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Piagentini, Nejma Danielle.
The science and policy that compels the wetland mitigation of phosphate-mined lands
h [electronic resource] /
by Nejma Danielle Piagentini.
[Tampa, Fla] :
b University of South Florida,
ABSTRACT: The State of Florida ranks fifth in the world's production of phosphate. The phosphate industry relies on surface mining to withdraw the phosphate ore, and this process can devastate the natural environment. One of the most impacted natural resources is wetlands. Federal laws permit the legal destruction of wetlands providing their loss is compensated by the mitigation (i.e., the restoration, creation, or enhancement) of other wetlands, but the complexity of wetland ecosystems makes the mitigation process difficult. One of the goals of this thesis is to review the established Federal, State and non-regulatory guidelines for the development and maintenance of mitigated wetlands, evaluate their efficacy and present some underlying reasons for successful versus unsuccessful mitigation projects.The environmental repercussions of phosphate mining are not only pertinent to Florida or the United States. Wetland mitigation has become a global issue.^ ^Laws and programs that facilitate specific countries do not benefit wetland ecosystems on a landscape level. It is important to remain cognizant of the ramifications of wetland destruction and avoid piecemeal solutions to a wide-spread problem. Thus, my second objective is to investigate the progress and status of international wetland preservation. I will examine how different countries and international organizations are addressing the environmental impacts of mining, and underscore the relevant methods and protocols. I will also supplement this review by proposing the use of soil microbial communities as bioindicators of wetland development and sustainability. I will describe the laboratory and field procedures necessary to evaluate the various biological and physical aspects of mitigated wetlands, thereby offering mangers an effective monitoring technique. My intention is to confirm that microorganism development and preservation are critical to wetland health and longevity.^ My final objective is to document the relevant literature on environmental policy, and provide current scientific and policy review for researchers, managers and legislators. This thesis will synthesize the diverse and often contradictory theories, and suggest possible methodologies to bridge the science-policy gap.Overall, I intend to supply researchers, managers, and government agencies with a source of publications that can assist in evaluating, managing and monitoring wetland mitigation projects.
Thesis (M.S.)--University of South Florida, 2006.
Includes bibliographical references.
Text (Electronic thesis) in PDF format.
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Adviser: Henry R. Mushinsky, Ph.D.
t USF Electronic Theses and Dissertations.
The Science and Policy that Compels the Wetland Mitigation of Phosphate-Mined Lands by Nejma Danielle Piagentini A thesis submitted in partial fulfillment of the requirements for the degree of Master of Science Department of Biology College of Arts and Sciences University of South Florida Co-Major Professor: Henry R. Mushinsky, Ph.D. Co-Major Professor: Earl D. McCoy, Ph.D. Valerie Harwood, Ph.D. Date of Approval: March 10, 2006 Keywords: Florida, phosphate mining, soil microbiology, environmental policy, reclamation, bioindicators Copyright 2006, Nejma Danielle Piagentini
Acknowledgements I would like to express my unending gratitude to my mother, Joy Petit. I was always encouraged by my mother to pursue a gradua te level degree, and I am thankful for her support. Thank you, Mom, for all of the time and patience you invested in the review of this thesis. Thank you also to my sister Kira, and brother Josh, who are sources of inspiration, true friends and loving siblings. And to my entire family, thank you for your love and continued support. I would like to thank Dr. Henry Mushinsky and Dr. Earl McCoy for allowing me to be a part of their laboratory, for th eir dedication to helping me find my way, and their efforts to ensure a well-written and successful thesis. I also extend sincere gratitude to Christine Smith for being my right hand in Tampa when familial circumstances led me to New Jersey I appreciate your constantly keeping in touch with me, answering my questions, and being a friend. Thank you to Dave Feigley for being the voice of reason, for your comments and your revisions of this work. I offer thanks and my heart to my husband, Ant hony. Thank you for supporting my graduate experience and granting me the time and resources to continue my studies. It is such an honor to be your wife.
i Table of Contents List of Tables iii List of Figures iv List of Acronyms v ABSTRACT vi Chapter 1: Wetlands Definiti ons, Impacts and Legislation 1 Introduction 1 Wetland Terminology 2 Characteristics of Wetlands 3 Wetland Classifications and Settings 10 Legislation for Wetland Protection and Mitigation 13 Federal Wetland Regulations 13 State and Local Legislation 17 Non-regulatory Approaches to Wetland Regulation 21 Wetland Mitigation 22 Mitigation Terminology 22 Case Studies of Successful Wetland Mitigation Projects 25 Case Studies of Unsuccessful Wetland Mitigation Projects 29 Chapter 2: Phosphate Mining and Wetland Alteration 33 Introduction 33 The Phosphate Mining Process 33 The Environmental Repercussions of Mining 36 Wetland Alterations from Phosphate Mining 39
ii Chapter 3: International Mini ng and Wetland Preservation 41 Introduction 41 The International Environmental Community 41 The International Mining Community 49 Chapter 4: Soil Microorganisms as Indicators of Functional Wetlands 58 Introduction 58 Wetland Soil Characteristics 59 Soil Microorganisms as Bioindicators 63 Soil Microorganisms Types and Mi crobial Community Analysis Techniques 67 Types of Soil Microorganisms 67 Techniques for Measuring Biodivers ity and Community Structure 75 Indices of Richness and Non-mol ecular Analytical Techniques 75 Molecular Techniques 77 Measurements of Microbial Activity Â– Analysis of the Biogeochemical Cycles 83 Measurements of Microbial Ac tivity Â– Soil Organic Matter 88 Conclusions 92 Chapter 5: Ecological Research and Environmental Policy 93 Introduction 93 Ecological Research and Policy 94 Proposed Solutions to the Science-Policy Gap 100 References 107 Appendices 128 Appendix A: Wetland Classification 129 Appendix B: Rules Pertaining To Landscape Restoration as Set Forth by the Bureau of Mine Reclamation 130 Appendix C: Wetland Mitigation Bibliography 133
iii List of Tables Table 3.1: Range of Natural Habitats within Different Temperate and Tropical Wetland Ecosystems 42 Table 3.2: Productivity of Se lected Wetland Ecosystems 43 Table 3.3: End Points of Terrestri al Monitoring and Corresponding Soil Ecosystem Parameters 54 Table 3.4: List of Microbial Indica tors for Soil Health Monitoring 56 Table 4.1: Comparison of Florida Sandy Soil and Phosphatic Clay 60 Table 4.2: Examples of Important Autotrophic So il Bacteria 68 Table 4.3: Examples of Important Heterotrophic Soil Bacteria 68 Table 4.4: Dominant Cultural Soil Bacteria 73 Table 5.1: Wetland Mitigation Ba nks of the United States 103
iv List of Figures Figure 1.1: Wetland Biodiversity 2 Figure 1.2: A Schematic of a Wetland System 9 Figure 1.3: Schematic Diagram of a Construc ted Wetland 25 Figure 2.1: Distribution of Mines in Florida 34 Figure 2.2: Schematic Diagram of Phosphate Mining Process 36 Figure 3.1: The Mining Life Cycle 50 Figure 4.1: A Typical Soil Profile 61 Figure 4.2: The Carbon Cycle 70 Figure 4.3: The Nitrogen Cycle 71 Figure 4.4: Sulfur and P hosphorus Cycling 72 Figure 4.5: Organic Matter Cycle 89 Figure 5.1: The Distribution of Mitigation Banks in the United States 104
v List of Acronyms BMR Florida Department of Environmen tal Protection Bureau of Mine Reclamation CARL FloridaÂ’s Conservation and Recreation Lands Program CDA Coordinated Developed Area CE United States Corps of Engineers CWA FWS Clean Water Act EMAS European Union Eco-Management and Audit Scheme EPA United States Environmental Protection Agency ERP Environmental Resource Permit EU European Union EWRA The Emergency Wetlands Resources Act FDEP Florida Department of Environmental Protection FWS United States Fish and Wildlife Service GATT General Agreement of Tariffs and Trade HGM Hydrogeomorphic Approach IFA International Fertilizer Industry IHN The Integrated Habitat Network IIED International Institute for Environmental and Development IWWR Interagency Workgr oup on Wetland Restoration MMPA Marine Mammal Protection Act MOA Memoranda of Agreement NAFTA North American Free Trade Agreement NMFS United States National Marine Fisheries Service NRC National Research Council NRCS Natural Resource Conservation Service RPMC Resource Planning a nd Management Committees U.S. United States UNEP United Nations Environment Program VNIRS Visible-near infrared reflectance spectroscopy WET Wetland Evaluation Technique WRP Wetlands Reserve Program WWF The World Wide Fund for Nature
vi The Science and Policy that Compel the Wetland Mitigation of Phosphate-Mined Lands Nejma Danielle Piagentini ABSTRACT The State of Florida ranks fifth in the worldÂ’s production of phosphate. The phosphate industry relies on surface mining to withdraw the phosphate ore, and this process can devastate the natural environmen t. One of the most impacted natural resources is wetlands. Federal laws permit the legal destruction of wetlands providing their loss is compensated by the mitigatio n (i.e., the restoration, creation, or enhancement) of other wetlands, but the complexity of wetland ecosystems makes the mitigation process difficult. One of the goals of this thesis is to review the established Federal, State and non-regulatory guidelines for the development and maintenance of mitigated wetlands, evaluate their efficacy and present some underlying reasons for successful versus unsuccessful mitigation projects. The environmental repercussions of phos phate mining are not only pertinent to Florida or the United States. Wetland mitigation has become a global issue. Laws and programs that facilitate specific countries do not benefit wetland ecosystems on a landscape level. It is important to remain cognizant of the ramifications of wetland destruction and avoid piecemeal solutions to a wide-spread problem. Thus, my second objective is to investigate the progress and status of interna tional wetland preservation. I will examine how different countries and inte rnational organizations are addressing the environmental impacts of mining, and underscore th e relevant methods and protocols. I will also supplement this review by proposing the use of soil microbial communities as bioindicators of wetland develo pment and sustainability. I will describe the laboratory and field procedures necessa ry to evaluate the various biological and physical aspects of mitigated wetlands, ther eby offering mangers an effective monitoring
vii technique. My intention is to confirm th at microorganism devel opment and preservation are critical to wetland health and longevity. My final objective is to document the rele vant literature on environmental policy, and provide current scientific an d policy review for re searchers, managers and legislators. This thesis will synthesize the diverse and often contradictory theories, and suggest possible methodologies to bridge the science-policy gap. Overall, I intend to supply researchers, managers, and government agencies with a source of publications that can assist in evaluating, managing and monitoring wetland mitigation projects.
1 Chapter 1: Wetlands Definiti ons, Impacts and Legislation Introduction Wetlands are among the most productive ecosystems in the world, comparable to rain forests and coral reefs. A variety of species of microorganism s, plants, insects, amphibians, reptiles, birds, fish, and ma mmals comprise wetland ecosystems (Figure 1.1). Physical and chemical features su ch as climate, landscape shape (topology), geology, and the movement and abundance of wate r help to determine the flora and fauna that inhabit each wetland (U.S. Environmen tal Protection Agency [EPA], 2006). The functions of a wetland and the values of th ese functions to human society depend on a complex set of relationships between the wetland and the other ecosystems in the watershed. Wetlands cover almost 30 percent of the State of Florida and account for just over 10 percent of the remaining we tland area in the lower 48 United States. Over the past 200 years, Florida has lost an estimated 10 million acres of wetland, about half of the total area thought to exis t in the 1780's (Clark, 2004). So me of these remaining wetlands are well known, like the Florida Everglades while others may be small and unassuming. All play a vital role in flood protection, water quality and wildlife habitat. This first chapter will discuss wetla nds, their delineation characteristics, classification schemes, and importance. In addition, Federal, State and non-regulatory wetland regulations for natural and mitigated we tlands will be provided, as well as case studies of successful and unsuccessful mitigation projects.
2 Figure 1.1: Wetland Biodiversity (U.S. Envi ronmental Protection Agency [EPA], 2006). Wetland Terminology Today there are more than 50 definitions of the term Â“wetland.Â” Each definition is based on the spatial, temporal political, economical, and soci al differences among states and federal agencies. In 1979, Cowardin, Cart er, Golet and Laroe proposed the following definition, which was later adopted by the Unite d States (U.S.) Fish and Wildlife Services (FWS): Wetlands are lands transitional between terrestrial and aquatic systems where the water table is usually at or near the surface or the land is covered by shallow water. For purpos es of this clas sification, wetlands must have one or more of the fo llowing attributes: (1) at least periodically, the land supports predominately hydrophytes, (2) the substrate is undrained, hydric soil, and (3) the substrate is non-soil and is saturated with water or covered by shallow water at some time during the growing season of each year (Cowardin et al., 1979, p.3). Today, the U.S. Environmental Protection Agency (EPA), U.S. Army Corps of Engineers (CE), and the U.S. Departme nt of Agriculture, National Resources Conservation Service (NRCS) colle ctively refer to wetlands as:
3 Those areas that are inundated or sa turated by surface or ground water at a frequency and duration sufficient to support, and that under normal circumstances do support, a prevalence of vegetation typically adapted for life in saturated soil conditions. Wetlands generally include swamps, marshes, bogs, and similar areas (Environmental Laboratory, 1987, p. 9). States often incorporate fede ral language into their definitions. For example, the State of Florida defines wetlands in Secti on 373.019 (17) of the Florida Statutes, and Section 62-340.200 (19) of the Florid a Administrative Code, as follows: (Wetlands are)Â…areas that are inundated or saturated by surface water or groundwater at a frequency and a dur ation sufficient to support, and under normal circumstances do suppor t, a prevalence of vegetation typically adapted for life in saturated soils. Soils present in wetlands generally are classified as hydric or alluvial, or possess characteristics that are associated with reducing so il conditions. The prevalent vegetation in wetlands generally consists of facultative or ob ligate hydrophytic macrophytes that are typically adapted to areas having soil conditions described above. These species, due to morphological, physiological, or reproductive adaptations, have the ability to grow, reproduce, or persist in aquatic environments or anaerobic soil conditions. Florida wetlands generally include swamps, marshes, bayheads, bogs, cypress domes and strands, sloughs, wet prairies, riverine swamps and marshes, hydric seepage slopes, tidal marshes, mangrove swamps and other similar areas. Florida wetlands generally do not include longleaf or slash pine flatwoods with an understory dominat ed by saw palmetto. All of these definitions express thre e primary delineation, or diagnostic, characteristics used to determine wetland presence and viability. Characteristics of Wetlands Although wetland types and locales vary, they a ll possess environmental characteristics that distingui sh them from upland or othe r aquatic ecosystems. These characteristics are used as guidelines fo r the Army Corps of Engineers (CE) 1987 Wetland Delineation Manual, which is still us ed today. Wetland delineation involves establishing the boundary be tween wetlands and uplands (or non-wetland areas).
4 Wetlands are characterized by unique hydrologic, soil (substrate), and biotic conditions. The CE describes these characteri stics in the following manner: Wetland Hydrology constitutes the areas inundated either permanently or periodically, or the soil is saturated to the surface at some time during the growing season of the prevalent vegetation. Wetland Soils are classified as hydric; possess ch aracteristics that are associated with reducing soil conditions. Wetland Vegetation includes macrophytes that are typically adap ted to areas having the hydrologic and soil conditions described above (Environmental Laboratory, 1987). The Delineation Manual provides guidance for identifying jurisdictional wetlands. States have traditionally relied on vegetati on for wetland identification, but recently have either adopted the Federal approach, or deve loped tiered methods of analysis consisting of first considering vegetation and, when n ecessary, considering th e additional diagnostic characteristics of hydric soil properties and hydrologic indica tors. The Wetland Delineation Manual has been updated th rough a series of guidance documents memoranda from the CE Wetland Research Program, and an electronic edition is available online at http://el.erdc.usace.army.mil/wetlands/pdfs/wlman87.pdf. Positive indicators of hydrophytic vegetati on, hydric soils, or wetland hydrology must be present during some portion of the grow ing season for an area to be considered a natural wetland. Exceptions to the delineation rules do exist, however, because wetland systems are often subject to alteration. Na turally occurring events can promote the creation or alteration of wetla nd systems. For example, b eaver dams can impound water, thereby shifting the hydrology and vegetation of a wetland system. A wetland may be considered Â“atypicalÂ” when one or more of the field indicators have been obscured by some recent change. If the change took plac e after 1977 it may be determined to be a wetland and subject to the Clean Water Act. Problem area wetlands are wetlands that are inherently difficult to identify because field indicators of one or more wetla nd parameters may be absent or misleading, at least at certain times of the year. For ex ample, wetlands that have been purposely or
5 incidentally created by human activities may not possess the appropriate wetlands indicators compared to local, natural we tlands. The substrate or hydrology may be different due to the resources available to th e wetland developers. In addition, during the dry season, highly seasonal wetlands often l ack the hydrology and vegetative indicators typical of non-seasonal wetlands (E nvironmental Laboratory, 1987). Of the three delineation characteristics, wetland hydrology is considered the most important variable for determining the de velopment and maintenance of wetlands on a landscape level, because it is deemed the proce ss that drives the other ecological elements of the system. Furthermore, understandi ng the hydrologic aspects of wetlands helps explain their landscape diversity (Winter, 1988, 1992). Bedford (19 96) contends that replacement wetlands cannot be considered eq uivalent to natural wetlands unless their hydrologic regimes are comparable. The surface water components of wetlands, such as the hydroperiod (the period of time during wh ich the wetlands is covered by water) and water depth gradients, have been the subjec t of extensive research (Menges & Waller, 1983; Lieffers, 1984). The hydrologic regime of the ecosystem is determined by the flow, duration, amount, and frequency of water on a s ite (Interagency Workshop on Wetlands Restoration [IWWR], 2 003). Brinson (1993) describes thre e basic wetlands flow types: (1) vertical fluctuations of the water tabl e that result from evapotranspiration and subsequent replacement by precipitation or gr oundwater discharge into the wetland, (2) unidirectional flows that range from strong channel-contained curre nts to sluggish sheet flow across a floodplain, and (3) bidirectional, surface or near-surf ace flows resulting from tides or seiches. Snodgrass, Komoroski, Bryan and Burger (2000) studied hydroperiod duration and evaluated models of lentic communities to determine the relationship between hydroperiod length and species sustainability Their results indicated that shortÂ– hydroperiod wetlands support a unique group of species. Snodgrass et al. (2000) concluded that hydroperiod lengt h should be a primary criteri on for the development of wetland regulations, and advocated the conser vation of a diversity of wetlands that represent the entire hydroperiod gradient (termed th e Â“Landscape ApproachÂ”).
6 The hydrologic complexities of wetlands ha ve prompted the establishment of field indicators, which are used to identify the hydrological characteristic s of natural wetland systems. Examples of these indicators include: Standing or flowing wate r observed on the area duri ng the growing season; Waterlogged soil during the growing season; Watermarks on trees or other erect objec ts. Such marks indicate that water periodically covers the area to the depth shown on the objects; Drift lines, which are small piles of debris oriented in the direction of water movement through an area. These lines often occur along cont ours and represent the approximate extent of flooding in an area; Debris lodged in tress or piled ag ainst other objects by water; and Thin layers of sediment deposited on l eaves or other objects. Sometimes these become consolidated with small plant part s to form discernible crust on the soil surface (Environmental Laboratory, 1987). The readily available water in wetlands s upports soils (called hydric soils) that remain saturated for part or all of the year The upper part of hydric soils (6-12 inches) becomes anaerobic as water stimulates the gr owth of microorganisms, which use up the oxygen in the spaces between so il particles (IWWR, 2003). Anaerobic soils have unique chemical attributes. For example, the decom position of organic matter is slowed down as a result of the anaerobic natu re of wetlands. As such, we tlands trap carbon as soil organic matter instead of releasing it into the atmosphere (as carbon dioxide) and are considered carbon sinks (Sims, 1990). The thickness and permeability of wetland soils are essential to wetland formation (Winter, 1988). The low permeability of organic soils, as a result of their underlying silts and clays, generally restrict the vertical moveme nt of water to or from the peat layer, yet allow local groundwater discharg es and recharges to occur. Thus water can persist in wetland systems for longer periods of time (Bedford, 1996). The NRCS has listed approxima tely 2,000 types of soils in the United States that may occur in wetlands. This list, as well as a list of the Field Indicators of Hydric Soils,
7 is available on their website (http://soils.usda .gov). Field indicator s offer on-site means to confirm or reject the presence of hydr ic soil. Some of the indicators include: An odor of rotten eggs (sulfuric smell), Mottled soils, which are grayish subso ils with orange or reddish mottles, reflecting a fluctuating water tabl e but prolonged saturation; and Predominantly sandy soil that has dark stai ns or dark streaks of organic material in the upper layer below the soil surface. These streaks are decomposed plant material attached to the so il particles. When soil from these streaks is rubbed between the fingers, a dark st ain is left on the fingers (Environmental Laboratory, 1987). With the permission of the Soil Survey staff, the Florida Department of Environmental Protection (FDEP) generated a do cument on their website that outlines the hydric soils in Florida, and offers soil criteria and field indicators (http://www.dep.state.fl. us/mainpage/default.htm). In addition, a detailed description of the hydric soil indicators developed for Florida is provided in Soil and Water Relationships of FloridaÂ’s Ecological Communities Plants that can grow, compete and re produce amidst the waterlogged, anaerobic conditions of wetlands ar e facultative or obligate hydrophytic macrophytes, or hydrophytes. Hydrophytes have developed mo rphological and physical adaptations and distinct colonization strategies that allow alternate metabolic pathways for successful oxygen exchange in otherwise toxic environmen ts (Elliot, 2004). Wetland plants can be grouped based on their expected frequenc y of occurrence in the wetland system. Obligate wetland plants (OBL) must live a nd grow in wetlands and deep water habitats in order to survive, and occur in wetlands 99 percent of the time; Facultative wetland plant species (FACW) usually occur in wetlands 67 to 99 percent of the time, but are o ccasionally found in non-wetlands; Facultative plant species (FAC) are equa lly likely to occur in wetlands or nonwetlands (34 to 66 percent of the time); Facultative upland plants (FACU) exist predominantly in non-wetland habitats (occurring 1 to 33 percent of the time), but can survive in wetlands;
8 Obligate upland plants (UPL) almost occur more than 99 percent of the time in uplands (U.S. Fish and Wildlife Service [FWS], 2006); A positive (+) or negative (-) sign, when used with in dicators, more specifically defines the frequency of occurrence in wetla nds. The positive sign designates, Â“slightly more frequently found in wetlandsÂ” and the negative sign indicates, Â“slightly less frequently found in wetlands.Â” (FWS, 2005). Of all the vascular plants that grow in the United States, only one-third can tolerate the unique conditions associated with most wetlands. Of t hose that do occur in wetlands, only 26 percent are obligate hyd rophytes (FWS, 2005). Typical wetland vegetation includes emergent plants (those with leaves that grow through the water column, such as cattails, sedge s, and rushes), submerged pl ants (pondweeds, eelgrass), floating-leaved plants (such as water lilies and duckweed), trees (cypress, red maple and swamp oak), shrubs (such as willows and bayberry), moss, and many other vegetation types. The most common wetland plants are cattails, bulrushes, co rd grass, sphagnum moss, bald cypress, willows, mangroves, sedges, rushes, arrowheads and water plantains. Some field indicators of wetland plants are: Trees having shallow root systems and/or swollen trunks (e.g. bald cypress, tupelo gum); and Roots found growing from the plant st em or trunk above the soil surface (Environmental Laboratory, 1987). Several studies have shown that the di versity of plant communities within and among wetland ecosystems is related to patte rns of groundwater flow and groundwater chemistry, which result from the hydrogeologic setting of the wetla nd, the interplay and abundance of water and nutrients in the soil a nd factors such as the salinity, water depth and water level fluctuation, type of soil, temperature, local elevation, and landscape (Siegel, 1983; Siegel &Glaser, 1987 ; Glaser, Janssens & Seigel, 1990). The interplay of wetland hydrology, hydric soil and hydro phytes allow for a complex ecosystem. Wetlands improve th e quality of surface water by filtering, absorbing and settling out soil particles, organic matter and some nutrients such as phosphorus (Mitsch & Gosselink, 1993). These removed materials provide an excellent
9 environment for algae and bacteria, which ac tively degrade organic pollutants, such as pesticides, and decompose chemicals arrivi ng from upstream natural and human sources (Mitsch & Gosselink, 1993). Microbial commun ities (on plant roots and in the soil) convert organic nitrogen in to ammonium nitrogen and pe rform denitrification (the conversion of nitrates in the so il to less toxic atmospheric nitrogen). The dense wetland vegetation also traps pollutants that attach to the soil particles, and controls erosion by reducing wave and current ener gy, binding and stabilizing the soil, and helping the soil recover from flood damage (Maier, Peppe r & Gerba, 2000). Figure 1.2 depicts the various components and interactions within wetland systems. Figure 1.2: A Schematic of a Wetland System (Maser, 2006). Wetlands serve as transitional areas between terrestrial and aquatic systems, and support a variety of microand macrofauna. More than one-third of the threatened and endangered species in the United States live primarily in wetlands, and nearly half use wetlands at some point in their lives (EPA, 1995). Numerous species of invertebrates, fish, and reptiles depend on wetland water cycles for reproduction and survival. Approximately 75 percent of all commercial marine fish species depend on estuaries, which in turn rely on nearby wetlands to maintain these productive ecosystems (IWWR, 2003). Nearly all amphibians and at least 50 percent of migrator y birds use wetlands
10 regularly. The breadth of species that de pend on wetlands for survival, reproduction and breeding habitat is given on the EPA website (www.epa.gov) and is available in the 1995 EPA resource entitled Â“AmericaÂ’s Wetlands: Our Vital Link Between Land and Water.Â” In addition to serving as habitats for various wetland species and filtering surface water, wetlands have a multitude of other benefits, such as providing: Healthy fisheries; High Biological productivity; Biodiversity protection; Erosion control; Flood damage reduction and sediment control; Aesthetic and recreational opportunities; and Water quality improvement (FWS, 2005; EPA, 1995). Wetland Classifications and Settings Wetlands occur in a wide variety of settings due to differences in geological characteristics, such as surface relie f and land surface slope, the thickness and permeability of the soils, and the compositi on, statigraphy and hydraulic properties of the underlying geological materials (Bedford, 1996). Winter (1992) outlined eight physiographic settings in which wetlands deve lop: (1) terraces and scarps within coastal lowlands, (2) terraces within riverine vall eys, (3) steep slopes (mountains) adjacent to narrow lowlands, (4) depressions in large extensive lowlands (playas), (5) morainal depressions, (6) dune fields, (7) sinkholes an d (8) permafrost. As all physiographic settings do not occur in all climatic regimes, the number of ideal locales is limited. Further wetland occurrence is restricted by the intricacies of local surface and groundwater flows (hydrodynamics). Cowardin et al. (1979) identified five ma jor wetland systems Â– Marine, Estuarine, Riverine, Lacustrine and Palustrine. Thes e systems are distinguished by a variety of hydrologic, geomorphic, chemical, and bi ological characteris tics such as: Marine open ocean and its associated shoreline
11 Estuarine estuaries where mixing of salt and fresh water occur Riverine channels of rivers and streams Lacustrine lakes, reservoirs and deep ponds Palustrine mostly freshwater wetlands and shallow ponds plus inland saline wetlands (see Appendix A). The FWS have developed a wetland classification system based on the determinations of Cowardin and his colla borators. The FWS identified 55 different classes of wetland and deepwate r habitats used to map we tlands throughout the United States. Wetlands and deepwater habitats ar e first divided into five ecological systems (listed above), then grouped by subsystems a nd classes. Most of the nationÂ’s wetlands are Palustrine wetlands. At the class level, wetlands are separated by vegetation (where more than 30 percent of the area is vegetated) or by the predominant substrate for other wetlands. Vegetated classes include aquati c bed, emergent wetland, scrub-shrub wetland, forested wetland, and moss-lichen wetland. N on-vegetated classes include rock shore, unconsolidated shore (e.g. mudflat and sandy beach), streambed, rock bottom, and unconsolidated bottom. Wetlands may furt her be described by various modifiers including water regime (hydrology), water ch emistry (pH and salinit y/halinity), soil (organic or mineral), and other special m odifiers (e.g. partly dr ained, diked/impounded, excavated and artificial) (FWS, 2005). The National Research CouncilÂ’s 1995 repor t entitled Â“Wetlands: Characteristics and BoundariesÂ” outlines a second classification scheme. This report lists several major classes of wetlands and typi cal associated vegetation: Freshwater Marsh grasses, sedges herbs; Tidal Salt and Brackish salt tolerant grasses, rushes; Prairie Potholes grasses, sedges, herbs; Fens sedges, grasses, shrubs; Bogs sphagnum moss, shrubs trees; Swamp Bottomland cypress, gum, red maple; and
12 Mangrove Forest black, red white mangroves (National Research Council [NRC], 1995). Brinson (1993) defines wetland classifi cation in terms of the Hydrogeomorphic Approach (HGM) to assessing wetland functio ns. The HGM Approach assesses wetlands based on three fundamental factors that influence ho w wetlands function: (1) hydrodynamics Â– the direction, flow and strengt h of water movement within the wetland, (2) geomorphic setting Â– the topographic loca tion of the wetland w ithin the surrounding landscape, (3) hydrology Â– precipitation, surf ace flow and groundwater discharge (i.e. that water source.). The HGM Approach first classifies we tlands based on their differences in functioning, defines the functions that each class of wetlands performs, and uses reference wetlands to estab lish the range of functioning of the wetland. Regional assessment models are then developed based on the functional profil e that describes the physical, biological and chemical characterist ics of a regional we tland subclass (Brinson, 1993). A supplementary classification of wetlands is based on the ecosystemÂ’s hydrogeologic setting, which is the position of the wetland in the landscape with respect to the flows of surface water and ground water. Several authors have summarized this classification method (Winter, 1992; Winter & Woo, 1990). For the complete national wetland classification standards, please vi sit: http://wetlands.fws.gov/PubsReports/pubs. Wetlands types include tidal marshes, pr airie potholes, seagrass beds, forested wetlands and seasonally ponded sites, such as ve rnal pools. Some of these wetland types, such as seasonal wetlands that are dry most of the year, may not al ways appear to be wetlands. Each system is unique and offers distinct structure, ecological function, aesthetic use and recrea tional benefits. For example, sm aller wetlands near flood-prone residential areas may offer greater flood prot ection than a larger wetland more suitable for extensive wetland habitat (Bi ngham, Clark, Haygood & Leslie, 1990).
13 Legislation for Wetland Protection and Mitigation Federal Wetland Regulations Federal, State and Local wetland regulations were developed for the protection of wetlands through permitting requirements, which control discharges and hazardous waste. These regulations also control use and/or alteration of wetlands as they pertain to other resources (e.g., wildlife) that depe nd on proper functioning wetlands (Dennison & Berry, 1993). The advent of legislation crea ted to protect the na tionÂ’s waterways began in 1899 with the Rivers and Harbors Act prohi biting the creation of obstructions to the navigable capacity of an y of the waters of the United Stat es without specif ic approval of the Chief Engineer of the CE. In 1972, Congress passed the Federal Wate r Pollution Control Act Amendments, also known as the Clean Water Act, to restore and maintain the chemical, physical, and biological integrity of the NationÂ’s waters (Lewis, 2001). The Act defined Â“navigable watersÂ” as Â“waters of the United States.Â” In 1977, the CE redefined navigable waters to include Â“isolated wetlands and lakes, interm ittent streams, prairie potholes, and other waters that are not part of a tributary system to interstate waters or to navigable waters of the United States, the degradation or destru ction of which could affect interstate commerce.Â” This definition remains in effect today (Lewis, 2001). Provisions in the Food Security Act ( FSA) of 1985 lent suppor t to the wetland preservation fight. Secti on 3821, Subchapter III (Wetland Conservation) of Chapter 58 (Erodible Land and Wetland Conservation and Re serve Program) of the Act states that individuals draining wetlands after December 1985 would not be eligible for many of the farm and loan programs of the U.S. Departme nt of Agriculture (USDA). This is the called the Swampbuster provision of the 1985 Act. Swampbuster discourages the conversion of wetlands to cropland use. Section 3822 of the Act deals with the delineation of wetlands, resulti ng wetland delineation maps, and other issues such as the appeal process if a land owner does not agree th at features delineated as wetlands on their farms are indeed wetlands (Dennison & Berry, 1993).
14 In 1988, as part of his election campai gn, then Vice-President George Bush pushed for a Â“no net lossÂ” of our nationÂ’s wetl ands as a national goal. He endorsed the recommendations of the National Wetland Poli cy Forum, an intergovernmental, public and private coalition convened at the reco mmendation of the EPA (Dennison & Berry, 1993). After the 1988 election, the Bush Administration upheld their decrees and the FWS released the National Wetlands Prio rity Conservation Plan in 1989, which implemented the Federal Emergency Wetlands Resources Act of 1986 (FWS, 1989). The Bush campaign recognized and acknowledged wetland losses across the United States and developed procedures for the acquisiti on of wetlands by Federa l, State and local governments. Unfortunately, over time, all good intentions began to wane as legislation was being passed that contradicted previous effo rts. The most noted examples were the Memorandas of Agreement (MOA) in 1989 an d 1990 between the CE and EPA regarding wetland mitigation requirements. This agr eement authorized a less than one-to-one replacement of wetland acreage for degraded or lost wetlands. Many environmentalists believed that this was a major step in the overall weakening of wetland protection (Dennison & Berry, 1993). Â“N o net lossÂ” remained federa l policy under the Clinton Administration, but today seems to have been relegated to nothing more than a catch phrase. Of all of the past wetland protection legi slation, the most cont roversial is Section 404 of the Clean Water Act (CWA). State, federal and private entries are required to obtain a permit from the CE before depositing dr edged or fill materials into the waters of the United States, which includes wetlands. Under Section 404, wetlands may legally be destroyed, but some form of reclamation or mitigation must compensate their loss. Several sections of the CWA pertain to regulating impacts to wetlands: Section 101 specifies the objectives of this Act which are implemented largely through Title III (Standards and Enforcement). Section 301 (Prohibitions) specifies that th e discharge of dredged or fill material into waters of the United States is s ubject to permitting spec ified under Title IV
15 (Permits and Licenses) of this Act a nd specifically under S ection 404 (Discharges of Dredge or fill Material ) of the Act. Section 401 (Certification) specifies a dditional requirements for permit review particularly at the state level. Section 309(a) allows the EPA to i ssue administrative compliance orders requiring a violator to st op any ongoing illegal discha rge activity and, where appropriate, to remove the illegal discha rge and otherwise restore the site. Section 309(g) stipulates that the EPA and the CE can assess administrative civil penalties of up to, but not exceeding, $125,000 per action. Sections 309(b) and (d) and 404(s) give the EPA and CE the authority to take civil actions, seek restoration and other type s of injunctive relief, as well as civil penalties. The agencies also have authority under Section 309(c) to bring criminal judicial enforcement actions for knowingly or negligently violating Section 404 (Lewis, 2001; De nnison & Berry, 1993). Administration of the CWA falls to several ag encies each with sp ecific tasks: the CE issues permits; the EPA reviews permit applications, prepares guidelines for disposal sites, denying or restricting the use of a ny defined area as a disposal site, general enforcement; the FWS manages fish and w ildlife game species and protects the threatened and endangered species within wetlands; and the U. S. National Marine Fisheries Service (NMFS) pr ovide consultation and assist ance under Fish and Wildlife Coordination Act (Dennison & Berry, 1993). In addition to jointly implementing th e CWA Section 404 program, the EPA and CE share Section 404 enforcement authority. Se ction 404 violations stem from failure to comply with the terms or conditions of a Se ction 404 permit, or discharging dredged or fill material to waters of the United States without a permit. Although EPAÂ’s Section 404 enforcement program relies on voluntary co mpliance, enforcement tools are provided under Sections 309 and 404 of th e CWA (Dennison & Berry, 1993). In 1989, the EPA and CE entered into a M OA on enforcement to ensure efficient and effective implementation of their shared authority. Under the MOA, the CE has the lead on Corp-issued permit violation cases. For unpermitted discharges, the EPA and the
16 CE determine the appropriate lead agency, based on criteria in the MOA. Decisions regarding enforcement consider many factors when determining the need and type of enforcement actions, such as the amount of fill, the size of the water body (acres of wetlands filled and the environm ental significance), the dischargerÂ’s previous experience with Section 404 requirements, and the disc hargerÂ’s compliance history (EPA, 1990). Though the provisions of Section 404 are th e most discussed, no single piece of legislation entirely expresses the federal go als for wetland protection and management. In reality, a series of goals that relate dire ctly or indirectly to wetland management and protection are embedded in several pieces of federal legislation. Executive Order 11990 (Protec ting of Wetlands) sets goa ls for federal agency action, and Executive Order 11988 (Flood Plain Ma nagement) requires federal agencies to avoid direct or indirect support for floodplain development wherever there is a practical alternative (Dennison & Berry, 1993) In 1973, the EPA adopted a Â“Statement of Policy on Protection of th e NationÂ’s WetlandsÂ” which lists the values of wetlands and describes the AgencyÂ’s deci sion-making processes. The EP A has also developed an implementation plan through its Wetlands Strate gic Planning Initiative, which lists five goals for wetland management. In addition, th e Endangered Species Act authorizes the FWS to prepare and implement a Habitat Cons ervation plan in certain situations to protect endangered species (Dennison & Berry, 1993). Wetland language is also found in direc tives such as the Emergency Wetlands Resources Act of 1986, which states that exis ting federal, state, and private cooperation should be strengthened in order to mini mize further losses and to assure their management for future generations (De nnison & Berry, 1993). In 1987, the National Wetlands Policy Forum was convened at the request of the EPA to address major policy concerns on protection and ma nagement of wetland resources. In November 1988 the Forum released Â“Protecting AmericaÂ’s We tlands: An Action AgendaÂ” which contains over 100 recommendations for improving wetlan ds conservation. Recently, the EPA and the CE announced the release of a compre hensive, interagency National Wetlands Mitigation Action Plan, as well as an improved Wetlands Mitigation Regulatory
17 Guidance Letter, to further the achievement of the goal of Â“no net lossÂ” of wetlands (Dennison & Berry, 1993). Additional federal environmental laws that may impose additional permitting and regulatory compliance requirements on activit ies undertaken in wetland areas include: The Coastal Zone Management Act; The Endangered Species Act; The National Historic Preservation Act; The Preservation of Historical and Archaeological Data Act; The Marine Mammal Protection Act; The National Environmental Policy Act; The Clean Air Act; The Coastal Barriers Resources Act; and The Marine Sanctuaries Act. State and Local Legislation The Â“State Wetland Protection Program sÂ”, prepared in 1986 for the EPA, identified 14 states with specific inland wetland laws (Dennison & Berry, 1993). Local governments may have ordinances that sp ecifically address wetland management, particularly in regions where states delegate some wetlands regulatory responsibilities to local governments. Goals for wetland mana gement may also be found in planning documents of local government s or regional planning bodies. Some states have consistency requirement s that apply to inland as well as coastal areas. Florida has a number of innovative pr ograms for coordinati ng state, regional and local planning, and assigning authority to the most appropriate level of government. For example, Resource Planning and Manageme nt Committees (RPMC), comprised of interested parties and land managers, identi fy resource issues and develop management plans that are later incorporated into local comprehensive plans and land use regulations. The RPMC process promotes advanced planni ng and provides a vehicle for introducing state goals and policies to the lo cal level (Dennison & Berry, 1993).
18 The Florida Department of Environmenta l Protection (FDEP) and the five water management districts (South Florida, Suwa nnee River, St. JohnÂ’s River, Southwest Florida and Northwest Florida) share the respons ibility for protecting wetlands in Florida. Wetlands are regulated by Chapter 373 of the Flor ida Statutes; Part IV of the Statutes is the StateÂ’s surface water regulatory program, which is jointly implemented by the State and the water management dist ricts using rules a dopted by the FDEP for each specific water management district (Morgan Lewis, 2001). Impacts to wetlands (including dredging or filling) require an Environmental Resource Permit (ERP) from the FDEP or one of the water management districts. The ERP program and Sovereign Submerged Lands Program regulate ac tivities throughout Florida, except for the Florida Panhandl e. The Wetlands Resource Permit program covers the Panhandle area and re gulates any alterations of wa ter, and wetlands that are connected (either naturally or artificially) to Â“named waterÂ” such as Gulf of Mexico, bays, sounds, estuaries, lake s and streams (FDEP, 2006a). Active mining throughout Florida has sp arked several legislative changes. Florida has mandated that all land mine d for phosphate after July 1, 1975 must be reclaimed (defined in the Wetland Mitigation Terminology section) (Chapter 211, Florida Statutes; Chapter 378, Florida Statutes). Prior to 1975, more than 149,000 acres of land were mined and disturbed by phosphate mining operations. Taxes collected on each ton of phosphate rock mined today are used to he lp pay for the reclamation areas mined prior to 1975, most of which are now in private ow nership. These taxes also help pay for preserving environmentally sensitive lands through the FloridaÂ’s Conservation and Recreation Lands (CARL) Program. Currently, about 15 percent of the Bone Valley district of Polk and eastern Hillsborough and Manatee counties are overlain by wetlands. Native wetlands include both the herbaceous ecosystems (wet prairi es and marshes), and forested or shrubdominated ecosystems (cypress domes, c ypress swamps, mixed cypress-hardwood swamps, bay forests and swap thickets). The phosphate i ndustry has been required by law to reclaim wetlands since the adoption of the FDEP rules (Appendix B), which state
19 that the wetlands affected by mining operations are to be restored to at least pre-mining surface areas, acre-for-acre and t ype-for-type (FDEP, 2006b). The FDEP Bureau of Mine Reclamati on (BMR) was developed in the late 1980s to preserve, manage and protect state la nds another than park s, recreational and wilderness areas. The Integrated Habita t Network (IHN) and Coordinated Developed Area (CDA) plans, prepared by the BMR, provide guidelines for t reclamation and permitting throughout the central Florida phosphat e mining districts (Florida Department of Environmental Protection Bureau of Mine Reclamation [FDEP-BMR], 2002a). Before any mining proceeds, the land reclamation process begins with the submittal of a reclamation plan to the BMR and other local, state and federal agencies. The BMR reviews the plans for consistency with adjace nt mines and land uses, and with its regional conceptual plan. For a reclamation plan to be approved, it mu st provide detailed information on the identified land (includi ng vegetative cover, animal species, topography, watersheds, land use and potential archaeology site s) both before the mining process and after reclamation (FDEP-BMR, 2002a). The reclamation projects are actively managed by the phosphate mining companies, but sites are also field inspected quarterly by FDEP staff to ensure that permit conditions are achieved. Evaluations are cons idered on a case-by-case basis, according to the conditions mentioned in the permit, becau se no standardized, quantitative assessment of phosphate mine reclamation exists. On ce reclamation had been satisfactorily completed in accordance with permit requirements, the mining company can be Â“releasedÂ” from further obligation to pe rform mitigation, and the BMR no longer has jurisdiction over the area (FDEP, 2006b). According to the December 2002 BMR Â“Rate of Reclamation Report,Â” 175,752 acres of land in Florida land have been mined since the mandatory reclamation law passed in July 1975. Of those acres, 53,821 acres have been reclaimed and released, and another 58,138 acres have been reclaimed th rough vegetation for industrial use, making a total of 111,959 reclaimed sites. This amount is 64 percent of what has been mined from 1975-2004 (FDEP-BMR, 2002b). These statistic s sound impressive, but we must recognize that many factors ar e necessary for successful reclamation.
20 Reclamation is difficult to define because it depends on the reclamation goals. For example, the Strategic Regional Policy Plan for South Florida specifies several measures of success including: (1) improving the quality and quantity of the wetlands in the region, (2) increasing the pe rcent of the restored/enhance d wetlands, and (3) reducing the loss of wetlands (Southwest Florida Regional Planning Council, 2004). Success also depends on improvements being made to the existing reclamation process and policies that emphasize the importance of ecosystem fu nctionality. The numerous challenges in wetlands restoration (e.g., crea ting and maintaining the appropriate hydrologic regimes, ensuring water quality, cont rolling invasive or opportunistic plant species, soil development, and native vegetation establis hment) seem to limit our advancements. Additional research is needed to furthe r address the ecological query of wetland sustainability. As a result, the Florida In stitute of Phosphate Research (FIPR) was developed to conduct and fund research aimed at addressing the environmental, economic and health impacts of FloridaÂ’s phospha te mining and fertilizer industry. FIPR is financed with funds from the state severance tax on phosphate rock (Florida Institute of Phosphate Research [F IPR], 2004a). Studies include those designed to develop better techniques fo r reclaiming land, evaluate th e chemical processing of phosphate rock, the beneficiation process, publ ic health issues, and emergent mining and processing technologies (FIPR, 2004a). Technical advisory committees, made up of stakeholders from government, industry, environmental gro ups and academics, review projects. Recently, research has focused on the potentia l of clay settling areas as the setting for wetlands creation. In 1984, FIPR funded a 5year restoration research project that centered on integrated lands cape restoration (Brown & Tighe, 1991). The project was designed to provide the phosphate industry w ith information about the structure and organization of watersheds (s izes, slopes, upland/wetland ratios, spatial organization, etc.) and ecosystems (dominant species in all strata, soils, t opography, and hydrologic regimes). The project team also studied mine d lands to characterize abiotic conditions in order to cross reference mine d lands with appropriate ecosystem types (Brown & Tighe, 1991). The authors considered the work a restoration cookbook, as they outlined the
21 design principles for watersheds, the characte ristics and design guidelines for floodplains, and the structural characteristics of natural wetlands. As the phosphate ore of the northern portions of the southern di strict is mined out, the industry is moving southward. Local oppositi on arises from the persistence of noise and air pollution, the alterati ons of the regional hydrology, a nd the clay settling basins, which are estimated to occupy as much as 40 to 60 percent of the landscape after mining (Brown, 2005). Current research funded by FIPR is addressi ng these topics, as well as modeling landscape scenarios of future impacts. Non-regulatory Approaches to Wetland Regulation Federal, state and local governments ha ve also adopted several non-regulatory programs for wetland protection. The principal owners and managers of wetlands at the federal level are the U.S. Department of the Interior, including the FWS, the National Park Service and the Bureau of Land Manage ment; the U.S. Department of Commerce, specifically the National Oceanic and At mospheric Administra tion; and the U.S. Department of Agriculture, mainly the U.S. Fo rest Service. Other Federal agencies, such as the Bureau of Reclamation in the Depa rtment of the Interior, the CE and the Department of Defense, also mana ge wetlands (Sunding & Zilberman, 2003). Some non-regulatory programs involve th e government acquiring full or partial title of the land, while others provide fina ncial inducement, tax credits or government payments to land owners for protecting and managing the wetlands in a desired manner. Many of the programs, however, have broader goals and the protection of the wetland is only part of that broader purpose (Goldfie ld & Clark, 1990). Overall, most programs encourage protecting existin g wetlands, and promoting re storation and enhancement efforts. Land acquisition has proven to be an e ffective program for wetland preservation. The Emergency Wetlands Resources Act of 1986 (EWRA) contains a provision for establishing a National Wetlands Priority Conservation Plan, whic h guides acquisition efforts on all levels. The FW S has designed a 10-year acqui sition plan which aims to protect more than 2.6 million acres of waterf owl habitat through the purchase of land or
22 conservation easements (FWS, 1988). For land acquisition, funds can be generated through appropriations from general revenues (u sually based on property tax receipts), special purpose bonds, and intergovernmen tal transfers (Sunding & Zilberman, 2003). The EWRA also authorizes entrance fees at certain national wildlife refuges with 70 percent of the proceeds going to the Migrat ory Bird Conservation Fund. Other funding sources include bonds, general re venue (from state taxes), stam p taxes, severance taxes, and tourist impact taxes, with additional funds coming from other taxes and revenues, as from state forests and hunting pe rmits (Sunding & Zilberman, 2003). The majority of the nationÂ’s wetlands are privately ow ned, thus private landowners are encouraged to manage/p rotect these resources through economic incentives such as subsidies and tax breaks. Wetland Mitigation Mitigation Terminology The terms mitigation and reclamation are often interchanged. Wetland mitigation refers to the restoration, creation, or enhancement of wetlands as compensation for permitted wetland losses under Section 404 of the Clean Water Act (Lewis, 1989). The term mitigation is synonymous with Â“compensatory mitigationÂ” because this process reduces environmental damage by avoiding, minimizing, and compensating for activities that damage protected resources. Mitiga ted wetlands can be reclaimed, restored, enhanced or created (construc ted). Reclamation can be us ed in a number of ways, including describing: 1) th e conversion of mined or other disturbe d lands into economically productive properties, such as grazi ng lands or orchards; 2) the filling in of wetlands or shallow coastal waters to create land, usually for housing or urban infrastructure, but also for agriculture in some parts of the world; and 3) the conversion of disturbed lands to natural or semi-natural habitat (Streever & Peck, 2002). Both terms are used throughout this document. The National Research Council, in its 1992 report Â“Restoration of Aquatic EcosystemsÂ”, defined restoration as the retu rn of an ecosystem to a close approximation of its condition prior to disturbance, and th e reestablishment of pre-disturbance aquatic
23 functions and associated physical, chemical a nd biological characte ristics. Restoration encompasses a plethora of terms applied to projects undertaken to re place or repair the ecosystem function. This process can be achieved by rehabilitation or reestablishment processes. Rehabilitation is us ed to describe aesthetic impr ovements made to an existing, natural resource. Rees tablishment is the restoration of a former wetland, which now may be degraded (NRC, 1992). Site restoration requires the recontouri ng of spoils, the replacement of native habitat and plant communities, and the creation of a landscape that includes areas of low topographic relief where surface and groundwat er can develop appropriate hydrologic regimes (Brown, 2005). Two techniques have been widely used to introduce wetland vegetation to created wetland areas: mulc hing and hand-planting. In many projects, organic mulch from wetlands permitted for mi ning is hauled to th e restoration site complete with the roots, sprouts and seeds of wetland plant species intact. The mulch is then spread over the subsoil to form an organic layer that may vary from a few centimeters to tens of centimeters in thickness. As the water table returns to pre-mining levels, germination of the seed material in the mulch begins the revegetation process, and later, trees and other species are planted to complete the suite of plant species required by the restoration permit (IWWR, 2003). Another mitigation option is enhancement. This technique involves increasing one or more of the functions performed by an existing wetland beyond what currently or previously existed in the wetland (Lewis 1989). Often there is an accompanying decrease in other functions. For example, adding more water to a wetland may create better habitat for fish, but it will decrease the ab ility of the wetland to hold flood waters. This trade-off is particularly true for e nhancement in relatively undisturbed wetlands (Lewis, 1989). Mitigation through preservation involves the protection of wetlands through appropriate legal and physical mechanisms, such as conservation easements and title transfers. Preservation can include protecti on of upland areas adjacent to the wetland, or enhancement of the existing system (Lewis, 1989).
24 Some mitigation requires the creation (or construction) of a new wetland system that is designed to replace the existing acr eage, structure or func tion of a pre-existing wetland. Constructed wetlands ar e created by excavating upland soils to elev ations that can support wetland species and the appr opriate wetland hydr ology. Hammer (1989) described constructed wetlands as man-made complexes of sa turated substrates, emergent and submergent vegetation, animal life and water that simulate natural wetlands for human use and benefits. Archit ecturally, constructed wetlands are a series of rectangular plots filled with soil or gr avel and lined to prevent waste from leaching into the groundwater (Figure 1.3). Constructed wetlands have di fferent basic designs, and th us have different flow characteristics. Kadlec (1987) described the diffe rent flows as: (1) horizontal flow systems with the wastewater level above the soil surface, (2) horizontal subsurface flow systems with the wastewater level below the soil surface, (3) and ve rtical flow systems with upstream or downstream characteristics, and continuous or intermittent loading. The soil is the main supporting material fo r plant growth and microbial films, and promotes the physiochemical and biological process that occur between the plants, microorganisms, and pollutants of the wetland ecosystem. The vegetation grown in these plots offer a root mass for filtration, and provide oxygen and carbon fo r water treatment. The roots also offer attachment sites for microorganisms, which consume the available oxygen in the process of breaking down pollutants (Stottmeister et al., 2003). Constructed wetlands can serve as wastew ater treatment plants. The operation and maintenance costs are lower than conven tional treatment plants, and less energy and supplies are needed. These facilities can surv ive with periodic on-site labor, rather than the continuous, full-time a ttention required by conventiona l plants (EPA, 2004). In addition, constructed wetlands can also serve as wildlife sites, designed to attract various animals and provide needed wetland habitat.
25 Figure 1.3: Schematic Diagram of a Constructed Wetland (EPA, 2004) Constructed wetlands can be used to treat domestic, agricultural and mine drainage wastewater, as well as leachacte contaminated groundwater and industrial effluents (EPA, 1988). With the increased use of constructe d wetlands, government agencies are devising appropria te regulations to pr otect public health and safety without unduly burdening constructed wetland designers and operators (EPA, 2000). A constructed wetland bibliography, compiled by the United States Department of AgricultureÂ’s Natural Resources and C onservation Service (formerly the Soil Conservation Service) and the Water Qua lity Information Center at the National Agricultural Library, consists of more than 600 citations outlining the use, maintenance and legislation of constructed wetlands. On e hundred and sixty one of these citations have abstracts. A second edition, Constr ucted Wetlands and Water Quality Improvement (II), created in June 2000, is now available. Both of these files are accessible online at http://www.nal.usda.gov/wqic/C onstructedwetlandsall. Case Studies of Successful Wetland Mitigation Projects Land managers often focus on ways to id entify functional wetlands in the field. Some of the documented indicators involve plants, soils and wildlife species. Though much has been written about wetland indicators, our ability to successfully mitigate these resources remains unacceptable. Glubiak, Nowk a and Mitsch (1986) pr edicted that while protection of wetlands has been offered th rough the Section 404 permitting process, this
26 legislation is not able to prevent wetlands lo sses as effectively as it should. However, mitigation success stories do exist. The St ate of Florida and Federal government are conducting large-scale ecosystem restorations such as the South Florida Everglades Ecosystem Restoration. This restoration pr oject was developed to address the excess nutrient loading in the Florida Everglades and re-establish the natural hydrology of the area (Chimney & Goforth, 2001). The Everglades dominate the landscape of south Florida and are considered an extremely important natural ecosystem. As a result of agricultural and urban development, eutrophication resulting from st ormwater runoff, changes in hydrology and the invasion of exotic species, the biotic inte grity of the Everglades is now threatened. To protect this valuable resource, approxi mately 17,000 hectares of treatment wetlands, referred to as Stormwater Treatment Areas, are being built to treat su rface runoff before it is discharged into the Everglades (Chimney & Goforth, 2001). Overall, only a very limited number of wetland community types have been created in reclaimed phosphate mining areas. Of the 30 classified types of wetlands occurring in Florida, 11 types were observe d in the pre-mining landscape within the Florida phosphate mining districts. Of th e numerous plant communities present in wetlands pre-mining, only about 6 types have been generally recreated in mitigation projects (Erwin, Doherty, Brown & Best, 1997). Increasing uncertainty exists about the appropriate time frames and best criteria for evaluating wetland reclamation. Many techniques can be used to monitor the progress of mitigated wetlands; observing vegetation growth is considered the easiest and most common method (Wentworth, Johnson & Kologiski, 1988; Jarman, Doberteen, Windmiller & Lelito, 1991; Atkinson, Perry, Smith & Carins, 1993). Vegetative cover may be an easy measure of success but is of ten a poor indicator of function. Magurie (1985) used area, vegetative cover and implem entation of permit conditions to estimate mitigation success in Virginia and found th at only 50 percent of the 23 mitigated wetlands were successful. An additional study by Reimold and Cobl er (1985) conducted for the EPA gave similar results. Odum (1989) investigated the Â“self organizationÂ” capacity of nature to
27 both recruit species on its own and to make choices from those species introduced by humans. The author introduced as many speci es as possible, under the assumption that the system would assist in its own design by choosing those most appropriate. This approach emphasized the importa nce of natural col onization of species in ecosystem. Reinartz and Warne (1993) compared 11 create d wetlands in southeastern Wisconsin that were naturally colonized with 5 wetlands in the same region where 22 species were introduced by seeding. Results indicated that success was achieved primarily through multiple seeding, multiple transplanting an d establishment of hydrologically open systems, which allowed nature to participat e in the wetland design. A large numbers of mitigation projects were undertaken in the 1970s and 1980s (reviewed and summarized in Kentula, Brooks, Gwin, Holland and Sherman, 1992) and analysis of these projects revealed that establishment of wetland vegetation was prim arily used as a measure of wetland success. As it is so difficult to monitor and mana ge every aspect of a wetland ecosystem, the Indicator Species Concept was proposed whereby a single species would be monitored and/or protected. Debate over this initiative arose because there is no consensus on what the indicator should indi cate, and which species provide the best indicators (Simberloff, 1998). Other mon itoring techniques include managing Umbrella Species (a species that needs such large tracts of habitat that saving it will automatically save many other species) and Flagship Species (charismatic large vertebrate used to anchor conservation campaigns since they can arouse public intere st and sympathy). Both of these concepts have been used for monitoring, but remain controversial. A review of the methods used to measur e success in previous restoration efforts by McCoy and Mushinsky (2002) specified the value of multi-metric methods for measuring the success of wildlif e community restoration when the restoration covers a broad geographic area. Using multi-metrics species were ranked according to the magnitude of the difference betw een distributions at restored and reference sties. McCoy and Mushinsky applied this species ranki ng method to data from 30 restored sites (phosphateÂ–mined land) and 30 reference sites in central Florida, and results indicated that ranking enables well-founded decision-mak ing about which species can be excluded,
28 if necessary, without seriously compromising th e restoration goal. This technique also overcame the problem that a single species ma y inordinately influence the result (McCoy & Mushinsky, 2002). The concept of ecosystem management also addresses the problems of singlespecies management. According to this view point, if a healthy ecosystem is maintained, the component species will all thrive. Some conservationists see ecosystem management as a Â“Trojan HorseÂ” that would allow continue d environmental destruction in the name of modern resource management (Simberloff, 1998). The amount of time required for successful mitigation is also a consideration. Petranka and colleagues (2003) used amphibi ans to assess the ecosystem function of compensatory wetlands created at a mitigation site in western North Carolina. Petranka, Murray and Kennedy (2003) studied pond co lonization and longterm community dynamics in ponds to examine whether land scape variables influenced the initial colonization of 22 constructed ponds (ten constructed and ten reference ponds) over seven breeding seasons. Results indica ted that the constr ucted ponds supported significantly more species than reference ponds. In addition, the authors concluded that post-restoration monitoring for 2-3 years may be sufficient to char acterize species and communities that will utilize ponds for the first decade or so after pond creation. While soil properties have not been traditionally used to measure wetland restoration success, Whited et al. (1999) recently suggested that, in addition to return of hydrology, plant communities, and wildlife use, some soil properties could be used as criteria for assessing wetland restoration. Nair, Graetz, Reddy and Olila followed in 2001 with a study designed to compare created wetlands to native wetlands in phosphatemined areas from central and north Florida us ing soil development analyses. The criteria selected for evaluation of soil samples fr om the 184 sites included soil compaction, bulk density, organic matter (carbon) and nitrogen content, the carbonÂ–n itrogen ratio, and available and total nut rient contents. Based on the a bove-mentioned parameters, results indicated that the reclaimed wetlands were sl owly developing into Â“typicalÂ” wetlands. Nair et al. (2001) mentione d that the rate of deve lopment could possibly be increased by minimizing soil compacti on, incorporating organic matter, or by
29 fertilization. Some of the substrates used in constructing a wetland in phosphate-mined areas include overburden, sand tailings, and/ or clay, which provide both mechanical support and growth media for plants (Nair et al., 2001). It is al so important to stress that the characteristics of soils in created wetla nds can influence the levels of available resources, which in turn will affect th e composition and diversity of above-ground vegetation and soil microflora and fa una (Chambers, Brown & Williams, 1994). The fact of the matter is, it is time to do something concrete, monumental and decisive. If when a wetland is reclaimed cer tain species survive while others do not, we still need to consider the succe ssful facets of the endeavor. We have to look at the value of mitigating wetlands and not get distracted by the faltering aspects which propagate the need for new, unproved innovations. Case Studies of Unsuccessful Wetland Mitigation Projects The reasons for reclamation shortcomings are numerous. Primarily, wetlands are hard to define. The physical and biologica l characteristics of wetlands are dynamic Â– plant boundaries can change in a season, water levels in hours. This lends to inconsistencies in terminology and interpreta tions. For example, reclamation success can have different connotations. Compliance succes s is determined by evaluating permitting requirements, whereas functiona l success is determined by a ssessing the restoration of ecological function to an altered ecosystem (Kentula, 2000). Historically, success was defined by individual projects, but ecologi sts are now considering the success of restoration on a landscape scale. Landscape success is a measure of how restoration (or management in general) contributes to the eco logical integrity of th e region or landscape and the maintenance of bi odiversity (Kentula, 2000). One of the greatest challenges is wetland creation, especially if it involves putting a wetland where one did not ex ist before. The primary quandari es in creation projects are bringing water to a site where it does not natu rally occur, and establishing vegetations on soils that are not hydric. While creation is po ssible, it typically re quires significantly more planning and effort than other restorat ion projects, and the out come is difficult to predict. Developers often atte mpt to convert uplands to we tlands, but their efforts result
30 in ecosystems that are distinct from natural wetlands in the area and provide limited wetland functions. Creating wetlands from ope n water is less difficult with respect to establishing a water source, but it often requires placing dirt or other fill into existing aquatic habitats, which means destroying one kind of aquatic habita t to create another. While this trade-off sometimes can be ju stified ecologically, the engineering and regulatory challenges of these projects are so complicated th at professional expertise and oversight are almost always required. The outcome of a creation and enhancemen t project is often difficult to predict because these projects essentially try to produce a new ecosystem. With restoration projects, results are more predictable, although there may still be uncertainty depending on the type of wetland, extent of degradati on, and many other factors. Under certain circumstances, creation or enhancement may be the best option but for the most part, restoration is more likely to have a posit ive outcome in terms of improving wetland resources. Another mitigation dilemma is the func tional equivalency of mitigated wetlands to natural reference wetlands. Mitigation ch anges the spatial distribution of wetland ecosystems, therefore Bedford (1996) advocates taking a landscape ap proach to defining hydrologic equivalence, so that potential cumu lative effects of wetla nd diversity can be identified. Mitigation alters natural config uration and spatial distribution of wetland systems not only in immediate areas, but al so over larger, geographic scales. Decisions regarding which sites to permit are often made in isolation, w ithout consideration of other projects or the potential ch anges to the wetland system. The majority of the nationÂ’s wetlands are privately owned, and landowners may manage their property in ways contradict ory to wetland preservation. Thus mitigation legislation should continue to encourage priv ate landowners to manage and protect their valuable wetland resource. Additional problems stem from the ideas th at wetlands are Â“lost.Â” In reality, the land itself remains although th ere has been a change in the complex nature of the functions and benefits the original wetland provided. Also, in academic arenas, knowledge and expertise about the resources are scattered and c onflicting. There are
31 differences in the reporting of the rate at which wetlands currently are being altered, the causes of these losses, ways to measure wetland function, and th e types of existing wetlands. Statistics of unsuccessful wetland mitigating focus on the acreage loss rather that the function omitted as a result of the loss. A measure of the damaged ecological functions would allow for a be tter insight into the wetlandÂ’s importance in a particular area. Some wetland functions can be mimicked with engineered structure, but engineered methods typically do not provide the maximum ecological benefit. For example, instead of re-establishing native vegetation on wetla nd edges to control erosion, a cement wall could be used to armor the bank. A cement wall could limit the erosion for a time, but it does not provide the other ecosystem benefits of wetlands, such as filtering pollutants and providing fish habitat (IWWR, 2003). Developing guidelines for wetland rest oration, creation, a nd enhancement is difficult because the goals of people undertak ing wetland projects vary widely and these goals influence the kind of activities best su ited to a particular site (IWWR, 2003). Erwin (1991) found that of the 40 mitigation proj ects in South Florida involving wetland creation and restoration, only about half of the required 430 h ectares of wetland had been constructed, and that 24 of th e 40 projects (60 percent) were judged to be incomplete or failures. Two of the contributing factors were improper water levels and incorrect hydroperiod. Some may consider the presen ce of non-native species in restored landscapes as unauthentic restoration. However, because the soils of mined lands have been altered and in most cases do not resemble those characteristics of native ecological communities, the likelihood of developing ecosystems with the same suite of species found in the native communities is diminished. Additionally be cause today there are more introducedÂ” species in the Florida landscape than in the past, it is probable that restored phosphate mined lands will contain many non-natives. For these reasons, if restored lands are not highly maintained, the presence of non-nati ves may be impossible to control (Brown, 2005). Petranka et al., (2003) conducted an 8-yr study to examine the demographic responses of the wood frog (Rana sylvatica ) and spotted salamander ( Ambystoma
32 maculatum ) to wetland creation at a mitigated bank in western North Carolina. Drought and outbreaks of a pathogen (Ranavirus) we re the primary causes of low juvenile production from 1998 to 2002, but despite site perturbations, the br eeding population of A. maculatum remained relatively stable. This study shows the need to be prepared for non-success as a result of natural phenomena, and emphasizes that even natural events can impact populations. Additional factors that can limit the su ccess of a mitigation project include: The need for consistent definitions; The high spatial and temporal va riability of project sites; The limited number of reference wetlands in the area available for comparison. In addition, comparisons are 1:1 regardless of similarities or lack thereof between reference sites; The limits of power of any statis tical comparisons due to sampling inconsistencies; Insufficient replication; Land managers looking for an Â“endpoi ntÂ” and not acknowledging that endpoints can vary; No pre-alteration comparisons of site; and Natural disturbances or stress levels th at may preclude functional equivalency (as with Petranka et al., 2003). Though we strive to discontinue anthr opogenic alterations of our nationÂ’s wetlands, unacceptable, natural losses in many parts of the country may persist. For example, LouisianaÂ’s coastal wetlands cont inue to dwindle because of flood control systems built along the Mississippi River du ring past decades. Effective wetland management policies of the future will have to address potential activities, as well as the activities of the past which have at tributed to continual wetland loss. In Florida, phosphate mining has affected both the quantity and quality of local wetlands. The following chapter will investigate the mining process and reveal how our need for phosphate is conflicting with our desi re to preserve environmental health.
33 Chapter 2: Phosphate Mining and Wetland Alteration Introduction Phosphate mining occurs in Idaho, Louisian a, Mississippi, North Carolina, Texas, Wyoming and Florida, but singularly the St ate of Florida provi des approximately 75 percent of the nationÂ’s supply of phosphate fe rtilizer and about 25 percent of the world supply (FIPR, 2004b). Most of the phosphate mi nes in Florida are located in the northern and central portion of the state, which include an area called Â“Bone Va lleyÂ” that stretches from Polk County southward through portions of Hillsborough, Manatee, Hardee, Sarasota and DeSoto counties (FIPR, 2004b). About 95 percent of the phosphate rock recovered in Florida is used in agriculture (90 percent is used to produce chemical fertilizer and 5 percent is used as feed s upplements for livestock). The remainder is added to a variety of consumer products incl uding soft drinks, toothpaste, bone china, film, light bulbs, vitamins, flame-resistant fabr ics, optical glass and shaving cream (FIPR, 2004b). Chapter 2 will describe the phosphate mining process and the subsequent environmental repercussions. This chapter will also illustrate how the mining process affects wetland ecosystems, and how the phosphate companies are participating reclamation. The Phosphate Mining Process The local geology plays a major role in mine location; mines are typically situated in areas where the commodity is shallow e nough to economically extract (FDEP, 2005). There are four types of surface mining Â– open-pit mining, strip mining, mountaintop mining and dredging. Open-pit mining involves extracting rock or minerals from the earth from an open pit or borrows. In Florida, phosphate is mined using this technique. The following map (Figure 4) illustrates the dist ribution of mines in Florida. The colored
34 dots represent different mineral commodities, and each dot may represent more than one mine. Figure 2.1: Distribution of Mines in Florida (FDEP, 2006c) Mine development involves: (1) prospectin g and exploring to lo cate and delineate the ore resource; (2) economic, environmental and technical feasibility assessment of the ore body; (3) planning and design ing the mine layout, site infr astructure and the mining sequence; (4) obtaining relevant government permits and approvals; and (5) constructing and commissioning the operation. Phosphate ore is found 15 to 50 feet beneath the ground in a mixture of phosphate pebbles, sand and clay known as the phosphate matrix. Dragline excavators (draglines) remove approxi mately 15 to 20 feet of overburden (soil on top of the matrix), pile it in nearby spo il piles, and transport the underlying phosphate matrix to shallow containmen t areas (FIPR, 2004b).
35 Next, high-pressured water is used to make a slurry, which is sent through pipes to the benefaction plant. The beneficia tion (or concentration) process removes contaminants and barren material prior to further processing. The naturally occurring impurities contained in phosphate ore depend heav ily of the type of deposit (sedimentary or igneous), associated minerals, and the ex tent of weathering. Impurities can include organic matter, clay and other fines, sili ceous material, carbona tes, and iron bearing minerals (International Fertilizer Indus try Association and the United Nations Environment Program [IFA-UNEP] 2001). Iron minerals may be present in the form of magnetite, hematite and goethite, and are typically removed through scrubbing, size classification, or magnetic se paration. The con centration of thes e characteristics influences the beneficiation pr ocesses used (IFA-UNEP, 2001). At the beneficiation plant, the slurry is moved thr ough a series of vibrating screens that physically separate the cla y, sand and pebble-sized particles. Other beneficiation processes include crushing/gr inding, water washing, and hydrocyclones. The clay particles are then pumped through pi pelines into storage ponds (clay settling ponds) where they sink to the bottom. The sand and phosphate concentrate are moved to a flotation plant where reagents are added to separate the two. The sand-tailings are used to back-fill the mined areas, and the phosphate concentrate is sent to dewatering tanks (FIPR, 2004b). Following beneficiation, the con centrated phosphate rock is converted to phosphate fertilizer in chemical plants, producing a gypsum by-product that is stacked near the plant. After the mining process ha s been completed, the final land uses include reclaimed land (about 50 60 percent of the landscape), clay settling areas and chemical plants (40 percent of landscape), and tran sportation lines and gypsum stacks (about 10 percent of the land). The mining proce ss is illustrated below (Brown, 2005).
36 Figure 2.2: Schematic Diagram of P hosphate Mining Process (Brown, 2005). The Environmental Repercussions of Mining Almost every facet of phosphate mining affects the native environment. Extraction and beneficiation change the natural landscape by re moving topsoil and vegetation, excavating and depositing over burden, disposing processing wastes, and underground mining-induced surface subsid ence. Surface and groundwater may be adversely affected by the release of proce ssing water, the erosion of sediment, the leaching of toxic chemicals from overburde n and processing wastes, and by dewatering operations. Intensified local ai r pollution may result from the release of emissions, such as dust and exhaust gases (IFA-UNEP 2001). In their 2001 report, Â“Environmental Aspects of Phosphate and Potash Mining,Â” the International Fe rtilizer Industry and United Nations Environment Program (IFAUNEP) outlined the following additional impacts: Changes in air quality by emissions of: o Dust; o Exhaust particulates and exhaust gases such as carbon dioxide (CO2); carbon monoxide (CO), nitrogen oxides (NOx), and sulfur oxides (SOx);
37 o Volatile organic compounds (VOC s) from fueling and workshop activities; and o Methane released from some geological strata. Changes in water quality by: o The release of slurry brines and contaminants into process water; o The erosion of fines from disturbe d ground such as open-cut workings; overburden dumps and spoil piles and waste disposal facilities; o The release or leakage of brines; and o The weathering of overburden contaminants, which may then be leached. Changes in social goods and intangible va lues (such as community lifestyles, land values and the quality of the ecosystem in the vicinity of the mine site) by factors such as: o Modification of the landscape; o Noise and vibration from activities su ch as blasting and the operation of equipment; and o Changes in wildlife habitat (IFA-UNEP, 2001). In Florida, phosphate mining began in the mid 1960s. In 1990, there were 11 phosphate companies operating in Florida; by 2004, as a result of the changes in ownership and corporate buyouts, there were three (FIPR, 2004b). Each company has numerous mines, ranging in size from 4,500 acres to about 21,000 acres with the average of about 10,000 acres. Currently, there are tw o active mining areas in Florida known as the Northern and Southern Phosphate Dist ricts, where about 5,000 acres of land are mined each year (Brown, 2005). The dominant producer in todayÂ’s Bone Valley is the Mosaic Company, created in 2004 when IMC Global combined with Ca rgillÂ’s crop-nutrition business (Hartley, 2005). Cargill had been a relative newcomer to Florida, buying several large phosphate operations in Bone Valley and the Tamp Bay ar ea in the 1990s. Mosaic also continues to mine former IMC properties at Hopewell and Kingsford. Other than Mosaic, Bone Valley's contending producer is CF Industries (Hartley, 2005).
38 As the scale and extension of operations grow, concern continues. The raw phosphate is essential for sustaining the worl dÂ’s agricultural output As there is no synthetic form of phosphate, the only supply is from natural deposits located beneath the earthÂ’s surface in certain parts of the worl d. In general, the phosphate companies in Florida have a good record of reclaiming th eir strip-mined lands. IMC-Agrico (now under The Mosaic Company) has won several environmental aw ards for their reclamation projects. However, a series of catastroph ic incidents involving several companies has raised doubts about the phosphate in dustry's environmental image. In June 1994, a 15-story-deep sinkhole ope ned in an 80 million-ton pile of phosphogypsum waste (known as a gypsum stack ) at IMC AgricoÂ’s New Wales plant (Satchell, 1995). The cave-in dumped 4 to 6 million cubic feet of toxic and radioactive gypsum and wastewater into the Florida aquife r, which provides 90 percent of the StateÂ’s drinking water. The company has voluntarily spent $6.8 million to plug the sinkhole and control the spread of contaminants in the groundwater (Satchell, 1995). Since 1990, at least 6 dams in Florida ha ve failed, three of them since October 1994 (Satchell, 1995). Billions of gallons of effluent have inundated nearby land, polluting streams and killing fish and other aquatic life. The biggest spill occurred in October 1985, when an IMC-Agrico dam burst and released 1.8 billion gallons of water. A 482 million-gallon torrent engulfed H illsborough County, and two people nearly drowned (Satchell, 1995). In December 1997, a breach of a phosphogypsum stack owned by the Mulberry Cooperation resulted in the rel ease of 50 to 60 million-gallons of toxic wastewater into the Alafia River in Polk County, causing massive aquatic deaths (Roth, Novey & Regalado, 2005). The IFA-UNEP maintains th at the mining industry has improved over the decades, although they acknowledge that challenges remain. Odum et al. (1983) documented the inte ractions between th e phosphate industry and wetlands on several sites in and around the Northern and Ce ntral Phosphate Districts. The authors evaluated landforms created by mi ning, the use of detritus as a seed source, tree planting on waste clay sites and the use of tree cores from cypress trees to monitor stress in wetlands. The first section of this report demonstrates how data concerning tree
39 growth can be used to design reclaimed la ndscapes and reveals that the major factor controlling vegetational succession appears to be proximity to natural seed sources (Odum et al., 1983). Rushton (1988) followed with a study fo cusing on the wetland establishment on waste clay sites and documented the following results: a pattern of arrested succession emerged that seemed to be related to the dist ance to seed source; the soils characteristic of mined lands tended to have higher concentrat ion of clay than native soils, and thus had lower air and water movement; and soils and water in phosphate mined area were higher in phosphorus and tended to naturally favor low diversity ecosystem types characterized by early successional colonizing species. Wetland Alterations from Phosphate Mining Wetland ecosystems are defined by tw o aspects; (1) abiotic conditions (temperature, soil type, terrain, disturbance re gime, etc.) and resource s (water, nutrients, sunlight ) and (2) the trophic structure or f eeding relationships among the species that shape the flow of energy through the ecosystem (Brown, 2005). An alteration to a wetland system may disrupt the abiotic components (s oil structure might be changed, as with mining) as well as the trophic structure (the loss of primary pr oducers, as when land is cleared). Several authors have suggested that in the past 200 years between 30 Â– 50 percent of wetlands within the conterminous United States have been lost, and many of the remaining wetlands are degraded (Bingham et al., 1990; Dahl & Johnson, 1991; Kentula, 1992). Wetland alterations result from bot h natural and anthropogenic phenomena. Indirect impacts from pollutants, urban r unoff, and invasion by non-native species can degrade wetlands as effectively as the direct impact of an excavat ion. Alterations can occur as discrete events (i.e. excavation activi ties, drainages, spills, fire and clearing) or transpire gradually (changes in hydrology, water level, sediment loads, concentration of chemicals, or the composition of biological communities). These alterations may induce subtle changes to the ecosystem or may devastate the landscape.
40 Wetlands altered by chemical contaminati on or biological changes may appear functional because the alteration may not affect the wetlandÂ’s physical appearance, but in reality the system has been affected. Though alterations to a we tland are localized, the cumulative effects of small-scale wetland alterations can have regional or national effects (Bingham et al., 1990). The entire surface mining process ravages the natural environment both directly and indirectly. Direct effect s include changes in water qua lity and sediment deposition, in addition to destruction of we tland habitat. Indirect alterations can consist of alteration to the natural hydrology or the release of se diment and particle-lad en waters into the wetlands and waterways (Kusler & Kentula, 1990). Mining may disconnect wetlands from natural water sources or from outfall structures which discharge excess water out of the wetland (Bedford, 1996). Increased turb idity within the wa ter column due to discharge or spills dur ing mining may affect the amount of light reaching vegetation below the surface. Increases in turbidity may result in mortality of the aquatic plants which provide habitat and food for other organisms (Kusler & Kentula, 1990). The natural landscape suffered as a result of years of mining, until laws were passed to rectify the situation. Beginning in 1899 and cu lminating in the 1970s, federal, state and local laws came into effect which required the reclamation of impacted wetlands. Although the land reclamation pro cess has been outlined and detailed, the increasing amounts of reclamation literature suggest that these guidelines are outdated and lack enforcement. Wetland restoration and creation are rela tively new fields; few engineers are trained in ecology and few ecologists have e xperience in engineering methods (Mitsch & Wilson, 1996). Failures are attributed to lack of understanding basic yet dynamic principles of wetland ecology. A need for a landscape point of view has prompted my investigation of internationa l mining and how different count ries are addressing wetland reclamation.
41 Chapter 3: International Mining and Wetland Preservation Introduction The environmental repercussions of phosphate mining and wetland reclamation are not only pertinent to central Florida. Of the numer ous countries producing phosphate, the United States produces 44,600 metric tons (FIPR, 2004b). Extensive transportation and industrial infrastructure s, which enhance the producti on and exportation of the product, allow the United States to genera te this enormous amount. Morocco and China rank a close second and third, respectively. MoroccoÂ’s phosphate reserves are estimated to be nearly six times that of the United States, which may indicate a potential shift in production power in the future. Additional active phosphate mining occurs is Russia, Tunisia, Jordan, Brazil, Israel South Africa, Syria, Togo, Senegal, Australia and Canada (FIPR, 2004b). All of the phosphate mini ng countries suffer the envir onmental ramifications of surface mining and wetland degradation. Laws and programs aimed at specific areas and in specific parts of the country do not bene fit wetland ecosystems on a landscape level. In order to avoid piecemeal solutions to a wide -spread problem, my second objective it to investigate international phosphate mining to examine how different countries and international organizations are addressing th e environmental implications. Chapter 3 will investigate the international environm ental and mining communities and will offer some international solutions to wetland preservation. The International Environmental Community Wetlands are as varied internationally as locally. Table 3.1 outlines the range of natural habitats within different temp erate and tropical wetland ecosystems.
42 Table 3.1: Range of Natural Habitats within Different Temperate and Tropical Wetland Ecosystems (Bacon, 1997). Climate (Representative Country ) Wetland Ecosystem Present Rock bottom Unconsolidated bottom Aquatic bed Rocky shore Unconsolidated shore Emergent marsh wetland Temperate Climate (such as The United States)1 Forested wetland Saturated forested wetland Tidally saturated forested wetland Permanently flooded emergent herbaceous wetland Seasonally flooded emergent herbaceous wetland Intermittently exposed unconsolidated shore Semi-permanently flooded aquatic bed Channels and pools Tropical Climate (such as Trinidad and Tobago)2 Marginal terra firma Stagnant brackish and hypersaline pools Drying up lagoons Tidal lagoons Soft tidal mudflats High Fiddler-crab zone of tidal mudflats Firm and tough clay banks Lower foreshore sandy beach Oceanic/Coastal Systems (such as Surinam)3 Back slope sandy beach 1Cowardin et al., 1979; 2 Bacon, 1988; 3Swennen & Spaans, 1985 Wetlands variability stems from seasonal fluctuations in rainfall or inundation patterns. The United States National List of Plant Species include s 6,728 species. Under more extreme conditions (such as the Ar tic Tundra, high mountain peat bogs and hypersaline salt marshes in the dry tropics) this diversity is lower, even though a range of highly specialized plants will often be present (Bacon, 1997). Wetland ecosystems are highly productive. The availability of water, which transports nutrients and removes waste produc ts, and the interaction between plant roots and microscopic organisms able to use nitroge n, allow wetland plants to grow rapidly and produce large quantities of or ganic matter. With tropical wetland plants, such as
43 mangroves, primary production can go on all year and reach levels comparable to the most intensively mechanized agricultural production (Bacon, 1997). Table 2 lists the productivity of various ecosystems. Table 3.2: Productivity of Select ed Wetland Ecosystems (Bacon, 1997). Wetland type Location Annual production, tons per hectare per year (above ground only) Estuarine mangrove Sri Lanka 12 Tidal salt marsh Louisiana, USA 14 Riparian forest Louisiana, USA 14 Freshwater (reed) marsh Denmark 14 Freshwater (Papyrus) marsh Kenya 30 Freshwater (reed) marsh Wisconsin, USA 34 Tropical seagrass bed Caribbean 70 The species diversity and hi gh production levels of wetla nd plants support very diverse animal communities. Wetlands of th e Ebro Delta in Spain support 48 resident species of fish, 29 mammals and 46 resident or migratory bird s (MAPA, 1991). Among the 104 species recorded in the Black River Morass, Jamaica, were 11 seabirds, 36 waterfowl, 7 birds of prey, a kingfish and 49 forest birds (Bacon, 1987). Approximately 251 species have been found in the Cache Ri ver Basin in Illinois, USA (FWS, 1994). According to Weigers (1990), 30 species of birds commonly breed in the somewhat restricted wetland forests in Western Europe. The livelihood and culture of large number s of people, in almost every country of the world, depend on wetland resources. A major portion of fisheries production, most hunting, much forest production and a significant part of ecotourism, as well as elements of heritage and environmental quality have evolved from nearby wetland ecosystems (Bacon, 1997). Twenty percent of commercial fish in Aust ralia are caught in mangrove resources, while 35 percent are food for commercial marine species (Robertson & Duke, 1987). Other commercial and economic uses of wetlands include: Fuel such as firewood, alcohol, charco al, wood (curing fish, smoking rubber and firing bricks), peat;
44 Medicines (from fruit, sap, bark, leaves) such as diuretics, purgat ives, astringents, vitamins (mainly b group), treatments for: arthritis, leprosy, catarrh, rheumatism, skin rashes, hemorrhoids, snake bite, tuberculosis; Agricultural, horticultural and aquaculture products such as fodder, fish feeds, green manure, peat/compost/fiber landscape plants, planting for coastal protection, ornamental pond plants, insect repellent (Bacon, 1997). It is important to stress, however, that it is not sufficient just to protect the populations of plants and animals that are dir ectly exploited Â– their health, survival, and sustainability depend on maintaining the whole complex of biodiversity that characterizes wetland ecosystems. The protection of such ecosystems is one of the goals of international environmental organizati ons such as the Ramsar Convention. The Convention on Wetlands of Intern ational Importance Especially as Waterfowl Habitat, signed in Ramsar, Iran in 1971 (referred to as the Ramsar Convention) is among the oldest global multila teral environmental agreements. Since its inception, its 138 Contracting Parties (encompa ssing at least 4 countr ies, including the United States) are committed to the conservation and wise use of all wetlands, to the designation and management of wetlands of in ternational importance (Ramsar sites), and to international cooperation in conserva tion implementation (Rodriguez, 2004). A variety of inland, coastal and near-shore mari ne wetland ecosystems are considered in these proceedings. The Convention owes its origins to c oncerns in the 1960Â’s over increasing destruction and degradation of wetlands, and the impact to migratory waterfowl (Hails, 1997). As such, approximately half of the curr ent Ramsar sites have been designated for the bird populations. The Convention works closely with non-governmental environmental organizations, scientific expert networks, water bird ecologist and general researchers. The four non-government organi zations that work closely with the Ramsar Convention, which are recogn ized as the Â“partner organizations,Â” are Birdlife International, the World Conservation Uni on, Wetlands Internationa l and the World Wide Fund for Nature (Pritchard, 1996).
45 Agencies such as the World Bank also play an important role in conservation efforts, and there are publications from th e Organization for Economic Cooperation and Development which contributes to the wealth of information being shared (Pritchard, 1996). In 1997, the International Institute for Environmental and Development produced a valuable directory of 150 sets of national, international an d sectoral guidelines, and at least twice this number is known to exist elsewhere (Roe, Dalal-Clayton & Hughes, 1995). Literature searches reveal that altho ugh the Ramsar Convention requires regular monitoring of wetland sites to detect changes in ecological character, the examples of successful long-term monitoring, especially in tropical wetlands, are limited. Bennun (2001) remarked that monitoring schemes run in to three kinds of diffi culties conceptual, logistical and political. The author empha sized that the sole goal on any monitoring scheme should be to assess deviation from the determined reclamation recommendation, so that appropriate corrective action can be taken. This assumes that adequate baseline information exists, and Bennun suggested that the global community does not have such data. Amezaga, Santamaria and Green (2002) ma intained that international wetland conservation approaches are based on the preser vation of isolated sites considered to be of special importance, thereby creating a Â“f ragmentedÂ” world view. The distribution of wetlands is critical to determining wetland co nnectedness even in the absence of direct hydrologic links. The authors hypothesized th at the effect of wetland loss on regional diversity might be much larger than previously expected, and thus call for more research linking local species richness. In a simila r paper, Amezaga and Santamaria (2000) concluded that the shortcoming in the curr ent management of European wetlands stem from the contradiction of creating highl y protected natural reserves surrounded by unprotected non-natural territory. Thus, as is suggested for the United States, a regional and landscape approach to wetland manage ment (within a contin ental context) was recommended. Some countries have attempted to use trade agreements to force less environmentally-advanced countries to deve lop wetland protection po licies, especially
46 when they are close enough to have a large impact on one another. An example is the North American Free Trade Agreement (NAF TA) between Canada, United States and Mexico (Marshall, 2000). While the three pa rticipating parties were negotiating NAFTA, they also addressed environmental concerns that they all shared. Prior to this agreement, all three countries had deve loped other environmental agreements focusing on border resources that the parties shared. Unde r NAFTA, economic and environmental joint decisions were achieved. Another example of the success of trade agreements is the General Agreement of Tariffs and Trade (GATT) The intent of GATT is to promote fair trade between the membersÂ’ parties (Marshall, 2000). GATTÂ’s ability to limit impacts on the environment was tested when the United States tried to impose trade sanctions on Mexico in an attempt to enforce the Marine Mammal Protection Act (MMPA). Mexico was killing dolphins in their tu na nets; the issue resulted in two cases that were brought before the courts (Marshall, 2000). Relatively few large-scale inventories of the worldÂ’s wetlands exist because of the difficulties of spatial scale, associated cost, and multiple objectives that drive classification. Kingsford et al. (2004) eval uated the extent of wetlands across New South Wales in Australia. Results indicated approximately 5.6 percent (4.5 million hectare) of the area is wetland. Currently, conservation rese rves only protect 3 percent of this area. Several books have been written about the various wetlands all ove r the world. The authors discuss the wetlands in isolati on, however, with each re ferencing only their particular areas. Unlike thes e authors, the editors of Wetlands of the World bring together most of the basic information present on ma ny types of wetlands in various locales. The first volume of this work covers wetlands in Africa, Australia, Canada and Greenland, the United States, th e Mediterranean, Mexico, Papua New Guinea, South Asia and tropical South America. The second a nd third volumes include Central America, Western, Northern and Central Europe, Northe rn and Western Asia, the Middle East, and Indonesia, the Far East, and New Zealand (Whigham, Dyky jova & Hejny, 1993). Each chapter describes and discusses the geolog ic, geomorphic and hydrologic controls on wetland formation within the region; the distribution of wetlands; their flora, fauna and
47 ecological characteristics; human impacts; and recommendations for conservation and management of these wetlands. Some of the reviews are more complete th an others, and some of the factors are discussed more than others (Whigham et al., 1993). For example, the chapters on South Asia and Australia cover geology, geomor phology and hydrology only in sparse detail, while the chapters on the United States focu s almost exclusively on hydrological factors and ignore the geologic and geomorphic condi tions that produce the regional hydrology. The differences in emphasis reflect the diffe rences in political structure and wetland classification. Rather than attempt a single, unified classification, the authors subscribed to the wetland schemes and definitions put forth by each region, thereby making interregional comparison difficult. All of the authors did agree on a series of recommendations: Universal adoption of the Ramsar conve ntion for international conservation; More complete national wetlands inventor ies, with emphasis on remote sensing; Detailed assessments of human impa cts and introduced species; and The development of regional management and conservation plans that focus on wetland preservation and limiting their de struction (Whigham et al., 1993). International wetland organizations are al so addressing the lack of cohesive wetland literature. The World Wide Fund fo r Nature (WWF) Australia is conducting annual reviews of the quality of public envi ronmental reports released by companiesÂ’ signatory to the Australian Mineral Industry Code for E nvironmental Management (IFAUNEP, 2001). The Fund recognizes the importance of environmental reporting in communicating environmental and social perfor mance to stakeholders The reports are assessed using nine performance indicators, wh ich were selected to assist stakeholders using environmental reports as tools fo r assessing the environmental and social performance of the company. The stakeholders believe that environmental reports should reflect the companyÂ’s performance, define goals, address stakeholder concerns arising from operations and provide independent verifi cation of the contents (IFA-UNEP, 2001). The WWF emphasizes the importance of exte rnal third party ve rification to assess company performance. Other international organizations participating in wetland
48 research include Wetlands In ternational, which was formed in 1995 when three organizations joined to create a wetland alliance. Nations that previously did not embrace wetland conservation are now addressing wetland degradation socially and politically. Within the recent adoption of The Federal Policy on Wetland Conservation, the Government of Canada has decreed to achieve the following goals: Maintain the benefits derived from wetlands throughout Canada; Achieve Â“no net lossÂ” of wetland func tions on federal lands and waters; Enhance and rehabilitate wetlands in areas where the continuing loss or degradation of wetlands or their func tions have reached critical levels; Recognize wetland functions in resour ce planning, management, and economic decision-making with regard to all fede ral programs, policies, and activities; Secure wetlands of significance to Canadians; and Recognize sound, sustainable management prac tices in sectors such as forestry and agriculture that make a positive c ontribution to wetlands conservation while also achieving wise use of wetland resources (Rubec, 1992). Following suit, the Costa Rican government adopted a Policy for Wetland Protection, the first of the ki nd in Central America, which promotes the conservation and wise use of the countryÂ’s wetland ecosy stems (Union Mundial para la Naturaleza [IUCN], 2001). While some nations are coping with the environmental problems of wide-scale mining and looking ahead to reclamation project s, some countries are just beginning to recuperate. In Tanzania, soil s are of low quality and have deteriorated to a level that heavy fertilization is essent ial if they are to be productive (Mwalyosi, 1988). Large phosphate mines are locat ed at Minjingu Hill in Lake Manyara Catchment basin. The life span of the mine was to be between 10 to 15 years, but the low production may increase the life span to 50 years (Mwalyos i, 1988). A resource conflict has arisen between the mining industry, the indigenous people and the state o fficial because the mine is located near the wildlife conser vation areas of Manyara and Tarangire, and
49 disrupts the major dispersal route of the larg e herbivores of the Tarangire National Park (Mwalyosi, 1988). Further complications arise because the mi ning community is expected to increase exponentially, and residents are occupying th e mining area faster than accommodations and social services can be situated. The mi ne is expected to attract small businesses, shops, and hotels to cater to the growing population. The incr eased population will require more land for agriculture and liv estock, and land-use disagreements are inevitable. Meanwhile, the animals around the mi ning site are at risk from ecological and environmental changes introduced by the mining industry (Mwalyosi, 1988). Mwalyosi (1988) urges the development of an environmental monitoring team and recommends that other agencies, such as the Tanzania National Environmental Council and the State Mining Cor porations, cooperate to minimi ze the inevitable impacts. In addition, the author calls for an integr ated resource management plan that could evaluate the ecological and socioeconomic f actors associated with the rise of the international phosphate industry. The International Mining Community Currently, most phosphate rock production worldwide is extracted using opencast dragline or open-pit/shovel excavator mining methods. This method is employed widely in parts of the United States of Ameri ca, Morocco and Russia (IFA-UNEP, 2001). Underground mining methods are currently used in Tunisia, Morocco, Mexico and India. The area of land surface affected by these opera tions is generally small and limited to the area immediately adjacent to the access decline or shaft (IFA-UNEP, 2001). The mining of flat-lying thin orebodies, as found in Florida, affects a fa r wider area of land than the mining of thicker or steeply dipping or ebodies as found in Brazil and Idaho. Some international companies are begi nning to adopt life-cycle techniques of mining. The industry has moved from Â‘endof-pipeÂ’ solutions, towards a pollution prevention strategy. This strategy requires an integrated, holistic vi ew of activities (IFAUNEP, 2001). Tools have b een developed to assist ma nagement, including cleaner production, life cycle assessments and industria l ecology. Each aspect of the mine
50 development looks at the life cy cle of the product or servic e (see Figure 3.1), to identify where the major environmental issues or problems may arise, and where the most costeffective solutions can be deve loped. Planning for the life of the mine, including closure and site rehabilitation, allows for a more effi cient and environmentally effective outcome. As an example, WMC Australia participates in environmental reporting and records their input and output flow amounts, including emissions, wastes a nd net land disturbance, to document their environmental performance (IFA-UNEP, 2001). Figure 3.1: The Mining Life Cycle (IFA-UNEP, 2001). The concept for the schematic is drawn from the Natural Resources Canada report Â“Sustainable Development and Minerals and Metals.Â” The ISO 14000 Environmental Management System and the European Union Eco-Management and Audit Scheme (EMA S) standards provide frameworks and guidelines that companies can use to deve lop their own environmental management systems (IFA-UNEP, 2001). Many phosphate mining companies all around the world are looking at the ISO 14000 standard s to guide the development of systems that are specific to their needs.
51 In Jordan, the Jordan Phosphate Mi nes Company Ltd. attained the ISO 14000 Environmental Management System and IS O 9000 Quality Assuranc e certification for most of their operations (IFA-UNEP, 2001). Today, the company acknowledges the benefits these systems offered in increasi ngly more demanding consumer markets. In Brazil, Fosfatados Fertilizantes adopted the Â“P rogram 5SÂ” quality management system in 1995. This system, which originated in Japa n, is based on the five principles of: Seiri Â– Sense of selection; Seiton Â– Sense of tidiness; Seisou Â– Sense of cleanliness; Seiketsu Â– Sense of health; and Shitsuke Â– Sense of self-discipline. Based on these principles, Fosfatados Fertilizantes developed the following company objectives for improving their work environment: (1) prevent accidents; (2) provide incentives for creativity; (3) elimin ate waste; (4) promot e team-work; and (5) improve the quality of service and products (I FA-UNEP, 2001). Success of this system is accomplished by changing behaviors at all levels of the company to improve the working environment, welfare and quality of the produc ts and services. The performance of the system is visible: the company system ha s won the Brazilian mining safety award for seven years running. Additionally, the economic performance of the company has improved in recent years (IFA-UNEP, 2001). Along with process improvements, financ ial compensation is offered to areas ravaged by historical mining activities. The Wellington mining company agreed to pay $8.5 million U.S. dollars in compensation to the Pacific island nation of Nauru for environmental damage caused by phosphate mi ning before 1968. It joins Britain and Australia in contributing to a compensation package that totals $76 million dollars (Â“Compensation Offered,Â” 1994). Currently, countries vary significantly in their approach to wetland monitoring and protection. Some developi ng countries have yet to establish wetland protection policies. Based on literature searches, it a ppears that compared to other nations, the United States has developed more advanced inventory and comprehe nsive study of their
52 wetlands. Even with an advanced inventory system, however, studies continue to show losses of United StatesÂ’ wetlands to be estimated at over 50 percent of what they were in the early 1600s. Canada and Europe have also experienced losses over 50 percent although neither has monitored wetland status as long or as thoroughly as the United States. Other parts of the world vary as to how many wetlands have been lost or degraded (Marshall, 2000). Most wetland successes originate from Ramsar participants. The Contracting Parties are required to report on their progress of implementing their commitments under the Convention by the subm ission of triennial National Reports to the Conference of the Contracting Parties. Th e National Reports then become part of the public record. Thus, internationally progr ess is being made. There does seem to be a common philosophy between our national approach a nd the international approach, but the international standards seem deficient. The international aren a does not have an international Clean Water Act for guideline unif ication, and countries need the equivalent of a global EPA to maintain some level of enforcement power. Treaties that may be binding under the United Nations (U.N.) Charter are not alwa ys adequate. Economic sanctions imposed through the U.N. are difficu lt given the wide range of opinions of the importance of wetland ecosystems. The fact that international guidelines are distinct from the United States is not unexpected. The hydrology, geology, climate, polit ical structure, social values and the value of wetlands are different depending on the country involved. Globally, however, the overall lack of wetland preservation is astonishing. The commitment to environmental protection and the capacity to set forth the necessary enforcement should be standard. It is difficult to suggest a cohesive form of environmental conservation. The framework of sound scientific analysis is nece ssary for appropriate legislation to develop. The myriad of unsuccessful wetland mitigati on points to a need for more comprehensive methods of evaluation. One such method has be en outlined by Danish scientists. In the 2002 National Environmental Research Institu te (NERI) report, Â“Microorganisms as Indicators of Soil Health,Â” Nielsen and Winding (2002) promote the use of microbial
53 indicators in terrestrial m onitoring programs, and provide recommendations for such indicators to be implemented into a Danish terrest rial monitoring program. In addition, the European UnionÂ’s COST Action 831 (www. isnp.it/cost/cost.htm), a cooperative project by international scientis ts, is encouraging the use of microorganisms as indicators of environmental impacts in soil monitori ng. A European Monitoring and Assessment Framework on soil has subsequently been proposed to provide policy-makers with relevant information on soil, an d to bring together the wealth of soil information derived from current national soil monitoring programs (Nielsen & Winding, 2002). Microorganisms respond quickly to change s, hence they can rapidly adapt to environmental conditions. The microorganisms that are best adapted, flourish. This adaptation allows microbial analyses to be discriminating in soil health assessment, and changes in microbial populat ions and activities may ther efore function as excellent indicators of changes in soil health (Kenne dy & Papendick, 1995; Pankhurst et al., 1995). Microbes also respond quickly to environmental stress as they have intimate relationships with their surroundings, due to their high surface to volume ratio (Nielsen & Winding, 2002). In some instances, changes in micr obial populations (or activity) can precede detectable changes in the physic al and chemical properties of the soil, thereby providing an early sign of soil improvement or an ea rly warning of soil degradation (Pankhurst et al., 1995). Because microorganisms are involved in many soil processes, they may also give an integrated measure of soil healt h, an aspect that ca nnot be obtained with physical/chemical measures alone. In terms of benefits for vege tation, their role in binding sediment particles to im prove rooting medium structure is critical (Kadlec & Knight, 1996). Microbial presence in the r oot zone of plants is also necessary for sustained plant growth. The activities of microbial communities in sediments are important for sediment quality, and the biological transformations of major nutrients infl uence plant growth (Kadlec & Knight, 1996). Soil microorganisms also affect the physical properties of the soil, such as the water holding capacity, infiltration rate, erodibility, and susceptibility to compaction. Production of extra-cellular polysaccharides and other cellular debris by microbes maintains soil structure, as these ma terials function as cementing agents that
54 stabilize soil aggregates (N ielsen & Winding, 2002). Microbe s also contribute to the enhanced biotransformation of toxic chemi cals in wetlands (Daane, Harjono, Zylstra and Haggblom, 2001). Given the multiple functions of soil micr obes (e.g., soil formation, toxin removal, contributions to the elemental cycles), it is difficult to identify any one property as a general indicator of soil heal th. Instead, Nielsen and Wi nding (2002) suggest using Â“end pointsÂ” to determine the soil ecosystem functions of interest (Table 3.3). Table 3.3: End Points of Terrestrial Monitoring and Corresponding Soil Ecosystem Parameters (Nielsen & Winding, 2002). End Point Soil Ecosystem Parameter Atmospheric balance C-cycling Soil ecosystem health Biodiversity C-cycling N-cycling Microbial biomass Microbial activity Key species Soil microbial community health Biodiversity Ccycling Microbial biomass Microbial activity Bioavailability Leaching to groundwater or surface run-off N-cycling Bioavailability Plant health N-cycling Key species Animal health Microbial biomass Bioavailability Human health Key species Bioavailability Microbial indicators of soil health cover a diverse set of microbial measurements, reflecting the multi-functional properties of mi crobial communities in the soil ecosystem. In the NERI report, microbial indicators incl uded bacteria, fungi a nd/or protozoa. The authors grouped the indicators according to different soil ecosystem parameters (identified in Table 3.3), and this assembla ge is presented in Table 3.4 (Nielsen and
55 Winding, 2002). These tables can be used to specify the indicator and parameter best suited to a particular mitigati on site. Generally, indicators sh ould be selected on the basis of ease of measurements, reproducibility, a nd their sensitivity towards the variables identified as pertinent to soil health for the selected area. Th e selection of indictors can be broad to give managers an overview of the wetland system, or detailed to better explain underlying trends. For example, if biodive rsity was chosen as the parameter for measuring microbial community health, four techniques (Table 3.4) can be used. This list is not complete, but offers a large number of techniques that ar e presently used.
56 Table 3.4: List of Microbial Indicators for Soil Health Monitoring (Nielsen and Winding, 2002). Soil Ecosystem Parameter Microbial Indicator Methods Genetic diversity PCR-DGGE Functional diversity BIOLOG Marker lipids PLFA Biodiversity Soil respiration CO2 production or O2 consumption Soil enzyme activity Enzyme assays Methane oxidation Methane measurements Methanotrophs MPN PLFA C-cycling Heterotrophic activity CO2 evolution N-mineralization NH4 + accumulation Nitrification NH4 + oxidation assay Denitrification Acetylene inhibition assay N-fixation: Rhizobium Pot test N-cycling N-fixation: Cyanobacteria MPN Nitrogenase activity Microbial biomass: Direct methods Microscopy PLFA Microbial biomass: Indirect methods CFI,CFE SIR Microbial quotient Cmicro / Corg Fungi PLFA or Ergosterol Fungal-bacterial ratio PLFA Protozoa MPN Bacterial DNA synthesis Thymidine incorporation Bacterial protein synthesi s Leucine incorporation Community growth physiology CO2 production or O2 consumption Bacteriophages Microscopy for biomass Microbial biomass Mycorrhiza Microscopy Pot test Human pathogens Selectiv e plating and/or PCR Suppressive soil Pot test Key species Biosensor bacteria Remedios Microtox Plasmid-containing bacteria Gel electrophoresis Antibiotic-resistant bacteria Selective growth Bioavailability Incidence and expression of catabolic Genes Selective growth
57 Wetland soil analysis is also being achie ved in other countries. Soil microbial parameters are being developed for the Domi ngos mine, an abandoned cupric pyrite mine in Southeast Portugal. A study designed by Pereira, Sousa, Ri beiro and Goncalves (2006) was initiated to evalua te the long-term effects of heavy-metal contamination on microbial community activity, and subse quently on important soil functions (e.g., nutrients cycling, decomposition of organic matter). Total metal contents, as well as physical and chemical parameters (organic matter moisture, pH a nd conductivity), were measured. The most sensitive microbial pa rameters were dehydroge nase activity (DHA) and potential nitrification (POTNI), which we re highly depressed in the milling area. Results suggested that the high metal contamin ation levels and low pH adversely affected the biomass and the activity of soil microorgani sms in the mining area, and thus affected the development of the wetland ecosyst em (Pereira et al., 2006). In their study of bacterial community structure in Scotland, Ellis, Morgan, Weightman and Fry (2003) inve stigated soil samples taken from sites with patchy metal contamination, and assessed dive rsity with a variety of appr oaches. The authors contend that the most appropriate methods for determ ining the effect of the contamination on soil health are assessing microbial activity, rather than the presence or absence of different cell types. Although culture-depen dent approaches to microbi al community analysis are often criticized for their selectivity, it is perhaps this discrimination that makes them useful for determining the impact of anthropogenic activity (Ellis et al., 2003). The field of soil microbiology is relativel y new, but the advances made and the plethora of laboratory technique s available is impressive. Th e next chapter will address the potential of microbial analysis for mitigated wetlands in the State of Florida.
58 Chapter 4: Soil Microorganisms as I ndicators of Functional Wetlands Introduction Hydric soils promote a very complex asse mblage of microorganisms, which break down and transform a wide variety of substances. Wetland vegetation provides attachment sites for the microorganisms th at make up the microbial communities, and when these plants die and accumulate as li tter, they create additional material and exchange sites, as well as sources of car bon, nitrogen and phosphorous to fuel microbial processes (Paul & Clark, 1989). Considerable interest has been focuse d on the functional assessment of wetlands, especially those created as a re sult of state mitigation policy. Several studies exist on the hydrology (Bedford, 1996; Niswander & Mits ch, 1995), seed banks (McKnight, 1992), vegetation ecology (Reinartz & Warne, 1993), microorganisms and mine waters (Ledin & Pederson, 1996), and wildlife (Leschisin, Williams & Weller, 1992) of mitigated wetlands, but comparable data is lacking fo r wetland soil microbiologyÂ’s participation in mitigation success. Our understanding of the complex processes involving plants, microorganisms, soil matrix, and how these factors of the ecosystem interact is incomplete. The literature suggests several reasons fo r our inability to successfully mitigate wetlands. The Â“successfulÂ” one s are those that remain part of the landscape for long periods of time, but these wetlands may not be healthy or properly developed. Recent hydrologic data shows that created wetlands ar e not achieving structural replacement, a minimum requirement for functional success (Cole & Brooks, 2000). Galatowitsch and van der Valk (1994) outlined provisions for the term Â“functional success.Â” Thus, although created and mitigated wetlands may me et jurisdictional requirements, measures of their ecological Â“healthÂ” do not suggest a thriving system.
59 My intent is to supplement this review by suggesting the i nvestigation of the microbial communities of mitigated wetlands. I will review literature describing the laboratory and field procedures necessary to evaluate the proper development of mitigated wetlands. This chapter will provide managers with a source of references that can serve as a framework for future wetland m itigation projects, and will offer effective tools for mitigation monitoring (i.e. comparing the microbial activity of their sites to natural wetlands). Chapter 4 will discuss wetland soil characteristics (preand post-mining) and evaluate the role of microorganisms as bioi ndicators in the mitigation process. In addition, research projects designed to study the microbial aspects of wetlands will be summarized. Hopefully, in the near future, a microbial standard can be developed and applied to various national sites to suppor t the equivalence of mitigated wetlands and natural wetland systems. This review can also help policy-makers establish an inference of wetland success. Wetland Soil Characteristics Wetland soil development is critical for the re-establishment of native vegetation after phosphate mining. Some of the substrat es used in construc ting a wetland in phosphate-mined areas include overburden, sand tailings, and clay (Nair et al., 2001). Reclamation of phosphate-mined lands usually begins on the clay settling areas. There are an estimated 160,000 acres of clay settling areas in Cent ral Florida (Stricker, 2000). Clay soils are problematic for wetland development, however, because they have: Highly compacted soils due to th e bulk density of clays; A lower amount of free oxygen, which is necessary to support the microorganism community; Elevated pH levels (very high levels of calcium); Very little soil organics (~0.5pe rcent of soil organic carbon); Severe deficiencies in nitrogen available to plants; Soil structure instability; Extreme seasonal variations in top-zone moisture;
60 Relative sterility because of the lack of native flora seed banks; and Poor drainage (Stricker, 2000). On a positive note, clay settling area so ils are very high in potassium (K) and phosphorus (P) two key components of productive soils (Stricker, 2000). The following table describes the differences in mineral c ontent and pH of sandy so ils and clay settling area soils in Florida. Table 4.1: Comparison of Florida Sandy So il and Phosphatic Clay (Stricker, 2000). Soil Component Sandy Soil (ppm) Phosphatic Clay (ppm) Calcium 63 5,923 Magnesium 10 2,569 Potassium 67 332 Phosphorus 166 586 Zinc 1 5.6 Manganese 0.3 6.3 In a soil profile (Figure 4.1), the capital letters O, A, B, C and E identify the master horizons, and the lowercase letters ar e used for further distinctions of these horizons (United States Department of Agri culture National Res ources Conservation Service [NRCS], 2006). Most soils have three major horizons -the surface horizon (A), the subsoil (B), and the substratum (C). So me soils have an organic horizon (O) on the surface, but this horizon can also be buried. The master horizon, (E), is used for subsurface horizons that have a significant lo ss of minerals (eluviation). Hard bedrock, which is not soil, uses the letter R (NRCS, 2006).
61 Figure 4.1: A Typical Soil Profile (United St ates Department of Agriculture National Resources Conservation Service [NRCS], 2006). In the field, soil color is evaluated by th e hue and chroma content. The color is derived from the iron and manganese present, and depends on the soilÂ’s current reduced (redox) state. The term redox comes from the two concepts of reduction and oxidation. O xidation describes the loss of an electron by a molecule, atom or ion, whereas r eduction describes the gain of an elec tron by a molecule, atom or ion (Maier et al., 2000). The U.S. Soil Conservation Service offers the follo wing physical soil tr aits (Environmental Sciences and Resour ces [ESR], 2006): Mineral Matter (50 percent of the soil composition) o Derived from parent rocks; o Solid framework of the soil; o Contains inorganic material; o Can appear mottled or gleyed: Mottled soils have a matrix of one color with blotches and flecks of other colors, and occur where the water table fluctuates and soils may only be saturated for part of the year. In mottled hydric soils, the matrix color is usually grayish
62 (chroma of 2 or less), and the mo ttles are usually red, brown or yellowish; Gleyed soils are usually light grey with a bluish or greenish tint (chroma of 1 or less). They o ccur under conditions of long term saturation where essentially all the iron and manganese are reduced. Organic Matter (5 percent of the soil composition) o Contain carbonaceous substances; o Made of organic compounds produced by current/past metabolism in soil; o Include remains of living organisms; o Types of organic soils: Saprists (muck) comprised of 2/3 decomposed material and 1/3 plant fibers; Fibrists (peat) comprised of 1/3 decomposed material and 2/3 plant fibers Hemists (muck/peat) conditions be tween saprists and fibrists Air (20 percent of the soil composition); and Water (25 percent of the soil composition). Doran and Safely (1997) consider soil h ealth essential to sustaining biological productivity, promoting the quality of air and water environments, and maintaining plant, animal and human health. A healthy soil functions as an environmental filter for removing undesirable solid and ga seous constituents from air and water. The extent to which a soil immobilizes or chemically alters substances that are toxic, thus effectively detoxifying them, reflects the degree of soil health (Doran & Safely, 1997).
63 Soil Microorganisms as Bioindicators Whether discussing evaluation techni ques or timeframes, land managers, legislators and ecologists disa gree on the most effective way to evaluate the functional equivalency of mitigated wetlands. One of the reasons for unsuccessful mitigation is inadequate monitoring. Si ngular evaluations of plant species, bioaccumulations of contaminants, sedimentation rates, population de mographics, habitat structure, acreage or hydrology do not address the complexity of the whole wetland ecosystem. Rather these factors segregate the wetland and address mitigation in a piecemeal fashion. Traditionally, wetland communities were anal yzed by indices or other analytical metrics. Examples of such techniques incl ude: (1) Similarity (Comparative) Indices; (2) Cluster Analysis and Ordination; (3) Food Web Analysis; (4) Tolerance Indices; (5) Guild Analysis; and (6) Indices of Bio tic Integrity (Adamus & Brandt, 2004). Similarity Indices are metrics that refl ect the number of sp ecies of functional groups in common between multiple wetlands or time periods. The metrics may be weighted by relative abundance, biomass, ta xonomic dissimilarity, or caloric content of the component species. These techniques include Jaccard coefficient, Bray-Curtis coefficient, rank coefficients, overlap indi ces, and the community degradation index (Ramm, 1988; Adamus & Brandt, 2004). Cluste r Analysis and Ordina tion are procedures that detect statistical patterns and associa tions in community data and can be used to hypothesize relationships to a stressor. Su ch techniques include principal components analysis, reciprocal averaging, de-trende d correspondence analysis, TWINSPAN, and canonical correlation (Pielou, 1984). Food Web Analysis measures the length of food chains, number of trophic levels, the ratio of the number of tr ophic species to tr ophic links, and similar measures (see Turner, 1988). As yet, this type of analys is has not been tested on stressed wetland systems. Tolerance indices are metrics that reflect the proportionate comparisons of tolerant versus intolerant taxa. Â“Tolerance, Â” in this case, means tolerance to organic pollution. Examples of tolera nce indices include saprobic indices, macroinvertebrate
64 EPT (Ephemeroptera, Plecoptera and Trichopt era) indices, and the Hilsenhoff index (Micacchion, 2003; Adamus & Brandt, 2004). The procedures for Guild Analysis involve assigning individual species functional groups (species assemblages) based on the similar facets of their: Life history; Habitat preference; Trophic level, assumed niche breadth; Size, biomass, caloric content; Toxicological sensitivity; Behavioral characteristics; Phenological characteristics; Sensitivity to human presence; Status as an exotic or indigenous species; Resident vs. migrant status; and Harvested vs. protected stat us (Adamus & Brandt, 2004). Finally, indices of biotic integrity are composites of weighted metrics, which describe richness, pollution tolerance, trophic levels, abundance, hybridization, and deformities, and are often used in stream fish studies (Karr, 1981). Decisions concerning selec tion of which resources, uses or functions to protect (or enhance) are inevitably complex, since the criteria for protecting one resource may be contradictory to protecting another (Adamus & Brandt, 2004). Section 404 of the Clean Water Act has outlined a generalized list of we tland functions or uses that might be the focus of protection or restorati on (33 CFR 320 (b) (2)), such as: Food chain production; General habitat; Research, education, and refuges; Hydrologic modification; Sediment modification; Wave buffering and erosion control;
65 Food storage; Ground water recharge or discharge; Water purification; and Uniqueness/scarcity. If these parameters have been identifie d, why look elsewhere for guidance? The answer lies in the continued unsuccessful m itigation projects. Although it may appear from the quantity of literature citations in this review that inland wetlands in some regions have been extensively studied, in r eality, relatively little is known about wetland biological response to anthropogenic stressors. Anthropogenic and/or natural stressors are often cumulative and interactive, thus complicating the development of individual indicators. Compared to the monitoring of streams and lakes, sampling wetlands on a recurrent or comparative regional basis ha s been limited, partly due to lack of government sponsorship of wetland biomonitori ng programs (Adamus and Brandt, 2004). Furthermore, the response of a wetland co mmunity to anthropogenic stress depends not only on the taxa present and th e severity of the stressor, but also on the geomorphic, physical, and chemical environment of the we tlands (Adamus, Clairain, Smith & Young, 1987; Adamus et al., 1991). In August 1988, EPAÂ’s Wetlands Research Program sponsored a workshop in Easton, Maryland, to identify organisms and metr ics that might be useful in indicating wetland ecological healt h. Findings were summarized in the EPA report, Â“Wetlands and Water Quality: EPAÂ’s Research and Monitori ng Implementation Plan for the Years 1989 1994Â” (Adamus, 1989). Although the repo rt acknowledged that there are numerous situations where traditional indicators of ecosystem or wetland function (e.g., plant species, bioaccumulation of contaminants, sedimentation rates, population demographics, and habitat structure) are cost-effective a nd can successfully reflect the ecological condition and sustainability of the wetland, the EPA report requested the synthesis of existing regional literature in ways that would allow candidate bioindicators to be identified and available data to be nume rically compiled (Adamus, 1989). One such bioindicator can be the mi crobial content of soil.
66 In a recent survey of 156 created we tlands on phosphate-mined lands, 83 were created on graded overburden, while 38 sites were created on a mixture of overburden and sand tailings. Wetlands created on sand tail ings alone constituted 8 sites, while 4 of the created sites used sand/clay as the substrate (Ervin, D oherty, Brown & Best, 1997). Construction information on the remaining wetla nds in the Ervin et al. (1997) survey was not available. Management of these m itigated sites included the monitoring of vegetation (53 sites), water quality (20 sites) macroinvertebrates (15 sites), wildlife (12 sites), and hydrology (10 sites). Of these 156 sites, only four sites had some soils monitoring and the monitoring protocols were no t discussed in depth (Ervin et al., 1997). None of these sites had microbial community sampling or analysis. More recently, Kelly and Tate (1998) studied the size, activity, an d structure of microbial communities from remediated and natural soils in the vicinity of a zinc smelter. Results of their work suggested that the microbial community may be a useful indicator of changes in soil quality. Based on a synopsis of the bioindicator literature (funded by the EPA), it was determined that indicators of ecosystem health should be: Socially relevant and eas ily understood as an indicato r of ecological integrity and/or health; Capable of evaluating the effectiveness of regulations, control, or management strategies; Able to give a maximum amount of information for a minimum cost, and thus be cost-effective and fiscally attractive; Capable of being generated from accessible data sources; and Capable of providing a warning in time to avoid widespread or irreversible damage of the resource (Adamus & Brandt, 1990). In order to incorporate these criteria and offer the greatest inference ability, a truly innovative approach is needed to evaluate the success of th e management of the wetland systems, and assess the overall mitigation pr ogram. Techniques are needed that can both quantitatively address the permit requirements, and provide evidence of how, and to what degree, the resource is changing. Thus, I offe r the following case studies to substantiate
67 the use of soil health and soil microbial co mmunities as indicators of thriving wetlands mitigation projects. Soil Microorganisms Types and Microbi al Community Analysis Techniques Types of Soil Microorganisms Microorganisms (Tables 4.2 and 4.3) contribu te to proper soil he alth as they are involved in many important f unctions such as soil formation, toxin removal, and the elemental cycles of carbon, nitrogen, phosphor ous, and sulfur (Borneman et al., 1996). Autotrophs (e.g., plants, algae, photosynthetic bact eria, lithotrophs, and methanogens) use carbon dioxide (CO2) as a sole source of carbon for growth, and reduce the molecule to organic cell material (CH2O). Heterotrophs require organic ca rbon for growth, and ultimately convert CH2O back to CO2 (Maier et al., 2000). Additional essential components of the soil microbial community are symbiotic arbuscular mycorrhiza fungi, which colonize the roots of the majority of terrestrial plants, and rhizospheric bacteria promoting plant growth, which inhabit the plant rhizosphere, the root surface and intercellular root spac es (Tate & Klein, 1985).
68 Table 4.2: Examples of Important Autotroph ic Soil Bacteria (Mai er et al., 2000). Organism Characteristics Function Nitrosomonas Gram negative, aerobic Converts ammonia (NH4 +) nitrite (NO2 -); 1st step of nitrification Nitrobacter Gram negative, aerobic Converts NO2 nitrate (NO3 -); 2nd step of nitrification Thiobacillus Gram negative, aerobic Oxidizes sulfur (S) sulfate (SO4 2-); sulfur oxidation Thiobacillus denitrificans Gram negative, facultative anaerobe Oxidizes S SO4 2(functions as a denitrifier) Thiobacillus ferrooxidans Gram negative, aerobic Oxidizes Fe2+ Fe3+ Table 4.3: Examples of Important Heterotr ophic Soil Bacteria (Maier et al. 2000). Organism Characteristics Function Actinomycetes (e.g., Streptomyces ) Gram positive, aerobic, filamentous Produce geosmins Â“earthly odor,Â” and antibiotics Bacillus Gram positive, aerobic, spore former Carbon cycling, production of insecticide and antibiotics Clostridium Gram positive, anaerobic, spore former Carbon cycling (fermentation), toxin production Methanotrophs (e.g., Methlosinus ) Aerobic Methane oxidizers that can co-metabolize trichloroethene (TCE) using methane monooxygenase Alcaligenes eutrophus Gram negative, aerobic 2,4-D degradation via plasmid pJP4 Rhizobium Gram negative, aerobic Fixes nitrogen symbiotically with legumes Frankia Gram negative, aerobic Fixes nitrogen symbiotically with nonlegumes Agrobacterium Gram negative, aerobic Important plant pathogen, causes crown gall disease
69 Microorganisms are essential to ecosyst em development and maintenance. Microbes affect the physical properties of th e soil (the water holding capacity, infiltration rate, erodibility, and susceptibil ity to compaction); the soil structure (the production of extra-cellular polysaccharides and other cellular debris help maintain soil structure, as these material function as cementing agents that stabilize soil aggreg ates); and initial soil development (Nielsen & Winding, 2002). After a site has been mined, the impor ted substrate (necessary for initial reclamation) may at first be devoid of or ganic carbon and nitrogen. The development of biota and soils on mitigation sites require: 1. Carbon-fixation from carbon dioxide (CO2) by photosynthesis; 2. Nitrogen-fixation by symbiotic or non-sy mbiotic micro-organisms, or slow accumulation by the retention of small amounts of fixed N deposited from the atmosphere; 3. Phosphorus, which generally is derived from the weathering of the substrate, and therefore is generally present in newly exposed material; and 4. Sulfur weathered from the substrate or accumulated from the atmosphere (Bolin et al., 1981). Microorganisms participate in the development and maintenance of the biogeochemical cycles of carbon (C), nitrogen (N), phosphorous (P) and sulfur (S). In soil habitats, micr oorganisms Â“fixÂ” carbon dioxide (CO2) from the atmosphere in the synthesis of carbohydrates, which constitute the carbon and energy used by the majority of organisms that do not phot osynthesize. Biodegradation is the reverse process the decomposition of organic material (CH2O) back to CO2, water (H2O) and hydrogen (H2). Fungi and prokaryotes (e.g. actinomycetes, clostridia, bacilli, arthrobacters and pseudomonads) play a significant role in bi odegradation. Decomposition involves the initial degradation of biopolymers (cellu lose, lignin, proteins, polysaccharides) by extracellular enzymes, followed by oxidation (fermentation or respiration) of the monomeric subunits (Maier et al., 2000). The end products include carbon dioxide (CO2), water (H2O), hydrogen (H2), ammonia (NH3) and sulfide (H2S). These products
70 are then used by lithotrophs and autotrophs, and all of these processes and products contribute to the carbon cycle (Figure 4.2). Figure 4.2: The Carbon Cycle (Maser, 2006) In the nitrogen cycle (Figure 4.3), ammonia (NH4 +) is oxidized to nitrites (NO2 -) by nitrosifying bacteria (Nitrosomonas and Nitrosococcus ). Nitrites are then converted to nitrates (NO3 -) by nitrifying bacteria ( Nitrobacter ) in a process known as nitrification. In denitrification, microorganisms reduce nitrates and nitrites to ni trogen-containing gases (Maier et al., 2000). Nitrogen fixati on is the process by which nitrogen (N2) in the atmosphere is converted into nitrogen compounds useful for other chemical processes (such as, notably, ammonia, nitrate and nitrog en dioxide). Nitrogen fixation is performed naturally by a number of di fferent prokaryotes, including bacteria, actinobacteria, and certain types of anaerobic bacteria. Micr oorganisms that fix nitrogen are called diazotrophs (Maier et al., 2000). Nitrogen-fixers add organic matter with a re latively low carbon to nitrogen ratios to the site; the organic matter accumulates and increases the waterand cation-holding capacity of the soil, and the level of availabl e nitrogen in the site rises. The rates of
71 nitrogen fixation are eventually reduced as th e availability of nitrogen increases in the soil, or when phosphate or sulfur become limiting to the nitrogen-fixers. Once the availability of nitrogen in the soil is high, other pl ants without nitrogen fixing capabilities can grow (Bolin et al., 1981). Figure 4.3: The Nitrogen Cycle (Maser, 2006) Like nitrogen and carbon, microbes can tran sform sulfur from its most oxidized form (sulfate or SO4) to its most reduced state (sulfide or H2S). In the sulfur cycle (Figure 4.4), anoxygenic photosynthetic purple and green sulfur bacteria oxidize H2S as a source of electrons for cyclic photophosphorylation. Since SO4 and sulfur may be used as electron acceptors for respiration, sulfate reducing b acteria produce H2S during a process of anaerobic respiration analogous to denitrification. The use of SO4 as an electron acceptor is an obligatory process that take s place only in anaerobic environments. The process results in the distinctive odor of H2S in anaerobic bogs, soils and sediments where it occurs. Sulfur is assim ilated by bacteria and plants as SO4 for use and reduction to sulfide. Animals and bact eria can remove the sulfide gr oup from proteins as a source of S during decomposition. These processes co mplete the sulfur cy cle (Maier et al., 2000).
72 Plants, algae and photosynthetic bact eria can absorb the phosphate (PO4) dissolved in water, and incorporate the minera l into various organic forms, including such molecules as nucleic acids and the phospholipids of cell membra nes. When the plants are consumed by animals, the organic phosphate in the plant becomes the organic phosphate in the animal and in the bacteria that liv e with the animal. Animal waste returns inorganic PO4 to the environment and also organic phosphate in the form of microbial cells (Figure 4.4). Dead plants and animals, as well as animal waste, are decomposed by microbes in the soil. The phosphate eventually is mine ralized to the soluble PO4 form in water and soil, to be taken up again by photos ynthetic organisms (Maier et al., 2000). Figure 4.4: Sulfur and Phosphor us Cycling (Maser, 2006) The soil-plant-microorganism cycle is strengthened by rapidly growing pioneer plants that invade disturbe d sites and take advantage of the available light, water, nitrogen, and phosphate to grow. The site then returns towards its predisturbance state, drawing carbon and some nitrogen and sulf ur from the atmosphere, and phosphate, nitrogen and sulfur from the soil (Bolin et al., 1981). Autotrophic and heterotrophic microorgani sms dominate global biogeochemistry, accounting for roughly half of global photosynthesis and mo st of the organic matter
73 decomposition, nitrification, deni trification, and methane produc tion (Maier et al., 2000). Several aspects of the biogeochemical cycles can be monitored for soil and microbial development. Though microorganisms are abundant in so ils, identifying microorganisms from soil samples is difficult because most bacteria in natural environments cannot be cultured with current techniques (for reviews, s ee Staley & Konopka, 1985; Roszak & Colwell, 1987a, 1987b; Amann, Ludwig, & Schleifer, 1995). Estimates reveal that only approximately 1 Â– 10 percent of the microorga nisms in soil have been identified (Amann et al., 1995). Table 4.4 lists some of the identif ied cultural bacteria a nd their functions in soil. Table 4.4: Dominant Cultural Soil Bacteria (Maier et al., 2000). Organism Characteristics Function Arthrobacter Heterotrophic, aerobic, gram variable. Up to 40% of culturable soil bacteria. Nutrient cycling and biodegradation. Streptomyces Gram positive, heterotrophic, aerobic actinomycete. 5-20%of culturable soil bacteria. Nutrient cycling and biodegradation. Antibiotic production, e.g., Streptomyces scabies. Pseudomonas Gram negative heterotroph. Aerobic or facultatively anaerobic. Posses wide array of enzyme systems. 10-20% of culturable soil bacteria. Nutrient cycling and biodegradation, including recalcitrant organics. Biocontrol agent. Bacillus Gram positive aerobic heterotroph. Produce endospores. 2-10%of culturable soil bacteria. Nutrient cycling and biodegradation. Biocontrol agent, e.g., Bacillus thuringiensis The biodiversity of soil microbial communities can be measured with mathematical indices that emphasize richness (t he number of species), equitability (the evenness of allocation of i ndividuals among the various speci es), or combinations of these two. Microbial biodiversity includes the number and dist ribution of species, as well as the functional diversity and redundancy. Func tional diversity is the number of distinct processes or functions that are carried out by a community, whereas functional
74 redundancy is a measure of the number of diffe rent species within the various functional groups or guilds (Gaston, 1996; Bei et al., 2000) Bacteria (and fungi) can be evaluated by viable count methods involving plating on nut rient-rich agar. Dire ct counting involves counting individual organisms with the naked eye or with a microscope. Plate counting calculates the number of bacteria l or fungal colonies that grow from a soil sample (Bei et al., 2000). Phenotypic methods of microbi al identification include microbial morphology, Gram staining, enzyme activities, and the utilization of se veral substrates as sole carbon and energy sources. Diversity can also be determined by meas urements of activity levels, such as the amount of by-products (e.g., CO2) generated in the soil, or the disappearance of substances, such as plant residue or methan e used by portions of the community or by specific groups of organisms. These meas urements reflect the total Â“workÂ” the community can do. Total biologica l activity is the sum of act ivities of all organisms, though only a portion is active at a particular time (Maier et al., 2000). Additional activity level measurements include: Determining the respiration rates by measuring CO2 production. This method is limited, as it does not distinguish which organisms (plants, pathogens, or other soil organisms) are generating the CO2. Determining the nitrification rates by m easuring the activity of those species involved in the conversion of ammonium to nitrate. Determining the decomposition rates by m easuring the speed of disappearance of organic residue or stan dardized cotton strips. In addition, evaluating the cellular cons tituents of the microorganisms can also suggest diversity. The total biomass of soil or ganisms or specific ch aracteristics of the community can be inferred by measuring: The amount of nutrients (carbon, nitrogen, or phosphorus) in the cells, which can then be used to estimate the total biomass of organisms. Chloroform fumigation is a common method used to estimate the amoun t of carbon or nitr ogen in all soil organisms;
75 The enzymes in the cells or those attached to the soil. Assays can be used to estimate potential activity or to ch aracterize the biological community; The Â“fingerprintÂ” of the community, us ing phospholipids and other lipids to quantify the biomass of groups such as fungi or actinomycetes; The various attributes of the individual nucleic acids (DNA and RNA) (Stephen et al., 1999a, 1999b; Maier et al ., 2000, Wright & Reddy, 2001). All of these techniques (mathematical indices of richness and equitability, nonmolecular and molecular meas urements of activity levels, and cellular constituents) can be used to as monitoring criteria for mitigation projects, and investigation into these procedures will be examined. Techniques for Measuring Biodiver sity and Community Structure Indices of Richness and Non-mol ecular Analytical Techniques The diversity of a community is expresse d as the species richness and the relative contribution each species makes to the total num ber of organisms present. The diversity of a microbial community is often desc ribed by the Shannon-Weaver index (HÂ’) (Shannon & Waever, 1949). The number of species has traditionally been determined by taxonomic classification studies, but as th ese are sub-optimal for microorganisms, molecular and biochemical techniques of estimating abundance and number of each species must be applied. The benefit of a hi gh genetic diversity is currently under debate because it is not always corre lated to functional diversity (Nielsen & Winding, 2002). Ibekwe, Lyon, Leddy and Jacobson-Meyers (2006) studied mi crobial community composition and water quality changes with in free water, surface constructed wetland cells that contained various plant densities and compositi on. The authors used the Shannon-Weaver index (HÂ’) to determine microbi al diversity and resu lts indicated that diversity was higher in the wetland cells with 50 percent plant density, rather than those with 100 percent plant divers ity. This study provided evid ence that wetlands with 50 percent plant cover can promote the growth of diverse microbial communities that facilitate decomposition of chemical pollutants in surface water, and improve water
76 quality. Thus showing that land managers s hould focus more on soil development, rather than planting techniques. Possible horizontal variations of commun ities in a natural oligotrophic fen, were analyzed by Galand, Fritze, an d Yrjl (2003) using metha nogens from two well-defined microsites: Eriophorum lawn and Hummock. The community structures were studied at two different layers of the fen, showing, respectively, high and low methane production. Phylogenetic analyses revealed six different clusters of seque nces grouping w ith only two known orders of methanogens. Upper layers of Hummock were dom inated by sequences clustered with members of Methanomicrobiales and the sequences that dominated the upper part of the Eriophorum lawn were members of the order Methanosarcinales Novel methanogenic sequences were f ound at both sites at both de pths. The authors concluded that the vegetation characterizing the microsites influenced the microbial communities in the layers of the fen where methane was produced. An additional non-molecular technique fo r measuring microbial diversity is a multi-stage dispersion and differential cent rifugation technique suggested by Hopkins MacNaughton and OÂ’Donnell (1991), used to sample non-filamentous microorganisms from soil. The soil aggregates were disp ersed and the microorganisms were dissociated from the soil particles. The released micr oorganisms were then separated by low-speed centrifugation. The biomass of the microorganisms were de termined using traditional methods such as direct microscopic cell counts, adenosine tr iphosphate (ATP), phospholipid, lipopolysaccharide, ergosterol conten ts and viable counts. When compared to traditional methods (phenotypic methods of microbial identification such as morphology, Gram staining, and enzyme ac tivities), the cent rifugation method yielded extracts that were more enriched wi th microorganisms (Hopkins et al., 1991). One approach for indexing the functional diversity of microbial communities has been the use of substrate or metabolic fi ngerprinting (Bei et al., 2000). Microbial communities are typically screened for their ability to utilize selected carbon substrates by using MicroPlates. Bei et al. (2000) developed a cultur e-independent strategy to examine bacterial functional redundancy and tested its use on soils collected along a vegetation gradient on reclaimed mine spoils in Rondonia, Brazil. The data suggested that
77 bacterial functional redundanc y increased in relation to the re-growth of plant communities, and therefore represented an important aspect of the restoration of soil biological functionality to reclaimed mine spo ils. The authors reiterated the importance of bacterial functional redundancy as an indi cator of soil quality and ecosystems functioning. Molecular Techniques Molecular microbial identification can be conducted in the laboratory ( in vitro) or in the field ( in situ). In vitro measurements may involve incubation of a soil sample in the laboratory under standardized conditions. Interpretation of in vitro measurements in relation to soil health can be difficult, however, because the re sults depend on the incubation conditions which may not be comparable to field conditions. Examples of in vitro measurements are soil respiration, nitrogen-miner alization, nitrification, denitrification, most probable number dilutions and other growth-based methods (Brock, Smith & Madigan, 1984). In situ measurements are based either on direct measurements in the field or fixed samples analyzed in the laboratory. These t ypes of measurements are sensitive to spatial and temporal variation, however, and may unde restimate the variability in soil health status (Brock et al., 1984). Examples of in situ measurements are phos pholipid fatty acid (PLFA), organic matter decomposition, thymidin e and leucine incorporation, short-term enzyme assays and most molecular methods (Brock et al., 1984). Phospholipid fatty acid (PLFA) analysis has been used by several authors for microbial analysis. PLFA provides information about soil microorganism biomass, fungal-bacterial ratios, biodive rsity and the occurrence of ke y species (Stephen et al., 1999a, 1999b). Fatty acids are extracted from microorganisms, characterized, and the ratios generated are compared to known ratio s of fatty acids present in cell membranes and other microbial structures (Stephen et al, 1999a, 1999b). The introduction of polymerase chain reaction amplification-dena turing gradient gel electrophoresis (PCRDGGE) to microbial ecology has provided a s econd valuable molecular fingerprinting technique for studying microbi al community structure (Henckel, Friedrich & Conrad,
78 1999; Heuer, 1997; Muyzer 1998). Denaturing gradient gel electrophoresis separates and distinguishes deoxyribonucleic acid (DNA) fragments by establishing a denaturing gradient that separates DNA into discrete ba nds than can then be excised and sequenced for identification. DGGE allows large numbers of samples to be analyzed simultaneously, thus this technique is id eally suited for monitoring the dynamics of microbial communities influenced by environmental ch anges (Kelley & Hentzen, 2003; Henckel et al., 1999). A study conducted at a wetland site in S outhern Illinois curre ntly undergoing restoration prior to leveeing and agricultura l use used PLFA and PCR-DGGE to identify the dominant bacterial populations of the Spunky Bottoms aquatic wetland system (Kelley & Hentzen, 2003). Water samples were collected and analyzed by gas chromatography/mass spectrometry to determine PLFA profiles for each water sample. The PCR-DGGE DNA fragments were excised, sequenced, and predominant microbial species were characterized based upon sequence homology to previously identified sequences contained in the National Cent er for Biotechnology and Ribosomal Data Project databases (Kelley & Hentzen, 2003). Results identified diverse microbial communities, including microorganisms th at may substantially contribute to biogeochemical cycling of elements, incl uding nitrogen and phosphorus. The sequence analysis of DNA fragments showed si milarity indices of 0.8 to 1.0 to a -Proteobacteria, Flavobacterium, Flexibacter-Cytophaga-Bact eroides groups, Gram-positive and Gramnegative bacteria, Cyanobact eria, and Chloroplasts (K elley & Hentzen, 2003). The authors emphasized that microbial communiti es (and populations) could be used to provide a basis of comparison for future research. Bossio, Fleck, Scow and Fujii (2006) used PLFA analysis to access the microbial community changes as a site previously under agricultural ma nagement transforms to a permanently flooded wetland. The PLFA anal ysis indicated that the active microbial community of the wetland was different fr om that of the agricultural field. Databases of PLFA profiles have been ge nerated for a variety of microorganisms to facilitate their identifi cation. Phospholipid fatty acids analysis has been used to examine microbial communities and their changes for a variety of marine, aquatic, and
79 terrestrial systems (Stephen et al., 1999a 1999b). Additionally, PCR-DGGE analysis has been used to isolate microbial DNA and thereby identify predominant microbial populations in a variety of sy stems (Kelley & Hentzen, 2003). Synthesis of DNA is a prereq uisite for bacterial cell division and, as such, an indicator of bacterial growth. The incorporation of H3or C14-thymidine into bacterial DNA can be used to determine DNA synthesis, as thymidine is a nucleoside that is used in DNA synthesis. This method has several requirements: (1) DNA synthesis has to be linearly correlated with the cell growth (balanced growth ); (2) all bacteria must take up thymidine through the cell membrane; (3) thym idine should not be catabolized; and (4) the radioactive label (H3) should not exchange with other molecules (e.g. proteins). It has been shown that only 5 Â– 20 percent of the H3-thymidine incorporated into total macromolecules is incorpor ated into DNA (Baath, 1998, Mich el & Bloem, 1993). The procedure involves incubating th e soil samples with radio-labeled thymidine for a short time followed by filtration to measure the am ount of radiolabel in the cells. Additional approaches to de scribe the diversity of natural soil or sediment communities have included DNA re-association experiments (Torsvik, Goksoyr & Daae, 1990), 16S ribosomal DNA (rDNA) retrieval a nd analysis (Borneman et al., 1996), fractionation of total bacterial DNA by G 1 C content (Holben & Harris, 1995), and crosshybridization of bacterial DNA from two co mmunities (Ritz & Griff iths, 1994). All of these techniques have pointed to highly comp lex assemblages of bacterial populations. For example, Torsvik et al (1990) estimated that 103 to 104 different genomic equivalents were present in 1 gram of soil, while Borneman et al. (1996) recovered 124 previously undescribed 16S rRNA gene sequences from an agricultural soil. These studies of genetic diversity have provided valuable and quali tative insights into microbial community composition. Recombinant DNA and molecular phylogene tic methods have recently provided means for identifying the types of organisms th at occur in microbial communities without the need for cultivation (see Hugenholtz & Pace 1996 for re view). Hugenholtz and Pace compiled a table that summarizes the environm ental distribution of sequences by habitat type, derived from most of the available 16S rRNA-based clonal analyses (86 studies
80 contributing nearly 3, 000 sequences). An ex panded version of this table, which details division-level representation in the individual studies, is available at http://crab2.berkele y.edu/pacelab/176.htm. The amount of ribonucleic acid (RNA) in individual cells can represent protein synthesis and, thus, microbial activity. The number of active cells can be detected by fluorescent in situ hybridization (FISH) (Amann et al., 1995). By this method, individual cells carrying high concentra tions of rRNA are quantifie d by fluorescence microscopy and/or reverse transcriptase polymerase chai n reaction (RT-PCR), where rRNA extracted from soil is detected by creating a DNA copy and separating by gel electrophoresis (Duineveld et al., 2001). The polymerase chain reaction (PCR) technique geometrically increases concentrations of specific DNA fragments to provide enough genetic material for identification by techniques such as gel elec trophoresis. PCR allows a small amount of the DNA molecule to be amplified many times, in an exponential manne r, and is used to amplify a short, well-defined parts of the DNA strand. PCR re quires several basic components: The DNA template; Two primers, which determine the beginning and end of the region to be amplified; Taq polymerase, which copies th e region to be amplified; Deoxynucleotides-triphosphate from wh ich the DNA-polymerase builds the new DNA; and Buffers, which provide a suitable chemical environment for the DNApolymerase The PCR technique was used by Henckel et al. (1999) to investigate the structure of the methanotrophic community in rice field soil. The soil microbial community was monitored by performing DGGE during the oxid ation process with different PCR primer sets based on the 16S rRNA gene and on f unctional genes. A universal small-subunit (SSU) rDNA primer set and 16S rDNA prim er sets specifically targeting type I methylotrophs (members of the subdivision of the class Proteobacteria [ -
81 Proteobacteria ]) and type II methylotrophs (members of the Proteobacteria ) were used. Functional PCR primers targeted the genes for particulate methane monooxygenase ( pmoA ) and methanol dehydrogenase ( mxaF ), which code for key enzymes in the catabolism of all methanotrophs. The yield of PCR products amplified from DNA in soil that oxidized CH4 was the same as the yield of PCR products amplified from control soil when the universal SSU rDNA primer set was used, but was significa ntly greater when primer sets specific for methanotrophs were used. The DGGE patterns and the sequences of major DGGE bands obtained with the universal SSU rDNA primer set showed that the community structure was dominated by non-methanotrophic populations related to the genera Flavobacterium and Bacillus and was not influenced by CH4. The structure of the methylot roph community, as determined with the specific primer sets, was less complex; this community consisted of both types I and II methanotrophs re lated to the genera Methylobacter Methylococcus and Methylocystis DGGE profiles of PCR products amplif ied with functional gene primer sets that targeted the mxaF and pmoA genes revealed that there were pronounced community shifts when CH4 oxidation began. High CH4 concentrations stimulated both types I and II methanotrophs in rice field soil with nonsat urated water content, as determined with both ribosomal and functi onal gene markers (Henckel et al., 1999). Zhou et al. (2002) identified possible de terminants for microbial community structure in soil by characterizing microbia l communities from 29 soil samples using a SSU rRNA-based cloning approach. The au thors concluded that while microbial communities in lowcarbon, saturated, subs urface soils showed dominance, microbial communities in low-carbon surface soils sh owed remarkably uniform distribution. Analysis using the SSU technique has been us ed for agricultural sites (McGarvey, Miller, Sanchez & Stanker, 2004), rice field soils (Fey & Conrad, 2 000), and nitrifying bacteria diversity in wastewater (Princic et al., 1998). Overall, the use of nucle ic acid-based methods for soil microbial community iden tification has revealed high prokaryote diversity (Bintrim et al., 1997; Borneman & Triplett, 1997; Felske, Wolterink, Lis & Akkermans, 1998; Dunbar et al., 1999). In Eur ope, a database (http://rrna.uia.ac.be/ssu) has compiled all complete or nearly co mplete SSU RNA sequences. Additional
82 information (i.e., literature reference, taxonomy, secondary structure models and nucleotide variability maps) is also available at this website. Microbial diversity can al so be measured using a process known as BIOLOG. This technique characterizes and identifies mi crobial isolates by ex amining their carbon source utilization profiles (N ielsen & Winding, 2002). BILOG test s the ability or inability of microorganisms to metabolize (i.e. degrad e via oxidation) a larg e and diverse set of chemicals substrates. Identification of the mi croorganism is generated by a computerized matching of the metabolic proper ties of the isolate to a large database of patterns (Nielsen & Winding, 2002). Although molecular techniques are powerful, they show some limitations in the case of diversity studies. One shortcomi ng of PCR is that it involves sampling a heterogeneous matrix of soils, a point which fr equently receives little consideration in study design (Amann et al., 1995). Grundma n and Gourbiere (1999) designed a soil sampling procedure that considered bacterial spatial distribution within the soil matrix. The results of their st udy suggested that the micro-fragme ntation of soil (fr actions of soil, sampling of specific habitats, dissection of mi nute pieces of soil) prior to culturing can help minimize cell interactions during sample processing for isolation. Tilman (1994) supported this finding and added that so il fragmentation decreases interspecies competition, allowing growth of minor popul ations in enrichment cultures. Molecular tests do not have to be time consuming or costly. When samples are collected for multiple tests, the soil samples can be divided into three parts one portion of the sample is used for determining moistu re content, one for organic content, one for microbial community analysis based on extr actable lipid profiles or microbial DNA (Ogram, 1987). The moisture content of soil can be determined by weighing out a wet soil sample then drying it overnight in an oven at 105 C. Lipids can be extracted from each sample using a Bligh-Dyer extraction buffer (Ogram, 1987). In addition to using microbial lipids as a biochemical marker for community structure, microbial genomic DNA can be extracted from soil samples then subjected to PCR. By designing different PCR primers, the whole community (or a subset of the community) can be analyzed.
83 Measurements of Microbial Activity Â– Analysis of the Bi ogeochemical Cycles Changes (natural or anthropogenic) in soil chemistry and stru cture can disrupt the requisite biogeochemical cycles and prolong wetland development. For example, the impairment of soil carbon cycling can trap the element in wetland soils. Because nitrifying bacteria (e.g., Nitrosomonas and Nitrobacter) are sensitive to acidity and require aerobic conditions, waterlogged so ils can become anoxic and may not support nitrification. Nodulation of legumes can be impeded by some kinds of pollutant molecules and thus the symbiotic nitrogen fixation phenomenon can be disrupted. Phosphorus cycling could conceivably be im peded by anything that interferes with mycorrhizal fungi (Sims, 1990). Instead of testing for the pollutant, the microbial facets of the soil can be evaluated. Nitrogen-fixing and ammonia-oxidizing bacteria were studied by Yeager et al. (2005) to examine the effects of forest fire on these importa nt groups of nitrogen-cycling bacteria. Analysis of nifH and amoA PCR amplicons was performed on DNA samples from unburned, moderately burned, and severely burned soils of a mixed conifer forest following the Cerro Grande Fire (an inte nsive crown fire) near Los Alamos, New Mexico. The results indicated that a decrea sed microbial biomass and shift in nitrogenfixing and ammonia-oxidizing communities was still evident in fire-impacted soils collected 14 months after the fire. The si te was not showing signs of deficiently, but these authors showed that the bacterial com ponents of the system had been affected and remained compromised months later. The authors also suggested long-term monitoring for successful rehabilitation. Key biogeochemical processes such as organic matter decomposition, pollutant degradation, and non-symbiotic nitrogen fixa tion occur at acceler ated rates in the rhizosphere and greatly influence ecosys tem functions (Anderson, Guthrie & Walton, 1993; Daane et al., 2001). Re cent work has indicated that the stimulation of microbial activity in the rhizosphere of plants can al so stimulate biodegrada tion of various toxic organic compounds (Anderson et al., 1993). The rhizosphere soil is the zone of soil under the direct influence of plant roots, which usually extends a few millimeters from the root surface and is a dynamic environment for microorganisms (Daane et al., 2001).
84 The rhizosphere microbial community is co mprised of microorganisms with different types of metabolic and adaptive responses for various environmental conditions. The production of mucilaginous material and the exudation of a variety of soluble organic compounds by the plant root play an important part in root colonization and maintenance of microbial growth in the rhizosphere (Daane et al., 2001). Bacteria of the genera Rhizobium ar e abundant in soil and form symbiotic associations with legume roots. The b acteria reside in nodules where they fix atmospheric nitrogen and provide the plant with elemental nitrogen for growth. In return, the plant provides the bacteria with organi c substrates for growth (Hungria, Chueire, Coca & Megias, 2001). Number s of Rhizobium has previous ly been proposed as an indicator of soil health based on the organismsÂ’ sensitivity to pesticides and heavy metals (Hungria et al.. 2001). The frequency and diversity of Rhizobium in soil can be determined by a simple pot test, where a divers e set of legume seeds are sowed in the test soil and the number of nodules formed are dete rmined after a specif ic growth period (Nielsen & Winding, 2002). Alternatively, the bacter ia may be quantified by direct isolation from soil using selective growth media together with morphological and physiological characterizations (Hungria et al., 2001). A study was conducted to isolate the pol ycyclic aromatic hydrocarbon (PAH)degrading bacteria from the rhizos phere of the salt marsh grasses S. alterniflora, J. geradi, D. spicata and S. airoides by enrichment using napht halene, phenanthrene, or biphenyl as the sole source of carbon and ener gy (Daane et al., 2001). The authors found both gram positive (predominantly nocardioform) and gram-negative (predominantly pseudomonad) bacteria, and a pasteurization te chnique prior to enri chment, resulted in the isolation of spore-forming bacteria (exclusively Paenibacillus sp.). These results demonstrated the wide variation between the PAH-degrading isolates indicating that the rhizosphere of S. alterniflora contained a diverse population of PAH-degrading bacteria (Daane et al., 2001). Polymerase chain reaction (PCR) pr imers that amplify parts of the narG and nirK genes can be used to create a community profile that focuses on the denitrification process (Maier et al., 2000) When cloned and sequence d, these PCR products can be
85 used to identify the bacteria involved. Afte rwards, the bacterial growth rate (number of cells formed per unit time) can be calculated. Soil respiration can be determined by either carbon dioxide (CO2) production or oxygen (O2) consumption. Measurements of CO2 concentration are more sensitive, because the atmospheric concentration of CO2 is only 0.033 percent versus the 20 percent of O2. Determination of CO2 production from soil samples can be made in the laboratory by simple and inexpensive techniques based on alkaline CO2 traps followed by chemical titration, or by more sophisticated auto mated instruments based on electrical conductivity, gas chromatography or infrared spectroscopy (Alef, 1995). Combined with automated sampling from test soil samples, automated instruments make it possible to determine CO2 production as a function of time for several days. (Zibilske, 1994). Respiration is highly influenced by temperatur e, soil moisture, nutrient availability, and soil structure. Pre-conditioning and standardization of the soil before measuring respiration is necessary to minimize the e ffect of these variables. Soil respiration measurements have been used as an indicat or of pesticide and heavy metal toxicity (Brookes, 1995). Wright and Reddy (2001) studied the influence of phosphorous loading on aerobic and anaerobic heterotr ophic microbial activities. The authors measured CO2 and methane (CH4) production in detritus and soil collected from a Water Conservation Area in the Everglades. This field st udy involved the measurement of CO2 production under drained and flooded soil conditions. Schott medi a bottles with screw-type lids containing 10 grams of drained, moist soil and a sodium hydroxide (NaOH) trap were sealed under an atmosphere of 21 percent O2 to facilitate aerobic condi tions. A flooded (anaerobic) treatment containing 10 grams of wet samples of flooded soil with an atmospheric nitrogen headspace was also included. Fo r substrate induced respiration (SIR) measurements, both drained and flooded treat ments were supplemented with glucose at an excess concentration of 25 mg C g soil 1, based on results of previous experiments (Wright & Reddy, 2001). Results revealed that both CO2 and CH4 production significantly correlated with soil P pa rameters and microbial biomass.
86 In a study by Rooney-Varga, Richar d, Robert & Hines (1997), phylogenetic diversity and community compos ition of sulfateÂ–reducing bacter ia in salt marsh sediment and in the rhizosphere of Spartina alterniflora were investigated. The authors chose a community of sulfate-reducing bacteria (SRB) inhabiting a salt marsh sediment. Sulfatereducing activity correlates with plant gr owth stages, suggesting that plant-SRB interactions in the S. alterniflora rhizosphere play an impor tant role in salt marsh biogeochemical cycles (Hines, 1991; Hine s, Knollmeyer & Tugel, 1989). While molecular studies of soil-sediment microbial communities have suggested extremely high complexity, with up to 104 species present in a gram of soil (Torsvik et al., 1990), the community structure or quantit ative distribution of indivi dual phylotypes remains poorly understood. In this study, the results sugg ested that while the overall sediment community may be highly diverse, there were a small number of well-adapted species in the sediment habitat that play a signifi cant role in microbial community dynamics (Rooney-Varga et al., 1997) Wetlands are considered important site s for methane oxida tion, as these areas receive a high input of organic material Methane is produced by methanogenic Archaea and consumed by aerobic methane oxidizing bact eria, the methanotrophs (Maier et al., 2000). Methane oxidation is measured by sp iking a soil sample with methane and incubating the sample in a closed jar in the laboratory. Loss of methane is subsequently determined by gas chromatography. Methanotr ophs can be quantified directly in soil by fluorescent in situ hybridization (FISH) (Amann et al. 1995) or standard growthdependent most-probable-number (MPN) co unts. Analyses of methanotrophic communities can also be done with PCR-DG GE using methanotrophs-specific 16S rDNA primers (see Ritchie, Edwards, McDonald & Murrell, 1997). Methanogens possess several potential biomarkers, including coenzyme F420, isoprenoids, and lipid ethers in the cell membra nes, as well as coenzyme M (CoM) (Elias, Krumholz, Tanner & Suflita, 1999). The previously developed techniques for assaying CoM include a bioassay and high-perform ance liquid chromatography (HPLC)-based method which measures fluorescent isoindole deri vatives of thiols (Elias et al., 1999). The latter method was not standardized fo r biomass determination. Both of these
87 methods are cumbersome and time-consuming, require strictly anaerobic conditions and were not designed for use with sediments. E lias et al. (1999) modified the HPLC-based procedure to quantify CoM with in hours of sample collectio n without a requirement for anoxic conditions. The authors also standardi zed the technique with pure cultures so that it could be used to quantify methanogen bioma ss in a variety of environmental matrices. The MPN assays used three-tube assays to a 10-8 dilution. The method resulted in an efficient and relatively simple method to estimate methanogenic biomass. In addition to analyzing aspects of the biogeochemical cycles, Tate and Klein (1985) suggested investigating soil microbial bi omass. The authors stress that microbial biomass is fundamental to the transformati on and flow of carbon, nitrogen, sulfur and phosphorus. The microbial biomass is the frac tion of the soil respons ible for the energy and nutrient cycling and the regulation of organic matter transformation (Nielsen & Winding, 2002). Microbes transform organic n itrogenous and sulfuric compounds to forms that can be utilized more readily to plants. Wetland floras are dependant on the microbial oxidation of these compounds by micr oorganisms for survival. In addition, soil microorganisms also immobilize nutrients that would be otherwise lost to leaching (Tate & Klein, 1985). Determination of soil microbial bioma ss by direct methods (microscopy or PLFA analysis) provides results that very closely represent the in situ soil conditions. Although the methods are time-consuming, they are currently used for soil monitoring purposes (see Bloem, Bolhuis, Veninga & Wieringa 1995). Indirect methods are generally cheaper, faster and easier to use than the dire ct methods. Results obtained by the indirect methods have been documented to be very clos e to the direct measur ements (Carter et al., 1999), thus providing confidence in th e utility of indirect methods. Analytical techniques for microbial biom ass determination include the chloroform fumigation incubation method (CFI) and the chloroform fumigation extraction method (CFE) (Carter et al., 1999). In both cases, th e chloroform vapor kills the microorganisms in the soil, and subsequently the size of the killed biomass is estimated either by quantification of respired CO2 over a specified period of in cubation (CFI) or by a direct extraction of the soil immediately after the fumigation followed by a quantification of
88 extractable carbon (CFE). The release of CO2 after fumigation is the result of germinating microbial spores utilizing th e carbon substrate prov ided by the killed microbial cells (Carter et al., 1999). Another common indirect method is substr ate induced respiration (SIR). This method measures only the meta bolically active portion of th e microbial biomass (Carter et al., 1999). Substrate induced respiration measures the initial change in the soil respiration rate as a result of adding an easily decomposable substrate (e.g. glucose). Soil microbial biomass is subsequently calculated using a conversion factor (Carter et al., 1999). Measurements of Microbial Activity Â– Soil Organic Matter Many soil properties impact soil quali ty, but organic matte r deserves special attention. It affects several critical soil functions, can be manipulated by land management practices, and is important in mo st agricultural setti ngs across the country (McCauley, Jones & Jaconsen, 2003). Becau se organic matter enhances water and nutrient holding capacity and improves soil structure, managing for soil carbon can enhance productivity and envir onmental quality, and can reduce the severity and costs of natural phenomena, such as drought, flood, and disease. In addition, increasing soil organic matter levels can reduce atmospheric CO2 levels that contribute to climate change (McCauley et al., 2003). Because soil microorganisms require ex ternal sources of organic carbon, the activity and sustainability of the microbial community depends on whether this organic matter is replaced after mining. The rate at which stable organic matter accumulates on disturbed systems will be largely determined by the biological characteristics of the mine site (McCauley, Jones & Jaconsen, 2003). The soil organic matter (SOM) content is equal to the net difference between the am ount of SOM accumulated and the amount decomposed. Soil organic matter cycling consis ts of four main processes carried out by soil microorganisms (Figure 4.5): 1. Decomposition of organic residues; 2. Nutrient mineralization;
89 3. Transfer of organic carbon and nutrients from one SOM pool to another; and 4. Continual release of CO2 through microbial respira tion and chemical oxidation (McCauley, Jones & Jaconsen, 2003). Figure 4.5: Organic Matter Cycl e (McCauley et al., 2003). This figure is modified from Brandy and Weil, 1999. The three main pools of SOM, determined by their time for complete decomposition, are active (1-2 years), slow (15-100 years), and passive (500-5000 years) (Brandy and Weil, 1999). Soils high in clay and silt (as found at mi ning sites) are generally higher in SOM content than sandy soils. This is attributed to aeration in finer-tex tures soils reducing the rate of organic matter oxidation, and the bi nding of humus (a dark brown, porous, spongy material that provides a car bon and energy source for soil mi crobes and plants) to clay particles. This process further protects th e soil from decompositi on (McCauley et al., 2003). The SOM content is a key indicator of so il quality and is correlated to a number of important soil processes th at occur in wetlands such as respiration, denitrification, and phosphorus sorption. To better understand the differences in the SOM content of created, restored, and natural wetlands, 11 created/ restored and natural wetland pairs were
90 sampled in North Carolina (Bruland & Richardson, 2006). Brul and and Richardson hypothesized that the SOM content of paired created/restored and na tural wetland would be similar. Results indicated that all of the individual sites had significantly different SOM contents, with created/restored wetla nds having significan tly lower mean SOM than their paired natural wetland on average. Th e authors thus concluded that if there is a choice in mitigation options (restoration or cr eation), wetlands should be restored rather than created (Bruland & Richardson, 2006). Noyd, Pfleger, Norland and Sadowsky (1995) studied the effect of mitigation treatments on actinomycetes, fungi, free-liv ing nitrogen-fixing ba cteria, and aerobic heterotrophic bacteria when compared in fiel d plots in coarse taconite tailing. The authors concluded that incorpor ation of a moderate rate of organic matter can ameliorate the stressful conditions of co arse taconite tiling, and ca n enhance the initiation of a functional soil ecosystem able to support th e establishment of seeded native prairie grasses. Noyd et al. also suggested that such amendments could provide a long-term solution to the reclamation of taconite tailing. When soils are disturbed, organic matter previously protected from microbial action is decomposed rapidly because of changes in water, air, and temperature conditions, and the breakdown of soil aggregat es accelerates erosion; a soil with high organic matter is more productive than soil where much of the organic matter has been reduced through tillage, excavation and poor ma nagement practices and transported by surface runoff and erosion (McCauley et al., 20 03). Nair et al. (2001) consider organic matter accumulation to be one of the most impo rtant indicators of functional wetlands. In a study conducted in 2001, Nair and colleagues compared the soil characteristics of created wetlands and na tive wetlands in phosphate-mined areas in central and north Florida. The author s took 151 samples (with soil cores 20-30 centimeters deep) along the hydrologic gradients of the created site, adjacent created and natural wetlands of the same wetland type, wetlands with different plant growth, and wetlands of different construction ages. The method of sampling and sampling strategy may differ per mitigation site, but the authors emphasize that it is important to get a representative sample of the microbial commun ity in the mitigation area. Results of the
91 Nair et al. (2001) study s howed that organic matter acc umulation increased across transect going from uplands to ward the center of the wetla nd, and thus showed promise of the constructed wetlands retu rning to Â“typicalÂ” wetlands. Nair et al. also called attention to th e use of reference wetlands. Wetland soils reflect the surficial geology and ch aracteristics of parent material This fact is critical for soil maintenance and important when comparing mitigated wetlands to natural, reference wetlands. Several studies have used eco-regions (areas di fferentiated by regional soil differences) as the framework for evaluating re storation ad creation projects (Abbruzzese, Allen, Henderson & Kentula, 1988; Brooks & Hughes, 1988). Overall, information on soils of wetland ecosystems is essential in order to re-establish native vegetation after phosphate mining. The advantages of a reference wetland approach include: Allowing for explicit goals for compensatory mitigation through the identification of reference standards deri ved from data that typify sustainable conditions in a region; Providing templates from which rest ored and created wetlands can be designed; and Establishing a framework whereby a d ecline in functions (resulting from adverse impacts) or recovery of func tions (following restoration) can be estimated both for a single project over a larger area accumulated over time (Brinson & Rheinhardt, 1996). To establish reference standards, conditi ons inherent to highly functioning sites must be identified for classes of wetlands th at share similar geomorphic settings, are of about equal age and have similar sedimentar y regimes and vegetation densities (Brinson & Rheinhardt, 1996). This is because mi crobial communities respond strongly to changes in sediment organic matter, which us ually accumulates with wetland age. Pratt and Cairns (1985) found that recently distur bed ponds had fewer microbe species than did natural and older reclaimed ponds on a surface-mined site. However, the microbial communities in ponds more than two years old we re indistinguishable from those in older reclaimed, unreclaimed, and natural ponds despit e differences in wa ter quality. Other factors that could be impor tant to standardize among collections of microbial
92 communities include: light penetration (wat er depth, turbidity, shade), temperature, sediment oxygen, baseline chemistry of wate rs (particularly pH and conductivity), detention time, current velocity, vegetati on density, dominant vegetation species, and moisture (Adamus & Brandt, 2004). Conclusions The incorporation of soil microbiology can ensure a functional reclaimed wetland. Legislation thus far has identified three ch aracteristics that comprise a wetland Â– its hydrology, vegetation and soil. Th e soil component is being thoroughly investigated and is proving relevant and pertinent. Microo rganisms in wetland ecosystems play a much more vital role than the legislation acknowle dges. Microbial indi cators should not serve as a Â“quick fixÂ” for reclamation problems, but instead should be considered a major part of the planning and implementation of recl amation projects. The science may seem complicated, but it is essential to include all of the components of natural wetlands and mitigate wetlands that are functionally, biologi cally and physically equivalent to natural systems.
93 Chapter 5: Ecological Research and Environmental Policy Introduction The abundance of wetland literature should refl ect the diversity a nd totality of the mitigation policies, but it does not. In term s of the interface between science and policy, the discrepancies continue to increase and the guidelines re main outdated and selective. One explanation for this Â“gapÂ” is the value of certain ecological systems is difficult to determine, and thus remains subjective. Mu rtaugh (1996) suggests the use of ecological indicators to determine environmental quality, and the establishment of statistical models to evaluate the usefulness of the indicator. Some authors suggest that ecologists should think more like policy makers and some suggest the opposite (Pouyat, 1999). A point of contention lies at the interf ace between scientists and no nscientists, and what ensues when they attempt to communicate (Rykiel 2001). Weber and Word (2001) emphasizes that non-scientists a ssumes that scientists are advocating a posi tion, while scientists believe that they are only providing objectiv e information. These authors recommend that the education of biologists and ecologists should include a communications component aimed at understanding the multifaceted intera ctions between what scientists say and what nonscientists hear (Weber & Word, 2001). Although critiques of the wetland mitigation process are available to scientists and policy-makers (i.e. Race, 1985), the guidelines remain inadequate. Government agencies have the dual res ponsibility of safeguarding the nationÂ’s wetland and providing avenues for development. Debate continues as federal policy mandates permits that are to be given only if impacts to wetlands are Â“unavoid able,Â” but in reality mitigation is used to compensate for most wetland habitat destructi on. On a landscape level, the fragmentation of these wetlands also translates into frag mented habitats and th e disruption of natural systems (Bedford, 1996). According to National Research CouncilÂ’s reports, mitigation efforts cannot yet claim to have duplicated lost wetland functional values (NRC, 1992).
94 Based on preliminary literature reviews, it seems that most ecologists are calling for a reordering of priorities. They envisi on policy-makers realizing the need for improved compliance, generating thriving wetlands and mainta ining a true baseline. My final objective is to reinforce the wealth of mitigation literature by documenting the relevant literature and highlighting the out standing arguments. In doing so, I can provide a more distinct scientific and policy revi ew for researchers, land managers and legislators. Chapter 5 begins with an ou tline the current literature, continues by synthesizing the varying point of view, and concludes with some possible ways to bridge the science-policy gap. Ecological Research and Policy Congressional policymaking has had a signi ficant impact on the management of wetlands in the United States. Tzoumis (1998) examined 240 congressional hearings and 1,569 witnesses who testified from 1789 to 1995 on wetlands. His findings showed three distinct eras exist in wetland policymaking: Era I (1789-1945), which was charac terized by a dominant monopoly of agricultural and development issu es in congressional policymaking Era II (1946-1965), which was the beginning of public attention to environmental issue Era III (1966-1995), which marked a peri od of wetland policymaking that was characterized as Â“contentious,Â” and no one issue controlled legislation. Unlike Era I, todayÂ’s congressional policym aking reflects many different issues. The growth of wetland science and continued pu blic interest presents a challenge to the dominant agriculture and development in terests that once controlled congressional policymaking. Wetland policy debates now include issues such as leisure, environmental protection, private property, and economics. A series of Local, State, Federal and pr ivate programs are available for wetland protection, but implementation of these pr ograms has been unsuccessful, therefore a continued loss of natural wetlands remains (Whigham, 1999). Even though Section 404 of the Clean Water Act establishes the major Federal regulatory programs protecting
95 wetlands, no explicit goals for their mana gement have been set (Dennison & Berry, 1993). In terms of biodiversity, current wetla nd protection policies fail because restored and created wetlands are often distinct from natural wetlands. Wetland policies often do not address the preservation or restoration of critical ec ological processes, such as nutrient cycling, because they mostly focus on individual wetland and ig nore the fact that wetlands are integral parts of landscapes (Whigham, 1999). Wetland mitigation projects often result in the exchange of one type of wetland for another and thus result in a loss of wetland functions at the landscape level. The most striking weakness in the current national wetlan ds policy is the lack of protection of dryend wetlands. From an ecological perspect ive, dry-end wetlands (such as isolated seasonal wetlands and riparian wetland associat ed with first order streams) may be the most important landscape elements. They often support a high biodiversity and are impacted by human activities more than ot her types of wetlands (Mitsch & Gosslink, 1993). The failings of current wetland protectio n and mitigation policies also reflect the lack of ecologically sound wetland assessm ent methods for guiding decision making processes. Examples of current wetland functional methodologies used for assessment are the Hydrogemorphic Classification Model (HGM) (Brinson, 1993; Smith, 1995), and the Wetland Evaluation Technique (WET) (Ada mus, 1991). The HGM is based on three fundamental factors that influence how wetlands function: (1) the position in the landscape (geomorphic setting) (2) hydrology, and (3) the fl ow and fluctuation of the water once in the wetland or hydrodynamics (Brinson, 1993). This assessment technique is described further in Chapter 1, Wetlands Classifications and Settings. For the WET approach, eleven functions and values are addressed: Ground water recharge; Ground water discharge; Flood-flow alteration; Sediment stabilization; Sediment/toxicant retention; Nutrient removal/transformation;
96 Production export; Wildlife diversity/abundance; Aquatic diversity/abundance; Recreation; and Uniqueness/heritage (Adamus, 1991). The WET approach also provides a procedur e to evaluate habitat suitability for 14 waterfowl species groups, 4 freshwater fi sh species groups, 120 species of wetlanddependent birds, and 133 species of saltwater fish and invertebrates (Adamus, 1991). The method evaluates the probability that a func tion will occur as high, moderate, or low. In addition, Fonseca et al. (2000) devel oped the Habitat Equivalency Analysis Technique which computes the amount of habitat to be restored to compensate for on-site lost ecological services. The Wetlands Re serve Program (Wetlands Reserve Program [WRP], 1993) provides a bib liography of other wetland miti gation evaluation techniques (Appendix C). Political jurisdictions do not necessar ily correspond to ecol ogically significant boundaries. Attempts at engaging more than on e agency or organization to jointly set goals for a particular location have remained mixed. The breadth of wetland issues, the various agencies and private landowners, and the saturation of disjointed wetland research make it difficult to generalize a bout any one case or answer any specific questions. Thus the planning process is hi ndered and questions about the planning process remain unanswered (Dennison & Berry, 1993). One solution is adaptive management. Adaptive management involves continua lly improving management policies and practices by learning from the outcome s of operational programs. (Holling, 1978; Walters, 1986; Walters & Holing, 1990). Th is form of management requires communication and cooperation, and that al l groups understand the overall goal (Brown, 2005). In all, the stakeholders, including scientists, miners, regulators and local citizens, are involved in a program of co-evolution of science, management and policy that somewhat parallels the pract ice of modern resource mana gement known as adaptive management (Holling, 1978; Walter & Holling, 1990; Gunderson & Holling, 2002).
97 The fate of wetland preservation depends on scientists, the phosphate industry and governmental regulators all sharing commo n goals. Ongoing research is needed to provide information and valuable insight into the necessary corrections that can be made and new approaches to defining success, bu t the information becomes scattered and disjointed. Because it is difficult for government al officials to keep up with the latest development and technical details of environmen tal research, they rely on review articles, government summary reporters and information from professional scientists with whom they interact (Race, 1985). In addition, ma ny published reports are misleading and an inaccurate picture of the status of restorati on attempts is projected (Race, 1985). The time has come to start synthesizing all of the materi al and use it to genera te answers instead of sparking more discussions. Regulators and managers routinely use simp le indicators of success. Normally 3-5 years of onceor twice-per-year monitoring is required, with easily measured parameters such as plant lists, animal identification, and percentage of vegetation cover as the overall indicators. Assessing success is then based on comparing these parameters with a simple set of criteria that were stipulated in the orig inal permit for the project, but these criteria may or may not accurately reflect we tland function (Mitsch & Wilson, 1996). The problem is that the natural f actors are not used cohesively in mitigation projects. Managers are also unable to critically an alyze the success of restoration projects because of the great variability in nature and scope, and the inconsistencies in definitions and classification schemes in the literature. Not only are th ere several definitions for wetlands, but there are various definitions fo r how to monitor a reclaimed wetland. Of the various monitoring techniques, the two types that are particularly relevant to restoration are implementation and effectiveness monitori ng. Implementation monitoring is used to assess if a directed management action has been fulfilled. Implementation monitoring quantifies changes immediately after treatment s, and evaluates whether treatments were done as prescribed (Noss & Cooperrider, 1994). Effectiveness monitoring is used to determine whether the action achieved the ultim ate objective. This type of monitoring requires response variables to be clearly ar ticulated so that they can be measured accurately and precisely. Typical response va riables for wildlife are related to speciesÂ’
98 habitats or populations (Noss & Cooperrider, 1 994). If the goal of restoration is not clearly stated and monitoring procedures do not consider the ecosystem functions necessary for wetland restoration, monito ring becomes nothing more than literally Â“watching the grass grow.Â” In addition, c onsultants and landscape architects only work for 1 year, so effective and consistent monitoring becomes a problem. In addition, the land structural component s of wetlands, rather than the dynamic processes necessary for wetland developm ent (hydrology, sedimentology, etc.), are usually of primary concern in mitigation proj ects. Short-term pr ojects cannot predict complex, long-term ecological processes. Some as pects of a project (such as the taxa or richness of epibenthic organisms, fish and birds) may be indicative of system maturity, but most (like sediment orga nic content) need longer time (7 or more years) for comparable results (Zedler, 1988). The ef fectiveness of the short-term indicators depends on the predictors used; the predictors ha ve to be able to describe patterns, trends and variability in natural wetlands as the sy stem matures and responds to disturbance and natural variability. In a study to test the predictability of the long-term processes and the short-term expediency of managers, Simenstad and T hom (1996) analyzed 16 ecosystem functional attributes of the Gog-Le-Hi-Te Wetlands in the Puyallup River Estuary in Puget Sound, Washington which were historically modified by industrial development. The remaining wetlands were small, fragmented and cont ained industrial wastewater. The authors monitored topography, sediments, vegetation, water chemistry, emergent plant growth and survival, benthic and inve rtebrate composition, and fish and bird species density. Over the 7-year study, results indicted th at only a few factors showed functional trajectories toward equivalency with natural wetlands (epibenthic or ganisms and fish in high density), and some factors (e.g., sedime nt levels) indicated an immature system (Simenstad & Thom, 1996). The restoration process is influenced by the type of wetland being restored, the region of the country, the ecological functio ns of interest, the type and degree of degradation, the surrounding la nd uses, and the ability to establish and maintain the appropriate hydrology. Attributes of a si milar, nearby wetland can be used for
99 comparison, but success depends on the choice of the reference site. Managers have to look at the wetland type, size, ecological setting, land use, and position in the watershed. Thus by comparing populations of wetla nds, an assessment of the cumulative effectiveness of restoration can be made. In addition, managers should be flexible in their mitigation parameters. For example, if the wetland system is in poor condition, but represent a rare wetland type, a manager s hould choose to improve rather than replace. The EPA inventoried the CWA Section 404 permit records from the 1970s and 1980s from 8 states (Oregon, Washington, Ca lifornia, Texas, Arkansas, Alabama, Mississippi, and Louisiana), and documented 724 permits in the cumulative record, with 898 wetlands impacted and 745 compensatory wetlands required (Zedler, 1996). Not only has the distribution of wetlands within th ese states been altered, but the ecosystem dynamics have become diminished because la nd mangers tend to replace the same types of wetlands. Overall, the EPA concluded that mitigation goals depend on the parameters defined for assessment and the local and regi onal landscape structure, and in these 8 states, the mitigation projects were largely uns uccessful in meeting their stated goals. Several authors have criticized that the Unites States has based legislation on the premature adoption of ideas. For example, Ra ce (1985) contends that the Army Corps of EngineersÂ’ experiments to establish vegetati on on dredged soil was extrapolated by some to mean that entire habitats could be create d. Then, on the presumption that experimental projects would succeed, policy makers adopted the technology as a mitigation measure in the permitting process (Race, 1985). The nation is thus basing our policies on the success of a few projects. On the academic side, published reports do not provide the variability of the site in comparison to ot her sites; consequently it is difficult to distinguish between failures, limited success and ongoing projects (Race, 1985). We are at the point that a framework n eeds to be established for successful and effective mitigation, and we should work togeth er to use upcoming data to modify, refine, and improve policy. The number of permits gr anted is small, but the repercussions of these permits are astounding. The goal of wetland restoration should be to get wetland area back, reverse loss of area, reestab lish natural ecological processes (hydrology, geochemical) and reestablish wetland function. The importance of the resource is not in
100 question, but despite extensive efforts, it remains difficult to quantify the range of benefits provided by wetlands, mainly due to the dynamic nature of the resource. Additionally, a discrepancy exis ts between the short life sp an of monitoring programs (3 to 5 years) and reclamation su ccess. All of these factors co ntribute to an overall feeling of unrest and dissatisfaction w ith current wetland policy. Proposed Solutions to the Science-Policy Gap My first recommendation is the incorpor ation of microbial analysis in wetland mitigation. Microorganisms have proven to be a steadfast indicator of ecosystem health, and the procedures involving laboratory a nd field tests have proved (and continue to prove) very valuable (see Chapter 4). A second recommendation is the generation of annual reviews of the progress of mitigation projects. The subsequent reviews would then refer those previously completed for baselines, but would report new findings. In any given year, a tremendous amount of literature is produced, so annual summaries co uld prove valuable. Academic summaries can exist, as well as those that summarize projects from local, national government or private organizations. Multiple reviews woul d allow for different points of view or emphasis. These reviews could be submitted to legislative bodies for inclusion into new policy. International summaries would also need to be completed; we need to stay abreast of international problem s and solutions because there is a need to start addressing ecological problems with more of a global approach. A third recommendation is the inclusi on of academic researchers (and/or scientists) and wetland organizations into the process of mitigation banking. Land owners or managers that need to compensate for authorized impacts to wetlands associated with development activities may chose to purchase credits from an approved mitigation bank, rather than reclaiming wetlands on or near the development site. The value of a mitigation bank is determined by quantifying the wetland values restored or created in terms of Â“creditsÂ” (Brown & La nt, 1999). Various techniques are used to establish credits for mitigation banks.
101 Banking can provide more cost effective mitigation and reduce uncertainty and delays for qualified projects, especially when the project is associated with a comprehensive planning effort. This techniqu e eliminates the temporal losses of wetland values that typically occur when mitigation is initiated during or after the development impacts occur. Consolidation of numerous small, isolated or fragmented mitigation projects into a single large parc el of may have greater ecological benefit. In addition, the money generated from mitigation banks can then be used for acquisition of land suitable for future restoration made in anticipati on of technology and knowledge to come. A mitigation bank can bring scientific and pl anning expertise and financial resources together, thereby increasing the likelihood of success in a way not practical for individual mitigation efforts. It is my opinion that to compel managers to develop ecosystems that they do not understand is futile, and only c ontributes to poor reclamation results and wasted mitigation dollars. Disadvantages of mitigation banking include the on-site versus off-site debate surrounding wetlands. Mitigated banks are ofte n placed in areas outside the scope of the wetland damaged and thus no longer benefit th e local ecosystem. This transplanted wetland may also not fulfill functions present in the degraded wetland. For example, migratory birds can be displaced and vulne rable other wetland species dislocated. Mitigation banking is not a perfect solution and in some cases may not be the best solution, and we have to look for answers w ithin the framework and resources we have available. Federal support exists for mitigation banking and interagency guidance for the establishment and use of mitigation banks is currently being developed. As of 2002, there were 219 active mitigation banks in the United States, encompassing approximately 120,000 acres in 29 states (Spieles, 2005). Table 5.1 lists some of the mitigation banks in the United States and Figure 5.1 shows their distribution. The Environmental Reorganization Act of 1993 directed the FDEP and state water management districts to adopt rules governing mitigation banking. The FDEP admits that although mitigation banking encourages restoration and promotes interconnected tracts of wetlands, it allows the destruction of sm aller wetlands that provide important habitats for certain species such as the wood stork (FDEP 2006b). For
102 more information on mitigation banking in Florida, please visit the DEP mitigation banking website (http://www.dep.s tate.fl.us/water/). Spieles (2005) conducted a study to determ ine if mitigation banks successfully support native wetland vegetation and if succe ss differs by mitigation method (created, restored, or enhanced), geomorphic class, age, or area. Three measures of ecological status were evaluated: (1) the prevalence of wetland vegetation; (2) the pervasiveness of non-native species; and (3) plant species richness. Sites ranged from less then 2 acres to over 1,300 acres, and included 17 created wetlands, 19 restored wetlands, and 9 enhanced wetlands. Spieles (2005) concluded that the plant community homogeneity increased with age, which indicated a period of self -organization and a pot ential trend toward vegetative equivalence with natural wetlands. Another study examined whether 68 wetland mitigation banks in existence in the United States in 1996 had achie ved Â“no net lossÂ” of wetl and acreage nationally and regionally (Brown & Lant, 1999). Although 74 pe rcent of the individual banks achieve Â“no net lossÂ” by acreage, overall, wetland mitig ation banks are projected to result in a net loss of 21,328 acres of wetlands. The author s stated that alt hough most wetland mitigation banks are using appropriate compensation methods and ratios, several of the largest banks use preservation or enhancement, instead of rest oration or creation. Ten of these banks, concentrated in the Wester n Gulf Coast Region, accounted for over 99 percent of the wetland lost. Even with these results, Brown and Lant advocate mitigation banking as Â“a sound environmental policy a nd planning toolÂ” bu t only if applied according to recently issued guidelines that en sure Â“no net lossÂ” of wetland functions and values (Brown & Lant, 1999, p. 333).
103 Table 5.1: Wetland Mitigation Banks of th e United States (Brown & Lant, 1999).
104 Figure 5.1: The Distribution of Mitigation Banks in the Unit ed States (Brown & Lant, 1999). Allowing wetland ecologists to panel the ex ecutive board of the mitigation banks may offer more effective results. In addi tion, organizations, such as FIPR, should be involved. FIPR can fund projec ts to continuously develop more efficient methods of monitoring and their link to the banking proces s will allow a grass-roots approach to the future policy making. A fourth suggestion to enhancing the sc ience and policy gap is the adoption of newer technology for wetland analysis. (Mathiyalagan, Grunwald, Reddy & Bloom, 2005) have developed a centralized repositor y and mechanism to share geospatial data, information and maps of Floridas wetlands and adjacent agricultural ecosystems. They
105 have also developed an inte ractive WebGIS and geodataba se for FloridaÂ’s wetlands, which provides map and data services. Managers can use ArcIMS (commercially available software) which can be extended using the MSAccess database, Java, Visual Basic and Active Server Pages. The Interactive website of FloridaÂ’s wetlands is accessible at http://www.gi swetlands.ifas.ufl.edu. Cohen, Prenger and Debusk (2005) used visible-near infr ared reflectance spectroscopy (VNIRS) to detect changes in so il quality. Soils are scanned under artificial illumination using a laboratory spectrometer. A multivariate data technique was then used to relate the post-processed reflectance spectra to laboratory observations, such as pH, organic content, total nitrogen, carbon and phosphorus, and extracellular enzyme activity. The VNIRS technology can provide a low-cost, nondestructiv e, rapid analysis method of land use assessment that retains hi gh analytical accuracy for numerous soil performance measures (Cohen et al., 2005). Re search has primarily targeted agricultural applications, but implications for assessing ecological systems are significant. In June 1988, FIPR founded the Centra l Florida Regional Planning Council to develop a geographical database for the sout hern mining district. The database was designed to depict present land use and pr ojected land use to 2010; recognizing that by 2010, most of the minable phosphate ore will have been extracted from the northern portions of the district and the industryÂ’s emphasis will shift almost exclusively to restoration (Brown, 2005). Currently, in Europe, an effort to monito r and assess the environmental impact of mining has produced a project known as MINE O (Chevrel, Reinhard & Klemens, 2002). The objective of the project was to develop hyperspectral remote sensing methods than can be used to measure and monitor mini ng and pollution, and to provide baseline standards across the European Union (EU). Hyperspectral imaging sensors produce data that can characterize the chemical and/or mineralogical composition of the imaged ground surface. The advantages of this imaging technique are the reduction in hazardous, time-consuming and expensive field sampling methods, and the capability to gather repeat data and monitor mining pollution (C hevrel et al., 2002). This technique was tested at the Steirischer Erzberg iron ore mi ne located near the city of Graz in the
106 province of Styria, Austria. Several intricate maps were generated that targeted that landscape degradation that had occurred du e to mining activities, the changes in vegetation, and the levels off contaminati on from carbonatic iron ore. This technology could greatly benefit Florida mining sites and sh ould be investigated for future use. A few publications (Costanza, 1993; Mitsch, Straskraba & Jorgensen, 1988) advocate use of simulation modeling in wetland ecology but this tool has some difficulty predicting what would be considered Â“mitiga tion success.Â” A final component would be to develop connections between structur al measurements (e.g., vegetation density, diversity, productivity) and functional com ponents (e.g., organic sediment acceleration, nutrient retention). Simply having a list of pl ant species is inadequate for lawmakers and managers to effectively estimate the success of mitigated wetlands. All of these technologies and recommenda tions can be applied to the phosphatemined reclaimed wetland mitigation, and I hope that in the future they will.
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Appendix A: Wetland Classification Wetlands are classified by the United States, Department of the Interior, Fish and Wildlife Service in a comprehensive hierarchical method that includes five systems and many subsystems and classes. The method is explained in the Classification of Wetlands and Deepwater Habitats of the United States (Cowardin et al., 1979). This classification method includes the marine system and the estuarine system which are ocean based systems and beyond the scope of this document. The other systems are the riverine, Lacustrine and Palustrine systems. The riverine system includes freshwater wetlands associated with stream channels, while the Lacustrine system includes wetlands associated with lakes larger than 20 acres. The Palustrine system includes freshwater wetlands not associated with stream channels, wetlands associated with lakes of less than 20 acres and other wetlands bounded by uplands. Most forested wetlands are in the palustrine system.
130 Appendix B: Rules Pertaining To Lands cape Restoration as Set Forth by the Bureau of Mine Reclamation in Chapt er 62c-16, Florida Administrative Code. B.1. 62C-16.0051 Reclamation an d Restoration Standards This section sets forth the minimum criteria and standards which must be addressed in an application for a program to be approved. (1) Backfilling and contouring : The proposed land use after re clamation and the types of landforms shall be those best su ited to enhance the recovery of the land into mature sites with high potential fo r the use desired. (a) Slopes of any reclaimed land area shall be no steeper than 4 ft horizontal to one foot vertical to enhance slope st abilization and provide for the safety of the general public. (2) Soil zone (a) The use of good quality topsoils is encour aged, especially in ar eas of reclamation by natural succession. (b) Where topsoil is not used, the operator shall use a suitable gr owing medium for the type vegetative communities planned. (3) Wetlands which are within the conceptu al plan area which are disturbed by mining operations shall be rest ored at least acre-for -acre and type-for-type. (4) Wetlands and water bodies : The design of artificially created wetlands and water bodies shall be consistent with health and safety practices, maximize beneficial contributions within local drainage patterns, provide aquatic and wetlands wildlife habitat values, and maintain downstream water quality by preventing erosion and providing nutrient uptake. Water bodies should incorporate a variety of emergent habitats, a balance of deep and shallow water, fluctuating wate r levels, high ratios of shoreline length to surface area and a variety of shoreline slopes. (a) At least 25% of the highwater surface ar ea of each water body shall consist of an annual zone of water fluctuation to encour age emergent and transition zone vegetation. This area will also qualify as wetlands under the requirements of subsection (4) above, if requirements in paragraph 62C-16.0051(9)(d) are met. In the event that sufficient shoreline configurations, slopes or water le vel fluctuations cannot be designed to accommodate this requirement, this deficien cy shall be met by constructing additional wetlands adjacent to and hydrologically connected to the water body.
131 (b) At least 20% of the low water surface sha ll consist of a zone between the annual low water line and 6 ft below the annual low wa ter line to provide fish bedding areas and submerged vegetation zones. (c) The operator shall provide either of the following water body perimeter treatments of the high water line: 1. A perimeter greenbelt of vegetation consistin g of tree and shrub species indigenous to the area in addition to ground cover. The greenbe lt shall be at least 120 ft wide and shall have a slope no steeper than 30 ft horizontal to one foot vertical. 2. A berm of earth around each water body which is of sufficient size to retain at least the first one inch of runoff. The berm shall be set back from the edge of the water body so that it does not interfere with the ot her requirements of subsection (5). (5) Water quality (a) All waters of the state on or leaving th e property under control of the taxpayer shall meet applicable water quality standards of the Florida Department of Environmental Protection. (b) Water within all the wetlands and water bod ies shall be of sufficient quality to allow recreation or support fish and other wildlife. (6) Flooding and drainage (a) The operator shall take all reasonable steps necessary to eliminate the risk that there will be flooding on lands not controlled by the operator caused by silting or damming of stream channels, channelization, slumping or debris slides, uncontrolled erosion or intentional spoiling or diking or other similar actions within the control of the operator. (b) The operator shall restore the original dr ainage pattern of the area to the greatest extent possible. Watershed boundaries shall not be crossed in restoring drainage patterns; watersheds shall be restored within their original boundari es. Temporary roads shall be returned at least to grade where their exis tence interferes with drainage patterns. (7) Revegetation : The operator shall develop a revege tation plan to achieve permanent revegetation which will minimize soil erosi on, conceal the effects of surface mining, and recognize the requirements for appropriat e habitat for fish and wildlife. (a) The operator shall develop a plan for th e proposed revegetation, including the species of grasses, shrubs, trees, aquatic and wetla nds vegetation to be planted, the spacing of vegetation and, where necessary, the program fo r treating the soils to prepare them for revegetation.
132 (b) All upland areas must have established ground cover for 1 year after planting over 80% of the reclaimed upland area, excluding ro ads, groves or row crops. Bare areas shall not exceed one-quarter (1/4) acre. (c) Upland forested areas shall be establis hed to resemble pre-mining conditions where practical and where consistent with propos ed land uses. At a minimum, 10% of the upland area will be revegetate d as upland forested areas w ith a variety of indigenous hardwoods and conifers. Upland forested area s shall be protected from grazing, mowing or other adverse land uses to allow establishment. An area will be considered to be reforested if a stand density of 200 trees/acr e is achieved at the end of 1 year after planting. (d) All wetland areas shall be restored a nd revegetated in accordance with the best available technology. 1. Herbaceous wetlands shall achieve a ground cover of at least 50% at the end of 1 year after planting and shall be protected from grazing, mowing or other adverse land uses for 3 years after planting to allow establishment. 2. Wooded wetlands shall achieve a stand density of 200 trees/acre at the end of 1 year after planting and shall be protected from grazing, mowing or other adverse land uses for 5 years or until such time as the trees are 10 ft tall. (e) All species used in revege tation shall be indigenous sp ecies except for agricultural crops, grasses and temporary ground cover vegetation
133 Appendix C: Wetland Mi tigation Bibliography