F> 12.0, p<.0001) at the artif icial reef hab itat types (RM and LB) as compared to the natural hardbottom/reference (R) habitats. An additional seven co mmercial species displayed no significant differences among habitats. 95 Figure 4.1. The location of GNGS Lime stone Boulder (LB) and Reef Module (RM) sites in the eastern Gulf of Mexico. Reference (R) sites were located in close proximity. 104 Figure 4.2. Example of 1 m2 photo-mosaic from stat ion #84 (~17 m). Four photos were combined to produce composite images that were used in point-count analyses. 106 Figure 4.3. Environmental parameters taken along a 24 kilometer east-west transect positioned 5-7 kilometers north of the Gulfstream Natural Gas Systems artificial reefs. 112 Figure 4.4. Multidimensional scaling ordination of 1 m2 benthic quadrats during the four sampling periods. Th e arrows depict the theoretical temporal trajectory of community response. 113 Figure 4.5. Temporal changes in total fi sh species richness at the LB, RM, and Reference sites. The Summer 2005 (S05) census was conducted prior to the red-tide event; Winter 2005 (W05) data were collected during and immediately after the event. 128 Figure 5.1. Percent cover of coral at GNG S artificial reefs. Data are shown for individual 1 m2 photo-quadrats captured during each of the 4 sampling times (March 2005, August 2005, June 2006, and March 2007). 137 Figure 5.2. Percent cover of poriferans at the GNGS artificial reefs; source of data as in Fig. 5.1. 137 Figure 5.3. Percent cover of other li ving fauna (primarily echinoderms and ascidians) at GNGS artif icial reefs; source of data as in Fig. 5.1. 138 Figure 5.4. (A) Macroalgal percent cove r at GNGS artificial reefs which varies inversely with (B) bare substrate cover. Samplings (March 2005,
ix August 2005, June 2006, and March 2007) do not display seasonal trends. 139 Figure 5.5. (A) Macroalgal percent cove r at natural ledges (FW=FWRI1 and M=Mastedon Tabletop) which va ry inversely with (B) bare substrate cover. 140 Figure 5.6. Bubbleplot depicting the average (n=3) perc entage of phototransects containing at least one juvenile (<2 cm) coral over the 22month sampling period. 143
x Ecological Dynamics of Livebottom Ledges and Artificial Reefs on the Inner Central West Florida Shelf Jennifer Maria Dupont ABSTRACT The West Florida Shelf (WFS) is one of the largest and most diversely-used continental shelf/slope systems in the worl d. The presence of paleoshorelines and scarped hardbottom outcrops (up to 4 m in re lief) along the inner shelf (10-30 m depth) provide important habitat for a variety of infaunal, epifaunal, and fish assemblages that contribute to the productivity of the re gion. This dissertation will present a comprehensive overview of the geological, physi cal, and chemical settings of the inner West Florida Shelf, with particular fo cus on biological and ecological community dynamics of epibenthic macroinvertebrates, al gae, and fish assemblages. Baseline and comparative data sets are presented in the form of historic and mode rn species lists, with focus on seasonal and intra-annual variations. Quantitative effects of disturbances (e.g., hurricanes, thermal stresses, and red tides) and subsequent recovery ra tes are discussed as they periodically perturb inner-shelf syst ems and can have significant effects on community structure. Benefits of and r ecommendations for using artificial reefs as restoration tools along the inner shelf, as mitig ation for future disturbances, are presented.
1 1. Overview of the Inner Central West Florida Shelf 1.1. Introduction The West Florida Shelf (WFS) is one of the largest and most productive continental shelf/slope systems in the world. It covers 170,000 km2 and extends more than 200 km west from the intertidal zone to the 200 m isobath across a very gentle slope (<<1) of ancient limestone platforms (Okey et al. 2004). The WFS is characterized by a range of seafloor morphologies, gradients, sediment types, biotic communities, reefal structures, and paleo sea-level indicators. Due to the importance of continental shelf resources to the State of Flor ida, including the prolific finfish and shellfish fisheries, offshore petroleum and natural gas explorati on, and tourism industries, the WFS has been the subject of numerous studies that addr ess the unique physical oceanographic regimes, chemical influences, and geologic features of the dynamic area. Despite the robust collection of WFS works, there is a large gap in knowledge of spatial distributions of benthic fauna and flora (epifaunal and infaunal), and temporal changes in these communities. This is su rprising considering that a number of the prolific finfish that populate the WFS (and which are the main ta rgets of economicallyimportant commercial and recreational fisheries) utilize virtually every portion of the broad continental shelf at some point in th eir life history. For example, gag grouper, Mycteroperca microlepis (one of the most valuable finf ishes in the southeastern United States and a ubiquitous staple at Florida seafood restaurant s), aggregate and spawn in
2 deep (>70 m) shelf waters (Coleman et al. 1996). Following spawning, females move into shallower waters (<30 m) while males drift off into deeper waters (McGovern et al. 1998). The larvae drift inshore and postlarv ae recruit to seagrasses, mangrove creeks, and coastal estuaries and lagoons where they remain for 3-5 months. They then move offshore to reefs and ledges along the WFS (Ross and Moser 1995) where the long-lived, slow-growing, protogynous hermaphrodites matu re between year 5 and 6 to repeat the process. Mycteroperca microlepis is one of a number of economically-important fish species that spend time traversing the ledge s of the inner WFS. Through various life stages, they rely on limestone outcrops th at support diverse livebottom (reef-like) communities and demersal fish assemblages, and which occupy approximately 50% of the inner WFS (Locker et al. 2003; Obrochta et al. 2003; Hi ne et al. 2008). Although tropical reef development is absent al ong the inner WFS (Jaap 1984), the extensive systems of scarped hardbottom provide habi tat (up to 4m relief) and support an association of hardy corals a nd other biota. The hardbotto ms also provide structure, protection, and abundant food sources for demers al and pelagic fish species that inhabit the areas. Shallow inner WFS livebottom ledges are biotic oases along the otherwise monotonous, quartz-sand dominated inner WFS. Scientists, managers, conservationists, and fishermen need to understa nd the spatial and temporal dyna mics that operate in these areas, as they are inextricably linked to the productivity of the region. Disturbances, including hurricanes, ther mal stresses, and harmful algal blooms, frequently affect areas along the WFS, whic h are already marginal with respect to a number of first-order determinants for reef assemblages including temperature, nutrient,
3 light, and aragonite saturation regimes. Acute disturbanc es, combined with chronic marginal conditions, are important comm unity-structuring forces along the WFS livebottom ledges. Baseline conditions for WFS benthic and fish assemblages must be defined, especially as reefs continue to be stressed by global change including ocean acidification and ri sing sea level. In this introductory chapter, I will pres ent a review of the general geological, physical, and chemical processes that influen ce the biological assemblages of the WFS, with particular focus on inner shelf areas (1030 m depth) along central west Florida. The studies on natural ledges and artificial reefs (15-20 m dept h) presented in this paper were initiated in response to a massive red tide ( Karenia brevis ) and associated hypoxic/anoxic event in 2005 that affected approximately 5,600 km2 of benthic communities and fish assemblages along the inner west central Florida shelf (FWRI 2005), causing substantive economic losses in the ar ea. Mass mortalities of invertebrates, fish, and marine mammals, along with advers e human effects (e.g., re spiratory issues), quickly raised interest in understanding th e dynamics and effects of the harmful algal bloom events that regularly affect the area wi th varying severity. Interest in mitigating the harmful algal blooms has peaked signifi cantly in response to the massive 2005 red tide, and a number of studies and experi ments are currently under way seeking to eliminate these Â“problem bloomsÂ”. The real problem is, however, that there are few studies that quantitatively document the eff ects of red-tide events and the subsequent recovery processes that take place along the WFS. Vargo et al. (1987) showed that red tides have the potential of contributing greatly to the primary production and annual carbon input along the WFS, and may be essent ial in ensuring the continued productivity
4 of the region. In addition, although natura l disturbances such as red tides can be detrimental to individuals and communities at large spatial scales (10-1000 km), new substratum becomes available at smaller tem poral and spatial scal es (Connell 1978). Patches of opportunity are opened for renewa l, development, and community succession (Holling 1996) and the current di versity of scarped hardbottoms and their associated fish assemblages, may depend on the red-tide events. In the chapters that follow this introd uction, I will present a data set on the seasonal dynamics of shallow inner WFS livebottom ledges (abi otic and biotic data). I will quantify the impacts of the 2005 red-ti de event on artificia l reef communities (epibenthic and demersal fish), and discuss th e use of artificial reef s as restoration and conservation tools along the WFS in context of future disturbances including, but not limited to, red tides. 1.2. Geological Setting There is a robust body of knowledge on th e formation and current geological setting of the extensive WFS a nd I refer readers to review sources including papers in Marine GeologyÂ’s Special Issue #200 (2003) and Hine et al. (2008), which provide a comprehensive set of papers that discuss shelf origin, sand ridges, transverse bars, sediment distribution, and many other topics in great detail. The west-central coast of Florida, exte nding from Anclote Key in the north to Cape Romano in the south, is an estuarine, barrier island, inner shelf system of marked contrasts, contradictions, and significant char acteristics (Hine et al. 2003). I will focus on describing areas along the central west Florid a Shelf extending from offshore (~40 km)
5 Clearwater Beach south to approximately 50 km west of the mouth of Tampa Bay (Fig. 1.1), since sites along this area are discussed in future chapters. Figure 1.1. Offshore (40-50 km) sites survey ed extending from Clearwater Beach south to the mouth of Tampa Bay. FWRI1, MT, and Station B are natural hardbottom ledges and the GNGS sites are a set of artificial reefs deployed in 2001. The central WFS is situated between the siliciclastic sand-dominated northwest shelf off the Florida Panhandle, whic h is significantly infl uenced by rivers and river deltas, and a carbonate-dominated shelf off the southwest Florida Peninsula, which is characterized by reefs, inner shelf carbona te muds, outer shelf skeletal sands, and lithified submerged calcarenitic (oolitic/skelet al grainstones) paleoshorelines (Hine and Locker 2006). The central WFS is a vast tr ansition zone that has been starved of both siliclastic and carbonate sediments, and is therefore characterized by extensive outcrops of karstified-deformed, biologically-eroded Ne ogene-Quaternary limestone surfaces. The
6 outcrops occupy approximately 50% of the sh elf seaward of 5 km (Locker et al. 2003) and support a diverse benthic community cove ring a surface that has sinkholes, elevated terraces, rock ledges, and scarps (Hine and Locker 2006). The outcrops can provide as much as 4 m of relief and are veritable oases of biotic productivity (i.e., epibenthic and fish assemblages) surrounded by mobile sediment. The surficial sedimentary pattern has been reported to consist of a nearshore band of fine-grained, quartz -rich (>75% quartz) sand, shifting offshore into coarsegrained, carbonate-rich (>75% CaCO3) sand and gravel (Doyle and Sparks 1980), although patchy distributions of other sedime nt types are common al ong the central inner WFS (Brooks et al. 2003). The complex and pa tchy distribution of sediments represents multiple sediment sources. The shallow inne r shelf areas (10-30 m depth) that are the focus of the rest of this dissertation are dom inated by fine-grain, quartz-rich sands which form a thin veneer (<3 m) over the kars tic limestone surface (Doyle and Sparks 1980; Holmes 1981). Biogenically-derived carbonate sediments, primarily of the mollusk-rich foramol assemblage characteris tic of non-tropical carbonate sy stems, are also present in association with hardbottom outcrops. The car bonate component reflects the influence of the living assemblages along the WFS (Brooks et al. 2003). In addition, phosphorite-rich sands that exist as a thin veneer on a major ity of the hardbottoms are likely the product of the reworking of underlying phospha te-rich strata. Input rates of all the sources to the surface sediment are unknown although it is unlike ly that they are very high as evidenced by the thin and patchy sediment cover (Brooks et al. 2003) The distribution of hardbottom outcrops and movement of sediments are important aspects in determining where WF S livebottoms can develop. Though there are
7 a number of abiotic factors that affect the distribution and abundance of benthic invertebrates (i.e., temperatur e, salinity, turbidity, current s, wave shock, and dispersal barriers), the availability of suitable substrate is the singl e most important factor in offshore areas along the WFS wher e abiotic parameters are less variable as compared to nearshore areas (Collard and DÂ’Assaro 1973; Lyons and Collard 1974). Although temperature and salinity fluctuations tend to be limiting in estuarine environments, such parameters become more constant in offshore areas, where bottom substrate and overlying water mass characteristics become criti cal factors. The availability of suitable substrate (in the form of emergent hardbottom) may be particularly important for larval stages of corals, which depend on the abil ity of the larvae to identify a suitable substratum for settlement where they can metamorphose and grow colonies (Richmond 1997). Bare substratum may occasionally be generated (i.e., by shifting sediments) and the frequency and duration of exposure of suitable substrate is one limiting factor in epibenthic macroinvertebrate recruitmen t and colonization. Biological community structuring forces (i.e., predation, compe tition, physiological tolerance, and population attributes) are also essentia l in determining th e abundance and distribution of benthic invertebrates. Brooks et al. (2003) studied the pattern s and control of surface sediment distribution along the west-centr al Florida inner shelf and determined that the patchy distribution of sediments indica tes that a majority of the sand grains reside in close proximity to where they were origina lly deposited (e.g., phosphorite-rich sand is consistently found surrounding hardbottoms and th e source is attributed to the underlying phosphatic limestone). Typical large-scale sediment distribution mechanisms (storms
8 and tides) do not appear to regularly influen ce the regional distribution of sediment. The distribution pattern is more a reflection of sediment source than the mechanism of transport (Brooks et al. 2003). Small-scale, pe riodic events that mobilize and redistribute sediment (e.g., storms and tides) along the WF S have been reported by Twichell et al. (2003) and are locally important in distribu ting sediments. These events could be expected to periodically e xpose or cover local low-relief hardbottom outcrops, affecting the sessile and slower-moving flora and fauna th at inhabit the areas (e.g., corals, Porifera, algae, echinoderms, etc.). The small-scale disturbance events are discussed in conjunction with larger-scale di sturbances such as hurricanes and red-tide events in Chapter 2. 1.3. Physical Oceanography The west Florida coast is a low-energy coast with mean annual wave heights of 10-25 cm (Tanner 1960) and tidal ranges < 1 m (Davis 1989). As discussed above, these processes are incapable of re gional-scale sediment redistri bution. Circulation along the WFS is very complex, driven to different degrees by winds, tides, and buoyancy fluxes, and is also influenced by the prominent GOM circulation feature to 1000 m depth, the Loop Current (He and Weisberg 2003). The Loop Current connects the Yucatan Current to the Florida Current via its northern, cloc kwise flow into the GOM. Variability in the Loop Current and spin-off eddies have been studied and modeled ex tensively (Hurlburt and Thompson 1980; He and Weisberg 2003) and may be important factors in determining the ecological diversity of bent hic and fish communitie s throughout the Gulf of Mexico (GOM). Along the eastern GOM, variations in the Loop Current and the
9 separation of anticyclonic eddies or rings can strongly influence benthi c habitats in terms of larval and nutrient supply (Richards et al. 1993). Tidal currents are relatively weak al ong the WFS (He and Weisberg 2002) and subtidal sea level and current variations are correlated with synoptic-scale wind variations (Mitchum and Sturges 1982). M onthly mean currents mid-shelf suggest a seasonal cycle with along-shore flows oriented southeast in the spring and northwest in fall (He and Weisberg 2002). The southeastward spring com ponent advects river waters, including from the Mississippi River, formi ng a low salinity tongue that often carries a chlorophyll plume southward along the WFS. There appears to be a distinct separati on between shelf-break currents (controlled by the Loop Current) and the inner-shelf cu rrents (controlled by local winds), although the unique geometry of the WFS does lend itse lf to intermittent Loop Current intrusions into shallower isobaths, shoreward of the shelf break (He and Weisberg 2003). These types of intrusions, during upwelling-favorab le winds, contribute deep, nutrient-rich waters to areas of the WFS and have been im plicated in stimulating blooms of harmful algae such as Karenia brevis (see biological section below) He and Weisberg (2003) speculate that bottom topogra phy and coastline geometry ar e important in generating regions of convergence and divergence along the WFS, and may create upwelling centers. The Florida Big Bend region (whe re the shelf break is 20 m deeper thereby requiring less upwelling for the deep waters to broach the shallower shelf) is one area that may serve as a communication center between deep GOM waters and the WFS (He and Weisberg 2003). There are numerous other areas that could also contribute to mean seasonal upwelling, greatly influencing nutrient concentr ations and productivity all along the WFS.
10 1.4. Chemical Oceanography The GOM is traditionally classified as an oligotrophic system (El-Sayed et al. 1972; Biggs 1992), although Muller-Karger et al. (1991) did show, via satellite measurements, that strong seasonal change s in shelf production occur along the WFS. There is evidence, however, that although the open-ocean, pelagic GOM is oligotrophic, the waters that overlie the shallow inner WFS are not. The intermediate nutrient conditions are discussed fu rther in Chapter 2. Gilbes et al. (2002) attempted to statisti cally explain cross-shelf and along-shelf differences in nutrients, suspended sediments, and optical properties (diffuse attenuation coefficients) as they related to phytopla nkton production and the development of a seasonal plume. They sampled along a transect (3 stations: nearshore, mid-shelf, and offshore) leading southwest from the mout h of Tampa Bay, and combined the results with data from the northwestern GOM. Alt hough their results were obtained from only one cruise, some general tre nds were presented. Nearshor e stations along the WFS were characterized by high nutrient concentrations, low salinities, high su spended sediments, and high diffuse attenuation coefficients. These stations reflected th e influence of river discharge from the nearshore, coastal areas. On middle-shelf and offshore stations, an increase in salinity was accompanied by a d ecrease in nutrients, suspended sediments, and diffuse attenuation coefficients, along w ith surface pigments, indicating that these areas are less influenced by river inputs. Th e stations southwest of Tampa Bay (situated closest to those sites discussed in this dissertation) were characterized by NH4 + concentrations between 0.0 and 0.3 M, NO2 + NO3 levels between 0.2 and 0.4 M, dissolved organic nitrogen (DON) between 6.0 and 9.0 M, total dissolved nitrogen
11 (TDN) between 6.0 and 10.0 M, total partic ulate nitrogen (TPN) between 1.0 and 2.0 M, PO4 between 0.05 and 0.12 M, dissolved organic phosphorous (DOP) between 0.10 to 0.17 M, total dissolved phosphorous (TDP) between 0.15 and 0.25 M, total particulate phosphorous (TPP) between 0.2 and 0.4 M, and Si at 0.0 M (Gilbes et al. 2002). Nitrogen, not phosphorous, is generally th e limiting nutrient along the WFS as the shallow shelf is situated in a broad phos phatic province. Normal background nitrate levels in the GOM are <0.1 M although the combination of upwelling-favorable west winds and the complex physical oceanographi c dynamics along the WFS, have caused significant increases in nitrat e concentrations (up to 3.31 M in near-bottom stocks) along the 20 m isobath from the Big Bend area to southeastern regions (Walsh et al. 2003). Nitrogen limitations can also be lifte d through actions of diazotrophs such as Trichodesmium (Lenes et al. 2001) and the effects of such blooms are discussed in the biological section below. The effects on benthic communities and fish populations, as well as on phytoplankton assemblages, vary from year to year but can be significant in structuring WFS biotic assemblages. 1.5. Biological Oceanography 1.5.1. Faunal Zones Much of the WFS may be considered eco tonal between the temperate Carolinian and tropical Caribbean (or West Indian) z oogeographic benthic inve rtebrate provinces (Hedgpeth 1957; Lyons and Collard 1974), meani ng that hardy constitu ents of both occur across the WFS. One of the primary source s of detailed, exploratory information into WFS biological diversity is compiled in a se ries of reports titled Â“Memoirs of the
12 Hourglass CruisesÂ” (FWRI 2005). Project H ourglass was a 28-month program conducted between August 1965 and November 1967. The systematic sampling (fixed locations, gear, and interval) of the Hourglass cruises was designed to provide extensive biological information on organisms in offshore waters in the GOM. Stations were sampled on a monthly basis in an hourglass pattern west of Egmont Key and Sanibel Island in depths of approximately 7, 20, 40, 60, and 80 m. Dr edging, exploratory tr apping, fishing, nightlighting, plankton and nekton tows, water sa mpling, Secchi disk measurements, and redtide sampling were among the techniques used to sample the WFS during the Hourglass cruises. Lyons and Collard (1974) used the data fr om >700 dredge and trawl tows from the Hourglass cruises, supplemented with a number of additional collections and SCUBA observations, to tentatively delineate fauna l variation zones along the WFS. I say Â“tentativelyÂ” as there are no clear-cut faunal boundaries in th e eastern GOM, particularly in offshore deeper waters where temperature a nd salinity extremes factor less into benthic invertebrate distributional lim its as compared to nearshore (estuaries and bays) areas. Substrate is the single most important variable in the distribution of GOM invertebrates. Lyons and Collard (1974) suggest five z ones along the WFS based on the degree of faunal change (Figure 1.2). The shoreward zone (0-10 m) extends from the land-water interface to the mean depth where rocky outcr ops become important substrate elements. Salinities fluctuate in response to runoff from nearby rivers and bays, and nutrient concentrations are generally higher than t hose of the rest of the GOM. This zone includes areas such as the Ten Thousand Is lands where mangroves and seagrass beds transition into the offshore rocky outcrops. Both tropical and temperate species can be
13 found in this zone, although the latter tend to be more common. Astrangia poculata an ephemeral coral species that encrusts molluscan shells and shell fragments, occurs in the shoreward zone. Coastal Barrier Islands, the Big Bend, and the Cape San Blas areas are subdivisions of the shoreward zone. Figure 1.2. Faunal zone delineatio ns of the WFS (not to scal e) as proposed by Lyons and Collard (1974). Chapters 2-5 will concentrate areas along the inner shelf zone (10-30 m). The second faunal zone is th e shallow inner shelf zone (10-30 m depth) where rock substrate allows establishment of a number of tropical species including scleractinians, mollusks, and crustaceans that are common in the shallower waters of the
14 Florida Keys. Sediments along the shallow sh elf are composed primarily of quartz sands with percentages of biogenica lly derived carbonates increasing seaward. A number of coral genera, including Stephanocoenia, Siderastre a, Cladocora, Solenastrea, and Oculina, are often observed along the shallow inne r shelf. These shallow inner shelf communities are the focus of this paper and historic data from these depths/areas are chosen to represent the historic communities in comparative studies with modern surveys. The middle shelf I (30-60 m depth) is separated from the shallow shelf by the widespread presence of carbonate sediments and an overlying mass of offshore, blue water. Widespread outcrops, including t hose of the Florida Middle Grounds, support diverse communities of Loggerhead sponges, cora ls, and tropical algae. The middle shelf II (60-140 m depth) sediments are almost entire ly carbonate, composed of coralline algae, bryozoan and molluscan fragments, with pl anktic foraminiferal tests beginning to contribute to sediment composition. The se ssile epifauna is mainly composed of scattered poriferans, bryozoans, ascidians, and alcyonarians attached to small rocks and shells, along with Agaricia spp. assemblages, whose light requirements tend to limit them to approximately 80 m depth. This zone is frequently impacted by the Loop Current. The last zone is the deep shelf (140-200 m) which overlaps greatly with the middle shelf II. Species diversity tends to decrease with depth and Pequegnat (1970) noted a number of brachyurans most common in depths co rresponding to the d eep shelf zone. Beyond 200 m, the molluscan-dominated calcareous sands give way to planktonic foraminiferal sands and coccolith muds (calcareous oozes) as the continental slope drops steeply (2003200 m) to the floor of the GOM.
15 Like the benthic macroinvertebrate fa unal assemblages, the fish assemblages along the WFS are rich, including both Caribb ean (tropical) species and warm-temperate (subtropical) species, with the majority of species along the central WFS belonging to the latter group (Springer and Woodburn 1960; Brig gs 1973). The mobility of fish species renders faunal zone designation nearly impossible, although Darovec (1995) used similarities in inverse cluster analyses from two different types of sampling gear and two different years to demonstrate the possibility that depth-related fish community structures may exist along the WFS. His work s uggests that nearshore/estuarine and middle/offshore shelf fishes may have more restricted ranges whereas shallow inner shelf fishes are more widely dispersed among stati ons. Further analyses of abiotic parameters indicate that salinity and temperature ranges generally decreases w ith increasing depth. Darovec (1995) concludes that there is evidence to support the hypothesis that depth, through its effect on bottom temperature and sa linity, may be responsible for some of the fish distributions observed by the Hourgla ss study. This evidence does not, however, preclude support for other hypotheses. The longitudinal faunal zones of Lyon and Collard (1974) are cross-cut by latitudinal zoogeographic divisi ons. I propose that the inne r WFS can be divided into three latitudinal zones. The first, most sout herly zone, is a tropical stenopic zone that extends from the Straits of Florida-Dry Tortuga s to the Content Keys (24 45Â’ N). Coral genera including Acropora, Diploria, and Colpophyllia are iconic presences in these coral-reef communities. The second zone, the transition zone, extends from the Content Keys to Naples, FL (26 05Â’ N). Determini ng the exact extent of the transition zone is difficult, as it is characterized by the gra dual overlap of both tr opical and subtropical
16 flora and fauna; discharges from adjacent ri vers (Caloosahatchee River and San Carlos Bay estuary) often affect the development of epibenthic communities. The remaining northern extent of the WFS is a warm temper ate-subtropical zone that includes the rocky outcrops and hardbottom communities that are the focus of this dissertation. 1.5.2. The 2005 Red Tide During summer of 2005, a persistent red tide (harmful algal bloom) and subsequent hypoxic/anoxic conditions negati vely affected epibenthic hard-bottom communities in the GOM off west central Florida (Heil 2006). The event was the impetus behind the in-depth temporal and sp atial analyses of WFS shallow inner shelf assemblages presented in this paper. The resu lts serve as baseline indices (natural reefs) and potential avenues for mitigation (artificial reefs) in future events. Catastrophic events like the 2005 red tide have been documented since 1881, and observed for an even longer period of time In 1881, Ernest Ingersoll of the U. S. Fish Commission described the waters of the GOM during the 1878 red tide as Â“brownish, discoloredÂ…thick and glutinousÂ…they lay in streaks drifting with the ti de. Everywhere throughout this whole extent of coast [of Florida], except in the m ouths of rivers and in shallow bayous, all the forms of sea-life died as if st ricken with a plague fatal alik e to all, and we re drifted upon the beaches in long windrows so dense that near human habitations, men were obliged to unite in burying them to prevent pest ilential stenchÂ…Â” (Ingersoll 1881). The organism responsible for the Florida red tide was originally identified as Gymnodinium brevis (Davis 1948; Steidinger 1975) but was later changed to Karenia brevis (Daubjerg et al. 2000). Numerous questio ns still exist regarding the physical, chemical, and biological factor s that lie behind the red tide blooms and subsequent mass
17 mortalities of benthic animals and plants, fi shes, and marine mammals. One of the best documented red tides occurred in the mi d-eastern GOM during the summer of 1971. Qualitative observations before, during, and af ter the 1971 event provided insight into effects of a red tide bloom (Smith 1975). Af ter the red tide dissipated in September 1971, researchers assessed the impact on reef fish co mmunities. They estimated that 80-90% of resident reef fish species perished in the ev ent. On inshore reef s (13-18m), fewer than 26% of reef fishes survived the red tide (S mith 1975). Smith reported that invertebrate populations sustained even higher mortality than fish populations. Echinoderms, gastropod mollusks, decapod crustaceans, scleractinian corals, polychaetes, and poriferans all declined dras tically (based on qualitativ e observations). These observations indicate that red tides have the potential to greatly affect community structure and functioning thr oughout the affected areas. The GOM physical circulation patterns va ry annually and seasonally and there are numerous hypotheses for their contributions to Karenia brevis initiation, transport, and advection/dispersion. The inherent seasonal and annual variability of the Loop Current has been implicated in the in itiation of certain re d tides. The number and strength of Loop Current meanders varies annually and the eddi es and warm filaments associated with the meanders could have entrained within them nutrients and K. brevis cells that are then transported nearshore (Murphy et al. 1975; Te ster and Steidinger 1997) where increased nutrient concentration sustain the Karenis brevis growth. Another hypothesis is that nutrient-rich water from the Mississippi Rive r becomes entrained in the Loop Current as a result of seasonal variations in its northw ard penetration in the GOM (Huh et al. 1981; Gilbes et al. 1996; He and Weisberg 2003). The waters are transported along the WFS
18 where, again, nutrient limitations are lifted as a result of nearshore/coastal nutrient concentrations. A third hypothesis that may have specifi cally applied to the initiation of the 2005 K. brevis bloom is the Saharan dust/Iron fertilization/ Trichodesmium hypothesis (Lenes et al. 2001; Walsh et al. 2003; Wa lsh et al. 2006). Aeolian dust containing Fe is blown across the Atlantic from the Saharan desert during the summer months. Trichodesmium cyanobacteria are Fe-limited diazotrophs that inhabit the offshore ol igotrophic waters of the GOM; once their Fe levels are met, they fix atmospheric nitrogen into biologically available forms such as nitrite and nitrate, rendering them usable to dinoflagellates including K. brevis Walsh et al. (2006) discussed the mechanisms involved in the hypothesis and concluded that the amount of nitrogen fixed by the cyanobacteria was sufficient to have sustained the 2005 red tide. The model car ried with it a number of stipulations and further inves tigation is needed to confirm the experimental conditions. Karenia brevis has physiological adaptations that allows it to out-compete other phytoplankton once it emerges from its initiati on depth (if that mechanism is indeed responsible for bloom initiation). Karenia brevis is positively phototactic so it congregates at the surface and subsurface wa ters during the day but disperses downward towards higher nutrient levels at night. Karenia brevis is low-light adapted and can utilize blue and green light for photosynthesis, which gives it a competitive advantage at low light levels. During times of increased irradiance in surface waters, K. brevis is equipped with Â“sunscreensÂ” in the form of xanthophylls that may help protect it from increased UV light, which is ha rmful to other phytoplankton. Karenia brevis is auxotrophic and uses both dissolved inorga nic nitrogen (DIN) and dissolved organic
19 nitrogen (DON) sources such as urea and ur ic acid that are excr eted by zooplankton and fish. Diatoms and K. brevis both have low Ks values, but diatoms have much higher growth rates, therefore they outcompete K. brevis at times of high inorganic nitrogen loading, but K.brevis Â’s ability to utilize organic nitr ogen gives it a competitive advantage at times when ratios of DIN:DON are low. A number of physical and climatologi cal conditions were implicated in the development of hypoxic/anoxic waters at depth during the 2005 red-tide event, exacerbating biotic mortalities and evacuations (Heil 2006; personal observation). The summer of 2005 was extremely warm and the thermocline was very shallow in the GOM, with significant water column stratification. There were a number of major hurricanes that swept through the area, te mporarily mixing the stratifie d water column that could have brought K. brevis cells in the bottom waters to th e surface, where the supply of both DON and DIN (from the Trichodesmium bloom) and inherent physiological advantages of K. brevis over diatoms enabled the developm ent of a large-scale bloom. Karenia brevis cells that remained caught beneath the strongly established thermocline contributed two-fold to the increase in organic matter as: (1) the phytoplankton cells themselves, upon death, fell to the bottom and the organic matter was oxidized by bacteria, and (2) the organisms affected by the brevetoxin died and contributed to the organic matter and subsequent oxygen depletion. The events resulted in the creation of a zone of benthic anoxia that extended from Pinellas County south to Sarasota (FWRI 2005). Throughout the water colu mn, the rain of organic matter from the phytoplankton and dead fish was oxidized, with a hydrogen sulfide layer forming at the top of the thermocline at approximately 6 m (personal observation). Stratif ication of the water
20 column persisted throughout the summer and in to the winter months; associated anoxic conditions extirpated much of the living, sessi le biota and many slower moving fish and invertebrates. Mobile species evacuated sha llow inner shelf areas in favor of deeper waters, where both K. brevis and bottom-water an oxia were absent. The red-tide event of 2005 provided an opportu nity for a quantita tive study of the responses of the benthic invertebrate community and demersal fish assemblage to a major red tide disturbance. At the same time, I began to amass a comprehensive database on shallow inner WFS species diversity and rela tive abundances using bot h historic studies and modern surveys. The marg inality of modern-day reef assemblages, combined with work on artificial reefs along the WFS, in spired thoughts and analyses on the use of artificial reefs as recruitment enhancemen t tools along the WFS, in areas frequently impacted by disturbance such as red tides a nd hurricanes. These da ta are presented and discussed in subsequent chap ters of this dissertation. 1.6. Overview of Dissertation The main body of this document is compos ed of three chapters that are either published, in review, or pending submission for publication. Each chapter is treated as an independent scientific cont ribution containing its respec tive figures and tables. All references for the entire document are grouped at the end of the dissertation. A conclusion chapter summarizes and compares the information (natural ledges and artificial reef assemblages) and outlines the re levance and usefulness of the data in future conservation and management projects along the WFS. Chapter summaries are as follows:
21 Chapter 2 examines the components and dynamics of marginal reef/livebottom assemblages along the WFS Chapter 3 presents a refer eed paper published in the Coastal Management Journal (Dupont 2008) that discusses th e effectiveness of a set of artificial reefs deployed along the WFS. Chapter 4 discusses the results of a fo cused (2005 to 2007) monitoring study of recruitment and succession on artificial reef structures before and after the redtide disturbance. Chapter 5 compares aspects of the natura l ledge and artificial reef communities and discusses the use of low-relief arti ficial reefs as recruitment enhancement tools to increase resiliency of livebottom assemblages along the marginal WFS.
22 2. Central West Florida Shelf Natural Ledge Dynamics 2.1 Abstract The West Florida Shelf (WFS) is one of the largest, most productive, and heavilyused continental shelf/slope systems in the wo rld. The WFS is home to some of the most valuable commercial and recreational finfishes in the southeastern United States. Shallow inner WFS livebottom assemblages (10-30 m depth) support a number of finfish life stages by providing structure and protection from predator s, benthic primary production, and a variety of food sources (associated crustaceans, mollu sks, gastropods, and smaller fish). Livebottom assemblages (including sc leractinian corals, macroalgae, poriferans, and echinoderms) along the WFS occur in transitional environmental conditions between subtropical/tropical Caribbean and temperate Carolinian zoogeographic provinces. Temperature, nutrient, and light regimes are highly variable and the livebottom and fish assemblages are further stressed by periodic acute disturbances including harmful algal blooms (red tides) and hurricanes. This paper assesses the spatial and temporal (seasonal) trends of epibenthic macroinvertebrates, juve nile corals, macroalgae, and demersal fish species over a two-year time period followi ng a red-tide disturbance at two livebottom reefs along the shallow inner WFS. Data from modern surveys are combined with historic data from similar depths to genera te comprehensive species lists. The goal is to provide baseline data on the essential communities that can be used to assess future
23 disturbance impacts and recovery rates, partic ularly in the face of global environmental change. 2.2 Introduction The West Florida Shelf (WFS) is one of the largest and most productive continental shelf/slope systems in the world. Due to the importance of continental shelf resources to the State of Flor ida, including the prolific finfish and shellfish fisheries, offshore petroleum and natural gas explorati on, and tourism industries, the WFS has been the subject of numerous studies that addr ess the unique physical oceanographic regimes, chemical influences (from the Mississippi Ri ver and the numerous ri vers/estuaries that drain into the GOM), and the dynamic geologic features (see Chapte r 1 and references therein). However, a more detailed search into the biological at tributes of the WFS, specifically community ecology studies, turns up far fewer papers. In general, the existing biological papers fit into two categorie s. The first includes papers that address specific taxa or populations including, am ong others, Echinodermata (Hill and Lawrence 2003; Cobb and Lawrence 2005), zooplankton (H untley and Boyd 1984; Kleppel et al. 1996), phytoplankton, including harmful algae bl ooms (HABs--Vargo et al. 1987; Tester and Steidinger 1997; Lenes et al. 2001; Walsh et al. 2006), viral and bacterial assemblages (Hewson et al. 2006), and characte ristics and life cycles of individual fish species such as Epinephelus morio (Richardson and Gold 1997), Sardinella aurita (Kinsey et al. 1994), and Mycteroperca microlepis (Fitzhugh et al. 2001). The second category of biological articles focuses primarily on either deep-water or mid-shelf reef communities including Pulley Ridg e (Jarrett et al. 2005; Hine et al. 2008) and the Florida
24 Middle Grounds (Cheney and Dyer 1974; Hine et al. 2008), along with a few nearshore (<10 m depth) seagrass community studies (Dawes and Tamasko 1988; Zieman et al. 1989). While the above studies, combined with the extensive research on physical, chemical, and geological features, have cont ributed greatly to our understanding of WFS dynamics, there is a dearth of data desc ribing shallow inner sh elf (10-30 m depth) epibenthic macroinvertebrate and fish commun ities on the WFS. The broad nature of the gently-sloping WFS allowed for the extensive, lateral movement of the shoreline during sea-level cycles, and led to the development of diverse distributions of paleoshorelines and shallow-water hardbottoms (Hine et al. 2008 ). Geologic works (Locker et al. 2003; Obrochta et al. 2003; Hine et al. 2008) have shown that al ong the west-central Florida coast, at least 50% of the inner shelf seaward of 5 km consists of hardbottom, or lithified seafloor. Hardbottoms are common in shallow carbonate and siliclastic marine settings, but are generally poorly desc ribed and documented (Obrochta et al. 2003). Although tropical reef development is absent along the inner WFS (Jaap 1984), most likely limited by excess nutrients and the associated high bi oerosion rates (Hallock and Schlager 1986; Hallock 1988), the extensive systems of scarpe d hardbottom provide re lief (up to 4m) and important habitat that support an association of hardy corals and other biota. Much of the WFS may be considered ecotonal between the temperate Carolinian and tropical Caribbean (or West Indian) zoogeographic be nthic invertebrate provinces (Hedgpeth 1957; Lyons and Collard 1974), meaning that ha rdy constituents of both occur across the WFS.
25 While the body of literature on physical a nd geological characteristics of the WFS has grown in recent years and numerous hardbottom areas have been mapped (Locker et al. 2003; Obrochta et al. 2003; Hine et al. 2008), there is little information on the abundance and diversity of the epibenthos (l ivebottom) and demersal fish assemblages that inhabit the WFS ledges on the inner sh allow shelf, and even less information on temporal (seasonal and interannual) changes in community structure. Jaap (1984) described the inner WFS habitats as critical ha bitats that should be provided with rational management due to their association with ex tensive fisheries of the eastern Gulf of Mexico including numerous important grouper and snapper species. As demands on the WFS resources increase, includ ing proposed offshore oil and natural gas exploration and production, detailed information on the life hi stories and ecology of marine organisms that inhabit these areas is essent ial to resource management. This paper first will summarize availabl e information from historic work on epibenthic communities and demersal fish assemblages along the inner WFS. Then it will present seasonal data from monthly sampling of two hardbottom areas along the WFS over a two-year period. The larger goa l is to provide baseline information on ecological attributes of the WFS, which can be used in future works that assess impacts from disturbances that are common in the GOM, including hurricane s, tropical storms, winter cold events, HABs, and hypoxic/anoxic events, as well as potential offshore petroleum exploration projects.
26 2.3 Historic Data One of the primary sources of detail ed, exploratory information for WFS biological diversity is compiled in a series of reports titled Â“Memoirs of the Hourglass CruisesÂ” (FWRI 2005). Projec t Hourglass was a 28-month program conducted between August 1965 and November 1967. The systematic sampling (fixed locations, gear, and interval) of the Hourglass cruises was designed to pr ovide extensive biological information on organisms in offshore waters in the GOM. Stations were sampled on a monthly basis in an hourglass pattern west of Egmont Key and Sanibel Island in depths of approximately 7, 20, 40, 60, and 80 m. Dr edging, exploratory tr apping, fishing, nightlighting, plankton and nekton tows, water sa mpling, Secchi disk measurements, and redtide sampling were among the techniques used to sample the WFS during the Hourglass cruises. The full suite of Hourglass data are available online; in this paper I will focus on Station B epibenthic and fish communities as the depth and location are most similar to my surveys of natural ledges (discussed in th is chapter) and artifi cial reefs (Chapters 3 and 4) along the WFS. The coordinates for th e my natural ledge sites, Station B, and the artificial reefs are shown in Table 2.1 and a map of their locations with respect to one another is presented in Figure 2.1. Table 2.1. Coordinates of sites sample d 1965-1967 (Station B) 2005-2007 (Artificial Reefs), and 2006-2007 (FWRI1 and MT) Site Latitude Longitude FWRI1 27 54' 47.16"N 83 06' 19.80"W Mastedon Tabletop (MT) 27 54' 48.95"N 83 06' 21.24"W Clearwater Wreck (CW) 27 54' 06.48"N 83 06' 29.16"W Station B (Hourglass Program) 27 37'N 83 07'W Artificial Reefs 27 34'N 83 05'W
27 A number of other reports and manuscripts have utilized the information gathered during Project Hourglass, and ha ve contributed subse quent information to descriptions of WFS biology. Interestingly, the majority of these works were produced in response to proposed outer continental shelf oil explora tion and production activ ities in the eastern GOM during the 1970Â’s and 1980Â’s Local stakeholders, including the scientific community, and Federal agencies such as th e Minerals Management Service initiated studies of eastern GOM ecosystems, recogni zing that there was a scarcity of basic environmental information for the area and that the increased demand for domestic energy sources, combined with the distinct po ssibility that oil might exist beneath the WFS, could open certain tracts for lease (a possibility that is again under consideration today). Basic works on the biological/faunal zones of the WFS (Lyons and Collard 1974; Lyons and Camp 1982) have designated five faunal zones: the shoreward zone (0-10 m depth), the shallow inner shelf (10-30 m), th e middle shelf I (30-60 m), the middle shelf II (60-140 m), and the deep shelf (140-200 m).
28 Figure 2.1. Map (Google Earth) of the study site s located along the shallow inner west Florida shelf. This paper will focus on sites within the shallow inner shelf (10-30 m) faunal zone (Fig. 2.1) where the presence of rock substrate supports a numb er of tropical biota including scleractinians, codiaceans, mollusk s, and crustaceans that are common in the shallower waters of the Florida Keys. Se diments along the shallow shelf are composed primarily of quartz sands with percentages of biogenically derived carbonates increasing seaward. The benthic communities are diverse and generally concentrated on the shoreward-facing (lee) side of the scar ped hardbottom (Obrochta et al. 2003). Halimeda spp. meadows cover the upper flat hardbottoms pr oximal to the scarp (Fig 2.2) while red calcareous algae, boring mollusks ( Lithophaga spp.), boring poriferans ( Cliona spp ) and echinoderms occupy both upper flat and scar ped surfaces (Obrochta et al. 2003). A
29 number of other benthic flor a and fauna have also been documented on the shallow inner WFS (Dawes and Lawrence 1990). Figure 2.2. Examples of hardbottom ledge co mmunities along the shallow inner west Florida shelf. Though numerous works have been published on WFS faunal zones, the majority have used data from one source: The Hourgl ass Cruises. The H ourglass Cruises were comprehensive in both spatial and temporal sampling scales and the benthic collection methods (otter trawl and box dredge), provi ding valuable perspective on faunal zones along the WFS. However, the data collected during the Hourglass Cruises are now over 40 years old, and a methodological review reveals limitations th at influence data interpretation. The otter trawl and box dre dge data are binary (presence/absence) for most taxa, and are insensitive to relative abund ance patterns. Equal weighting of rare and common species also contribute to biases in station/site descriptions. The continued characterization of important resources al ong the WFS requires robust sampling methods and modern data sets. My goal in this chapte r is to present quantitative approaches to characterizing areas along the shallow inner central WFS. The methods can then be
30 expanded, in conjunction with technologic adva nces such as submersibles and ROVs, to shelf-wide community surveys. 2.4 Background and Ration ale for Current Study This chapter focuses on sites located in the inner shallow central WFS (10-30 m depth) faunal zone of Lyons and Collard (1974) and Lyons and Camp (1982). Monthly surveys of epibenthic community and fish assemblage data were conducted from February 2006 through December 2007, represen ting almost two years of data. The study was initiated in response to a massive Karenia brevis bloom (red tide) that persisted in the area during the majority of 2005 (Heil 2006 ). Reports of mass benthic mortalities, along with in situ dissolved oxygen measurements and K.brevis cell counts (FWRI 2005), confirmed that the development of an intense thermocline, combined with the rain of decomposing organic matter from the alga l bloom, led to the development of hypoxic/anoxic conditions and mass mortalitie s in patches of bottom waters along the WFS. Reports indicated that deeper, offshore areas (>30 m) were relatively unaffected by the K. brevis and anoxia, while shallower area s displayed mass die-offs of scleractinian corals, poriferans, echinoderms, mollusks, and crustaceans, as well as a number of fish species. Previous reports (S mith 1975; 1979) have qualitatively assessed the impacts and recovery rates of both epibenthic macroinvertebrates and fish assemblages in response to red tide events. These studies, while informative, lack an attention to quantitative de tails on shallow shelf comm unity composition (species diversity and abundances) as well as temporal (seasonal) fluctuations particularly in the case of epibenthic macroinvertebrates.
31 My study began in February 2006 as a comb ined effort between University of South Florida (USF), Florida Aquarium (FL AQ), and FWRI scientists and divers to document the benthic mortalit ies associated with the 2005 K. brevis bloom and hypoxic/anoxic bottom-water conditions. Although the K. brevis bloom dissipated in late 2005, sampling was not begun until February 2006 due to logist ical and weather-related issues. Despite the lack of quantitative information befo re, during, and immediately after the K. brevis bloom and associated hypoxia/anoxia, a two-year data study was initiated to assess post-disturbance conditions and to track the recovery of epibenthic macroinvertebrate and fish species. Quantitative assessments of red-tide effects are presented in Chapter 4, from a set of artif icial reefs deployed along the shallow inner shelf. 2.5 Methods 2.5.1. Site Characteristics Sampling effort was focused at two site s, FWRI1 and Mastedon Tabletop (MT) (Table 2.1). FWRI1 and MT were chosen for a vari ety of reasons. First, they were located approximately 40 km west of Clearwa ter, Florida, at 1820 m depth, and were situated along the shallow inner WFS (10-30 m depth). Second, while they were located north of the Hourglass Program Station B, they were situated at approximately the same longitude (and depth), thereby allowing for qualitative compar isons between historic and modern surveys. Third, both FWRI1 and MT had been sampled during an FWRI red-tide sampling cruise from August 10-12, 2005. Water samples and in situ diver-collected data confirmed the presence of medium to high concentrations of K. brevis cells in surface waters (>100,000 cells L-1), hypoxic (<2 mg L-1) bottom waters, and benthic mortalities,
32 indicating that both s ites were affected by the red-tid e bloom. Fourth, both sites had typical WFS hardbottom features, including up to 2 m of scarped hardbottom, undercut by bioerosional forces (Obr ochta et al. 2003), and provi ded suitable substrate for epibenthic macroinvertebrate and fish associat ions. Fifth, the sites were located in close proximity to one another and were easily accessed by boat from Clearwater, FL, providing access to sample the si tes on a monthly basis. 2.5.2. Benthic Community Data During each sampling, one of the two site s (MT or FWRI1) was chosen as the target site. The captain of the boat naviga ted precisely to the c oordinates and a buoy was dropped marking the site. The anchor was th en deployed in close proximity to the buoy and divers descended down the anchor line to the site. This precise navigation was necessary through the first year (2006). During the second year of sampling (2007), a temperature logger was deployed at each of the three sites. The logge r was attached to a cinderblock along with a sub-surface buoy situated five meters from the bottom, allowing for exact location of the study site. Upon reaching the bottom, diver teams began to survey the fish assemblage utilizing the Bohnsack and Bannerot (1986) method described below while I conducted photographic benthic transect surveys. Three 15 m transect lines that trended ledge-parallel (northwest to southeast) were surveyed at random distances from one anot her (random number of fin-kicks chosen a priori). Photographs were captured at each 0.5 m mark along the transect line using a Canon Powershot A550 with the camera set to the underwater scene for best contrast. The camera was kept at a fixed distance of 50 cm from the bottom, providing a total of thirty 48 cm x 38 cm photographs per transe ct. The distance ensured that there was
33 maximum coverage of the transect with no overlap between photos, and enabled the identification of many organisms to genus a nd species levels duri ng the post-processing analysis of images. Substrate and biological cove r attributes of the bent hic photographic transects were assessed using point-count analysis (e.g., Curtis 1968; Bohnsack 1979; Carlton & Done 1995; Jaap and McField 2001; Jaap et al. 2003). Twenty randomly generated points were superimposed on each image in Coral Point Count v.3.4 (Kohler & Gill 2006), and the benthic component under each point was identified to provide an estimate of benthic cover (Hackett 2002). Seven major biological and substrate categories (Coral, Porifera, Macroalgae, Dead Coral with Algae, Bleached Coral, Bare Substrate, and Other Living Fauna) were included in the assessmen t, with subcategorie s (including specific coral and algal species) also being identif ied when possible. One advantage of the program, Coral Point Count v.3.4, is that sub categories are linked to a major category, thereby providing researchers with the ability to describe organisms to species level when possible, without sacrificing th e description in the major cat egory. This capability is especially important when analyzing phot os in the eastern GOM, where seasonal visibility can vary, often precl uding accurate identification of organisms to species level. The data are reported as percent cover (% co ver) values, and are averaged over the three transects (n=3) for each sampling. This was the maximum number of transects possible, while remaining within dive limits. Adult coral species were identified in all phot ographs for comparison with 19651967 coral data, in terms of species richne ss (simple presence/abs ence enumeration). Juvenile corals were also easily seen in th e photographs and include d in the analyses, as
34 very little is known about the spawning and recruitment patterns of the hardy coral species that inhabit the hardbottom outcrops on the WFS. A criterion of minimum size was used to distinguish sexually-produced juve nile corals from adult coral and isolated fragments (Miller et al. 2000; Irizarry-Soto 2006). Because the majority of the adult corals on the WFS ledges were small (<20 cm), including species such as Siderastrea radians and Stephanocoenia intersepta isolated colonies less than 2 cm in diameter were considered sexually-produced juveniles (Irizarry-Soto 2006). Although numerous juveniles were observed in photos the inherent properties of ne w recruits (i.e., small size, propensity for burial under sediment) make true quantitative reports through photo documentation very difficult. Instead of coun ting individual recru its, as is often done in situ using quadrats, the simple presence/absence of juveniles in photos was recorded as a percentage of photos containing juveniles pe r 15 m transect (i.e., the number of photos with at least one juvenile/30 pictures). This allowed me to observe general patterns of recruitment on a 2-year time scale, with pa rticular focus on the seasonal influences on coral spawning and recruitment in the eastern GOM. The monthly variations of the seven majo r categories and juve nile photo-transect percentages were plotted as boxplots displaying interquartile ranges, medians, means, and outliers. These data reveal insight into seas onal trends and sample distributions of major component categories over the two-year peri od. The major benthic component data were right-skewed and subsequently transformed using a Log(x+ 1) transformation. One-way Analysis of Variance (ANOVA) was used to te st for significant s easonal differences in the major categories followed by the TukeyÂ’s post-hoc multiple comparison procedure to determine pair-wise differences. A Principa l Components Analysis (PCA) was used to
35 reduce the dimensionality of the data, dete rmine important gradients, and spatially display the data. 2.5.3. Fish Assemblage Data During each sampling trip, 2-4 research divers conducted fish surveys in accordance with the Bohnsack-Bannerot fish count protocol (Bohnsack and Bannerot 1986). The Bohnsack method provides sta ndard quantitative data on reef-fish assemblage structure over a variety of habita ts in an efficient and effective manner. Observers position themselves on the center poi nt of the census area, and wait for three minutes prior to recording. The waiting period allows for the dampening of any disturbance and fishes can acclimate to dive r presence. Divers attempt to count all individuals and species of fish in an imaginary 5 m radius cylinder extending from the bottom to the surface. New species are listed while rotating in one direction and scanning the field of view. The observer remains stat ionary except for rotation. Five minutes was chosen as an optimum counting time because it allows for most fish to habituate to the diver, but minimizes the time for mobile speci es outside the cylinder to accumulate. The observers were usually able to conduct betw een 3 and 5 surveys per dive, yielding from 6-20 fish surveys per dive. A number of these surveys did, however, have to be eliminated from use in the study as the observe rs were either practic ing fish-identification skills or did not pass the GOM fish identificati on test administered before the dives. Due to the mobile nature of the fish and the close proximity of the two sites, the data were analyzed as a group, with no differentiation between FWRI1 and MT data. The grouping of data yielded a database that was more chr onologically consistent than if the two sites had been analyzed separately. Species numbers and assemblage composition from the
36 2006-2007 data were compared to lists from comparable depths (Station B) of the Hourglass Cruises. Data were al so compared to artifi cial reef sites situated in comparable depths/locations in the GOM (Chapter 3). Fish species abundance data from the 2006-2007 surveys (pooled) were entered into a matrix worksheet and an Anderson-Da rling test was used to test for normality within samplings. The Anderson-Darling test p-values i ndicated that, at >0.02, there is evidence that most samples did not follo w a normal distribution. Therefore, nonparametric multivariate analyses were conducted using the Primer-ETM (Clarke &Warwick 2001) package of software appl ications to analyze assemblage-wide changes/differences among samplings. Abundan ce data were square-r oot transformed to focus attention on patterns within the whole assemblage, mixing contributions from both common and rare species (Clarke & Warwick 1994). Multivariate distances were calculated using the Bray-Curtis similarity co efficient (Bray & Cur tis 1957) and plotted using a non-metric multi-dimensional scali ng (MDS) ordination. The MDS finds a nonparametric monotonic relationship between dissim ilarities in the item-item matrix and the Euclidean distance between the items, and plots the location of each item in lowdimensional space. MDS ordination stress le vels <0.15 signify a useful representation (i.e., configuration closely repr esents the rank order of dissimilarities in the original triangular matrix), while stress levels > 0.20 signify a random arrangement of samples, bearing little resemblance to the original ranks (Clarke 1993). Fact ors were added to the original data to view and determine optim al spatial arrangements among groups. Factors included site, year (2006 or 2007), season, and sampling. Second-level procedures (Clarke & Warwick 2001), including Analysis of Similarity (ANOSIM) and Similarity
37 Percentages (SIMPER) tests, were used to test for significant differences in fish assemblage structure between those samples/groups that appeared to separate spatially in the MDS. 2.5.4. Abiotic Data Abiotic parameters, including salinity and Secchi disk depths, were measured sporadically throughout the st udy and average values are re ported in this paper. Consistent bottom temperature data were co llected beginning in mid-February 2007. Temperature loggers were depl oyed throughout the year from February to November at three sites, FWRI1, MT, and CW. The loggers were affixed to a cinder block with a subsurface buoy and were deployed for periods of 1 to 3 months, at which time they were swapped out and taken back to the lab for data download. Th e loggers were set to record temperature data at either 5 or 10 minute interv als. The data are presented in this paper, and represent one of the first high-resolution benthic temperature databases for the central shallow inner WFS. Secchi depths were converted into light attenuation coefficients (k-values) and percent surface light reaching bottom at 17 m depth was calculated using the BeerLambert Law: Iz/I0 = e-kz. 2.6 Results 2.6.1. Historic Data Historic data from the Hourglass Program Station B are summarized in this paper in the form of species lis ts (Tables 2.2 through 2.5) and used as qualitative comparative baselines to my surveys, which are also includ ed in the tables. Coral species are listed in Table 2.2, echinoderms in Table 2.3, benthic algae in Table 2.4, representing those
38 epibenthic flora and fauna that were readily identified in digital phot o-transect surveys. Although a number of other H ourglass reports were genera ted on various taxa (not presented here), this paper will focus on benthic species lists that either: (1) fit into one of the seven benthic major categories (e.g., corals and macroalgae) or (2) are epibenthic macroinvertebrates that may have fallen into the Â“Other Living Fa unaÂ” category, but are easily distinguished in photographs (e.g., Echi nodermata). Table 2.5 lists fish species from all studies.
39 Table 2.2. Coral species recorded at shallow hardbottom sites on the inner west Florida shelf. Data are compiled from three data sets and species presence is denoted by the corresponding number: (1) 19651967 (Hourglass Program), (2 ) 2006-2007 (Natural Reef Ledges) and (3) 2005-2007 (Ar tificial Reefs). Reproducti ve modes are also listed. N/K=Not known. Species Reproduction Data Set Cladocora arbuscula N/K 1,2,3 (LeSueur, 1821) Phyllangia americana N/K 1,2,3 Milne-Edwards & Haime, 1849 Solenastrea hyades Broadcast 1,2 (Dana, 1846) Manicina areolata Brooding 1 (Linnaeus, 1758) Siderastrea radians Brooding 1,2,3 (Pallas, 1766) Oculina robusta Broadcast 1,2 (Pourtals, 1871) Stephanocoenia intersepta Broadcast 1,2 (Esper, 1795) Scolymia lacera Brooding 1 (Pallas, 1766) Balanophyllia floridana N/K 1 De Pourtals, 1868 Porites divaricata Brooding 1 LeSueur, 1821 Millepora alcicornis Budding 1 Linnaeus, 1758 Astrangia poculata N/K 1 (Ellis & Solander, 1786) Isophyllia sinuosa Brooding 1 (Ellis & Solander, 1786)_______________________________ *Occurred at 10% frequency in the Hourglass Program
40 Table 2.3. Echinoderm species recorded on shallow hardbottom sites on the inner WFS. Data are compiled from three data sets and species presence is denoted by the corresponding number: (1) 19651967 (Hourglass Program), (2 ) 2006-2007 (Natural Reef Ledges) and (3) 2005-2007 (Artificial Reefs). Echinodermata Data Set Arbacia punctulata 1,2,3 (Lamarck, 1816) Lytechinus variegatus 1,2 (Lamarck, 1816) Clypeaster prostratus 1* Ravenel, 1848 Clypeaster subdepressus 1 (Gray, 1825) Mellita quinquiesperforata 1 (Leske, 1778) Encope aberrans 1* Martens, 1867 Encope michelini 1 Agassiz, 1841 Plagiobrissus grandis 1* (Gmelin, 1788) __________________________ Occurred at 10% frequency in the Hourglass Program
41 Table 2.4. Benthic algae species recorded on shallow hardbottom sites on the inner WFS. Data are compiled from three data sets and species presence is denoted by the corresponding number: (1) 19651967 (Hourglass Program), (2 ) 2006-2007 (Natural Reef Ledges) and (3) 2005-2007 (Artificial Reefs). Benthic Algae Data Set Cyanophyta Calothrix confervicola 1* C. Agardh 1824 Lyngbya bergei 1* Kellerman 1893 L. confervoides 1 Umezaki (1961) Chlorophyta Avrainvillea levis 1* Howe 1905 A. longicaulis 1 G.Murray & Boodle Anadyomene stellata 1* C. Agardh 1822 Caulerpa mexicana 1,2,3 Kntzig ex Sonder 1849 C. paspaloides 1 Weber-van Bosse 1898 C. peltata 1* (Weber-van Bosse) Reinke 1900 C. prolifera 1*,2,3 Lamouroux 1809 C. sertularioides 1,2,3 Howe 1905 Codium isthmocladum 1,2,3 Vicker 1905 C.i. subsp. clavatum 1 Vicker 1905 C. repens 1 P. and H. Crouan Ex Vickers Cystodictyon pavonium 1* Lambert Halimeda discoidea 1,2,3 Decaisne H. opuntia 1* (Linnaeus) Pseudotetraspora antillarum 1* Howe 1905
42 Udotea conglutinata 1,2,3 J. V. F. Lamouroux 1812 U. cyanthiformis 1* J. V. F. Lamouroux 1812 U. flabellum 1,2,3 Howe Valonia macrophysa 1* C. Aghardh 1823 Phaeophyta Cladosiphon occidentalis 1* Kylin Colpomenia sinuosa 1 Derbes and Solier Dictyopteris delicatula 1* J.V. Lamouroux 1809 D. membranacea 1* Batters Dictyota dichotoma 1,2,3 Nizamuddin 1981 D. divaricata 1* J.V. Lamouroux 1809 Ectocarpus elachistaeformis 1* Heydrich (1892) Giffordia sp. 1* G. Hamel Rosenvingea intricata 1 Brgesen R. sanctae-crucis 1* Brgesen Sargassum filipendula 1,2,3 Grunow 1916 S.f.v.montagnei 1,2,3 Steidinger & Van Breedveld 1969 S. natans 1,2,3 Gaillon 1828 Sporochnus bolleanus C.Agardh 1824 1 S. pedunculatus 1 Lucas 1936 Rhodophyta Acrochaetum antillarum 1* Farlow, W. G. 1876
43 A. flexuosum 1* Naegeli 1858 A. thurettii 1* Born Agardhiella ramosissima 1 Schmitz (1896) A. tenera 1 Schmitz (1896) Amphiroa rigida v. antillana 1* Lamouroux 1816 Asterocytis ramosa 1* Tanaka 1944 Botryocladia occidentalis 1 Kylin Brongniartella mueronata 1* H Woods 1897 Callithamnion halliae 1* Collins, Holden & Setchell 1900 Ceramium fastigiatum 1 Celan & Serbanescu 1959 C. leptozonum 1* Howe 1918 C. rubrum 1* Hudson Champia parvula 1,2,3 Harvey Chondria floridana 1 M.A. Howe C. tenuissima 1* C. Agardh 1817 Chrysymenia enteromorpha 1 Harvey (1853) C. ventricosa 1* J. Agardh (1842) Crouania attenuata 1* J. Agardh, 1842 Dasya collinsiana 1* M. Howe D. corymbifera 1 J. Agardh 1841 D. pedicellata 1 C. Agardh 1824 D. rigidula 1* Ardissone 1878 Digenia simplex 1*
44 Wulfen 1803 Erythrocladia sp 1* Rosenvinge 1909 Eucheuma acanthocladum 1,2,3 J. Agardh 1847 E. isiforme 1,2,3 J. Agardh 1847 Fosliella atlantica 1 Harvey 1836 Gracilaria armata 1* Greville 1830 G. blodgettii 1 Harvey 1853 G. cervicornis 1* J. Agardh 1852 G. cylindrica 1 Brgesen 1920 G. debilis 1 Borgesen G. ferox 1 J. Agardh 1852 G. foliifera 1 Brgesen 1932 G.f.v. angustissima 1 Taylor G. mammillaris 1 M.A. Howe 1918 G. sjoestedtii 1* Kylin G. verrucosa 1 Papenfuss 1950 Halymenia agardhii 1* C. A. Agardh 1817 H. bermudensis 1* Collins and Harvey H. floresia 1 C. A. Agardh H. gelinaria 1 Collins & Howe 1916 H. pseudofloresia 1 Collins and Howe Jania adherens 1 Lamouroux 1816 J. capillacea 1* Harvey 1853
45 Laurencia gemmifera 1 Harvey 1853 L. intricata 1* Lamouroux L. obtusa 1* Lamouroux L. poitei 1 Howe Lithothamnium incertum 1 Dakwix 1854 L. occidentale 1 Lemoine 1917 Lomentaria baileyana 1 Farlow Peyssonnelia rubra 1* J. Decaisne 1841 Polysiphonia hapalacantha 1* Harvey 1853 P. subtilissima 1* Mont 1840 Spyridia filamentosa 1 Harvey Wrightiella blodgettii 1* Schmitz Wurdemannia miniata 1 Feldmann & Hamel 1952 Angiospermae Halophila baillonis 1 Aschers 1874 Thalassium testudinum 1 Keough 1986 ______________________________________ *Occurred at 10% frequency in the Hourglass Program
46 Table 2.5. Fish species recorded on shallow hard bottom sites on the inner WFS. Data are compiled from three data sets and speci es presence is denoted by the corresponding number: (1) 1965-1967 (Hourglass Program), (2) 2006-2007 (Natural Reef Ledges) and (3) 2005-2007 (Artific ial Reefs). Species Common Name Data Set Abudefduf saxatilis Sergeant major 3 Acanthostracion quadricornis Cowfish/boxfish 1,3 Acanthurus chirurgus Doctorfish tang 2 Anisotremus surinamensisi Black margate 2 Anisotremus virginicus Porkfish 3 Antennarius ocellatus Ocellated frogfish 1 Apogon quadrisquamatus Sawcheek cardinalfish 1 Archosargus probatocephalus Sheepshead 2,3 Arius felis Hardhead catfish 1 Balistes capriscus Grey triggerfish 1,2,3 Bothus robinsi Twospot flounder 1 Calamus bajonado Jolthead 3 Calamus penna Sheepshead porgy 2,3 Calamus spp. Porgy 1,2,3 Caranx crysos Blue Runner 2,3 Caranx ruber Barjack 2,3 Centropristis ocyurus Bank seabass 1 Centropristis striata Black seabass 1,2,3 Chaetodipterus faber Atlantic spadefish 1,3 Chaetodon ocellatus Spotfin butterflyfish 2 Chaetodon sedentarius Reef butterflyfish 3 Chaetodon striatus Banded butterflyfish 2 Chasmodes saburrae Florida blenny 3 Chilomycterus schoepfi Striped burrfish 1 Citharichthys macrops Spotted whiff 1 Cosmocampus hildebrandi Dwarf pipefish 1 Cyclopsetta fimbriata Spotfin flounder 1 Decapterus punctatus Round scad 1,2,3 Diodon hystrix Spot-fin porcupinefish 3 Diplodus holbrookii Spottail pinfish 2,3 Diplectrum formosum Sand perch 1,2,3 Diplogrammus pauciradiatus Spotted dragonet 1 Epinephelus itajara Goliath grouper 3 Epinephelus morio Red grouper 1,2,3 Equetus lanceolatus Jackknife fish 1,3 Equetus punctatus Spotted drum 2,3 Etropus crossotus Fringed flounder 1 Etropus rimosus Gray flounder 1 Ginglymostoma cirratum Nurse shark 3
47 Gobiosoma horsti Yellowline goby 1 Gobiosoma macrodon Tiger goby 1 Gymnothorax nigromarginatus Blackedge moray 1 Gymnothorax saxicola Ocellated moray 1 Haemulon aurolineatum Tomtate 1,2,3 Haemulon flavolineatum French Grunt 3 Haemulon plumieri White grunt 1,2,3 Halichoeres bivittatus Slippery dick 2,3 Halichoeres maculipinna Clown wrasse 3 Halieutichthys aculeatus Pancake batfish 1 Harengula jaguana Scaled sardine 3 Hippocampus erectus Seahorse 1,3 Holacanthus bemudensis Blue angelfish 2,3 Holacanthus ciliaris Queen angelfish 3 Holacanthus townsendi Townsend angelfish 3 Holocentrus adscensionis Squirrelfish 3 Hypoplectrus unicolor Butter Hamlet 3 Khyphosus sectatrix Chub 3 Lachnolaimus maximus Hogfish 2,3 Lagodon rhomboides Pinfish 2,3 Leiostomus xanthurus Spot croaker 1 Lutjanus apodus Schoolmaster 2 Lutjanus griseus Mangrove Snapper 2,3 Lutjanus synagris Lane snapper 3 Mulloidichthys martinicus Yellow goatfish 2,3 Mycteroperca bonaci Black grouper 2,3 Mycteroperca microlepis Gag grouper 2,3 Mycteroperca phenax Scamp grouper 2,3 Nicholsina usta Emerald parrotfish 1,2,3 Ocyurus chrysurus Yellowtail snapper 2,3 Ogcocephalus radiatus Polka-dot batfish 1 Ophidion spp. Cusk-eel 1 Opissthonema oglinum Threadfin herring 3 Opsanus pardus Leopard toadfish 1,2,3 Orthopristis chrysoptera Pigfish 1,3 Parablennius marmoreus Seaweed blenny 1,2,3 Paralichthys albigutta Gulf flounder 1,3 Paralichthys lethostigma Southern Flounder 3 Pareques umbrosus Cubbyu 1,2,3 Pomacanthus arcuatus Grey Angelfish 3 Pomacanthus paru French angelfish 3 Porichthys plectrodon Atlantic midshipman 1 Prionotus longispinosus Bigeye searobin 1 Prionotus martis Barred searobin 1 Prionotus ophryas Bandtail searobin 1
48 Prionotus roseus Bluespotted searobin 1 Prionotus tribulus Bighead searobin 1 Pristigenys alta Short bigeye 1 Ptereleotris calliura Blue goby 2,3 Rhinobatos lentiginosus Atlantic guitarfish 1 Rhomboplites aurorubens Vermilion snapper 1 Rypticus maculatus Whitespotted soapfish 1,2,3 Rypticus saponaceus Greater soapfish 1 Sardinella aurita Spanish sardine 1,3 Scartella cristata Molly miller 3 Scomberomorus maculates Spanish mackerel 2,3 Scorpaena brasilinesis Barbfish 1 Scorpaena calcarata Smooth-head scorpionfish 1 Seriola dumerili Greater amberjack 2,3 Serranus subligarius Belted sandfish 2,3 Serranus tigrinus Harlequin bass 2 Sp.? Filefish 1,2,3 Sp.? Orange blenny 3 Sp.? White goby 2,3 Sphoeroides spengleri Bandtail puffer 1,2,3 Sphyraena barracuda Great barracuda 2,3 Stegastes leucostictus Beaugregory 2,3 Stegastes variabilis Cocoa damsel 2,3 Syacium papillosum Dusky flounder 1 Symphurus urospilus Spottail tonguefish 1 Synodus foetens Inshore lizardfish 1,2,3 Synodus intermedius Sand diver 1,2 Thalassoma bifasciatum Bluehead wrasse 3 Trachinocephalus spp. Snakefish 1 2.6.2. Benthic Community Data Seven major categories (Coral, Porifera, Macroalgae, Dead Coral with Algae, Bleached Coral, Bare Substrate, and Other Li ving Fauna) were identified in the digital photo transects. Juvenile corals are consider ed as an eighth major category for a number of the analyses. Boxplots of the eight categories are disp layed in Figures 2.3 through 2.5 and 2.7 through 2.11. The data are displayed in two panels on each graph, each
49 corresponding to percent cover of the category at one of the two study sites, FWRI1 or MT. Note that the y-scale changes in each graph. Average transect coral cover (Fig. 2.3) varied from a low of 1.3% (July 2006, MT) to a high of 6.1% (June 2007, FWRI1). The data in the boxplots represent three replicates per sampling time, with ranges of coral cover varying within sampling times from as high as 7% (April 2006, MT) to as low as 0.23% (July 2006, MT). Adult coral species were identified and compared (presen ce/absence) to a list of species collected at Station B during the Hourglass cr uises (Table 2.2). Seven species, four of which were observed infrequently ( 10% of the time) at Station B, were not detected in photographs from the modern surveys. Percent cover of both macroalgae (Fig. 2.4) and bare substrate (Fig. 2.5) were highly variable. Regression analysis re vealed a significantly (p<0.05) negative correlation, with months of high macroalg al cover (e.g., May) corresponding to low percentages of bare s ubstrate (Fig. 2.6).
50 Figure 2.3. Boxplot of percen t coral cover at FWRI1 and MT from February 2006 to December 2007. The bars represent the in terquartile ranges, sample means are designated by a diamond and medians by a horizontal line.
51 Figure 2.4 Boxplot of macroalgal cover at FWRI1 and MT from February 2006 to December 2007. Figure 2.5. Boxplot of bare s ubstrate cover at FWRI1 and MT from February 2006 to December 2007.
52 Figure 2.6. Regression analysis of macroalgae versus bare substrate percent cover at natural ledges. Juvenile corals were detected in photo-t ransects in all sampling times (Fig. 2.7). Percentages of transects with juvenile cora ls varied from 9.6% (May 2007, MT) to a high of 47% (June 2007, FWRI1). Poriferans (2.8) also displayed major ranges and seasonal changes in average percent cover with a lo w of 0.0% (May 2006, FWRI1) and a high of 7.1% (FWRI1, November 2007). The remaini ng categories (Other Living Fauna, Dead Coral with Algae, and Bleached Coral; Figs. 2. 9 to 2.11) contributed very little to overall percent WFS cover (<2.5% during all samplings).
53 Figure 2.7. Boxplot of percentage of transect photos in which juvenile corals were identified at FWRI1 and MT from February 2006 to December 2007.
54 Figure 2.8. Boxplot of Porifera cover at FWRI1 and MT from February 2006 to December 2007. Figure 2.9. Boxplot of other living fauna cover at FWRI1 and MT from February 2006 to December 2007.
55 Figure 2.10. Boxplot of dead coral with alga e cover at FWRI1 and MT from February 2006 to December 2007. Figure 2.11. Boxplot of bleached coral cover at FWRI1 and MT from February 2006 to December 2007.
56 One-way ANOVAs revealed that only f our of the eight categories displayed significant seasonal changes (Tab le 2.6). Macroalgae and Bare Substrate cover values in fall were significantly different (F=10.6 and F=9.9, respectively; p=0) from both spring and winter values. Porifera percent cover va lues were significantly lower in the spring (F=7.7; p=0) than fall. Juvenile coral presence percentages we re significantly lower (F=10.5; p=0) in the spring as co mpared to all other seasons. Table 2.6. ANOVA and T ukeyÂ’s post-hoc comparison resu lts for the 4 categories that displayed significant seasonal differences. Re sults from the TukeyÂ’s test first list the categoryÂ’s determining season followed by the seas on(s) that it differs significantly from. Category ANOVA Results Diffe ring Seasons (TukeyÂ’s) Macroalgal Cover F=10.6; p=0 Fall Â– Spring & Winter Bare Substrate Cover F=9.9; p=0 Fall Â– Spring & Winter Coral Juvenile Presence F=10.5; p=0 Spring Â– Summer, Fall, & Winter Porifera Cover F=9.9; p=0 Spring Â– Fall A PCA (Fig. 2.12) of the eight categories (i ncluding juvenile corals) indicates that the first component is positively related to per cent cover of poriferans, bare substrate, and juvenile coral counts and negatively related to the percent cover of macroalgae. The second principal component is positively related to coral pe rcent cover and negatively to dead coral with algae. The first two components (eige nvalues of 2.7 and 1.2, respectively) cumulatively explain about 50% of the total variance, indicat ing that a number of other factors (or components) ar e involved in WFS benthic community dynamics.
57 Figure 2.12. Principal Components Analysis of Log (x+1) transformed data from seven major categories plus juvenile corals. Eigen values for th e first two components are 2.7 and 1.2, respectively, and cumulatively explai n approximately 50% of the variance. 2.6.3. Fish Assemblage Data There were a total of 47 fish species obs erved over the 22-month study period at FWRI1 and MT. Hourglass Program collections enumerated 59 species at natural ledge Station B over the 28-month sampling pe riod between 1965 and 1967. I also observed 71 species at a set of artificial re efs (Chapter 3), located near St ation B in comparable depths. The complete list of species ( 113 total) is shown in Table 2.5. The mobile nature of fish makes it difficult to quantify spatial and tempor al changes in assemblage structure. No transformations were able to make the data conform to normality so non-parametric multivariate procedures were used to yiel d insights into some of the changes in assemblage structure in the 22-month data set from the pooled FWRI1 and MT data (justification for pooling provide d in methods section).
58 The 2-D non-metric MDS ordination of fish data is shown in Figure 2.13. The 2D stress level (=0.2) configura tion signifies that the spatia l arrangement of the MDS may be a near random arrangement of samples, bear ing little resemblance to the original ranks (Clarke 1993). However, when the data ar e viewed in a 3-dimensional MDS graph by season factor, the stress level decreases to 0.14. The 3-D version is not presented here due to the complicated visualization of group-se paration that it provides. The 3-D MDS does indicate that there is sp atial separation amongst the gr oups and renders second-level seasonal procedures valid. Seasonal differen ces among fish assemblages are confirmed by an ANOSIM Global R=0.3 (p=.001). ANOSIM and SIMPER pair-wise comparison results are shown for those seasons that diffe red significantly in their fish assemblage (Table 2.7).
59 Figure 2.13. Non-metric multidimensional scalin g ordination (2-D) of fish assemblage samples over the 22-month study period. Separation of seasonal groups [e.g., July-Sep data grouped towards upper ri ght is confirmed via a 3-D an alysis which yields a lower stress (0.14)]. The 3-D graph is not shown due to poor visual representation. A priori groupings were analyzed based upon those grouping. Summer fish assemblages differed from th e other three seasons, and the same top four species ( Halichoeres bivittatus, Diplectrum formosum, Haemulon plumierii, Serranus subligarius ) were implicated in the dissimilarities. Abundances of D. formosum continuously decreased from winter through spring to summer, and then increased again in the fall. Halichoeres bivittatus abundances responded in the exact opposite manner through the season s (first increasing from wi nter to spring to summer, then decreasing in the fall). No other linke d trends were observed in the species data, although H. plumierii and S. subligarius did display opposite trends in abundance between the winter to summer samplings a nd the summer to fall samplings. Seasonal
60 trends were generally less clear and therefor e less quantifiable among fish assemblages as compared to the benthic community data. Table 2.7. ANOSIM analysis a nd SIMPER pair-wise comparisons of fish assemblages during fall, winter, spring, a nd summer samplings. Only those seasons that differed significantly from one another (ANOSIM R>0.3; p<0.05) are shown. The SIMPER results list the four top speci es contributing to the dissimila rity between the two seasons. Their relative abundance change (+/-) is also listed. 2.6.4. Abiotic Data Bottom temperature was measured from March to November 2007. Ten-day averages were plotted for FWRI1 (Fig. 2.14). Temperature trends at MT were similar as the sites are located in close proximity. Ma ximum daily fluctuations were observed at FWRI1 during March with a one-day increase in temperature from 17.5C to 19.6 C (Fig. 2.15). The greatest monthly increase in temp erature occurred over the month of May as temperatures at the bottom rose from 21.1 C to 24.8 C. The greatest temperature decrease occurred in November as temper atures dropped from 25.7 C to 21.3 C. Seasons ANOSIM Results SIMPER Results Winter & Summer R=0.4; p=.001 Halichoeres bivittatus (+) Diplectrum formosum (-) Haemulon plumierii (-) Serranus subligarius (+) Spring & Summer R=0.7; p=.001 Halichoeres bivittatus (+) Diplectrum formosum (-) Haemulon plumierii (+) Serranus subligarius (+) Summer & Fall R=0.4; p=.001 Halichoeres bivittatus (-) Diplectrum formosum (+) Haemulon plumierii (+) Serranus subligarius (-)
61 10 15 20 25 30 353/10/07 3/24/07 4/7/07 4/21/07 5/5/07 5/19/07 6/2/07 6/16/07 6/30/07 7/14/07 7/28/07 8/11/07 8/25/07 9/8/07 9/22/07 10/6/07 10/20/07 11/3/07 11/17/07 12/1/07Temp (C) Figure 2.14. Ten-day average bottom temperat ure data at FWRI1 from March to December 2007.
62 16 16.5 17 17.5 18 18.5 19 19.5 20 20.5 212/28/07 3/2/07 3/4/07 3/6/07 3/8/07 3/10/07 3/12/07 3/14/07 3/16/07 3/18/07 3/20/07 3/22/07 3/24/07 3/26/07 3/28/07 3/30/07Temp (C) Figure 2.15. Daily fluctuations in bot tom temperature during March 2007 at FWRI1. Data were collected every 10 mi nutes and daily aver ages are plotted. Average salinity over the 2-year study period at FWRI1, MT, and CW was 34.3 at the surface and 34.0 at the bottom. Light atte nuation coefficients (k-values) calculated from Secchi depths varied from a minimum of 0.07 in April to a maximum of 0.21 in August, with an average k of 0.16. These correspond to percent su rface light reaching a depth of 17 m depth between 2.7% and 33%, wi th an average of 6.4%. A number of phytoplankton species were observed in water sa mples collected at th e three study sites, but only September 2006 samples contained any Karenia spp. (low cell counts<30,000 cells/liter), indicating that harmful algal blooms were not affecting the areas during my study.
63 2.7 Discussion 2.7.1. Benthic Communities Livebottom communities have both seasonal and non-seasonal components to their structure. The assessment of seasonality is influenced by the type of data available and by practical limitations on sampling adequ acy. The degree of seasonal environmental variation along WFS livebottom ledges depends on depth, latitude, and proximity to the shelf edge. Past studies have shown that seasonal species richness/abundance variation is generally greater at shallow, inner-shelf bent hic communities as compared to mid-shelf or outer-shelf areas (MMS 1985). Seasonal biotic variations de tected in historic studies (primarily from the Hourglass Cruises) were proposed to reflect seasonal abundance patterns of different algal groups, which are prolific along the WFS (Table 2.4). The majority of species in Table 2.4 represent co llection and microscopic analyses from the Hourglass Cruise collections, which are opt imal methods for studying algal diversity. I identified algal species, when possible, from the digital photographs, but in general they were placed in a genera l Â“MacroalgaeÂ” category. Results from the Macroalgae (Fig. 2.4) a nd Bare Substrate (Fig. 2.5) boxplots and one-way ANOVAs from the 2006-2007 data (Tab le 2.6) corroborate historic seasonal hypotheses as significant seasonal differences were observed in both the Macroalgae and Bare Substrate categories. The two categorie s were negatively correlated (Fig. 2.6); high values of one category (i.e., high macroalgal co ver in the spring or bare substrate in the fall) corresponded to low cover values of th e other category. Macroalgae tended to dominate in the spring months (April and May). Anomal ously high percent cover of macroalgae in the late winter of 2006 (Febru ary, ~67%) are due to the abundant growth
64 of Â“slimyÂ” unidentifiable alga or cyanobact erium that covered the subtrata (Fig. 2.16). The growth had disappeared by the April and May 2006 samplings, replaced by fleshy macroalgal species. The slimy growth was absent during the February 2007 sampling. The prolific growth of the uni dentified microorganism may ha ve been opportunistic after the dissipation of the red tide, when other co mpeting species had peri shed as a result of the K. brevis bloom. Follow-up work after a future red-tide event might yield more insight into the algae/cyanobacteria dynamics in benthic WFS communities. Figure 2.16. Photograph depicti ng conditions at FWRI1 in February 2006. Anomalously high percent cover of the Â“MacroalgaeÂ” category was attributed to this unidentified algal growth. The growth had disappeared by the April 2006 sampling. The coral in the picture is Solenastrea hyades Two other categories also displayed si gnificant seasonal differences. Percent cover of Porifera (i ncluding clionids and Dysidea. spp.) was significantly lower in the spring as compared to fall values (Fig. 2.8; Table 2.6). Poriferans appear to be particularly sensitive to red-ti de disturbances (Chapter 4) su ch as the event that occurred
65 during the majority of 2005, and dissipated a fe w months prior to th e commencement of this study. Low percent cover of poriferans in the spring of 2006 coul d be attributed to mass mortalities during the 2005 red tide. Popul ations began to recover in late 2006 and into 2007. By the fall 2007 samplings, average cover at both sites hovered between 5% and 7%. The majority of seasonal differences in cover of poriferans could be attributed to red-tide effects as well as the Â“maskingÂ” e ffects that fleshy macroalgae might have in the digital photographs taken in the spring. Table 2.2 displays the list of 12 scleract inian and 1 milleporid corals identified at Station B (1965-1976), FWRI1 and MT (20062007), and GOM artif icial reefs (20052007). Although 13 corals were identified in the Hourglass Program, 4 of these were sampled infrequently over the 28-month period ( Millepora alcicornis, Porites divaricata, Isophyllia sinuosa and Astrangia poculata ) and understandably do not appear in my digital photographs. There is further evidence that two a dditional species ( Manicina areolata and Scolymia lacera ) may no longer be present along shallow inner WFS ledges (W. Jaap pers. comm.). The only other species observed at Station B and not in my data set is Balanophyllia floridana The different collection te chniques are implicated in the discrepancies. Benthic-dredging techniques ut ilized in the historic work will tend to collect more species than digital photo transe cts, leading to the a ppearance of a more diverse coral community. Percentage of transect photos in which j uvenile corals were identified was the last major category to differ seasonally (Fig. 2.7). Percentages were lower in the spring as compared to other seasons, probably because high macroalgal cover during spring months effectively masked the recruits. Trends in juvenile corals ar e discussed further in
66 the section below, as the importance of r ecruitment processes to the resilience and survival of livebottom ledges, and those fi sh species that inhabit them, along the WFS can not be overemphasized. There appeared to be significant spatial variability in the be nthic data along the WFS ledges as evidenced by la rge ranges in percent cove r within sampling times. Randomly-placed transects were used instead of permanently-fixed stations to maximize spatial coverage. However, it is impo rtant to note that ledges along the WFS are livebottom areas, meaning that their biotic c over is patchy in nature as compared to Â“traditionalÂ” coral reefs. The difference in percent cover of the major categories varied immensely from transect to transect during certain samplings, indi cating that a greater sampling effort (more random transects) would benefit future studies seeking to quantify livebottom communities along the WFS. 2.7.2. Juvenile Coral Recruitment Over the last few decades, studies on reproduction and ecology of reef corals have elucidated the sensitivity of these processes to natu ral and anthropogenic stresses (Hughes 1994; Wolanski et al. 2003; Bellwood et al. 2004). The recruitment of sexual and asexual individuals plays a major role in the dynamics of reef ecosystems and can ultimately play a role in both the short and long-term recovery (or decline) of a reef system. Much more study of coral larvae a nd recruitment is need ed on WFS livebottom habitats, including orig ins/reservoirs, spatial and tem poral recruitment scales, and juvenile survival rates, as there is little known about these processes. In the Caribbean, several studies have s hown that reef topography, depth gradient, oceanographic and environmental processes, as well larval dispersion (i.e., life histories),
67 contribute to the abundance, survivorship a nd distribution of cora l recruits (Bak and Engel 1979; Chiappone and Sullivan 1996; Edmunds et al. 2004). There are a number of upstream sources of larvae to the WFS, resulting from the inherent interconnectivity of the GOM via the Loop Current and its associated eddies and spin-offs (Berger et al. 1996; Sahl et al. 1997; Walker et al. 1997; Nowlin et al. 1998). Lugo-Fernandez et al. (2001) demonstrated that the Flower Garden Ba nks (FGB) in the northern GOM contain a repository of coral species that may function as a regional source of la rvae. More likely, larvae come from other coral-i nhabited ledges along the WFS. Although adult forms of cora ls are relatively easy to identify in digital photographs, it is very difficult to distinguish juvenile coral speci es. Photo-quadrats and transects are not optimal methods for recru itment studies, and tend to underestimate the number of juveniles as compared to in situ visual survey methods (Edmunds et al. 1998). Therefore, I made no attempt to identify the individual coral species in the photographs. Instead, monthly juvenile distributions we re assessed as a function of the average percentage of photographs in the transects (n=3 ) that contained at least one juvenile coral (<2 cm in size). The sizes of the juveniles varied from a pproximately 2 mm in diameter up to the 2 cm limit, as smaller juveniles were indistingu ishable in the photogra phs. Juvenile corals were present in all three 15 m transects dur ing all sampling times at both FWRI1 and MT, although an analysis of photographs from Novemb er and June indicate that juvenile coral sizes in June photographs were, on average, larger than November sizes. Many GOM and Caribbean coral species spawn after a fu ll moon or in concert with maximum water temperatures from July through September (Szmant 1986; de Graaf et al. 1999). Since
68 most coral larvae are competent within 3-10 days and competence periods can last as long as 120 days (Fadalllah 1983; Wilson and Harrison 1998), settling of some larvae along WFS ledges could begin immediately after spawning as early as July, particularly for brooding species whose larvae tend to settle in close proximity to the adult colonies. Larval settling could last through January for the larvae that travel long distances (i.e., from the FGB). Given the ge neral growth rate of 12 mm yr-1 for small corals (Bak and Engel 1979; Van Moorsel 1988), it is plausible that detection of juveniles in photographs could occur within 2-3 mont hs of spawning, which corresponds to the November sampling (assuming that settling of larvae o ccurred some time around August). Smaller size classes of recruits (between 2 and 4 mm) in November samples as compared to June (8 to 10 mm) corroborate these recruitmen t time scales and correspond to the peak spawning times of a number of the br ooding and broadcastspawning species. The lower numbers of juveniles in the spring months coincided with times of increased macroalgal cover (Figs. 2.4 and 2.7), particularly in May when macroalgal cover exceeded 60% at both site s. Accordingly, the cover of bare substrate decreased in the month of May and increased again towards th e end of the year (Fig. 2.5). The growth of macroalgae could obscure the juvenile corals, again indicati ng that photographic methods have limited resolution, especially in areas where seasonal changes in benthic cover are substantial. The percentage of photos c ontaining juvenile corals ha s a similar range at both sites although there are differences during certain samplings (Fig. 2.7). The random placement of transects and the small number of transects limits data resolution. More
69 detailed analyses of juvenile patterns were precluded by the methodology, although the data did unveil some interesting ideas and hypotheses which are discussed below. Rezak et al. (1990) proposed that the in stallation of thousands of oil and gas platforms along the northwest and central shelf of the GOM could provide stepping stones for corals to advance eastward across the Gulf, extendi ng their range through areas where substrate had previously been unsuita ble for settling and growth. The natural ledges along the WFS, with their limestone ou tcrops, function as stepping stones in their own right, as they provide suitable substrat e for the larvae of hardy coral species that originate in coral repositorie s such as the FGB, the Flor ida Middle Grounds, and other ledges along the WFS. WFS natural ledges ar e an essential link in the GOM basin-wide system connecting Caribbean co ral larvae entrained in Loop Current rings (Biggs 1992) to the FGB and on to the rest of the southeaste rn GOM (Lugo-Fernandez et al. 2001). It is important to note that the basin-wi de larval interconnec tivity studies are not limited to just coral larvae. Lee et al. (1992) demonstrated that eddies also remove and displace fish larvae in the Fl orida Current, affecting Florida KeysÂ’ species, particularly those that spawn in the water column. The mechanisms that prevail in the GOM have basin-wide implications for all larval organism s. The availability of substrate and habitat, as well as suitable environmental conditions and biological forces (e.g., predation and competition), are the keys to successfully recruiting new, sexually-produced larvae. Lugo-Fernandez et al. (2001) proposed that the strategic placement of artificial structures along the dispersing routes from the FGB could strengthen coral strongholds, and decrease the distance between sources of coral larvae in the GOM. Dupont (2008; see Chapter 3 of this dissertation) evaluated a se t of artificial reefs designed to mimic natural
70 WFS ledge relief deployed in approximately 20 m depth in previously unsuitable (sand over limestone) habitat. Within 4-5 years of deployment, a robust epibenthic community of corals (smaller species such as Cladocora arbuscula and Phyllangia americana ), poriferans, echinoderms, ascidians, and algae had developed on the ar tificial substrate. Seventy-one species of fish (demersal and pela gic) were found to be associated with the structures, providing evidence that artificia l reef placement in the GOM may be an effective way to boost larval survival between upstream and downstream sources and sinks. Lugo-Fernandez et al. (2001) suggest th at if coral populations of the FGB continue to thrive, they could contribute to the long-term recovery of damaged reefs of the southern GOM (Tunnel 1992) and the Flor ida Keys (Porter and Meir 1992) or perhaps become a coral refuge or repository. Larval supply, recruitm ent, and survival are important steps for resisting phase shifts to de graded alternate states and provide valuable information on the reproductive succ ess of species (Bellwood et al. 2004). An understanding of coral recruitm ent patterns and juvenile surv ival over time are essential to understanding ecological and physical pr ocesses that control population growth, distributions, and variability of community struct ures in time and space. It will also help us better understand how these systems fare after a natural or anthropogenic disturbance event. 2.7.3. Fish Assemblages Fish assemblage data are not oriously difficult to collect and analyze, particularly when resources, manpower, and logistics prev ent the collection of large, statistically robust data sets. The Bohnsack method (B ohnsack and Bannerot 1986) was chosen for
71 its ability to collect standard quantitative da ta on reef-fish assembla ge structure over a variety of habitats in an effi cient and effective manner. Howe ver, statistical robustness of the data depends on large sample sizes whic h, due to logistical issues (e.g., time and personnel constraints, dive limits), were often impossible to collect in my study. Therefore, the data presented in this paper represent an overview of shallow inner shelf fish assemblages but are by no means comprehensive. A total of 47 fish species were obs erved over the 22-month study period at FWRI1 and MT. Hourglass Program collections enumerated 59 species at natural ledge Station B over the 28-month sampling pe riod between 1965 and 1967 (Table 2.5). Only 12 of the species were observed in both the historic study and my study. Sampling techniques likely account for these observed differences. During the Hourglass Program, a flat trynet and a balloon trynet were dragged along the bottom for 15-30 minute intervals. The resulting collections repres ent a community heavily skewed to demersal species such as flounders (e.g., Bothus robins, Cyclopsetta fimbriata, Etropus crossotus, Etropus rimosus and Syacium papillosum ) and searobins (e.g., Prionotus spp.), while Bohnsack surveys tend to account for pelagic sp ecies, along with demersals. Although the sampling techniques were very different, th e combination of survey data yield a more comprehensive species list for shallow inne r WFS ledges. The list can function as an ecological baseline for scientis ts and marine managers. Additional data from artificial reefs (des ignated with a number Â“3Â”) at comparable depths/locations are also displayed in Table 2.5. Seventy-one species were observed at artificial reefs with 24 exclusively observed at the artificial reefs. The majority of the 24 additional species are tropical/subtropical sp ecies and include various angelfish (e.g.,
72 Holacanthus ciliaris, Holacanthus townsendi, Pomacanthus arcuatus and Pomacanthus paru ) and smaller tropicals (e.g., Calamus bajonado, Chaetodon sedentarius, Thalassoma bifasciatum, Hypoplectrus uni color, Scartella cristata and Chasmodes saburrae ). The artificial reefs are located south of FWRI 1 and MT (Fig. 2.1) and experience warmer water temperatures, as evidenced by averag e temperatures in March 2007 of 19.1C and other consistently warmer months. A multidimensional scaling ordination of samplings at FWRI1 and MT, relating their respective fish assembla ges (Fig. 2.13), tentatively groups the samples by season. An ANOSIM test was employed to test for significant seasonal differences. Results indicate that the fish assemblage during the summer differed from all other seasons (Table 2.7). Abundances of Halichoeres bivittatus, Diplec trum formosum, Haemulon plumierii, and Serranus subligarius were consistently the t op four contributors to seasonal dissimilarities. The reproduc tive habits, low population doubling times, mobility, and resiliency of these species may contribute to seasonal differences, as they quickly evacuate and re-populat e areas in response to chan ging environmental conditions and biological forces (e.g., a ggregate spawning, food supply, or pr edator to prey ratios). Halichoeres bivittatus are protogynous hermaphrodites that form leks (mating arenas) while spawning; peak spawning occurs in May or June (Allsop and West 2003). Haemulon plumierii peak spawning activity has also been reported in May (Murie and Parkyn 1999), although spawning has also been show n to occur year-round in this species, particularly in its southernmost distribution (Munro et al. 1973). Diplectrum formosum and S. subligarius are synchronously hermaphroditi c, with short population doubling times (<15 months), hence are highly re silient (Froese and Pauly 2008). Benthic
73 invertebrates (mollusks, crabs, worms, sh rimp, gastropods, and crustaceans) and other smaller fishes tend to be the major food sour ce for the four discriminating species. These four species may move from ledge to ledge along the WFS, following optimal environmental conditions and food sources as they become available and avoiding stressful environmental conditions su ch as hypoxia as they occur. Future work on fish assemblages alo ng the WFS ledges should incorporate a number of sampling methods a nd collection gears. The combination of data from three studies in comparable depths within a sma ll, geographic area revealed a diverse (113 species) fish community. Further studies are needed to quantify populations of fishes and determine their spatial and temporal distri butions. Smaller, non-commercial species (including the majority of those enumerated in this paper) may be important sources of food for the commercially-important species th at utilize natural le dges along the WFS throughout their life cycles. Managers and c onservationists should consider these areas as inherently important to sustaining the ec onomically-important fisheries of the eastern GOM. 2.7.4. Marginal West Florida Shelf Assemblages and Disturbance In this section, I discuss th e inherent marginal or tr ansitional conditions of the eastern GOM and the effects of regular disturba nces as they pertain to the development of livebottom assemblages along the WFS ledges. Marginal reef assemblages reflect the effects of steady-state or l ong-term average environmental limitations (Guinnotte et al. 2003). The WFS ledges are situated where fi rst-order determinants of global reef distribution (temperature, salin ity, nutrients, light, and ara gonite saturation state) are marginal (Kleypas et al. 1999).
74 The definition of Â“marginalÂ” with respect to coral reefs has been discussed in depth and Guinnotte et al. (2003) suggest that marginality may be defined in three ways: a) in a purely statistical sense, identifyi ng the subset of reef communities or conditions that are near the extreme of a particular su ite of environmental variables or community conditions; b) in terms of organism a nd community condition (cover, composition, diversity, health) or metabolism; c) on th e basis of proximity to an environmental condition known or reasonably assumed, ba sed on physiological or biogeographic evidence, to place an absolute limit on the o ccurrence of reef communities or key classes of reef organisms. Hardbottom outcrops a nd their associated livebottom assemblages along the shallow inner WFS can be defined as marginal under the second and third definitions. Under the second definition it is a pparent that although WFS livebottom assemblages are home to an abundance of bent hic flora and fauna including scleractinian corals and calcifying algae such as Halimeda spp, they are by no means comparable to accretional coral reefs where hi gh cover and diversity of zooxanthellate, scleractinian corals with hydrocorals and reef-associated calcifying algae epitomi ze the definition of a non-marginal reef community (Guinnotte et al 2003). Under the third definition, there is ample evidence that a number of first-orde r determinants (e.g., temperature, nutrients, salinity, light, and aragonite saturation state) defined by Kleypas et al. (1999) are at or near minimum or maximum limits for coralreef development along the eastern GOM. Guinnotte et al. (2003) defined high-temper ature, thermally stressed areas as those experiencing temperatures >31.1C. Temper atures >31.1 C were sustained for about a month between mid-August and mid-Septem ber at FWRI1 and MT in 2007. Increased
75 temperatures correspond to increased metabolic rates (N ichol 1967) and as a result, organisms inhabiting these areas may be part icularly sensitive to the development of hypoxic conditions during severe red-tide events. Those areas exposed to temperatures <18 C, especially for long periods of time, were also defined as thermally stressed by Guinnotte et al. (2003). Bottom temperatures at FWRI1 and MT were <18 C for the first 2 weeks in March 2007, and were also near this range throughout most of February. These results indicate that the inner W FS ledges experience marginal temperature conditions for reef development. Salinity ra nges at the sites are, however, within normal reef limits, although about 2 ppt lower than t ypical for Florida Keys reefs, which could negatively affect the aragonite satu ration state (discussed below). Nutrient concentrations along the WFS are also marginal for reef growth. Kleypas et al. (1999) averaged values across reef locations and found that 90% of reef locations have <0.60 mol L-1 nitrate and <0.20 mol L-1 phosphate. Ambient nitrate concentrations during non-bloom periods with in 5 km of the WFS coast are <0.5 mol L-1. Nitrate limitations along the WFS can, how ever, be alleviated when diazotrophs ( Trichodesmium spp.) bloom in response to iron-laden Saharan dust events (Lenes et al. 2001). Approximately 50-100% of the dissolved organic nitrogen (DON) excreted by the Trichodesmium is in the form of amino acids, whic h help mitigate nitrogen limitation for other members of the phytoplankton comm unity and the microbial loop, including Karenia brevis (red tide). Moreover, Florida is a phosphatic province and phosphorous species are rarely limiting, indicating that livebottom communities are near marginal nutrient limits in non-bloom conditions and can become inundated with nitrogen during
76 blooms, along with increased carbon input (Var go et al. 1987). In addition algae blooms can have a shading effect (Okey et al. 2004). Hallock and Schlager (1986) discussed the importance of water transparency and light intensity at depth as th ey pertain to coral-reef development and growth. Branching corals require approximately 60% of surface light, head corals require about 20%, and plate corals require 4%. Pe rcentage of light reaching liv ebottom assemblages along the shallow inner WFS (17 m depth) over my 22-month study averaged 6.4%, with a maximum of 33% and a minimum of 2.3%. The WFS assemblages are exposed to variable light intensities, with optimal light for photosynthesis occurring during the spring, when macroalgal cover increases, and lo wer light intensities dom inating in the fall. Aragonite saturation, the last of the first-order determinan ts of reef distribution as defined by Kleypas et al. (1999) covaries with temperature and salinity, from maximum values near the equator to minimum values ou tside the 20-30 latitude belt. FWRI1 and MT are situated between 27 and 28 latitude, and therefor e near the lower aragonite saturation limits but not outside of them. The lower salinity (~2 ppt < Florida Keys reefs) likely decreases aragonite saturation and thereby contributes to marginality for scleractinian corals among the WFS livebottom communities. West Florida Shelf hardbottom communities are, for the most part, exposed to conditions that are above the Â“lower limitsÂ” of salinity, light, and aragonite saturation for reefs, yet these are not reef-forming areas. This suggests that other factors prevent these communities from building reefs. The first pos sibility is that second-order determinants play an important role in limiting reef growth in this region. Second-order determinants include biological variables (i .e., species diversity and la rval sources) and hydrodynamics
77 (i.e., wave and tide action, sediment movement) which act on a region al scale (Kleypas et al. 1999). Larval availability, recruitment, and survival do not seem to be limiting factors in the development of livebottom communities, as long as suitable settling substrate is available. Hydrodynamic influences are ve ry limited along the shallow inner WFS as these ledges are situated at sufficient depths to avoid strong wave or tide action and resulting sediment movement. The patchy se diment distribution a nd close proximity of sediment types to their source, suggests that storms are not responsible for the large-scale sediment redistribution on the west central inner Florida shelf, but may be locally important (Brooks et al. 2003). Small-scale, periodic mobilization a nd redistribution of sediment by storms has been shown by Twic hell et al. (2003). However, it does not appear that second-order determinants are th e primary causes of the lack of reefs along the WFS. Another possibility is that the combination of thermal stress, abundant nutrients, and times of lowered light levels may cumulatively and synergistically prevent coral reef development. A third suggestion, specific to WFS hardbottoms, is that the chronic stresses imposed by the lower limits of certain first-order determinants, combined with acute disturbances such as red tides and storm/hurricane events, may restrict reef development, limiting livebottom species to thos e that are hardy, wee dy (quick to recruit or migrate back after a disturbance), and tolera nt of persistent chr onic and repeated acute disturbances. The spatial scale of acute disturbance affects ecosystem resilience (Sousa 1985), along with factors such as the frequency and duration of the disturbance (Nystrom et al. 2000). Estimates from FWRI indicate that approximately 5600 square kilometers of
78 benthic communities may have been affect ed by the 2005 red tide and the hypoxic/anoxic conditions (FWRI 2005). Table 2.8 places red tide events and associated hypoxic/anoxic conditions into context with other natural disturbancesÂ’ spatial extent, frequency, and duration. It also describes the level within the ecosystem that is most affected (individual, population, community, or ecosystem) and th e primary disturbance mechanism(s). Although natural disturbances such as red ti des and hurricanes can be detrimental to communities at large spatial scales (10-1000 km ), new substratum becomes available at various temporal and spatial scales (Connell 1978), increasin g the chance of recruitment and survival at the individual/population level. Patches of opportunity are opened up for renewal, development, and evolution as a resu lt of periodic disturbances (Holling 1996). Regional conditions that are marginal be tween temperate and tropical provinces along with chronic and acute disturbances in the eastern GOM influence community structure on livebottom ledges. The episodi c occurrence of severe red tides, in conjunction with other stochastic factors such as fluctuating sea temperatures, turbidity, and hurricanes, likely prevents the development of coral reef assemblages. Should the frequency and severity of di sturbances decrease, different community structures might develop, possibly a more Â“coral reef-likeÂ” comm unity. At present, hardbottom ledges, with their marginal environments, select fo r hardy species that can either survive the persistent marginal conditions and interm ittent large-scale acut e disturbances (e.g., Solenastrea hyades corals which temporarily retract thei r polyps or bleach in response to a disturbance, but quickly recover after the di sturbance has been alleviated) and/or whose r-selected reproductive characteristics enable them to quickly r ecruit to available substrate and utilize open niches (e.g., Diplectrum formosum and Serranus subligarius ).
79 Table 2.8. Natural disturban ces acting on WFS livebottom ar eas. The Â“level influencedÂ” column specifies whether benthic (B) or fi sh (F) levels are most influenced by the respective disturbance process. Process Spatial Extent Frequency DurationLevel Influenced Mechanism Predation and grazing 1 m DaysMonths MinuteDays Individuals (B,F) Mortality Bioerosion 1-10 m MonthsYears DaysWeeks Individuals (B) Communities (B) Creation and collapse of scarped hardbottom Bleaching/Disease 1 m MonthsYears DaysWeeks Individuals (B) Physiological weakening, mortality Storm events 1-102 km Months Hours Individuals (B) Populations (B) Communities (B) Sediment movement-burial and exposure Hurricanes 10-103 km MonthsDecades Days Communities (B) Physical disturbance Seasonality (temperature, light, etc.) Regional Annual Months Individuals (B,F) Populations (B,F) Light limitation, algal blooms, energetics Red tides 10-103 km MonthsYears Months Individuals (B) Populations (F) Brevetoxin effects Severe red tides resulting in anoxia 10-103 km YearsDecades MonthsYears Individuals (B) Populations (B,F) Communities (B,F) Brevetoxin and anoxic effects Â– exposure of bare substratum Sea-level or temperature change Global 104 -10 5 years 103 -10 4 years Communities (B,F) Ecosystems (B,F) Chronic stress Â– thermal, light, aragonite saturation etc. Kleypas et al. (2001) suggest that future aragonite saturation state reductions will gradually lead to less carbonate accumulation, slower coral extension rates and weaker skeletons, and possibly to reduced cementation and reef structure stabilization. This
80 suggests that non-framebuilding communities will become more common. They also postulate that if these area s are exposed to episodic extremes (i.e., thermal stresses, nutrient pulses), increases in mortality coul d be expected to occur. The livebottom assemblages of the WFS may represent the fu ture state of wester n Atlantic/Caribbean coral reefs that are currently at or near th eir marginal limits. The good news is that WFS ledge organisms, as individuals and popul ations, appear to have acclimated to intermittent episodic disturbances, giving some hope to the survival of other communities that reach the Â“tipping pointÂ” over to non-fr amebuilding reefs, after some period of acclimatization and selection. Recruitment of sexually or asexually-produced individuals is very important to the recovery of livebottom and coral-reef assemblages after a disturbance. Recruitment enhancement plans, such as the placement of artificial reefs al ong the WFS, should be seriously considered as a mechanism to enhance both epibenthic and fish communities along the natural ledges. Enhancing comm unities from the benthos up will increase productivity at the upper trophic levels, ensuring the preservation of important commercial and recreational fisheries in th e eastern GOM. The importance of benthic communities to the overall productivity of th e WFS should not be ignored or, worse yet, negatively affected by activities at the surf ace or in the pelagic zone. Appropriate considerations for livebottom areas, which occu py >50% of the shallow inner WFS, must be incorporated into any construction, management, and conservation plans.
81 3. Artificial Reefs as Restoration Tools: A Case Study on the West Florida Shelf 3.1 Abstract Artificial reefs are one of a number of tool s that should be cons idered by scientists and managers when planning coastal zone restor ation and/or mitigation projects. In this paper, the details of one proj ect from the West Florida Shel f are presented. Two types of artificial reefs were used to mitigate pipe line construction impact s on natural hardbottom ledges in the eastern Gulf of Mexico. The projectÂ’s primary objective was to avoid the paradigm of building artificial reefs as fish attraction device s, and to instead implement a design that would mimic, not augment, natu ral hardbottom conditions. Fish assemblage parameters (species richness and commercia l fish abundances) were compared between the artificial habitats and natural hardbottom re ference sites. Results indicate that species richness trends are similar among artificial and natural reefs, wh ile certain commercial fish abundances are significantly higher on th e artificial reefs. Recommendations for future restoration/mitigation projects using artificial reefs are discussed.
82 3.2 Introduction The modern era of artificial reef-building is barely a ha lf-century old, but its brief history has included extraordinary advancemen ts in structural de signs, technologies, and techniques, as well as changes in uses, stak eholder interests, management schemes, and evaluation criteria. Artificial reefs are define d as one or more objects of natural or human origin that are purposefully submerged to influence biological, physical, or socioeconomic processes related to marine resour ces (Jensen 1997; Seaman 2000). Artificial reefs have been used most prominently fo r fisheries harvest enhancement though they have been employed globally in a variety of other coastal management schemes including aquaculture in the Adriatic S ea (Fabi et al. 1989), enhancem ent of recreational diving and tourism opportunities throughout the United St ates (Milon 1991; Ditton et al. 1999), habitat rehabilitation in th e Maldives (Clark and Edwards 1994), and prevention of trawling in Europe (Reilini 2000). One of the more recent applic ations of artificial reefs has been for environmental mitigation purposes, especially in coasta l areas where physical damage by storms, exposure to toxic phytoplankt on blooms, destructive fishi ng practices, construction and dredging projects, and chemical pollutant contamination are among a few of the many natural and anthropogenic causes of habitat de gradation. In the rest oration of ecosystems after such damage, especially where physic al structure provides added benefits (e.g., habitat or shelter) to the ecosystem, artific ial reefs represent one potentially useful restoration tool (Pickering et al. 1998). P hysical structure in an ecosystem can be achieved in a number of ways, and definitive progress has been made since the early 1900s when artificial reefs were built as Â“a hit-or-miss dumping operation of unsightly
83 scrap materialÂ” (Dean 1983) such as tires and car bodies. While numerous studies have reported on the effectiveness of artificial reefs in aggregating fish (Randall 1963; Buchanan 1973; Stone et al. 1979), relative ly few studies discuss the design, location, planning, and evaluation of artif icial reefs (Bohnsack et al. 19 94) in relation to specific project objectives such as mimicking natural habitat for mitigation purposes or enhancing targeted species and their supporting community structures. It is increasingly becoming recognized that this is one of the major area s where further work is needed: determining the relative benefits of different de signs for production purposes (Bohnsack and Sutherland 1985; Seaman and Sprague 1991; Pickering and Whitmarsh 1997) and in meeting stated project objectives. In this paper, a detailed overview of one mitigation/restoration project on the central West Florida Shelf in the eastern Gulf of Mexico will be pr esented. The goal of the paper is to discuss the pertinent informati on of an artificial reef study as defined by Baine (2001) including details on the projectÂ’ s objectives, reef site, environmental conditions, design, monitoring, re sults and performance evaluation, and legal framework. Conclusions will be drawn regarding the success of the design and planning of the particular set of artificial reefs in fulf illing management goals and objectives, and explanations will be given for observed fa ilures in project execution. Recommendations for future mitigation/restorati on projects using artific ial reefs in the Gulf of Mexico, and in coastal areas worldwide, will be discussed.
84 3.3 Artificial Reef Projec t Background and Objective In 2001, Gulfstream Natural Gas System, L. L.C. (heretofore referred to as GNGS) constructed a 90-cm diameter pipeline acro ss the Gulf of Mexico (GOM) to transport natural gas from plants in Mississippi and Al abama to markets in central and southern Florida. Under the Federal Mitigation Pla n, GNGS was required to measure, mitigate, and monitor construction impacts to hard/live bottom benthic habitats in the GOM. The overall objective of the mitigation sites was to mimic the natural habitats (fish and benthic communities) that were either di rectly impacted by pipeline construction activities or indirectly affected by increased water column turbidity and sedimentation. A secondary objective was to evaluate the e fficacy of two different reef designs in achieving the primary goal. 3.4 Artificial Reef Site Description Compensatory mitigation for livebottom impacts caused by pipeline construction included the installation of six artificial reef sites on the seafloor (16-20 m depth) in Federal Waters, 19-25 km west of the mouth of Tampa Bay, FL (coordinates of sites have not been publicized to ensure that natura l community development occurs without the impacts of recreational diving a nd fishing activities). Three of the six sites were created by dispersing approximately 13,000 metric tons of limestone boulders (>1 m diameter) in 150 m x 150 m areas. These will be referred to as limestone boulder (LB) sites. Three additional sites consisted of grouped placemen t of pre-fabricated 1.8 m wide x 2.7 m long x 1.8 m tall reef modules (Fig. 3.1; de signed by H. Hudson, U.S. patent #5215406) constructed of limestone in a concrete ma trix in 150 m x 150 m areas. A total of 153
85 modules were dispersed among the three sites and will be referred to as reef module (RM) sites. All six artificial reef sites were de ployed near natural live bottom areas and on sand bottom that did not exceed a thickness of 0.6 m. Figure 3.1. Artificial reef module designe d by H. Hudson (U.S. patent #5215406) and constructed of limestone in a concre te matrix (each module occupies 8.7 m3). The cavity passes through the entire lengt h of the module. Three groups of 17 modules were placed at each of the three Reef Module (RM) artificial r eef habitats. Ten Reference (R), or control, sites were established in close proximity to the artificial reefs, in unimpacted livebottom ar eas. These sites were monitored consistently along with the artificial reefs sites and the data will be used for comparative analyses as no comparative, quantitative community data ar e available from the impacted sites prior to the advent of construction ac tivities. The projectÂ’s lack of pre-construction data should be noted and remedied in future mitigation efforts. Consistent measurements of abiotic parameters, habitat character istics, and biotic data s hould be performed prior to construction activities for a sufficient durat ion of time (length of time will vary among
86 projects and should be determined prior to start of project) as these data allow for effective evaluation of a projectÂ’s progress to wards achieving the stated objective. 3.5 Environmental Conditions Seasonal (summer and winter) abiotic parameters (temperature and Secchi depth), and habitat characteristics (rugosity and de pth) were measured and the results are displayed in Table 3.1. Surface roughness and ve rtical complexity were measured using a Rugosity Index calculated as th e ratio of a fixed length of chain (9.6 m) to the linear distance traversed by the chain. Rugosity m easurements were significantly different among all three habitat types (Kruskal-Wallis H=24.3, p=0.001) with LB sites consistently displaying highest surface r oughness and R sites the lowest. All other abiotic parameters were not significantly di fferent among the sites within the respective sampling.
87 Table 3.1. Abiotic and habitat characterization data summary presented as mean (n=10) values ( S.E). Data were recorded during each of the 5 sampling times a Parameter Limestone Boulders (LB) Reef Modules (RM) Reference Stations (R) Water Depth (m) 19.5 (1.1) 16.9 (0.2) 18.5 (0.9) Rugosity Index 1.53 (0.04) 1.36 (0.08) 1.13 (0.03) Water Temperature (C) Summer 2005 Winter 2005 Summer 2006 Winter 2006 Summer 2007 26.5 (0.5) 22.2 (1.0) 29.1 (0.3) 18.2 (0.7) 26.7 (0.8) 27.3 (0.1) 22.8 (0.7) 29.6 (0.2) 18.2 (0.6) 26.3 (0.5) 26.9 (0.2) 20.7 (1.2) 29.4 (0.4) 18.1 (0.4) 25.9 (0.5) Secchi Depth (m) Summer 2005 Winter 2005 Summer 2006 Winter 2006 Summer 2007 13.9 (0.5) 9.0 (0.7) 17.6 (2.1) 12.6 (0.8) 10.5 (0.7) 10.9 (0.4) 7.3 (0.3) 9.6 (0.3) 12.3 (0.4) 9.4 (0.7) 8.5 (0.4) 9.8 (0.5) 12.9 (1.7) 12.8 (0.7) 9.8 (0.6) a Data are adapted from 5 GNGS re ports (GNGS, 2005a; GNGS, 2005b; GNGS 2006a; GNGS 2006b; GNGS, 2007) 3.6 Artificial Reef Size and Design The influence of artificia l reef size and structure on species abundance and richness is an ongoing debate, as is the debate over whether increases in artificial reef fish biomass are a result of simple attraction to the structure versus new production (Pickering and Whitmarsh 1997). Results from numerous studies indicate that larger reefs, with greater habitat heterogeneity, te nd to attract a greater number of persistent species and a higher biomass (Campos and Gamboa 1989; Bohnsack et al. 1994; Moffitt et al. 1989; Pickering and Whitmarsh 1997). Bohnsack et al. (1994) attributed the higher biomass densities on large reefs to larger but fewe r individuals which out-competed or preyed
88 upon smaller individuals (including juveniles). They suggested that larger reefs may be better for aggregating large adult fishes, whereas smaller reefs are better for overall recruitment, as significantly higher settler mo rtality was observed at the larger reefs due to increased competition and predation from larger resident populations and larger individual fish (Bohnsack et al., 1994). The GNGS reefs are all large ar tificial reefs, as compared to reefs in the Bohnsack et al. (1994) study and other work (Rounsefell 1972). The GNGS reefs were designed to mitigate pipeline effects in an equivalently-si zed area, and provide habitat for adult fishes that may have been displaced due to cons truction activities. The six GNGS reefs each covered a similar sp atial area (22,500 m2), but the design and layout of the reefs (LB and RM) were very different, allowing for st atistical comparisons between species colonization trends and assessment of the e fficacy of the two different reef types in mimicking natural trends. The LB sites were created by lowering approximately 13,600 metric tons of boulders (> 1 m diameter) into the 22,500 m2 areas. The boulders were strategically overlapped and stacked during deployment to provide various swim-through holes, crevices, and sheltered areas. The boul ders were spread contiguously throughout the area, as opposed to the RM sites which consisted of the ordered placement of 51 modules per site (three groups of 17 modul es), with approximately eight meters separating each module. Areas in between modules consisted of bare, unconsolidated substrate. Each module occupied 8.7 m3 and was designed with one crevice cut into the limestone at the top of the c oncrete matrix that extended th rough the entire module (Fig. 3.1).
89 The LB site design is more representativ e of natural substrates in the GOM (and other hardbottom or reef areas) which are comp rised of various rock types with different physical relief, modified by the provision of secondary substrate (Pickering and Whitmarsh 1997). Small and large crevi ces are dispersed throughout the area and different orientations of the physical structure alter water flow regimes in the area. Though LB sites displayed greater structural complexity (Table 3.1) both the LB and RM sites were designed as low vertical relief (<3 m) artificia l habitats to mimic the relief of the natural inner continen tal shelf which consists of 50% exposed hardbottom, superimposed with ledges or scarps up to 4 m in relief (Hine et al. 2003 ). It is important to note that, in this particular case st udy, the objective was no t to enhance fishery harvests, but instead to mitigate losses to na tural habitat through restoration of equivalent fish and benthic invertebrate populations. The use of low vert ical relief designs in the GNGS work is not typical of other coastal artificial projects whose main objective is to attract and aggregate fish for fishing purposes through use of large, heterogeneous artificial reefs. 3.7 Monitoring Methods GNGS collected data on epibenthic and fish communities as part of the monitoring portion of the Federal Mitigati on Plan. This paper will focus on fish assemblage data, with further emphasis on co mmercial species, as reef fish abundances and diversities demonstrate a significant dependence on availabl e habitat (Sale 1978; Moffitt et al. 1989; Pratt 1994). A qualitative overview of epibenthic macroinvertebrate and macroalgal communities are given in this paper and will be discussed in detail in
90 future publications, as their contributions to ar tificial reef performa nce evaluation is very significant. Ten point-count censuses of the fish communities were conducted at each of the three habitat types (LB, RM, and R) in Summer 2005 (June), Winter 2005 (November 2005-January 2006), Summer 2006 (July-August), Winter 2006 (December 2006-March 2007), and Summer 2007 (June 2007). Census es were conducted using the Bohnsack Point Count Method (Bohnsack and Bannerot 19 86; Bohnsack et al. 1994). The data are summarized in five separate GNGS re ports (GNGS 2005a; GNGS 2005b; GNGS 2006a; GNGS 2006b; GNGS 2007), but the author was granted access to th e individual excel data files to further analyze temporal and sp atial trends through inter-sampling and interhabitat statistical analyses. Total fish assemblage data were averaged (n=10) within each of the three habitats during the five sampling times. Species ric hness values were compared to determine whether temporal changes in fish assemblages followed similar patterns at the artificial reefs and reference sites. Commercially important species were anal yzed separately due to their importance to Gulf of Mexico fisheries management conservation, and ec onomy. Commercially important species were identified using Th e Gulf of Mexico Fishery Council (2005) commercial fishing species list. Temporal tr ends were ignored as the data were pooled with respect to habitat (n=50) with the goal of determining whether artificial reef sites differed significantly from reference sites. Analysis of Variance (ANOVA) followed by TukeyÂ’s post hoc comparison tests were used to test for significant differences among each commercial fish speciesÂ’ abundance m eans over the three type s of habitat.
91 3.8 Results and Performance Evaluation A total of 71 fish species were observed at the artificial reefs and reference sites over the course of the study. An index of sp ecies richness values at each of the three habitat types is shown in Figure 3.2. The data are part of a larger study that assesses the impacts of a massive Karenia brevis (red tide) bloom that passed through the GNGS areas shortly after the Summer 2005 sampling (Heil 2006; Dupont Chap ter 4). Although the macroinvertebrate and fish communities at the artificial reefs and adjacent natural, reference sites were negatively affected by both the dinoflagellate toxin as well as the development of a bottom anoxic layer, the K. brevis bloom did provide a literal Â“blank slateÂ” for comparing recolonization and recr uitment patterns of communities at the artificial and reference sites. These comparative data are very useful in assessing and evaluating the artificial reef sÂ’ efficacy in achieving the GNGS projectÂ’s primary goal: mimicking natural reef biotic composition a nd patterns. The secondary objective can also be accomplished as the tw o types of artificial reefs (LB and RM) can be statistically compared to the reference sites and to one another.
92 Figure 3.2. Species richness trends over the five samplings: Summer 2005 (S05), Winter 2005 (W05), Summer 2006 (S06), Winter 2006 (W06), and Summer 2007 (S07). The data are part of a larger study that assessed the effects of a Karenia brevis (red tide) bloom that passed through the area imme diately after the Winter 2005 sampling, extirpating a majority of th e benthos and altering the fi sh assemblage. Recovery trajectories of species richne ss at the three habitat type s (RM Â– Reef Modules, LB Â– Limestone Boulders, and Ref Â– Re ference) are very similar. Fish species richness trends were very similar as declines of 50-65% were observed between the Summer 2005 and Winter 2005 samplings at LB, RM, and R sites (Figure 3.2). The number of sp ecies recovered to 80-124% of their original values by the Summer 2007 sampling, with similar recovery tr ajectories displayed at all three habitat types. The similarities between species ri chness patterns is a promising observation, as the LB and RM recovery trajectories appear to effectively mimic the R sites. LB sites did display consistently higher species numbers and abundances as compared to RM and R
93 sites during all samplings. These trends c ould be explained by the greater habitat heterogeneity and vertical relief inherent with in the LB site design (Table 3.1). The high concentration of boulders in the LB areas (i n contrast to the e qually spaced groups of modules at RM sites), coupled with the pres ence of numerous protected areas and swimthrough holes due to boulder overl ap and stacking, would be e xpected to cater to a more diverse fish community as opposed to the lowe r relief, less spatially complex nature of the RM and R sites. Vertical relief within a structure varies tur bulence patterns, water flow, sedimentary regimes (Pickering a nd Whitmarsh 1997) and larval settlement patterns, all of which promote a diverse comm unity structure. The structural complexity of reefs, particularly the pr esence and variety of crevices (Luckhurst and Luckhurst 1978; Anderson et al. 1989), the proximity of ne ighboring modules, and the provision of secondary biotic space through bio-fouling (Palmer-Zwahlen and Aseltine 1994) have been shown to contribute significantly to sp ecies composition, coloni zation patterns, and biological productivity of reefs. It should, however, be not ed that other studies have shown that certain fish do prefer less co mplex structures (Risk 1972; Sale and Douglas 1984) and if management plans call for the rest oration or enhancement of these particular target species, rather than the overall fish assemblage, then vertical relief should be varied accordingly. Of the 71 observed fish species, 12 are listed as commercially important or protected species according to The Gulf of Mexico Fishery Council (2005). These species are: Mycteroperca microlepis (Gag grouper), Mycteroperca bonaci (Black grouper), Mycteroperca phenax (Scamp grouper), Epinephelus itajara (Goliath grouper), Epinephelus morio (Red grouper), Seriola dumerili (Greater amberjack), Balistes
94 capriscus (Grey triggerfish), Lachnolaimus maximus (Hogfish), Lutjanus synagris (Lane snapper), Lutjanus griseus (Mangrove/Grey snapper), Scomberomorus maculates (Spanish mackerel), and Ocyurus chrysurus (Yellowtail snapper). ANOVA analyses and post hoc tests indicated that only five speciesÂ’ abundances ( M. microlepis, M. phenax, L. griseus, B capriscus, and L. maximus ) were significantly higher (ANOVA F >12, p<.0001) at one or both of the ar tificial reef habitat types. The five speciesÂ’ sampling distributions at the three hab itat types, including median, interquartile range, upper and lower limits, and outliers, are plotted in Fi gure 3.3. The remaining seven species were observed infrequently, and the data were c onsistently classified as outliers with no significant differences observed among the thr ee habitats. Several of the species that showed no variation in a bundances among habitats (e.g., S. dumerili, S. maculates, O. chrysurus ) are pelagic, migratory species that disp lay less site fidelity and dependence on benthic habitats. These species are expect ed to benefit less from the placement of artificial reefs as compared to demersal, phi lopatric, habitat-limite d, territorial, and/or reef-dependent species (Pic kering and Whitmarsh 1997).
95 Figure 3.3. Sampling distributi on plots (including median, interquartile range, upper and lower limits, and outliers) of the five commercially important fish species whose abundances were significantly higher (ANOVA F> 12.0, p<.0001) at the artificial reef habitat types (RM and LB) as compared to th e natural hardbottom/refe rence (R) habitats. An additional seven commercial species displayed no significant differences among habitats. The five species that preferred the ar tificial reefs (Fig. 3.3) are euryphagic carnivores, with a number of them feeding on benthic invertebrate s and smaller fishes. The epibenthic communities on the artificia l reefs were typically dominated (percent cover values > 60%) by algal sp ecies including cyanophytes a nd other filamentous algae, as well as rhodophytes ( Gracilaria and Eucheuma spp), chlorophytes ( Caulerpa mexicana, C. racemosa, Halimeda spp., and Udotea spp. ), and other undistinguishable macroalgae. Sessile and slow-moving macroinve rtebrates included scleractinian corals
96 ( Cladocora arbuscula and Phyllangia americana ), poriferans ( Cliona spp. ), and Echinoderms ( Diadema antillarum and Astrophyton muricatum ) which, along with bare substrate (rubble), accounted for the majority of the remaining benthic percent cover on the artificial reefs. The reference (R) sites, in contrast, were relative ly devoid or sparsely colonized by epibenthic inve rtebrates and macroalgae. Although mobile invertebrates, including mollusks, gastropods, and crustaceans, were not enumerated in the study, it is assume d that a diverse sessile macroinvertebrate and algal community on the artificial reefs c ould provide secondary substrate and shelter for mobile invertebrate fauna, hence e nhancing food source availability for both carnivorous and herbivorous fish species. The preferential association of the five commercial species, as well as many of the ot her non-commercial fishes, with the vertical relief and substrate provided by the artificial reefs (particularly the LB sites) can be attributed to: 1) the provision of shelter and habitat (attract ive to both juvenile and adult stages), and 2) the development of seconda ry substrate and epib enthic invertebrate communities that provide food sources. The results from the fish assemblages i ndicate that the artificial reefs were effective in mimicking the natural hardbotto m areas in the GOM in terms of speciesrichness trends and abundances of most comm ercial fish species. The GNGS artificial reefs were built as low relief structures to increase the environmental carrying capacity and biomass of the previously vacant ar eas, while maintaining and mimicking the integrity of natural hardbottom areas in the GOM as a means of environmental mitigation (i.e., avoiding the paradigm of artificial reefs as pure fish attraction devices).
97 3.9 Ecological and Legal Framework The GNGS reefs were successful in fulf illing the project objectives due to a number of opportune situational cond itions, within both the ecological and legal/management contexts. From an ecologi cal standpoint, the abiotic conditions of the uncolonized substrates in the GOM are ideal fo r artificial reef placement and recruitment of thriving epibenthic and fish communities. The LB and RM sites were chosen in areas with less than 0.6 m of unconsolidated substr ate (mostly fine quart z and biogenic sands), underlain with calcitic or dolomitic limestone (Obrochta et al. 2003) to minimize sinking of the artificial reef habitats The depths of the reefs (>16 m) were sufficient to avoid displacement or movement due to wave a nd wind action during tropical storms and hurricanes as reported by divers that examined the reefs after a series of severe hurricanes passed through the GOM in 2004 and 2005. The complicated physical oceanographic dynamics of the region (wind and wave fo rces, tidal currents, Loop Current eddy intrusions, tropical storms, Ta mpa Bay influences, etc.) conne ct the GNGS sites to areas throughout the GOM as well as coastal and near shore Florida state wa ters, in terms of larval supply, anthropogenic influences, nutri ent supply, fluctuations in temperature, salinity, dissolved oxygen, and other important abiotic parameters that affect the health of ecosystems. The proximity of the GNGS artificial reefs to natural hardbottom areas provided an initial supply of adult fish that quickly colonized th e areas, but future diversity patterns may vary considerably given the dynamics of the region, and should continue to be investigated. From a legal or policy stance, the deploymen t of artificial reef s is often dependent on a complex array of permits and authori zations with the outcome a compromise
98 between the mandates and agendas of nume rous local, national, and international agencies (Pickering et al. 1998). The ability to award a permit or l ease and the conditions attached to the award, depend on how the proposed project fits within the legislative and policy frameworks governing the actions of the agencies involved and the use of coastal areas (Pickering et al. 1998). Three permits, w ith subsequent modifi cations, were issued to GNGS for construction and operation of the pipeline. Permits for Federal waters were issued by the U.S. Army Corps of Engineer s and Minerals Management Service. The Florida Department of Environmental Protecti on issued a permit for operations in Florida State waters. Extensive discussion, collabo ration, and compromise among agencies was required but the resulting permits, and subs equent modifications, ensured that proper construction, operation, mitigation, and mon itoring activities ensued in the GOM and along the West Florida Shelf. The site pl acement of the GNGS reefs (i.e., in Federal waters and at sufficiently deep depths) avoided many of the issues that usually accompany deployment of artificial reefs in coastal areas including provisions for the safety of navigation, cables and pipelines, coastal defense, and development control (Pickering et al. 1998). The issue of regula tion of fishing and diving activities on the GNGS reefs has been avoided, beacause the coor dinates for the sites are not available to recreational anglers, dive charters, or the general public, thereby allowing the artificial reef sites to be monitored without the influence of anthropogenic pressures. 3.10 Conclusion Coastal managers along the West Florida Sh elf, and other coasta l areas worldwide, should consider artificial reefs to be one of a number of available management tools in
99 future mitigation and restoration projects, although it is essential that projects be evaluated on a case-by-case basis so that coas tal managers define and execute a plan that fits their specific areaÂ’s e nvironmental, economic, and social needs as well as the available resources. Case studies of various artificial reef projects are important in elucidating planning schemes, personnel a nd agency structures, resource needs, management, monitoring, and enforcement techniques that make certain projects successful, where others fail. Baine (2001) attri buted the failure of most artificial reefs in meeting project objectives to seven major issu es: siting, size, stabil ity, cost, inadequate monitoring, unmanaged local use, and the influen ce of external climatic factors. Of these seven proposed causes of failure, only the influe nce of external clima tic factors is beyond the scope of general planning and manageme nt. The other six factors are essential components to a coastal management scheme th at employs the use of artificial reefs, and it is essential that managers consider and evaluate these in detail.
100 4. Ecological Impacts of the 2005 Red Tide on Artificial Reef Epibenthic Macroinvertebrate and Fish Communities in the Eastern Gulf of Mexico 4.1 Abstract A harmful algal bloom (red tide) and a ssociated anoxic/hypoxic event in 2005 resulted in massive fish kills and comparab le mortality of epibenthic communities in depths less than 25 meters along the central we st Florida shelf. Th ere is a robust body of information on the etiology of red tide and human health issues; however, there is virtually no quantitative information on the effects of red tide on epibenthic macroinvertebrate and demersal fish communities. An ongoing monitoring study of recruitment and succession on artificial reef structures provided a focused time series (2005 to 2007) before and afte r the red tide disturbance. Radical changes in community structure were observed after the red tide. Scleractinian cora ls, poriferans, and echinoderms were among the epibenthos most affected. Fish species richness declined by >50%, with significant reductions in the abundances of most species. Successional stages were monitored over the next two y ears; stages tended to follow a predictable progression and revert to a prered tide state, corroborating pr evious predictions that the frequency of disturbance events in the sh allow eastern Gulf of Mexico may limit the effective species pool of colonists. The data indicate that recovery times may be shorter than predicted in previous studies.
101 4.2 Introduction Blooms of the toxic dinoflagellate Karenia brevis have been documented along the west Florida shelf since the late 1800Â’s (In gersoll 1881). These re d tides have varied in location, size, duration and intensity, and exposure to the brevetoxins has been shown to affect vertebratesÂ’ (fish, marine mamma ls, and humans) central nervous systems by alteration of sodium channels (Kirkpatrick et al. 2004). There is no single known cause of the red tides, though several factors have been suggested to pl ay a role, including eutrophication (Dixon and Steidinger 2004; Brand and Compton 2007), upwelling and current regime (Tester and Steidinger 1997) a nd iron fertilization (W alsh et al. 2006). Minor K. brevis blooms (< 105 cells L-1 as defined by the Florida Fish and Wildlife Research Institute, St. Petersburg) of limited duration and associated fish kills may be an annual, natural phenomenon in coastal waters of the Gulf of Mexico (Steidinger and Ingle 1972; Walsh et al. 2006), although occasionally large blooms (>105 cells L-1) cause mass mortalities of fish, marine mammals, and ot her marine life (Lands berg 2002; Flewelling et al. 2005). Given the prevalence of minor and majo r bloom events, surprisingly few studies have investigated the effects of red tides on benthic invertebrate and demersal fish communities on the west Florida shelf. In a qualitative study of the impacts of the 1971 red tide, Smith (1979) reported that 77% of shallow-water (12-18 m) resident fish perished. Echinoderms, gastropod mollusks, decapod crustaceans, sc leractinian corals, polychaetes, and poriferans sust ained even higher mortalities. Post-impact recolonization studies by Smith (1975, 1979) indicated that major red tides might result in near-
102 extirpation of shallow-water livebottom biot as, requiring a decade or more for benthic communities to recover to pre-red tide conditions. With respect to the infaunal bentho s, Simon and Dauer (1972) conducted a quantitative study of communities in the nor thern Tampa Bay estuary, also during the 1971 red tide event. They compared the co mmunity structure befo re and after the red tide and quantified the sustained losses. Only 5 of the 22 most abundant species remained on the intertidal fl at after the 1971 event. Repopulat ion of the polychaete fauna and reestablishment of the benthic commun ity following the natural defaunation were quantified and modeled in th e subsequent years (Dauer and Simon 1976; Simon and Dauer 1977). Recovery rates of infaunal comm unities in Tampa Bay were much faster than those predicted by Smith (1975) for easte rn Gulf of Mexico livebottom systems. The studies outlined above represent the ma jority of the information available on the ecological effects of red tides on benthi c and demersal communities in the eastern Gulf of Mexico. Typically, studies on ha rmful algal blooms have focused on the acute effects of algal toxins, rather than ecologica l impacts of chronic exposure to algal toxins (Van Dolah et al. 2001). At the lower trophic levels, acute exposure to algal toxins has been shown to produce deleterious effect s on zooplankton, includi ng reduced feeding, growth, and egg production (Gill a nd Harris 1987; Turner and Test er 1997). It is virtually unknown how chronic exposure to algal toxins may impact population dynamics of other lower trophic level species, and how changes in these dynamics may ultimately affect important commercial and recreational fish populations over time-scales spanning years to decades (Van Dolah et al. 2001).
103 The objectives of this study were to quantify the impacts of the 2005 red tide/hypoxia disturbance (Heil 2006), which pers isted for over one year along the west Florida shelf, on artificial reef epibenthic macroinvertebrate and fish communities. We also monitored successional stages of and te mporal changes to the communities for two years post-event. The artificial reef communities were chosen due to the availability of a Â‘before-impactÂ’ database that provided us w ith an important ecological baseline. The unique baseline is used to assess the imme diate red tide impact and examine recovery trajectories of benthic and fish communities in a specific habitat area. These data can be used in future quantification of seasonal a nd annual changes that result from natural or anthropogenic disturbances. Data greatly augment the limited database of communityscale ecological impacts of red tides in th e Gulf of Mexico and represent the first quantitative, multi-year study of epibenthic macroinvertebrate and demersal fish community dynamics after a t oxic red-tide disturbance. 4.3. Methods 4.3.1. Study Area Characteristics In 2001, as mitigation for construction of a natural gas pipeline, Gulfstream Natural Gas Systems (GNGS) installe d artificial reef structures at six sites in U.S. Federal Waters, 19Â–25 km west of Tampa Bay and in water depths of 18Â–25 m (Fig. 4.1). Three sites consisted of limestone boulders (>1 m diameter) haphazardly dispersed to provide some overlap and habitat structure (here after referred to as Â“LB site sÂ”). Pre-fabricated 1.8 m x 2.4 m reef m odules (H. Hudson TM) were installed at the other three sites (here after referred to as RM sites). A total of 153 modules were constructed of limestone in a concrete matrix and dispersed among the thr ee sites. The six mitigation sites were
104 deployed on sand bottom that did not exceed a thickness of 0.6 m. Under the Federal Mitigation Plan, GNGS was required to monito r the development of benthic and fish communities at these LB and RM sites, as well as at three adjacent undisturbed Reference (control) sites (here after referred to as Â“R sitesÂ”). Abiotic parameters (temperature and Secchi depth) and habitat characteristics (rugos ity and depth) were recorded during each of the five fish censuses (described below). Figure 4.1. The location of GNGS Limestone Boulder (LB) and Reef Module (RM) sites in the eastern Gulf of Mexico. Reference (R) sites were located in close proximity.
105 In addition to the regular sampling of abiotic parameters during each fish census, scientists from the Florida Fish and Wildlife Research Institute (FWRI) in St. Petersburg, FL sampled temperature, salinity, dissolve d oxygen, and relative fluorescence along an east-west transect during the peak of the red tide bloom in August 2005 (FWRI unpubl. data). The cruise track extended out 24 km from Bunces Pass and passed through areas located 5-7 km north of the GNGS artificial reef sites. Satellite images (MODIS and SeaWIFS) were examined along the west Fl orida shelf, but the presence of clouds precluded determining whether a unified wate r mass extended throughout the 5-7 km area. The FWRI data are presented as representativ e of the parameters that were observed at depth during the height of the 2005 red tid e bloom in the general area of the GNGS artificial reefs. 4.3.2. Benthic Communities Eight 1 m2 photostations were digitally photogr aphed in March 2005 (prior to the red tide), August 2005 (during the red tide ev ent), July 2006, and March 2007. Three of the photostations were located at LB sites and five at RM sites. The center of each photostation was marked with a 0.67 m-long st ainless steel rod and a uniquely numbered plastic tag. Photographs were captured us ing an Olympus 5060 series digital camera encased in an underwater housing. The camer a was attached to an apparatus that maintained a 50 cm distance from the substrat um. Four photographs were taken adjacent to the center of the photostation, each capturing an area of 0.25 m2. The four photos were processed with CanvasTM to create a seamless 1.0 m2 mosaic that was used for analysis (Fig. 4.2). The eight photostations were chos en as they had been photographed prior to the red tide event. Because of the small sample size, ten random 0.25 m quadrats were
106 photographed throughout each of the artificial reef sites (LB and RM) during August 2005, July 2006, and March 2007 samplings to provi de a statistical assessment of spatial differences at each of the reefs, and to a ssess whether the small number of photostations could accurately portray the community that de veloped at each of the sites. There were no significant differences among quadrats within each of the two types of artificial reefs during each sampling, and the photostations we re deemed sufficiently representative of the relatively uniform benthi c community development. Figure 4.2. Example of 1 m2 photo-mosaic from station #84 (~17 m). Four photos were combined to produce composite images that were used in point-count analyses. Substrate and biological c over attributes of the be nthic photostations were assessed using point-count an alysis (e.g., Curtis 1968; Bohns ack 1979; Carlton and Done 1995; Jaap and McField 2001; Jaap et al. 2003). One hundred random points were superimposed on each image in Coral Point Count v.3.4 (Kohler and Gill 2006), and the 5cm
107 benthic component under each point was identifi ed to provide an estimate of benthic cover (Hackett 2002). Twenty biological and substrate cate gories were included in the assessment. Important species were identi fied where possible. Six of the categories included a particular specie s or phylum (the coral, Cladocora arbuscula [LeSueur 1821] the urchin, Arbacia lixula [Linnaeus 1758], the phylum Porifera) and their respective Â“bleachedÂ” or Â“deadÂ” counterparts. Normal-appearing C. arbuscula did not display any signs of bleaching, whereas the bleached categor y includes all corals displaying partial or full bleaching. Normal-appearing Porifera included Cliona spp while the dead/diseased Porifera category refers to organisms whose position in the photostation mosaics corresponded to those of their health y counterparts in earlier mosaics ( i.e., August 2005 mosaics were compared to March 2005 mosaic s to determine locations of previously healthy animals). The remaining 14 categories consist of Leptogorgia virgulata [Lamarck 1815], Astrophyton muricatum [Lamarck 1816], ascidians, rock/rubble, unknown, and nine algal categories. Algae were divided into five distinguishable algal genera or species ( Acetabularia spp., Halimeda spp., C aulerpa mexicana, C. prolifera, and C. racemosa ) and 4 general algal classifications. Gene ral classifications include rhodophytes (e.g., Eucheuma and Gracilaria spp. ) and chlorophytes (e.g., Udotea spp. ); if identification proved impossible due to poor quality of photograph, exce ss sedimentation, etc., the algae were grouped into macroalgae a nd turf algae/cya nophyte categories. Multivariate analyses were conducted using the Primer-ETM (Clarke and Warwick 2001) package of non-parametric software appl ications, as data di splayed significant nonnormality. Point-count values were square-root transformed to draw information from across the whole assemblage (Clarke and Gr een 1988). Multivariate distances were
108 calculated using the Bray-Curtis similarity coefficient (Bray and Curtis 1957) and plotted using a non-metric multi-dimensional scali ng (MDS) ordination. The MDS finds a nonparametric monotonic relationship between dissim ilarities in the item-item matrix and the Euclidean distance between the items, and plots the location of each item in lowdimensional space. MDS ordination stress le vels <0.15 signify a useful representation (i.e., configuration closely repr esents the rank order of dissimilarities in the original triangular matrix), while stress levels > 0.20 signify a random arrangement of samples, bearing little resemblance to the original ra nks (Clarke 1993). Second level procedures (Clarke and Warwick 2001) were used to te st for significant diffe rences in benthic community structure between those samples/groups that separated spatially in the MDS. An analysis of similarity (ANOSIM) test was run to detect significant community differences among sampling times. The ANOSIM is analogous to the multivariate analysis of variance (MANOVA) but is used preferentially in this paper because the probability distribution of count s could not be normalized by any transformation due to the dominance of zero values. Data for all eight photostati ons were grouped together ( n = 8) for each of the four survey periods, as a two-way ANOSIM reve aled no differences between the benthic communities at the two types of artificial stru ctures (LB and RM) within sampling times. The similarity percentages (SIMPER) procedur e was utilized to dete ct the biological or substrate categories that contri buted significantly to changes in cover between surveys. Discriminating categories satisf y the two conditions of (1) co ntributing significantly to the average dissimilarity between time periods and (2) contributing consistently (small standard deviation) to the average dissimilarity.
109 4.3.3. Fish Communities Censuses of the fish communities were conducted in summer 2005 (June), winter 2005 (November 2005-January 2006), summ er 2006 (July-August), winter 2006 (December 2006-March 2007), and summer 2007 (June 2007) at randomly chosen LB, RM, and R sampling stations. Censuses we re conducted using a modified Bohnsack visual fish-census method (Bohnsack and Bannerot 1986; Bohnsack et al. 1994), with observersÂ’ fish identification skills evaluated pr ior to the surveys. Once in the water, the divers rotated and counted fish within a 5 m radius cylinder extending from the surface to bottom for 5 minutes. Ten surveys were c onducted at LB, RM, and R sites during each sampling period (30 total). These data are summarized in five separate GNGS reports (GNGS 2005a, 2005b, 2006a, 2006b, and 2007). Species-richness values were plotted to depict temporal trends before, during, and after the red-tide event at th e three types of ha bitat (LB, RM, and R). Abundance data were fourth-root transformed to focus atten tion on patterns within the whole community, mixing contributions from both common and rare species (Clarke and Warwick 1994). ANOSIM tests were performed within each sampling time to determine whether fish assemblages were significantly different betw een LB, RM, and R sites. The three types of habitat differed significantly from one anothe r in terms of fish-com munity structure, so subsequent analyses were performed on separated data. Multivariate tests in cluded ANOSIM and SIMPER analyses. For certain analyses, fish species were classified according to thei r predominant habitat: pelagic or demersal. For analysis purposes, those fish that are not considered demersal, but feed on benthic organisms, were included in the demersal cat egory. These distincti ons were important as
110 certain pelagic species (e.g., Haemulon aurolineatum ) are seasonally schooling species that can skew statistical analyses with abunda nce values three orders of magnitude higher during summer samplings. Separate analyses were performed with these species removed from the data set to assess the influence of other rarer species. Habitat classifications were based upon species descriptions from Robins and Ray (1986), McEachran and Fechhelm (1998), Froese and Pauly (2005) as well as the five GNGS reports. 4.4 Results 4.4.1 Study Area Characteristics Mean depth, temperature, and Secchi de pths at the sites during each sampling time are displayed in Table 4.1. Depths at the sites range d from 16.9 to 19.5 m; rugosity was greatest at the LB sites ( 1.53) and least at the R sites ( 1.13). Seawater temperatures ranged from 18.1 C in the wint er to 29.6C in the summer, w ithin the nominal values for the area (Joyce & Williams 1969). Secchi dept h measurement ranged from 7.3 m to 17.6 m, with considerable variabili ty to the measurements.
111 Table 4.1 Abiotic and habitat characterization data summary presented as mean values ( S.E). Data were recorded during each of the five fish censuses a Parameter Limestone Boulders Reef Modules Reference (LB) (RM) (R) Water Depth (m) 19.5 (1.1) 16.9 (0.2) 18.5 (0.9) Rugosity Index 1.53 (0.04) 1.36 (0.08) 1.13 (0.03) Water Temperature (C) Summer 2005 26.5 (0.5) 27.3 (0.1) 26.9 (0.2) Winter 2005 22.2 (1.0) 22.8 (0.7) 20.7 (1.2) Summer 2006 29.1 (0.3) 29.6 (0.2) 29.4 (0.4) Winter 2006 18.2 (0.7) 18.2 (0.6) 18.1 (0.4) Summer 2007 26.7 (0.8) 26.3 (0.5) 25.9 (0.5) Secchi Depth (m) Summer 2005 13.9 (0.5) 10.9 (0.4) 8.5 (0.4) Winter 2005 9.0 (0.7) 7.3 (0.3) 9.8 (0.5) Summer 2006 17.6 (2.1) 9.6 (0.3) 12.9 (1.7) Winter 2006 12.6 (0.8) 12.3 (0.4) 12.8 (0.7) Summer 2007 10.5 (0.7) 9.4 (0.7) 9.8 (0.6) a Data are adapted from: GNGS 2005a, 2005b, 2006a, 2006b, and 2007 The extended temporal and spatial scale of the 2005 Karenia brevis bloom prompted a focused sampling effort by the FWRI in August 2005. Water samples from areas west of Tampa Bay indicated that medium to high concentrations (>105 cells L-1) of K. brevis were present in both the surface waters and at depth in areas within the 30 m isobath (FWRI unpubl. data). On August 3, 2005 a cruise transect passed through areas in close proximity to the GNGS artificial reefs and the depth-correlated parameters are shown in Figure 4.3. Dissolved oxygen levels (mg L-1) decreased at depth at the offshore sites, declining from >9 mg L-1 at depths <5 m to 0.8 mg L-1 at depths > 17 m. Relative fluorescence of chlorophyll increased at offshore sites up to 1.7 g L-1. FWRI data are representative of conditions th at prevailed along areas of th e west-central Florida shelf
112 where high concentrations of K. brevis occurred during 2005. Diver observations indicate that similar hypoxic/ anoxic conditions were pres ent at depth at GNGS reefs during the August 2005 sampling time, there by negatively affecting both benthic macroinvertebrate and fish communities. Figure 4.3. Environmental parameters taken along a 24 kilometer east-west transect positioned 5-7 kilometers north of the Gulfstream Natural Gas Systems artificial reefs 4.4.2. Benthic Communities The close proximity and similar abiotic c onditions at the GNGS artificial reefs led to the development of relatively uniform epibenthic communities, despite the difference in substrate type at LB and RM sites. A two-way ANOSIM was run between the habitat types within the sampling times and confirmed that there were no significant differences in benthic community structure at the LB and RM sites. Th e benthic data (n=3 at LB sites and n=5 at RM sites) were pooled for subsequent analyses. 0 5 10 15 20 25 30 35 40 134569101112131517 Depth (m) Temp (C) Sal (PSU) DO (mg l-1) Relative Fluor Chl (ug l-1)
113 A non-metric MDS plot spatially grouped the samples according to similarities in benthic composition (Fig. 4.4). Samples groupe d relatively well into four distinct Figure 4.4. Multidimensional scaling ordination of 1 m2 benthic quadrats during the four sampling periods. The arrows depict the theo retical temporal trajectory of community response. groupings, corresponding to the four sampling times. Samples are overlain with the dendrogram similarity results (40% and 60% intervals). Arrows depict the temporal trajectory of benthic community succession. The trajectory pro ceeds in a clockwise circle with the initial Ma rch 2005 samples grouped at th e bottom of the MDS. The August 2005 samples (taken during the peak of the red tide) group farthest to the left, July 2006 samples grouped towards the top right, and March 2007 samples are
114 interspersed between the July 2006 and March 2005 samples, as the community proceeded to return to the baseline state. An ANOSIM indicates that, when analy zed in chronological order, the benthic compositions differed significantly betw een March 2005 and August 2005 (ANOSIM R =0.8, p = 0.2%), between August 2005 and Ju ly 2006 (ANOSIM R =0.8, p = 0.2%), and between August 2005 and March 2007 (ANOSIM R = 0.9, p = 0.1%). Table 4.2 lists the major components contributing to differences, as well as their averag e changes in percent cover. Percent cover and c ondition of poriferans tended to have a strong influence on temporal trends, as they c ontributed to significant commun ity differences in all three temporal pairings. Poriferans were ne gatively affected by the August 2005 red tide; percent cover of dead Porifera increased by 6.5% between March 2005 and August 2005. Cladocora arbuscula, the only scleractinian coral enum erated in the photographs, was also severely affected by the re d tide; percent co ver of bleached C.arbuscula increased by 6.1%, with an accompanying 5.1 % percent cover decrease in normal C.arbuscula between March 2005 and August 2005. During post red-tide samplings (July 2006 and March 2007), the predominant contributors to community differences were algal taxa. Turf-algae cover increased 28% by July 2006 and increased by an additional 7.0% by March 2007. Percent cover of various r hodophytes and chlorophytes increased by March 2007, with an accompanying decrease in th e rock/rubble category as the algae encroached upon the vacated sp aces. Declines in the d ead Porifera category were observed in the last two samplings, but there is a noticeable lack of recovery of the normal Porifera category. Percen t cover of normal (non-bleached) C. arbuscula increased
115 during both the July 2006 (+0.6%) and March 2007 (+8.0%) samplings, with many of the colonies regaining their symbiotic zooxanthellae. Table 4.2. Results from the SIMPER test to dete rmine discriminating benthic categories for pairs of sampling times that differed significantly (ANOSIM R 0.5, p 0.5%). Discriminating categories satisfy the condi tions of contributing significantly and consistently to the average dissimilarity. The average percent cover change of each category is shown in the last column. Sampling Times Benthic Categories Average Percent (% Dissimilarity) Cover Change March 2005 & August 2005 Dead Porifera +6.5 Bleached C. arbuscula +6.1 Normal C. arbuscula -5.1 Normal Porifera -3.3 August 2005 & July 2006 Turf Algae +28 Rock/Rubble -19 Dead Porifera -6.3 C. mexicanus +5.3 (C. arbuscula)a (+0.6) August 2005 & March 2007 Rhodophytes +10 Turf Algae +7 Chlorophytes +7 Dead Porifera -6.5 (C. arbuscula)a (+8) aCladocora arbuscula values are displayed, though they are not among the top 4 discriminating species, due to their importance as potential bioi ndicators of stress during red-tide events. 4.4.3. Fish Communities Table 4.3 provides the fish species list (71 species total) for all samplings from summer 2005 to summer 2007 at the GNGS LB, RM, and R sites. Primary habitat (demersal or pelagic) is listed for each species, and commercially important species are identified (Gulf of Mexico Fishery Manage ment Council 2004). In some cases, divers identified fish only by common names that could not be matched with species (i.e.,
116 filefish and wrasse). These were rare speci es, whose contributions to overall community assemblage are considered to be non-significant. Table 4.3. List of the 71 fish species observe d during census activites from March 2005 to March 2007 at GNGS artifi cial reef sites including limestone boulder (LB), reef module (RM), and reference s ites. Commercial importance and primary habitat are noted. Common Name Scientific Name Commercially Primary (Species Author) Important Habitat Atlantic Spadefish Chaetodipterus faber Pelagic (Broussonet, 1782) Bandtail Puffer Sphoeroides splengeri Benthic (Bloch, 1785) Barjack Caranx ruber Pelagic (Bloch, 1793) Beaugregory Stegastes leucostictus Benthic (Muller & Troschel, 1848) Belted Sandfish Serranus subligarius Benthic (Cope, 1870) Black Grouper Mycteroperca bonaci Yes Pelagic/Benthic (Poey, 1860) Black Seabass Centropristis striata Benthic (Linnaeus, 1758) Blue Angelfish Holacanthus bemudensis Benthic (Goode, 1876) Blue Goby Ptereleotris calliura Benthic (Jordan & Gilbert, 1882) Bluehead Wrasse Thalassoma bifasciatum Benthic (Bloch, 1791) Blue Runner Caranx crysos Pelagic (Mitchill, 1815) Butter Hamlet Hypoplectrus unicolor Benthic (Walbaum, 1792) Chub Khyphosus sectatrix Benthic (Linnaeus, 1758) Clown Wrasse Halichoeres maculipinna Benthic (Muller & Troschel, 1848) Cocoa Damsel Stegastes variabilis Benthic (Castelnau, 1855) Cubbyu Pareques umbrosus Benthic (Jordan & Eigenmann, 1889)
117 Emerald Parrotfish Nicolsina usta Benthic (Valenciennes, 1840) Filefish Sp.? Pelagic Florida Blenny Chasmodes saburrae Benthic Jordan & Gilbert, 1882 French Angelfish Pomacanthus paru Benthic (Bloch, 1787) French Grunt Haemulon flavolineatum Benthic (Desmarest, 1823) Gag Grouper Mycteroperca microlepis Yes Pelagic/Benthic (Goode & Bean, 1879) Goliath Grouper Epinephelus itajara Protected Pelagic/Benthic (Lichtenstein, 1822) Great Barracuda Sphyraena barracuda Pelagic (Edwards, 1771) Greater Amberjack Seriola dumerili Yes Pelagic (Risso, 1810) Grey Angelfish Pomacanthus arcuatus Benthic (Linnaeus, 1758) Grey Triggerfish Balistes capriscus Yes Benthic Gmelin, 1951 Gulf Flounder Paralichthys albigutta Benthic (Jordan & Gilbert, 1882) Hogfish Lachnolaimus maximus Yes Benthic (Walbaum, 1792) Inshore Lizardfish Synodus foetens Benthic (Linnaeus, 1766) Jackknife Fish Equetus lanceolatus Benthic (Linnaeus, 1758) Jolthead Calamus bajonado Benthic (Bloch & Schneider, 1801) Lane Snapper Lutjanus synagris Yes Benthic (Linnaeus, 1758) Leopard Toadfish Opsanus pardus Benthic (Goode & Bean,1880) Mangrove/Grey Snapper Lutjanus griseus Yes Benthic (Linnaeus, 1758) Molly Miller Scartella cristata Benthic (Linnaeus, 1758) Nurse Shark Ginglymostoma cirratum Yes Benthic (Bonnaterre, 1788) Orange Blenny Sp.? Benthic
118 Pigfish Orthopristis chrysoptera Benthic (Linnaeus, 1766) Pinfish Lagodon rhomboides Benthic (Linnaeus, 1766) Porgy Calamus sp. Benthic Porkfish Anisotremus virginicus Benthic (Linnaeus, 1758) Queen Angelfish Holacanthus ciliaris Benthic (Linnaeus, 1758) Red Grouper Epinephelus morio Yes Pelagic/Benthic (Valenciennes, 1828) Reef Butterflyfish Chaetodon sedentarius Benthic (Poey, 1860) Round Scad Decapterus punctatus Pelagic/Benthic (Cuvier, 1829) Sand Perch Diplectrum formosum Benthic (Linnaeus, 1766) Scaled Sardine Harengula jaguana Pelagic Poey, 1865 Scamp Grouper Mycteroperca phenax Yes Pelagic/Benthic (Jordan & Swain, 1884) Scrawled Cowfish Acanthostracion quadricornis Benthic (Linnaeus, 1758) Seahorse Hippocampus sp. Benthic Seaweed Blenny Parablennius marmoreus Benthic (Poey, 1876) Sergeant Major Abudefduf saxatilis Pelagic/Benthic (Linnaeus, 1758) Sheepshead Archosargus probatocephalus Benthic (Walbaum, 1792) Sheepshead Porgy Calamus penna Benthic (Valenciennes, 1830) Slippery Dick Halichoeres bivittatus Pelagic/Benthic (Bloch, 1791) Southern Flounder Paralichthys lethostigma Benthic Jordan & Gilbert, 1884 Spanish Mackerel Scomberomorus maculates Yes Pelagic (Mitchill, 1815) Spanish Sardine Sardinella aurita Pelagic Valenciennes, 1847 Spot-fin Porcupinefish Diodon hystrix Benthic Linnaeus, 1758
119 Spottail Pinfish Diplodus holbrookii Benthic (Bean, 1878) Spotted Drum Equetus punctatus Benthic (Bloch & Schneider, 1801) Squirrelfish Holocentrus adscensionis Benthic (Osbeck, 1765) Threadfin Herring Opissthonema oglinum Pelagic (Lesueur, 1818) Tomtate Haemulon aurolineatum Pelagic Cuvier, 1830 Townsend Angelfish Holacanthus townsendi Benthic (Nichols & Mowbray, 1914) White Goby Sp.? Benthic White Grunt Haemulon plumierii Benthic (Lacapede, 1801) Whitespotted Soapfish Rypticus maculates Pelagic/Benthic Holbrook, 1855 Yellow Goatfish Mulloidichthys martinicus Benthic (Cuvier, 1829) Yellowtail Snapper Ocyurus chrysurus Yes Benthic (Bloch, 1791) A two-way ANOSIM among habitats within sampling times indicated that there were significant differences in fish comm unities (diversity and abundance) among the LB, RM, and R sites. Subsequent analyses were performed within individual habitat types (n=10). Numbers of species present at the two artificial reef habitat types and the reference sites were substantially lower imme diately after the red tide event (Fig. 4.5). Prior to the red tide event, th e highest number of fish speci es was observed at LB sites, with R sites having the lowest number. Immedi ately after the red tide all sites exhibited a sharp decline in species numbers: 50% for LB sites, 65% for RM sites, and 60% for R sites. Diversity trended upward in all sites from Summer 20 06 through Summer 2007,
120 with RM and R sites reaching their origin al (Summer 2005) levels by the Summer 2007 sampling. Significant temporal changes in fish assemblages were determined by an ANOSIM test (Table 4.4). S ites and samplings that displa y significant differences are denoted by an Â“SÂ” while those that are sim ilar are deemed non-significant and denoted by an Â“NÂ”. Further analyses (SIMPER) were run on the significant samples to determine species that contributed to the dissimilar ity; average changes in individual species abundances were calcula ted (Table 4.5). Table 4.4. Matrix of significant (S) and nonsignificant (N) temporal fish-assemblage trends at LB, RM, and R sites, respectivel y. Summer 2005 (S05) sampling occurred prior to the peak of the red-tide event. Summer 2007 (S07) represents the final sampling in the focused two-year time series. S05 W05 S06 W06 S05 W05 SSS S06 SNS NSN W06 NNN NSN NNN S07 SSS SSS NNN SSS
121 Table 4.5. Results from SIMPER analys es performed on significantly different assemblages (determined by ANOSIM in Tabl e 4.4 above) to determine discriminating species and their average change (+/) in abundance between time periods. AP = Archosargus probatocephalus, SS = Serranus sub ligarius, SV = Stegastes variabilis, MP = Mycteroperca phenax, LM = Lachnolaimus maximus, CS = Chasmodes saburrae, HB = Halichoeres bivittatus, CC = Caranx cr ysos, DF= Diplectrum formosum, HP = Haemulon plumierii, CX = Calamus sp., CB = Calamus bajonado, LR = Lagodon rhomboides. Sampling Times R P (%) Discriminating Species Habitat Type (LB, RM, or R) (Average Change) Summer 2005 & Winter 2005 LB 0.7 0.1 ^AP (+2), SS (-8), SV (-2) RM 0.6 0.1 ^SS (-18), SV (-2), MP (-2) R 0.5 0.1 SS (-1), DF (+1), HP (-1) Summer 2005 & Summer 2006 LB 0.3 0.1 MP (-2), CC (+12), SS (-7) R 0.4 0.1 LR (+46), HP (+1), SS (-2) Summer 2005 & Summer 2007 LB 0.3 0.1 ^CS (+5), CC (+13), SS (-2) RM 0.5 0.1 ^HB (+9), CS (+6), DF (-10) R 0.6 0.1 HB (+9), CX (+3), HP (+1) Winter 2005 & Summer 2006 RM 0.5 0.1 ^SV (+3), SS (+6), LM (+1) Winter 2005 & Winter 2006 RM 0.5 0.1 SV (+2), SS (+7), LM (+1) Winter 2005 & Summer 2007 LB 0.7 0.1 ^SS (+6), CS (+5), SV (+3) RM 0.9 0.1 ^HB (+9), CS (+7), SV (+3) R 0.6 0.1 HB (+9), SS (+6), CX (+1) Winter 2006 & Summer 2007 LB 0.5 0.1 ^SS (+6), CS (+5), LM (-3) RM 0.6 0.1 ^HB (+9), CS (+6), MP (+1) R 0.4 0.1 HB (+9), CX (+4), CB (-10) ^Indicates the placement of Haemulon aurolineatum when included in the analyses.
122 Serranus subligarius, Stegastes variabilis, and Mycteroperca phenax abundances all declined during or immedi ately after the red tide even t at the LB and RM sites, contributing significantly to differences between the Summer 2005 and Winter 2005 samplings. Archosargus probatocephalus abundances increased at the LB sites, indicating that although this sp ecies may have initially evacu ated the area during the red tide, it was among the first to return to the LB sites immediately after its dissipation. Diplectrum formosum displayed a similar trend at R site s, as it was the only species to increase in number from Summer 2005 to Winter 2005. Fish assemblages were significantly different only at LB sites between Winter 2005 and Summer 2006, with increased abundances in the three primary discriminating species ( S. variabilis, S. subligarius, and Lachnolaimus maximus ). Summer 2006 and Wint er 2006 samplings had similar fish assemblages, followed by a shar p increase in most sp eciesÂ’ abundances by the Summer 2007 sampling. Only L. maximus and Calamus bajonado abundances were lower at LB and R sites, respectively. The remaining four pairings in Tabl e 4.5 detail the changes in abundances between Summer and Winter samplings, with various changes in fish assemblages occurring among sites. The fifth pairing in the table compares Summer 2005 data to Summer 2007 data to assess ove rall changes over the two-year sampling period. Both the LB and RM sites saw overall increases in Chasmodes saburrae abundances. LB sites also experienced an increase in Caranx crysos abundances whereas RM sites experienced an increase in Halichoeres bivittatus. Both sites saw declines in certain discriminating species as Serranus subligarius and Diplectrum formosum abundances decreased at LB and RM sites, respectively. All discriminating species ( Halichoeres bivittatus, Calamus
123 sp., and Haemulon plumierii ) increased in abundance over th e two-year sampling time at the R sites. Attempts were made to use a non-metric MDS ordination to spatially examine the fish abundance data but, unlike the bent hic data where rank dissimilarities among temporal groups were significantly higher th an those within samples in a group (as determined by an ANOSIM), fish data did not separate into distinct temporal groups. High stress values in the MD S ordinations (>0.2) indicated that interpretations based on the ordination are not useful as the samples are essentially randomly placed, bearing little resemblance to the original similarity ra nks in the triangular matrix (Clarke 1993). Ordinations are not displayed in this paper. 4.5 Discussion 4.5.1. Benthic Communities Disturbance is defined by Connell (1997) as an event that damages or kills residents at a given site. Dist urbances can be either acute (short-term) or chronic (longterm) with direct effects on the physical/b iological environment (e.g., a storm alters community topography) or indire ct effects (e.g., a disease k ills corals and indirectly reduces physical/biological comple xity of the community). Re d tide events are classified as acute, indirect, episodic disturbances that have the ability, through exposure to brevetoxin or hypoxic/anoxic conditions, to alter community structure by negatively impacting the benthic, demersal, and water-column communities. The spatial scale of the disturbance a ffects ecosystem resilience (Sousa 1985), along with factors such as the frequency and duration of the disturbance (Nystrom et al. 2000). Estimates indicate that approximately 5,600 km2 of benthic communities may
124 have been affected by the 2005 red tide and the anoxic/hypoxic conditions (FWRI unpub. data). Although natural distur bances such as red tides can be detrimental to individuals and communities at large spatial scales (10Â–1000 km2), the emigration/extirpation of organisms vacates substratum, making it availa ble at various temporal and spatial scales (Connell 1978). This provides opportunity for renewal, development, and succession of the community (Holling 1996). Recovery time s after a disturbance can vary greatly among communities and within populations depending on levels of adult dispersal/encroachment and competition, larval supply, selective forces acting on the planktonic larval stages, selectivity of la rvae for different types of substrate, and predation effects on la rvae (Thorson 1950, 1955, 1957, 1966). An important conclusion in SmithÂ’s (1975) original qua litative study on the impact of a severe red-tide event on west Florida shelf communities was that major events may result in the near-extirpation of liv ebottom biotas and that recovery rates may be on the order of years to decades. However, the data presented in this focused two-year time period indicate that communities may r ecover more quickly than originally predicted, particularly on artificial reef s. The data indicate that while benthic communities were significantly impacted by th e red-tide event and related anoxic bottom conditions, the two-year recovery trajectory is towards a prered tide community structure (Fig. 4.4). The initial, pioneer ing species that recruited to the sites included dense mats of cyanophytes and other small turf algae. The cyanophytes and turf algae became subdominant to recolonizing filame ntous algae (rhodophytes such as Gracilaria and Eucheuma spp. ), chlorophytes ( Caulerpa mexicana, C. racemosa, Halimeda spp., Udotea spp. ), and other macroalgae (phaeophytes and others) by March 2007.
125 In addition to recruiting alga l communities in the oneto two-year period after the red tide, many Cladocora arbuscula colonies survived the stre ss, despite having bleached during the height of the red tide and anoxic conditions. Large, healthy colonies >10 cm in diameter were observed in July 2006 and March 2007. Because growth rates of C. arbuscula are on the order of 5 cm per year (W.C. Jaap, personal comm.), larger colonies must have survived the 2005 red tide. These fi ndings are consistent with those of Rice and Hunter (1992), who found that C. arbuscula are among the scleractinian corals most resistant to environmental stress. The per cent of rubble/bare subs trate increased from July 2006 to March 2007 as the opportunistic algal species became sub-dominant and herbivorous fish populations began to rec over. Small numbers (representing <3% of benthic cover) of echinoderms (primarily Arbacia lixula ) were also present during the July 2006 and March 2007 sampling; these echi noderms could graze on algae and expose substrate. This evidence suggests that r ecovery from a major red ride, including hypoxia, can occur on the order of y ears, rather than decades. Results from the Simon and Dauer (1977) study in Old Tampa Bay, Tampa, FL, indicated that, although a marked loss of benthic infaunal in vertebrates did occur as a result of the 1971 red tide and reported anoxia, the fauna made a rapid recovery in terms of species numbers and composition within two years. These recovery rates are similar to those I observed, and both sets of data indicate that communities may recover much faster than predicted by Sm ith (1975), although certain populations may take much longer to fully recover. Colonization rates of certain taxa are rapid (e.g ., polychaetes in infaunal communities and algal species in ep ifaunal communities), while other taxa appear to have longer recovery periods, greatly influenced by the time of year when a
126 perturbation in the community occurs (e.g. mollusks, amphipods, and other crustacea in infaunal communities, as well as Porifera a nd echinoderms in epifaunal communities). I agree with the assertions by Simon and Daue r (1977) that benthic community analyses are essential when assessing the effects of disturbances (anthropogenic and natural), as opposed to single taxon studies. The variety of colonization rates suggests that certain taxa may be considered Â“rapid response and recoveryÂ” organisms (various algae, polychaetes), while other taxa might be more useful in determining whether a community has reached an Â“equilibriumÂ” level of species (mollusks and echinoderms). The benthic-community data presented here are limited in spatial scale and are focused only on artificial reef structures Benthic-community dynamics could be very different at natural livebottom/rocky-le dge communities in the Gulf of Mexico Natural livebottom communities in the eastern Gulf of Mexico have much lower relief but more diverse coral assemblages (including Oculina diffusa, Solenastrea hyades, Siderastrea spp., Stephanocoenia intersepta, and others) than the artificial reefs I studied. Comparative responses of the artificial reefs and livebo ttom ledge communities will define whether there are differe nces between the two types of habitat and provide insight into the efficacy of artificial reefs as mitigation structures. Natural livebottom areas and comparative processes will be the focus of future publications. 4.5.2. Fish Communities The mobile nature of most fish species (p articularly migratory or pelagic species) allows them to respond quickly to acute distur bances such as red tide events. The patchy nature of most red-tide blooms may provide areas of refuge am idst the anoxic/toxic conditions, meaning that there are four basic responses of fish species to a red-tide
127 disturbance: (1) they may permanently relocate (emigrate) to another area not affected by the toxin/anoxia, (2) they may remain in an area affected by the red tide, where they either survive the bloom conditions or they pe rish, (3) they may temporarily evacuate an unsatisfactory area, but return again upon bloom dissipation, and (4) new species may immigrate in response to the presence of newly vacated habitat in the area or to escape the encroaching red tide bloom as it is advected along the shelf. The first three responses likely accounted for the significant reduction in fish species richness (Fig. 4.5) observed after the 2005 red tide. Below I discuss species that displayed the f our responses outlined above.
128 0 5 10 15 20 25 30 35 40 45 50S05W05S06W06S07Species Richness LB RM Ref Figure 4.5. Temporal changes in total fish sp ecies richness at the LB, RM, and Reference sites. The Summer 2005 (S05) census was conduc ted prior to the red-tide event; Winter 2005 (W05) data were collected during a nd immediately after the event. Eight species were recorded during the Summer 2005 sampling, which preceded the red-tide event, but were not obs erved in any subsequent samplings: Khyphosus sectatrix Holacanthus ciliaris Harengula jaguana Acanthostracion quadricornis Abudefduf saxatilis Scomberomorus maculatus Opissthonema oglinum and Ocyurus chrysurus Three of these ( H. jaguana S. maculatus and O. oglinum ) are pelagic species that may have evacuated the area during the sampling times and simply had not returned to the sites within th e study period (response #1). The re maining five species are semisedentary demersal species that occupy a pa rticular ledge for extended periods, if not their entire life. The absence of adults or juveniles suggests that extirpation from the
129 area, rather than emigration, has occurred (response #2, mortality). The failure of these five demersal species to recolonize may be a result of their low fecundity, lack of larval supply, high planktonic mortality, lack of se ttlement in the area, low competitive success, or any combination of the above. Five species were observed at all site s during all sampling times, although their abundances varied greatly (response #2, survival): Serranus subligarius Balistes capriscus Diplectrum formosum Haemulon aurolineatum and Haemulon plumierii These species survived the red tide as re mnant populations or returned soon after its dissipation as they were observed during th e pre-event sampling (Summer 2005) as well as all subsequent samplings. Other surviv ing remnant populations at two out of three sites include Lachnolaimus maximus (LB and RM), Lutjanus griseus (LB and RM), Archosargus probatocephalus (LB and RM), Synodus foetens (LB and R), and Calamus bajonado (LB and R). Two of the remnant species ( L. maximus and L. griseus ) are mobile, commercially important species. Ar tificial reef sites a ppear to have been effective in retaining or recruiting th ese species after the red-tide event. Two other commercially important species ( Epinephelus morio and Mycteroperca phenax ) displayed response #3, as they were present during the Summer 2005 samplings, absent during the Winter 2005 sampling, but were again present in subsequent samplings at all sites. These species may have moved offshore to escape the detrimental red tide conditions, but then returned to utilize the artificia l reef habitat. Other species that displayed this response were Chasmodes saburrae, Rypticus maculates, Stegastes variabilis, and Sardinellla aurita. Surprisingly, all of these species, except S. aurita are classified as demersal, reef-associated and w ould not be expected to move from the reefs
130 during unfavorable conditions, so they may have been hidden within the reef habitat and escaped notice during the fish census. All five species are highly resilient with population doubling times < 15 months (Froes e and Pauly 2005), so populations could be expected to recover quickly after acute disturbances. Six species were observed regularly afte r the red tide, but were not recorded during the initial Summer 2005 sampling (response #4): Holacanthus bemudensis Thalassoma bifasciatum Scartella cristata Decapterus punctatus Pareques umbroses, and Diplodus holbrookii This suggests that they ar e opportunistic species with the ability to colonize new niches opened due to the emigration/extirpation of other species. Reproductive characteristics, such as protogyny and group-spawning in T. bifasciatum up-current of settling areas (Warner 1984), coul d make them successful colonizers after a disturbance, provided that suitable food sources a nd habitat are available. Fish abundances and community composition differed significantly between the artificial reefs, with a small number of disc riminating species consis tently contributing to the majority of temporal dissimilarities (Table 4.5). Discriminating species were characteristically highly resilient sp ecies with population doubling times 18 months (Froese and Pauly 2005). Adults that surviv ed the red tide and relocated to other livebottom areas produced a stea dy supply of planktotrophic larvae that found favorable conditions at the artificial sites and, less abundantly, at reference sites. LB and RM sites were generally more successful in retaining or recruiting commercial fish species during and after the red tide than the Reference sites. This may be due to the higher rugosity at the artificial sites, which in turn provides greater diversity of shelter and feeding sites (Bel l and Galzin 1984). Observati ons indicate that structures
131 placed in the Gulf of Mexico are effective in retaining/recruiting commercial species. Further studies including size -distribution measurements could provide insight into whether artificial structures are contributi ng to overall fisherie s biomass or simply attracting fish that are already present in this area of the Gulf of Mexico. 4.5.3. Red Tides as a Community Structuring Force The data presented here contribute to a quantitative database of ecological impacts of red tides and associated hypoxi c/anoxic events on West Florida Shelf communities. Smith (1979) proposed that ea stern Gulf of Mexico reef-fish communities develop according to predictabl e, rather than chance processe s. In this view, ultimate stability in species richness and composition represents the attainment of a Â“climaxÂ” community, as opposed to a dynamic species equilibrium predicted by MacArthur and Wilson (1963). Smith attributed the deve lopment of a climax community to the inhospitable nature of the Gulf of Mexico which reduces the eff ective species pool of colonists. Hardy species (or species that produce hardy planktotrophi c larvae) recruit (or settle) during the early stages of coloni zation and are difficult to displace. These characteristics, combined with observations that benthic communities in the Gulf of Mexico are not isolated Â“isl ands,Â” may make it difficult to apply the MacArthur-Wilson species equilibrium model to either benthi c or fish communities along the inner West Florida Shelf. My benthic data agree with SmithÂ’s assertions, as communities progressed towards a pre-red tide state with few cha nges in species composition. Successional stages appear to follow a trajectory towards the pre-red tide state, corroborating SmithÂ’s application of the intermediate disturba nce hypothesis (Connell 1978). However, I
132 choose to forgo use of the term and concep t of a Â“climax community.Â” The proposed episodic occurrence of red tides in conjunction with other st ochastic factors such as fluctuating sea temperatures, tu rbidity, and hurricanes, likely prevents the development of complex climax communities. Instead, the te ndency to recruit equi valent species and revert to the pre-red tide state may be an intermediate stage in a prolonged multi-staged succession that never reaches a Â“dynamic equilibriumÂ” as proposed by MacArthur and Wilson (1963). Should the frequency and severi ty of disturbances decrease, different community structures may develop. Red tides in the Gulf of Mexico have been and will continue to be important in structuring epib enthic and fish communities. Mitigation for red tides should therefore focus on the quick restoration of communities through regulation of fisheries and placement of more artificial structures, and not on the process of eliminating the K. brevis bloom, which is a fundamental ecological process in the eastern Gulf of Mexico.
133 5. Enhancement of Natural Ledge Substrat e Via Deployment of Artificial Reefs Along the West Florida Shelf 5.1 Introduction A discussion on artificial r eefs often incites vi gorous debate, with the core of the argument focused on the well-rehearsed Â“at traction versus productionÂ” argument (Bohnsack et al. 1997; Pickering and Whitmar sh 1997). On the one hand, opponents of artificial reefs have come to regard them with alarm, considering them mere fish aggregating devices (FADs) that concen trate fish populations and render them increasingly susceptible to exploitation by fishermen. On the other hand, proponents view artificial reefs as impor tant habitat and recruitment-en hancement tools, arguing that the substrate provided by appropria tely-placed structures attracts larval recruits that might not otherwise find appropriate substratum. It is clear that the attr action/production debate is central to the issue of arti ficial reef deployment and it mu st be satisfactorily addressed by local or regional scientists and managers before extensive deployment of artificial reefs can be considered as part of a re storation, mitigation, or conservation plan. In Chapter 3, I evaluated a specific set of WFS artificial reefs deployed with the goal of mimicking natural ledge habitat. No w I shall expand on the comparisons between artificial reefs and natural ledges, using data from Chapters 2, 3 and 4. I will address the potential for future use of artificial reefs along the WFS, including their contribution to the resolution of the attraction/production debate.
134 5.2 Defining Current Artificial Reef Data Issues Productivity, as it pertains to artificial reefs, relies on the assumption that artificial reefs provide additional critical habitats that increase the environmental carrying capacity and thereby the abundance and bi omass of reef biota (Polovin a 1994; Bortone et al. 1994). While this definition encompasses all Â“reef biotaÂ” without solely focusing on fish assemblages, most papers that discuss artifi cial reef usefulness/efficacy discuss only the associated fish assemblages, with only mi nor mention of benthic communities (Randall 1963; Beets 1989; Bohnsack 1989; Beets and Hixon 1994; Carr and Hixon 1997; Rilov and Benayahu 2000, among others). This has focused thinking among resource managers and scientists that artificial reefs are prim arily deployed to restor e, protect, enhance, concentrate, or aggregate (depending on whether you are an opponent or proponent) fish populations, and only fish populations. Accordingly, most artificial reef opponents cite the lack of definitive data from artificial reef fish popul ations in their arguments against th e use of artificial reefs. Topics that lack Â“definitiveÂ” data include: (1) disc erning whether fishes that settle on or are attracted to artificial r eefs would have found suitable substrate elsewhere, (2) understanding whether fish survival and growth ra tes are higher at arti ficial reefs than in natural habitat, (3) determining whether fo raging success and food web efficiency is improved by artificial reefs, and (4) knowi ng whether other habitat was vacated by fish moving to artificial substrat e (Bohnsack et al. 1994). B ohnsack (1989) proposed that proof of artificial reefs in creasing production would require direct evidence such as an increased total regional catc h or standing stock in some proportion to the amount of material deployed, while accounting for fish ing effort, recruitment from surrounding
135 areas, and changes in year class strengt h. A comprehensive study encompassing the above parameters has not yet been attempted (a nd may be virtually impossible) and so the Â“lack of definitive dataÂ” argument continues to be employed to discourage the use of artificial reefs. To summarize the two main problems with ar tificial reef studies to date: (1) they overemphasize the contributions and importan ce of fish populations to reef biomass calculations and underemphasize benthic comm unity contributions and (2) they do not provide definitive data on the preferences or movements of individual fish. I shall address the first issue using data from the WF S artificial reefs, with particular emphasis on the development of a robust epibenthic community and bottom-up production effects. The second problem is a bit more esoteric a nd difficult to address using data from the WFS artificial reefs, as my data can not be classified as Â“definitiveÂ”. I will instead present reasons why this lin e of thinking should be dismissed in many situations, particularly in current coasta l restoration and habitat conser vation projects al ong the WFS. The results and arguments presented here can be debated and evaluate d in other regions, where applicable, as certain areas may be similar to the WFS conditions while others differ drastically. 5.3 Problem #1: Rationale for Including Benthic Communities in Production Calculations 5.3.1. Artificial Reef Contributions Pickering and Whitmarsh (1997) revealed interesting insights into the services that artificial reefs provide outsi de the usual realm of fishes. They state that the artificial reef (when properly construc ted and deployed) potentially provides: (1) substrata for benthic fauna and, thereby, additional food a nd increased feeding efficiency, (2) shelter
136 from predation or tidal currents (Collins et al. 1991; Spanier 1996) and (3) recruitment habitat for individuals that would otherwise be lost fr om the population (fishes and benthic invertebrates). Thes e three topics will be discus sed as they pertain to the Gulfstream Natural Gas Systems (GNGS) ar tificial reefs deploye d along the shallow inner WFS, west of Egmont Key (see Chapter 3 for background on construction, deployment, monitoring, and evaluation of the reefs). 5.3.2. Substrate, Benthic Fauna, and Increased Food Availability The deployment of artificial reefs in the eastern GOM increases the biomass of sessile benthic invertebrates and macroalgae substantially when compared to surrounding quartz-dominated sand ridges and associated in faunal assemblages. Epibenthos include corals ( Cladocora arbuscula and Siderastrea spp.), poriferans, echinoderms, ascidians, and mollusks. Bubbleplots displaying th e assessed categories at the GNGS artificial reefs are shown in Figures 5.1 to 5.3, and repr esent varying abundances of the relatively diverse epibenthic community over the four sampling times. Samplings prior to a redtide event (March 2005) displayed coral cove r of up to 21% in certain quadrats. Similarly, the March 2007 samples had coral co ver up to 24%. Poriferans and other living fauna (echinoderms, ascidians) also cont ributed greatly to perc ent cover values. Macroalgal percent cover data are shown in Figure 5.4 and are negatively correlated to bare substrate cover (Fig. 5.4) in the same manner as observed along natural livebottom ledges (Fig. 5.5). It is difficult to discern whethe r seasonal macroalgal trends at the artificial reefs mimic those of natural livebottoms as the samplings are less highly resolved.
137 Figure 5.1. Percent cover of coral at GNGS artificial reef s. Data are shown for individual 1 m2 photo-quadrats captured during each of the 4 sampling times (March 2005, August 2005, June 2006, and March 2007). Figure 5.2. Percent cover of porif erans at the GNGS artificial re efs; source of data as in Fig. 5.1.
138 Figure 5.3. Percent cover of ot her living fauna (primarily ec hinoderms and ascidians) at GNGS artificial reefs; source of data as in Fig. 5.1. A transition from an infaunal communitiy to an epifaunal community generally increases the areaÂ’s biomass as demonstrated by Foster et al. (1994). They compared biomass of infauna prior to artificial reef emplacement in Delaware Bay to epifaunal biomass after. They found that biomass va lues had increased by 148 to 895 fold in the shift from the infaunal to epifaunal communities. These enhanced biomass figures reflect the expanding available surface area for benthi c biota. Trapping of plankton and other resources by the structure, increased sedime ntation of suspended particles, reef waste products, and detached organisms may also cont ribute to increased biomass (Foster et al. 1994). Sessile invertebrates and al gae serve to attract fish (Dudley and Anderson 1982; Wallace and Benke 1984) and, as gut content surveys have demonstrated, provide an essential food source (Johnson et al. 1994).
139 Figure 5.4. (A) Macroalgal percent cover at GNGS artificial reefs wh ich varies inversely with (B) bare substrate cover. Sa mplings (March 2005, August 2005, June 2006, and March 2007) do not displa y seasonal trends. B A
140 Figure 5.5. (A) Macroalgal percent cover at natural ledges (FW=FWRI1 and M=Mastedon Tabletop) which vary inversel y with (B) bare substrate cover. 5.3.3. Provision of Shelter The GNGS artificial reefs, by virtue of their design (eit her reef modules with the cavity cut through the limestone matrix or ar rangement of limestone boulders), provide shelter for macroinvertebrates and fish species. Their de sign, in conjunction with the development of secondary substrate which alters reef topogra phy and heterogeneity, provides essential shelter for ju veniles and adult organisms seek ing refuge from predation, wave forces, and sediment movement (Hi xon and Brostoff 1985; Relini et al. 1994). A B
141 However, growth on artificial reefs de pends on the length of immersion. The GNGS artificial reefs were depl oyed in 2001, allowing for only 46 years of growth at the time of data collection. Therefore, the epib enthos on the artificial reefs was much less diverse (although percent cover values of biota were sim ilar) than on the natural livebottom ledges. Only three species of co rals were observed at the artificial reefs ( Cladocora arbuscula, Phyllangia americana, and Siderastrea radians ) as compared to the 6 species observed at natural ledges (s ee Chapter 2). The same held true for macroalgal species (personal observation), as the same types of fleshy macroalgae ( Sargassum spp.) and Halimeda spp. were observed on a ll artificial reefs. Shelter was truly provided for fish species as evidenced by the 71 species that were observed at the artificial reefs as comp ared to 47 species at the natural ledges. A number of the species observed at the artifici al reefs were typical of more tropical regions (i.e., Holacanthus ciliaris Pomacanthus paru and Thalassoma bifasciatum ), and the more southerly location (50 km south) and slight ly warmer waters coul d be a factor in the higher species richness observed at the artifici al reefs. Abundances of most fish species were higher at the artificial re efs; future work should make an effort to assess size-class distributions to more effectively contribu te data to resolve the production versus attraction debate. 5.3.4. Recruitment Habitat Larva numbers often far exceed the numbers able to settle on a reef (Sale 1980) which, with food eliminated as a direct factor (Shulman 1984), leaves habitat as the likely limiting factor for reef populat ions. According to Collard and DÂ’Assaro (1973) and Lyons and Collard (1974), the availability of suitable s ubstrate is the single most
142 important recruitment/community-structuring factor in offshore areas along the WFS where abiotic parameters (temperature and sali nity ranges) are less variable as compared to nearshore areas. The shelter provided by a hab itat type is critical for settlement and the reduction of predation mortality among newl y settled juveniles (Shulman 1984; Doherty and Sale 1986). This pertains to both macr oinvertebrates (e.g., cora ls) and fish species. Valuable commercial fish species, including Mycteroperca microlepis utilize structures provided by scarped hardbottom ledges during a number of their life stages. It is plausible, and even probable, that placement of more structures like the GNGS artificial reefs, which mimic scarped ledges, would enha nce juvenile and smalle r-adult survival of commercially important fish species (see Chapte r 3 data on commercial fish preference of artificial reef versus reference habitat), as well as invertebra tes. In addition, Chapter 4 of this dissertation discusses the recruitment of juvenile corals to available substrate along natural ledges, where bare limestone s ubstrate provides optimum settling conditions. Figure 5.6 depicts the trends in juvenile co rals along the natura l ledges over the samplings, and reveals that continuous recru itment may be occurring, as long as substrate is available. Placement of mo re artificial-reef structures along the WFS could enhance fish and other epibenthic recruitment.
143 Figure 5.6. Bubbleplot depicting the averag e (n=3) percentage of photo-transects containing at least one juve nile (<2 cm) coral over the 22-month sampling period. 5.4 Problem #2: Rationale for Dismissing Â“Lack of Definitive DataÂ” Argument Against Artificial Reef Use Along the West Florida Shelf Artificial reefs are by no means universal tools that should be deployed in all marine restoration or conservation projects. Thorough analyses of biotic and abiotic parameters must be conducted in an area before artificial reefs can be considered as one option in a suite of alternatives. Bohnsack (1989) pointed out a num ber of factors that should be considered contra -indicative to artific ial reef deployment. The attraction hypothesis is likely to hold for locations wher e natural reef habitat is abundant, fishing mortality is high, recruitment is limited, and mo st species are pelagic, highly mobile, and non-reef dependent. Artificial r eefs would be ineffective, an d even deleterious, in these areas. Increased production is li kely at locations isolated fr om natural reefs, with low fishing pressure, and dominated by habitat-lim ited, demersal, philopatric, territorial, and obligatory reef species (Bohnsack 1989). If su fficient data (abiot ic and biotic) are available from an area and the data indicate that production, not attraction, processes will
144 prevail, then artificial reefs should be cons idered in restoration/conservation plans. Unfortunately, Â“sufficientÂ” data are not the same as Â“definitiveÂ” data. However, as I will explain below, there are times when logical a nd rational decisions can be made to restore, enhance, or conserve an area without possessing truly definitive data. The WFS, with its expansive quartz-domi nated sand ridges, intermittent limestone outcrops, and associated livebottom assemblages, is a perfect candidate for artificial reef construction and deployment. The patchy dist ribution of natural livebottom habitats and assemblages could be enhanced by deployment of low-relief, limest one structures in areas where a thin veneer of sand overlies limestone bedrock. Optimal placement would be between, but not close to, natural ledge substrate and or iented in a northwest to southeast (ledge-parallel) directi on in accordance with BohnsackÂ’s (1989) recommendations. Bombace et al. (1994) further confirmed the importance of adhering to this recommendation through their work with artific ial reefs in the Adriatic Sea. Catches at reefs deployed far from natu ral reefs showed a gradual in crease in fish abundance, species richness (both mean and total) a nd diversity. Evidence for this was the appearance and/or the increase in catches of some hard-substrate species of fish and mollusks which were rare or completely absent in the original sand-plain habitat. The increase of these species seemed to be direc tly correlated to the reef dimensions in terms of volume of immersed materials and of area covered. The Bombace et al. (1994) results indicate that the spatial s cale of artificial reef pla cement along the WFS must be sufficient to yield the desired effect of increased productiv ity. Calculation and modeling
145 of optimal spatial scales are beyond the scope of this paper but shoul d be considered in future work. Particular attention should also be paid to ensuring that the ar tificial reefs are not only located at optimal distances from natural substrates, but also c onstructed in a manner that mimics the natural substrates. Carr a nd Hixon (1997) compared fish assemblages at natural and artificial reefs and found that artificial reefs w ith structural complexity and other abiotic and biotic features similar to those of natural reefs would best mitigate inkind losses of reef fish populat ions and assemblages from na tural reefs. The GNGS reefs are good examples of sound construction and de ployment as they effectively mimicked natural livebottom assemblages along the sh allow inner WFS. Future deployments should evaluate whether artifi cial reefs would be more effective oriented in a ledgeparallel (northwest to southeas t direction) or ledge-perpendicu lar (east to west direction). A series of ledge-parallel reefs could provi de stepping stones and areas of refuge for mobile species during a red-tide/hypoxic even t. Deeper areas in the eastern GOM were populated by fish during the shallow-water hypoxia of 2005, and it is plausible that the placement of artificial reefs could enhance evacu ation and survival in the future. Ledgeperpendicular set-ups could provide a conti nuous evacuation route for mobile species and direct their movement back into shallow wate rs after dissipation of the red tide/hypoxia. Again, economic and ecological models woul d be helpful in determining optimum orientations of artificial reefs. Fishing pressure along the WFS is high. Many commercially and recreationallytargeted finfishes, includi ng those of the valuable Groupe r/Snapper complex, inhabit the area. For artificial reefs to be successful along the WFS, th ey must be protected from
146 fishing, at the very least during the early stages of recruitment, much like the GNGS reefs. When the original plans for the GNGS pipe line construction and mitigation activities were released, the route for the pipelin e was published in navigation charts, but coordinates for the artificial reefs remained unpublicized. Now, after a few years of deployment, many fishermen have learned the locations of the reefs and have begun to target them but not in sufficient numbers to alter fish abundances Pitcher and Seaman (2000) take this recommendation one step furthe r and state unequivocal ly that artificial reefs should be protected as no-take areas. Variations to this theme could include opening a small number of reefs to license d fishing so that local fishermen would understand the effects and assist in monitoring. It is essential that fishermen are educated on the uses of artificial reefs for productio n/enhancement purposes as they will most likely reap the benefits in the future, but only if the reefs ar e left alone during the initial community-development phases. The deployment of GNGS artificial r eefs led to the development (through both initial attraction and subsequent production) of a thriving re ef-like habitat. Although the benthic assemblage was less diverse than natural substrate assemblages, the fish assemblage was much more diverse, as a number of tropicals and commerciallyimportant species were frequently counted in the area. The ma jority of the fish species are demersal, reef-dependent species that provide bottom-up support for the pelagics and mobile species that frequent the areas. Once again, the proper ties of the WFS are amenable to artificial-reef deploymen t, consistent with BohnsackÂ’s (1989) recommendations.
147 5.5. Conclusion The state of the EarthÂ’s aquatic ecosystem s is in turmoil. Synergistic impacts including overfishing, pollution, ocean acidifi cation, warming, habitat destruction, and introduction of new species are transformi ng once complex and productive systems such as coral reefs into monotonous level bottom with lim ited ecological value (Jackson 2008). Action needs to be taken now to boost resilien cy of all reef assemblages, as marginal environmental conditions for reef distribution become more widespread (Guinnotte et al. 2003). What role could artifi cial reefs play in future mitigation, restoration, and conservation activities? Pitche r and Seaman (2000) suggest th at protected artificial reefs have a role to play as hedges against extinc tion. Artificial reefs al ready sustain regional commercial and local artisanal fishing in some areas (Pitcher and Seaman 2000) and their expanded use could be employed to enhance fi sh stocks and benthi c production, restore critical habitats, and provide refugia from wh ich recolonization can take place. This is not to imply that artificial reefs should be used in every restor ation or conservation program (for reasons stated in section 5.4 above). But in areas where abiotic and biotic parameters appear conducive to deployment, resource managers and scientists should not hesitate to construct and deploy artificial reefs to meet their production goals. There have been recent efforts to expand offshore aquacu lture along the WFS, and while caged structures may effectively grow fish, they are not l ong-term, sustainable solutions. The impacts of aquaculture fac ilities on benthic communities can be very detrimental as organic matter concentrations are elevated and the potential for benthic mortality via sedimentation and hypoxia/anoxia development is high. Instead of the shortterm investment in large offshore aquacultu re infrastructure, resource managers and
148 fisheries scientists should consider a bottom-up enhancement of fish stocks via deployment of low-relief, natu ral substrate structures. Pitcher and Seaman (2000) emphasize, and I concur, that for artificial reefs to produce maximum benefits, they must be afford ed some type of early-stage protection in the form of designation as a Marine Prot ected Area (MPA) or no-take zone. The protection would allow a complex community to recruit and establish, providing major enhancement to fishery catch. There are a variety of ways to go about designating artificial reefs as MPAs, but cons tituents (local stakeholders) mu st be part of the process. Stakeholders should be educat ed about the utilit y of artificial reefs and perhaps given access and fishing rights at certain reefs, while self-enforcing notake zones at other reefs. Although the task of enlisting the support of local stakeholde rs may seem daunting, there have been cases where unexpected support fo r no-take areas has b een expressed. The task is difficult but by no means impossible and may even prove to be enjoyable once a rapport with local WFS stakehol ders has been established. Actions need to be taken as soon as possible to save existing reefs and livebottoms. These days, we should not hesita te to employ methods to restore and protect todayÂ’s depleted ecosystems even without defi nitive scientific eval uation (Clark 1996). The quest for robust scientific data should absolutely continue but it should not preclude restorative actions, or else we risk lo sing these valuable ecosystems forever.
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About the Author Jennifer Maria Dupont was born and rais ed in New York. She earned her BachelorÂ’s degree, Magna cum laude, in Biology with a French minor from the University of Richmond in 2003. While at the University of Richmond, she spent a semester studying and conducti ng research at the Duke Un iversity Marine Laboratory and Bermuda Biological Stati on for Research, where she became an Advanced Open Water Diver and decided to pursue a gra duate degree in Marine Science. She started her work at the University of South Florida College of Marine Science Reef Indicators Lab in August 2004, under the di rection of Dr. Pamela Hallock Muller. While at USF, she worked as an environm ental consultant and conducted surveys along natural gas pipeline routes and ship groundings She also worked as a science mentor and divemaster with the SCUBAnauts, Internationa l. She was awarded her Ph.D. in Marine Science in 2009.