Applied wetlands science and technology

Kent, Donald M. “Frontmatter” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

Library of Congress Cataloging-in-Publication Data Kent, Donald M. Applied wetlands science and technology / edited by Donald M. Kent.—2nd ed. p. cm. Includes bibliographical references. ISBN 1-56670-359-X (alk. paper) 1. Wetland conservation. 2. Ecosystem management. 3. Wetlands. 4. Water quality management. I. Kent, Donald M. QH75 .A44 2000 333.91'8—dc21

00-030927 CIP

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© 2001 by CRC Press LLC Lewis Publishers is an imprint of CRC Press LLC No claim to original U.S. Government works International Standard Book Number 1-56670-359-X Library of Congress Card Number 00-030927 Printed in the United States of America 1 2 3 4 5 6 7 8 9 0 Printed on acid-free paper

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Preface Compared to other ecosystems, wetlands have received an unprecedented amount of attention. Much of this attention has occurred as a result of, and subsequent to, passage of Section 404 of the U.S. Clean Water Act in 1977. The Act recognizes the importance of wetlands to the societal good. That is not to suggest that wetland values were institutionally unrecognized prior to this time. Beginning in the 1930s, wetlands were recognized as valuable for the production and protection of wildlife, especially waterfowl and furbearers. In the 1960s, wetlands were recognized as important for attenuating floodwaters. Now, wetlands are recognized for providing these and other functions, including nutrient and contaminant retention and transformation, groundwater recharge, and production export. Ironically, and coincident with this recognition of wetland value is an awareness that wetlands are disappearing at an alarming rate. Between 1780 and 1980, an estimated 47 million ha of wetlands were lost in the contiguous United States. Of the north central wetlands 75 percent, mostly prairie potholes, were lost between 1850 and 1977. Bottomland hardwood forests were cleared at a rate of 67,000 ha per year between 1940 and 1980. Gulf coast wetlands disappeared at a rate of 10,000 ha per year. Of those wetlands that remain, many are degraded from channelization, damming, and agricultural and urban runoff. Remaining wetlands are fragmented or isolated. This awareness of wetland loss and degradation, and the promulgation of laws and regulations for protecting wetlands and regulating their use, has spawned the development and growth of the wetland professions. This book is for working wetland professionals and nonprofessionals alike. It is intended for managers, regulators, consultants, and developers responsible for effective decision making. The book also is intended for anyone interested in how wetlands function, how wetlands can be protected, and how wetlands can be managed. In some ways, this is a “how to” book, in that it is a guide for working with wetlands. However, we understand that each and every wetland, and each and every situation, is unique and requires a unique solution. As such, the book seeks to provide the guidelines for effective decision making. This second edition of the book, as was the first, was written by practicing wetland professionals. In this manner, the most relevant, up to date information on applied wetland science and technology is available. Each chapter is fully referenced, providing the reader with an opportunity to seek out more detailed information. The book has 14 chapters—2 fewer than the first edition. The reduction in the number of chapters is due to consolidation, and not a reduction, in material. In fact, three new chapters have been added, and several chapters appearing in the first edition have been partly or completely revised. Several new authors have participated in the revision. Chapter 1 provides an introduction to wetland management, including definition and classification. The chapter also discusses legislation and regulation in the United States to provide the context for subsequent chapters. Chapter 2, Wetland Identification and Delineation, is a consolidation of Chapters 2 and 3 from the first edition. ©2001 CRC Press LLC

Carl Tammi has updated the subject in response to recent legislation and policy. Wetland functions and values are discussed in Chapter 3. The subject has been totally revised from the first edition, with an emphasis on evaluating wetland functions and values, including economic values. Chapter 4 is an updated version of the ecological risk assessment material. As in the first edition, David Kent and his coauthors discuss evaluating wetland impacts from anthropogenic chemical, physical, or biological stressors. Chapter 5 discusses avoiding and minimizing impacts to wetlands from anthropogenic activities. Wetlands impacted by anthropogenic activity will require remediation. Enhancement, restoration, and creation of wetlands are methods for remediating impacts. John Zentner, author of Chapter 6, has consolidated the freshwater and coastal enhancement, restoration, and creation chapters from the first edition. Another remediation option is mitigation banking. Chapter 7—Wetland Mitigation Banking, authored by Mike Rolband, Ann Redmond, and Tom Kelsch, is a chapter new to this edition. Chapter 8, Monitoring Wetlands, is a carryover from the first edition. Monitoring is an important component of remediation efforts, as well as fundamental to treatment wetland programs. Tom and Bill DeBusk have written Chapter 9 on treatment wetlands. The chapter is completely new and consolidates the discussions of three distinct chapters in the first edition. Chapters 10 through 13 discuss wetland management. Chapter 10 consolidates Chapters 13 and 14 from the first edition in discussing design and management of wetlands for wildlife. The chapter addresses design issues based upon modern conservation principles and specific management techniques for existing wetlands. Coastal Marsh Management (Chapter 11) continues to be an important issue, and Robert Buchsbaum’s chapter from the first edition is again included in this edition. A new chapter on Watershed Management is included in the second edition. Effective management of wetlands is constrained without consideration of surrounding activities. The theme of broadened perspectives is expanded even further in Chapter 13, Managing Global Wetlands. This new chapter, coauthored by Annette Paulin, recognizes that some wetlands are of international importance and discusses mechanisms for managing these wetlands. The final chapter, Wetlands Education, is the anchor for the book. Karen Ripple has completely revised the chapter. Nevertheless, the intent of the chapter remains the same as that of the first edition—effective wetland management and regulation depends ultimately upon educating the general populace on the value of wetland functions. In closing, I am indebted to the contributing authors. Their contributions are invaluable to those who may read this book and hopefully will influence the way wetlands are managed. I am pleased and honored to have worked with each author and proud to share this book with them. As always, a work of this sort is the result of numerous discussions both past and present. To all those who have stimulated and influenced my thinking, thanks. Donald M. Kent, Ph.D.

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About the Authors Dr. Robert Buchsbaum is Massachusetts Audubon Society’s coastal ecologist and is responsible for applied research on coastal habitats and providing technical analysis on coastal issues. He has published technical papers on a variety of topics including herbivory by Canada geese on saltmarsh plants, nitrogen dynamics in decomposing marsh plants, and on eelgrass wasting disease, and nontechnical papers for the lay public. His recent work includes studies of human impacts on wildlife and water quality in salt marshes and investigations into the role of the environment in the ability of eelgrass to resist disease. Thomas A. DeBusk is the President of Azurea, Inc. and DB Environmental Laboratories, Inc., environmental consulting and research firms located in central Florida. He has 23 years experience with the use of aquatic plants and wetlands for water treatment. Dr. William F. DeBusk is Assistant Professor in the Department of Soil and Water Sciences, University of Florida, Gainesville. He specializes in elemental cycling, wetland ecology, and analysis of ecosystem processes along spatial gradients. Dr. James F. Hobson is Senior Project Advisor for Arcadis Geraghty & Miller in Millersville, MD. He has significant experience in environmental toxicology and testing programs in support of existing and new chemical product registrations under U.S. federal, state, Canadian, and European regulations. He is a diplomate of the American Board of Toxicology and has frequently spoken at and chaired industrial and academic conferences on environmental toxicology issues. Dr. Kenneth D. Jenkins is Director of the Molecular Ecology Institute at California State University and principal in the consulting firm of JSA Environmental in Long Beach, CA. He has been directly involved in numerous ecological risk assessments, especially for hazardous waste sites, and is widely published in the field of ecological assessments. Tom Kelsch is Director of the Conservation Education Initiative which he joined in 1998. Previously, he worked for 8 years as an environmental scientist with the U.S. Environmental Protection Agency’s Office of Wetlands, Oceans and Watersheds in Washington, D.C., including 3 years as Chief of the Wetlands Regulatory Policy Section. He also has extensive experience as an environmental planner for a private consulting firm. He earned a master’s degree in Environmental Studies from Yale University and holds a bachelor’s degree in Landscape Architecture from Michigan State University. David J. Kent is currently an environmental toxicologist and consultant with THE WEINBERG GROUP INC. in Washington, D.C. He formerly was manager of the aquatic toxicology laboratory and project manager on field assessments for ©2001 CRC Press LLC

International Technology Corporation. He has extensive experience in performing and managing ecological assessments for a wide variety of regulatory programs including ecological risk assessments for the pesticide industry and RCRA and Superfund sites. He has numerous scientific presentations and publications to his credit in the areas of aquatic toxicology and ecological risk assessment. Dr. Donald M. Kent is Principal Technical Staff Director with Walt Disney Imagineering Research and Development’s Environmental Science and Technology Group and a consultant. His broad experience includes ecological and environmental research, conduct of functional ecological assessments and impact analyses, design and implementation of mitigation projects, training of international decision makers, and expert technical review. He continues to investigate and develop various applied ecological and biotechnological techniques for integrating land use and natural resource protection. Kevin McManus is Manager of Technical Services for the Massachusetts Water Resources Authority’s Toxic Reduction and Control Program. He has 17 years experience in the environmental field, working in both the public and private sectors. He has specialized in environmental impact assessment, wetlands permitting and mitigation, facility siting, NEPA compliance, pollution prevention, and oil spill response technologies. Annette M. Paulin is a Project Manager with Azurea, Inc., where she conducts original research and develops and implements environmental management and training programs. Her research experience includes studies of nuisance aquatic plants, contaminant removal by natural systems, and copper impacts on aquatic communities. She has also developed training programs for international decision makers. Presently, her duties include managing a volunteer-based water quality monitoring program in central Florida. Ann Redmond is Vice President of Development for Wetlandsbank, Inc. She is responsible for Wetlandsbank’s expansion into new markets, as well as its regulatory and legislative activities. She was with the Florida Department of Environmental Protection for 12 years as the agency’s expert on mitigation and mitigation banking, leading the development and implementation of rules and legislation in these areas. She has worked in ecosystem management initiatives, the conceptual framework for restoration planning, development of functional assessment methods, and sustainability of mitigation projects. She also has worked with a regional water management agency and as an environmental consultant. She earned her bachelor and master of science degrees in Biological Science from the Florida State University with emphases on botany and ecology. Karen L. Ripple is Education Director for Environmental Concern Inc., a nonprofit nonadvocacy corporation devoted to wetland education, wetland research, and the development and application of technology in the construction, restoration, and enhancement of wetlands. She has extensive experience as a public school ©2001 CRC Press LLC

environmental educator and national teacher trainer specializing in wetlands. She has authored several articles and coauthored a book of wetland activities for educators. Prior to her concentration on wetland education, she contributed to a variety of fisheries research projects. Michael Rolband is the president of Wetland Studies and Solutions, Inc. which specializes in water resource issues. He is a member of the Virginia Chesapeake Bay Local Assistance Board, serves on the Fairfax County Stormwater Utility Advisory Group, and is on the boards of the Northern Virginia Chapter of the National Association of Industrial and Office Properties and the National Mitigation Banking Association. He has taught courses on wetland mitigation, permitting, and Chesapeake Bay Preservation Ordinances. He has a bachelor of science and master of engineering degree in Civil and Environmental Engineering, and a master of business administration degree from Cornell University. Carl E. Tammi is a Senior Wetlands Project Manager in the Water Resources Department of ENSR, Inc. He has over 12 years of professional experience in wetland assessment, delineation, permitting, mitigation and restoration design, construction, monitoring, and treatment design. He manages a group of wetland scientists who routinely work throughout the United States on capital projects, remediation and restoration projects, biological monitoring, and applied wetland science research. He is a Certified Professional Wetland Scientist, and received his Provisional Delineator Certification from the U.S. Army Corps of Engineers, Baltimore District. John Zentner is a principal with Zentner and Zentner, a professional consulting firm with offices in California. Zentner and Zentner specializes in planning and restoration throughout the western United States. He specializes in wetland science, land planning, permit processing, and restoration of natural resources. Among other accomplishments, he is President of the Western Chapter of the Society of Wetland Scientists and has participated in various federal, state, and local wetland working groups.

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Contents Chapter 1 Definition, Classification, and U.S. Regulation Donald M. Kent Chapter 2 Wetland Identification and Delineation Carl E. Tammi Chapter 3 Evaluating Wetland Functions and Values Donald M. Kent Chapter 4 Ecological Risk Assessment of Wetlands David J. Kent, Kenneth D. Jenkins, and James F. Hobson Chapter 5 Avoiding and Minimizing Impacts to Wetlands Donald M. Kent and Kevin McManus Chapter 6 Wetland Enhancement, Restoration, and Creation John Zentner Chapter 7 Wetland Mitigation Banking Michael S. Rolband, Ann Redmond, and Tom Kelsch Chapter 8 Monitoring Wetlands Donald M. Kent Chapter 9 Wetlands for Water Treatment Thomas A. DeBusk and William F. DeBusk Chapter 10 Design and Management of Wetlands for Wildlife Donald M. Kent Chapter 11 Coastal Marsh Management Robert Buchsbaum ©2001 CRC Press LLC

Chapter 12 Watershed Management Donald M. Kent Chapter 13 Managing Global Wetlands Annette M. Paulin and Donald M. Kent Chapter 14 Wetlands Education Karen L. Ripple

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Kent, Donald M. “Definition, Classification, and U.S. Regulation” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

1

Definition, Classification, and U.S. Regulation Donald M. Kent

CONTENTS Definition Classification Classification of Wetlands and Deepwater Habitats of the United States A Hydrogeomorphic Classification for Wetlands U.S. Regulation Regulating Agencies Army Corps of Engineers Environmental Protection Agency Fish and Wildlife Service National Resource Conservation Service National Marine Fisheries Service References

DEFINITION Wetlands are defined directly or implicitly in a variety of ways. Several factors, including personal perspective, position in the landscape, and wetland diversity and function, contribute to the tractable nature of the definition. Each individual or group brings to the definition its own perspective based upon cumulative experience and personal needs (Figures 1 to 4). For example, the lay person asked to define wetlands may envision a deep-water marsh teeming with

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ducks, or alternatively a dark swamp. To an engineer a wetland may be a place that will require a specialized construction design to accommodate poorly drained soils. The scientist likely has a functional perspective, defining a wetland as a place where anaerobic processes occur, and plants are adapted specially for living in saturated or inundated conditions. Finally, those charged with regulating wetland use are likely to have a structural perspective, defining wetlands by characteristic soil, hydrology, and plants so as to facilitate permit decision making.

Figure 1

Salt marsh is an emergent, interdial estuarine wetland system characterized by persistent plant species such as cordgrass (Spartina alterniflora). Saltmarsh has a fringe geomorphic setting, and the water source and hydrodynamics are predominantly surface or near surface bidirectional flows.

Defining wetlands is further complicated by their position in the landscape. Wetlands are transitional habitats in the sense that they are neither terrestrial nor aquatic, but exhibit characteristics of both. Their boundaries are part of a continuum of physical and functional characters, and may expand or contract over time depending upon factors such as average annual precipitation, evapotranspiration, and modifications to the watershed. The transitional nature of wetland characteristics and the shifting of wetland boundaries renders precise identification of wetland boundaries difficult if not impossible. The diversity of wetland types also contributes to the tractable nature of the definition. Wetlands include such familiar habitats as marsh and swamp, as well as less familiar seasonal wetlands such as vernal pools and intermittent streams. They may be tidal or nontidal, saline or fresh, lotic or lentic, permanent or impermanent. Vegetation may consist of herbaceous or woody species, or there may be no vegetation.

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Figure 2

This marsh is an emergent, palustrine wetland with a depressional setting. It is subject to vertical fluctuations in water level, and derives its water from precipitation and groundwater discharge.

Figure 3

Palustrine wetlands may also be wooded as illustrated by this hardwood swamp. Swamps frequently have a riverine setting and unidirectional, surface or nearsurface flows.

Wetlands also defy a unifying functional definition. Each wetland is unique with respect to its size, shape, hydrology, soils, vegetation, and its position in the landscape. As such, wetlands exhibit a wide range of functional attributes, including provision of aquatic and wildlife habitat, retention of sediments and toxicants, flood

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Figure 4

This vernal pool in California is inundated or saturated for a short time in early spring. It occurs in a depressional setting and is dependent upon precipitation for water.

attenuation, nutrient metabolism, groundwater recharge, and production export. Individual wetlands may exhibit some of these attributes, all of these attributes, or in rare instances, none of these attributes. Moreover, individual wetlands of similar attributes are likely to provide functions to differing degrees. Despite the difficulty in singularly defining wetlands, several formal definitions have been proposed. The earliest definition was for managers and scientists, particularly those concerned with waterfowl and wildlife (Shaw and Fredine, 1956). Largely a structural definition, it uses language understandable to the lay person. The term wetland … refers to lowlands covered with shallow and sometimes temporary or intermittent waters. They are referred to by such names as marshes, swamps, bogs, wet meadows, potholes, sloughs, and river-overflow lands. Shallow lakes and ponds, usually with emergent vegetation as a conspicuous feature, are included in the definition, but the permanent waters of streams, reservoirs, and deep lakes are not included. Neither are water areas that are so temporary as to have little or no effect on the development of moist-soil vegetation.

The definition established two parameters essential for a habitat to be a wetland: the presence of surface water and the development of moist-soil vegetation. At a workshop of the Canadian National Wetlands Working Group, 23 years later, a definition evolved that recognized a third parameter, hydric soils, and which noted the functional attributes of wetlands (Tarnocai, 1979). Furthermore, it expanded the previous definition of wetland to include not only those habitats with surface water but also those having saturated soils. ©2001 CRC Press LLC

Wetland is defined as land having the water table at, near, or above the land surface or which is saturated for a long enough period to promote wetland or aquatic processes as indicated by hydric soils, hydrophilic vegetation, and various kinds of biological activity which are adapted to the wet environment.

That same year, the U.S. Fish and Wildlife Service adopted a definition that also recognized wetland hydrology, hydric soils, and hydrophytic vegetation as defining parameters (Cowardin et al., 1979). Intended for wetland scientists, the definition is distinguished from the Canadian definition in that a wetland need not exhibit characteristics of all three parameters. 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 purposes of this classification wetlands must have one or more of the following three attributes: (1) at least periodically, the land supports predominantly hydrophytes; (2) the substrate is predominantly undrained hydric soils; and (3) the substrate is nonsoil and is saturated with water or covered by shallow water at some time during the growing season each year.

The three-parameter approach developed for scientists and managers is reflected in Section 404 of the Clean Water Act, forming the basis for regulatory decision making. The term “wetlands” means those areas that are inundated or saturated 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.

CLASSIFICATION Classification is the act or process of classifying; systematically arranging in groups or categories according to established criteria. For example, Linnaean taxonomy is a binomial nomenclature providing for orderly classification of plants and animals according to their presumed natural relationships. Classification schemes may be hierarchical or nonhierarchical. If hierarchical, the classification scheme may be divisive or agglomerative, and monothetic or polythetic. Hierarchical, divisive, polythetic classifications are common and lend themselves to a wide range of scientific applications. Regardless of the scheme used, the effect of classification is to provide a common language. Many wetland classification schemes have been developed (e.g., Martin et al., 1953; Stewart and Kantrud, 1971; Golet and Larson, 1974). The classification system developed by Cowardin et al. (1979) and the classification system developed by Brinson (1993) have received wide acceptance by scientists, policymakers, and managers. These classifications can be applied across broad geographic areas and in large part encompass many other classification schemes. ©2001 CRC Press LLC

Classification of Wetlands and Deepwater Habitats of the United States In 1974, the U.S. Fish and Wildlife Service directed its Office of Biological Services to design and conduct a national inventory of wetlands. Existing classification systems were considered too simplistic (Martin et al., 1953) or too geographically limited (Stewart and Kantrud, 1971; Golet and Larson, 1974; Odum et al., 1974; Zoltai et al., 1975) to satisfy the requirements of a national inventory. Cowardin et al. (1979) developed a classification of wetlands and deepwater habitats of the United States to support the national inventory. The objectives of the classification were to describe ecological units with homogeneous natural attributes, to arrange these units in a system that would aid resource management decisions, to furnish units for inventory and mapping, and to provide uniformity in concepts and terminology throughout the United States. Their classification of wetlands and deepwater habitats of the United States is hierarchical, divisive, and polythetic (Table 1). Wetland and deepwater systems are made up of subsystems, subsystems of classes, and classes of subclasses. Dominance types (plants and animals) are attributed to subclasses. Water regime, water chemistry, and soil modifiers are applied to classes, subclasses, and dominance types. For example, according to the classification, a cordgrass (Spartina sp.) saltmarsh would be estuarine (system), intertidal (subsystem), emergent (class), and persistent (subclass). Systems are a complex of wetlands and deepwater habitats that share hydrology, geomorphology, chemistry, and biology. The classification has five major systems: marine, estuarine, riverine, lacustrine, and palustrine. Subsystems are more specific categories of systems and provide hydrological information. Marine and estuarine systems encompass subtidal and intertidal subsystems. The riverine system includes a tidal subsystem, as well as lower perennial, upper perennial, and intermittent subsystems. Limnetic and littoral subsystems comprise the lacustrine system. The palustrine system is not divided into subsystems. The class describes the general appearance of the habitat. Classes are defined by either the dominant vegetation form or the physiography and composition of the substrate. Examples include rock bottom, aquatic bed, emergent wetland, and forested wetland. Cowardin et al. (1979) intended that the classes be discernible without extensive biological knowledge, and in many cases recognizable by remote sensing. Finer differences in vegetation form or substrate are recognized at the subclass level. For example, rock bottom is divided into bedrock and rubble, and emergent wetland is divided into persistent and nonpersistent. Dominance type is the most precise category and reflects the dominant plant species or dominant sedentary or sessile macroinvertebrate. Class, subclass, and dominance type levels of the classification are more fully described by modifiers. Water regime modifiers describe hydrological characteristics and require detailed knowledge of duration and timing of surface inundation. Salinity in all habitats, and pH in freshwater habitats, are water chemistry modifiers. Soil modifiers are mineral and organic. The classification also includes special modifiers to describe man-made wetlands, and wetlands modified by the activity of persons or beavers. ©2001 CRC Press LLC

Table 1

A Classification of Wetlands and Deepwater Habitats of the United States (Cowardin et al., 1979)

System Marine

Subsystem Subtidal

Intertidal

Estuarine

Subtidal

Intertidal

Riverine

Tidal

Lower perennial

Upper Perennial

Lacustrine

Intermittent Limnetic

Littoral

Palustrine

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Class Rock bottom Unconsolidated bottom Aquatic bed Reef Aquatic bed Reef Rocky shore Unconsolidated shore Rock bottom Unconsolidated bottom Aquatic bed Reef Aquatic bed Reef Streambed Rocky shore Unconsolidated shore Emergent wetland Scrub–shrub wetland Forested wetland Rock bottom Unconsolidated bottom Aquatic bed Rocky shore Unconsolidated shore Emergent wetland Rock bottom Unconsolidated bottom Aquatic bed Rocky shore Unconsolidated shore Emergent wetland Rock bottom Unconsolidated bottom Aquatic bed Rocky shore Unconsolidated shore Streambed Rock bottom Unconsolidated bottom Aquatic bed Rock bottom Unconsolidated bottom Aquatic bed Rocky shore Unconsolidated shore Emergent wetland Rock bottom Unconsolidated bottom Aquatic bed Unconsolidated shore Moss–lichen wetland Emergent wetland Scrub–shrub wetland Forested wetland

Cowardin et al. (1979) designed the classification for use over a broad geographic area, and for use by individuals and organizations with varied interests and objectives. Information about the wetland or deepwater area to be classified is obtained directly from site inspections, or indirectly from maps, aerial photographs, and other sources. The type and extent of information necessarily determine the level to which the area is classified. Cowardin et al. (1979) intended the classification to be openended and incomplete below the class level. Users are expected to identify additional dominance types and add subclasses as necessary. A Hydrogeomorphic Classification for Wetlands A hydrogeomorphic classification for wetlands presented by Brinson (1993, 1995) emphasizes the abiotic characteristics of wetlands. The abiotic characteristics are intended to be indicators of wetland functions. Brinson (1993) suggests that a functionally based classification of wetlands is needed for two reasons. First, the concept of wetlands needs to be simplified to improve communication among researchers, managers, and the public, and this communication is best achieved by focusing on processes that are fundamental to sustained existence. Second, paradigms that clarify the relationship between ecosystem structure and function must be developed to support monitoring of ecosystem health. A hydrogeomorphic classification for wetlands was developed to support ongoing efforts at assessing the physical, chemical, and biological functions of wetlands subject to U.S. Army Corps of Engineers permit. It has its scientific origins in a coastal ecosystem classification system based upon biological, geological, chemical, and physical factors (Odum et al., 1974); a mangrove classification system relying on vegetation and on water source, quality, and flow (Lugo and Snedaker, 1974); and a study by Brown et al. (1979) that found that differences among freshwater forested wetlands are attributed to the amount of water flow. Kangas (1990), who developed a wetland classification system based upon energy flow and landscape properties, also influenced the classification. The classification is nonhierarchical and has three interdependent components: geomorphic setting, water source and its transport, and hydrodynamics (Table 2). Geomorphic setting is defined as the topographic location of the wetland within the surrounding landscape. The seven geomorphic settings tend to have distinct combinations of hydroperiod, dominant direction of water flow, and zonation of vegetation. An individual wetland may have characteristics of more than one geomorphic setting category. Also, a small wetland (e.g., less than or equal to 1 ha) may be difficult to assign to a geomorphic setting category if it is part of a larger wetland complex. Water sources include precipitation, groundwater discharge, and surface or nearsurface flow. The latter include flooding from tides, overbank flow from stream channels, and interflow or overland flow from higher potentiometric surfaces. Precipitation is, of course, a contributor to all wetlands and its relative importance is a function of the contribution of groundwater and surface water. The relative importance of the three sources of water can be determined by construction of a water budget, or more quantitatively by interpreting a hydrograph of the wetland. Bogs

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Table 2

Hydrogeomorphic Classification for Wetlands (Brinson, 1993, 1995)

Geomorphic setting Riverine Depressional Slope Mineral soil flats Organic soil flats Estuarine fringe Lacustrine fringe Water source Precipitation Groundwater discharge Surface or near-surface inflow Hydrodynamics Vertical fluctuations Unidirectional flows Bidirectional surface or near-surface flows

and pocosins are heavily influenced by precipitation, fens and seeps by groundwater discharge, and riverine and fringe wetlands by surface and near-surface flow. Hydrodynamics refers to the direction of flow and strength of water movement within the wetland. There are three qualitative categories of hydrodynamics: vertical fluctuations, unidirectional flows, and bidirectional flows. Astronomic tides, wind, or a combination of both generate the latter. The classification makes inferences about hydrodynamics based upon velocity of flow, rate of water table fluctuations, particle size distribution of bedload sediments, and the capacity to replace soil moisture deficits created by evapotranspiration. These correspond, respectively, to the depressional, riverine, and fringe geomorphic settings. Implementing the classification requires use of indicators of wetland ecological significance, profile development, and comparison to reference wetlands. Indicators are observed in the field, or derived from maps, photographs, water quality data, or other sources. Representative indicators of ecological significance include water characteristics such as suspended sediment and salinity, water color, pH, nutrient status, and soil or sediment characteristics (Table 3). Ecological significance, as determined from direct observation of wetland function or from indicators, is then used to develop a profile for the wetland. A profile may be either narrative or tabular, and is the end point of the classification. Brinson (1993) recommends that profile development be tied to the establishment of reference wetlands that would be used to facilitate assessment, training, and mitigation. The classification emphasizes the interpretation of wetland ecological significance based upon geomorphic setting, water source, and hydrodynamics, and illustrates how fundamental knowledge about water flows and sources can reveal ecological functioning. It does not allow the user to take it to the field for the purpose of matching indicators with functions (Brinson, 1993). Also, the interpretation of ecological significance results in the description of wetland function without necessarily placing the wetland into a discrete category. Brinson (1993) expects that the ©2001 CRC Press LLC

Table 3

Hydrogeomorphic Classification for Wetlands Indicators of Ecological Significance (Brinson, 1993) Water characteristics Suspended sediments Salinity Color Clear Black pH Acid Circumneutral Nutrient status Low Medium High Soil or sediment characteristics Mineral Organic sediments

classification will be used in conjunction with locally recognized wetland names or with classification of wetlands and deepwater habitats of the United States developed by Cowardin et al. (1978).

U.S. REGULATION Numerous federal statutes have been enacted which impact activities in and around wetlands (Table 4). These statutes encompass regulation, acquisition, and restoration of wetlands, incentives and disincentives to use of wetlands, and other programs. Far from being a cohesive collection of synergistic laws, the statutes have been initiated and enacted piecemeal over the years by various federal agencies. Federal authority to regulate wetlands derives principally from Section 404 of the Federal Water Pollution Control Act of 1977 (33 U.S.C. 1344). Section 404 requires landowners and developers to obtain permits prior to dredge and fill activities in navigable waters. The definition of navigable waters has been extended to include adjacent wetlands. However, the Water Pollution Control Act exempts normal agriculture, silviculture, and ranching activities, provided these activities do not convert areas of U.S. waters to uses to which they were not previously subject, do not impair the flow or circulation of such waters, or do not reduce their reach. Due in part to the aforementioned exemptions, Section 404 regulates only about 20 percent of the activities that impact wetlands. The Food Security Act of 1985 (P.L. 99–198 Statute 1354) compensates, in part, for this lack of regulation through three provisions: Swampbuster, the Conservation Reserve Program, and the Wetlands Reserve Program. Swampbuster denies federal farm program benefits to producers who plant an agricultural commodity on wetlands that were converted after December 23, 1985. Although Swampbuster is the only legislative provision that directly affects eligibility for other federal benefits, the policy allowed ©2001 CRC Press LLC

producers to plant a commodity crop in wetlands when prices were high enough to make federal farm program benefits unnecessary, and to plant converted wetlands with a noncommodity crop in years when federal program benefits might be needed. To overcome this deficiency, the Food, Agriculture, Conservation, and Trade Act of 1990 was enacted to make noncomplying producers ineligible for federal benefits for that year and all subsequent years. By contrast, the Conservation Reserve Program authorizes the federal government to enter into contracts with agricultural producers to remove highly erodible cropland from production for 10 to 15 years in return for annual rental payments. Producers are required to implement a conservation plan that usually includes planting cover such as grass or trees to hold soil in place and to reduce erosion. The program was expanded in 1989 to make cropped wetlands eligible for enrollment. Similarly, the Wetlands Reserve Program offers financial incentives for landowners to enhance wetlands in exchange for retiring marginal agricultural land. Landowners may sell a conservation easement or enter into a cost-share restoration agreement to restore and protect wetlands. The landowner continues to control access to the land, and may lease the land for undeveloped recreational activities (e.g., hunting, fishing). Another program that removes land from use is authorized by The Water Bank Act of 1970 (16 U.S.C. 1301). The Water Bank Program provides funds to purchase 10-year easements on wetlands and adjacent areas for the purpose of preserving, restoring, and improving the wetlands. Private landowners enter into agreements with the federal government in which they promise not to drain, fill, level, burn, or otherwise destroy wetlands, and to maintain ground cover essential for the resting, breeding, or feeding of migratory waterfowl. Implementation of the program is concentrated in the prairie pothole region of the United States. Two other programs having a significant affect on the preservation of wetlands are the Migratory Bird Hunting and Conservation Stamp Act of 1934 (16 U.S.C. 718) and the Coastal Barrier Resources Act of 1982 (16 U.S.C. 3501). The Stamp Act uses proceeds from duck stamps to preserve wetlands and adjacent uplands important to waterfowl through purchase or perpetual easement. The Coastal Barrier Resources Act attempts to minimize the loss of human life, wasteful expenditure of federal revenues, and damage to fish, wildlife, and other natural resources by prohibiting federal expenditures and financial assistance for development of coastal barriers. Regulating Agencies Army Corps of Engineers The Corps is responsible for issuing Section 404 permits authorizing dredge or fill activities in waters of the United States and adjacent wetlands. Approximately 15,000 project-specific permit applications and 40,000 minor activities associated with regional and nationwide general permits are evaluated each year (United States General Accounting Office, 1991). Wetland determinations and delineations are also the responsibility of the Corps, including verification of the accuracy of delineations performed by consultants for permit applicants. Public interest reviews are conducted to determine the efficacy of permits. Consideration is given to economics, aesthetics, ©2001 CRC Press LLC

Table 4

Significant Federal Wetlands-Related Legislation Legislation

Date

Effect on Wetlands

Section 10, Rivers and Harbors Act

1899

Migratory Bird Hunting and Conservation Stamp Act Federal Aid to Wildlife Restoration Act

1934 1937

Fish and Wildlife Act

1956

U.S. Fish and Wildlife Coordination Act

1958

Land and Water Conservation Fund Act

1965

National Wildlife Refuge System Administration Act National Flood Insurance Act

1966 1968

National Environmental Policy Act

1969

Water Bank Act Endangered Species Act

1970 1973

Resource Conservation and Recovery Act

1976

Section 402, Federal Water Pollution Control Act Section 404, Federal Water Pollution Control Act Coastal Barrier Resources Act

1977

Requires permits from U.S. Army Corps of Engineers for dredge and fill activities in navigible waterways and wetlands Proceeds of duck stamps used to acquire habitat Assistance to states and territories for restoring, enhancing, and managing wildlife Established U.S. Fish and Wildlife Service Requires all federal projects and federally permitted projects to consider wildlife conservation Purchase of natural areas at federal and state levels Established National Wildlife Refuge System Requires communities to develop floodplain management programs Requires Environmental Impact Statements for federal actions Purchase easements on wetlands Prohibits federal agencies from undertaking or funding projects which threaten rare or endangered species Controls disposal of hazardous waste, reducing threat of contamination to wetlands Authorized national system for regulating sources of water pollution Regulates dredge and fill activities

1977 1982

Food Security Act Swampbuster Conservation Reserve Program Wetland Reserve Program

1985

Emergency Wetlands Resources Act

1986

Agricultural Credit Act

1987

Everglades National Park Protection and Expansion Act North American Wetlands Conservation Act

1989

Food, Agriculture, Conservation and Trade Act

1990

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1989

Prohibits federal expenditure or assistance for development on coastal barriers Discourages farming on wetlands Removes erodible crop land from use Restores and protects wetlands on private property Promotes conservation through intensified cooperation and acquisition efforts Preserves land reverting to Department of Agriculture’s Farmers Home Administration Increased water flow to park and acquired more land Increased protection and restoration of wetlands under the North American Waterfowl Plan Strengthened Swampbuster, Conservation Reserve Program, and the Wetland Reserve Program

Table 4 (continued) Significant Federal Wetlands-Related Legislation Legislation

Date

Effect on Wetlands

Water Resources Development Act

1990

Coastal Wetlands Planning, Protection and Restoration Act

1990

Coastal Zone Management and Improvement Act

1990

Requires federal agency development of action plan to achieve no-net loss Restoration of coastal wetlands and funds North American Waterfowl Management projects Sets guidelines and provides funding for state coastal zone management programs

historic value, fish and wildlife value, value for attenuating floods, navigation, recreation, water supply and quality, and other needs and welfare of the public. Compliance inspections are conducted following permit issuance to ensure permit conditions are met. The Corps has the authority to seek civil or administrative remedies for violation of permit conditions, or for other unauthorized discharges into wetlands. Environmental Protection Agency The Environmental Protection Agency has statutory enforcement authority to deal with unpermitted dredge and fill activities. In addition, the Agency determines the scope of navigable waters and interprets the scope of exemptions under the Section 404 Program. In consultation with the Corps, the Agency developed the guidelines for selection of sites for disposal of dredged or fill materials. The Agency has veto authority under Subsection 404(c) if disposal of dredged or fill material will have an unacceptable adverse effect on municipal water supplies, shellfish beds and fishery areas, and wildlife or recreational areas. Fish and Wildlife Service The Fish and Wildlife Service is an advisor to the Corps with regard to the Section 404 Program, making recommendations for approval or disapproval of permits, and recommending conditions for permits to be approved. In addition, the Fish and Wildlife Service is active in a number of programs designed to protect, restore, and enhance wetlands. The Fish and Wildlife Service helps the National Resource Conservation Service map agricultural wetlands and select and manage wetlands protected under the Farmers Home Administration Conservation Program and the Wetlands Reserve Program. The Fish and Wildlife Service also implements restoration projects on highly erodible cropland under the Conservation Reserve Program and assists the Department of Agriculture and individual farmers in designing wetland conservation plans necessary to qualify for Farm Bill incentives. Other activities of the Fish and Wildlife Service include management of the National Wildlife Refuge System, research, and development of National Wetlands Inventory maps.

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National Resource Conservation Service The U.S. Department of Agriculture National Resource Conservation Service administers and enforces Swampbuster, the Conservation Reserve Program, and the Wetland Reserve Program. This includes providing wetlands information to producers and third parties, identifying and delineating wetlands, monitoring compliance with regulations, responding to public complaints and producer’s appeals of decisions, and dealing with violations. Operations are carried out in conjunction with other federal and state agencies and county committees. National Marine Fisheries Service The National Marine Fisheries Service is active in coastal wetland issues and makes recommendations to the Corps regarding Section 404 permits under authority of the Fish and Wildlife Coordination Act. Field staff also work closely with state fish and wildlife agencies and water quality agencies.

REFERENCES Brinson, M. M., A Hydrogeomorphic Classification for Wetlands, Technical Report WRP-DE-4, U.S. Army Engineer Waterways Experiment Station, Vicksburg, MS, 1993. Brinson, M. M., The HGM approach explained, Natl. Wet. Newsl., November/December, 7, 1995. Brown, S., Brinson, M. M., and Lugo, A. E., Structure and function of riparian wetlands, in Strategies for the Protection and Management of Floodplain Wetlands and Other Riparian Ecosystems, Johnson, R. R. and McCormick, J. F., Tech. Coord., Forest Service General Technical Report WO-12, U.S. Department of Agriculture, Washington, D.C., 1979, 17. Cowardin, L. M., Carter, V., Golet, F. C., and LaRoe, E. T., Classification of Wetlands and Deepwater Habitats of the United States, U.S. Department of the Interior, Fish and Wildlife Service Biological Services Program FWS/OBS-79/31, 1979. Golet, F. C. and Larson, J. S., Classification of Freshwater Wetlands in the Glaciated Northeast, U.S. Fish and Wildlife Service, Resource Publication 116, 1974. Kangas, P. C., An energy theory of landscape for classifying wetlands, in Lugo, A. E., Brinson, M. M., and Brown, S., Eds., Forested Wetlands, Elsevier, Amsterdam, 1990, 15. Lugo, A. E. and Snedaker, S. C., The ecology of mangroves, Ann. Rev. Ecol. Syst., 5, 39, 1974. Martin, A. C., Hotchkiss, N., Uhler, F. M., and Bourn, W. S., Classification of Wetlands of the United States, U.S. Fish and Wildlife Service, Special Scientific Report, Wildlife 20, 1953. Odum, H. T., Copeland, B. J., and McMahan, E. A., Eds., Coastal Ecological Systems of the United States (4 Volumes), The Conservation Foundation, Washington, D.C., 1974. Shaw, S. P. and Fredine, C. G., Wetlands of the United States, Their Extent, and Their Value for Waterfowl and Other Wildlife. U.S. Department of Interior, Fish and Wildlife Service, Circular 39, Washington, D.C., 1956. Stewart, R. E. and Kantrud, H. A., Classification of Natural Ponds and Lakes in the Glaciated Prairie Region, U.S. Fish and Wildlife Service, Resource Publication 92, 1971.

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Tarnocai, C., Canadian wetland registry, in Proceedings of a Workshop on Canadian Wetlands Environment, Rubec, D. D. A. and Pollett, F. C., Eds., Canada Land Directorate, Ecological Land Classification Series, No. 12, 1979, 9. United States General Accounting Office, Wetlands Overview: Federal and State Policies, Legislation and Programs, GAO/RCED-92–79FS, Washington, D.C., 1991. Zoltai, S. C., Pollett, F. C., Jeglum, J. K., and Adams, G. D., Developing a wetland classification for Canada, Proc. N. Am. For. Soils Conf., 4, 497, 1975.

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Tammi, Carl E. “Wetland Identification and Delineation” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

2

Wetland Identification and Delineation Carl E. Tammi

CONTENTS Off-Site Wetland Identification Identification Resources Interpreting Resources U.S. Geological Survey (USGS) Topographic Maps U.S. Fish and Wildlife Service (USFWS) National Wetland Inventory Maps U.S. Department of Agriculture Natural Resources Conservation Service Soil Surveys and the Hydric Soils of the United States List Comparison and Corroboration Aerial Photographs U.S. Geological Survey Surficial Geologic Maps Individual State Wetland Maps On-Site Wetland Delineation Wetland Hydrology Hydrological Field Indicators Hydric Soils Hydric Soil Field Indicators Hydrophytic Vegetation Indicators of Hydrophytic Vegetation Identifying and Delineating Wetlands Undisturbed Areas Disturbed Areas Difficult Areas Aids to Delineation References

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Wetland identification and the science of delineation are regulatory-driven activities that are commonly required in land-use development, planning, exploration, and a host of related activities involving future site expansion. Although federally mandated wetland regulatory statutes have been in existence for over 25 years, the science of identifying and delineating the extent and types of wetlands has been consistently evolving. As the science has evolved, a greater awareness of the functions and values wetlands provide has occurred with the resultant development of extensive wetland identification and delineation resources within the last 10 years. Today, the land-use planner, wetland scientist, and manager have a range of tools in print, graphic, and electronic format available to assist in making wetland determinations and defendable jurisdictional delineations. Typically, the science of identifying and delineating wetlands is a two-tiered process. An initial office-based off-site assessment is conducted for identification purposes. A legally binding jurisdictional determination requires an on-site field assessment called a wetlands delineation. Identifying the location and determining the areal extent of jurisdictional wetlands is an important consideration for those involved in land use management, development, remediation, or assessment. Today, defining wetland limits and boundaries is primarily driven by comprehensive federal and, where applicable, state and local land-use laws and regulations. Section 404 of the Clean Water Act is the principal tool that the U.S. Army Corps of Engineers and the U.S. Environmental Protection Agency use to regulate the discharge of dredged or fill material into waters of the United States, including wetlands (33 CFR 320–330). At the federal level, wetlands are further defined from a regulatory viewpoint as, “Those areas that are inundated or saturated by surface or groundwater 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” (33 CFR 328.3). In identifying and delineating federal jurisdiction wetlands, three essential technical criteria or factors are applied: the presence of wetlands hydrology through surficial or groundwater; a prevalence of wetland vegetation (hydrophytes) that typically has specialized morphological and physiological adaptations to tolerate saturated or inundated conditions; and wetland soils (hydric soils), which in their undrained condition exhibit characteristics of somewhat poorly drained, poorly drained, or very poorly drained soils. Other major federal legislation that drives wetland identification includes Section 401 Water Quality Certification (delegated to the individual states), Section 10 of the Rivers and Harbors Act of 1899 and the National Environmental Policy Act. Many states have promulgated and adopted wetland protection legislation for inland, and where applicable, coastal wetlands. Identification and delineation techniques vary slightly from state to state, although most have adopted the principles of the federal methodology (to be described in greater detail later). Given the regulatory framework behind wetland protection, it is incumbent upon project proponents and land-use managers to determine, locate, and identify wetland resources on a subject parcel. Furthermore, it is important to adequately and accurately determine the location and approximate areal extent, as well as the predominant wetland cover type, early in project planning stages to avoid wetland impacts ©2001 CRC Press LLC

and resultant time-consuming permit decisions. This action can streamline the permitting process during more advanced stages of project design through avoidance and minimization of wetland impacts. An off-site macroscale wetland determination makes a positive or negative wetland determination for a subject parcel, and determines the approximate location of wetland and deepwater areas. It also determines the approximate areal extent and distribution of wetland and deepwater areas, and the predominant wetland cover type (Cowardin et al., 1979). Finally, an off-site macrosite wetland determination assesses the need for continued analysis and approximate level of effort associated with any analyses. In some instances, information relative to the potential presence of hydric soils, surficial hydrology, and site disturbance can be determined from off-site wetland determinations. Historical and current land use as it pertains to wetland resources can also be ascertained in many circumstances. By making initial determinations and preliminary conclusions regarding the aforementioned factors, a project proponent can make informed decisions, save valuable time and expense, and determine if detailed on-site investigations are necessary. The level of effort to conduct off-site investigations can vary greatly, and can be tailored to suit individual site permitting or project requirements.

OFF-SITE WETLAND IDENTIFICATION For the purposes of this chapter, off-site identification of wetlands is defined as assembling and interpreting readily available natural resource mapping and reports and other documents, both published and unpublished, from existing sources, for the sole purpose of identifying, locating, and describing wetland resources on a given site or parcel of land. By applying existing resource document information, the researcher can make initial determinations relative to the perceived presence or absence of one, two, or sometimes three of the parameters necessary for an area to be considered a jurisdictional wetland. In instances where on-site inspection is not necessary or is beyond the scope of the investigation (e.g., National Environmental Policy Act wide range alternatives analyses, or limited environmental assessments), off-site wetlands determinations may be the only source of information for environmental planning decisions. The overall accuracy of off-site wetland determinations is a function of the quality of the information (sources) used and the ability of an individual(s) to interpret the data. The keys to conduct of an effective and technically valid analysis include the following: • • • •

Define the project scope and goals prior to conducting the analysis. Ensure that a wide range of available sources are investigated and used. Emphasize comparison and corroboration between different sources for the same site. Obtain recent data, but also data that cover many different years to assist in understanding the site history. • Understand individual resource document symbols and interpretation keys. • Understand regulatory requirements for documentation. ©2001 CRC Press LLC

The primary objective of off-site wetland determinations is identifying and determining whether wetlands exist on a parcel, followed by the approximate distribution and areal extent. In determining and quantifying these parameters, the key is corroboration between different sources. That is, not only locating wetlands on a subject parcel from a single source, but corroborating the identification through multiple sources. Another important objective of off-site determinations is documenting the dominant wetland cover type on parcels that have been preliminarily determined to have wetlands within their boundaries. Depending on the source, an interpreter can determine whether the wetlands are forested, scrub–shrub, emergent, aquatic bed, or open water. Detailed interpretation requires a greater level of effort and expertise but can result in greater detail, such as evergreen forest vs. deciduous forest, or persistent emergent vs. nonpersistent emergent, or artificially created vs. naturally occurring. Classification schemes can be tailored to an individual state’s system, or the widely accepted federal system developed by the U.S. Fish and Wildlife Service (Cowardin et al., 1979) and now recognized as the Unified Federal Classification Scheme (Federal Geographic Data Committee, 1995). Site soil characterizations and surficial hydrological features can also be recognized and described from off-site resources. Published sources exist which reveal site soils mapping to varying levels of detail and accuracy. Determining the hydrological regime, or simply the hydrology of a wetland, is a significant feature in determining the areal extent of wetlands both in the field and from mapped sources. Off-site interpretation can reveal a wetland’s hydrological source, as well as its drainage features. Identification Resources The first step in offsite wetland interpretation studies is identifying and obtaining readily available sources of information. Resources are generally diverse, with varying levels of accuracy. Also, resources have been dramatically expanded in recent years with many new tools available to the interpreter. These resources are generally available and provide a baseline of information from which to work. • U.S. Geological Survey (USGS) Topographic Maps, Standard Edition and Provisional Edition (7.5 minute or 15 minute quadrangles, scales 1:24,000 or 1:25,000, continental United States, 1:20,000 Puerto Rico, 1:63,360 Alaska), U.S. Department of the Interior Geological Survey National Mapping Division. • U.S. Department of the Interior/Fish and Wildlife Service (USFWS) National Wetland Inventory Maps (scale 1:24,000, continental United States, 1:63,360, Alaska), interpreted and adapted from High Altitude Aerial Photography and superimposed on U.S. Geological Survey Topographic Maps. • U.S. Department of Agriculture Natural Resources Conservation Service County Soil Surveys, in cooperation with individual state agriculture experiment stations; used in conjunction with the hydric soils of the United States, 1991, National Technical Committee for Hydric Soils, U.S. Department of Agriculture Natural Resources Conservation Service.

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• Aerial photography (stereo-paired, black and white, color, color infrared; positive transparency/aero negative; various scales and dates), Federal, State, and Commercial Suppliers. • U.S. Geological Survey Surficial Geologic Map Quadrangles (7.5 minute quadrangles, scale 1:24,000), U.S. Department of the Interior Geological Survey. • Individual state wetland maps (limited coverage and level of accuracy).

Interpreting Resources This section describes in more detail the analysis and interpretation of the resources listed above. Although the level of detail and accuracy varies with each source, a first-time evaluator should be able to extract sufficient information to reasonably determine if wetlands are present on-site, and the approximate historical or current location and extent of wetlands. U.S. Geological Survey (USGS) Topographic Maps The U.S. Department of the Interior, Geological Survey National Mapping Division generates 7.5- and 15-minute topographic maps through the National Mapping Program. Available are two separate editions, the Standard Edition Maps and the Provisional Edition Maps, each produced at 1:24,000 (English units) or 1:25,000 (metric units) for the continental United States. Standard Edition Quadrangles represent a finished product with the earth’s topographic relief depicted by contours. Provisional Edition Quadrangles represent an updated draft format, including hand lettering and limited descriptive labeling of some physical features. Stereoplotting field verified, high altitude aerial photographs produce both editions. Some quadrangles are mapped by a combination of orthophotographic images and map symbols, with orthophotographs derived from aerial photographs by removing image displacements owing to camera tilt and terrain relief variations (USGS, 1991). The use of USGS Topographic Maps for off-site wetland identification is often the first step to evaluate a site’s physical features. In addition to topographic, hypsographic, infrastructure, and other physical features, the USGS Topographic Maps provide detailed information relative to vegetation cover types, surface features, coastal features, hydrographic features such as rivers, lakes, and canals, and submerged areas and bogs. Figure 1 is a section of an USGS Quadrangle and depicts some of these features. The section includes wooded marsh or swamp in the western and southern parts of the site, perennial ponds or lakes in the central part of the site, and perennial streams associated with cranberry bogs in the northeast part of the site. Although most of the wetland and open water interpretation keys that accompany the maps are self-explanatory, the individual submerged areas and bogs keys require a little elaboration to distinguish among different wetland cover types. Marsh or swamp designations are wetlands characterized by saturated soil conditions in the root zone (as opposed to inundation), with emergent, herbaceous, or aquatic bed vegetation as the dominant cover class. An example would be a rush (Juncus spp., Scirpus spp.) and sedge (Carex spp.) dominated wet meadow. ©2001 CRC Press LLC

Figure 1

A U.S. Geological Survey topographic map.

Submerged marsh or swamp designations indicate an inundated root zone condition with emergent, herbaceous, or aquatic bed vegetative dominants. A typical example is a broad-leaved cattail (Typha latifolia) or pickerelweed (Pontederia cordata) marsh. Wooded marsh or swamp is a wetland characterized by saturated soil conditions with shrub, sapling, or mature forest as the dominant cover class. A saturated red maple (Acer rubrum) swamp is an example. Submerged wooded marsh or swamp indicates root zone inundation (ponding) as the dominant water regime with shrub, ©2001 CRC Press LLC

sapling, or mature forest as the dominant cover class. A bottomland hardwood forest dominated by cypress (Taxodium spp.) trees is an example. Land subject to inundation can be floodplain and flood-prone areas that may support wetland hydrology and wetland vegetation (hydrophytes). Rice fields and cranberry bogs are examples of anthropogenically influenced wetland areas. Some of the advantages to using USGS Topographic Maps include the relative accuracy of the topographic contours in undisturbed areas, photointerpretation documentation is groundtruthed at regular intervals, and individual quadrangles are periodically photorevised which assists in chronological evaluation of a site’s history. The limitations in using USGS Topographic Maps include interpretation problems associated with the small scale (1 in. equals 610 m) of the maps, and smaller wetlands often are frequently unmapped. In some parts of the country, quadrangles may be too outdated to be of use. U.S. Fish and Wildlife Service (USFWS) National Wetland Inventory Maps The USFWS initiated the National Wetland Inventory (NWI) program and mapping in 1975 to assess, measure, and characterize the extent of wetlands and open water areas throughout the United States. The NWI Maps are produced from photointerpretation of high altitude, stereo, aerial photographs. High altitude aerial photographs were selected over satellite imagery because of the problems satellite imagery had in capturing optimum water conditions for wetland detection, detecting smaller wetlands, and identifying forested wetlands (Tiner and Wilen, 1983). The NWI Maps are developed from 1:60,000 color-infrared aerial photographs. Photointerpretation of the aerials provides a three-dimensional image, thus allowing the interpreter to identify trees from shrubs, while considering shade and slope. Wetland and open water types are differentiated based on their characteristic photographic signatures. NWI Maps are developed according to a comprehensive evaluation process (Tiner and Wilen, 1983). The preliminary field investigations and photointerpretation of high altitude aerial photographs is the initial step, with review of existing wetland information and quality control of the interpreted photographs completing the first phase. Draft map production is initiated with a subsequent interagency review of draft maps and final map production. Available are two series of NWI Maps: the 1:100,000/1:250,000 scale and large scale 1:24,000. The USGS Topographic Map Quadrangle is used as a base map with wetland and deepwater areas depicted as overlays. A new wetland classification was developed by USFWS to correspond with the NWI Maps. Classification of wetlands and deepwater habitats of the United States (Cowardin et al., 1979) describes individual wetland ecological attributes and arranges them in a hierarchical system that facilitates resource management and inventory. The three key components in the ecological hierarchy are hydrophytes, hydric soils, and hydrology. The use of NWI Maps in off-site wetland identification typically provides the greatest level of detail and accuracy with the least amount of interpretation effort ©2001 CRC Press LLC

and expertise. The USFWS Classification System is a comprehensive and progressive inventory that groups wetlands into one of five major systems, marine, estuarine, lacustrine, riverine, and palustrine, based upon hydrologic, geomorphologic, chemical, and biological factors (Cowardin et al., 1979, see Chapter 1). The hierarchy progresses through subsystems, classes, and subclasses that further refine and describe specific wetland structural (vegetation, hydrology, dominant life form, etc.) components. Figure 2 depicts a representative section from an NWI Map that corresponds with Figure 1, and that indicates several different wetland classes within the palustrine system. Comparing Figure 2 with Figure 1, and using the interpretive key that accompanies the map, the interpreter is able to determine that the wooded swamp or marsh of the USGS Map has been further defined as palustrine forested broad-leaved deciduous wetland. An interpreter can become familiar with this system with a little practice resulting in quick characterizations of site conditions relative to wetland types. The accuracy of NWI Maps varies between systems and classes, with the highest degree of accuracy occurring for large marine, lacustrine, and estuarine systems. Less accurate are smaller mapped units for palustrine wetlands, specifically palustrine forested wetlands. The latter can be misstated owing to photointerpretation difficulties encountered as a result of leaf-in periods, when the interpreter cannot accurately describe the forest floor (MacConnell et al., 1989). NWI Maps provide the greatest diversity of all off-site references, with the possible exception of aerial photographs. However, the latter require a greater degree of photointerpretation expertise, and the NWI maps were prepared for the express purpose of identifying and classifying wetlands. Through use of the USFWS Classification system, an interpreter can characterize a wetland’s system, the dominant vegetative structural life form (e.g., forested, emergent, aquatic bed), its hydrological regime (e.g., intermittent vs. perennial), and substrate (e.g., rock bottom or unconsolidated bottom). The taxonomy also has provisions for documenting anthropogenic influence on created or farmed wetlands (e.g., palustrine farmed cranberry bogs and palustrine open water artificially excavated). The limitations of NWI Maps for off-site wetlands identification include the small scale (1 in. equals 2000 ft), errors associated with photointerpretation of select cover types (principally deciduous forest), limited field verification, and the lack of photorevision since initial production. NWI Maps are beneficial as a qualitative reference and are one of the only federally produced and readily available documents for the sole purpose of identifying, inventorying, and characterizing wetlands. U.S. Department of Agriculture Natural Resources Conservation Service Soil Surveys and the Hydric Soils of the United States List The U.S. Department of Agriculture Natural Resources Conservation Service produces County Soil Surveys in cooperation with the individual state’s agricultural experiment station. Programs have mapped individual soil series based on comprehensive field investigations conducted by Natural Resources Conservation Service and State soil scientists. To produce the maps, soil scientists observe the steepness, length and shape of slopes, the size and velocity of streams, the kinds of native ©2001 CRC Press LLC

Figure 2

A National Wetland Inventory map.

plants and rocks, and evaluate soil profiles (U.S. Department of Agriculture Natural Resources Conservation Service, 1978). Soil profiles are examined to the depth of the parent material and are compared to soil profiles examined in other counties for the purpose of comparing and contrasting known soil series. A unified soil taxonomy, the U.S. Department of Agriculture Soils Classification, is used across the nation to characterize and classify soil types. Soil series and soil phase are the most common terms used in describing individual soil types. A soil ©2001 CRC Press LLC

series is a grouping of soils that have similar profiles and major horizons and are named after the town in which the series was first discovered (U.S. Department of Agriculture Natural Resources Conservation Service, 1978). Phases further refine the series based on the texture in the surface layer, slope, or stoniness (U.S. Department of Agriculture Natural Resources Conservation Service, 1978). Soil mapping units and boundaries are depicted as overlays on high altitude aerial photography and are originally drafted by the field soil scientists. These boundaries are further refined following laboratory analysis of soil properties. The finished product indicates soil boundary delineations, soil series descriptions, and biophysicochemical properties. Additional sections of the soil survey provide information about recommended use and management of the soils, soil properties, and soil formation. Use of soil surveys for off-site wetland identification is limited to the identification and distribution of hydric soils. Hydric soils, one of the three essential characteristics of a federal jurisdictional wetland, have unique physical properties that set them apart from nonhydric soils. A hydric soil is defined as “a soil that is saturated, flooded, or ponded long enough during the growing season to develop anaerobic conditions in the upper part” (U.S. Department of Agriculture Natural Resources Conservation Service, 1991). The development of hydric soils is ultimately driven by the presence of wetland hydrology, and under sufficiently wet conditions (root zone saturation and inundation), hydric soils support the growth of hydrophytic vegetation. The U.S. Department of Agriculture Natural Resources Conservation Service has developed a list of hydric soils of the United States by applying criteria (e.g., drainage class, organic vs. mineral, etc.) of the National Technical Committee for Hydric Soils (U.S. Department of Agriculture Natural Resources Conservation Service, 1991). Included in this list are most of the somewhat poorly drained soil series, and all of the poorly drained and very poorly drained soils. Hydric soil classifications have been developed based on taxonomic and morphologic features. Tools subsequently have been developed for assisting in the field determination of drainage classes (New England Hydric Soils Technical Committee, 1998). Through the use of individual soil surveys, an interpreter can determine all mapped soil series on-site and then cross-reference the list of soils with Hydric Soils of the United States. Figure 3 is a soil survey map of the site depicted in previous figures. Examination of the site and cross-reference with the hydric soils list indicates that Sanded Muck (SB), Peat (Pe), Freshwater Marsh (Fr), Scarboro Sandy loam (ScA), Muck (Mv), and Au Gres and Wareham loamy sands (AuA) are hydric soil. The occurrence of hydric soils corresponds with the wooded swamp or marsh, perennial lakes or ponds, and cranberry bogs of the USGS Map, as well as the palustrine forested broad-leaved deciduous, palustrine farmed wetlands, and palustrine open water of the NWI Map. Although soil mapping involves an intensive field effort, the accuracy of the soil maps is quite variable, and areas mapped as hydric soils (a hydric soil series) can contain inclusions of nonhydric soils. Conversely, areas of nonhydric soils may contain hydric inclusions. The soil mapping information is best used as a macroscale assessment tool and should not be used for definitive boundaries of hydric soils. An

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Figure 3

A U.S. Department of Agriculture Soil Conservation Service soils survey.

advantage in using soil surveys and the list of hydric soils is that the interpreter does not need to spend time learning the U.S. Department of Agriculture Soils Classification and taxonomy system to be able to locate areas of mapped hydric soils on a site, although it is recommended that the interpreter understand the basic principles underlying the criteria for listing a soil as a hydric soil. ©2001 CRC Press LLC

Comparison and Corroboration USGS Topographic Maps, USFWS NWI Maps, and U.S. Department of Agriculture Natural Resources Conservation Service Soil Surveys are typically the most accessible resources and require the least amount of technical knowledge to interpret. The effectiveness and level of accuracy in conducting off-site wetlands interpretation studies using these resources is a function of the time needed to obtain each source and interpreting the information. Emphasis should be placed on comparing and contrasting individual sources: that is, comparing and corroborating the results from USGS to NWI to the Soil Survey, being sure to consider the years each source was initially produced. For example, the USGS indicates wooded marsh or swamp throughout the western and southern sections of the reference site (Figure 1), the NWI indicates palustrine forested broad-leaved deciduous wetland in the western and southern sections of the site (Figure 2), and the U.S. Department of Agriculture Natural Resources Conservation Service indicates hydric soils in the western and southern sections of the site (Figure 3). Therefore, the interpreter can be reasonably confident that wetlands, and most likely forested wetlands, exist at the site. In an effort to further refine the information already obtained from these resources, additional resources can be consulted and evaluated. Aerial Photographs Aerial photography has been used since the 1860s for remote sensing land-use patterns and activities through the use of hot-air balloons (McKnight, 1987). These actions spawned the development of photogrammetry, the science of obtaining reliable and defensible measurements from photographs and mapping from aerial photographs (Ritchie et al., 1988). Historically, the interpretation of aerial photography has fallen under the term remote sensing, with the net result being that aerial photographs were the only tool used in remote sensing. Contemporary views have altered the term to include a wide range of tools and analytical devices. One recent definition stated remote sensing is the measurement of reflected, emitted, or backscattered electromagnetic radiation from the earth’s surface using instruments stationed at a distance from the site of interest (Roughgarden et al., 1991). Nonetheless, aerial photography is used today as a powerful source for remote sensing land use, including wetland identification, characterization, and perturbation as a result of anthropogenic activities. Stereo-paired vertical contact prints provide the most useful and scientifically defensible information as a three-dimensional image of the earth’s surface is presented to the interpreter via a stereoscope. This three-dimensional image is obtained through photographs taken in stereo pairs with end overlap. The photographs are interpreted with a stereoscope, which allows the interpreter to closely examine site conditions by adding depth of field, and provides the ability to distinguish wetland cover types. Different cover types have characteristic signatures that can be quantified based on observable color, tone or hue, shadow, texture, and depth of field. Color signatures can be further refined using the Inter-Society Color Council and

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National Bureau of Standards (ISCC-NBS) method of color description, using Centroid Color Charts (Smith and Anson, 1968). Permanently and seasonally inundated forested, scrub–shrub, emergent, and open water areas are relatively easy to distinguish from upland habitats using aerial photographs. Surface water has a different reflectance pattern than dry areas. The difficulty comes when trying to identify seasonally saturated wetlands, particularly seasonally saturated forested wetlands. Aerial photographs are available in many formats, scales, and geometry (Figure 4). The most common photograph types used for wetland identification are vertically oriented black and white, color, and color infrared using aerographic film. Aerial photograph geometry (i.e., basis of the angular relationship to the earth’s surface) can be divided into two major categories: oblique (including low oblique and high oblique) and vertical (see Figure 5 and Hudson and Lusch, 1990). Oblique aerial photographs have an advantage to the interpreter, in that ground features can be interpreted from a familiar point of view (McKnight, 1987). In addition, owing to the orientation of the camera for oblique photos, the stereoscopic effect is reduced, thereby negating the three-dimensional effect. However, due to measurement inadequacies and scale development, vertically oriented photographs are used for determining quantitative information and provide more useful and defensible information for wetland identification. Combining aerial photographs with groundtruthing can be a very effective and cost-efficient means to identifying and delineating wetland boundaries on large parcels. Film types are also an important consideration in identifying wetlands. Panchromatic black and white photographs are the least expensive and most common. However, color infrared photographs have a demonstrated and significant advantage. Color infrared discriminates between living and dead vegetation, enhances open water areas, and discriminates natural features from man-made features. Larger scale, low-altitude aerial photography is recommended over smaller scale, high-altitude aerial photography for clarity, ease of interpretation, distinguishing ground features, and general resolution. The most useful and accurate interpretations are made through chronological analysis of a site. Aerial photographic coverage and availability significantly limit its use for offsite wetland identification. Federal and State agencies have inventories with spotty coverage, which may not be available for purchase. The National Archives (in Utah) maintains an active database and inventory of aerial photographs (black and white, color infrared, stereo-paired) covering the entire continental United States for select years, and these photographs are available for purchase. They are generally highaltitude, small-scale aerial photographs. Private commercial suppliers often maintain inventories, and usually specialize within a region (i.e., New England, Northwest, Southeast). Coverage is largely unpredictable varying from complete chronological coverage over several years to no coverage at all. Purchase costs can be quite high, with some firms charging access fees for database reviews. Because aerial photography provides the base map and framework for most other off-site resources, it is apparent how important it can be when used in its unrevised form. Stereo aerial photographic interpretation has evolved into a scientific and technical discipline of its own and, in some instances, requires considerable expertise ©2001 CRC Press LLC

Figure 4

A vertical black and white aerial photograph.

to extract valid information. Nonstereo paired aerial photographs can be used to supplement the other off-site information in a qualitative manner. There is much variability among nonstereo aerial photographs, especially relative to geometry, ©2001 CRC Press LLC

HIGH OBLIQUE

LOW OBLIQUE

VERTICAL

90o

Figure 5

Aerial photography angular orientation.

scales, film types, and coverage. When conducting preliminary qualitative off-site wetland reviews, interpretation of aerial photographs may not be necessary. However, if the goal is to quantify wetland site conditions over time, aerial photographs may prove to be an indispensable tool. U.S. Geological Survey Surficial Geologic Maps The U.S. Geological Survey produces surficial geologic maps primarily for use as indicators of geologic zonation above bedrock. These maps provide some detail relative to soil and subsurface composition and are helpful in locating and identifying swamp deposits, alluvium, surface water bodies, and other wetland features. Figure 6 ©2001 CRC Press LLC

illustrates the location of surface waterbodies, swamp deposits (Qs), and cranberry bogs on the reference site. The swamp deposit designations are indicative of organic matter, clay, silt, and sand accumulating in swamps (USGS, 1967).

Figure 6

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A U.S. Geological Survey surficial geologic map.

Individual State Wetland Maps Several individual states have produced wetland maps for use in macroscale planning, and at least in one instance (MacConnell et al., 1989), for jurisdictional purposes. Most of the maps are developed based on interpretation of stereo-paired aerial photography, similar to the process used by the National Wetland Inventory. Representative states that have produced wetland maps include Maine, Vermont (based on NWI), Massachusetts (Wetlands Restriction Program), New York, and New Jersey. Coverage, interpretation keys (classification systems), and accuracy are variable from state to state.

ON-SITE WETLAND DELINEATION The ability to document wetland site conditions without detailed on-site investigations has a demonstrated need from a natural resources planning perspective as well as from a jurisdictional perspective. Documentation of anthropogenic influence on wetlands is another demonstrated need for using off-site materials for wetland identification. By using the resources and methods discussed above, a reviewer can generally make a positive or negative determination regarding the presence or absence of wetlands, estimate the areal extent of wetlands, and, in some cases, determine major cover types (although a follow-up site inspection is always recommended for full confirmation). However, off-site wetland identification is not a substitute for on-site wetland delineation when the goal is a definitive demarcation or delineation of wetlands for site development and project planning purposes. The need to identify jurisdictional wetlands and delineate wetland/upland boundaries in the United States is principally driven by Section 404 of the Clean Water Act by state and municipal wetland protection statutes. The wetland protection statutes, and associated regulatory policies, dictate that wetland boundaries be established and, in many cases, confirmed, prior to site development and related land management activities. Therefore, project proponents are required to characterize and quantify the differences between wetlands and uplands so that the boundary can be identified with some certainty and repeatability. This is accomplished by field assessment of vegetation, soils, and hydrology. Through interagency consensus, three characteristics or parameters have been selected to distinguish wetlands from uplands. First, wetlands are characterized by the presence of water, typically from a surface or groundwater source. Water levels in wetlands are typically dynamic, with the frequency of saturation and inundation varying among wetland types and varying temporally within wetland types. Second, wetlands are characterized by the presence of unique soils that are diagnostic of wetland conditions. These soils display properties that indicate anaerobic conditions in the root zone resulting from prolonged saturation or inundation. Finally, wetlands are characterized by the presence of wetland vegetation that possesses morphological

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adaptations that enable them to tolerate frequent root zone saturation or inundation and anaerobic conditions. Earlier in this chapter, the regulatory definition of wetlands was provided. More recently, the National Academy of Sciences (National Research Council, 1995) proposed another definition of wetlands. A wetland is an ecosystem that depends on constant or recurrent, shallow inundation or saturation at or near the surface of the substrate. The minimum essential characteristics of a wetland are recurrent, sustained inundation or saturation at or near the surface and the presence of physical, chemical and biological features reflective of recurrent, sustained inundation or saturation. Common diagnostic features of wetlands are hydric soils and hydrophytic vegetation. These features will be present except where specific physicochemical, biotic, or anthropogenic factors have removed them or prevented their development.

There are many factors that affect the presence of these features. Topographical relief and overall landscape position are significant physical factors that often dictate the source for wetland hydrology. For example, topographical depressions often correspond closely with water table elevation in glaciated wetlands of the northeastern United States. Headwaters of streams and rivers are often the result of sheet runoff from watersheds that emanate from mountainous regions. Palustrine and riverine wetlands are often associated with these surface water bodies. Other factors that influence the formation of the three wetland factors include stratigraphy, surficial and bedrock geology, and watershed and climatological conditions including precipitation and evapotranspiration. In identifying and delineating wetlands, it is important to establish in advance the overall goal and scope of the wetland investigation. The investigator should determine if it is necessary to conduct a comprehensive on-site delineation of the entire wetland and upland boundary or simply confirm the presence or absence of wetlands. A wetland determination is the process by which the evaluator makes a positive or negative assumption that wetlands are extant on a site. This assumption is based on identifying whether or not wetland characteristics are present anywhere within the site’s boundaries. Wetland delineation is the process by which the investigator identifies and locates wetlands, then qualitatively or quantitatively assesses the areal extent of wetlands on the site. This is accomplished through consideration of hydrological field indicators, soil profiles, and vegetation sampling and inventory. Wetland delineation techniques and methodologies vary from place to place in response to local and state jurisdictional requirements. Nevertheless, almost without exception, local and state mandated procedures are predicated upon parameters defined by the federal agencies. These factors, and their associated technical criteria, are expressed in the Corps of Engineers Wetland Delineation Manual (Environmental Laboratory, 1987), the Wetland Identification and Delineation Manuals, Volumes I and II (Sipple, 1988), the Federal Manual for Identifying and Delineating Jurisdictional Wetlands (Federal Interagency Committee for Wetland Delineation, 1989), the National Food Security Act Manual (U.S. Department of Agriculture, 1994), and the National Research Council (1995).

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As federal legislation has evolved, resultant modifications to the jurisdictional definitions and limits have occurred, and likely will continue to occur, with ensuing performance standard modifications (U.S. Army Corps of Engineers, 1991; U.S. Army Corps of Engineers, 1992). Other sources (Tiner, 1991; Sipple, 1992) discuss the importance of the dynamic nature (e.g., seasonality, degree of wetness) of wetlands as it relates to the individual factors and the ability to effectively recognize wetland boundaries. Tiner (1993) proposed an innovative approach to delineating wetlands based on identifying primary indicators of hydrophytes and hydric soils in undrained wetlands. Also in 1993, the National Academy of Sciences was tasked by Congress to conduct an independent assessment of current wetland delineation practices, with their findings published in 1995 (National Research Council). The National Research Council report emphasized regionalizing standard approaches and evaluation criteria. This section discusses the three factors associated with wetlands, with an emphasis placed on conducting field assessments that evaluate and characterize each factor. Techniques are presented which assist the investigator in assessing each factor, including recognizing field indicators of wetland hydrology, hydric soils, and hydrophytes. In conducting these analyses, it is important that the methodology be practical, reproducible, efficient, and cost effective. The following sections describe each factor in depth, and an evaluation procedure for assessing all these in conducting a wetland delineation. Wetland Hydrology Wetland hydrology is the single greatest impetus driving wetland formation (Mitsch and Gosselink, 1986; Federal Interagency Committee for Wetlands Delineation, 1989; Tiner and Veneman, 1989; Tiner, 1993), and has historically been the most controversial (U.S. Environmental Protection Agency, 1991; Environmental Law Institute, 1991). Wetland hydrology is characterized by permanent, temporary, periodic, seasonal, or tidally influenced inundation or soil saturation within the root zone. This water may derive from surface water (e.g., streams, rivers, ponds, lakes, ocean), groundwater, overbank flooding, precipitation, sheet flow, or tidal flooding. Wetland hydroperiod is a term used to characterize the hydrological condition of a wetland and is a function of flood duration and flood frequency. All wetlands are dynamic systems from a hydrological viewpoint. The hydrology of perennial wetlands varies irregularly on an annual basis. Ephemeral wetlands such as vernal pools typically have seasonally varying hydroperiods. Tidally influenced wetlands experience daily periodic hydrologic fluctuations. Keeping in mind this inherent variation, a wetland’s net hydroperiod can be represented by the following equation (Mitsch and Gosselink, 1986). V = Pn + Si + Gi – Et – So) – Go ± T

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where V equals the change in volume of water storage. Pn equals the net precipitation. Si equals the surface inflows (stream flow and sheet flow). Gi equals the groundwater inflows. Et equals the evapotranspiration. So equals the surface outflows. Go equals the groundwater outflow. T equals the tidal inflow (+) or outflow (–).

Wetland hydrology drives the development and distribution of the other two indicator parameters. Permanent or periodic inundation or saturation of the root zone during the growing season results in the development of anaerobic conditions, which is one of the chief determinants of hydric soil conditions. Root zone saturation is, in turn, responsible for the occurrence and distribution of vegetation that can withstand these conditions. There has been much debate regarding the duration of time required for anaerobic conditions to develop from root zone saturation or inundation. Until recently, the U.S. Army Corps of Engineers required a minimum of 2 weeks of continuous saturation or inundation within the root zone for an area to be considered to possess wetland hydrology (U.S. Army Corps of Engineers, 1991). Presently, the Corps stipulates that the root zone be saturated or inundated for more than 12.5 percent of the growing season for an area to have wetland hydrology (U.S. Army Corps of Engineers, 1992). Areas inundated or saturated for 5 to 12.5 percent of the growing season may or may not have wetland hydrology. The National Research Council, in their 1995 study, Wetlands: Characteristics on Boundaries, proposed that the threshold for duration of saturation can be approximated as 14 days during the growing season in most years. Historically, the definition of the root zone and its maximum vertical extent has been critical to determining jurisdictional limits. This vertical extent is a direct function of soil series drainage class and permeability. The Corps of Engineers Wetlands Delineation Manual defines major portions of the root zone to be that area within 30 cm of the soil surface (Environmental Laboratory, 1987). The ability to identify and delineate wetlands relative to the hydrology parameter relies on the investigator’s ability to recognize field indicators of wetland hydrology. Generally, quantitative studies and hydrological modeling are not required for onsite wetland identification and delineation. Although many of the technical aspects of the scientific discipline of hydrology are widely applicable to wetland formation, only a rudimentary knowledge of the underlying principles is crucial to wetland delineation. Proficiency in recognizing wetland hydrological indicators, however, is required. Hydrological Field Indicators On-site wetland inspection and delineation require the ability to recognize and distinguish wetland hydrological field indicators (Table 1). The indicators provide ©2001 CRC Press LLC

visual or assumed evidence of soil saturation or surface inundation. Certain indicators provide strong evidence of the frequency and duration of saturation or inundation and can be interpreted to support wetland boundary determinations. Field indicators include the direct observation of wetland hydrological conditions (during the growing season), soil characteristics, morphological plant adaptations, and evidence of water movement (Environmental Laboratory, 1987; Federal Interagency Committee for Wetland Delineation, 1989). Table 1

Wetland Hydrological Field Indicators Widely Used for Identifying Wetlands, and Delineating the Wetland and Upland Boundary

Direct observation of inundation Direct observation of soil saturation Wetland drainage patterns Plant morphological adaptations Adventitious roots Aerenchyma Hypertrophied lenticels Leaf adaptations Multiple trunks Oxidized rhizospheres Pneumatophores Shallow roots Stooling Tree buttressing Water marks and drift lines Surface scouring and water-borne sediment deposits Field confirmed hydric soils Water-stained leaves

The strongest field indicator of wetland hydrology is the direct observation of surface inundation or soil saturation within the root zone during the growing season (Figure 7). This indicator represents a real-time visual observation of wetland diagnostic conditions. Inundation is a valid indicator of wetland hydrology; however, the boundary or areal extent of a wetland may, indeed, extend far beyond the limit of the inundated area. Typically (especially in groundwater discharge wetlands), as topographic elevation increases, the areal extent of inundation decreases, with soil saturation within the root zone becoming the dominant defining characteristic. The depth to which soil saturation is present, and its persistence, determine whether the area is a jurisdictional wetland. Depth is also related to the hydric soil criterion, specifically its drainage class and permeability. Generally, the soil should be saturated within 0 to 46 cm of the surface during the growing season. Soil saturation can be field confirmed through an auger boring to a depth of 46 cm to determine the water table elevation. Saturated soils will occur below this elevation and slightly above due to capillary action (Environmental Laboratory, 1987). Careful attention needs to be paid to extremely fine textured soils (silty clays and clays) which can have a very high capillary fringe and not be good indicators of the actual water table elevation. Soils can also be squeeze tested to extract free water to determine saturation. Wetland drainage patterns occur along riverine, estuarine, palustrine, and some lacustrine wetlands. They are typically associated with moving or fluctuating water systems such as streams, rivers, creeks, etc. Visual indicators include drainage channels, eroded areas, the absence of litter, litter deposits, and characteristic meandering. ©2001 CRC Press LLC

Figure 7

The strongest field indicator of wetland hydrology is the direct observation of surface inundation or soil saturation.

Many wetland plants have developed specialized morphological adaptations that enable them to survive and proliferate with their roots in an anoxic environment. These adaptations have developed in response to root zone saturation and are, therefore, treated as wetland hydrology indicators (Federal Interagency Committee for Wetlands Delineation, 1989). The adaptations typically are specialized structures enabling the plants to capture molecular oxygen and transport it to stems and roots. This is true of buttressed trees (swollen base of the trunk), pneumatophores (above ground root structures), adventitious roots (above ground roots), shallow roots, multiple trunks and stooling, hypertrophied lenticels (exaggerated lenticels on stems), aerenchyma (air-filled tissue), and leaf adaptations (floating and polymorphic leaves). An oxidized rhizosphere is also indicative of wetland hydrology. Evidenced by channels that have developed along the roots for the transport of oxygen, rhizopheres are difficult to observe and clearly identify unless iron oxide concretions are present. Watermarks are visible indications of inundation on woody vegetation and other permanent structures within wetlands. These are considered strong indicators of the presence of wetland hydrology owing to their confirmation of inundated conditions. Drift lines are visible lines indicating the extent of hydrologically driven actions. Tidal marshes typically display drift lines with debris and driftwood extending up to the spring high tide elevation. Drift lines also accumulate along riverbanks and floodplains. Surface scouring occurs in wetlands with widely fluctuating hydroperiods. Scouring also occurs in areas subject to storm or tidal forces that expose soils, remove leaf litter, and cause surficial erosion. Sediment deposits on vegetation also indicate that inundated conditions have occurred in the recent past. ©2001 CRC Press LLC

Observations of wetland hydrology are typically performed at the time of the on-site inspection. However, in cases where long-term observation may be necessary to confirm precise wetland boundaries, wetland hydrology should be assessed for longer periods such as entire growing seasons. Long-term hydrological monitoring can be a labor intensive and costly process unless historical groundwater and surface water elevation data are available. In situations where the jurisdictional limits of a wetland have been debated or questioned and precise boundaries are required (e.g., real estate transactions), hydrological monitoring can typically be accomplished in a cost-effective manner through the installation and monitoring of hand-set groundwater monitoring wells. Typically, 5 cm diameter slotted screen PVC monitoring wells, 60 to 75 cm long, are installed at random locations along a transect line extending from within the recognizable wetland area out to recognizable upland. Hand augers are sufficient to install wells, and borings should be backfilled with washed sand to allow unrestricted passage of groundwater. Well locations and elevations should be surveyed and plotted on a site plan, and monitoring of groundwater elevations should be conducted weekly throughout the growing season. Water level recording can be accomplished using a sounder mechanism and an incremented cord or tape rule. Precipitation should be recorded throughout the monitoring period. Regulatory agencies may require several growing seasons’ worth of data, which can be impractical from the standpoint of project timing and cost. Nevertheless, long-term hydrological information represents the strongest evidence for the extent of wetlands. Hydric Soils The soils found in wetlands have unique morphological and other observable properties that differentiate them from upland soils. A hydric soil by definition is a soil which is saturated, flooded, or ponded long enough during the growing season to develop anaerobic conditions in the upper part (U.S. Department of Agriculture Natural Resources Conservation Service, 1991; New England Hydric Soils Technical Committee, 1998). Hydric soil properties are a direct function of the frequency and duration of saturation and inundation, specifically in the root zone. Soils that display flooded and saturated conditions for an extended period (2 weeks or more) during the growing season create an environment where free oxygen is deficient and, ultimately, unavailable to plants. As a result of this saturation and inundation, hydric soils display observable field indicators that are diagnostic of wetland conditions. The U.S. Department of Agriculture Natural Resources Conservation Service has developed a classification system that provides criteria for listing a hydric soil as well as categories of the listed hydric soils. This list, Hydric Soils of the United States, categorizes hydric soils into two major groups: organic soils and mineral soils (U.S. Department of Agriculture Natural Resources Conservation Service, 1991). Generally, soils with at least 46 cm of organic matter in the upper part of the soil profile are considered organic soils, or histosols (Tiner and Veneman, 1989). Organic soils are divided into groups based on the degree to which plant fibers and material are decomposed. Fibrists (peats), hemists (mucky peats and peaty mucks), and saprists (muck) are organic hydric soils listed in increasing order of plant ©2001 CRC Press LLC

material decomposition. Folists are the fourth group of organic soils, but they are not considered hydric soils because the organic component does not derive from long-term saturation or inundation (Tiner and Veneman, 1989). Mineral soils generally have less organic material in the upper part of the profile than organic soils and have different field indicators. Mineral hydric soils are also taxonomically arranged and include soils in Aquic suborders, Aquic subgroups, Albolis suborder, Salorthids great groups, and Pell great groups of Vertisols (shrinking or swelling dark clay soils, Federal Interagency Committee for Wetlands Delineation, 1989; U.S. Department of Agriculture Natural Resources Conservation Service, 1991). Mineral soils are considered hydric soils when any of several criteria are satisfied (Tiner and Veneman, 1989). Somewhat poorly drained soils with a water table less than 15 cm from the surface for a significant period during the growing season are hydric. Poorly drained or very poorly drained soils are hydric if the water table is at or less than 30 cm from the surface for a significant period during the growing season if permeability is equal to or greater than 15 cm/hr in all layers within the top 50 cm, or the water table is less than 46 cm from the surface for a significant period during the growing season if permeability is less than 15 cm/hr in any layer within the top 50 cm. Mineral soils are also hydric if water is ponded for a long duration (more than 7 days) or a very long duration (greater than 1 month) during the growing season. Mineral soils frequently flooded for a long duration (more than 7 days) or a very long duration (more than 1 month) during the growing season are also considered hydric. A significant period is defined as at least 15 consecutive days of saturation or 7 days of inundation during the growing season (U.S. Army Corps of Engineers, 1992). The 1987 Corps Manual, which is the current manual guiding federal jurisdictional technical delineation, defines growing season as that portion of the year when soil temperatures at 50 cm below the soil surface are higher than biologic zero (5˚C). For ease in determination, the growing period can be estimated to occur when air temperature exceeds 22˚C (U.S. Army Corps of Engineers, 1992). Drainage classes are a significant criterion when determining the presence of hydric soils, as the soils relate to individual taxonomic groups (New England Hydric Soils Technical Committee, 1998; Smith, 1973; Tiner and Veneman, 1989). In addition, field determination of drainage classes has been made easier through the use of manuals such as the Field Indicator for Identifying Hydric Soils in New England (New England Hydric Soils Technical Committee, 1998). All very poorly and poorly drained soils are hydric soils, assuming the soils have not been drained. A very poorly drained soil is a soil where water is removed from the soil so slowly that free water remains at or near the surface during most of the growing season. A poorly drained soil is a soil where water is removed so slowly that the soil is saturated periodically during the growing season or remains wet for long periods. Many somewhat poorly drained soils are also hydric. A somewhat poorly drained soil is one where water is removed slowly enough that the soil is wet for significant periods during the growing season. Soil mapping provided by the U.S. Department of Agriculture Natural Resources Conservation Service indicates the soil series drainage class which can be field confirmed using criteria developed by the U.S. Army Corps of Engineers (1991). ©2001 CRC Press LLC

Essential to the investigation of soils for hydric properties is an understanding of soil horizons. Soil horizons, or layers, develop in response to localized chemical and physical processes resulting from the activities of soil organisms, the addition of organic matter, precipitation, and percolation. Horizons can be distinguished based upon color, texture, and composition (Environmental Laboratory, 1987). However, the soil horizon is essentially a continuum and there is no clear distinction between one horizon and another. Soils typically have four major horizons: an organic layer (O) and three mineral layers (A, B, and C); see Figure 8. The O horizon is the surface layer and is composed of fresh or partially decomposed organic material. The A horizon is characterized by an accumulation of organic matter and the loss of clay, iron, and aluminum. Together, the O and A horizons constitute the zone of maximum biological activity. The B horizon is characterized by the accumulation of silicates, clay, iron, aluminum, and humus, whereas the C horizon contains weathered material either similar or dissimilar to the parent material. Soil colors also provide critical information on soil wetness, and the degree of saturation and inundation. The Munsell Color Chart standardizes three aspects of color: hue, value, and chroma (Kollmorgen Corporation, 1975). Hue describes the soil based on its relation to the spectral colors (i.e., red, yellow, green, blue, purple, or a mixture of these colors), value describes the degree of lightness, and the chroma indicates the strength or purity of the color. The Munsell Color Chart depicts individual hues on separate pages, with the corresponding value on the vertical axis and the chroma on the horizontal axis. Soil color is reported as hue, value/chroma — for example, 7.5 YR 2/1. A relatively new standard soil color is the earth tones book that has an increased number of soil color chips and is being used more commonly for hydric soil determinations. Soil colors are an important field indicator for field verification of mapped hydric soils as well as hydric soil inclusions in mapped, nonhydric soils. Both soil matrix, the predominant color derived from the parent material or deposits, and redoximorphic features and concretions, contrasting spots within the matrix, should be characterized. Soil colors are diagnostic of wetland conditions when the matrix chroma immediately below the A horizon is two or less in redoximorphic features (e.g., mottled soils), or one or less in soils without redoximorphic features (Environmental Laboratory, 1987). In some soils, though, it is now recognized that carbon soils with a chroma of three and the presence of redoximorphic features indicate a hydric soil. Hydric Soil Field Indicators Hydric soils display certain field indicators that provide clues to the degree of saturation and inundation. These clues are, in turn, indicative of the presence or absence of wetlands hydrology (e.g., indicate that the root zone has experienced saturation or inundation). The field indicators presented in Table 2 are nationally accepted and were developed based on U.S. Department of Agriculture Natural Resource Conservation Service factors (Environmental Laboratory, 1987; Federal Interagency Committee for Wetland Delineation, 1989; Tiner and Veneman, 1989). Since that time, there have been several attempts to develop regional hydric soil ©2001 CRC Press LLC

Figure 8

A generalized soil profile. Actual horizons represent a continuum and will not appear this distinct in the field.

field indicators (New England Hydric Soils Technical Committee, 1998). Recognizing and applying these field indicators will enable the wetland scientist to field confirm or refute the hydric soil criterion. Most of these field indicators are used when attempting to recognize hydric mineral soils, and some, such as a high organic content at the surface, subsurface streaking, and wet spodosols, are particularly useful when attempting to recognize sandy hydric mineral soils. As mentioned previously, all organic soils except folists are considered hydric soils. Organic soils are readily identifiable in the field due to the relatively high percentage of poorly decomposed organic matter such as vegetative material and litter in the soil matrix. Organic soils also often possess dark hues and very dark chromas (7.5 YR 2/0, 10 YR 2/1, etc.).

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Table 2

Field Indicators of Hydric Soils

Organic soils (histosols) Histic epipedons Sulfidic material Aquic or peraquic moisture regime Reducing condition Soil color Mottling or concretions High organic matter content in surface horizon Subsurface streaking Spodic horizon Note: Indicators are applied to the soil profile within the root zone.

Histic epipedons are thick (20 to 40 cm) organic surface layers overlying hydric mineral soils. They develop as a result of prolonged saturation within the root zone, at least 30 consecutive days or more in most years (Environmental Laboratory, 1987). The presence of sulfidic material is a strong indicator of permanently saturated soil conditions. Sulfidic material is evidenced by a rotten egg-like odor, typically following soil disturbance. The odor occurs as a result of sulfates being reduced to sulfides under anaerobic conditions associated with prolonged root zone saturation. This indicator is especially common in tidal marshes. Elevated groundwater levels characterize aquic and peraquic moisture regimes. Aquic moisture regimes are mostly reducing, and are nearly free of dissolved oxygen as a result of saturation due to groundwater or the capillary fringe. A part or all of the root zone is reduced and totally deficient in dissolved oxygen during the growing season. Peraquic moisture regimes occur when groundwater is always at or near the soil surface. Reducing conditions can be observed in hydric mineral soils of high iron content using a calorimetric test kit. In a reducing environment, ferrous (reduced) ions are detected using a-adipyridil (Environmental Laboratory, 1987). It should be understood that this test only reveals reducing conditions during the time of the field investigation and may not reflect conditions typical of a significant part of the growing season. Repeated sampling over an extended temporal scale is required to satisfy the hydric soil criterion. Soil colors typically provide the most useful and diagnostic field indicators for hydric mineral soils. When water is at or near the soil surface for most or all of the year, soils may become gleyed and take on a lighter dull gray or bluish color. The color arises from the conversion of oxidized (ferric) iron to its reduced (ferrous) state, and the subsequent removal of the latter from the soil. Oxidizing conditions are apparent only along root channels in gleyed soils. The formation of redoximorphic features (e.g., mottles, concretions, and depletions) occurs in soils where the water table fluctuates periodically. Saturated, anaerobic conditions cause the solubilization of iron compounds. Subsequent depressions in the water table create an aerobic environment and iron ions are converted to insoluble oxides that appear as orange or red mottles in the soil matrix. Relatively longer periods of soil saturation and anaerobic conditions followed by aerobic conditions convert manganese ions to ©2001 CRC Press LLC

oxides. The manganese oxides are deposited as concretions and are evidenced by small, dark brown or black nodules interspersed with the soil matrix. Both mottles and concretions persist for long periods after formation and may not reflect current conditions. However, a combination of redoximorphic features and a low matrix chroma is generally indicative of hydric conditions. Accumulations of iron and manganese are now generally referred to as redox concentrations (New England Hydric Soils Technical Committee, 1998). Redox depletions are areas within the soil with low chromas without iron and manganese oxides. Soil color is a poor indicator of hydric condition in sandy soils and other indicators must be used. A high organic matter content in the surface horizon is indicative of hydric conditions because it is caused by prolonged inundation or saturation. This prolonged inundation or saturation greatly reduces oxidation of organic matter, leading to its accumulation. Subsequent lowering of the water table may cause streaking of the subsurface horizons as organic matter is moved downward through the sand. Eventually, this downward moving organic matter may accumulate at a point coinciding with the most commonly occurring depth to groundwater. The organic matter may become cemented with aluminum, forming a less permeable, hardened, spodic horizon. Each of these characteristics, high organic matter content in the surface horizon, streaking, and a spodic horizon, is indicative of hydric conditions. When examining the soil component during wetland identification and delineation, minimum equipment needs include a hand auger, Munsell Color Chart, delineation data sheets, pencil, and camera. Borings need to be hand augured at locations along the perceived upland and wetland sides of the boundary, keeping hydrological and topographical indicators in mind. Borings should extend to depths between 30 to 46 cm to adequately characterize O, A, and B horizons. Diagnostic observations include depth to groundwater or depth of inundation, and examination for each of the field indicators for mineral and organic hydric soils. The hydric soil boundary should be identified based upon the wetland side having observable characteristics of hydric soils and the upland side having a general lack of these characteristics. Visible hydrology will influence the number of borings needed to adequately define the soil boundary. Hydrophytic Vegetation Probably the most visible and easily recognizable diagnostic feature of wetlands is hydrophytic (water-loving) vegetation. Hydrophytic vegetation is the sum total of macrophytic plant life that occurs in areas where the frequency and duration of inundation or soil saturation are of sufficient duration to exert a controlling influence on the plant species present (Environmental Laboratory, 1987). Hydrophytes possess anatomical and physiological adaptations that allow them to survive and thrive in saturated or inundated soils, where oxygen depletion is the primary factor limiting vegetation occurrence (Figure 9). Adaptive structures of hydrophytes include aerenchyma, adventitious roots, stooling, hypertrophied lenticels, and buttressing. Plant species with a demonstrated ability to achieve maturity and reproduce where the root zone is inundated or saturated during the growing season are listed in the National List of Plant Species that Occur in Wetlands (Reed, 1988, 1995). ©2001 CRC Press LLC

Figure 9

Hydrophytes, such as this skunk cabbage (Symplocarpus foetidus ) possess anatomical and physiological adaptations that allow them to survive in saturated soils.

Approximately 7000 species of hydrophytes are listed. Originally designed as an appendix to Classification of Wetlands and Deepwater Habitats of the United States (Cowardin et al., 1979), the list was modified to assist in determining the probability that a vegetation community is a wetland (Wentworth and Johnson, 1986). This is accomplished by assigning each plant to an indicator category which reflects the probability, expressed as a frequency of occurrence, of a species occurring in wetland or upland (Table 3). Modifiers (+ or –) are also assigned to the categories to further refine individual species affinity toward wet conditions. A ‘+’ indicates an increased probability of wetland tolerance, and a ‘–’ indicates an affinity for drier conditions (Reed, 1988). The categories should not be equated to degrees of soil wetness. For example, many obligate wetland species occur in permanently flooded wetlands, whereas other obligate plants occur in seasonally or temporarily flooded wetlands. Indicators of Hydrophytic Vegetation Dominance by hydrophytes is the most reliable indicator for determining whether hydrophytic vegetation is present. It has its basis in the current federal definition of wetlands which includes the phrase a prevalence of vegetation typically adapted for life in saturated soil conditions. Vegetation is said to be prevalent when it is dominant, that is, the species contribute more to the character of a plant community than other species (Environmental Laboratory, 1987). If the dominant species are hydrophytes, the community consists of hydrophytic vegetation. The standard for determining if the vegetation is hydrophytic is the 50 percent rule—more than 50 percent of the dominant vegetation is OBL, FACW, or FAC (excluding FAC–) on the appropriate

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Table 3

Indicator Category and Frequency of Occurrence in Wetlands for Plants Listed (in Reed 1988) Frequency of Occurrence in Wetlands > 99 percent

Indicator Category Obligate Wetland (OBL) Facultative Wetland (FACW) Facultative (FAC)

67–99 percent 34–66 percent

Facultative Upland (FACU) Obligate Upland (UPL)

1–33 percent < 1 percent

Representative Species Cattail (Typha spp.) Button bush (Cephalanthus occidentalis ) Soft rush (Juncus effusus ) Speckled alder (Alnus rugosa ) Red maple (Acer rubrum ) Sweet pepperbush (Clethra alnifolia ) American beech (Fagus grandifolia ) White pine (Pinus strobus ) Black oak (Quercus veltina ) Staghorn sumac (Rhus lyphina )

Note: Upland species are not listed.

list of plant species that occur in wetlands (Environmental Laboratory, 1987; U.S. Army Corps of Engineers, 1991; U.S. Army Corps of Engineers, 1992). Various ecologically based methods are acceptable for determining dominant plant species. Hays et al. (1981) provides operational descriptions of several techniques suitable for quantitatively measuring habitat variables. The Corps of Engineers recognizes various techniques that consider individual vegetation strata (Table 4). Consistent with the Corps of Engineers (Environmental Laboratory, 1987; U.S. Army Corps of Engineers, 1992), trees and lianas within a 9.1 m (30 ft) radius of a sample point are identified and measured for basal area at breast height (1.4 m or 4.5 ft). The basal area of all individual tree species and all individual liana species are summed and tree species and liana species separately ranked in descending order based upon total basal area. As an alternative to measuring the basal area of lianas, the number of individual stems of each species may be counted. As many lianas branch, stem counts should be conducted at ground level. Lianas are then ranked in descending order of dominance based upon number of stems. Table 4

Categories Used in the Assessment of Vegetation Communities

Story Woody overstory Woody understory

Stratum Tree Liana Sapling Shrub

Herbaceous understory

Seedlings and herbs Mosses and liverworts

Definition Woody, nonclimbing, at least 12.7 cm dbh and at least 6.1 m tall Woody vines, climbing Woody, nonclimbing, at least 1 cm dbh but less than 12.7 cm dbh and at least 6.1 m tall Woody, nonclimbing, at least 0.9 m tall but less than 6.1 m tall Woody, less than 0.9 m tall, or nonwoody and any height Small, green, nonflowering

Other strata are typically evaluated based upon percent areal coverage. According to the Corps of Engineers (Environmental Laboratory, 1987; U.S. Army Corps of Engineers, 1992), saplings and shrubs are assessed within a 4.5 m (15 ft) radius, ©2001 CRC Press LLC

and seedlings and herbs are assessed within a 1.5 m (5 ft) radius, of a sample point. For each stratum, the species are ranked in descending order of dominance based upon percent cover. Mosses and liverworts are only considered when they constitute an important component of the vegetation community. In determining whether hydrophytic vegetation is present, the Corps of Engineers (Environmental Laboratory, 1987; U.S. Army Corps of Engineers, 1992) suggests using the three dominant species from each vegetation stratum and five species if only one or two strata are present. The indicator status of each species is recorded. If the majority of dominant species are OBL, FACW, or FAC (excluding FAC–), then the vegetation is said to be hydrophytic and, therefore, indicative of wetlands. A variation of the dominance method requires calculation of a prevalence index (Federal Interagency Committee for Wetland Delineation, 1989). Each indicator category is assigned an index value: OBL equal to 1.0, FACW equal to 2.0, FAC equal to 3.0, FACU equal to 4.0, and UPL equal to 5.0. Dominant vegetation is identified based upon its relative frequency of occurrence or relative areal coverage. An index value is assigned to each species and weighted by its relative dominance in the community. If the sum of the index value is less than 3.0, then the vegetation is considered to be hydrophytic. Although dominance by hydrophytes is the most reliable indicator of hydrophytic vegetation, other indicators exist. In general, these other indicators should be applied only after application of the dominant species method. Nevertheless, visual observation of plant species growing in areas of prolonged inundation or soil saturation, particularly if those species have been observed in other wetland areas, suggests hydrophytic vegetation. This approach may be applied with some reliability for OBL, and to a lesser extent FACW species, but it is considerably less reliable for FAC species. The presence of standing water or saturated soil is, in many cases, insufficient evidence that the observed species are capable of tolerating anaerobic conditions for an extended period. When in doubt, determine hydrology and soil characteristics. Morphological adaptations to inundation or soil saturation such as buttressed tree trunks, adventitious roots, and shallow root systems are also suggestive of hydrophytic vegetation (see Table 1). Most individuals of the dominant species should exhibit morphological adaptations for this indicator to be reliable. However, not all hydrophytic species have obvious morphological adaptations. Conversely, apparent morphological adaptations may develop in response to factors other than soil wetness.

IDENTIFYING AND DELINEATING WETLANDS Undisturbed Areas When a site is relatively undisturbed, identifying and delineating wetlands are accomplished by simultaneously applying the criteria for wetland hydrology, hydric soils, and hydrophytic vegetation. Typically, the best time to identify or delineate wetlands is during the growing season, when dominant vegetation (especially ©2001 CRC Press LLC

annuals) are evident and hydrology can be directly observed. Nevertheless, identification and delineation of wetlands can be reasonably accomplished at any time of the year, although the process is made infinitely more difficult by snow or ice cover. To facilitate wetland identification and delineation, it is helpful to establish a working baseline. The baseline should extend parallel to any major watercourses or waterbodies or to any potential wetland areas (Figure 10). It is also helpful to establish the baseline perpendicular to the topographical gradient. Transects are established perpendicular to the baseline, with the number of transects depending upon the length of the baseline and the diversity of plant communities on the site. Minimally, at least one transect should sample each plant community type. The Corps of Engineers (Environmental Laboratory, 1987) offers guidelines for establishing transects. A number of observation points should be established along each transect. Again, the Corps of Engineers (Environmental Laboratory, 1987) offers guidelines for establishing observation points. An experienced investigator will minimize effort by establishing one observation point in recognizable wetlands and a second point in recognizable upland. Using observed hydrological, soil, and vegetation characteristics at these two observation points as references, the investigator can then focus his or her efforts on areas that are less readily discernible.

Figure 10

Establishment of a baseline parallel to observed watercourses and perpendicular to the topographical gradient facilitates wetland identification and delineation. Transects are established perpendicular to the baseline, and at least one transect should sample each plant community type.

At each observation point, hydrology, soil, and vegetation are simultaneously characterized and a determination is made as to whether the area constitutes wetlands or uplands. More precisely, the observation point is inspected for indicators of wetland hydrology, indicators of hydric soil, and hydrophytic vegetation. Under normal conditions, indicators of wetland hydrology, hydric soil, and hydrophytic vegetation must all be present for an area to be considered a wetland. A determination ©2001 CRC Press LLC

that the observation point constitutes a wetland is sufficient if the purpose of the exercise is only identification of wetlands. For the purpose of delineation, subsequent observation points along the transect, upgradient of the wetland observation point, must be characterized until wetland indicators are absent. Conversely, initial identification of an upland observation point requires subsequent investigation of downgradient observation points until all wetland indicators are present. Disturbed Areas When a site is disturbed, identifying and delineating wetlands through application of the criteria for wetland hydrology, hydric soils, and hydrophytic vegetation may be impossible. The disturbance may have been intentional or inadvertent, the result of unauthorized activities or natural events. Site hydrology may be altered due to dam construction, water diversion, channelization, or groundwater withdrawal, resulting in a site that is wetter or drier than normal. Soil may be buried, removed, or mixed, whereas vegetation may be cut, burned, or converted to agricultural crops. In many cases, normal application of the methods discussed above will lead to the conclusion that the site is not a wetland because one or more of the wetland indicators are likely to be absent. If one or more wetland indicators cannot be assessed and the site clearly has been disturbed, it is appropriate to evaluate the remaining indicators and to identify other sources of information regarding the missing indicator(s). In general, as many sources as possible should be considered prior to any determination. There are several avenues of investigation for determining normal site hydrology. Recent aerial photographs of the site, taken during the growing season, may provide relatively conclusive information about site inundation and somewhat less conclusive information about soil saturation. In some instances, a stream or tidal gauging station may be located in close proximity to the site, allowing calculation of high and low water elevations. Certain field indicators of wetland hydrology, such as plant morphological adaptations or high water marks, may be visible. Less reliable but of some usefulness are historic records, floodplain management maps, and the personal knowledge of local officials or residents. If fill has been placed on the site, buried soils can reasonably be examined for hydric soil indicators. Conversely, the removal of surface soil layers may allow for examination of the exposed subsurface horizons. Plowed soils can be examined immediately beneath the plow zone. Although inconclusive, review of soil survey maps, or examination of adjacent, undisturbed soils of the same series may be useful. Information relative to the preexisting vegetation community varies in its reliability. The most reliable information comes from examination of vegetation communities only partially removed or from review of recent documentation of site vegetation. Somewhat less reliable is examination of recent aerial photography and examination of vegetation in adjacent, undisturbed areas. Least reliable are review of soil survey plant community descriptions, review of National Wetland Inventory maps, and the recollection of local officials or residents.

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Difficult Areas Some wetlands are inherently more difficult to identify and delineate than other wetlands. The difficulty arises because one or more of the wetland indicators may be impermanent. Slope wetlands in glaciated wetlands have thin soils over relatively impermeable till or varying hydraulic conditions that produce groundwater seepage. Hydric soils and hydrophytic vegetation are generally readily apparent, but wetland hydrology is present only during wetter parts of the growing season. Seasonal wetlands such as vernal pools also lack wetland hydrology during all but the wettest part of the growing season. Prairie potholes have standing water for the majority of the growing season in normal or wet years, but they may be dry in years of below average precipitation. Vegetated flats are dominated by annual plant species during the growing season, but they are devoid of vegetation outside of the growing season. Almost without exception, identification and delineation for each of these wetland types, as well as other difficult wetlands, can reasonably be accomplished by conducting the evaluation at the appropriate time of the year or in a year of normal precipitation. Aids to Delineation In the last several years, many tools have been developed or improved which have facilitated the technical procedures associated with wetland delineation. Differential Global Positioning System (GPS) software has widely expanded reference or landscape point location capability, including wetland boundaries. Use of GPS technology as it applies to wetland jurisdictional delineations has largely involved the ability to electronically locate wetland boundaries and data points (flags or transect locations) and place these data points on a site plan or other underlay in an efficient and reproducible manner. Use of GPS for locating wetland boundaries allows an operator to collect significant line and point features, and attribute data for objects or points in the natural landscape, in a relatively short period of time. GPS data collection is often undertaken using a backpack unit consisting of handset data logger and an antenna with differential beacon receiver allowing the operator to cover large areas and areas with difficult access to record the jurisdictional wetland boundary. GPS raw data typically needs to be downloaded onto computers for storage and differential correction but can be readily integrated with Geographic Information Systems (GIS) software to create site plans or base maps. The most common limitation of GPS for wetland assessment is human operational error. Inexperienced operators should understand basic settings and how to maximize the GPS’s efficiency. In order to maximize the effective use of GPS, a digital plan should be utilized with specific horizontal and vertical control. Other limitations include signal obstruction from the satellite constellation that impairs the ability of the receiver to collect a position. This can occur under heavy forest cover, against buildings, or when the sky is effectively blocked. However, satellite positions change in relation to the operator’s location on the earth so returning to a position at a later time may solve the problem.

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Technical resources and wetland delineation software also facilitate wetland delineation. A number of federal, state, and private vendor internet sites have made it much easier to acquire wetland delineation technical information. This information includes access to the various wetland delineation manuals, various technical publications, maps and soil surveys, and delineation software (Table 5). Table 5

Internet Sites with Wetland Delineation Technical Information Site

Information

www.fws.gov/fgdcwet.html www.nwi.fws.gov/qanda.html www.statlab.iastate/edu/soils/hydric www.nwi.fws.gov www.wssinc.com www.libweb.wes.army.mil

U.S. Fish and Wildlife Service publications Wetland classification Hydric soil list National Wetland Inventory maps, plant list WetForm software U.S. Corps of Engineers Waterways Experiment Station publications U.S. Department of Agriculture publications

www.itis.usda.gov/itis/index.html

Wetland Studies and Solutions, Inc. have developed specialized wetland delineation software, WetForm. A windows-based application, WetForm provides to field delineators a U.S. Army Corps of Engineers-accepted data form consistent with the data collection needs of the 1987 Corps Manual. Electronic data entry fields for observation plot, vegetation, soils, hydrology, and wetland determination are included and also are consistent with the Corps of Engineer Policy guidance issued subsequent to the 1987 Manual. The software package includes a vegetation database and all numerical calculations (e.g., vegetation dominance, percent cover, dominance ratio) are automatically calculated. Vegetation data fields include common and scientific names of wetland species from the applicable USFWS regions in the continental United States. The soils database is not comprehensive, but it does have inputs for matrix and redoximorphic Munsell color, texture, and structure. Soils series from the U.S. Deparment of Agriculture Soil Conservation Soil Surveys are also in the database. Hydrological indicator fields are essentially identical to the Corps of Engineer forms, but related keys are triggered when answers to consistent indicators are given.

REFERENCES 33 CFR 320–330, 1986 Code of Federal Regulations, U.S. Army Corps of Engineers, 1986. Cowardin, L. M., Carter, V., Golet, F. C., and LaRoe, E. T., Classification of Wetlands and Deepwater Habitats of the United States, U.S. Department of the Interior, Fish and Wildlife Service Biological Services Program FWS/OBS-79131, 1979. Environmental Laboratory, Corps of Engineers Wetland Delineation Manual, Technical Report Y-87–1, U.S. Army Engineer Waterways Experiment Station, Vicksburg, MS, 1987. Environmental Law Institute, What is a jurisdictional wetland? Nat. Wet. Newsl., September/October, No. 5, 1991. Federal Geographic Data Committee (Wetlands Subcommittee), Proposal to Establish the Cowardin System as the Federal Wetland Classification Standard, 1995. ©2001 CRC Press LLC

Federal Interagency Committee for Wetland Delineation, Federal Manual for Identifying and Delineating Jurisdictional Wetlands, U.S. Army Corps of Engineers, U.S. Environmental Protection Agency, U.S. Fish and Wildlife Service, and U.S. Department of Agriculture Natural Resources Conservation Service, Washington, D.C., 1989. Hays, R. L., Summers, C., and Seitz, W., Estimating Wildlife Habitat Variables, U.S. Department of the Interior Fish and Wildlife Service, FWS/OBS-81/47, 1981. Hudson, W. D. and Lusch, D. P., Airphoto/satellite imagery—an introduction, in Proceedings of the Remote Sensing and GIS Applications to Nonpoint Source Planning Conference, sponsored by the U.S. Environmental Protection Agency and the Northeast Illinois Planning Commission, Chicago, IL, 1990. Kollmorgen Corporation, Munsell Soil Color Charts, Baltimore, MD, 1975. MacConnell, W., Goodwin, D., Jones, K., Stone, J., Foulis, D. B., Springston, G., and Swartout, D., Wetland Restriction Program: Mapping Standards for Wetlands Restriction Maps in Massachusetts, Technical Memorandum published by the Massachusetts Department of Environmental Protection, 1989. McKnight, T., Physical Geography: A Landscape Appreciation, 2nd ed., Prentice-Hall, Englewood Cliffs, NJ, 1987. Mitsch, W. J. and Gosselink, J. G., Wetlands, Van Nostrand Reinhold, New York, 1986. National Research Council, Wetlands: Characteristics and Boundaries, National Academy Press, 1995. New England Hydric Soils Technical Committee, Field Indicators for Identifying Hydric Soils in New England, 2nd ed., New England Interstate Water Pollution Control Commission, Wilmington, MA, 1998. Reed, P. B., Jr., National List of Plant Species that Occur in Wetlands: National Summary U.S. Fish and Wildlife Service, Biological Report 88(24), Washington, D.C., 1988. Reed, P. B., Jr., National List of Plant Species that Occur in Wetlands: National Summary, U.S. Fish and Wildlife Service, Biological Report, Washington, D.C., 1995. Ritchie, W., Wood, M., and Wright, R., Surveying and mapping for field scientists, in Aerial Surveying Techniques, Longman Scientific and Technical, 1988. Roughgarden, J. S., Running, W., and Matson, P. A., What does remote sensing do for ecology? Ecology, 72(6), 1918, 1991. Sipple, W. S., Wetland Identification and Delineation Manual, Vol. 1, Rational, Wetland Parameters, and Overview of Jurisdictional Approach, U.S. Environmental Protection Agency, Office of Wetlands Protection, Washington, D.C., 1988. Sipple, W. S., Wetland Identification and Delineation Manual, Vol. 2, Field Methodology, U.S. Environmental Protection Agency, Office of Wetlands Protection, Washington, D.C., 1988. Sipple, W. S., Time to move on, Nat. Wetl. Newsl., 14, 4, 1992. Smith, G. D., Soil moisture regimes and their uses in soil taxonomies, in Field Soil Water Regime, Bruce, R. R., Flach, K. W., and Taylor, H. M., Eds., Soil Scientist Society of America, Madison, WI, 1973. Smith, J., Jr. and Anson, T. A., Eds., Manual of Color Aerial Photography, 1st ed., American Society of Photogrammetry, VA, 1968. Tiner, R. W., Wetland delineation, in Proceedings of the 1991 Stormwater Management/Wetlands/Floodplain Symposium, Aron, G. and White, E. L., Eds., Pennsylvania State University, Department of Civil Engineering, University Park, PA, 1991. Tiner, R. W., The primary indicators method—a practical approach to wetland recognition and delineation in the United States, Wetlands, 13(l), 50, 1993.

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Tiner, R. W. and Wilen, B. O., The U.S. Fish and Wildlife Service’s National Wetland Inventory Project, Technical Memorandum, USFWS, Washington D.C. and Newton Corner, MA, 1983. Tiner, R. W. and Veneman, P. L., Hydric Soils of New England, University of Massachusetts Cooperative Extension, Revised Bulletin C-183-R, Amherst, MA, 1989. U.S. Army Corps of Engineers, Questions and Answers on the 1987 Corps of Engineers Manual, CECW-OR 7, October 1991. U.S. Army Corps of Engineers, Technical Memorandum: Clarification and Interpretation of the 1987 Manual, CECW-OR, USCOE/Washington, D.C., March 1992. U.S. Department of Agriculture Natural Resources Conservation Service, Soil Survey of Kennebec County, Maine, USDAISCS Cooperative Publication with the Maine Agricultural Experiment Station, 1978. U.S. Department of Agriculture Natural Resources Conservation Service, Hydric Soils of the United States, National Technical Committee for Hydric Soils, Publication No. 1491, Washington, D.C., 1991. U.S. Department of Agriculture Natural Resources Conservation Service, National Food Security Act Manual, 3rd ed., 180-V-NFSAM, 1994. U.S. Environmental Protection Agency, Proposed revisions to the Federal Manual for Delineating Wetlands, Fed. Regist., 58, 157, August 14, 1991. U.S. Geological Survey, Surficial Geologic Map, Hanover, MA Quadrangle, 1967. U.S. Geological Survey, Topographic Map, Hanover, MA Quadrangle, 1974. U.S. Geological Survey, Topographic Map, Duxbury, MA Quadrangle, 1975. U.S. Geological Survey, Topographic Map Index, 1991. Wentworth, T. R. and Johnson, G. P., Use of Vegetation for the Designation of Wetlands, U.S. Fish and Wildlife Service, Washington, D.C., 1986.

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Kent, Donald M. “Evaluating Wetland Functions and Values” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

3

Evaluating Wetland Functions and Values Donald M. Kent

CONTENTS Functions and Values Aquatic and Wildlife Habitat Educational and Scientific Venue Elemental Transformation and Cycling Flood Flow Alteration Groundwater Recharge Particle Retention Production Export Raw Materials Recreation Soil Stabilization Evaluating Functions and Values Representative Evaluation Techniques Expert Opinion Wetland Evaluation Technique Rapid Assessment of Wetlands (RAW) Wetlands Integrated Monitoring Condition Index (WIMCI) Hydrogeomorphic Assessment (HGM) Habitat Evaluation Procedures (HEP) Virtual Reference Wetlands (VRW) Economic Valuation Economic Valuation Methodologies Direct Economic Valuation Indirect Economic Valuation

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The Value of the World’s Ecosystem Services and Natural Capital References

As do all ecosystems, wetlands have functions and values. Functions are processes that are inherent to a wetland. They derive from the wetland’s hydrological, geological, biological, and chemical characteristics. For example, groundwater recharge is a wetland function that occurs when water in the wetland, derived from precipitation, surface runoff, or both, infiltrates downward through permeable soils to the groundwater table. Wetland functions occur regardless of whether there are people present to benefit from these processes. Wetland values are functions that prove useful or are important to people. The aforementioned wetland functioning to recharge groundwater will possess a groundwater recharge value only if the recharged groundwater is used by local or regional populations. Values may be provided within the confines of the wetland, for example, recreation, or beyond the wetland boundaries, for example, flood protection. Another characteristic of wetland values is that they vary with time and circumstances. Again returning to the example of a groundwater recharge wetland, a downstream community drawing drinking water from a surface impoundment does not view the wetland as valuable to its drinking water supply. Should the surface water supply diminish or become contaminated, and groundwater withdrawal become necessary, that wetland now takes on value. Clearly, wetland functions and values are inextricably linked. Values cannot be provided without there first being a function. Conversely, a function has no value until someone exploits that function. Recognizing the confounding nature of the relationship between wetland function and value, many functions and values have been attributed to wetlands (Amman et al., 1986; Mitsch and Gosselink, 1993; Adamus et al., 1987; Reimold, 1994; Brinson, 1995). Some of the commonly recognized functions and values of wetlands are listed in Table 1 and described briefly below. Table 1

Wetland Functions and Values

Aquatic and wildlife habitat Educational and scientific venue Elemental transformation and cycling Flood flow alteration Groundwater recharge Particle retention Production export Raw materials Recreation Soil stabilization

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FUNCTIONS AND VALUES Aquatic and Wildlife Habitat All wetlands, with the exception of those that have been severely degraded, provide habitat for wildlife. And wetlands with seasonal or permanent surface water support fish and other aquatic vertebrates and invertebrates. Many threatened and endangered species are associated with wetlands. The type and degree to which aquatic and wildlife habitat is provided is dependent upon local and landscape characteristics including water depth and permanence, vegetation type and cover, habitat size, and the nature of the surrounding environment (Forman and Godron, 1986; Kent, 1994). Educational and Scientific Venue Numerous public and private organizations exist for the purpose of educating people about the importance of wetlands. Educational topics include awareness, regulations and legislation, conservation and planning, and science and management (Drake and Vicario, 1994). Wetlands provide an opportunity for studying fundamental biological and ecological principles including energy flow, biogeochemical cycling, population biology, and community structure. As well, wetlands are the focus of more specific studies related directly to inherent functions and values such as pollutant removal, habitat provision, and flood attenuation. Elemental Transformation and Cycling Wetlands serve as sinks, sources, or transformers of many inorganic and organic chemicals, including those of ecological and socioeconomic importance such as nitrogen and phosphorus, carbon, sulfur, iron, and manganese. Chemicals enter the wetland through hydrologic pathways such as precipitation, surface or groundwater, tidal exchange, or alternatively through biotic pathways including photosynthetic fixation of atmospheric carbon and bacterial fixation of nitrogen, respectively. Wetlands export or lose chemicals by burying in the sediment, outflow in surface or groundwaters, denitrification, atmospheric loss of carbon dioxide, ammonia volatilization, or methane and sulfide release. While within the wetland, chemicals may become part of the litter, remineralized, translocated in plants, or transformed by changes in redox potential or biotic components. Flood Flow Alteration Wetlands have the potential for reducing downstream peak flows and delaying the timing of peak flows. Water from precipitation, overbank flow, overland flow, and subsurface flows may be detained in wetlands by depressions, plants, and debris, or as the result of the wetland slope. Alternatively, water may be retained in the wetland, infiltrate, and recharge surficial groundwater. The importance of wetlands

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for reducing downstream flooding increases with an increase in wetland area, distance the wetland is downstream, size of the flood, closeness to an upstream wetland, and the lack of other upstream storage areas (Ogawa and Male, 1983, 1986). Coastal wetlands also have the capacity to alter flood flows as well as reduce flood wave severity. In this case, salt marshes and mangrove forests absorb the energy of coastal storms, thereby protecting inland areas. Groundwater Recharge Wetlands with pervious underlying soils recharge underlying materials, groundwater, or aquifers. Recharge is thought to occur primarily around the edge of wetlands, making groundwater recharge relatively more important in smaller wetlands. As most wetlands are thought to have impervious underlying soils, the majority of wetlands may not exhibit this function and value (Larson, 1982; Carter and Novitzki, 1988). Particle Retention Wetlands trap and retain sediments, nutrients, and toxicants, primarily through physical processes. Reduction in water velocity causes sediments, and chemicals sorbed to sediments, to settle. Dissolved elements and compounds are retained with inorganic and organic particulates after sorption, complexation, precipitation, and chelation. In contrast to chemical transformation and cycling, incoming particles are subject to long-term accumulation or permanent loss from incoming water sources through burial in the sediments or uptake by vegetation. Production Export Some wetlands, especially those with high primary productivity, export dissolved and particulate organic carbon to downslope aquatic ecosystems. Plant material and other organic matter are leached, flushed, displaced, or eroded from the wetland, providing the basis for microbial and detrital food webs. Raw Materials Wetlands are a source of plants and animals that serve as raw materials for various domestic, commercial, and industrial activities. Forested wetlands, for example, bottomland hardwoods and cypress swamps, are a source of lumber. Lower quality timber and woody shrubs are used for the production of other wood products, paper pulp, or firewood. Marsh vegetation is used for food (e.g., rice), fodder, thatch for roofs, and other commodities. Wetland wildlife, fish, and shellfish are consumed as food, and wildlife skins are used for clothing and related items. Because of the extractive nature of this function and value, the provision of raw materials is likely to have serious impacts on other wetland functions and values. Sustainable practices can minimize these impacts.

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Recreation Wetlands provide passive and active recreation including fishing, hunting, birdwatching, hiking, canoeing, photography, and others. The opportunity for recreation is related to access and landscape heterogeneity. Recreation can at times be incompatible with other functions. Soil Stabilization Vegetated wetlands have the potential for stabilizing underlying soils. Stems, trunks, and branches dissipate water energy through frictional resistance and reduce erosive forces. Roots bind soil. The dissipation of erosional forces and binding of soil affords protection to nonwetlands in coastal and in riverine areas.

EVALUATING FUNCTIONS AND VALUES The white and gray literature is replete with methods for evaluating wetland functions and values. Differences among the methods are reflected in the precision, accuracy, and reliability of conclusions. Critical factors to consider when selecting or interpreting evaluation methods are whether functions and values are measured directly or implied through indicators, whether evaluated data are qualitative or quantitative, whether the evaluation was conducted off-site or on-site, and whether assumptions and limitations are clearly stated. In general, a method should be selected based upon the type and level of information desired, available labor and economic resources, and the required time scale. Several representative evaluation methods are described below. In many circumstances, a combination of two or more of these methods, or development of an original method, may be most appropriate. Representative Evaluation Techniques Expert Opinion Expert opinion is perhaps the simplest, quickest, and least expensive technique for evaluating wetland function and value. The technique is most applicable when a functional assessment is required on short notice, when money is a limiting factor, and a precise or accurate evaluation is not essential. However, when a group of experts is convened and empirical information is available, expert opinion can represent actual function and value with fair accuracy. At its simplest, expert opinion consists of the professional judgment of an individual conversant with wetland ecological processes. Individual professional judgment should not be the entire basis for decision making when the decision can have serious consequences. More commonly, expert opinion consists of a convention of experts that come together with the goal of reaching consensus. The

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consensus opinion can be given more weight, and more reasonably be used in critical decision making. The two expert opinion techniques that have enjoyed widespread use are the Nominal Group Technique and the Delphi Technique (Delbecq et al., 1975). The two techniques are similar, except that the Nominal Group Technique requires faceto-face meeting(s) of participating experts, whereas the Delphi Technique is typically conducted through correspondence. The Nominal Group Technique is the quicker and more cost-effective technique if the convening experts are proximally located. Conversely, the Delphi Technique may be less costly and less time consuming for participants that interact poorly or are geographically disjunct. In general, the Delphi Technique will require more time to conduct and complete. The Delphi Technique is described to illustrate the Nominal Group and Delphi Techniques process. Delphi was the meeting site in Greece where Oracles met to discuss matters of the time and issue opinions. In modern times, the Delphi process consists of a discussion among knowledgeable individuals with the goal of reaching an agreeable conclusion (Pill, 1971). There are two assumptions fundamental to the process: 1. Expert opinion is sufficient input to decision making when absolute answers are unknown. 2. The collective decision of a group of experts will be more accurate than the professional judgment of an individual.

Involved in a Delphi process are three separate groups: the decision makers, a moderator, and experts (Turoff, 1970). Decision makers initiate the process by posing a question, and then seek an individual or group to moderate the process. The moderator identifies experts, designs the initial and follow-up questionnaires, and summarizes the expert responses. The experts respond to the question posed by the decision makers and transmitted by the moderator. Generally, the experts are polled, responses are tabulated, analyzed, and returned to the experts, and the experts respond again based upon the aggregate responses. The process is repeated until a consensus is reached. The identity of the experts may remain hidden to all parties except the moderator throughout the process. Delbecq et al. (1975) indicated that the quality of Delphi responses is strongly influenced by the interest and commitment of the experts. One area in which the Delphi Technique has been applied with some success is in the development, habitat suitability curves for fish (Crance, 1985, 1987a, 1987b). Habitat suitability curves describe the relationship between a habitat variable (e.g., water temperature or bottom substrate) and the probability that a fish will use a habitat with that particular characteristic. Crance (1987b) has offered guidelines for developing habitat suitability curves, based in part upon general recommendations by Delbecq et al. (1975). The guidelines are believed to be applicable to terrestrial species as well. The number of experts is governed by the number of respondents needed to constitute a representative pooling of judgments, and the information processing capabilities of the monitor. A total of 8 to 10 experts are likely an optimal number, ©2001 CRC Press LLC

although more or less may be sufficient. Crance (1987b) develops a list of 15 to 20 experts, and then prioritizes the list based upon best knowledge of the species’ habitat requirements, geographical coverage, and enthusiasm. The experts should represent a diversity of knowledge about the habitat use by the species, and overrepresentation by any single stakeholder group should be avoided. Experts are mailed an information packet that reiterates the purpose of the exercise and provides guidelines for responding. A response time of about 10 days is established. A second information packet is sent after 4 to 6 weeks which summarizes the results of the first round and includes the preliminary suitability index curves for each variable and life stage considered to be important, new questions, and instructions for the second round. Experts review the preliminary suitability index curves and indicate their agreement or disagreement. Disagreement with a preliminary curve requires sketching of a new curve and providing explanatory comments. Responses to the second round are summarized by the monitor and returned to the experts for further review and comment. The process continues until an acceptable level of agreement is reached. A final report is generated which includes feedback to the experts, and which summarizes exercise goals, process, and results. Crance (1985) has concluded that Delphi exercises are not a replacement for empirical curve development, but provide a more updated and interactive exchange of scientific information than can be achieved with a literature search. Also, Delphiderived curves tend to represent average values of habitat quality for a species and, therefore, will be useful only for predicting average suitability indices. Wetland Evaluation Technique The Wetland Evaluation Technique (WET, Adamus et al., 1987) was developed upon recognition that professional expertise may not always be available, and can be difficult to reproduce. WET’s objectives are to assess most recognized wetland functions and values, be applicable to a wide variety of wetland types, be rapid and reproducible, and have a sound technical basis in the scientific literature. There are 11 functions and values assessed by WET (Table 2). In addition, WET assesses the suitability of wetland habitat for 14 waterfowl species groups, 4 freshwater fish species groups, 120 species of wetland-dependent birds, 133 species of saltwater fish and invertebrates, and 90 species of freshwater fish. Adamus et al. (1987) suggest that WET can be used to compare different wetlands, estimate impacts from wetland modification, prioritize wetlands for acquisition or more detailed study, develop conditions for permits, and compare enhanced, restored, or created wetlands with reference wetlands. Geographically, WET is designed for use in the contiguous United States. Users should, at a minimum, have an undergraduate degree in biology, wildlife management, environmental science or a related discipline, or have several years of experience in one of these areas. Knowledge of the Fish and Wildlife Service classification system (Cowardin et al., 1979, see Chapter 1) and an ability to delineate wetlands are also recommended. WET evaluates functions and values in terms of social significance, effectiveness, and opportunity. Social significance assesses the value of a wetland to society due to its special designations, potential economic value, and strategic location. For ©2001 CRC Press LLC

Table 2

Functions and Values Assessed by the Wetland Evaluation Technique (WET, Adams et al., 1987)

Groundwater recharge Groundwater discharge Floodflow alteration Sediment stabilization Sediment/toxicant retention Nutrient removal/transformation Production export Wildlife diversity/abundance Aquatic diversity/abundance Recreation Uniqueness/heritage

example, a wetland would have a high social significance value for groundwater recharge if it were a sole source aquifer, Class II Groundwater, or had wells, and if it were used as a source of water by a nearby population. Effectiveness assesses the capability of a wetland to perform a function owing to its physical, chemical, or biological characteristics, and opportunity assesses the opportunity for a wetland to perform a function to its level of capability. For example, wetlands with a high effectiveness and opportunity for recharging groundwater would have permeable substrata, a negative discharge differential, and no outlet or a restricted outlet. Functions and values are characterized based upon physical, chemical, or biological processes and attributes. Characterization is accomplished by identifying threshold values for predictors—simple or integrated variables that directly or indirectly measure the physical, chemical, or biological processes and attributes of a wetland and its surroundings. Predictors are chosen for ease of measure or evaluation and vary in directness and accuracy. Threshold values for predictors are established by answering questions, and the responses to the questions are analyzed in a series of interpretation keys. The interpretation keys define the relationship between predictors and functions and values based upon information found in the technical literature. Functions and values are assigned a qualitative probability rating of high, moderate, or low. The ratings are not direct estimates of the magnitude of a wetland function or value, but are an estimate of the probability that a function or value will exist or occur. In practice, WET requires three steps: preparation, question response, and interpretation (Figure 1). Preparation includes obtaining resources, establishing the context, and defining the assessment and surrounding areas. Type and level of evaluation are also determined at this time. The Social Significance Evaluation has two levels: the first level has 31 questions and can be completed in 1 to 2 hr. The second level refines the probability rating for Uniqueness/Heritage function and value, and requires several hours to several weeks to complete depending upon the availability of information.

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Figure 1

Evaluation process for the Wetland Evaluation Technique (WET, Adamus et al., 1987).

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Effectiveness and opportunity are evaluated concurrently at three levels. Each level consists of a series of questions, and successive levels build upon previous levels to develop an increasingly detailed characterization. The level selected depends upon available time and information, and the desired confidence in the evaluation results. Level 1 can be conducted off-site in 1 to 2 hr. Level 2 requires a site visit and 1 to 3 hr. Level 3 requires a site visit and detailed physical, chemical, and biological monitoring data. The second level is recommended as an appropriate level in most circumstances. Interpretation is accomplished through a series of keys; each key consists of a series of boxes. Within each box are coded references to a question or group of questions, and each coded reference is followed by a specified answer of “yes” or “no.” A “true” or “false” arrow leads from each box to either another box or to a probability rating. The user proceeds through each key until a probability rating has been assigned to each function and value for each type of evaluation. There are social significance keys for 11 functions and values, effectiveness keys for 10 functions and values, and opportunity keys for 3 functions and values. The Habitat Suitability Evaluations are accomplished in the same manner using answers to questions in Effectiveness and Opportunity Evaluations 1, 2, and 3. Dougherty (1989) evaluated the applicability of WET to high elevation wetlands in Colorado. Two subalpine wetland complexes, Cross Creek and Willliams Fork, at similar elevations but of differing hydrologic regime, size, and geomorphology were studied in 1985 to evaluate WET’s ability to distinguish between similar wetlands and to compare WET’s evaluations to collected data. At both sites, data were available on groundwater level and surface water stage, groundwater and surface water quality, vegetation cover and standing crop, and stream gauging. The wetlands were assessed using WET Social Significance Levels 1 and 2, Effectiveness and Opportunity Levels 1, 2, and 3, and Habitat Suitability. The evaluation indicated differences in ratings between the two wetland complexes and differences between the WET probability ratings and empirical data. Of the 24 WET probability ratings, 13 were considered questionable, 4 were supported by empirical data, and 7 were rated moderate and thus considered neutral by Dougherty (1989). The 13 questionable ratings centered on three issues. First, WET is insensitive to the degree of overbank flooding which is a major hydrologic distinguishing characteristic of montane and subalpine wetlands in Colorado. Second, WET does not consider snowmelt which drives high elevation wetland hydrology. Third, WET’s heavy reliance on locality (“a relatively small political or hydrologic area”) as a predictor of social significance functions and values may have artificially applied different probability ratings for floodflow alteration and nutrient removal/transformation to the two wetland complexes. In the opinion of the author, WET was most accurate in instances where more detailed data were available (e.g., groundwater measurement) to support the Effectiveness and Opportunity Level 3 assessment. However, the application of WET to this situation is limited by questions that do not appear to be well suited for assessment of high elevation wetlands. In part, this may be attributed to the reliance of WET on the technical literature which is sparse for this region. In closing,

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Dougherty (1989) cautions that WET should be considered as a broad-brush tool for the organization of information and decision making, and not an end in itself. Rapid Assessment of Wetlands (RAW) Kent et al. (1990) developed a macroscale wetland function and value assessment technique to facilitate preliminary land-use planning efforts. The technique was designed to assess widely recognized wetland functions and values, provide expeditious field application, apply to a variety of wetland types, and to be reproducible. The function and value assessment incorporates functions and values, and critical criteria identified in WET (Adamus et al., 1987) and in the “Method for the Evaluation of Inland Wetlands in Connecticut” (Amman et al., 1986). There are 11 functions and values, identical to those of WET, assessed using available resources (e.g., U.S. Geological Survey maps, soil surveys, National Wetland Inventory maps, etc.) and one or more field visits. Each function and value is assessed relative to critical criteria (Table 3), which are used to determine whether or not a wetland potentially provides a function and value under consideration. A wetland is presumed effective for a function and value if the assigned criteria are satisfied. Conversely, a wetland is presumed ineffective for a function and value if the assigned criteria are not satisfied. The cumulative value of a wetland is determined by summing the number of positive responses, dividing this sum by the number of functions and values being assessed (less than or equal to 11), and multiplying the resultant quotient by 100. Wetland systems with percents ranging from 1 to 20 are assigned a poor value, 21 to 40 a below average value, 41 to 60 an average value, 61 to 80 a high value, and 81 to 100 a very high value. The State of Connecticut Department of Transportation, Bureau of Planning, conducted a macroscale delineation and function and value assessment of wetlands at Bradley International Airport in 1990. The purpose of the delineation and assessment was to develop a wetland resource map that would provide guidance to the Department in the development of the Bradley Master Plan, and to facilitate future planning by identifying areas requiring more detailed investigation at a later date. The area covered by the determination and assessment was approximately 405 ha (1000 acres). Identified were 18 separate wetland complexes, the majority of which were broad-leaved, deciduous forested wetlands. Wetlands at the airport were assessed using RAW in the spring of 1990. The majority of the wetlands were assessed as poor value. These wetlands were primarily small, isolated wetlands effective only for flood storage and groundwater recharge. Two larger, contiguous wetlands were assessed as high value, effective for all wetland functions and values except for aquatic diversity and abundance, recreation, and uniqueness and heritage. Intermediate size wetlands with a hydrological connection to other wetlands were assessed as below average and average value. RAW adequately satisfied the planning goals of the Connecticut Department of Transportation, Bureau of Planning. The assessment was conducted in a relatively short period of time and at low cost. Wetland experts with the Bureau of Planning reviewed the assessment and found its conclusions consistent with their opinions of site functions and values. Nevertheless, the authors caution that this technique is ©2001 CRC Press LLC

Table 3

Functions and Values and Assessment Criteria for the Rapid Assessment of Wetlands (RAW, Kent et al., 1990) Function and Value

Aquatic diversity and abundance

Flood flow alteration or flood storage

Groundwater discharge Groundwater recharge Nutrient removal and transformation Production export

Recreation Sediment stabilization Sediment and toxicant retention Uniqueness and heritage Wildlife diversity and abundance

Criteria Permanent open water; open water and vegetation interspersion, water quality suitable for aquatic organisms Regulated outflow, perceived outflow less than perceived inflow, greater than 200 acres and at least 70 percent vegetation coverage Pervious substrate, nonfringe wetlands with outlet only Pervious substrate, permanent inlet and no outlet, impermanent inundation Sediment retention, well-vegetated, low water flow velocity Permanent outlet, high primary productivity, potential erosive conditions, permanent or periodic high water flow velocity Public use permitted Potential sediment sources, reduced water inflow velocity, well-vegetated Potential sediment and toxicant source, absent or constricted outlet, well-vegetated Critical habitat for threatened and endangered species, historical or archaeological site Large and vegetatively diverse, moderate-sized oasis, floodplain

applicable only to planning efforts, and that there is no scientific basis for assigning a cumulative wetland value. Wetlands Integrated Monitoring Condition Index (WIMCI) The Wetlands Integrated Monitoring Condition Index (WIMCI) was intended to provide a framework for cost effective, scientifically responsive monitoring of enhanced, restored, and created wetland functions and values, particularly those associated with local, state, and federal permit activities (Kent et al., 1992). The authors recognized that the standard for measuring success was based largely on structural parameters related to vegetation, and that functional approaches for monitoring wetland ecosystems or addressing impacts were largely nonexistent (Kusler and Kentula, 1989). WIMCI was designed to directly assess the majority of wetland functions, to be flexible, simple to use, produce repeatable results, relatively inexpensive, and to be accomplished in a reasonable period of time. WIMCI assessed eight functions (Table 4). Values, such as uniqueness, heritage, recreation, and education, for which insufficient published literature suitable for objective assessment was lacking, were excluded. So, too, consumptive functions and values (e.g., agriculture, forestry) inconsistent with the intended use of the index were also excluded. The eight assessed functions and values are measured directly and expressed as a fraction of a reference wetland function and value. Individual ©2001 CRC Press LLC

functions and values can then be averaged to produce the WIMCI. Condition index values range from zero for a monitored wetland that does not provide any functions, to one for a monitored wetland that functions at a level comparable to the reference wetland. Instances in which the monitored wetland functions and values exceed that of the reference wetland are also assigned a value of one. Table 4

Wetlands Integrated Monitoring Condition Index (WIMCI) Functions, Values, and Measurements (Kent et al., 1992)

Function and Value

Measurement

Aquatic habitat Flood attenuation Groundwater recharge Nutrient metabolism Production export Sediment retention Toxicant retention

Animal species list Flood storage capacity Recharge volume Total nitrogen or total phosphorus Total suspended organics Total suspended solids Heavy metals, volatile organics, and petroleum hydrocarbon analysis Animal species list

Wildlife habitat

As the overall WIMCI averages individual functions and values, it is compensatory in nature. A high function and value can offset an absent or low individual function and value. The converse is also true. WIMCI does not allow one function and value to be an absolute limiting factor to overall wetland condition. The WIMCI approach is flexible in two ways. First, WIMCI can use more or less functions than described above. Second, WIMCI can be easily modified to assign weights to individual functions and values. The authors note that WIMCI has not been applied in its entirety but is based upon commonly used measurement and assessment techniques. Hydrogeomorphic Assessment (HGM) The HGM assessment was designed to accurately and rapidly measure the net change in wetland function resulting from degradation and restoration (Brinson, 1995, 1996). Changes in function are measured by comparing an impacted or degraded wetland with reference wetlands. The assessment differs from many other evaluation techniques by measuring only functions and not values and by requiring that functions be sustainable. A number of regional guidebooks presenting examples of how to assess wetlands are in press, in development and testing, or being planned (Brinson et al., 1997). HGM first classifies wetlands into functional categories based upon the position of the wetland in the landscape, dominant sources of water, and the flow and fluctuation of the water. A total of seven wetland classes have been recognized: riverine, depressional, slope, mineral soil flats, organic soil flats, estuarine fringe, and lacustrine fringe (see Chapter 1). The classification is intended to simplify the development and conduct of the assessment.

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For a given wetland class, a series of indicators and variables are developed to reflect the measurable properties of the wetland’s functions. Variables are either hydrologic, representing inflow processes, or structural, representing removal processes, and are derived from measurements and visual indicators of wetland function (Table 5). Level of functioning is estimated by combining variables in equations. Variables scale from zero, indicating no function, to one, indicating function equivalent to a reference standard. Focal wetland functions exceeding the reference standard are also assigned a value of one. Unlike other evaluation approaches, the user has the option of reducing the estimated value if the function is determined to be nonsustainable. Table 5

Hydrogeomorphic (HGM) Assessment Functions (Brinson et al., 1995)

Hydrologic Short-term surface water storage Long-term surface water storage Maintenance of high water table Biogeochemical Transformation, cycling of elements Retention, removal of imported substances Accumulation of peat Accumulation of inorganic sediments Habitat and food web support Maintenance of characteristic plant communities Maintenance of characteristic energy flow

Reference wetlands are the most critical component of the HGM assessment. Reference standards are established based upon observations and measurements of wetland sites of the same class. The reference wetlands are intended to be selfsustaining and representative of the highest levels of overall performance. Reference wetlands must be established, at a minimum, for regional subclasses in each physiographic province. Because professional judgment and local knowledge are required to select appropriate reference wetlands, assessment teams comprised of members with complementary scientific skills are recommended. The assessment concludes with a calculation of replacement ratios. To demonstrate its use, Brinson et al. (1995) developed a guidebook for applying the HGM approach for functional assessment to riverine wetlands. To develop the guidebook, wetland sites were studied in the glaciated northeast, Gulf coastal plain, Southwest, Rocky Mountains, Olympic Peninsula, and Puget Sound. The guidebook is to be used by teams of individuals who adapt the information to riverine wetlands of their physiographic region. The information in the guidebook must be modified, calibrated, and tested to determine its effectiveness under local and regional conditions. The guidebook identified 15 functions of riverine wetlands in 4 major categories: hydrologic, biogeochemical, plant habitat, and animal habitat (Table 6). A total of 44 variables were identified: 14 for hydrologic functions, 16 for biogeochemical

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functions, and 27 for habitat functions. From two to nine measurements are required to determine the value of an individual variable. Table 6

Riverine Wetland Functions (Brinson et al., 1995)

Hydrologic Dynamic surface water storage Long-term surface water storage Energy dissipation Subsurface storage of water Moderation of groundwater flow or discharge Biogeochemical Nurtient cycling Removal of imported elements and compounds Retention of particulates Organic carbon export Plant habitat Maintain characteristic plant communities Maintain characteristic detrital biomass Animal habitat Maintain spatial structure of habitat Maintain interspersion and connectivity Maintain distribution and abundance of invertebrates Maintain distribution and abundance of vertebrates

HGM limits include an inability to compare wetland functions between wetland classes. Also, the development of a reference wetland set can require tens of thousands of dollars and several months time, limiting the applicability of HGM to relatively large or controversial projects (Magee, 1996). Criticisms of the HGM assessment approach include poorly defined terminology (e.g., self-sustaining ecosystems, sustainable, ecological stability), and an inability to quantify levels of functions greater than that of reference wetlands (Hruby, 1997). Additional development may be necessary before HGM can be widely used (Opheim, 1996). Habitat Evaluation Procedures (HEP) HEP was developed in response to a need to document the nonmonetary value of fish and wildlife resources and is based upon the assumption that habitat quality and quantity can be numerically described (U.S. Fish and Wildlife Service, 1980a). The assessment procedure is a species-habitat approach to impact assessment. Habitat quality for selected evaluation species is documented with the Habitat Suitability Index (HSI) which is derived from an evaluation of the species’ key habitat components. HSI values are multiplied by area and aggregated to obtain a Habitat Unit (HU). HUs provide the basis for comparison of the relative value of different wetlands at the same point in time, or the relative value of the same wetland at two different points in time. The time and costs associated with a HEP analysis are highly variable, and depend upon size of the study area, number of cover types, number of evaluation species, and the number and types of proposed actions. ©2001 CRC Press LLC

The steps required to conduct a HEP analysis are depicted in Figure 2 (U.S. Fish and Wildlife Service, 1980b). Study limits are defined initially. This includes defining the study area, delineating cover types, and selecting evaluation wildlife species. The study area should include areas where direct or indirect biological changes are expected to occur. The level of delineation of cover types depends upon mapping constraints and the level of detail required for the analysis. A cover type classification applicable to the region is recommended. Evaluation wildlife species can be a single species, a group of species, a species life stage, or a species life requisite. There are two approaches to the selection of species: selection of species with high public interest or economic value and selection of species providing a broad ecological perspective of the assessment area. Examples of the latter include species sensitive to land-use changes, keystone species, and species that represent guilds. HUs are determined next. HUs are the product of the total area of available habitat and the HSI. The total area of available habitat includes all areas that can be expected to provide some support to the evaluation species. The recommended method of describing HSI values is through the use of models which may be in word or mathematical form. Calculating an HSI requires establishing HSI model requirements, acquiring an HSI model, and determining HSI for available habitat. Index values are an estimate of habitat conditions in the subject area relative to a standard of comparison, typically optimal habitat conditions. Values range from 0.0 (no habitat) to 1.0 (optimal habitat). Ideally, HSI models produce an index with a proven, quantified, positive relationship to carrying capacity. Model development is described in Habitat Evaluation Procedures (U.S. Fish and Wildlife Service, 1980b). Figure 3 is an example of an HSI. Schroeder (1982) developed a HSI model for the yellow warbler (Dendroica petechia) to facilitate consideration in HEP evaluations. The model is applicable to the breeding range, season, and habitat of the species. Breeding habitat was determined to be deciduous shrubland and deciduous scrub–shrub wetland (U.S. Fish and Wildlife Service, 1981). There are three variables that describe the suitability of breeding habitat for the yellow warbler: percent deciduous shrub crown cover (variable 1), average height of deciduous shrub canopy (variable 2), and percent of deciduous shrub canopy comprised of hydrophytic shrubs (variable 3). Optimal habitat is considered to exist at shrub densities of 60 to 80 percent, shrub heights of 2 m, and 100 percent hydrophytic vegetation. Other levels of suitability are described by the equation (variable 1 × variable 2 × variable 3)1/3. Baseline HUs describe the existing ecological conditions which facilitate landuse planning and alternatives analysis. Baseline HUs can be compared among areas or to predictions of future HUs for a single area following some change in land use. The latter application constitutes an impact assessment. Application of HEP to impact assessments requires predictions of changes in physical, vegetative, and chemical variables. The same HSI models must be used to determine habitat value at all points in time. HEP allows the incorporation of value judgments through the use of a Relative Value Index (RVI). RVIs are applied as weighting values to the HUs calculated for each evaluation species. Calculation of RVIs requires defining the perceived significance of RVI criteria, rating each evaluation species against each criterion, and

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Figure 2

Habitat Evaluation Procedures (HEP) evaluation process (U.S. Fish and Wildlife Service, 1980b).

transforming the perceived significance of each criterion and each evaluation species’ rating into a RVI. HUs are no longer directly related to carrying capacity after adjustment by RVIs. Unavoidable HU losses can be offset by the development of compensation plans that apply specified management measures to existing habitat to effect a net increase in HUs. Compensation may be in kind, in which the goal is to offset the HU loss for each evaluation species. Alternatively, the compensation plan may have a goal of equal replacement, in which HU losses may be compensated by a gain of an equal number of total HUs, regardless of individual evaluation species’ HUs. ©2001 CRC Press LLC

Figure 3

Suitability index graphs for the yellow warbler (Schroeder, 1982).

Virtual Reference Wetlands (VRW) VRW (Kent et al., 1999) offer an alternative or supplemental approach to assessing wildlife. An idealized standard, the VRW, is established by compiling a list of ©2001 CRC Press LLC

all wildlife species occurring in regional wetlands. Compilation can be accomplished using field guides, agency lists, scientific publications, the gray literature, or expert opinion. The VRW can represent wildlife of all wetland types or can be partitioned by wetland type. Also, the VRW can represent wildlife of all seasons or any one season. Wildlife of the focal wetland, determined by direct observation, is compared to the VRW. The approach is applicable to assessing the success of enhancement, restoration, and creation efforts, or establishing the relative value of a population of wetlands as part of an alternative analysis. The VRW approach is repeatable and, because it is based upon direct observation, allows for evaluation of actual rather than potential function and value. The use of an idealized standard eliminates the need to identify an appropriate reference wetland and facilitates comparison and extrapolation to other wetlands. VRW also provide for calculation of various metrics including relative richness, similarity, and the proportion of upland–wetland species and habitat generalists–specialists. The metrics can be applied at several temporal and spatial scales. The most serious limitation of the VRW approach is the inability to compare wildlife abundance. Kent et al. (1999) applied the VRW approach to evaluation of wildlife in a created wetland. The created wetland wildlife community was found to be relatively rich, due in part to the occurrence of upland birds. The wildlife community most closely resembled that of freshwater marshes of the region consistent with the wetland’s physical characteristics. Habitat specialists were just as likely to occur in the wetland as habitat generalists.

ECONOMIC VALUATION There has been recognition in recent years that wetlands provide ecosystem services, and that these ecosystem services can be economically valued (e.g., Thibodeau and Ostro, 1981; Danielson and Leitch, 1986; Costanza et al., 1997). The total economic value of a wetland consists of its use value and its nonuse value (see Table 7 and Pearce and Moran, 1994). Use value represents the value arising from the actual use of the wetland and is further divided into direct use, indirect use, and option values. Direct use values include recreation, fishing, and forest products. Indirect use values derive from ecosystem functions such as water quality renovation. Option values represent an individual’s willingness to pay for the option to use a value at a later date. Table 7

Categories of Wetland Economic Values (Pearce and Moran, 1994)

Total Economic Value Use values Nonuse values Direct use Indirect use Option values

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Bequest values Existence values

Nonuse values, by contrast, accrue independent of on-site or off-site use of the wetland. Somewhat more difficult to estimate, nonuse values are divided between bequest values and existence values. Bequest values accrue to an individual from the knowledge that others might benefit from the wetland in the future. Existence values derive simply from the knowledge that the wetland exists, even if the individual never sees the wetland. Nonuse value may be quite large relative to use value (Brown, 1993) and has become increasingly accepted. For example, nonuse values have been recognized by the Comprehensive Environmental Response, Compensation and Liability Act (CERCLA, 26 U.S.C. 4611 et seq.) and the 1990 Oil Pollution Act (35 U.S.C. 2701 et seq.). In practice, total economic value is almost impossible to calculate. Unlike financial analyses, which consider whether a project will be profitable for investors, economic analyses must consider the value of a project to society as a whole. All economic analyses rely on the accuracy of costs and an accurate prediction of the future. In practice, many environmental costs are difficult, if not impossible, to estimate. Existing valuation techniques can reasonably distinguish use values from nonuse values but cannot reliably distinguish between option, bequest, and existence values (Pearce and Moran, 1994). Economic valuation is also unlikely to accurately account for underlying, holistic, ecosystem functions, and for intrinsic value (Pearce and Moran, 1994; Pimm, 1997). Economic Valuation Methodologies Direct and indirect approaches are used to estimate the economic value of environmental goods including wetlands (Pearce and Moran, 1994). Direct approaches establish values directly from individuals through surveys or experiments and can be used to determine both use and nonuse values. Indirect approaches establish values from actual, observed market-based information. Nonuse values cannot be determined from indirect approaches. All of the valuation techniques have advantages and limitations that should be understood prior to use. In general, users should select and use techniques that are institutionally acceptable, that consider the needs of the end user(s), and that balance costs against the level of information required (Pearce and Moran, 1994). Direct Economic Valuation Experiments offer the most reliable method for establishing the economic value of a wetland. For example, to determine the recreation value of a wetland, the wetland could be enclosed and an entrance fee charged. In practice, it is difficult to design and establish experiments to determine the economic value of a wetland. Individuals may be asked to rank their preferences as in the Contingent Ranking Method. More commonly the Contingent Valuation Method (CVM) is used and individuals are asked to state what they are willing to pay for some change in the provision, or what they are willing to accept to forgo a change in the provision, of a wetland function or value.

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The CVM has three parts. First, a hypothetical description of the terms under which the function or value is to be offered is presented to the respondent. Second, the respondent is asked questions to determine how much they value the function or value. Respondents may be asked whether or not they want to pay for a function or value if it costs a specified amount or how much they would be willing to pay for the function or value. And third, relating responses to socioeconomic and demographic characteristics tests response validity. CVM willingness to pay responses are analyzed for the frequency distribution of the responses to the valuation questions and relationships between willingness to pay and socioeconomic variables. The CVM is not without its shortcomings (Batie and Shabman, 1982; Diamond and Hausman, 1994; Pearce and Moran, 1994). The quality and presentation of information described to the respondent can have a significant effect on results. The method inherently establishes a hypothetical market which may or may not reflect the behavior of respondents if they had to make actual payments. Also, respondents may not be familiar with the commodity being valued, and responses are sensitive to the described method of payment. Arrow et al. (1993) offer guidelines which improve the reliability of the Contingent Valuation Method. Stevens et al. (1995) used the CVM to estimate the total value of preserving different types of wetlands in New England. A survey was mailed to 2,510 randomly selected New England residents. Each group was asked for their opinions about the importance of wetlands and about rules and regulations governing wetland preservation in New England. Respondents were also asked to rank four types of wetlands: 1. 2. 3. 4.

Wetlands Wetlands Wetlands Wetlands

that provide recreation containing rare species of plants that provide food that provide flood protection, water supply, and water pollution control

Surveyed individuals were asked one of five versions of a contingent valuation question about the amount of sales tax they were willing to pay to prevent wetland loss. The first version asked respondents if they would accept a sales tax of $N each year for the next 5 years for their highest priority wetland type. The second version asked for the respondents’ willingness to pay for all four types of wetland. Version three asked about the willingness to pay to preserve wetlands containing rare species of plants. Version four was the same as version two except less information was provided, and version five asked about willingness to pay for wetland restoration rather than preservation. The survey response rate was 34 percent and 90 percent of respondents said that wetland preservation was very important to them. Of the respondents, 48 percent gave top priority to wetlands that provide flood protection, water supply, and water pollution control; 38 percent gave top priority to wetlands containing rare species of plants; only 9 and 4 percent gave top priority to wetlands providing recreational opportunities and food, respectively. Of the respondents, 64 percent were willing to pay to preserve wetlands. The average respondent was willing to pay between US$73.89 and US$80.41 per year for wetlands that provide flood protection, water

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supply, and water pollution; between US$80.77 and US$96.07 for wetlands containing rare species of plants, and approximately US$114 to preserve all wetland types. The aggregate value of New England wetlands was estimated to be between US$242 and US$261 million per year for wetlands providing flood protection, water supply, and water pollution control, and between US$264 and US$313 million per year for wetlands containing rare species of plants. There was no significant difference in willingness to pay for wetland preservation compared to restoration. The majority of respondents who would not pay were opposed to the special sales tax. The authors concluded that a substantial economic value is associated with wetland preservation and restoration by New England residents. Nevertheless, the response rate, combined with the tendency for respondents to be better educated than the average resident, may have influenced the willingness to pay amount. Indirect Economic Valuation Indirect valuations elicit values from observed market-based information, that is, an individual purchases a good to which the wetland function or value can be related (Pearce and Moran, 1994). Obtained values will be sufficient for cost–benefit purposes. Indirect valuations can be divided into two categories: surrogate market and conventional market. Surrogate market techniques consider markets for private goods and services that are related to environmental goods and services. Individuals reveal the value they place on the environmental goods and services by the value they place on the private goods and services. Surrogate market approaches include household production functions and the hedonic pricing method. Household production functions use expenditures for goods or services that are substitutes for the environmental goods or services. The travel cost approach is perhaps the most commonly applied household production function and uses expenditures for travel to recreation sites. Money and time spent by individuals to get to a site are used to estimate willingness to pay for a site’s functions and values. Data requirements for the travel cost approach are extensive and include the number of visitors to a site, the visitor’s place of origin, socioeconomic characteristics, journey duration, direct travel expenses, and purpose of the visit (Pearce and Moran, 1994). Nevertheless, the travel cost approach can value the demand for recreation function and value. Another surrogate market technique, the hedonic pricing method, estimates an implicit price for an environmental good or service, that is, a use value, by looking at real markets in which those goods or services are traded (Pearce and Moran, 1994). For example, house or land values are used to establish a relationship between property prices and proximal environmental attributes. As with the travel cost approach, data requirements are substantial and may be difficult to obtain. Conventional market techniques use market prices, or shadow market prices, to establish the value of an environmental good or service. Recognized are two approaches: dose–response and replacement cost (Pearce and Moran, 1994). The dose–response technique establishes a dose–response function by relating damage to the environment (the response) to the cause of the damage (the dose). A monetary ©2001 CRC Press LLC

damage function is established by multiplying the dose–response function by the price or value per unit of damage. The approach is applicable to environmental changes that have impacts on marketable goods; for example, pollution impacts on fisheries, forestry, and agriculture. Cost of application ranges from inexpensive to expensive depending upon the availability of dose–response functions. The approach is not suited to valuing nonuse benefits. As the name implies, the replacement cost approach estimates the cost of replacing or restoring a damaged environment to its original condition. Replacement costs can be determined by observing actual spending on replacement efforts, or estimation of the cost of replacement. Incomplete assessment of actual damage, and the practical inability to replace all functions and values, can lead to underestimation. Conversely, use of replacement cost to assign value may underestimate ancillary benefits of the replacement. The Value of the World’s Ecosystem Services and Natural Capital Costanza et al. (1997) estimated the value of the world’s ecological services and natural capital stocks to make the range of potential values more apparent, to establish a first approximation of the relative magnitude of global ecosystem services, to establish a framework for further analysis, and to point out those areas in need of more research. Ecological services are defined as the “flows of materials, energy, and information from capital stocks that combine with manufactured and human capital services to produce human welfare.” Capital stocks are “a stock of materials or information that exists at a point in time.” Trees, minerals, ecosystems, and the atmosphere are examples of the latter. Ecosystem services were grouped into 17 major categories and applied to 16 biomes. Only renewable services were considered. The unit value of ecosystem services was estimated by synthesizing previous studies which used a variety of methods including direct observation, contingent valuation, and replacement cost. The areal extent of the 16 ecosystems was determined and multiplied by the unit value of the ecosystem services to estimate the total value of the world’s ecosystem services and natural capital. This value is estimated to be in the range of US$16–54 trillion per year, with an average of US$33 trillion per year. Wetlands were estimated to have a value of US$29,571 per hectare per year and a total value of US$4.9 trillion per year. This is approximately 15 percent of total global ecosystem services. For purposes of the study, the wetland biome consisted of freshwater wetlands (swamps, bogs, riparian wetlands, and floodplains) and coastal wetlands (tidal marshes and mangroves). Assigned to freshwater wetlands were 10 ecosystem services which had a total value of US$19,580 per hectare per year and a total global value of US$3.2 trillion per year (Table 8). Water supply and disturbance regulation (i.e., storm protection) were the most significant ecosystem services of freshwater wetlands. Assigned to coastal wetlands were 6 ecosystem services which had a total value of US$9990 per hectare per year and a total global value of US$1.6 trillion per year. Waste treatment was the most significant ecosystem service provided by coastal wetlands, accounting for 67 percent of the total value.

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Table 8

Average Global Value of Annual Wetland Services (Costanza et al., 1997)

Services Gas regulation Disturbance regulation Water regulation Water supply Waste treatment Habitat/refuge Food production Raw materials Recreation Cultural Total

1994 US$ per Hectare per Year Freshwater Coastal Total 265 7,240 30 7,600 1,659 439 47 49 491 1,761 19,581

1,839

6,696 169 466 162 658 9,990

265 9,079 30 7,600 8,355 608 513 211 1,149 1,761 29,571

Costanza et al. (1997) consider their estimate of the value of ecosystem services to be a minimum estimate because of a number of uncertainties, including the exclusion of the infrastructure value of ecosystems. Pimm (1997) agrees that the estimate is clearly an underestimate, in part because ecosystems are exceedingly complex and poorly understood and, therefore, cannot reasonably be replicated. Pimm (1997) also questions the moral right to place monetary values on sustaining the environment for future generations. Both Costanza et al. (1997) and Pimm (1997) recognize that the value of ecosystem services will increase exponentially as they become scarcer. Constanza et al. (1997) conclude that their (and subsequent) valuation of ecosystem services will help modify systems of national accounting and provide a basis for project appraisal.

REFERENCES Adamus, P. R., Clairain, E. J., Smith, R. D., and Young, R. E., Wetland Evaluation Technique (WET), Vol. II, Methodology, Operational Draft Department of the Army Waterways Experiment Station, Vicksburg, MS, 1987. Amman, A. P., Franzen, R. W., and Johnson, J. L., Method for the Evaluation of Inland Wetlands in Connecticut, Connecticut Department of Environmental Protection, Bulletin No. 9, 1986. Arrow, K., Solow, R., Portney, P. R., Leamer, E. E., Radner, R., and Schuman, R., Report of the NOAA panel on contingent valuations, Fed. Regist., 58(10), 4602, 1993. Batie, S. and Shabman, L., Estimating the economic value of wetlands: principles, methods, and limitations, Coastal Zone Manage. J, 10, 255, 1982. Brinson, M., The HGM approach explained, Nat. Wetl. Newsl., November/December, 7, 1995. Brinson, M., Assessing wetland functions using HGM, Nat. Wetl. Newsl., January/February, 10, 1996. Brinson, M. M., Hauer, F. R., Lee, L. C., Nutter, W. L., Rheinhardt, R. D., Smith, R. D., and Whigham, D., A Guidebook for Application of Hydrogeomorphic Assessments to Riverine Wetlands, Final Report to the U.S. Army Corps of Engineers, Waterways Experiment Station, Vicksburg, MS, 1995.

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Brinson, M., Lee, L., Ainslie, W., Rheinhardt, R. D., Hollands, G. G., Smith, R. D., Whigham, D. F., and Nutter, W. B., Common misconceptions of the hydrogeomorphic approach to functional assessment of wetland ecosystems: scientific and technical issues, Wetl. Bull., 14(2), 16, 1997. Brown, T., Measuring nonuse value: a comparison of recent contingent valuation studies, W-133 Benefits and Cost Transfer in Resource Planning, 6th Interim report, Department of Agricultural and Applied Economics, University of Georgia, Athens, 1993. Carter, V. and Novitzki, R. P., Some comments on the relation between ground water and wetlands, in The Ecology and Management of Wetlands, Vol. 1, Ecology of Wetlands, Hook, D. D., McKee, W. H., Jr., Smith, H. K., Gregory, J., Burrell, V. G., Jr., DeVoe, M. R., Sojka, R. E., Gilbert, S., Banks, R., Stolzy, L. H., Brooks, C., Matthews, T. D., and Shear, T. H., Eds., Timber Press, Portland, OR, 1988. Costanza, R., D’Arge, R., de Groots, R., Farber, S., Grasso, M., Hannon, B., Limburg, K., Naeem, S., O’Neill, R. V., Paruelo, J., Raskin, R. G., Sutton, P., and van den Belt, M., The value of the world’s ecosystem services and natural capital. Nature, 387, 253, 1997. Cowardin, L. M., Carter, V., Golet, F. C., and LaRoe, E. T., Classification of Wetlands and Deepwater Habitats of the United States, U.S. Fish and Wildlife Service Publication FWS/OBS-79/31, Washington, D.C., 1979. Crance, J. H., Delphi Technique Procedures Used to Develop Habitat Suitability Index Models and Instream Flow Suitability Curves for Inland Stocks of Striped Bass, U.S. Fish and Wildlife Service WELUT-85/WO7, 1985. Crance, J. H., Habitat suitability curves for paddlefish developed by the Delphi technique, N. Am. J. Fish. Manage., 7, 123, 1987a. Crance, J. H., Results on the use of the Delphi technique for developing category I habitat suitability criteria for redbreast sunfish, in Proceedings of Species Criteria Workshop, Fort Collins, CO, December 10–12, 1986, 1987b. Danielson, L. E. and Leitch, J. A., Private versus public economics of prairie wetland allocation, J. Environ. Econ. Manage., 13(1), 1986. Delbecq, A. L., Van de Ven, A. H., and Gustafson, D. H., Group Techniques for Program Planning—A Guide to Nominal Group and Delphi Processes, Scott Foresman, Glenview, IL, 1975. Diamond, P. and Hausman, J., Contingent valuation: is some number better than no number? J. Econ. Perspect., 8(4), 45, 1994. Drake, K. and Vicario, M., Wetlands education, in Applied Wetlands Science and Technology, Kent, D. M., Ed., Lewis Publishers, Boca Raton, FL, 1994, 363. Dougherty, S. T., Evaluation of the applicability of the Wetland Evaluation Technique (WET) to high elevation wetlands in Colorado, in Wetlands: Concerns and Successes, American Water Resources Association, 1989, 415–427. Forman, R. T. T. and Godron, M., Landscape Ecology, John Wiley & Sons, New York, 1986. Hruby, T., Continuing the discussion: scientific and technical issues regarding the hydrogeomorphic approach to function assessment of wetlands, Wetl. Bull., 14(3), 23, 1997. Kent, D. M., Designing wetlands for wildlife, in Applied Wetlands Science and Technology, Kent, D. M., Ed., Lewis Publishers, Boca Raton, FL, 1994, 283. Kent, D. M., Reimold, R. J., And Kelly, J. M., Macroscale Wetlands Delineation and Assessment, Technical report to the Connecticut Department of Transportation, Bureau of Planning, Metcalf and Eddy, Inc., 1990. Kent, D. M., Reimold, R. J., Kelly, J. M., and Tammi, C. E., Coupling wetland structure and function: developing a condition index for wetlands monitoring, in Ecological Indicators, McKenzie, D. H., Hyatt, D. E., and McDonald, V. J., Eds., Elsevier Applied Science, London, 1992. ©2001 CRC Press LLC

Kent, D. M., Schwegler, B. R., and Langston, M. A., Virtual reference wetlands for assessing wildlife, Fla. Sci., 62, 222, 1999. Kusler, J. A. and Kentula, M. E., Eds., Wetland Creation and Restoration: The Status of the Science, U.S. Environmental Protection Agency, Report No. 600/3–89/038a, 1989. Larson, J., Wetland value assessment—state of the art, in Wetlands: Ecology and Management, Gopal, B., Turner, R. E., Wetzel, R. G., and Whigham, D. F., Eds., National Institute of Ecology and International Scientific Publications, Jaipur, India, 1982. Magee, D. W., The hydrogeomorphic approach: a different perspective, Wetl. Bull., June, 12, 1996. Mitsch, W. J. and Gosselink, J. G., Wetlands, 2nd ed., Van Nostrand Reinhold, New York, 1993. Ogawa, H. and Male, J. W., The Flood Mitigation Potential of Inland Wetlands, Water Resources Research Center Publication No. 138, University of Massachusetts, Amherst, 1983. Ogawa, H. and Male, J. W., Simulating the flood mitigation role of wetlands, J. Water Res. Plan. Manage., 112, 114, 1986. Opheim, T., HGM: the beast is unleashed, Nat. Wetl. Newsl., May/June, 2, 1996. Pearce, D. and Moran, D., The Economic Value of Biodiversity, Earthscan Publications Ltd., London, 1994. Pill, J., The Delphi method: substance, context, a critique and an annotated bibliography, Socio-Econ. Plan. Sci., 5, 57, 1971. Pimm, S. L., The value of everything, Nature, 387, 231, 1997. Reimold, R. J., Wetlands functions and values, in Applied Wetlands Science and Technology, Kent, D. M., Ed., Lewis Publishers, Boca Raton, FL, 1994. Schroeder, R. L., Habitat Suitability Index Models: Yellow Warbler, U.S. Department of the Interior, Fish and Wildlife Service FWS/OBS-82/10.27, 1982. Stevens, T. H., Benin, S., and Larson, J. S., Public attitudes and economic values for wetland preservation in New England, Wetlands, 15(3), 226, 1995. Thibodeau, F. and Ostro, B., An economic analysis of wetland protection, J. Environ. Manage., 12(1), 19, 1981. Turoff, M., The design of a policy Delphi, Technol. Forecast. Soc. Change, 2, 149, 1970. U.S. Fish and Wildlife Service, Habitat as a Basis for Environmental Assessment, Division of Ecological Services, Department of the Interior, Washington, D.C., 1980a. U.S. Fish and Wildlife Service, Habitat Evaluation Procedures (HEP), Division of Ecological Services, Department of the Interior, Washington, D.C., 1980b. U.S. Fish and Wildlife Service, Standards for the Development of Habitat Suitability Index Models, 103 ESM, U.S. Department of the Interior Fish Wildlife Service, Division Ecological Services, 1981.

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Kent, David J. et al “Ecological Risk Assessment of Wetlands” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

4

Ecological Risk Assessment of Wetlands David J. Kent, Kenneth D. Jenkins, and James F. Hobson

CONTENTS The Human Health Risk Assessment Paradigm Ecological Risk Assessment The Ecological Risk Assessment Framework Problem Formulation Phase Choosing Biological Endpoints Spatial and Temporal Considerations Other Considerations Analysis Phase Exposure Characterization Ecological Effects Characterization Risk Characterization Phase Predictive Ecological Assessments Problem Formulation Exposure Characterization Ecological Effects Characterization Risk Characterization The Quotient Method for Risk Characterization Retrospective Ecological Assessments Problem Formulation Exposure Characterization Ecological Effects Characterization Risk Characterization Summary References

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Until recently, the term risk assessment generally was applied to the estimate of risk to human health, typically from chemical exposure. For example, a cancer risk assessment is an estimate of the risk to humans from carcinogenic compounds. Recently, however, the term risk assessment has been applied to ecological systems. An ecological risk assessment is an estimate of the adverse effect to an ecosystem from chemical, physical, or biological stressors resulting from anthropogenic activity. This recent interest in assessing ecological health is evidenced by several publications (Bartell et al., 1992; Cairns et al., 1992; Suter, 1993; Newman and Strojan, 1998; Lewis et al., 1999) including two documents produced by the U.S. Environmental Protection Agency (USEPA, 1992, 1998). The first of these USEPA documents, Framework for Ecological Risk Assessment (1992), was intended as the first step in a long-term program to develop guidelines for the performance of ecological risk assessments. The second document, Guidelines for the Ecological Risk Assessment, provided more detailed information and is the current guidance on the subject. The principles of ecological risk assessment can be applied to any ecosystem, although they may be particularly relevant to wetlands. The extent and rate of wetland loss, as well as the biologic, economic, and social importance of wetlands, are well documented (Mitsch and Gosselink, 1986). Moreover, the transitional nature of wetlands may make them especially sensitive to stress. Despite the uniform application of assessment principles to ecological systems, individual wetlands are sufficiently different in their spatial, temporal, and physiochemical characters to warrant site-specific sampling and analysis (Figures 1 and 2). These differences will influence the design and interpretation of the ecological risk assessment.

Figure 1

Risk assessment can be broadly applied to a variety of ecosystems. Nevertheless, individual wetlands are sufficiently different in their spatial, temporal, and physicochemical characters to warrant site-specific samping and analysis. A New England saltmarsh is shown here.

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Figure 2

An Arkansas riverine system wetland.

To understand the use of ecological risk assessment for wetland ecosystems, an introduction to the principles of risk assessment is necessary. The current basis for risk assessment is derived from the National Research Council Risk Assessment paradigm (1983). Following this, an in-depth discussion of the USEPA Ecological Risk Assessment Framework (USEPA, 1992) is presented as is a discussion of the challenges and strategies associated with carrying out assessments. Finally, examples of specific applications to wetland ecosystems are provided.

THE HUMAN HEALTH RISK ASSESSMENT PARADIGM The risk assessment paradigm has been used for some time to evaluate the chronic impacts of environmental pollutants on human health. This strategy was initially conceptualized by the National Research Council Risk Assessment Panel (NRC, 1983) and formalized by the USEPA in its 1986 Guidelines (USEPA, 1986a–e). This risk assessment paradigm consists of several components:

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• Hazard identification: does a chemical contaminant represent a specific threat to human health? Establishment of cause–effect relationships is central to this component. • Defining dose–response: what is the relationship between the magnitude of the exposure and the probability of an adverse health effect? • Exposure assessment: what is the potential for human exposure to the chemical of concern? • Risk characterization: what is the potential magnitude of risk to human health given the predicted exposure and dose–response data? What is the uncertainty associated with this risk estimate?

Standard methodologies are employed to evaluate potential threats to human health. The methodologies usually involve determining all relevant effects and then summing those effects to get a total effect value.

ECOLOGICAL RISK ASSESSMENT Ecological risk assessments (ERA) examine the probability that undesirable ecological effects are occurring or may occur as a result of exposure to a stressor or a combination of stressors. The term stressors is used here to reflect the broad range of anthropogenic factors that can result in ecological perturbations. Stressors may include any chemical, physical, or biological factor resulting from anthropogenic activities that can cause an ecological disturbance. Most often, however, the term stressor refers to toxic chemicals. ERAs can be used to address a wide range of issues and are generally classified as predictive or retrospective. Predictive ERAs are designed to assess the risks associated with proposed actions, such as the introduction of new chemicals into the environment and the establishment of new sources of stressors or hypothetical accidents (USEPA, 1991). Predictive risk assessments have usually followed the NRC human health paradigm relatively closely while emphasizing the choice of biological endpoints and related stressor–response data. In contrast, retrospective ERAs address the risks associated with stressors released due to current or previous anthropogenic activities. Examples of retrospective assessments include evaluating the ecological impact of hazardous waste sites and previous releases or spills. The goal of this type of risk assessment is to establish and define the relationship between the pollution source, the distribution of stressors, the exposure of biological endpoints, and the level of effects of this exposure on the ecosystem. Retrospective assessments often take advantage of field data to define contaminant sources and measure adverse biological effect. Various levels of data collection or site-specific assessment may be necessary to provide the information required to design and conduct the retrospective ecological risk assessment, and to achieve a given level of confidence. The challenge here is to establish cause–effect relationships between the source of stressors and any observed ecological effects. Some ERAs, such as those used for wetlands, may involve both predictive and retrospective aspects. For example, in an assessment of a hazardous waste site, the ©2001 CRC Press LLC

current status of the site may require a retrospective evaluation, but the long-term impact of various repetition scenarios would be addressed in a predictive fashion. As the type of risk assessment to be conducted is dependent upon the ultimate application of the results, a clear understanding of the objectives of the risk assessment is essential. The move to ERAs by the regulatory community is driven by a number of factors. From a legal standpoint, many of the underlying statutes require some form of evaluation of ecological risk. For example, the Superfund Amendment and Reauthorization Act (1987) specifies that the actual and potential risk to public health and the environment must be assessed for each hazardous waste site. The use of a basic risk assessment paradigm would provide a structural framework for ecological assessment and a consistent strategy for managing various types of risk. This issue is particularly important when comparing the sensitivity of ecological endpoints relative to the human endpoint. In some cases, nonhuman endpoints may prove to be more sensitive than human endpoints, and would, thus, drive the overall risk assessment. This type of comparison is facilitated by consistency in the strategies used to carry out risk assessments for both types of endpoints. Although human health risk assessment strategies provide a useful model, assessing ecological risk has proven to be more complex. A number of factors contribute to this complexity. • Multiple biological endpoints: these could include multiple species, and various levels of biological organization (e.g., subcellular, individual, population, community, and ecosystem). • More complex exposure pathways: these are determined by the biological endpoints of concern. • Indirect effects: indirect effects such as habitat impairment or disruption of intertrophic relationships may be more important than direct exposure to chemicals. • Evaluating impacts on ecosystems: ecosystems are complex, their function is often not tightly coupled to stressor inputs, and they show resilience and recovery to varying degrees of stress.

In spite of these complexities, there are some distinct advantages to estimating risk to ecological endpoints. Exposures and hazards can often be estimated directly on the species of concern or a closely related surrogate species. This may result in a more accurate estimate of risk, because there is no need for extrapolation from more distantly related species as is almost always the case in human risk assessment. This is particularly useful when multiple stressors or complex exposure matrices are involved. In ERAs, the biological endpoints of interest can often be tested directly against the specific stressor or mixture of stressors of concern, thus eliminating the need to estimate such factors as stressor interactions, chemical form (speciation), and bioavailability. Despite the aforementioned advantages, the added complexity associated with ERAs results in a higher degree of uncertainty than is normally associated with human health-based risk assessments. This complexity requires more effort in the initial planning stages so that the final assessment is well focused. Furthermore, ©2001 CRC Press LLC

some modifications of the basic NRC risk assessment paradigm are required. The Risk Assessment Forum within the USEPA has developed the framework (and now guidelines) for ecological risk assessment which addresses these issues, yet is conceptually consistent with human health risk assessment strategies. It is important to note that slight differences in terminology exist between the various ERA structures currently promoted. However, the elements are fairly analogous. For the sake of simplicity, the discussions in this chapter will utilize the USEPA framework terminology, although the concepts may be drawn from multiple sources. The Ecological Risk Assessment Framework Two major elements that form the basis of the ERA framework are the characterization of exposure and the characterization of ecological effects. Aspects of the two elements are considered in all phases of the framework process. While carrying this common thread throughout the paradigm, the framework is divided into three phases. The phases are problem evaluation, analysis, and risk characterization (Figure 3). Problem Formulation Phase The first step in this process requires defining the specific purpose of the ERA. Although this may seem trivial, many ERAs suffer from lack of clear focus and, as a result, may be ambiguous and misleading. Once the specific purpose of the assessment is defined, the specific goals that must be met to achieve this purpose are formulated. These goals provide a basis for establishing a precise conceptual study design. In undertaking the study design, a number of questions must first be addressed. • • • • •

Is the ERA to be predictive or retrospective? Is the ERA to be site-specific or generic? What type of ecosystem(s) is at risk? What types of stressors are involved? What are the potential source(s) for a given stressor or set of stressors?

Addressing these questions requires a rigorous review of available data. Where existing data are unavailable or incomplete, it may be necessary to carry out a preliminary study, particularly to establish the types of potential stressor. Most important, ecological risk assessment is often an iterative process. A given level of information is required for developing the design and objectives (i.e., the problem formulation). Additional data may be required for the complete ERA. Ultimately, the level of information needed and the extent of new data collection are dependent upon the objectives, such as the level of certainty desired. As well, the level of information is dependent upon the outcome of the previous iterations, for example, how much ecological impact has occurred at a given site or how hazardous a new chemical may be based on laboratory toxicity studies.

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Figure 3

From the Framework for Ecological Risk Assessment (USEPA, 1992).

Choosing Biological Endpoints An important issue at this stage is determining the appropriate suite of biological endpoints to be used in the evaluation. Biological endpoints should be carefully chosen specifically to address the overall goals of the assessment. The parameters to be considered in choosing these endpoints should include the ecological relevance of the endpoint and the spatial and temporal occurrence of the endpoint relative to the distribution of the stressors and potential biological receptors.

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In ecological assessments the distinction is often made between assessment endpoints and measurement endpoints (Warren-Hicks et al., 1989; Suter, 1993). Assessment endpoints represent the ultimate resource(s) or final environmental values that are to be protected. They should have social or biological relevance, be quantifiable, and provide useful information for resource management or regulatory decisions. For example, successful reproduction of a species in danger of extinction is an appropriate assessment endpoint. For populations that are valued for commercial or sport uses such as anadromous fish (e.g., salmon) or estuarine or marine shrimp, the important assessment endpoints could be growth, reproduction, or overall productivity. Other potential assessment endpoints include yield and productivity, market or sport value, recreational quality, and reproductive capability (see WarrenHicks et al., 1989). In general, assessment endpoints focus on population and community parameters because ecological risk assessments are usually concerned with protecting these higher levels of biological organization rather than individual organisms. However, when there is concern for endangered species, the assessment endpoints may indeed focus on the individual organism. In contrast, measurement endpoints represent the specific parameters that are to be measured in a given assessment. They are often chosen based upon practical considerations such as availability and ease of measurement. It is often impossible from a practical sense to measure some endpoints (e.g., endangered species) and other more obtainable measurement endpoints are chosen as surrogates for the actual assessment endpoints of interest. Such measurement endpoints should be well characterized and take into account exposure pathways and temporal factors. Ideally, measurement endpoints should be chosen so that the data from these endpoints can be linked directly or indirectly to appropriate assessment endpoints. This latter issue is often the most difficult to address in ERAs. At the level of the individual, measurement endpoints may include mortality, growth, and fecundity. Abundance and reproductive performance are measurement endpoints on a population level. Other measurement endpoints include species evenness and diversity on a community level and biomass and productivity on an ecosystem level. Assessment of endangered species poses a special problem. Because exposure for an endangered species cannot usually be directly assessed, the residues of a contaminant in a principal food item can be a measurement endpoint. Alternatively, if exposure to a chemical and its potential impact on an endangered species is the assessment endpoint, then a co-existing species with similar life history habits might be used as a surrogate measurement endpoint. Spatial and Temporal Considerations Ultimately, the problem formulation phase should result in the establishment of the study design which will form the basis of the assessment. The design must take into account a number of environmental and ecological factors that may affect the stressors or their potential impact on biological systems. Many of these factors have spatial or temporal components that must be taken into account in the design. Spatial

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factors include potential routes of exposure, other sources of stressors, location of sensitive biological resources, and factors that may modify contaminant mobility or availability. The changing composition of sediments is an example of the latter. Temporal factors may include seasonal changes in physical, chemical, or biological aspects of ecosystems that may influence the magnitude or form of the stressor or its potential to cause a biological effect. For example, increased surface water movement in the wet season in riparian habitats can dramatically affect contaminant migration and the potential for exposure. Also, seasonal variation in physical or chemical parameters such as temperature or pH can modify the bioavailability of contaminants, thus changing the nature of the exposure. The biological characteristics of the exposure matrix may also change temporally. Species may be present only seasonally as is typical of migratory waterfowl, anadromous fishes, or species that migrate seasonally within a local area. Exposure to a given species, population, or community may vary with seasonal changes in life history habits such as seasonal feeding patterns or reproductive cycles. Thus, temporal changes in the biological components of an ecosystem may influence the distribution of the stressor or the availability of the biological endpoint within the ecosystem. Temporal variation would affect the ultimate risk assessment. Therefore, these temporal patterns must be addressed in designing the sampling scheme and data collection for the risk assessment. Temporal considerations may also include long-term historical or predicted trends in stressor influence and potential seasonal variation in stressor impact. Historical or predicted trends are very important to understanding the overall impact of stress on an ecosystem and may be important to the application of risk information in risk management decisions such as remediation plans, wetland restoration, and registration of a new pesticide. Other Considerations Another factor to consider is the presence of biological resources that are either sensitive to stressor impact or may be of particular economic or social importance. Wetlands are nursery areas for many species that inhabit adjacent terrestrial and aquatic communities. Other resources may include populations of economically important species such as anadromous fish populations or endangered species which are protected by law. In completing the problem foundation phase of the ERA, these issues must be addressed in rigorous and systematic fashion. The ultimate goal is the development of a conceptual model that will serve as a basis for the ERA. The model should address the spatial and temporal distributions of potential stressors, appropriate biological endpoints, as well as probable routes and levels of exposure. Finally, the model should be precisely linked to the goals and purposes of the risk assessment. A well-conceived conceptual model is essential to the effective implementation of the subsequent component of the risk assessment.

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Analysis Phase The analysis phase consists of two distinct activities: characterization of exposure and characterization of ecological effects. It is here that the two main elements of an ecological risk assessment are most prominent and most closely interrelated. Exposure Characterization The goal of exposure characterization is to develop an accurate appraisal of the potential for exposure to important biological resource or receptors, as represented by the measurement endpoints, to the identified stressors. Spatial and temporal factors are particularly important in this stage of the ERA. Issues associated with multiple routes of exposure must also be addressed at this time. Both the stressor and the ecosystem must be characterized with regard to the distribution and pattern of change. This is accomplished through the use of modeling or monitoring data, or both, depending on whether the ERA is predictive or retrospective. Stressors can be characterized from direct sampling, laboratory or field testing, or through remote sensing. At this stage, it is important to evaluate the means by which stressors can be modified in the ecosystem through biotransformation or other environmental fate processes such as photolysis, hydrolysis, and sorption. Transport and fate models are often employed, including interactive mass balance models like the Exposure Analysis Modeling System (EXAMS) developed at the USEPA research laboratory in Athens, GA (Burns et al., 1982). Characteristics of the ecosystem that can affect exposure also need to be evaluated. These can include habitat requirements, food preferences, reproductive cycles, and seasonally influenced activities. After stressor and ecosystem characteristics are defined, the spatial and temporal distributions are combined to evaluate exposure. Finally, the magnitude and distribution patterns are quantified for the scenarios developed during the problem formulation phase of the ERA. All sources of uncertainty should also be quantified to the degree possible for the input into the ERA. Recently, probabilistic techniques have been used to characterize uncertainty under various exposure conditions (Burmaster and Anderson, 1994; Power, 1996). Ecological Effects Characterization The goal of this portion of the analysis phase is to identify and characterize any adverse ecological effects that are associated with exposure to a particular stressor or stressors. To the extent possible, these effects should be quantified and any cause and effect relationships evaluated. Initially, an evaluation must be made of all effect data that are relevant to the stressor. The types of data that are relevant will depend upon the characteristics of both the stressor and the ecological component and also on whether the ERA is to be predictive or retrospective. Commonly used data types include aquatic toxicity tests, computer models, quantitative structure activity relationships (QSARs), microand mesocosms, species diversity analyses, artificial substrate comparisons, the ©2001 CRC Press LLC

Wetland Evaluation Technique (Adamus et al., 1987), and others (Suter, 1993; USEPA, 1992; McKim et al., 1987; Barnthouse et al., 1986). Data from laboratory and field observations, as well as from controlled studies, may be used. These data are considered based upon their relevance to the measurement and assessment endpoints. Evaluating the quality of the data, that is, the adequacy of sampling and statistical design, is an important component of this stage. The next step is to quantify the ecological response in terms of the stressor–response relationship. The aim is to describe the relationship between the magnitude, frequency, and duration of the stressor, and the magnitude of the response. The dose–response curve in laboratory aquatic toxicity tests is an example of this type of analysis. Determining relative differences in productivity, biomass, species composition, and diversity between contaminated and reference sites is also a commonly used means of defining this relationship. Any extrapolations required between the stressor and measurement endpoints must be evaluated as well as any extrapolations between measurement and assessment endpoints. The strength of the causal relationship must be quantified to the extent possible. Finally, as with the exposure analyses, quantitative estimates of uncertainty such as natural variability in ecological characteristics and responses should be included in the analyses. Often this analysis of uncertainty can be done probabilistically (Moore, 1996). However, it should be emphasized that although the goal is to provide quantification of the stressor–response relationship, in many cases, it can only be described qualitatively and, therefore, professional judgment plays an important role in most ERAs. Risk Characterization Phase The ultimate purpose of the ERA is to estimate the risk that unacceptable adverse ecological effects will occur due to exposure to anthropogenic stressors. The risk characterization phase provides an estimate of the likelihood that an adverse impact has occurred or will occur. It also should address the relative ecological consequences that are associated with the various levels of risk. Ideally, risk characterization should be quantitative and include estimates of uncertainty. However, the complexity of ecological risk assessments may often preclude precise quantification. Under these circumstances, qualitative estimates of risk are often employed. In the risk characterization phase a hazard-exposure matrix is developed. This matrix represents a fusion of the exposure predictions developed in the exposure characterization phase and the estimates of effects developed in the ecological effects characterization phase. Ultimately, the risk characterization phase should evaluate the ecological consequences or ecological significance of predicted or observed effects. This is simpler in some cases for retrospective risk assessments, because the changes can be documented and quantified in situ. Also, the ecological effects, that is, the resulting changes at higher levels of biological organization or in the ecosystem as a whole can be monitored directly with the proper sampling design and appropriate reference sites. In predictive ecological risk assessments, these effects or changes in ecosystems must be predicted. The degree of uncertainty for such predictions is relatively large in most situations. As the historical database for ecotoxicology and ©2001 CRC Press LLC

ecological risk assessment increases, especially field data sets, these predictions will be made with higher levels of confidence. Especially important is the field validation of risk assessments based on laboratory-based hazard assessments and modeled exposure assessments. As a result of the inherently different goals of predictive and retrospective ecological risk assessments, the remainder of this chapter will provide illustrations of how ERAs may be applied for each of these types with regard to wetland ecosystems. The first type of ERA to be discussed will be the current tiered assessment design, a predictive approach employed by the USEPA in the pesticide regulation program. The second describes a retrospective approach more common in the assessment of hazardous waste sites or spills. Predictive Ecological Assessments A practical example of a predictive risk assessment can be seen in evaluating the potential ecological impact of the introduction of a new herbicide. Under the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) and its amendments, all new pesticides and herbicides must be registered for each specific use. The Office of Pesticide Programs of the USEPA prepared the Standard Evaluation Procedure for Ecological Risk Assessment (SEP) (Urban and Cook, 1986) to define the procedures to be used for the ecological risk assessment process. In short, this document presents a tiered approach to determining the potential for unreasonable adverse effects on the environment as a result of the use of pesticides and herbicides. Additional information for the testing of new pesticides and herbicides can be found in the Pesticide Assessment Guidelines (USEPA, 1982a, b). To illustrate this example, consider a new herbicide for control of weeds in rice paddies adjacent to riparian and emergent wetlands. The goal here is to assess the potential risk of the new herbicide on the wetland ecosystem based on the toxicology and environmental fate of the pesticide under the proposed use pattern. Problem Formulation The first activity in this predictive ERA is to define the scope of the problem. The planned use pattern for the herbicide requires its application to fields that have a great dependence on water flow and, thus, have a large potential for transport of chemicals into adjacent aquatic habitats. It is important that the inherent toxicity of herbicides be assessed in relation to the likely exposure to various ecological compartments. In simple terms, the questions to be answered are how toxic is it and will there be enough exposure for the toxicity to be realized? The typical rice field (Figure 4) may be 10 to 60 ha in size and laser-leveled to control water flow. The fields are not allowed to get too dry and, if sufficient rainfall is lacking, they may be flushed with water to depth of 2.5 cm or so on a fairly regular basis. Once the rice is adequately established, the field is typically flooded to a depth of 7.5 to 15 cm, which is then maintained until harvest. The temporary flush and permanent floodwaters are drained to an adjacent ditch that runs within 100 m to a small bayou. Emergent and riverine wetlands are associated ©2001 CRC Press LLC

with the bayou. Approximately 400 ha of rice drainage enter into the bayou above and below our hypothetical field. For this example, let us assume that the herbicide has a maximum application rate of 2.7 kg (5 lb) active ingredient per hectare and is applied only once a year, just prior to the permanent flood.

Figure 4

The typical rice field may be 10 to 60 ha in size and laser-leveled to control water flow. Flooding and drainage are accomplished through use of the ditch in the foreground.

Exposure Characterization Essentially, there are four means by which the herbicide could enter the adjacent wetland habitats. First, exposure may occur if the pilot applying the material oversprays the field and deposits some of the herbicide directly on the wetland. Second, small droplets of the substance have the potential to drift in the wind during application. Third, owing to the nature of rice culture, periodic releases of water are made from the fields. And finally, severe storms could potentially create an overflow of paddy water into nearby wetlands. Each of these routes should be examined in the ERA. In most cases, measured concentrations of the new herbicide in the field will not be available at this point. Therefore, the potential exposure must be estimated. The endpoint of interest in the exposure characterization is the determination of an estimated environmental concentration (EEC). The EEC provides the risk assessor with a concentration of the herbicide that may be present in the environment. A tiered approach is commonly used to determine the EEC, moving from simple to more complex depending on the need for additional information. ©2001 CRC Press LLC

The simplest way to determine the EEC is to assume that the herbicide is applied directly to the water in the wetlands adjacent to the rice fields. Although this is unlikely, it is possible if, for example, the pilot applying the herbicide overshoots the field. In any case, it provides an extreme worst-case exposure scenario. At the given application rate, and assuming a direct application to water that is 0.6 m deep, the EEC immediately following application would be 919 ppb (Urban and Cook, 1986). Comparing this EEC to ecological effects would allow a cursory assessment of risk to be made. However, this simple method of determining the EEC does not consider variables that could modify the actual concentrations in the adjacent wetland system. For example, normal use practices of the herbicide would preclude direct application to water bodies. Rather, the application would be directly to the field and transport of the chemical would be a function of water flow off the field. In addition, physiochemical and environmental fate characteristics of the herbicide may allow it to partition to the sediments or be metabolized or degraded into either more or less persistent compounds. This method represents a screening level estimation of the worst case environmental concentration. Should some concern remain following the screening level EEC determination, then a more realistic determination is required which may incorporate additional tiers of evaluation. For example, a second level may incorporate additional variables such as drainage basin size, surface area, and percent runoff into the simple equation. A still more complex EEC determination can be made utilizing computer models such the Pesticide Root Zone Model (PRZM) and EXAMS (Burns et al., 1982) to predict herbicide concentrations in aquatic systems resulting from runoff. Finally, in the event that some concern still exists, the last tier EEC determination may include conducting an actual field residue monitoring study in which the herbicide is applied at the maximum label-allowed rate in a test field and residues are measured periodically around the field. The spatial and temporal factors associated with the introduction of a pesticide are important to consider. For the herbicide example, these can include the application and use rates, water flow dynamics in the rice field (e.g., is water released immediately after application or held for a significant time period before release), mobility and persistence of the herbicide (e.g., water solubility, hydrolysis/photolysis rates, metabolism rates, etc.), and potential for bioaccumulation. Other factors include the timing of application with respect to seasonal biological cycles and the type and number of organisms that could potentially be affected. Ecological Effects Characterization The purpose of this segment of the ERA is to characterize any adverse impacts that may be associated with the pesticide. Because this is a herbicide, it is designed to be toxic to certain broadleaf weeds. Therefore, the effects characterization should be limited to nontarget flora and fauna that may be exposed to the herbicide either during use or following runoff. To test all of the species that may come in contact with the herbicide at all of the potential sites of application is unrealistic. Therefore, a series of standard surrogate species are typically tested. ©2001 CRC Press LLC

As with the determination of the EEC, a tiered approach is usually taken to quantify the toxicity of the herbicide to nontarget organisms. Results from the first tier testing determine the need for second tier testing, and so on. Each tier normally examines increasingly sensitive endpoints. Tier I addresses acute toxicity to warm water and cold water fish, as well as to aquatic invertebrates and plants (World Wildlife Fund, 1992). Acute oral and dietary dosages to game birds such as bobwhites and mallards are also evaluated for toxicity. Additional tests of toxicity to estuarine species, honeybees, and others may be required depending on the use pattern and environmental fate of the compound. Tier II testing is sometimes required to determine a no-observable-effect concentration (NOEC) for effects such as reduced survival, growth, and reproduction. Tier II tests for toxicity to fish early life stage or invertebrate life cycle can be triggered by acute toxicity where the median lethal concentration (LC50) is less than 1.0 ppm. These tests can also be triggered by high EEC levels (where EEC is greater than 0.01 of the lowest Tier I LC50 value), chemical persistence in the environment, continuous or recurrent exposures, or high potential for bioaccumulation or reproductive effects. Tier II studies are automatically required for any aquatic use of herbicides. Tier II avian reproduction studies may also be triggered if Tier I avian tests indicate toxicity. Tier III full life-cycle tests with fish may be required if the EEC is more than 0.1 of the NOEC from the Tier II fish early-life-stage study, or if data suggest that there is a potential for impairment of fish reproduction. Finally, Tier IV field or mesocosm studies may be invoked if the expected concentrations in the environment exceed the Tier III NOEC determination. Tier IV field testing with avian or mammalian species may also be required. It is important to note that recent guidance by the USEPA suggests that decisions on the reregistration of existing pesticides and herbicides can be made without requiring Tier IV studies (Fisher, 1992). However, companies wishing to reregister existing pesticides and herbicides where a hazard for aquatic organisms has been presumed may want to conduct Tier IV testing in an effort to rebut this presumption. Risk Characterization Phase By comparing the estimates of environment concentrations developed in the exposure characterization section and the estimates of toxicity developed in the ecological effects characterization section, an assessment of risk of the rice herbicide to the adjacent wetlands can be generated. There are several methods for examining the relationship between the EEC and environmental effects (Barnthouse et al., 1982a, b). One of these is the quotient method of risk analysis. The Quotient Method for Risk Characterization The quotient method is often employed for predictive risk characterizations of single compounds such as pesticides and herbicides. This method is based on a simple comparison of the estimated exposure to a stressor response value (SRV) for a specific endpoint. The SRV represents a level of exposure that has been demonstrated to have unacceptable toxicological effects. When the estimated exposure ©2001 CRC Press LLC

derived from the exposure assessment exceeds the SRV, the associated risk is considered unacceptable. Clearly the choice of the SRV is critical to this approach. Usually SRVs are derived from laboratory toxicity studies and are estimates of threshold toxicity for appropriate endpoints, such as 1/10 the acute median lethal concentration (LC50) or the chronic NOEC. The use of the quotient method is straightforward and is useful for single chemicals. However, there are a number of disadvantages to this approach. The estimate of risk from the quotient method is only as good as the hazard and exposure estimates. The latter is often based on cursory environmental fate information and has a low confidence level. The hazard assessment data for ecological endpoints are often based on acute effects (i.e., mortality) rather than subacute effects such as inhibition of growth rates or impaired reproduction. Moreover, these so called aquatic hazard assessments are usually based on clean water laboratory tests which may not accurately reflect in situ exposure conditions. Finally, and most importantly, the quotient method is not useful in estimating risk from multiple stressors. Most ecological risk assessments, especially retrospective assessments, involve more than one stressor. Risk from multiple stressors cannot be accurately addressed by adding the stressor–response values due to stressor interactions (e.g., synergism, antagonism, potentiation). In these cases, the hazard assessment and the exposure assessment phases of the ERA must be modified and tailored to the suite of stressors. This assumes that the suite of stressors is known or identified in the preliminary characterization carried out as of the conceptual framework. The Ecological Effects Branch of the USEPA uses a similar methodology. However, where the quotient method uses the resultant quotient in a relative ranking to indicate adverse effects, the Ecological Effects Branch compares the EEC and an effect level (e.g., LC50) to regulatory risk criteria. Many of the risk criteria summarized in Table 1, incorporate safety factors derived from a toxicological model presented in the FIFRA regulations. These safety factors are included to allow for the differential variability and sensitivity among resident fish and wildlife species (Urban and Cook, 1986). Based on the results of the risk characterization, a determination is made as to whether to register the herbicide for use on rice, restrict its use (through required labeling) to certified applicators, restrict its use to areas with limited wetland habitat proximal to the fields, or deny registration of the herbicide (World Wildlife Fund, 1992). It should be noted that recent debate has centered on the development and use of probabilistic ecological risk assessment techniques. A cooperative effort between industry, environmental advocacy groups, and the USEPA, referred to as ECOFRAM, has prepared draft documentation of various probabilistic ERA tools (ECOFRAM, 1999a, b). The actual use of these tools by pesticide risk assessors remains to be seen. Retrospective Ecological Assessments An example involving an ecological assessment of wetlands adjacent to a hazardous waste site will demonstrate how the principles of ecological risk assessment ©2001 CRC Press LLC

Table 1

The Regulatory Risk Criteria* No Risk

Acute toxicity

Chronic toxicity

Risk Limited by Restricted Use

Unacceptable Risk Non-Endangered Species EndangeredSpecies

Mammals EEC 1/5 LC50

EEC > LC50

Birds EEC LC50

Aquatic EEC 1/2 LC50

EEC < chronic no effect level

NA

EEC > chronic effect levels

EEC or EEC EEC or EEC EEC or EEC EEC

> 1/10LC50 > 1/5 LC10 > 1/10LC50 > 1/5 LC10 > 1/20LC50 >1/10 LC10 > chronic effect levels

* The criteria listed in this table can be compared to the estimated environmental concentration (EEC) and the effect level to characterize risk. Adapted from Urban and Cook, (1986).

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are applied in a site-specific retrospective study. The principles developed in the USEPA framework and guideline documents (USEPA, 1992, 1998) can be applied to retrospective risk assessments as well as, and in some cases better than, the predictive ERAs (Pascoe et al., 1993). In addition, several other USEPA documents provide guidance applicable to ecological assessments at hazardous waste sites (USEPA, 1988, 1989a; Warren-Hicks et al., 1989). Finally, guidance specific to the Superfund program was issued in 1989 (USEPA, 1989b, c). This guidance, commonly referred to as RAGS (for Risk Assessment Guidance for Superfund), has become the main source of procedural information for both human and environmental health risk assessments. Problem Formulation For the purposes of this example, consider a hazardous waste site that is a longterm chemical manufacturing facility. The facility is located on 2 ha of commercial land that includes a nontidal marsh. In addition, a levee separates the site from an adjacent tidal marsh and a large estuarine bay. The primary objective is to assess the degree of risk that contamination from the site poses for the adjacent wetland ecosystem. Another objective of this sitespecific assessment is to provide information for the development and evaluation of an ecologically sound remedial action plan for the site. In the case of the nontidal wetland, these data can be used to help define the strategy required to address any effects of the contamination from the site on the nontidal wetland ecosystem. For the tidal marsh, these data would provide a basis for determining if contaminantrelated impacts are occurring and whether remediation is necessary. Moreover, if required, these data could be used to determine the appropriate type and extent of remediation. The following specific goals can be established for performing an assessment of risk to the tidal wetlands. • Assess the potential for exposure to the contaminants of concern for ecologically and socially important species from the tidal wetland. • Determine the potential for acute and chronic toxicological effects from accumulated contaminants on these species. • Evaluate the structure of the populations and communities of plants, benthic invertebrates, and small mammals in the tidal wetland. • Relate the exposure, bioaccumulation, and toxicity data obtained in this assessment with population, community, and ecosystem data to provide an integrated picture of the impact of contamination from the site on the adjacent tidal wetland. • Evaluate the source or sources of any stressors that may be identified by examining the distributional patterns of those stressors in the tidal wetland in relation to identified sources including the site and the bay. That is, do the stressors come from the site or from the other sources via the connection to the bay?

Exposure Characterization The retrospective approach to exposure characterization is usually more direct than in the predictive approach. In a retrospective ERA, stressors are already present ©2001 CRC Press LLC

in the environment. The goal is to determine which stressors are present and in what concentrations, rather than attempting to predict which stressors might enter the environment. The potential for exposure can be determined by measuring the actual accumulated dose of stressor in environmental matrices, including the tissues of native or transplanted organisms. The first step in the exposure characterization phase is to determine the nature and distribution of the stressors. As indicated above, this is accomplished by collecting samples of the various environmental matrices and analyzing them for the suspected contaminants. In our waste site example, sampling should be performed within the site proper, in the on-site nontidal marsh, and at near field and far field locations within the adjacent tidal wetland. Additionally, external reference sites should be sampled to provide a comparison to the zones within the site and adjacent wetlands (Figure 5).

Figure 5

For hazardous waste sites, initial sampling commonly requires extra protection for team personnel.

The precise sampling strategy employed will be dependent upon the exact goals of the ecological assessment, the structure and function of the ecosystem involved, and the nature of the contaminants of concern. One approach is a random sampling strategy that is designed to evaluate differences over relatively broad areas of the wetland. A grid system comprised of relatively large quadrats is set up within each zone to be examined. The use of randomly selected sampling stations within each grid facilitates statistical comparisons between zones, and the number of samples to be taken per unit area can be calculated (Warren-Hicks et al., 1989). Statistical comparisons of data from the nontidal marsh, the near field tidal marsh and far field tidal marsh grids allow for spatial comparisons of the different measurement endpoints ©2001 CRC Press LLC

relative to their distances from the site. This provides a basis for an internally controlled evaluation of the spatial effects of the site. Data on these three zones can also be compared to the remote reference sites. This strategy provides two independent methods for evaluating the broad scale effects of the site on the adjacent wetlands. A second approach for evaluating the impact of the site on the adjacent tidal wetlands can be used in parallel with the random sampling approach described above. In this approach, the sampling focus is on the potential route by which onsite stressors may be transported off-site and into the tidal wetland. In this nonrandom sampling strategy, a series of stations can be established in the major drainage slough to provide a spatial gradient of contamination moving from the site to the bay. The types of sample measurements that are commonly taken at stations such as those described should include concentrations of contaminants in the sediments, soils, and surface waters. Measurements of actual concentrations of the stressors in plants and animals are also helpful. The combined database from all of the sites will provide the basis for interpreting the extent and magnitude of the stressors on the site and in the adjacent wetlands. The exposure characterization discussion thus far has been limited to determining the nature and distribution of stressors on the site and neighboring wetlands. An equally important component of this phase of the ERA is to evaluate the potential routes by which organisms may be exposed to the stressor. In the example, routes of potential exposure include direct exposure in the sediments and surface waters and indirect exposure via the food chain. Soils and sediments commonly serve as sinks for chemical stressors and, thus, will need to be examined closely. The bioavailability of stressors sequestered in the sediments should also be evaluated because this could have a profound effect on the actual ecological effects observed. The degree to which the stressor is bioaccumulated or depurated is also very important to the evaluation of actual exposure. Finally, the temporal aspects of the stressors must be considered. For example, the standing waters of the nontidal wetland would likely recede during the dry season. Marsh plants commonly exhibit a rapid growth period at the end of the wet season and continue this growth into the dry season. As available water diminishes, plant growth subsides, and eventually the plants senesce and die. These and other seasonal changes could modify stressor distribution and bioavailability. Therefore, it is often necessary to perform the ecological assessments during both the growing and nongrowing seasons. Some inherent complicating factors must be addressed in order to establish cause–effect relationships between stressors in the environment and observed changes in biological endpoints. For example, in highly commercial areas, extensive anthropogenic activities can result in numerous stressors from multiple sources. Moreover, these stressors may have differential mobility within ecosystems and the ratios of one stressor to another may change with distance from the source. This complexity makes it difficult to relate any observed biological effects to specific stressors or to a specific source. The spatial and temporal sampling schemes described above should provide a basis for determining if the concentrations of chemical contaminants and biological parameters in the wetlands show significant correlation with one another and specific relationships to the site. ©2001 CRC Press LLC

Ecological Effects Characterization To accomplish the retrospective ecological assessment, it is necessary to determine if the stressors from the site have produced adverse effects on the structure and function of the adjacent wetland ecosystems. In the predictive ERA example, the emphasis was on performing laboratory studies to determine potential ecological effects, followed by increasingly higher tier modeling or field dissipation studies if warranted by the laboratory data. Conversely, with the retrospective ERA, the emphasis is on determining actual ecological effects on the site and in the adjacent wetlands. This is accomplished by emphasizing field evaluations of community structure and the toxicity of soils, sediments, and surface waters. It is important to co-ordinate the ecological effects sampling with the exposure sample collection on both a spatial and a temporal scale. In other words, samples for chemical analysis and ecological effects should be taken at the same time and place whenever possible. Examples of the types of ecological effects sample measurements commonly taken are provided in Table 2. Table 2

Retrospective Ecological Risk Assessment (ERA) Emphasizing Field Evaluations*

Bioaccumulation of contaminants in dominant plant species Bioaccumulation of contaminants in dominant benthic invertebrate species Bioaccumulation of contaminants in small mammals Bioaccumulation of contaminants in dominant fish species Sediment toxicity (pore water, elutriates, or solid phases) Water toxicity Plant population and community structure within each grid Benthic infaunal invertebrate population and community structure Histopathological evaluation of native small mammals or fish Fish population survey * Typical ecological effects sample measurements are listed in the table and should be collected at the same time and place as the chemical measurements.

As can be seen in Table 2, the emphasis is on measuring actual effects in the field whenever possible. Nevertheless, laboratory toxicity tests can play an important role in the evaluation of wetlands water and sediments affected by hazardous waste sites (Zimmer et al., 1988; Durda, 1993; Woodward et al., 1988; see Figure 6). The toxicity of field collected samples is used to define the extent of contamination with regard to direct biological significance. On-site studies such as cage studies may also provide valuable information in the delineation of ecological effects. Measurement endpoints should be chosen for relevant characteristics. Endpoints may be key ecosystem components that are sensitive to stressors, accumulate the stressor of concern, or are food resources for higher trophic levels. Other endpoints include key indicators of ecosystem structure and function or surrogates for stressor impacts to endangered species or other assessment endpoints that cannot be directly monitored. Species continuously exposed to contaminated sediment or water have the greatest potential for contaminant uptake and serve as excellent indicators of ©2001 CRC Press LLC

Figure 6

Ecological effects characterization emphasizes measurement of actual effects in the field. Nevertheless, laboratory toxicity tests, such as this sediment test, can play an important role.

bioaccumulation potential. In wetland sites, marsh plants, benthic invertebrates, and filter-feeding mollusks are particularly useful indicators of bioavailability. Transfer of contaminants along the food chain should also be considered. The biological effects of accumulated contaminants are also complex. Contaminants may be stored in an inactive form, mobilized, or excreted. Alternatively, they can interact with cellular macromolecules causing metabolic perturbations and cellular damage that impact the function of the organism. Thus, while bioaccumulation of contaminants reflects the potential for toxicity, supplemental approaches with greater resolution are needed to determine actual toxicity to the organism. Sublethal impacts on the organism can affect higher levels of biological organization through inhibition of such physiological processes as growth and reproduction. Changes in these parameters can impinge upon populations and communities. In addition, if the marsh plant populations are adversely affected, the ecological impact could result in habitat disturbance for birds and small mammals, including endangered species. As in the previous cases, impact at these higher levels of organization can be inferred or measured directly. Risk Characterization Comparison of the exposure and ecological effects characterizations is performed to provide an assessment of risk to both the on-site nontidal wetland and the tidal estuarine wetland adjacent to the site. Because of the complexity of the linkages ©2001 CRC Press LLC

between contaminant distribution and ecosystem effects, the ecological assessment must be designed to carefully examine the correlations between the concentrations of contaminants in sediment, soil and water, and biological function observed at any of several levels of biological organization. The collection of matched samples for measuring contaminant concentrations in the sediments, bioaccumulation, toxicity, and population and community status is important to allow a rigorous examination of the relationships among these endpoints. The specific procedures used in the risk characterization phase will depend on the design chosen to determine exposure and ecological effects, as well as the overall purpose for performing the ERA. In the retrospective example, the primary objective is to assess the degree of risk that contamination from a hazardous waste site will have on the on-site and adjacent wetland ecosystems. Data collected includes chemical concentrations in the sediments and surface waters, bioaccumulation of the stressors in dominant plant and animal species, sediment and surface water toxicity, fish and wildlife population surveys, and plant and benthic invertebrate population and community structure. Initially, the stressor concentration data can be tabulated and plotted on both the spatial and temporal scales to evaluate trends in concentration. Are the concentrations higher near the site and lower away from the site? Do concentrations appear to follow a gradient associated with one or more transport routes away from the site? Has the stressor partitioned into the sediments or has it been dispersed for some distance? Furthermore, how do these concentrations relate to water quality standards and criteria established for wetlands (USEPA, 1990)? Similarly, the sediment and surface water toxicity data should be tabulated and plotted to show spatial and temporal trends. Are any trends evident and, if so, do they correlate with the stressor concentration trends? Bioavailability can be established by comparison of contaminant levels in the abiotic matrices and biota exposed to these matrices. Ecological effects on the population and community level can be evaluated using several statistical procedures designed for these types of data (Ludwig and Reynolds, 1988). Species diversity, richness, and abundance can be compared using ANOVA and cluster analyses to determine if differences in these parameters can be correlated with stressor concentration. A picture should begin to emerge regarding the levels of exposure and the degree of effect after comparing all of the collected data. If the picture is sufficiently muddled because of extensive anthropogenic inputs into the systems under investigation, techniques such as the principal components analysis may be used to help define the key contaminants associated with specific effects. In addition, fate and transport processes can often be modeled based on data collected in the field and the physicochemical characteristics of the stressors of concern. Ultimately, the actual assignment of ecological risk will often depend on the professional judgment of the risk assessor.

SUMMARY ERA is a rapidly developing and increasingly important discipline that can be used as a management tool in the protection of wetlands. It is a process that estimates ©2001 CRC Press LLC

the impact of anthropogenic activities on ecosystems. The results of ERAs are used by managers and regulators to determine the appropriate action for the wetland or site in question. The application of the ERA paradigm as described provides a framework for assessing risk to ecological resources for a variety of scientific and regulatory purposes. The principles of ERA can be applied to a variety of situations including retroactive analysis of anthropogenic activities and predictive assessment of future actions. Additional definition of the principles and procedures of ERA is needed to help risk assessors standardize their methods as much as possible. With this objective in mind, the American Society for Testing and Materials (ASTM) has undertaken an effort to develop standards for the assessment and valuation of wetlands (Ethier, 1993). The use of a standard approach with standard terminology will facilitate communication and interpretation of results and concepts. Furthermore, this standard approach will allow for balanced risk management of human health and environmental concerns under a variety of regulatory programs. However, it is important to understand that even with increased standardization of the ERA framework, ERAs will still need to be designed on a site- or chemical-specific basis to address the specific concerns.

REFERENCES Adamus, P. R., Clairain, Jr., E. J., Smith, R. D., and Young, R. E., Wetland Evaluation Technique (WET), Vol. II, Methodology, Department of the Army, Waterways Experiment Station, Vicksburg, MS, 1987. Barnthouse. L. W., Bartell, S. M., DeAngelis, D. L., Gardner, R. H., O’Neill, R. V., Powers, D. D., Suter, G. W., Thompson, G. P., and Vaughan, D. S., Preliminary environmental risk analysis for indirect coal liquefaction, Draft Report, Oak Ridge National Laboratory, Oak Ridge, TN, 1982a. Barnthouse, L. W., DeAngelis, D. L., Gardner, R. H., O’Neill, R. V., Powers, C. D., Suter, G. W., and Vaughan, D. S., Methodology for Environmental Risk Analysis, ORNL/TM 8167, Oak Ridge National Laboratory, Oak Ridge, TN, 1982b. Barnthouse, L. W., Suter, G. W., Bartell, S. M., Beauchamp, J. J., Gardner, R. H., Linder, E., and Rosen, A. E., User’s Manual for Ecological Risk Assessment, Publication Number 2679, ORNL-6251, Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, TN, 1986. Bartell, S. M., Gardner, R. H., and O’Neill, R. V., Ecological Risk Estimation, Lewis Publishers, Chelsea, MI, 1992. Burmaster, D. E. and Anderson, P. D., Principles of good practice for the use of Monte Carlo techniques in human health and ecological risk assessments, Risk Anal., 14, 477, 1994. Burns, L. A., Cline, D. M., and Lassiter, R. R., Exposure Analysis Modeling System (EXAMS): User Manual and System Documention, EPA/600/3-82-023, 1982. Cairns, Jr., J., Niederlehner, B. R., and Orvos, D. R., Eds., Predicting Ecosystem Risk, Princeton Scientific, Princeton, NJ, 1992. Durda, J. L., Ecological risk assessments under Superfund, Water Environ. Technol., 45, 42, 1993. ECOFRAM, Aquatic draft report, The Ecological Committee on FIFRA Risk Assessment Methods, May 10, 1999. ©2001 CRC Press LLC

ECOFRAM, Terrestrial draft report, The Ecological Committee on FIFRA Risk Assessment Methods, May 10, 1999. Ethier, W. H., New wetlands standards: providing great tools for policymakers, Stand. News, 21, 26, 1993. Fisher, L., Memo re: decisions on the biological, fate and effects task force, Program Guidance on Ecological Risk Management, 1992. Lewis, M. A., Mayer, F. L., Powell, R. L., Nelson, M. K., Klaine, S. J., Henry, M. G., and Dickson, G. W., Ecotoxicology and Risk Assessment for Wetlands, SETAC Press, Pensacola, FL, 1999. Ludwig, J. A. and Reynolds, J. F., Statistical Ecology, John Wiley & Sons, New York, 1988. McKim, J. M., Bradbury, S. P., and Niemi, G. J., Fish acute toxicity syndromes and their use in the QSAR approach to hazard assessment, Environ. Health Perspect., 71, 171, 1987. Mitsch, W. J. and Gosselink, J. G., Wetlands, Van Nostrand Reinhold, New York, 1986. Moore, D. R. J., Perspective: using Monte Carlo analysis to quantify uncertainty in ecological risk assessment: are we gilding the lily or bronzing the dandelion? Hum. Ecol. Risk Assess., 2, 628, 1996. National Research Council, Risk Assessment in the Federal Government: Managing the Process, National Research Council, National Academy Press, Washington, D.C., 1983. Newman, M. C. and Strojan, C., Risk Assessment: Logic and Measurement, Lewis Publishers, Chelsea, MI, 1998. Pascoe, G. A., Blanchet, R. J., Linder, G., and Ingersoll, C. G., Assessment of ecological risks of contaminated wetland, a Superfund case study, Presented at the 1st SETAC World Congress, Lisbon, Portugal, March 1993. Power, M., Probability concepts in ecological risk assessment, Hum. Ecol. Risk Assess., 2, 650, 1996. Superfund Amendments and Reauthorization Act, Fed. Regist., 52, 13378, April 22, 1987. Suter, G. W., Ecological Risk Assessment, Lewis Publishers, Chelsea, MI, 1993. Urban, D. J. and Cook, N. J., Hazard Evaluation, Standard Evaluation Procedure, Ecological Risk Assessment, EPA 540/9-85-001, 1986. U.S. Environmental Protection Agency, Pesticide Assessment Guidelines, Subdivision E, Hazard Evaluation: Wildlife and Aquatic Organisms, EPA 540/9-82-024, 1982a. U.S. Environmental Protection Agency, Pesticide Assessment Guidelines, Subdivision N, Chemistry: Environmental Fate, EPA 540/9-82-021, 1982b. U.S. Environmental Protection Agency, Guidelines for carcinogen risk assessment, Fed. Regist., 51, 33992, September 24, 1986a. U.S. Environmental Protection Agency, Guidelines for mutagenicity risk assessment, Fed. Regist., 51, 34006, September 24, 1986b. U.S. Environmental Protection Agency, Guidelines for health risk assessment of chemical mixtures, Fed. Regist., 51, 34014, September 24, 1986c. U.S. Environmental Protection Agency, Guidelines for health assessment of suspect developmental toxicants, Fed. Regist., 51, 34028, September 24, 1986d. U.S. Environmental Protection Agency, Guidelines for exposure assessment, Fed. Regist., 51, 34042, September 24, 1986e. U.S. Environmental Protection Agency, Review of Ecological Risk Assessment Methods, EPW23O-10-88-O41, 1988. U.S. Environmental Protection Agency, Rapid Bioassessment Protocols for Use in Streams and Rivers: Benthic Macroinvertebrates and Fish, EPA/444/4-89-001, 1989a. U.S. Environmental Protection Agency, Risk Assessment Guidance for Superfund, Vol. I, Human Health Evaluation Manual, Part A, EPA/540/l-89/002, 1989b.

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U.S. Environmental Protection Agency, Risk Assessment Guidance for Superfund. Vol. 11, Environmental Evaluation Manual, 1989c. U.S. Environmental Protection Agency, Water Quality Standards for Wetlands: National Guidance, EPA 440/S-90-011, 1990. U.S. Environmental Protection Agency, Summary Report on Issues in Ecological Risk Assessment, EPA/625/3-91/018, 1991. U.S. Environmental Protection Agency, Framework for Ecological Risk Assessment. Risk Assessment Forum. EPA/630/R-92/001, 1992. U.S. Environmental Protection Agency, Guidelines for Ecological Risk Assessment, Risk Assessment Forum, EPA/630/R-95/002F, 1998. Warren-Hicks, W., Parkhurst, B. J., and Baker, Jr., S. S., Ecological Assessment of Hazardous Waste Sites: a Field and Laboratory Reference Manual, EPA/600/3/89/013, 1989. Woodward, D. F., Snyder-Conn, E., Riley, R. G., and Garland, T. T., Drilling fluids and the arctic tundra of Alaska, U.S.A.: assessing contamination of wetlands habitat and the toxicity to aquatic invertebrates and fish, Arch. Environ. Contam. Toxicol., 17, 683, 1988. Zimmer, R. D., Buchanun, G., Charters, D., Ferretti, J., and Kent, D. J., Aquatic toxicological evaluation of soils collected from a mid-western Superfund site, presented at the 61st Annual Conference of the Water Pollution Control Federation, Dallas, TX, 1988. World Wildlife Fund, Improving aquatic risk assessment under FIFRA: report of the Aquatic Effects Dialogue Group, Washington, D.C., 1992.

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Kent, Donald M. et al “Avoiding and Minimizing Impacts to Wetlands” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

5

Avoiding and Minimizing Impacts to Wetlands Donald M. Kent and Kevin McManus

CONTENTS Planning Design and Construction Design Construction Erosion and Sedimentation Nitrogen Loading Planning Guidelines Estimating Nitrogen Loads Stormwater Runoff Planning and Nonstructural Practices Structural BMPs Pretreatment Detention Basins/Retention Ponds Vegetated Treatment Infiltration Filtration References

Recent estimates of the extent of global wetlands range from 5 to 8.6 million ha (Mitsch, 1995). Increasing evidence suggests that the historic extent of global wetlands was substantially greater. For example, in Japan, 45 percent of tidal flats have been destroyed since 1945 (Hollis and Bedding, 1994). Northern Greece has lost

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94 percent of its marshland since 1930. In the conterminous United States, an estimated 47 million ha of wetlands have been lost over the last 200 years — an average rate of 235,000 ha per year (U.S. Office of Technology Assessment, 1984; Dahl, 1990; Hollis and Bedding, 1994). This rate of loss appears to have decreased dramatically in recent years, to about 32,000 ha per year, coincident with recognition of the importance of wetlands and a “no net loss” government policy (Heimlich and Melanson, 1995). Wetland losses are attributed to filling and draining, primarily in support of development and agricultural activities. An unknown number of wetlands, not filled or drained, have been otherwise impacted by changes in watersheds or adjacent land uses. Alterations to wetland plant communities lead to increased erosion and sedimentation. Construction of buildings, parking lots, and other impervious surfaces increases the quantity and decreases the quality of surface runoff to wetlands. Septic systems and fertilizers increase the concentration of nitrogen in groundwater flow to wetlands. Activities adjacent to wetlands can disturb wildlife. Wetland impacts, both direct and indirect, can be avoided or minimized by appropriate planning, design, and construction. In this chapter, planning is discussed as a means for avoiding or minimizing direct impacts to wetlands. Design and construction techniques are discussed as a means to avoid or minimize indirect impacts to wetlands. Discussed in some detail are three design and construction issues. They are erosion and sedimentation, nitrogen loading, and stormwater.

PLANNING Planning to avoid or minimize direct impacts to wetlands is fundamentally a three-step process. The first step is to identify the wetland resource. Discussed in detail in Chapter 2, this step requires applying hydrology, soils, and vegetation criteria to undeveloped areas. For large areas, off-site resource identification is an effective and appropriate approach for preliminary planning. Greater resource resolution, typically requiring on-site identification, is more appropriate for smaller areas and for detailed planning. Characterization and classification (e.g., palustrine forested wetland, emergent marsh; see Chapter 1) of wetland resources are also helpful at this stage. The second step in effective planning is to assign functions and values to identified wetland resources. Common techniques for determining functions and values include professional opinion, the use of indicators, direct measurement, and economic analysis (see Chapter 3). As with resource identification, off-site and less detailed approaches are most appropriate for large areas during preliminary planning, whereas on-site assessments are most appropriate for small areas and detailed planning. Assigning functions and values will facilitate prioritization in the event that not all resource areas can be preserved and reveal functions and values that need to be protected or replaced during construction and operation. Finally, wetlands identified and evaluated for functions and values are incorporated into a site selection process. Site selection typically includes identification of

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several alternative sites and development of site selection criteria. Alternative sites satisfy minimal, implicit criteria such as availability and location. At a minimum, the site selection process should consider the criteria listed in Table 1 (McManus, 1994). Direct and indirect impacts to wetland and other environmental resources should be identified. Other environmental resources include fish and wildlife, navigation channels, and recreation areas. Projects not dependent upon access to water should be sited elsewhere. The minimum size required to satisfy the project purpose should be determined and project configuration and layout evaluated to further reduce project size. Constructability refers to project topographic, slope, soil, and backfill requirements. Extensive grading, blasting, or filling are typically associated with environmental impacts and should be avoided. Proximity to supporting infrastructure, such as utilities and roadways, affects project size, configuration and layout, and cost. Cost prohibitive sites should be eliminated; thereafter, the costs of development should be weighed against the costs of environmental impacts. The opportunity for successfully satisfying the requirements of various international, national, regional, and local entities such as regulatory agencies and lending institutions should also be evaluated. Table 1

Representative Site Selection Criteria (Adapted from McManus, 1994)

Wetland impacts Other environmental impacts Water dependency Site size Constructability Supporting infrastructure Costs Regulatory/institutional issues

Larger and more complex projects will require a more detailed site selection process. In the United States, the National Environmental Policy Act (U.S. Congress/NEPA, 1978) provides guidance as to appropriate criteria for evaluating project impacts to wetlands and other environmental resources. In addition to environmental impacts, this approach considers impacts to human uses and the technical, economic, and institutional feasibility and merits of the site. Table 2 represents a hypothetical site-screening matrix consistent with the NEPA (McManus, 1994). In the example, Site 1 is technically and economically feasible, but will likely impact the environment and human use of the site, and is not publicly acceptable. Site 2 has no significant environmental, human use, or institutional constraints but has technical and economical issues. Site 3 is the preferred site, having no significant environmental or human use impacts, being technically and economically feasible and acceptable to the public.

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Table 2

A Hypothetical Site Selection Matrix (Adapted from McManus, 1994) ScreeningCriteria

Site 1

Site 2

Site 3

0 0 0 – – –

0 0 0 0 0 0

0 0 0 0 0 0

– – 0 0 –

0 0 0 0 0

0 0 0 0 0

0 – – – – 0 – 0 0 0 0

0 0 0 0 0 0 0 0 0 0 0

0 0 0 0 0 0 0 0 0 0 0

+ +

– – –

+ + +

+ + 0

– – –

+ + 0

– 0

0 0

+ 0

Environmental Aquatic Ecosystem Substrate Water quality Water circulation Normal water fluctuations Threatened and endangered species Other aquatic organisms and wildlife Special Aquatic Sites Sanctuaries/refuges Wetlands Mudflats Vegetated shallows Riffle and pool complexes Human Uses Water supplies Recreational and commercial fisheries Water–related recreation Aesthetics Parks, preserves, wilderness areas Archaeological or historical sites Compatibility with adjacent land uses Potential noise impacts Potential odor impacts Public health Traffic increase Technical Suitable foundation/soils conditions Adequate land area Access to existing roads and utilities Economic Land acquisition Operation and maintenance Capital cost—construction Institutional Public acceptance Compliance with existing regulations

Note: + indicates an expected positive impact; – is an expected negative impact; 0 is an insignificant or no impact.

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DESIGN AND CONSTRUCTION Design Once the site selection process has been completed, the focus can shift to design details, site layouts, construction methods, and other specific engineering requirements to minimize unavoidable wetland impacts. A reasoned assessment of the minimum economically and functionally viable size for a proposed structure(s) should be made, particularly if the project is not water dependent. Even for water dependent projects, such as marinas or dredging projects, project scope should be evaluated with an eye toward minimizing wetland impacts. The project should have an accurate wetland delineation line depicted on site plans to facilitate evaluation of layout options. For projects that may involve clearing of trees and other existing vegetation, care should be taken to minimize the limits of clearing to the minimum acreage needed for the project. Maintenance of existing vegetative buffers, particularly within wetland areas, is not only a valuable means of providing a visual and auditory buffer for the facility, but it also may reduce overall facility wetland impacts. This is particularly true along active coastal shorelines, such as eroding bluffs, beaches, and dune environments. The orientation and layout of a project are generally a function of its intended purpose and use. Many projects, such as railways, roads, and retaining walls, being linear features, have limited flexibility with regard to basic configuration. However, their actual alignment, relative to wetland areas, can often be optimized to reduce impacts to insignificant levels. Similarly, layouts of buildings and ancillary structures such as garages, walkways, and decks can be adjusted to minimize direct wetland impacts. Specific design details for a project can also be important factors in reducing wetland impacts. For example, use of the maximum safe slopes for site preparation will minimize incursions into wetland areas. Maximum safe slopes can be achieved using vertical retaining walls, cellular confinement, sheet piling, or gabion rock walls. Backfill and other construction materials should ensure good drainage and scour protection (Nelson, 1995). Another method for minimizing impacts is to use boardwalks supported by posts or post-like anchors. Waterway crossings offer another opportunity to minimize wetland impacts. Typically, culverts are used when crossing small waterways. Culverts should be designed to pass expected flows (e.g., 100-year flood event), and to avoid changes to flow velocity and increased erosion and scour. Bridges can minimize impacts to larger waterways, especially if construction is accomplished in midair using a crawler crane. Construction For many projects, such as subsurface water, sewage, and other utility pipelines, the primary impacts to wetlands occur during construction. The use of temporary

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access materials, specialized construction equipment, and the placement of staging areas can all affect the level of wetland impacts. Temporary pile-supported construction trestles can be used to significantly reduce direct wetland impacts through ecologically sensitive wetland areas such as estuarine and fresh-water marshes, beach or dune environments, and peat bogs (Figure 1). These trestles can be located either directly above, or directly adjacent to, the work area. Equipment can be brought to the work area using rail-mounted transport platforms, and the trestle can be constructed in stages to accommodate the construction schedule. Trestles provide a stable temporary work platform that directly impacts little wetland acreage.

PLAN VIEW pipeline wood planking upland

wetland

sheet piling construction trench PROFILE

wood planking sheet piling

pipeline

Figure 1

Temporary pile-supported construction trestles can be used to significantly reduce wetland impacts. The trestles may be located either directly above or immediately adjacent to the work area.

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Another effective construction technique uses steel sheet piling to isolate the active work area, and temporary wood decking placed directly on top of the sheet piling. This allows construction equipment to access the work areas without compacting wetland soils. Compacted wetland soils lose their original productivity and hydrologic functions. For smaller projects that may not warrant the use of sheet piling, geotextile fabric, clean granular material, and wood decking can be placed within the project alignment. Sheet piling can also be used in intertidal or shallow freshwater areas. Combined with siltation curtains, piling can prevent the release of sediment-laden water to surrounding wetlands and waterways. The use of barge-mounted equipment can also be used in intertidal and shallow freshwater areas to access sensitive sites. Work barges can be floated into place on rising tides, and grounded out to provide suitable access with minimal or no long-term impacts. For construction of trenches in wetland areas, utility workers have developed specialized, tracked, trenching vehicles that can operate on soft, unstable soils. The vehicles work directly within the project alignment. Wide, low-pressure tires on vehicles that distribute loads across wetland soils and vegetation also reduce vehicle impacts. For dredging within wetlands, waterways, and waterbodies, clamshell dredge equipment fitted with covers and watertight buckets minimizes sediment washout and turbidity. Large construction projects typically require staging areas. Staging areas should be located outside wetlands and their designated buffer zones and should be paved to minimize erosion and groundwater impacts. Also, staging areas should include stormwater management systems designed to trap suspended sediments and to contain accidental releases of fuel oil, lubricants, and other potentially hazardous releases from equipment. Scheduling can minimize temporary, construction-related impacts. As a general rule, wetland work in temperate climates should be scheduled during winter and early spring when plants are dormant and the soils are frozen or well consolidated. Soil compaction is minimized, and site cleanup and rehabilitation during the coming peak growing season are facilitated. Other seasonal restrictions are often applied for work within coastal environments based upon the expected presence of commercially and recreationally important fish and wildlife species. Species susceptible to illtimed construction include spawning and migrating anadromous fish and shrimp, overwintering groundfish, and migratory waterfowl. Another method for minimizing the impacts of construction within wetlands is proper work sequencing. For example, minimizing the extent of clearing in front of the active trenching operation will reduce the potential for soil erosion into adjacent wetlands and reduce impacts to wildlife using the existing vegetative cover. Wherever possible, work that is required within wetland areas should be completed as quickly as possible, without excessive delays between the initial disturbance and rehabilitation. Trenching should be conducted as a single, continuous operation, involving clearing, installation, backfilling, and soil restoration. An open trench can act as a channel to dewater adjacent wetland areas and increase erosion and runoff impacts.

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EROSION AND SEDIMENTATION Sedimentation of wetlands can be avoided or minimized by preventing soil erosion and controlling already eroded sediments. There are numerous methods for erosion and sedimentation control, all of which seek to isolate and contain, to the maximum extent possible, sediment-laden runoff generated during project construction activities. The performance of these various methods in the field varies considerably depending upon the type of soils, water flows, exposure, and other site specific factors. Figure 2 summarizes some of the more popular sedimentation control methods. Critical elements of effective erosion and sediment control plans are listed in Table 3 (Brown and Caraco, 1997). Erosion and sedimentation control methods can be used singly or in combination. By limiting the amount of incremental and total land clearing, and maintaining existing ground cover to the maximum extent possible, potential runoff, gully creation, rutting, and airborne dust formation can be reduced to acceptable levels. Cleared land produces as much as 2000 times more sediment than uncleared land (Paterson et al., 1993). Where feasible, a project site layout should take advantage of existing vegetation between the clearing limits and adjacent wetlands. Buffers of at least 25 m in width are the most effective in filtering sediment from construction site runoff (Woodward, 1989). Vegetative buffers should also be preserved for projects with shoreline frontage to protect structures from wave and flooding impacts. Installation of hay bales within shallow cut-off trenches upgradient of wetland areas can be an effective and inexpensive perimeter control method. Bales should be staked to the ground, without gaps between bales. Bales should be routinely monitored, and bales damaged, moved, or destroyed during construction should be repaired. Construction specifications should provide for regular checks of the condition and effectiveness of the hay bale protection systems. Geotextile siltation fences can be wrapped around hay bales and staked into the ground to provide an extra measure of protection against the release of fine-grained materials. Siltation fence efficiency ranges from 35 to 86 percent depending upon site conditions (Horner et al., 1990; W&H Pacific and CH2M-Hill, 1993). Siltation curtains can also be used effectively in both wetlands and open water environments. Curtains can be used to surround subaqueous dredging operations, particularly those occurring within sheet piling, to isolate trench water from the surrounding environment. Curtains with flotation can also be installed around shoreline construction projects and anchored in place to isolate the work area. However, the effectiveness of these structures decreases significantly in areas of strong river currents, tidal flows, and large tidal ranges, particularly if the curtain is installed perpendicular to the current flow. In such cases, the siltation curtain experiences rollover or submergence and is susceptible to damage from debris. Therefore, siltation curtains are most effective in ponds, lakes, and other sheltered water bodies with little or no variation in water height. In any construction project, regardless of the proximity to wetlands or other adjacent sensitive habitats, construction specifications should require prompt stabilization of newly exposed soils, including stockpiled soil. Seeding and sodding are relatively inexpensive, and up to 99 percent effective in reducing erosion (Brown ©2001 CRC Press LLC

Figure 2

Erosion and sedimentation control methods (McManus, 1994). Black indicates the method is suitable for use in the environment; gray indicates the method is suitable with limitations.

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Table 3

Critical Elements of an Erosion and Sediment Control Plan (Adapted from Brown and Caraco, 1997)

Minimize clearing and grading Protect waterways and stabilize drainage ways Phase construction to limit soil exposure Stabilize exposed soils immediately Protect steep slopes and cuts Install site perimeter control to filter sediments Use settlement traps and basins for larger volumes Use experienced contractors to implement the plan Tailor the plan to specific site conditions Assess plan effectiveness after storms

and Caraco, 1997). Seeding is the least expensive option and is appropriate when temporary stabilization is required. Seeds can be broadcast by hand or hydroseeded. The latter is a mixture of seeds, water, fertilizer, lime, and mulch sprayed onto the soil. Sodding is more appropriate for permanently vegetated areas and provides immediate cover and greater resistance to higher flow velocities. The construction schedule should allow time for vegetation to become reestablished prior to the end of the growing season. In cases where this is not possible, more expensive but generally less effective measures, such as mulching or covering exposed areas with erosion control blankets, jute mats, or geotextile mats, should be employed. Mulches, blankets, and mats protect seeds from erosion, dehydration, and animals until the next growing season (Brown and Caraco, 1997). Mulches, consisting of straw, hay, fiber, or wood chips, are effective on flat or gently sloping areas. Erosion control blankets consist of a mulch material held together by a plastic netting, and jute mats are sheets of woven jute fiber. Effective on relatively level ground, both the blankets and the mats are stapled to the ground after seeding and degrade over time. Geotextile mats are more appropriate for steeper slopes and channels. The mats are typically laid on the soil surface and covered with topsoil and seed. As previously discussed, isolation of the active work area in both wetlands and open water areas is an effective method to limit the horizontal extent of disturbance, particularly in areas where significant dredging is required. In such cases, dredging open trenches beyond 1 m in depth requires side slopes which can range from 3:1 to 5:1 or greater, meaning that a 3-m-deep trench would disturb a minimum 20- to 33-m width of sediments. Clearly, this size dredging operation would require the handling and disposal of large amounts of excess dredged material. Conducting this work within sheet pilings allows a vertical sidewall, thereby reducing the volume of material to be handled and isolating the silt-laden trench water from surrounding marsh and other wetland areas. Open dredging within or adjacent to wetland areas can produce significant amounts of turbidity. If typical dredging equipment is used, for example, a bargemounted crane with a clamshell dredge bucket, methods are available which can reduce turbidity. These include establishing requirements that all lifts of a clamshell dredge bucket through the water column are vertical, that dredge buckets be used ©2001 CRC Press LLC

with covers and gasket seals to prevent washout of sediments and suitable filtering of water released from stockpiled dredged material. Hydraulic dredging can also be used in certain unconsolidated sediments to reduce turbidity. With this method, sediments are removed and pumped as a slurry to a settling barge or disposal site. While initial turbidity at the point of dredging is minimal, large amounts of water must be filtered and removed from sediments at the disposal site, and pumping limitations require that disposal occur in close proximity to the point of dredging. Construction will often require temporary stockpiling of soils, and care should be taken to continually spray these piles with water, or cover them, in order to prevent wind erosion and transport of fines. Similarly, newly graded access roads should be frequently sprayed with water or dust suppressants to reduce dust formation. The construction schedule should attempt to minimize the period of time where exposed stockpiles or unpaved road surfaces are required. Site grading and excavation activities in areas already served by drainage systems are a potential concern for sedimentation. Many parking lots, roadways, and other facilities use stormwater drainage systems that discharge directly into adjacent wetland areas. In order to minimize the impacts from run-off of sediment-laden water, existing catch basins and storm drains should be completely ringed with staked haybales and a layer of filter fabric. Other inlet protection methods include concrete block wrapped with wire and stones and placing geotextile fabric and stones directly over the inlet (Brown and Caraco, 1997). These sediment traps will allow stormwater flow to pass through, but will filter out significant amounts of suspended sediments. These structures also provide protection in the event of an accidental fuel oil spill, hydraulic hose rupture, or other hazardous material release, providing some measure of initial containment upgradient of adjacent wetland areas. As with all hay bale structures, the sediment traps need to be maintained and periodically replaced to ensure their effectiveness. Excavation for foundations, utility trenches, and other facilities will often extend below the existing water table, resulting in collection of groundwater within the excavation. In order to dewater these areas and prevent discharge of sediment-laden water into surrounding areas, various types of settling basins and detention structures can be constructed. Sediment removal efficiencies generally range from 60 to 90 percent, with higher efficiencies associated with wet storage (Brown and Caraco, 1997). These structures allow particulate matter to settle and gradually discharge filtered runoff. Figure 3 is a schematic representation of a typical settling basin which can be constructed upgradient of a wetland area using filter fabric and clean rip-rap material to effectively filter silts and sediments at a construction site. Concrete or fiberglass settling basins are also available for use as sedimentation control structures during dewatering operations and are often used on barges during dredging operations to filter water discharged from stockpiled dredged materials. Geotextile wetland filter bags have also been developed to serve as sedimentation and erosion control devices on construction sites (Figure 4). The true test of any sedimentation and erosion control plan will occur during the first significant rainfall event during construction. Thus, it is recommended that on-site resident inspectors monitor the success of the installed erosion control devices ©2001 CRC Press LLC

Ground Slope

15' - 20 (Typ.) or as Directed

Suitable Device to Dissipate Velocity

To Natural Water Course

Sediment Laden Water Sediment Free Water Pump Discharge Flat Stone Ground Slope Approved Filter fabric Mat Baled Hay or Straw

10'-15' (Typ.) or as Direcled

Pump Discharge Line

Suitable Velocity Dissipator Clean Stones (If Required) Flat Stone

Approved Filter Fabric Mat Sediment

Baled Hay or Straw as Directed

TYPICAL SECTION

SEDIMENT TRAP Figure 3

Settling basins are used in conjunction with dewatering operations to prevent discharge of sediment-laden water into wetlands. The basins are constructed upgradient of wetlands using filter fabric and clean rip-rap material.

during and immediately after a rainstorm or snowmelt. The hay bales, siltation fences, and other structures should be observed on, at least, a weekly basis to detect damage from wildlife, machinery, or other activities on site. Equally important, resident inspectors should conduct frequent visual observations of the adjacent wetlands or open water bodies to detect turbidity plumes resulting from on-site runoff. For certain subaqueous activities, significant shortterm increases in turbidity are unavoidable. Nevertheless, attention should focus on the effectiveness of the siltation curtains, dredging methods, and dewatering

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Wetland Filter Bag

Flow Pipe Water flowing out from the bag

Figure 4

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A geotextile bag can be used on construction sites to remove sediments from site runoff.

practices to ensure that surrounding background levels of turbidity are not significantly increased. For deepwater areas, the use of a Secchi disk or similar device will provide a qualitative measure of the water clarity and amount of suspended sediments during construction.

NITROGEN LOADING Nitrogen occurs in wetlands in various inorganic and organic forms (Mitsch and Knight, 1997). Ammonia, nitrate, and nitrite are the most important forms for wetland processes. Ammonia is an important nutrient for most wetland plants and autotrophic bacteria and is a growth limiting compound in coastal waters. Coastal waters are the most highly fertilized ecosystems on earth (Nixon, 1986; Kelly and Levin, 1986). In natural waters, ammonia is readily oxidized resulting in oxygen consumption. Ammonia in its unionized form (NH3) is toxic to many forms of aquatic life at low concentrations (0.2 ppm). Nitrate is reduced to nitrite in oxygen-poor environments and is conservative in groundwater. In infants, nitrite combines with fetal hemoglobin preventing oxygen transport. This potentially fatal condition is known as methemoglobinemia. High nitrate concentrations have also been linked to carcinogenic effects (U.S. Environmental Protection Agency, 1990). The primary source of nitrogen, particularly in residential areas, is domestic wastewater (Cape Cod Commission Water Resources Office, 1992; Valiela et al., 1997). For residential septic systems, the concentration of nitrogen depends upon soil characteristics, loading rates, distance to the impervious stratum, distance to the water table, time of year, and depth below the leach field (Suffolk County Department of Health Services, 1983; Canter and Knox, 1985). Typical nitrate nitrogen concentrations range from 33 to 41 ppm (Cape Cod Planning and Economic Development Commission and U.S. Environmental Protection Agency, 1978; Nassau-Suffolk Regional Planning Board, 1978; Suffolk County Department of Health Services, 1983; IEP, 1988; Robertson et al., 1991). Nitrogen concentrations in nonresidential areas are less well known and vary widely in character and quantity. In general, nitrate nitrogen concentrations are higher when no gray water (i.e., sinks and showers) is present. Secondary sources of nitrogen include lawn fertilizer, atmospheric nitrogen, and runoff from impervious surfaces. Nitrogen levels from lawn fertilizer vary with soil type, application rate, precipitation, temperature, turf type, and nitrogen form. Typical fertilizer application rates range from 0.8 to 1.7 kg/100 m2 (1.7 to 3.8 lb nitrogen/1000 ft2) up to 4.7 kg/100 m2 (9.6 lb nitrogen/1000 ft2) for golf course greens (Nassau-Suffolk Regional Planning Board, 1978; Cape Cod Planning and Economic Development Commission, 1979; Eichner and Cambareri, 1990). Nitrogen leaching rates vary widely, from 0 to 60 percent (Nassau-Suffolk Regional Planning Board, 1978; Brown et al., 1982; IEP, 1988; Petrovic, 1990). Controlled application of fertilizer to healthy turf can eliminate or minimize leaching (Petrovic, 1990).

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Atmospheric nitrogen loading, largely from precipitation, is relatively minor compared to potential loading from wastewater and fertilizers. Precipitation concentrations in the United States range from 0.14 to 1.15 ppm nitrate nitrogen (Loehr, 1974). Dry deposition of nitrogen may double this concentration (Valiela et al., 1997). Nitrogen loading off of impervious surfaces is, however, significant, ranging from 0.41 to 1.75 ppm nitrate nitrogen and 1.13 to 10 ppm total nitrogen (IEP, 1988). Recharge rate off of impervious surfaces is poorly understood. The TR-55 stormwater modeling program assumes a recharge rate of 98 percent (Soil Conservation Service, 1986) and the Water Resources Office of the Cape Cod Commission (1992) assumes a 90 percent recharge rate. Planning Guidelines Based upon the threat of methemoglobinemia and cancer, the U.S. Environmental Protection Agency has established a limit of 10 ppm nitrate nitrogen in drinking water (U.S. Environmental Protection Agency, 1990). Studies on Long Island, NY, revealed that average nitrate nitrogen concentrations of 6 ppm led to violation of the 10 ppm criteria 10 percent of the time, and that average concentrations of 3 ppm led to violation of the 10 ppm criteria 1 percent of the time (Nassau-Suffolk Regional Planning Board, 1978; Long Island Regional Planning Board, 1986). Based upon this information, Long Island recommended that areas be sewered if the average nitrate nitrogen concentration exceeds 6 ppm. The Cape Cod Planning and Economic Development Commission (1978) and Cape Cod Commission Water Resource Office (1992) adopted a 5 ppm nitrate nitrogen standard. A second consideration in the establishment of nitrogen standards is protection of coastal embayments. Each embayment has a unique critical nitrogen loading rate dependent upon embayment morphology and flushing rate. For example, U.S. Environmental Protection Agency and the Massachusetts Executive Office of Environmental Affairs (1991) developed recommended nitrogen loading limits for Buzzards Bay (Table 4). Recommended nitrogen loads are lower in shallower embayments than deep embayments and in higher quality waters than lower quality waters. Based upon studies of Waquoit Bay, MA, Valiela et al. (1997) offer general recommendations that wastewater disposal within 200 m of shore be limited, that homes be required to use multiple leaching fields or septic systems, and that fertilizer use be controlled on near-shore lawns. Estimating Nitrogen Loads Several methods have been used to estimate nitrogen loading to groundwater and coastal embayments. For example, the Cape Cod Commission Water Resources Office (1992) recommends a site-specific mass balance analysis for relatively small sources and a cumulative loading analysis for proposed sources in groundwater recharge zones and relatively large sources. The former estimates nitrogen and water uses within the boundaries of a development, whereas the latter is a recharge zone-wide analysis for

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Table 4

Recommended Nitrogen Loading Limits for Coastal Embayments (Adapted from U.S. Environmental Protection Agency and Massachusetts Executive Office of Environmental Affairs, 1991) Embayment

Shallow Flushing < 4.5 Flushing > 4.5 Deep Select rate resulting in lesser annual loading

Outstanding Resource Areas3

SB Waters1

SA Waters2

350 mg/m3/Vr4 30 g/m2/yr

200 mg/m3/Vr 15 g/m2/yr

100 mg/m3/Vr 5 g/m2/yr

500 mg/m3/Vr 45 g/m2/yr

260 mg/m3/Vr 20 g/m2/yr

130 mg/m3/Vr 10 g/m2/yr

1

Excellent for fish, other aquatic life and wildlife, primary and secondary contact recreation, and shellfish harvesting with depuration.

2

Excellent for fish, other aquatic life and wildlife, primary and secondary contact recreation, and shellfish harvesting without depuration.

3

Outstanding socioeconomic, recreational, ecological, or aesthetic value.

4

Vollenweider flushing term; Vr equals r /1 +

r r equals flushing time (years).

existing and proposed conditions. The models estimate the nitrate nitrogen load by totaling the nitrogen inputs from wastewater, impervious surfaces (roof and paved), and fertilizer, and dividing nitrogen inputs by total water inputs. Figure 5 illustrates a mass balance analysis process for a hypothetical 20 house residential development. In this example, as typically occurs, the majority of nitrate nitrogen originates in wastewater. Valiela et al. (1997) developed a model to estimate atmospheric, fertilizer, and wastewater nitrogen loading to watersheds and receiving waters. Based upon data from the Waquoit Bay Land Margin Ecosystems Research Project and syntheses of published information, the model estimates nitrogen inputs to surfaces of the major types of land use within the landscape. Nitrogen losses in the various watershed compartments are then estimated. For example, atmospheric and fertilizer nitrogen are lost in vegetation, soils, the vadose zone, and aquifer. Wastewater nitrogen losses occur in septic systems and effluent plumes and during diffuse transport in aquifers. Nitrogen loss calculations are conducted separately for each major type of land cover. The model was developed for Waquoit Bay, MA, but is believed applicable to other rural to suburban watersheds underlain by unconsolidated sandy sediments. According to the model, the atmosphere is the largest contributor of nitrogen to the watershed, but wastewater is the largest source of nitrogen to receiving estuaries (Table 5). The model implies that estuary management should focus on wastewater disposal, particularly within 200 m of shore. The authors also suggest that installation of multiple conventional leaching fields or septic systems in high flow parcels could be beneficial. Other recommendations include control of fertilizer use on near-shore lawns and conservation of parcels of accreting natural vegetation. The latter effectively intercept atmospheric nitrogen.

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Model Parameters Impervious Surfaces Wastewater NO3 35 mg/L Roof Area 2,000 m2 Roof runoff NO3 0.75 mg/L Paved Area 4,000 Paved runoff NO3 1.5 mg/L Natural Area 20,000 m2 Fertilizer 1000 g/100 m2 Lawn Area 10,000 m2 Fertilizer leach rate 0.25 Wastewater 400 L/bedroom Impervious surface recharge rate 1 meter/year Natural area recharge rate 0.45 meter/year Roof 2,000 m2 × 1 m/yr × 1,011 L/m3 × 1 yr/365 day L runoff/day × 0.75 mg NO3/L =

5,539.7 L/day 4,154.8 mg/day

Paved 4,000 m2 × 1 m/yr × 1,011 L/m3 × 1 yr/365 day L runoff/day × 1.5 mg NO3/L =

11079.5 L/day 16,619.3 mg/day

Natural 20,000 m2 × 0.45 m/yr × 1,011 L/m3 × 1 yr/365

24,928.8 L/day

Lawn 10,000 m2 × 1000 mg/100 m2/yr × 1 yr/365 days 68.5 mg/day Wastewater 3.5 bedrooms × 400 L/bedroom × 20 bedrooms L wastewater per day × 35 mg NO3/L=

28,000 L wastewater per day 980,000 mg/day

Cumulative nitrate nitrogen load 4, 154.8 + 16, 619.3 + 68.5 + 980, 000 mg ------------------------------------------------------------------------------------------------------------ = 14.39 mg/L 5, 539.7 + 11079.5 + 24, 928 + 28, 000 liters Figure 5

Table 5

Nitrate nitrogen loading calculations for a hypothetical 20 house residential development with an average of 3.5 bedrooms per house.

Percent Nitrogen Input to the Watershed, Loss within the Watershed, and Input to Estuaries According to the Waquoit Bay Land Margin Ecosystems Research Model (Adapted from Valiela et al., 1997)

Source

Input to Watershed

Losses within Watershed

Input to Estuaries

Atmospheric Fertilizer Wastewater

56 14 27

89 79 65

30 15 48

STORMWATER RUNOFF Development is accompanied by an increase in impervious surfaces which increases stormwater runoff and decreases infiltration and evapotranspiration. This decreases the time for the runoff to reach wetlands and streams, increasing the

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frequency and severity of erosion and downstream flooding. During periods of prolonged dry weather, water tables and stream flows are reduced leading to a loss of wetland and aquatic habitats. Hydraulic and biological changes to streams occur when 10 to 20 percent of a watershed has impervious surfaces [Massachusetts Department of Environmental Protection and Massachusetts Office of Coastal Zone Management (DEP/CZM), 1997]. Typical impervious area percentages range from 20 to 40 percent for low density residential developments to 95 to 100 percent for business districts (Brach, 1989). Stormwater runoff is contaminated with a variety of pollutants that have various effects on wetland and aquatic habitats (Bingham, 1994; DEP/CZM, 1997). Nutrients from animal wastes, human wastes, and fertilizers induce algal growth and lower dissolved oxygen. Sedimentation also lowers dissolved oxygen as well as increasing turbidity and smothering aquatic life. Pathogens contaminate drinking water, swimming areas, and shellfish. Metals, hydrocarbons, organic chemicals, and salt increase the toxicity of the water column and sediments and may bioaccumulate in aquatic organisms. Stormwater impacts to wetland and aquatic habitats can be avoided or minimized through careful planning, use of nonstructural practices, and use of structural best management practices (BMPs). These measures reduce the volume of runoff, store runoff water, promote infiltration of stormwater to groundwater, and remove pollutants. Planning and Nonstructural Practices Effective stormwater management planning will minimize the size and cost of structural requirements. Planning and nonstructural practices can mitigate most stormwater impacts to wetlands and aquatic habitats for small developments. For larger developments, planning and nonstructural practices can significantly reduce the extent and, therefore, the cost of structural BMPs. Perhaps the most important planning technique available is minimizing impervious surfaces. Minimization of impervious surfaces is critical in recharge areas, especially those associated with drinking water supplies. Methods for minimizing impervious surfaces include the maintenance of natural buffers and drainageways. This allows infiltration of runoff, reduces runoff velocity, and removes suspended solids. Other methods include minimizing steep slopes, reducing building footprints and parking areas, limiting the width of roadways and the use of sidewalks, using shallow grassed roadside swales and parking lot islands, using turf pavers, gravel, and other porous surfaces, and maintaining as much predevelopment vegetation as possible (DEP/CZM, 1987). Other planning techniques are also available for managing stormwater. Developments can be “fit” to the terrain by designing road patterns that match the landform. Grassed waterways, vegetated drainage channels, and water quality swales can be constructed along roadways to channel runoff (DEP/CZM, 1987). Similarly, natural,

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vegetated drainageways can be preserved, helping to maintain predevelopment flood volumes, peak discharges, and base flows. Pollutants will be filtered by vegetation and will bind to underlying soils and organic matter. The planning process should also attempt to mimic predevelopment hydrologic conditions including peak discharge, runoff volume, infiltration capacity, base flow levels, groundwater recharge, and water quality. Several nonstructural techniques can also be effective in managing stormwater quantity and quality. Developments, especially commercial developments, will benefit from preparation of a pollution prevention plan that identifies potential sources of pollution and ensures implementation of practices that reduce pollutants in stormwater discharges (DEP/CZM, 1987). Example techniques include proper pesticide and fertilizer application, pet waste management, the proper storage, use, and disposal of hazardous chemicals, and proper operation and maintenance of septic systems. Areas accustomed to winter snowfalls use sand and salt to mitigate icy roadways. Street and parking lot sweeping can reduce total suspended solids like sand and salt by 5 to 80 percent. Vacuum sweepers tend to be more effective than mechanical sweepers. Much of the solids also ends up in catch basins, which benefit from regular cleaning. Salt (NaCl) toxicity can be minimized by using alternative de-icing compounds such as calcium chloride (CaCl2) and calcium magnesium acetate (CMA), and by designating “low salt” areas on roadways near wetlands and streams. Deicing compounds should be stored on sheltered, impervious pads, and stored snow should be placed where it can slowly infiltrate into the ground. Structural BMPs Structural BMPs are required when planning and nonstructural practices alone are insufficient to mitigate stormwater impacts. There are five major categories of stormwater structural BMPs: pretreatment, detention basins/retention ponds, vegetated treatment, infiltration, and filtration (Table 6). As with planning and nonstructural practices, the goals of a stormwater management design using structural BMPs should be to approximate predevelopment runoff rates and volume and to maximize pollutant removal. In many instances, the most effective design will incorporate several BMPs in series. Selecting a BMP requires consideration of the quantity of stormwater runoff to be produced, the water quality to be achieved, the proximity of critical areas (e.g., wetlands, aquatic habitats), maintenance requirements, aesthetics, cost, and site constraints (DEP/CZM, 1987; Schueler, 1987; Horner et al., 1994). In many instances, site physical characteristics limit or determine BMP selection. For example, sandy soils will inhibit the use of ponds but will facilitate infiltration BMPs. Pond BMPs require a relatively large contributing drainage area, whereas infiltration BMPs are restricted to a relatively small drainage area. A water table at or near the surface is essential for wetlands and wet ponds but will preclude infiltration BMPs. Swales and trenches are most effective when slopes are greater than 5 percent but less than 20 percent. Wet ponds and wetlands should not outflow to cold water streams, and infiltration BMPs should not be located close to foundations.

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Table 6

The Major Types of Structural Best Management Practices for Stormwater Pretreatment Sediment traps Water quality inlets Catch basins Detention/retention ponds Extended detention ponds Wet retention ponds Vegetated treatment Drainage channels Water quality swales Constructed wetlands Infiltration Dry wells Trenches Basins Filtration Basins

Pretreatment Sediment traps, water quality inlets, and catch basins remove debris, oil, and grease, and sediment and associated pollutants. Settling is the primary treatment mechanism. Sediment traps are on-line units, whereas water quality inlets and catch basins are off-line units. Pretreatment BMPs should only be used as pretreatment devices for other stormwater management technologies because they have limited storage capacity, short detention times, and do not remove soluble pollutants. Essential elements of detention basin/retention pond and vegetated systems, pretreatment BMPs are applicable to other BMPs as well. The longevity of pretreatment BMPs is high with frequent maintenance. A sediment trap is an excavated pit or cast structure 1 to 2 m (3 to 6 ft) deep. Typically, sediment traps can accommodate the 2- and 10-year storms. Water quality inlets and catch basins are chambered, underground retention systems. Water quality inlets have multiple chambers with permanent pools of water in the first couple of chambers. Floatable debris and sediments are trapped in the first chamber, oil and grease are trapped in a second chamber, and water is routed out of a third chamber into the storm drain or another BMP. Catch basins operate similarly, but only a single chamber is present. Water quality inlets and catch basins are particularly applicable to parking lots and other areas with substantial vehicular traffic. Detention Basins/Retention Ponds Detention basins capture and hold stormwater for 24 h or more, which permits solids to settle and downstream flooding to be reduced. Typically, a detention basin will have a lower stage capable of detaining smaller storms, and an upper stage capable of detaining larger, less frequent storms. One of the less expensive BMPs ©2001 CRC Press LLC

capable of controlling both stormwater quantity and quality, detention basins remove significant levels of sediment and sorbed pollutants. Detention basins are largely ineffective at removing soluble pollutants. Retention ponds use a deep, permanent pool of water to remove both solid and soluble pollutants (Figure 6). Soluble pollutants, such as nutrients, are removed by the biological activity of algae and fringing wetland. Retention ponds also have additional storage capacity to control peak discharge rates. A pool depth of 0.9 to 1.8 m (3 to 6 ft) is recommended to optimize particle settling.

Figure 6

Retention ponds use a deep, permanent pool of water to remove solid and soluble pollutants.

Both detention basins and retention ponds require a contributing watershed of at least 4 ha (10 acres). Inflow points should have energy dissipaters and a forebay or settling zone to trap course sediments. The original design should account for the gradual accumulation of sediment. A routine inspection should be conducted at least once a year, and sediment should be removed as necessary. Vegetated Treatment Treatment by vegetated BMPs ranges from simple drainage channels to constructed wetlands. Each vegetated BMP, to a degree, controls peak discharges by reducing runoff velocity and promoting infiltration and reduces pollutants by trapping, filtering, and infiltration. Both drainage channels and water quality swales are most effective when the percentage of impervious cover in the contributing area is relatively small and the slope is minimal (0 to 5 percent). Constructed wetlands require relatively large contributing drainage areas to maintain dry weather base flows. All vegetated BMPs benefit from inclusion of a sediment forebay to settle large particles. Drainage channels have grass or some other channel lining so runoff can be conveyed during large storm events without causing erosion. Channels are applicable to residential and other low to moderate density areas and can be used in parking lots. A minimum channel length of 30 m (10 ft) is recommended to optimize pollutant removal. ©2001 CRC Press LLC

Water quality swales are drainage channels enhanced to remove stormwater pollutants and are typically sized to accommodate the 10-year storm. The three major types of water quality swales are dry swales, wet swales, and grassed swales (DEP/CZM, 1987; Claytor and Schueler, 1996). Dry swales allow for filtering or infiltration through the bottom of the swale. Underlying soils should be permeable, and the seasonal high water table should not be within 0.7 to 1.3 m of the swale bottom. Wet swales are useful when the water table is at or near the soil surface, or soils are poorly drained (Figure 7). Sediment accumulation, filtration, and vegetation uptake remove pollutants. Grassed swales resemble dry swales in that underlying soils are relatively permeable, but they are vegetated with moisture-tolerant grass species that produce a fine, uniform, dense cover. Filtration, vegetation uptake, sediment accumulation, and to some extent infiltration are the pollutant removal mechanisms.

Figure 7

Water quality swales remove stormwater pollutants by filtration, vegetation uptake, sediment accumulation, and infiltration.

Constructed wetlands are the most complex, and most expensive, vegetated treatment BMP. Designed to mimic elements of natural wetlands, constructed wetlands reduce peak discharge and reduce the occurrence of downstream flooding, settle particulate pollutants, and facilitate the uptake of pollutants by vegetation. Constructed wetlands require relatively large contributing drainage areas to maintain dry weather base flows, and construction costs are relatively high. Chapter 10 discusses constructed wetlands for the treatment of stormwater and other waste waters. Infiltration Infiltration BMPs are aggregate-filled devices which capture stormwater runoff and gradually exfiltrate the runoff through the bottom of the device into the subsoil ©2001 CRC Press LLC

and groundwater. Examples include dry wells and infiltration trenches and basins. Infiltration BMPs require permeable underlying soils (minimum 1.3 cm/h) and a groundwater level at least 0.7 m (2 ft) below the bottom of the infiltration device. To avoid contaminating groundwater, infiltration BMPs should not be used when runoff is highly contaminated. Infiltration BMPs can be used to manage peak discharges and reduce runoff volume. Regular maintenance is required because infiltration BMPs are susceptible to clogging by sediments. Dry wells are small pits used for infiltrating relatively good quality water such as roof runoff. With a storage time of 48 to 72 h, dry wells typically have a contributing drainage area of less than 0.4 ha (1 acre). Infiltration trenches are applicable to sites with gentle slopes (5 percent or less), groundwater levels at least 1.2 m (4 ft) below the surface, and contributing drainage areas of 2 ha (5 acres) or less. In addition to reducing runoff volume and peak discharge, infiltration trenches remove soluble and particulate pollutants from runoff. Infiltration trenches require pretreatment by inlets, sumps, swales, or forebays to remove sediment and oil, and grease which may clog the trench. Infiltration basins resemble detention/retention ponds but are constructed over permeable soils. The contributing drainage area should be 6 ha (15 acres) or less, depth to seasonal high water table at least 0.7 m (2 ft), and the soil infiltration rate 1.3 to 6 cm (0.5 to 2.4 in.) per hour. A sediment forebay or other pretreatment device is necessary to capture coarse particulate pollutants. Infiltration basins are designed to have a retention time of 48 to 72 h. Filtration Filtration basins consist of sand, peat, or compost underlain by gravel and perforated underdrains. Filter fabrics may be installed at the top of the bed, and between the filter media and gravel bed, to minimize clogging. Pollutant removal is achieved by settling on top of the basin and by straining pollutants through the filtering media. A sedimentation chamber designed to remove coarse pollutants precedes the filter basin. Filter basins are applicable to small drainage areas of 0.2 to 2 ha (0.4 to 4 acres) for most development situations. A design filtration rate of 5 cm (2 in.) per hour is typical, and the filter should drain within 24 h.

REFERENCES Bingham, D., Wetlands for stormwater treatment, in Applied Wetlands Science and Technology, Kent, D. M., Ed., Lewis Publishers, Boca Raton, FL, 1995. Brach, J., Protecting Water Quality in Urban Areas: Best Management Practices for Minnesota, Minnesota Pollution Control Agency, Division of Water Quality, St. Paul, MN, 1989. Brown, K. W., Thomas, J. C., and Duble, R. L., Nitrogen source effect on nitrate and ammonium leaching and runoff losses from greens, J. Agron., 74, 947, 1982. Brown, W. E. and Caraco, D. S., Muddy water in - muddy water out?: a critique of erosion and sediment control plans, Watershed Prot. Tech., 2(3), 393, 1997.

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Canter, L. W. and Knox, R. C., Septic Tank System Effects on Groundwater Quality, Lewis Publishers, Chelsea, MI, 1985. Cape Cod Commission Water Resources Office, Nitrogen Loading, Technical Bulletin 91-001, Barnstable, MA, 1992. Cape Cod Planning and Economic Development Commission, Water Supply Protection Project, final report, Barnstable, Bourne, Brewster, Dennis, and Yarmouth, Barnstable, MA, 1979. Cape Cod Planning and Economic Development Commission and U.S. Environmental Protection Agency, Draft environmental impact statement and proposed 208 water quality management plan for Cape Cod, Barnstable, MA, 1978. Claytor, R. A. and Schueler, T. R., Design of Stormwater Filtering Systems, Center for Watershed Protection, prepared for Chesapeake Research Consortium, Solomons, MD, 1996. Dahl, T. E., Wetland Losses in the United States 1780s to 1980s, U.S. Department of the Interior, Fish and Wildlife Service, Washington, D.C., 1990. Eichner, E. M. and Cambareri, T. C., The Cape Cod Golf Course Monitoring Project, Cape Cod Commission, Water Resources Office, Barnstable, MA, 1990. Heimlich, R. and Melanson, J., Wetlands lost, wetlands gained, Natl. Wetlands Newsl., May–June, 17(3), 1, 23, 1995. Hollis, T. and Bedding, J., Can we stop wetlands from drying up?, New Sci., July, 31, 1994. Horner, R. R., Skupien, J. J., Livingston, E. H., and Shaver, H. E., Fundamentals of Urban Runoff Management: Technical and Institutional Issues, Terrene Institute, Washington, D.C., 1994. Horner, R. R., Guedry, J., and Kortenhog, M. H., Improving the Cost Effectiveness of Highway Construction Site Erosion and Pollution Control, Washington State Transportation Center, Federal Highway Administration, Seattle, WA, 1990. IEP, Water Resources Protection Study: Town of Yarmouth, Massachusetts, Sandwich, MA, 1988. Kelly, J. R. and Levin, S. A., A comparison of aquatic and terrestrial nutrient cycling and production processes in natural ecosystems, with reference to ecological concepts of relevance to some waste disposal issues, in The Role of the Oceans as a Waste Disposal Option, Kullenberg, G., Ed., NATO Advanced Research Workshop Series, Reidel, Dordrecht, The Netherlands, 1986, 154. Loehr, R. C., Characteristics and comparative magnitude of nonpoint sources, J. Water Pollut. Control Fed., 46, 1849, 1974. Long Island Regional Planning Board, Special Groundwater Protection Area Project for the Oyster Bay Pilot Area and Brookhaven Pilot Area, 1986. Massachusetts Department of Environmental Protection and Massachusetts Office of Coastal Zone Management, Stormwater management, Vol. 2, Stormwater Technical Handbook, MADEP, 1997. McManus, K., Wetlands avoidance and impact minimization, in Applied Wetlands Science and Technology, Kent, D. M., Ed., Lewis Publishers, Boca Raton, FL, 1994, 105. Mitsch, W., The world’s wetlands and SWS—a call for an international view, Wetlands Bull., September, 1, 4, 1995. Mitsch, W. H. and Knight, R. L., Treatment Wetlands, Lewis Publishers, Boca Raton, FL, 1997. Nassau-Suffolk Regional Planning Board, The Long Island Comprehensive Waste Treatment Management Plan, Hauppauge, NY, 1978. Nelson, K. A., Design considerations for segmental retaining walls in water environments, Land Water, July/August, 22, 1995.

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Nixon, S. W., Nutrients and the productivity of estuarine and coastal marine ecosystems, J. Limnol. Soc. S. Afr., 12, 43, 1986. Paterson, R. G., Luger, M. I., Burby, R. J., Kaiser, E. J., Malcolm, H. R., and Beard, A. C., Costs and benefits of erosion and sediment control: the North Carolina experience, Environ. Manage., 17(2), 167, 1993. Petrovic, A. M., The fate of nitrogenous fertilizers applied to turfgrass, J. Environ. Qual., 19(1), 1, 1990. Robertson, W. D., Cherry, J. D., and Sudicky, E. A., Groundwater contamination from two small septic systems on sand aquifers, Ground Water, 29(1), 82, 1991. Schueler, T. R., Controlling Urban Runoff: A Practical Manual for Planning and Designing Urban BMPs, Metropolitan Washington Council of Governments, Washington, D.C., 1987. Soil Conservation Service, Urban Hydrology for Small Watersheds, 2nd ed., Technical Release 55, U.S. Department of Agriculture, Washington, D.C., 1986. Suffolk County Department of Health Services, Pilot Plant Study: Nitrogen Removal in Modified Residential Subsurface Sewage Disposal System, Phase 2—Additional Investigations, Hauppauge, NY, 1983. U.S. Congress, National Environmental Policy Act, PL 91-190, 42 USC4321, Federal Register, Vol. 43, U.S. Government Printing Office, Washington, D.C. U.S. Environmental Protection Agency, 1990. U.S. Environmental Protection Agency and Massachusetts Executive Office of Environmental Affairs, 1991. U.S. Office of Technology Assessment, Wetlands: Their Use and Regulation, U.S. Congress, Washington, D.C., 1984. Valiela, I., Collins, G., Kremer, J., Lajtha, K., Geist, M., Seely, B., Brawley, J., and Sham, C. H., Nitrogen loading from coastal watersheds to receiving estuaries: new method and application, Ecol. Appl., 7(2), 358, 1997. W&H Pacific and CH2M-Hill, Demonstration project using yard debris compost for erosion control, Portland Metropolitan Service District, Portland, OR, 1993. Woodward, S. E., The Effectiveness of Buffer Strips to Protect Water Quality, Master’s thesis, University of Maine, Orono, ME, 1989.

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Zentner, John “Wetland Enhancement, Restoration, and Creation” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

6

Wetland Enhancement, Restoration, and Creation John Zentner

CONTENTS Site Selection and Analysis Topography Vegetation Association Mapping Site History and Current Status Hydrological Analysis Soil Analysis Cultural Constraints Adjacent Site Conditions The Use of Template Associations Small-Scale Experimental Construction Goal Setting Elements of a Goal Statement Goal-Setting Process Practicability Construction Design Geography Size and Shape Location Slope Adjacent Uses Hydrology Hydroperiod and Depth Water Supply Soil

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Vegetation Succession Planting Design Plant Selection Stock Selection Planting Density Weed Control Cultural Issues Mosquitoes Water Quality Implementation Construction Sequencing Protective Flagging Weed Removal Salvaging Grading Planting Water Supply Fencing As Builts Maintenance Weed Control Erosion Control Herbivory Plant Care Irrigation System Maintenance Litter Removal General Maintenance Frequency Minimizing Maintenance Efforts Research Needs References

Freshwater wetlands develop at elevations above open water aquatic habitats and below uplands. They are found in a wide range of hydrologic conditions, from permanently flooded (to a depth of 1 m) to seasonally saturated. Freshwater wetlands occur on a wide variety of soil types including both organic and mineral soils, as well as in nonsoil conditions. Most freshwater wetlands are either freshwater marshes or riparian woodlands. Freshwater marshes are dominated by herbaceous emergents and can be divided into three general categories reflective of hydrology (Figure 1). Wet meadows are temporarily or intermittently flooded and dominated by graminoids and Juncaceae. In the United States, seasonal marshes are seasonally flooded or saturated and dominated by Cyperaceae and Juncaceae. Perennial marshes are permanently or

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UPLAND

Hydrology Dominant Plants Figure 1

WET MEADOW

SEASONAL MARSH

PERENNIAL MARSH

Temporarily flooded or intermittently flooded

Seasonally flooded or saturated

Permanently flooded or semi-permanently flooded

Grasses, rushes

Sedges, rushes

Cattail, bulrush, tules

OPEN WATER

Freshwater marshes are dominated by herbaceous emergents and can be divided into three hydrological categories: perennial marsh, seasonal marsh, and wet meadow.

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semipermanently flooded and dominated by tall emergents such as cattails (Typha latifolia) or bulrush (Scirpus acutus). Riparian woodlands are dominated by shrubs and trees and are characterized by impermanent and varying periods of inundation or root zone saturation during the growing season. Compared to freshwater marshes, riparian woodlands occur on relatively permeable and well-oxygenated substrates. As with freshwater marshes, riparian woodlands can be categorized by hydrological regime (Figure 2). High terrace woodlands are temporarily flooded and dominated by a variety of species, especially oaks (Quercus spp.), that are typified by heavy seeds with relatively longer viability. Mid-terrace woodlands are seasonally flooded and generally dominated by green ash (Fraxinus pennsylvanica), sycamore (Platanus occidentalis), and other species with medium weight seeds. And low terrace woodlands are semipermanently flooded and generally dominated by willows (Salix spp.), silver maple (Acer saccharinum), and similar species with relatively light seeds of limited viability. These categories correspond to Categories V (higher hardwood wetlands), IV (medium hardwood wetlands), and III (lower hardwood wetlands), as described by Cowardin et al. (1979), Larson et al. (1981), and Clark and Benforado (1981). Coastal wetlands share many of the characteristics of freshwater wetlands and are generally defined as those wetlands that lie within the realm and effects, however minor, of tidal salt water. As such, coastal wetlands include saltmarsh, fresh and brackish tidal marsh, and, in tropical waters, mangrove. Saltmarsh is the most ubiquitous type of coastal wetland, occurring on all coasts where appropriate substrate and tidal regimes are present (Figure 3). Saltmarshes are distributed over a relatively broad salinity range, from the high intertidal zone to the oligohaline habitats upstream on tidal tributaries where salinity may never exceed 10 parts per thousand. In the United States, various species of cordgrass (Spartina spp.) are dominant, with rushes (Juncus spp.) also common. On the Pacific Coast, Spartina foliosa and pickleweed (Salicornia virginica) are common. On the Gulf Coast and in the southeast, other species of cordgrass (Spartina alterniflora, S. patens), saltgrass (Distichlis spicata), and black needlerush (Juncus roemerianus) are typical. Sawgrass (Cladium jamaicense) marshes, such as the Everglades, occur in more brackish water. On the Atlantic Coast and to the northeast, Spartina cynosuroides, S. alterniflora, J. roemerianus, and some other species are common. Wax myrtle (Myrica cerifera) and groundsel tree (Baccharis halimifolia) are typical shrubs associated with salt marshes from the Gulf Coast to New England. Oligohaline (salinity of 0.5 to 5.0 ppt) and tidal freshwater marshes (salinity is less than 0.5 ppt) are herbaceous wetlands located in tidally influenced rivers or streams. The plant community exhibits a diverse mixture of true marine species and typical freshwater taxa that tolerate low salinities (Cowardin et al., 1979; Lewis, 1990). Mangrove forests are limited in distribution to subtropical and tropical zones (Figure 4). In the United States, they occur predominantly in southern Florida, sparsely along the Gulf Coast to the Laguna Madre of Texas, and extensively in Puerto Rico (Kuenzler, 1974). Many species of trees of different families are called mangroves, with black mangrove (Avicennia germinans), red mangrove (Rhizophora mangle), and white mangrove (Laguncularia racemosa) common to the ©2001 CRC Press LLC

UPLAND

Hydrology Dominant Plants Figure 2

HIGH TERRACE WOODLAND

MID-TERRACE WOODLAND

LOW TERRACE WOODLAND

Temporarily flooded

Seasonally flooded

Semi-Permanently flooded

Oaks

Green Ash, Sycamore

Willows, Silver Maple

Riparian woodlands are dominated by shrubs and trees. As with freshwater marshes, riparian woodlands can be categorized by hydrological regime: semipermanently flooded, seasonally flooded, and temporarily flooded.

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Figure 3

Saltmarshes are the most widely distributed of coastal wetlands. They occur over a relatively broad salinity range from the high intertidal zone to oligohaline habitats on tidal tributaries.

United States. There are an additional 31 species worldwide (Tomlinson, 1986). All mangroves have features in common that are adaptations for survival in the saline conditions of the intertidal zone (Tomlinson, 1986). Morphological adaptations include the aerial roots (pneumatophores) of black mangrove that provide for gas exchange and the viviparous, floating propagules of red mangroves. In fact, nearly all species of mangroves are viviparous, with the white mangrove being an exception. Physiological mechanisms for dealing with salt excretion or exclusion are also characteristic of mangroves. Lewis (1990) recently defined wetland enhancement, restoration, and creation. Enhancement is an increase in values afforded to a specific vegetation association by construction actions. Restoration is the recreation of a specific vegetation association on a site where that association was once known to occur. Creation is the construction of a wetland from an upland or aquatic site. The term construction will be used to encompass enhancement, restoration, and creation in this chapter. The earliest wetland construction projects may have occurred many thousands of years ago with the manipulation and creation of freshwater and coastal wetlands to enhance rice or fish harvests. Waterfowl conservation and hunting organizations have been responsible for numerous wetland construction projects in this century. New York, especially, had numerous small freshwater marshes created in the mid1950s (Dane, 1959). Specific problems associated with wetlands, such as nuisance mosquitoes, have also resulted in a number of useful research and enhancement practices. More recently, interest in wetland construction is a response to regulatory requirements. Nevertheless, there is also a greater public interest and understanding

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Figure 4

Mangrove forests are found throughout the subtropics and tropics. The tree species which comprise mangrove forests have morphological and physiological adaptations for surviving in intertidal saline conditions.

of the environmental values of wetlands and a desire for the recreation of lost or diminished landscapes and values. This chapter discusses the construction of freshwater marshes, riparian woodlands, and coastal wetlands. The construction process is described in five steps inherent to designing, building, and maintaining wetland landscapes. These steps are site analysis, goal setting, construction design, implementation, and maintenance. Implicit in these steps are two tenets. First, specification of a target vegetation association (TVA) is the primary goal. A TVA provides the habitat for any plant or wildlife populations that may be desired on the project site and, because of its structural nature, facilitates maintenance and monitoring efforts. Second, in construction planning there is no substitution for directed observation of the project site and the TVA.

SITE SELECTION AND ANALYSIS The ideal construction site will most closely meet the requirements of the community to be constructed. Therefore, the goals of the project will have substantial bearing on final site selection. Construction projects inherently require understanding the initial cause of habitat differentiation or degradation and determining the probability that the habitat can be constructed and maintained. The question of habitat displacement inevitably needs to be addressed in the site selection process as well. Site analysis is the first tangible step in a wetland construction project. The analysis may be completed during a wetland delineation effort, as the first step in a ©2001 CRC Press LLC

neighborhood restoration program, or from many other perspectives. This step provides the basic framework for goal setting, identification of the TVAs, and development of performance standards. The site analysis typically includes a topographic assessment, wetland vegetation association mapping and analyses, and a review of historic conditions. The analyses will also include an assessment of hydrologic, soil, and cultural conditions, and examination of any nearby templates, or examples, of TVAs. Small-scale experimental wetland construction efforts should be initiated at this time if possible. Generally, the initial site analysis is completed within a period of three weeks to three months, with a relatively intensive effort in defining site topography, vegetation associations, soils, and historic and cultural conditions in the first few weeks. Less extensive effort is expended in monitoring hydrology and constructing and observing experimental wetlands over the remaining period. Topography Elevation and slope are two of the more critical factors in determining the success of wetland construction projects. It is difficult to recommend an absolute planting elevation for a given species because of local and geographic variability. Synergistic effects may also alter growth at given elevations. For example, reduced salinity may permit growth of smooth cordgrass at higher and lower elevations than those typically recommended for the species. The optimum elevation can be determined empirically by observing and measuring the lower and upper elevation limits of a nearby natural wetland. Lewis (1990) recommends that the lowest and highest points should be disregarded, and only the middle range used for planting. If reference information is unavailable, an adequate test program should be conducted prior to initiating construction. Some degree of slope is essential for proper drainage. A gradual slope will increase the area available for planting and will dissipate hydrologic energy over a greater area, thereby reducing the possibility of erosion (Broome, 1990). Slope should be toward water sources to minimize ponding as substrates settle following site preparation. Gentle slopes do not drain as extensively as steeper ones, which is generally beneficial to most wetland vegetation. However, gentle slopes are more susceptible than steeper slopes to ponding, which can be especially problematic for riparian woodland species unless the soils are also relatively permeable. Vegetation Association Mapping Wetland vegetation associations for the construction site and surrounding area can be identified, and their extent determined, by mapping from an aerial photograph. Aerial photographs are usually available as stock or library film from an aerial survey center. These centers fly important regions every two to three years, providing a backlog or library of film that can be reproduced at a variety of scales. Requesting a half-tone mylar as well as a print of the project site is important. The mylar can be used to make blueprints for field use, and blackline prints can be reproduced and included in reports. ©2001 CRC Press LLC

The borders of the vegetation associations are more precisely defined through on-site analysis after initial mapping from an aerial photograph. For marshes, randomly selected 1-m plots representative of each vegetation association can be sampled for species and cover. Randomly selected 0.25 hectare (ha) polygons (defined by tree cover and separation among woodland vegetation associations) are effective in woodlands. Table 1 provides an example of vegetation association mapping from a hypothetical freshwater wetland project site in California characterized by gently sloping hills surrounding a central creek. The result is a table that describes the vegetation associations and the absolute and relative cover of marsh and riparian woodland vegetation. A scan of the table and knowledge of the region reveal that three vegetation associations are present and that the site is dominated by nonnative species and species representative of disturbed areas. It is also appropriate at this time to identify and survey any upland areas that may be considered for wetland construction to ensure that valuable upland habitats are not inadvertently lost. Table 1

Example of Vegetation Association Mapping from a Hypothetical Project Site in California1

Vegetation Association

Plant Species

Wet meadow

Area (ha)

Cover (%)

1.4

80 40

Lolium perenne* Hordeum hystrix* Elymus triticoides Elymus glaucus Seasonal marsh

1.4 Lolium perenne* Juncus balticus Lotus corniculatus* Polypogon monspieliensis*

Perennial marsh

40 30 20 10 0.9

Typha latifolia** Scirpus acutus Low-terrace woodland

0.5 Salix lasiolepis**

30 20 10 50

100 80 20 100 100

* Nonnative species. **Species indicative of disturbance. 1

The table provides absolute and relative information for use in site analysis and target vegetation association selection.

Site History and Current Status Shisler (1990) states that if wetlands are not present there has to be a reason, especially for tidal wetlands. Determining this reason is fundamental to reviewing a site for potential wetland habitat construction. If possible, past functioning of the ©2001 CRC Press LLC

wetland should be evaluated, and all past uses of the site should be reviewed prior to habitat construction. Current and future land use, zoning regulations, and projected sea-level rise may also have a bearing on site selection. Historic aerial and other site photographs, and any available topographic or other maps, are reviewed to define the past conditions and boundaries of the site vegetation associations. Important issues to review with these sources include the location of historic wetland vegetation associations and their relationship to current vegetation associations, water sources, and cultural elements. Historic aerials are often available from the U.S. Geologic Survey (USGS) and local aerial photography sources. Historic maps are also found at USGS as well as local historical societies. Costs for historic aerials are generally less than current aerials, however, reproduction costs for older maps may be significant. Hydrological Analysis The quantity, quality, and timing of water entering the wetland are of critical importance to its survival. If the correct hydrology is not present, or is insufficient, the project will fail (Shisler, 1990). The initial site analysis provides an opportunity to define the hydrologic conditions that create the site or template vegetation associations. Generally, hydroperiod, the duration of inundation or saturation, and water depth best define the vegetation association for freshwater wetlands. Accordingly, hydrology stakes and shallow observation wells are placed in wetland and upland areas to define surface and groundwater conditions. The wells typically consist of 10 cm diameter PVC pipe inserted 1.2 to 2.4 m into the ground, with a surrounding backfill of gravel or other permeable material. Wells and stakes can be observed weekly during the analysis to define the depth of ponding and the surface approach of groundwater. However, short-term monitoring may not reflect typical groundwater levels. Long-term groundwater level data may be available from local water districts or farmers. For coastal wetlands, hydrologic factors that need to be evaluated during the site selection process include drainage features, such as channels and ponds, the tidal regime and range, wave intensity, and salinity. Obviously, some of these can be manipulated through design while others cannot. It is sometimes necessary to base site selection upon the most practical trade-off. In general, coastal wetland construction projects are more likely to be successful in sheltered locations where wave energy and current velocity are minimal. Salinity variation at the site should be monitored so that appropriate plant species can be selected. Adequate drainage is another requirement, because standing water is typically detrimental to the establishment of wetland vegetation. Wide, shallow channels that retain water at low tide maximize flushing and plant health. Water quality (e.g., dissolved oxygen, turbidity) is an especially important factor determining site selection for seagrass restoration projects. Other factors affecting hydrology are precipitation, surface flows, ground water, and evapotranspiration. Each of these factors affects habitat productivity and species diversity.

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Soil Analysis Soil analysis can be one of the most important steps in defining wetland construction opportunities and constraints, especially for the drier freshwater marshes and riparian woodlands. The soil analysis can identify the edaphic requirements of the vegetation, thereby defining the appropriate TVA. Soil analysis will also identify any restrictions to construction activities because of soil permeability or soil chemistry. Generally, the best source of information on freshwater wetland soil characteristics for an area is direct site observations. Direct observations can be supplemented by local, state, or federal information, such as Natural Resource Conservation Service (NRCS, formerly U.S. Soil Conservation Service) soil series maps. The latter are available at no cost from NRCS offices. The NRCS and associated services such as county agricultural offices often have aerial photographs for review as well. Where soil series maps are unavailable, NRCS field personnel may provide on-site assistance. Site observations typically consist of a series of soil pits, 30 to 50 cm in depth, dug by hand throughout the site for the purpose of defining surface conditions. Use of a backhoe to define conditions at depth is generally helpful and is essential if significant excavation is required. Test pits will help identify any significant variation in permeability among soil layers. It is also useful to identify similar soils on proximal, undisturbed sites using the NRCS soil survey so that natural vegetation associations can be observed and used as a template. Soil chemical properties can be analyzed in a laboratory. If the site requires the addition of soil, or if the site has been filled in the past, the possibility of contamination should be investigated. Generally, sandy soils are more amenable to grading and planting than are silt and clay. However, some degree of organic content is recommended as a nutrient source (Broome, 1990). The soil must be of sufficient depth to support planting. Garbisch (1986) recommends a minimum depth of 0.3 m for marshes. Trees for woodland construction will typically require a greater soil depth, depending on the species. Based on a literature review and field inspections of created wetlands in New Jersey, substrate preferences for coastal marshes were ranked as follows: 1. 2. 3. 4.

Natural marsh peat Clay and silty clay Estuarine sediments (dredge and fill) Sand (Shisler, 1990)

For riparian woodlands, however, the list is almost completely reversed. Some anoxic soils can become highly acidic upon exposure to air, such as during earthwork, and may warrant the addition of a calcareous material such as crushed shell as a buffering agent. At sites subject to tidal or riverine inundation, sedimentation processes should be carefully studied. Some degree of sedimentation may be desirable to stabilize initial plantings; however, burial of seedlings should be avoided.

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Cultural Constraints Few sites are unconstrained by human artifacts. Artifacts may range from prehistoric remains to buried sewer lines. Visible artifacts such as roads that affect the project site (e.g., drainage impediment) should be identified. Identification of subsurface artifacts is more difficult. Nevertheless, cultural resources can be preliminarily identified in the United States through record searches with the State Historic Preservation Office (SHPO) or local universities. Buried infrastructure may be identified through interviews and research at local Public Works Departments. Adjacent Site Conditions Conditions near the proposed construction site should also be considered. Sites adjacent to existing functioning wetlands offer the greatest chance for success. Conversely, an increasing wetland area in the watershed may alter hydrologic regimes. The presence of aggressive exotic species close to the proposed site may also be a problem. In many cases, a buffer zone should be included in the project design. A buffer zone may increase the diversity of flora and fauna present (Josselyn et al., 1990) and will protect the site from human and vehicular traffic. The buffer zone must be kept clear of exotics and other nuisance species that may invade the project site. Buffer zones with tall vegetation provide animals, especially birds and mammals, with refuge from cold winds, high temperatures, and high waters. The proximity or attractiveness of the potential construction site to predation should also be considered. Geese (e.g., Branta canadensis), muskrat (Ondatra zibethica), and nutria (Myocastor coypus) have all been identified as voracious consumers of newly planted marsh vegetation (Shisler, 1990; Broome, 1990). Conversely, placement of a well-designed created wetland near a disturbed wetland may offer refuge to dwindling populations of desirable species. Open water near potential sites should be assessed for boat traffic and offshore depth because both can affect wave energy. Shallow offshore water reduces the severity of waves reaching the shore (Broome, 1990). The degree of wave energy can be modified to a certain extent (see below), but sites facing more than 5.5 nautical miles of open water are generally not recommended for planting (Crewz and Lewis, 1991). The Use of Template Associations Recreation of an historic vegetation association will in many instances increase the probability of success. In the event that the TVA is missing or poorly represented on the project site, a proximally located vegetation association can be used as a template. The template association should have hydrology and soils similar to existing or intended conditions for the construction site. In addition to determining species composition, the template vegetation association should be studied to determine species evenness and distribution, and identify potential additions to the planting list and likely successional patterns. The specific TVA to be constructed may be difficult ©2001 CRC Press LLC

to predict at this stage of the project. However, initial review of several different vegetation associations may have a significant effect on goal setting. For example, participants in a riparian woodland construction project for a degraded waterway in Solano County, CA, wished to plant the California sycamore (Platanus racemosa) extensively due to its stature and appearance. However, the sycamore is adapted to alluvial flats with relatively permeable soils and would not do well on the project site's clay soils. Taking the participants to a regional park that included a wellpreserved riparian woodland native to this region and similar to the project site resulted in the selection of several alternative trees, all better adapted to the project site. Small-Scale Experimental Construction When time is available, small-scale experimental projects or pilot studies should be undertaken to identify soil and hydrology constraints and to refine planning alternatives. Appropriate studies include determining the water-holding capacity of unmodified soils, determining germination and salvage survival rates for target vegetation, refining transplanting techniques, and evaluating construction management abilities. The importance of small-scale pilot projects for building confidence and consensus among the planning group should not be underestimated.

GOAL SETTING The setting of specific goals is crucial to wetland construction projects. Goals help define monitoring elements and protocols, and establish standards for judging success. The latter allows others to objectively evaluate the project and serves as a basis for practitioner and general public educational efforts. Elements of a Goal Statement Goal statements are typically comprised of a substantive objective, for example, construct 1.3 ha of coastal salt marsh. Specification of a TVA may be the most appropriate substantive goal as noted in the introduction to this chapter. A TVA can be related to wetland functions (e.g., provision of wildlife or fisheries habitat), facilitates clear performance standards and monitoring, and can be readily measured. This is most easily accomplished through comparison with an existing or historically well-described vegetation association from the project site or a nearby site. One of the more difficult tasks is to keep the TVA general enough to be realizable, yet specific enough that the goal is meaningful. Typically, selection of a species type association, for example a rush-dominated wet meadow, is an effective approach. Other commonly used options for the substantive element of the goal statement include a specific wildlife or plant population. In this case, the measurement parameter would identify a particular population level as the target. Wildlife goals are generally less useful than TVAs because wildlife populations vary so greatly year to year as a result of regional conditions. Regional climate changes can have a much greater effect on wildlife populations, for example, than the maturation of the ©2001 CRC Press LLC

construction site. Moreover, the construction project will build a specific type of habitat or assemblage of habitats to produce the desired change in wildlife populations. Therefore, designation of the TVA(s) makes as much sense as designation of the wildlife goal and is more readily measured. A wetland function, such as water quality improvement, is also often used as the substantive element of the goal and the measurement parameter defined to include an indicator such as percent reduction in target pollutants. Again, functional goals are still dependent on a specific TVA. Equally important, and often overlooked or unstated, is the procedural element. The procedural element of a goal statement refers to the method by which the goal statement is completed. Too often, goals are developed without discussion with, or consensus among, all interested parties. Conversely, a large committee may develop goals that are nonspecific, thereby negating effective evaluation at any point. Goal-Setting Process Several points should be considered in the development of goals. Construction goals should address the concerns of all those meaningfully affected by the project. Affected parties will include landowners on or adjacent to the project site, regulating agencies, and operation and maintenance entities such as mosquito abatement and flood control districts. Table 2 lists parties potentially affected by wetland construction activities, and goals common to these parties. All parties should be contacted and their comments solicited on initial goal statements. For smaller-scale projects where the affected party’s interests are relatively similar (e.g., neighborhood groups reconstructing a local creek), comments should be used to develop a consensus on project goals. For most projects this is not practicable, and the consultative process will involve reviewing each comment and incorporating a response into the construction plan. Experience suggests that a flexible, two-step goal development process is most effective. The initial goals are developed following the initial site analysis and discussions with the affected parties. Final goals are selected at the conclusion of the construction design process. Even then, the goals may require modification during the construction, or even post-construction monitoring phase, as new opportunities and constraints arise. Practicability A substantial body of literature exists on the design and goals of wetland construction projects (Kusler and Kentula, 1989). However, a significant gap exists between those who develop the goals and write the plans and those who construct wetlands. This gap mirrors a similar chasm found between landscape architects and landscape contractors and is likely to exist because of education dissimilarities, professional practices, and even insurance requirements. This gap poses a significant danger to the development of ecological construction because it may ensure that important knowledge gained during construction is not incorporated into subsequent designs. ©2001 CRC Press LLC

Table 2

Parties Potentially Affected by Wetland Construction Projects and Goals Common to These Parties Affected Parties

Landowners Site landowner Adjacent landowners Adjacent residents

Permitting and commenting agencies Corps of Engineers Wildlife agencies Public land agencies

Operation and maintenance entities Flood control districts Park districts Mosquito abatement

Goals Minimize costs Increase aesthetics Eliminate off-site issues (e.g., mosquitos, odors, floods) Increase access Reduce vandalism Provide appropriate data Respond to all concerns Restore single species habitat Provide appropriate public access Ensure success Minimize maintenance costs Provide adequate flood conveyance Ensure compatibility with park needs Minimize mosquito production

Building a wetland shares many basic elements with any landscape construction project, whether it is gardening or reforestation. The most important issue in any such project is to determine what conditions create the TVA. Subsequent issues may arise that threaten vegetation association growth, whether one is planting corn, taro, or rushes. Wetlands are among the wettest landscapes and, accordingly, the quantity and quality of the water, as well as the use of soil types which ensure water retention, are extremely important. Also important are those conditions that result from the application of significant amounts of water to the soil such as erosion and salt buildup, which reduce the viability of the TVA. Wetland construction strives to achieve ecological function and presumes that the project, once constructed, will evolve in a natural fashion. Ideally, no maintenance will occur following an initial establishment period making an understanding of the successional aspects of the TVA very important. On the other hand, many wetland construction projects, especially those specifically for waterfowl habitat, will use sources of water that require continued operation and maintenance (Payne, 1992). Unless the project is built in an extremely remote area, human actions will influence the project. Conversely, the project may influence humans. For these reasons, wetland construction must consider the role of people in the design and, in some cases, identify methods for ensuring that potential interactions between the constructed habitat and people are mutually beneficial. These elements are considered in more detail in the Construction Design section of this chapter. However, they must also be considered in the goal setting stage. Ignoring these issues and proceeding without consideration of the practicability of the project may shorten the planning phase of the project but will likely decrease the probability of success.

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CONSTRUCTION DESIGN Geography Size and Shape Market or regulatory forces well in advance of any design analysis often determine the size and configuration of the construction site. Even when strict limitations exist, the designer must consider the appropriate extent and shape for the different vegetation associations to be constructed. Recent work stemming from MacArthur and Wilson’s (1967) research on island biogeography suggests that areas of at least 16 ha contain significantly greater numbers of species of birds, mammals, and invertebrates than do smaller areas (Tilghman, 1987; Faeth and Kane, 1978). Species richness is also positively correlated with edge length and habitat diversity (O’Meara, 1984). Location Exposure to wave energy is particularly important for coastal wetland construction projects. The relationship between wave energy and effects on cordgrass planting can be related to fetch (the distance traveled by waves). At less than one nautical mile of fetch, plantings will be unaffected. Between 1 and 3.5 nautical miles, some replanting is necessary, and over 3.5 nautical miles, some kind of wave-barrier structure is necessary to protect plantings (Crewz and Lewis, 1991). Orientation of the site with respect to colonizing (or invading) plants and animals is another factor influencing design. Exposure to desirable flora and fauna should be provided. Conversely, exposure to exotic and nuisance species should be restricted. Prevailing winds and currents may enhance or deter establishment of volunteers. The presence and condition of adjacent wetlands may also affect design specifics. For example, preserving isolated patches of desirable vegetation may inhibit water circulation by creating berms. Sound ecological judgment must be used in assessing whether existing vegetation impairs or contributes to developing desired functional attributes at the site (Crewz and Lewis, 1991). Slope Perhaps the most critical aspect in successful wetland construction design is grading the soil surface to the elevation that provides the optimal hydrologic regime for the TVA (Broome, 1990). It would not be an overstatement to consider accurate topographic surveys and grading the keystone to meeting wetland construction goals. Slopes should be as gradual as possible, while still allowing good surface drainage at low water conditions. Gentle slopes at the perimeters of the project site are recommended to reduce erosion and filter runoff reaching the site. If steep slopes are necessary, they should be stabilized to serve as a deterrent to exotic or nuisance species invasion. Gradual slopes at the perimeter of construction sites also serve to

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prevent compression of vegetative zones owing to sea-level rise, the rate of which is predicted to increase rapidly after the year 2025 (Crewz and Lewis, 1991). Adjacent Uses Significant opportunities exist for a larger native landscape or a corridor when adjacent uses include natural or restored lands. Whereas a great deal of information has been developed on the value of larger sites, the case for creating corridors is more intuitive. MacClintock et al. (1977), Fahrig and Merriam (1985), and Wegner and Merriam (1979) all provide study results that suggest that birds and small mammals actively use natural corridors, and that patches of native landscape connected by such corridors to larger habitats receive greater wildlife use than isolated patches. Construction projects should consider potential use of the wetland by people and the effects of the human landscape on the natural system. Adams and Dove (1989) document widespread interest in viewing and interacting with wildlife within suburban and urban settings. Providing a trail that offers views and interaction with the wetland is often the best means to ensure visitors do not impact sensitive areas. Leaving the edge of the wetland open so that views of the wetland are created may be the best means to ensure that the wetland is not used for trash and debris dumping. Glare is best treated with tall, rapidly growing trees as a buffer. Off-road vehicle access should be eliminated through the use of post and cable barriers between any public roadways and the wetland, unless steep vertical barriers or moats are provided. Domestic and feral cats and dogs are significant predators on ground-nesting wetland birds. Both can be restrained with fences, and open water is an effective barrier to cats. Hydrology Wetland hydroperiod, including the duration and frequency of inundation or saturation, and depth of inundation, are generally the major determinants of wetland vegetation associations. However, the manner in which the water is supplied to the construction site may be equally crucial. Artificial supply systems, such as drip or spray irrigation and pumps, require a significant amount of maintenance and reduce the naturalness of the wetland. However, these systems eliminate much of the uncertainty involved in water supply planning. Natural water supply systems such as open channels are less manageable but can result in a self-maintaining system that mimics natural systems. Use of natural water supply systems requires a good understanding of ecology and engineering and will require more effort in the investigation and planning stages. Hydroperiod and Depth Two models are useful in estimating wetland hydroperiod and depth of inundation in freshwater wetlands. Slope models are used to define the hydroperiod and depth of inundation for wetlands abutting a river or lake. Basin models are used to model wetlands isolated from riverine or lacustrine systems. For tidal wetlands, a variety of models have been used to describe tidal flows. In any case, tidal ranges ©2001 CRC Press LLC

and important elevations (mean high water, etc.) must be identified; these are generally derived from local tidal gauges and other records. Slope wetlands are characterized by seasonally varying water levels with no implied storage. The relationship between wetland vegetation associations and water surface elevations is well described (Dickson et al., 1965; Harris et al., 1975). However, water surface elevation analyses of streamflows for storms of various intensity or duration are complex. Channelized streamflow equations have been developed for flood control analyses. These equations consider the amount of flow, the crosssectional area and slope of the channel, and a friction element that defines channel roughness. The latter reflects the capability of channel vegetation to reduce flow velocity. The HEC-1 model developed by the Corps of Engineers (1981) is an example of this type of model. Used in conjunction with an accurate topographic map of a site, these models can predict the depth of water and duration of flooding at a specific elevation for a specific storm event. However, these models require accurate data that are typically very expensive to generate. Slope wetland hydrology can often be more simply defined by directed observations of the vegetation associations and of the channel during specific storm events. The locations of wetland vegetation associations in a slope environment relate to water level, which is related to topographic position. Identifying the elevation of a TVA relative to the channel and then predicting the creation of a new and similar association based on a similar channel position is crude but relatively effective (Figure 5). Observations during storm conditions are crucial for determining channel behavior during high flows. Widening it to pre-alteration conditions and thus reestablishing the TVA might easily restore a channelized stream. However, other factors may also be critical. For example, soil or underlying bedrock may be radically different on adjacent reaches. Greater or lesser permeability in the construction reach might result in less or more water than predicted. Other factors such as channel roughness in the form of dense vegetation or in-channel morphological features may slow flows and significantly reduce peak flows. A comparison of stormflow peak and duration in the template and construction reaches can eliminate most sources of error. Basin wetland hydrology is best defined by a water budget. That is, the water available to support the wetland is a function of water inflow minus water outflow. This is illustrated by S = P + SI + GI – ET – SO – GO where S equals the storage in the wetland basin. P equals the precipitation. SI equals the surface inflow. GI equals the groundwater inflow. ET equals the evapotranspiration. SO equals the surface outflow. GO equals the groundwater outflow or infiltration. ©2001 CRC Press LLC

Figure 5

Determining the appropriate elevation of a constructed slope wetland can sometimes be accomplished through observation of existing vegetation associations and the elevations at which the associations occur.

Water sources for wetlands include precipitation falling directly on the wetland, surface inflow, and groundwater inflow. Precipitation amounts should be identified on at least a monthly basis. Data are usually available from local weather stations or airports. In the United States, National Oceanic and Atmospheric Administration (NOAA) handbooks are also a good source of information. Surface inflow is the product of precipitation within the watershed less the amount lost to evapotranspiration and infiltration or inflow from adjacent lacustrine or riverine sources. For inflow from adjacent sources, a stage-discharge model may be used to determine surface inflow or the elevation of the flood flow can be estimated based upon corresponding vegetation association boundaries. Groundwater inflow is often extremely difficult to determine and is best accomplished through the use of piezometers. ©2001 CRC Press LLC

Water is lost from the wetland (and its watershed) through evapotranspiration, surface outflow, and groundwater outflow. Evapotranspiration is often described on a monthly basis in NRCS soil surveys, NOAA handbooks, and other sources. The low point in the basin will control the basin water surface elevations and surface outflow. The average storage capacity is then the remaining volume of water in the basin. Some water flowing into the basin will be lost to outflow through the soil, with the amount of outflow determined by the height of groundwater and the type of soil. Where groundwater levels are high or the soil underneath the basin is at field capacity, no or little loss will occur. Where groundwater levels are low or the soils beneath the basin are not at field capacity, water loss to the soil will be more significant. The amount of loss during these conditions is dependent upon soil type, with clay soils generally appropriating less water than sandy soils. The factors used in a water budget calculation are often dependent upon each other and interaction between the surface and groundwater components can confound the results of the budget analysis (Brown and Stark, 1987). Additionally, many other factors may affect the parameters or their measurement. In all cases, the models should be verified, validated, and calibrated through direct observation of existing or experimental wetlands. Water Supply Providing a relatively natural water supply system, or designing a wetland within a system such as a channel also used for flood conveyance, is relatively simple in tidal conditions but increasingly problematic with greater freshwater inflow. Much experience, often not positive, has been gained from the design of flood channels that also host wetland creation efforts. Concrete or earth-lined trapezoidal channels have been extremely popular in flood control design because they require little land, are relatively inexpensive to build, keep flood flows below existing ground levels, and are easy to maintain. As these systems were expanded to provide for wetland construction, however, the inherent problems associated with this channel design have become apparent. The substantive problem is the perennial saturation of the flat bottom of the channel, which encourages the development of a low-terrace riparian woodland. Woody species tend to grow densely and rapidly, creating a significant impedance to flood flows. One resolution of these maintenance conflicts would be the establishment of a wetland vegetation association with relatively low channel impedance. Riparian woodland associations vary tremendously in their effect on channel impedance but generally follow a pattern of greater impedance with greater affinity for water. That is, a low-terrace woodland impedes flood flows more than a mid-terrace association, which in turn impedes flow more than a high-terrace association. In a study of riparian woodland establishment patterns at several wetland creation projects in the Central Valley of California, Zentner and Zentner (1992) observed that the low-terrace woodland in this region occurs from the upper edge of the summer water level to approximately the mean annual flood line. The mid-terrace zone occurs from the mean annual flood line to about the 10- to 15-year storm line, and the high-terrace woodland

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occurs from the mid-terrace zone to the 100-year flood line. Therefore, to circumvent blockage of flood waters by dense low-terrace riparian growth while still providing for wetlands, broad low terraces could be designed that can accommodate flood flows despite the dense nature of the vegetation. Alternatively, the low-terrace zone could be restricted, and terraces would be constructed just above the mean annual flood line where mid-terrace riparian woodland, in friable soils, or marshes in indurated soils can be maintained (Figures 6 and 7). Additional hydrological factors, including water velocity, sediment loading, and erosion potential, will also be important. Table 3 is a checklist for reviewing channel design issues. The checklist can be used to review both upstream and downstream preconstruction and postconstruction conditions. In the western United States, and most likely for other regions as well, seasonal low water, mean annual flood level, and the 100-year flood level are the most pertinent water surface elevations. Seasonal low water level defines the upper limit of perennial marsh and the lower limit of low-terrace riparian woodland. Mean annual flow approximates the upper limit of seasonal marsh and low-terrace woodland. The 100-year flood level is an important cultural limit, and flows above that level resulting from a wetland construction project will require special consideration. The relevant tidal wetland elevations include mean tide level, mean higher water (MHW), and mean higher high water (MHHW), all important boundaries for vegetation associations (Lewis, 1994). Impedance associated with the vegetation association has an important effect on channel flow, as does channel bottom slope microtopography. Each of these factors, as well as channel soil characteristics, are important factors in erosion and sedimentation. Sandy soils erode more easily than clays under most conditions, and impediments to channel flow such as bridges increase scouring through local increases in flow velocity. Decreases in flow velocity, especially when there is a high sediment load, will result in silting-in of channel bottom basins. For coastal wetlands, some kind of proximal offshore structure may be needed for many sites exposed to moderate wave energy. The protection offered by berms or breakwaters can be substantial and are a critical factor in successful establishment of newly planted vegetation. At some sites, construction of a breakwater may be all that is necessary to stabilize a site, both subtidally and intertidally, and consequently facilitates significant habitat improvement. For example, a small spoil island in Clearwater Harbor, FL, that provided good wildlife habitat for local populations (especially birds) was experiencing severe shoreline erosion. Offshore seagrass beds were also showing signs of stress related to excess wave energy. In 1989, a 140-mlong breakwater of large rocks was constructed approximately 25 m offshore of the island, seaward of the grass beds. Following breakwater placement, the shoreline stabilized through formation of a small beach and the seagrass beds appear healthy and may be expanding in area. Large rocks and rubble are the material recommended for construction of breakwaters in most circumstances. Smaller material may be less stable, or can consolidate excessively, prohibiting adequate tidal flushing. Constructed properly, rock berms used in conjunction with mangrove restoration projects aid in trapping propagules, thereby reducing or eliminating the need for planting.

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Figure 6

To avoid blockage of flood waters by dense, low riparian growth, the low-terrace zone can be restricted and terraces can be constructed above the mean annual flood line.

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Figure 7

Another method for avoiding blockage of flood waters while providing for wetlands is to construct seasonal marsh in low-terrace areas.

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Table 3

A Design Checklist for Addressing Critical Hydrological Design Components of Channel Wetland Construction

Water surface elevations Seasonal low water level Mean annual flood level 100-year storm level Vegetation association High impedance Moderate impedance Low impedance Channel bottom slope On-site Upstream Downstream Erosion and erosion potential On-site Upstream Downstream Sedimentation and sedimentation potential On-site Upstream Downstream

Soil Soil properties have generally received little consideration in the design and construction of wetlands. Exceptions include wetlands where soil characteristics determine the wetland form, such as western vernal pools and bogs. In large part, this historic inattention to soils is due to the predominance of tidal and perennial marsh construction which can be supported to some extent on almost any soil as long as an appropriate hydroperiod and water depth are provided. However, constructing the drier freshwater wetland types, such as seasonal wetlands or wet meadows, has made soil considerations more important. Soil texture and its corollary soil permeability often strongly influence the type of marsh or the development of a woodland in place of a marsh. Hydroperiod and inundation depth being equal, fine textured, less permeable soils support a wetter vegetation association and a generally larger wetland than do course textured, more permeable soils. More permeable and well oxygenated soils (including various grades of rock) will support woodlands in addition to, or instead of, marshes. As noted above, tidal and perennial freshwater marshes can be established on almost any soil type. Soil depth is also an insignificant factor in many instances. Areas with indurated soil layers at the surface have established perennial marsh within three months of inundation as have areas of cobble (Zentner and Zentner, 1989). Nevertheless, wetland vegetation establishment is easier where significant levels of organic material are present to increase soil water retention capacity. In

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several cases, peat or other organic mulches have been added to the substrate and have increased the growth of perennial marsh species (Brown et al., 1984). Seasonal marshes generally require a surface horizon of 10 to 45 cm depth and a soil with some water holding capacity. This is particularly true when root zone saturation is infrequent. However, if the soil is extremely permeable, seasonal marshes can still be established where water levels are periodically within 15 to 45 cm of the soil surface. Generally, these conditions must exist for two weeks per month during the growing season. Seasonal marshes can even be established on cobble but to a lesser extent than perennial marshes. Wet meadows are greatly affected by soil type. Typically, wet meadows occur on clay or clay loam soil which is saturated for 4 to 8 weeks in an average year. Generally, inundation is infrequent but may be long-standing when it does occur due to the dense nature of the soil. Soil depth may also be important for many of the wet meadow species. Some rushes, for example, have a root depth to 1.8 m. Riparian woodlands may establish on a broad array of substrates, but, as with marshes, wetter vegetation associations are less affected by soil type than drier vegetation associations. Accordingly, low-terrace woodland species may establish on thin cobble layers above mud soils, a useful technique for stabilizing rock weirs in wetlands. However, mid- and high-terrace woodland species require a greater depth of permeable soil as noted above. In some instances, soil analyses of a potential construction site may show that wetland establishment is not feasible under natural conditions. Godley and Callahan (1984) document the construction of marsh and riparian woodlands in sandhills of Florida using a combination of a clay liner and organic soils to retain water. Pond wetlands with marsh and woodland components have been constructed in the xeric foothills of central California in cobble piles with extreme permeability using a mixture of clay and bentonite to retain water (Zentner and Zentner, 1990). Despite the apparent success of these examples, all sites must be examined with an understanding that wetland construction on such sites may best be conducted elsewhere. One of the primary functional objectives of any ecological construction is to minimize or eliminate maintenance after an initial establishment period. Projects that depend on continual maintenance are at best unnatural, and often unsuccessful in the long term. Vegetation At its most basic, plant ecology is the study of plants and plant associations and the relationship of plants and plant associations to various edaphic and landscape factors. Because wetland construction seeks to place a specific vegetation association in a specific location, it is basically applied plant ecology. Plant ecology encompasses an extremely broad range of subjects. It can be argued that a multitude of factors affect the landscape position of a plant or vegetation association and, therefore, affect construction design. Primary among these factors are soil and hydrology. Other factors having a significant effect on the establishment and persistence of the TVA are plant succession, planting design, and ruderal invasion.

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Succession Succession refers to the changes that a vegetation association may go through in response to autogenic or allogenic factors. Two forms of succession are recognized: phase and type. Phase succession is said to occur when the soil and hydrology conditions do not change, but the vegetation association changes, usually from a seral to a mature stage. Generally, most vegetation associations have seral and mature phases. The variation between seral and mature phases of marshes is relatively noticeable. Typically, readily seeded species such as cattails dominate the seral stage of a perennial marsh while species more adapted to vegetative propagation dominate the mature phase. Seed viability for these mature marsh species is often low or noncompetitive on bare ground compared with the more adventitious species. As such, plug planting may be the only practicable means of assuring growth of the mature phase during the early years of a construction project. Conversely, identifying the seral stage species of the TVA allows the ecologist the option of not planting these species but instead simply ensuring a seed source. Understanding this phase of the TVA is important and is best accomplished through observation of template sites or experimental plots. Seral and mature stages of woodland vegetation associations are often extremely difficult to identify. Successional models for woodlands suggest that the low-terrace woodland association is the seral stage for the drier vegetation associations. This may be due to the historic predominance of autogenic succession in woodlands that once assured episodic or periodic pulses of sediment through the system. Once these successional forces have been eliminated through flood control, relatively static vegetation associations may result. However, phase succession is still observable in the response of species to canopy openings and, in the propagule stage, to varying degrees of light intensity. Species that do well in full sun when planted in the field or in nursery conditions generally represent the seral stage of the TVA. Conversely, species that require some shade typically comprise the mature stage. A common problem in woodland construction involves the planting of a full complement of woodland species in a newly constructed site, that is, a site exposed to full sun, without consideration of the successional status of each species. Type succession is said to occur when the soil, hydrology, and vegetation association change. Classical models of succession provide for the eventual disappearance of a freshwater wetland as substrate surface elevations increase due to peat or sediment buildup and its replacement by an upland forest community (Clements, 1934). Similarly, riparian woodlands would shift from the low-terrace stage through the high-terrace stage and into an upland woodland or forest. More recent work suggests that succession in many wetlands is a combination of autogenic forces such as peat buildup, which direct it in a linear, Clementsian development pattern, and allogenic forces that may disrupt or reverse this pattern. To some extent, allogenic forces are more familiar to wetland scientists and are revealed by relatively persistent zonation which when disturbed exhibits a predictable series of changes (Mitsch and Gosselink, 1986). In wetland construction design, succession and the external forces that often seem to control wetland change are both important parameters. Succession in ©2001 CRC Press LLC

wetlands often follows the classical autogenic model. Under more or less natural conditions, riparian woodlands become isolated from waterways, or sediment buildup provides for drier conditions, and the wetter association is gradually (or rapidly in some cases) replaced with a less hydric association (McBride and Strahan, 1984). Similarly, perennial marshes may gradually increase in surface elevation to become seasonal marshes or wet meadows. However, the allogenic model also often holds true. Marshes may not transition through woodlands. For example, soil conditions adverse to woodlands, especially duripans or heavy clay layers, favor dense mats of marsh plants. The construction design should also be sensitive to those forces that may be periodic or episodic such as beavers, floods, fires, or debris flows. Cultural factors are also increasingly important in succession. When floods are controlled and the waterway form remains constant, or sediment inputs have otherwise been significantly reduced, wetland associations may remain in place without succeeding to more xeric associations. For example, Klimas (1989) found that highterrace, oak-dominated riparian woodlands were significantly underrepresented in the lower Mississippi River valley compared to the more mesic woodland associations. Typically, riparian woodland associations vary in seed size and weight in direct relationship to their successional stage and location relative to seasonal low water levels. Low-terrace woodland species have the lightest seeds and are the first of the woodland vegetation associations to establish, while the high-terrace species have heavy seeds and are a much later successional stage. Klimas theorized that levees acted as barriers to the dispersal of the heavier seeds, which were not transported by flood waters as easily as the lighter seeds. Similarly, aggregate-rich lands adjacent to the San Joaquin River, CA, were mined over an 80-year period resulting in a series of ponds supported by groundwater in a matrix soil of gravely loam. The soils are very permeable and little to no flooding occurs. A low-terrace willow-cottonwood association, a typical early sere, dominates all the ponds regardless of age (Zentner and Zentner, 1989). Without this broad understanding, a designer reviewing the woodland stands in both regions would tend to underrepresent the upper elevation woodlands (dominated by oaks in both cases) in the planting plan, thereby mimicking an altered vegetation association. The construction design must analyze which set of conditions will govern the construction site and then plan for these accordingly. Planting Design If a site is ideally suited to a particular type of habitat improvement, the decision of whether or not to install plants may present itself. Suitable tidal freshwater sites can rapidly become vegetated in as little as six months (Broome, 1990). On the north-central Gulf Coast, two general types of marsh habitat improvement projects are common. The first is the use of dredged material which may be confined by temporary or permanent dikes. Alternatively, dredged material may be deposited unrestricted in open water. Such projects are rarely or only experimentally planted and are usually allowed to vegetate naturally. The second less common approach to unplanted marsh construction is diversion of river flow. This process provides substrate and the introduction of fresh water counteracts the effects of saltwater intrusion ©2001 CRC Press LLC

and rising sea level. Again, natural colonization is allowed to occur in projects of this type (Chabrek, 1990). The major benefit of natural colonization is the automatic use of local gene stock. However, the time required for natural colonization varies considerably. Good plant establishment can occur within several months in freshwater tidal wetlands but may require several years in more saline conditions (Shisler, 1990). The presence of an adequate seed or propagule source should be verified if natural colonization is anticipated. Another option is the use of wetland soils salvaged from another site. The topsoil (sometimes referred to as mulch) removed from a wetland scheduled for destruction can be stockpiled and later used as a top layer in a wetland improvement project. The effectiveness of this soil as a seed bank depends on the plant species present (exotics and nuisance species may be excessive) and the viability of the seeds. The latter depends on the species, the length of time the topsoil has been stockpiled, and the conditions under which the topsoil was stockpiled. This technique is most effective if the soil is in place at the beginning of the growing season. Salvaged soils appear to be most useful for construction of seasonal marshes, probably due to their relatively high species diversity. At least 20 to 25 cm of soil and plants should be transplanted into the seasonal marsh zone and immediately supplied with the appropriate hydrologic regime. Plant Selection If planting is elected, the species selected will depend on the type of wetland desired. In Florida, one or two species are commonly used for coastal marsh habitat projects with the emphasis on smooth cordgrass. However, with the increasing emphasis on wetland function, complex plantings of multiple species may become more common. For regional guidelines on plant species selection, refer to the reviews in Kusler and Kentula (1989). Crewz and Lewis (1991) provide specific guidelines for species in Florida. As suggested above, construction of a freshwater perennial marsh in a riparian situation may require only inflow water and at least a thin, preferably organic, soil layer. A more complex task is ensuring that the marsh is not overwhelmed by weedy species such as cattails. Planting marsh species representative of the mature perennial marsh in the proposed establishment zone is cost-effective and provides for rapid establishment of a diverse flora. Marsh species should be salvaged in the form of plugs from nearby sites to supplement on-site salvage. Wet meadows may be established using common agricultural practices such as preparation of the seed bed and sowing of seed, or direct planting of small, container species (plugs) into unamended soil. Seed bed preparation (outside of the necessary hydrologic modifications) will consist of weed removal, disking or otherwise cultivating the soil, seeding, and harrowing to cover the seed. For species that do not produce highly viable seed, plugging rooted plants or hydromulching with stolons should be considered.

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Similarly, although it is often neglected in forested wetland construction work, the herb stratum is an important component of the riparian woodland. One to two month old plugs of native species should be planted at one to two plugs per square meter. Additionally, spreading detritus from an existing forest floor may be very helpful as is planting stock inoculated with the appropriate mycorrhizal fungi. Stock Selection The species to be planted and the form of the propagule need to be specified. Depending on the species and the location, seeds, seedlings, vegetative shoots, or transplants may be the most desirable planting unit. In the case of mangroves, propagules are recommended for red mangroves, but seedlings of black and white mangroves are recommended. Smooth cordgrass can be grown from seed but not throughout its range. However, plugs and bare-root culms are easily transplanted, and nursery-grown units from field-harvested stock are easy to cultivate, handle, and install. Seedlings grown from seeds of cordgrass and black needlerush are used in the southeast (Broome, 1990), although transplants of black needlerush are generally unsuccessful. Saltgrass can be grown from transplants, rhizomes, seeds, or plugs. In general, plants grown locally should be used because adaptations to local conditions may create ecotypes. This is especially important with regard to species with wide geographic ranges such as smooth cordgrass. Planting Density Installation on 1-m centers is the standard for most coastal marsh, mangrove, and seagrass restoration projects. However, this is probably not a realistic specification. Naturally colonizing species, mangroves for example, tend to initially occur at much higher initial densities. Density decreases over time, with differential survival of the hardiest plants. Crewz and Lewis (1991) recommend 25-cm centers for red mangrove propagules and 50-cm centers for 1-year-old seedlings. As increasing planting density also increases the cost of a project, there should be a distinction made for mitigation vs. restoration. The importance of obtaining functional equivalency in mitigation should require higher planting density, while restoration can be more economical, relying on natural colonization and a longer period of time allowed to achieve stated objectives. Planting density should be determined according to the desired rate of plant coverage which should be rapid in the case of mitigation for short-term wetland losses. The size of the site may also affect planting density because large sites are often planted at a lower density than a smaller one of the same wetland type, simply for reasons of economics. Reviews of riparian woodland plantings have found average densities of mature stands to vary from 200 to 300 trees per hectare, roughly equivalent to plantings on 2.5- to 3-m centers. Under natural conditions of course, the initial establishment rate is often much greater, especially for the low-terrace woodland species. A sparser planting rate assumes some maintenance or other care in planting that will result in a significantly higher survival rate than under natural conditions.

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Weed Control Weeds are nonnative, adventitious plants that replace desired components of the TVA. The term nonnative is used here to describe a species that occurs outside its historic range or landscape position. Weeds include species such as common reed (Phragmites communis), which has replaced native species in the New Jersey Meadowlands, or melaleuca (Melaleuca quinquenervia), which has invaded south Florida wetlands. Weeds do not include native species, such as cattails, whose growth and spread are promoted by disturbance but which are typically the natural, seral phase of a TVA. In some instances, weed control or elimination can be the sole action taken in a wetland construction project. The spread of adventitious nonnatives throughout the environment threatens many native vegetation associations. However, simply eliminating the exotics will not suffice to restore a TVA. Weed management should identify and address the conditions that allowed or encouraged the exotics to become established, remove the exotic, replace the exotic with the appropriate native plant, and maintain the TVA such that further opportunities for the exotic to become reestablished are limited. For example, a noxious pest in California, bristly ox-tongue (Picris echiodes) has very light seed that does well in heavily grazed, dense, annual grasslands. Grazing leaves small openings in the turf that allow the plant to germinate. Mulching bare soil or ensuring a dense carpet of native groundcover, and then reducing soil disturbance, can effectively eliminate this species. Cultural Issues Mosquitoes Marshes and mosquitoes, as well as the problems caused by mosquitoes, have been synonymous for millennia. The increase in wetland construction projects has been a cause of some alarm for mosquito abatement officials. Mosquito abatement officials will eventually modify marshes that generate mosquitoes that act as vectors for diseases or that annoy nearby residents. Mosquito abatement officials operate under a public health and safety mandate and are generally immune to wetland protection regulations. Mosquito abatement officials are, therefore, particularly interested in influencing wetland construction design so that postconstruction control is unnecessary. Mosquitoes have relatively short growth periods (generally 20 to 30 days) and can fly 2 to 3 km. Those mosquitoes that bite people do not breed in turbid waters, in waters that are too cold or hot, or rapidly flowing waters. Also, insect and fish predators can significantly reduce mosquito populations. Generally, optimal conditions for mosquitoes consist of well-vegetated, shallow (less than 20 cm) still water that is isolated from deeper water that would contain predators. Wetlands can be designed to limit mosquito production (Figure 8). Wet meadow and seasonal marsh areas can be designed to dry by the onset of warm weather. This can be accomplished using the basin model described above to compare evapotranspiration rates with the hydrologic requirements of the TVAs to define weir height. ©2001 CRC Press LLC

This, in turn, determines the water depth. Perennial marshes can be designed as a series of isolated stands adjacent to open water areas that contain mosquito predators. Ponded water areas would include a slide-gated culvert to allow drainage and maintenance of the site should a major outbreak occur. In general, designs should include significant amounts of wet meadow, seasonal marsh, and open water habitat but limited perennial marsh. Large expanses of perennial marsh, as are created for duck clubs, generally also need active annual drainage and intensive management to limit mosquito production if they are near residential areas.

Wet Meadow Seasonal Marsh Perennial Marsh Open Water

Berm/Dam

Berm/Dam

Winter Water Level Summer Water Level

SM WM

Figure 8

Slide Gate

OW PM

KEY KEY: Wet Medow (WM) Seasonal Marsh (SM) Perennial Marsh (PM) Open Water (OW)

Mosquito production can be limited by designing wet meadow and seasonal marsh areas to periodically dry out, by limiting perennial marsh, and by providing a permanent open water refuge for mosquito predators.

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Water Quality Untreated stormwater from urban areas is likely to contain pollutants, especially metals, in such concentrations that water quality objectives are exceeded in downstream receiving waters. Many, though not all, pollutants contained in urban runoff are sediment associated. Controls to remove sediments, such as settling or infiltration basins and filtration devices, are recommended in design manuals as the most practical manner to control discharges of pollutants in urban runoff (Stahre and Urbonas, 1989). Settling basins should be designed to retain a specific level of storm event in the basin for a 24- to 48-h period. Storms between the mean annual and 10–year storm levels are often used to define the basin size. Infiltration basins should be underlain by a soil that allows percolation at a rate sufficient to trap sediments in the interstitial area for the same storm level. Inflow structures for these basins should be designed so that erosion and short circuiting through the basin are minimized. Outflow structures should be dual low flow and high flow, allowing the passage of extreme events while ensuring that first flush waters receive the extended detention or infiltration time needed to allow settling. Low flow structures used in these basins should include trash racks to prevent clogging.

IMPLEMENTATION The project designer and the contractors should meet and agree upon a specific order of operations and responsibilities prior to beginning any construction project. The order of operations set forth in Table 4 is believed to be the most efficient method of constructing a woodland or marsh project. However, changes in the order of operations may be necessary to allow for weather conditions or other factors. It is imperative that the constructed wetlands, as well as any wetlands to be preserved, are not disturbed by construction activities. This includes using wetlands as right of ways to transport equipment or as equipment storage areas. Communication and coordination between wetland construction operations and any other construction operations are essential. Access, turning on or off utilities, stockpiling materials on project sites, or any situation that might affect either operation should be resolved in advance. Table 4

Recommended Sequence of Construction Events for Wetlands Construction

Install protective flagging and stakes to preserve or salvage areas Remove all invasive exotic plants and on-site garbage Salvage all specified plants Grade construction site to specifications Dispose of excavated materials at designated fill sites Prepare woodland planting basins, seed beds, and otherplanting areas Install water supply or irrigation system and test Plant in designated zones Provide protection of construction area Prepare as-builts and construction report

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An Ecological Monitor (EM) should be designated to ensure that the wetland is constructed in accordance with the approved construction plan. The EM should be responsible for certifying that the project has been built in accordance with the approved construction plan and permits, and should monitor all operations occurring in the wetland areas. The EM should also attend any preconstruction conferences and meetings during construction which are pertinent to the wetland. Among the factors that expedite project implementation is obtaining all necessary permits issued prior to scheduled site preparation and planting. There can be several regulatory agencies involved in the wetland construction project and the specific requirements of each must be met. These obligations typically include assurance that nearby habitats will not be disturbed and provision of means to control siltation during construction. The planting schedule needs to be coordinated with plant availability, regardless of whether plants are field-collected or nursery-grown. This also needs to be scheduled for the optimal time of the year for planting the species involved. Construction and final grading should take place well in advance of the optimum planting dates because several weeks are required for the settling of fill material (Broome, 1990). Generally, planting should occur when temperature, rainfall, and salinity are moderate. For example, in the southeast, planting should occur between the first of April and mid-June (Broome, 1990). In central and south Florida, this period can be extended to mid-March through September. Predicted tides and other hydrologic conditions should be checked to avoid excessively high or low water conditions. Construction should also be timed to avoid disturbance to seasonally nesting or migratory wildlife. Construction Sequencing Construction should proceed in an orderly sequence for most projects (Table 4). This sequence begins with the installation of protective flagging for any areas to be preserved and includes removing invasive species, salvaging desired species, grading, planting, and fencing. Production of as-built plans is the final operation. Construction of a water supply may be necessary some instances. Protective Flagging All existing habitats that will remain undisturbed, those which will be salvaged within the general work area, and all travel ways should be staked and flagged prior to any grading activities. Flagging should be highly visible and replaced as necessary so as to be continuously in place during construction activities. Stakes should be at least 1 m tall and painted an easily observable color such as fluorescent orange. Flagging should extend around the protected area perimeters or along their entire lengths to prevent any unwanted intrusions or damage. Accidentally damaged plants or areas should be repaired. All fencing and flagging should be removed at the end of the construction project.

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Weed Removal Mowing or scraping requirements for weed removal will vary from site to site depending on the amount and type of weeds and the amount of grading. Typically, areas of noxious weeds should be excavated and the infested soil deposited elsewhere. Alternatively, the weeds should be disked, watered to encourage seed growth, and disked again. The latter should be repeated until weed growth is extremely sparse. Chemical controls may also be used. Areas with little or no grading will require at least a thorough mowing so that debris will not hamper construction. A flail-type mower should be used and all debris thoroughly mulched. Salvaging All appropriate plants and soil on site should be salvaged for replanting. Salvaging of wetland soil, and hence the seed bank, can be extremely important for construction success. Dunn and Best (1983) found that a representation of plants from all successional stages may only occur where a seed bank has been placed in the marsh. Erwin et al. (1985) found much higher species richness and cover in marshes created with salvaged topsoil than with subsoil. However, seeds of woody species may be absent or underrepresented in seed banks (Leck 1989). The collection and broadcasting of woody seeds or the planting of propagules can effect timely development of riparian woodland associations. Salvaged plants should be kept in the shade and watered if excavation has not proceeded to the point where they can be planted immediately (within 48 h of excavation). The viability of salvaged plants declines significantly after 45 days. Additionally, marsh plants may be salvaged from selected locations outside the project site. Salvaging should occur manually or with a trenching or similar narrowbucket backhoe, and should remove no more than 20 percent of any one stand. Grading Studies of wetland construction projects completed between 1960 and 1980 show that improper grading and grading management are two of the most important factors contributing to failure (Garbisch, 1977; Zentner, 1989). The EM is most useful during this stage and will often be required to make quick decisions when unanticipated issues develop. For example, while excavating a basin for wetland construction in Folsom, CA, the EM noticed a slightly discolored soil area. The discoloration proved to be an unobserved sand lens which, if left untreated, would have resulted in complete drainage of the proposed marsh after inundation. Similarly, large organic debris may be encountered during excavation and should be incorporated by the EM into the project. The EM should recognize that few preconstruction soil surveys are sufficient to precisely define subsurface conditions. Grading requirements are defined almost entirely by measurement of elevation change. These measurements require reference to established benchmarks. Nearby structures may have recorded elevations that can be obtained from local government agencies. At least two benchmarks should be used when establishing a site-reference ©2001 CRC Press LLC

elevation. As an alternative, local plant populations can be used as a guide to proper elevations. Juveniles or seedlings of the desired species provide the most accurate information because they are more sensitive to elevation differences than adult specimens. The site should be surveyed after grading and prior to plant installation to ensure that the specified characteristics have been attained. Topographic surveys should be completed and reviewed before construction equipment is removed from the site, so that slope and elevation can be efficiently adjusted. Shisler (1990) recommends preparing a bathymetric or topographic contour map on 15 cm (0.5 ft) contours after fill has settled but prior to planting. The map also provides a base for as built surveys and monitoring efforts. If the site has been filled or excavated, sufficient time for the substrate to settle should be allowed prior to determining and adjusting final elevations. This is especially important for tidal sites. The site may be susceptible to ponding and resuspension (turbidity), and the substrate will not provide adequate support to young plants, if planting occurs too soon after earthwork. Planting As noted previously, plant material may be installed as container stock, bare root seedlings, cuttings, canes, plugs, or seed. The first four are typical for woodland plantings, plugs are most often used in marsh projects, while seed is common for both woodland and marsh. Container plants generally range in size from small seedlings to 3.75 l (1 gal) or similar-sized stock. Smaller stock is less expensive to grow and plant while larger material can provide an immediate presence to increase buffer capabilities and aesthetics. Wetland or wetland buffer conditions are rarely as protected as nurseries, and larger stock will require additional care during the first year after planting to reach its potential. All plant material used should be collected from within the immediate region. Considerations as to the proximity of the gene pools and planted species vary greatly from one plant species to another. Plant material should be free from disease, insects, and weeds, not be rootbound, and identified correctly as to genus and species. All plant species and quantities should correspond to the planting plan and should be inspected prior to planting. Container plants should be collected, propagated, and grown for at least one growing season prior to planting. Stock should be planted in a hole twice as deep and wide as the root ball (Figure 9). The hole should be backfilled with good quality soil so that the root collar of the plant is level with the surrounding soil level. Slowrelease fertilizer may be added to each planting hole prior to planting the seedling. A water-conserving polymer may also be useful in arid regions. The root ball should be covered with at least 1 cm of soil to prevent a wicking effect from drying out of the root ball. In many irrigated cases, construction of a basin will help hold water near the plant and ensure adequate irrigation. The basin typically consists of a 0.6 m diameter (or larger) water ring 5 cm deep with a surrounding berm 5 cm above grade centered around the plant. Shredded bark or similar mulch should be placed on watering basins to a depth of 5 cm making sure not to cover the crown of the rootball. A weed mat may be used to further reduce weed growth but should then be covered ©2001 CRC Press LLC

with mulch to prevent overheating of the near-surface soil. Weed mats generally reduce weed growth but increase construction costs significantly. In addition, the weed mats trap moisture and facilitate fungal growth that can lead to reduced plant viability. Where planting basins are eroded, dirt can settle on top of the weed mat and form a substrate for weed growth. When drip irrigation is used, a plant emitter should be placed directly on top of the root ball. The plant should be watered thoroughly after installation and checked 2 to 3 days later for settling and stress.

Figure 9

Survival of planted stock can be maximized through careful attention to planting, watering, and fertilizing.

Most container plants will benefit from an augured planting hole when riparian woodland planting occurs in low permeability soils. For example, 3.75 I propagules with deep tap roots generally require holes augured to a depth of 1.2 m using a 23 cm diameter auger. Holes should be responsive to the plant form. For example, deep tap root species will use deeper and narrower holes than shallow-rooted species. ©2001 CRC Press LLC

All holes should be backfilled by the end of the same workday and should be completely saturated to promote settling of excavated soil. Bare-root seedlings have been a staple of the fruit tree industry for some time and are now commonplace in constructed wetlands. Bare-root seedlings are essentially large container stock without the containers. Instructions for container stock are applicable to bare-root material. Special care should be taken that the material is planted in a hole deep enough and wide enough to prevent recurving or J-rooting the tap root. The use of dormant cuttings in low-terrace woodland construction is wellrecognized. Cuttings should be at least 2 to 5 cm in diameter and 1 m in length, although these dimensions will vary somewhat by region. At one extreme, construction work in the xeric southwest may involve auguring deep beneath the surface and the planting of poles up to 6 m long in attempts to reach subsurface water (Anderson et al., 1984). At the other extreme, planting the same species in the more mesic Pacific Northwest, which is characterized by less permeable soil, may require a pole cutting 30 cm in length or less. All cuttings should be taken from young wood, be free of disease, 2/3 of the cutting should be pounded into the ground, and the soil should be tamped around the cutting. The top should be cut off at a 45 degree angle to facilitate growth. Cuttings must be kept moist and either can be planted the same day they were cut or left in water for 4 to 5 days to harden off. Cuttings are probably the most inexpensive way to quickly establish riparian woodland cover. Canes are effective planting stock for many vines. They resemble cuttings in that they should be nearly dormant and leafless, 60 to 75 cm long, and in healthy condition. A shallow trench 3/4 of the cane length should be made and the cane laid in the trench and largely covered with one end exposed above the soil. The trenches should be oriented on the contour, that is, perpendicular to the slope. Plugs are small clumps of herbaceous plant material and associated soil. Plugs are easiest to salvage with backhoes and excavators, although they are also grown in containers. Each salvaged plug should be about 30 by 30 cm, and should retain as much substrate as possible (at least 10 to 20 cm deep). Containerized plugs are generally much smaller, 3 cm across and 6 to 8 cm deep. Plugs should be planted so that the bases of the stems are level with the ground surface in inundated to saturated soil. All seed used on the site should be certified as to germination percentage and should be weed and disease free. Seeding rates and ratios should follow specifications developed by the grower. The EM should inspect all seed, soil preparation, calibration ratios, and phases of seeding operations. Specific soil preparation and seeding techniques should be adopted for each site and outlined in each site’s planting plan. Generally, seeding will work best when it follows local agricultural practices. Extensive seed bed preparation (disking, cross-disking, harrowing, and cultipacking) may not be applicable on all sites but will provide good germination at relatively low cost. Hydroseeding or hydromulching with stolons may be applicable where dense plantings or site conditions make tractor or hand seeding and seedbed preparation impracticable. Hydroseeding is more costly and good access to the planting site is required.

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The optimal time for seeding operations is during the dormant period. Seed broadcast early in the dormant season generally results in greater seed emergence than seed broadcast later in the dormant season. In arid regions, one key to success is to establish healthy, growing roots prior to the summer drought, so that the roots may follow the subsurface moisture level downward through the soil as summer progresses. The actual means of plant installation depends on the habitat type, the planting units used, and the substrate. On intertidal sandy substrates, mechanical means (e.g., tobacco planter pulled by a tractor) can be used, while on soft or muddy substrates, planting must often be conducted by hand working off of wooden mats or rafts. Transplanted material should be removed from the donor site so conditions conducive to invasion by exotics or attractive to potentially destructive fauna are not created. For example, large open patches are attractive to Canada geese and invite root predation by avian and mammal populations (Shisler, 1990). The use of fertilizer varies by species. The use of a slow-release fertilizer for riparian woodlands has been noted above. The effects of fertilizer on mangroves are unclear, but fertilizer has been shown to accelerate the growth of cordgrass (Broome, 1990). Fertilizer and lime can be used to provide short-term adjustments to substrate nutrient status and pH. Although plants may respond positively to small amounts of fertilizer at planting, accommodation to predominant long-term substrate conditions is probably desirable (Crewz and Lewis, 1991). When used, a time-release fertilizer incorporated into the substrate at the time of planting is recommended. Broadcasting fertilizer over the site is expensive, often increases weed cover, and only delays the achievement of nutritional equilibrium. Some coastal projects will require temporary protection of newly installed plants from excessive wave action during a preliminary establishment period. This can be accomplished by using floating tire breakwaters, earthen dikes, sandbags, and erosion control fabric. Although unattractive, floating tire breakwaters have the advantage of being reusable once plant establishment has occurred. The Riley encased methodology (Riley et al., 1999) is a promising technique for establishing mangroves in moderate to high energy environments. The method places seedlings in slotted PVC pipes inserted into the substrate in the intertidal zone. Water Supply Construction and placement of water control structures are relatively simple provided certain issues are resolved. First, review the placement of the structure prior to emplacement. Design analyses are important, and care should be taken to ensure that the location and design make sense in the field before blindly following construction plans. Second, ensure that the structure is firmly sealed into the surrounding soil. Water will tend to erode soil around the structure if proper contact is not provided. Third, check the structure during daily, seasonal, and episodic flows. Finally, remember that any structure placed in a waterway will require continual maintenance. Plan ahead to provide the appropriate financing and maintenance reminders. Enclosed supply systems (drip irrigation) may be required during the initial establishment period to assure the success of new plantings. Spray irrigation is not ©2001 CRC Press LLC

recommended owing to its propensity to encourage weeds. Systems connected to a potable water supply will require a reduced pressure backflow assembly inspected and tested by a certified backflow technician. Nonpotable supplies will require filtering, usually several large filters in sequence. All irrigation lines should include a tracer line (#14 UF direct burial for easy identification of line location). Generally, drip irrigation systems consist of a main line of polyvinyl chloride (PVC) tubing 5 to 15 cm in diameter running to valves that control the flow. From the valves, secondary or PVC lateral lines, 2.5 to 5 cm in diameter, transport water to local points where polyethylene (poly) tubing (2.5 cm in diameter) provides water directly to the plants. Before passing into the poly tubing, all water should be filtered and the pressure reduced. Smaller poly tubing, known as spaghetti tubing, may be hooked into the larger poly lines for further movement of water. However, spaghetti tubing is difficult to maintain and not recommended for most construction projects. In buffer strips along streets and other visible areas, irrigation should be below grade. In more secluded areas, mainlines and valves should be below grade, but poly lines may be installed above grade. The latter reduces trenching costs and eases maintenance. Jute hooks or 15 cm staples should be used to secure poly tubing in place. Staples should be placed every 3 m on a run and 30 cm on either side of each planting. All lines should be thoroughly flushed before inserting emitters, and each planting receives an emitter. Emitters should be placed directly over the root ball because the density of the root ball and indigenous soil may be quite different. Pressure testing upon completion of the main line and valves should be required; lateral lines should be left on for a period of at least 2 h to visually inspect all leaks. Upon completing the irrigation system, the EM should make a final walkthrough to verify the system is operating correctly. Fencing All habitat areas adjacent to streets should be separated from the roadways by a post and cable or similar barrier. In areas where maintenance access for vehicles is necessary, locking bullards should be installed in accordance with local specifications. Fencing during the construction period may also be necessary. As-Builts The EM should be responsible for overseeing preparation of grading, planting, and irrigation as-builts when the project is completed. The planting of as-builts should include the location, genus and species, and container size of each propagule. The EM should also certify that the project has been built in accordance with the plan and transmit the as-builts and certification to the appropriate parties.

MAINTENANCE Popular thought would have it that native plants are pre-adapted to the growing conditions of a site and should not require substantial maintenance to reach maturity. ©2001 CRC Press LLC

However, site conditions in the United States and much of the rest of the world have changed so radically over the past two centuries that site conditions often no longer reflect the conditions under which native species developed. Good examples are the wet meadows of the Southwest. Prior to European settlement, these meadows were dominated by perennial graminoids. With the arrival of Europeans came an invasion of annuals that almost extirpated the native perennials. Over the past 200 years, invading annuals and introduced livestock have succeeded in producing a substantial seed bed of annuals, formed thick thatch-dominated near-surface conditions, and fostered the growth of small and large mammal populations. These changes in the western environment have made construction of native meadows much more difficult than simply clearing a patch of ground and spreading some seed. Vigorous maintenance efforts must be carried out to keep the annuals from re-invading the native meadow. Additionally, construction is not a natural process, but rather it entails artificially inseminating a site with plant material. Under natural conditions, the plants that reach 10 cm are the few survivors of several thousands of seeds in a favorable microsite at a time when conditions are relatively optimal for their growth. The project proponent rarely has the luxury to wait or the knowledge to define these conditions. Consequently, the species to be planted are unlikely to be propagated naturally and may not even be planted in a site for which they are now adapted. Maintenance activities will be needed to ensure that the vegetation association develops as a native vegetation association, without invasion by harmful exotic species. The array of tasks required for any one project may not match the list provided in Table 5, but each project is likely to require some form of maintenance. Although maintenance is required, it should be viewed as a short-term measure, not only owing to its costs but because long-term maintenance suggests that the TVA is ill-suited for site conditions. Consequently, effective wetland construction will entail a good design to ensure the best possible match between the selected species and a site, and manipulation of the site to improve planting and growth conditions. Maintenance, even intensive maintenance activities, should generally be limited to the first three years after the cessation of planting. If a site is not relatively self-sufficient after this time, either additional construction work will be required or the design must be reconsidered. Weed Control In most cases, careful design and preplanting weed control procedures will reduce the burden of weed abatement considerably. However, a freshly graded construction site is a haven for weeds that may be carried in from off site. The ecologist must make a distinction between ruderal species that are part of natural succession and those species that will impede development of the TVA. The latter, particularly nonnative, aggressive species can be addressed using manual or chemical controls. The ecologist must also ensure that the appropriate native species replaces the weed, and the conditions that resulted in weed growth will not reoccur. Manual controls include hand removal of the entire plant or cutting the plant above ground. Hand removal of the entire weed is labor intensive and is best suited ©2001 CRC Press LLC

Table 5

Wetland Construction Maintenance is an Essential Step for Ensuring Project Success Problem

Weeds Annual Perennial Woody Erosion Sheet Scour Slump replant Herbivory Post-planting Grazing Plant care Nutrients Form Water supply Pest/diseases Irrigation Mortality Litter Trash

Tools

Survey Frequency

Manual Chemical Hand removal

1 to 3 times per year

Redirect flow Regrade

Post-storm and annually

Screens

Once

Fertilizer Staking Basin rework IPM/spray

As needed

Replacement

Monthly

Patrol

Monthly

* Each project should be assessed for the problems listed and appropriate remediation should be initiated.

for woody plants. The advantage of hand removal is that the area of impact is usually quite small. Less specific weed control activities can create the type of disturbance which results in more weed growth. It is often much less expensive to cut the weed and then spot spray the stump with an herbicide. However, for conditions that militate against the use of chemical controls, an array of tools are available from suppliers that can be used to selectively remove even mid-sized trees without disturbing the adjacent land or plants. Cutting can include hand-pruning or chopping, weed whips or similar hand-held but gas-powered tools, and tractor-driven or floating mowers. Selection of a manual control method will depend on the type and extent of the weed and the project design. Hand pruning or chopping is labor intensive and costly unless the personnel consist of volunteers. Prior to selecting this method, the ecologist should personally weed a specific area for at least 2 h and develop an estimate of the amount of personhours required, multiplying this by the cost of the labor. Where weeding will occur in dense growth, or where only a few plants among many are to be eliminated, this method is the most productive. Weed whips and similar tools are very useful for clearing annual or soft perennial growth over large or small areas where tractors cannot be driven (e.g., slopes, marshes). However, these should not be used near planted trees and shrubs or aboveground irrigation lines as even an experienced practitioner can mistakenly girdle a plant or cut a poly line. Mechanical mowers are the most efficient means of cutting large areas of vegetation. However, they cannot run on steep slopes (although some have mowing booms that can reach out and cut limited areas) and require a certain distance between trees and shrubs to be preserved. ©2001 CRC Press LLC

These issues should be considered during the design phase. An easily maintained project is generally subject to less disturbance and may prove to be more successful. Timing is often crucial in manually controlling weeds, especially annuals. Because cutting does not remove the root systems, it is generally important to cut the plant just prior to seed set to maximize the likelihood of complete control. Even this technique may not always work. Mowing of the annual yellow star thistle (Centaurea solstitialis) on clay soils just before seed set almost eliminated this noxious weed on one site. By contrast, mowing of another population on rocky soils (where this species thrives) resulted in a second flowering several weeks later. Chemical controls are often the least expensive means of eradicating or reducing weed growth. However, very few chemicals are licensed for use in wetlands and a licensed technician is often required by local regulations. Further, environmental conditions (e.g., wind speeds) never seem to be as optimal as required and spraying often inflicts damage outside the target area. Chemical controls include pre-emergents that are applied to the ground prior to the emergence of the weeds in the spring and post-emergents that kill or stunt plants during their active growth stages. Preemergents are most useful when applied adjacent to newly planted propagules because they limit the growth of competing weeds. Post-emergent controls are most useful for broad areas completely dominated by weeds, areas where no plants are proposed (e.g., firebreaks), or spot-spraying of individual weeds. Surveys for weed control needs should start at the beginning of the first growing season following construction. Weeding frequency will depend on the weed species and seasonal growth cycles. Annuals may need controls applied as much as three times in a year, whereas woody plants can generally be removed once per year. Erosion Control Erosion control is best effected by good design. However, one of the truisms of any construction project is that site conditions are not completely predictable. Unseen channel irregularities, inclusions of different soil types, and similar events will require modifications to the design during construction in order to minimize subsequent maintenance efforts. Erosion in marshes or the channel banks of riparian woodlands often occurs as sheet erosion, scouring, or slumping. Sheet erosion takes place on relatively flat slopes and generally results in the loss of all or most of the new plantings. This typically is due to poor seed bed preparation, inadequate root penetration prior to the storm, a subsurface soil layer that inhibits root growth, or a combination of these factors. The remedy is to try again. If possible, channel the flowing water into a swale rather than across the slope. Scouring occurs on moderate to steep slopes where higher velocities rapidly remove soil to some depth. Scouring is usually the result of modifications that have increased velocities through the site or re-oriented high velocity flows. Remedies are cause specific, but look for forces that have increased upstream flows or reduced downstream elevations. Slumping occurs when the surface soil on a channel bank moves downward as a relatively intact unit, often because it has been undercut. The movement will take ©2001 CRC Press LLC

anything on top of the soil unit with it including plantings or irrigation lines. Remedies are, again, site specific, but may be corrected by light grading and replanting after the erosive stimulus has been corrected. It is not unusual for the design to underestimate the flow velocities at specific points in a channel. Water supply structures, especially those perpendicular to the channel, can be undermined or the adjacent side slopes eroded. Heavy rock interplanted with fast-growing trees may eliminate future erosion when only a few centimeter of soil are lost between the structure and the soil. More significant erosion will require replacing the structure or identifying and rectifying the cause of the erosion. Reviewing the project site for erosion should occur immediately after the first three typical storms and each time thereafter for storms of greater intensity. Erosion controls are best applied annually, either immediately following the storm event that caused the erosion or during the appropriate planting or grading period. Herbivory Grazing has been and may be used to manage a TVA. However, the effect of herbivores on target plants is not always beneficial. The most common remedy for deer, rabbit, or beaver grazing is to provide wire screens around the plants or planting area. These screens can be relatively costly to provide, adding as much as US $1.00 to the cost of each plant. Consequently, screening most efficiently might be applied after a problem has surfaced. Screens in or adjacent to waterways should be removed prior to high water levels or floating material will catch in the screen. This may create enough mass that the screen, debris, and the plant will wash away. Monitoring for herbivory should occur weekly during the spring of the first year after planting, and less frequently thereafter. Plant Care Container-grown trees and shrubs in woodland construction sites may be fertilized once in the spring of the first and second year after planting if nutrient levels are low on the site. Fertilizer should be dribbled into the basin around the propagule. Trees may also need to be staked if wind or other forces are warping or bending the trees unnaturally. To stake a tree, place two 2.5 m pressure-treated stakes, 5 cm in diameter, 60 cm into the ground, 30 cm on either side of the tree. The tree should be bound to the stakes using tree tie tape fastened in a figure-eight pattern, 30 cm below the top of the stake. Basins around propagules should also be maintained to supply sufficient irrigation water. Insect or disease infestations are common on habitat construction projects. Generally, all infestations should be treated using Integrated Pest Management (IPM) techniques. Plantings of a single species are especially susceptible to massive outbreaks that are not well treated by IPM, however, and commercial sprays may be required. Monitoring for these problems should be completed weekly during the first year of the project with remedial action taken immediately.

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Irrigation System Maintenance The irrigation system should be checked regularly and frequently (e.g., weekly) immediately following planting for malfunction and vandalism. All drip lines and filters should be flushed as needed. Emitters must be checked regularly for malfunctions, position of emitter over root ball, and vandalism. Flushing end caps must also be checked regularly. In colder climates, backflow devices and mainline systems should be completely drained at the end of the growing season to prevent rupture due to freezing. Irrigation for plantings is a temporary tool to help establish the constructed vegetation association. Water should be decreased every year, and ideally should not be needed by the fourth year. Caution should be exercised not to create lush plants that would become dependent on regular watering. Deep watering should be spaced at the longest time possible before plants show signs of stress. This type of watering will encourage plant roots to travel downward into the appropriate water zones in the soil. Litter Removal Wetlands are usually depressions in the landscape and seem to collect litter. Litter can impede the growth of propagules or seedlings and increase the cost of plant care maintenance efforts. Therefore, each site is thoroughly policed and all litter and debris removed from the site at least annually. Construction projects that are near or within residential areas generally collect the most trash. Over 181 kg of litter was collected from one project in central California in 2 months time, at a cost of 42 labor h and US $350 in dump fees. Experience suggests that projects in residential areas will require at least 1 to 2 h per week for litter collection. General Maintenance Frequency Most projects will require a walk through by an ecologist or maintenance technician every 1 to 3 weeks for the first 1 to 2 years. The precise number of visits depends on the size of the site, its proximity to disturbance, and the time of year. Visits reasonably may be required every week for the first year and twice monthly for the next 2 years. Smaller sites, for example those less than 4 ha, can reasonably be visited once every 2 to 3 weeks during the first year, and monthly in subsequent years. Minimizing Maintenance Efforts Local education can effectively reduce maintenance efforts associated with vandalism and litter. Efforts should be aimed at informing neighbors as to the project goals, values, and needs, and in enlisting their support for the long-term protection of the site. Conducting a replanting project for local third through sixth graders at least once every 2 years has proven helpful. Small brochures for each project that are mailed to local residents can also be beneficial. Preparation of a maintenance ©2001 CRC Press LLC

manual in the last year of monitoring for long-term maintenance of the project area also is effective.

RESEARCH NEEDS Many issues related to the enhancement, restoration, and creation of wetlands would benefit from the conduct of basic and applied research. This is especially true for coastal wetlands. These include an increased knowledge about species’ life requisites. The physical and chemical properties of wetlands, including topography, soils properties, subsidence rates (which vary locally), and hydraulic circulation, are often poorly understood. For wetlands designed to improve local water quality, information is needed on the effects of treatment on other wetland functions.

REFERENCES Adams, L. W. and Dove, L. E., Wildlife Reserves and Corridors in the Urban Environment. National Institute for Urban Wildlife, Columbia, MD, 1989. Anderson, B.W., Disano, J., Brooks, D. L., and Ohmart, R. D., Mortality and growth of cottonwood on dredge-spoil, in California Riparian Systems: Ecology, Conservation and Productive Management, Warner, R. E. and Hendrix, K. M., Eds., University of California Press, Berkeley, 1984. Broome, S. W., Creation and restoration of tidal wetlands of the southeastern United States, in Wetland Creation and Restoration: The Status of the Science, Kusler, J. A. and Kentula, M. E., Eds., Island Press, Washington, D.C., 1990, 37. Brown, M. T., Gross, F., and Higman, J., Studies of a method of wetland reconstruction following phosphate mining, in Proceedings of the Eleventh Annual Conference on Wetland Restoration and Creation, Webb, F. J., Ed., Hillsborough Community College, Tampa, FL, 1984. Brown, R. G. and Stark, J. R., Comparison of ground-water and surface-water interactions in two wetlands, in Wetland and Riparian Ecosystems of the American West, Mutz, K. M. and Lee, L .C., Eds., Proceedings of the Eighth Annual Meeting of the Society of Wetland Scientists, Society of Wetland Scientists, Wilmington, NC, 1987. Chabrek, R. H., Creation, restoration, and enhancement of marshes of the north central Gulf Coast, in Wetland Creation and Restoration: The Status of the Science, Kusler, J. A. and Kentula, M. E., Eds., Island Press, Washington, D.C., 1990, 125. Clark J. R. and Benforado, J., Eds., Wetlands of Bottomland Hardwood Forests, Elsevier, Amsterdam, 1981. Clements, F. E., The relict method of dynamic ecology, Ecology, 2, 39, 1934. Cowardin, L. M., Carter, V., Golet, F. C., and LaRoe, E. T., Classification of Wetlands and Deepwater Habitats of the United States, FWS/OBS-79/31, U.S. Fish and Wildlife Service, Washington, D.C., 1979. Crewz, D. W. and Lewis, R. R., An evaluation of historical attempts to establish emergent vegetation in marine wetlands in Florida, Florida Sea Grant College Technical Paper No. 60, Florida Sea Grant College, University of Florida, Gainesville, FL, 1991. Dane, C. W., Succession of aquatic marsh plants in small artificial marshes in New York State, NY Fish Game, 6(1), 57, 1959. ©2001 CRC Press LLC

Dickson, R. E., Hosner, J. F., and Hosley, N. A., The effects of four water regimes upon the growth of four bottomland tree species, For. Sci., 11(3), 299, 1965. Dunn, W. J. and Best, G. R., Enhancing ecological succession: seed bank survey of some Florida marshes and role of seed banks in marsh reclamation, in Proceedings of the 1983 Symposium on Surface Mining, Hydrology, Sedimentology, and Reclamation, University of Kentucky, Lexington, 1983. Erwin, K. L., Best, G. R., Dunn, W. J., and Wallace, P. M., Marsh and forested wetland reclamation of a central Florida phosphate mine, Wetlands, 4, 87, 1985. Faeth, S. H. and Kane, T. C., Urban biogeography: city parks as islands for Diptera and Coleoptera, Oecologia, 32, 127, 1978. Fahrig, L. and Merriam, G., Habitat patch connectivity and population survival, Ecology, 66, 1762, 1985. Garbisch, E. W., Recent and planned marsh establishment work throughout the contiguous United States: a survey and basic guidelines, U.S. Army Corps of Engineers, Waterways Experiment Station, Vicksburg, MS, 1977. Garbisch, E. W., Highway and wetlands: compensating wetland losses, Federal Highway Administration Report No. FHWA-IP-86–22, U.S. Department of Transportation, 1986. Godley, J. S. and Callahan, R. J., Creation of wetlands in a xeric community, in Proceedings of the 10th Annual Conference on Wetland Restoration and Creation, Webb, F. J., Ed., Hillsborough Community College, Tampa, FL, 1984. Harris, R. W., Leiser, A. T., and Fissell, A. T., Plant tolerance to flooding, University of California at Davis, Department of Environmental Horticulture, Davis, CA, 1975. Josselyn, M., Zedler, J., and Griswold, T., Wetland mitigation along the Pacific coast of the United States, in Wetland Creation and Restoration: The Status of the Science, Kusler, J. A. and Kentula, M. E., Eds., Island Press, Washington, D.C., 1990, 3. Klimas, C. V., Limitations on ecosystem function in the forested corridor along the lower Mississippi River, in Wetlands and River Corridor Management, Kusler, J. A. and Daly, S., Eds., Proceedings of a symposium presented by the Association of Wetland Managers, Charleston, SC, July 5–9, 1989. Kuenzler, E. J., Mangrove swamp systems, in Coastal Ecosystems of the United States, Vol. I, Odum, H. T., Copeland, B. J., and McMahan, E. A., Eds., The Conservation Foundation, Washington, D.C., 1974, 346. Kusler, J. A. and Kentula, M. E., Eds., Wetland Creation and Restoration: The Status of the Science, EPA/600/3–98/0, EPA, Corvallis, OR, 1989. Larson, J. S., Bedinger, M. S., Bryan, C. F., Brown, S., Huffman, R. T., Miller, E. L., Rhodes, D. G., and Touchet, B. A., Transition from wetlands to uplands in southeastern bottomland hardwood forests, in Wetlands of Bottomland Hardwood Forests, Clark, J. R. and Benforado, J., Eds., Elsevier, Amsterdam, 1981. Leck, M. A., Wetland seed banks, in Ecology of Soil Seed Banks, Leck, M. A., Parker, V. T., and Simpson, R. L., Eds., Academic Press, San Diego, CA, 1989. Lewis, R. R., Mangrove forests, in Creation and Restoration of Coastal Plant Communities, Lewis, R. R., Ed., CRC Press, Boca Raton, FL, 1982, 153. Lewis, R. R., The restoration and creation of seagrass meadows in the southeastern United States, in Proceedings of the Symposium on Subtropical Seagrasses of the Southeastern United States, Durako, M. J., Phillips, R. C., and Lewis, R. R., Eds., Florida Department of Natural Resources Mar. Res. Pub. No. 42, St. Petersburg, FL, 1987, 174. Lewis, R. R., Wetlands restoration/creation/enhancement terminology: suggestions for standardization, in Wetland Creation and Restoration: The Status of the Science, Kusler, J. A. and Kentula, M. E., Eds., Island Press, Washington, D.C., 1990, 417.

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Lewis, R. R., Enhancement, restoration, and creation of coastal wetlands, in Applied Wetland Science and Technology, Kent, D. M., Ed., Lewis Publishers, Boca Raton, FL, 1994. MacArthur, R. H. and Wilson, E. O., The Theory of Island Biogeography, Princeton University Press, Princeton, NJ, 1967. MacClintock, L., Whitcomb, R. F., and Whitcomb, B. L., Evidence for the value of corridors and the minimization of isolation in preservation of biotic diversity, Am. Birds, 31, 6, 1977. McBride, J. R. and Strahan, J., Establishment and survival of woody riparian species on gravel bars of an intermittent stream, Am. Midland Nat., 112(2), 235, 1984. Mitsch, W. J. and Gosselink, J. G., Wetlands, Van Nostrand Reinhold, New York, 1986. National Research Council, Restoration of Aquatic Ecosystems: Science, Technology, and Public Policy, National Academy Press, Washington, D.C., 1992. O’Meara, T. E., Habitat island effects on the avian community in cypress ponds, Proc. Am. Conf. SE Assoc. Fish and Wildlife Agencies, 38, 97, 1984. Payne, N. F., Techniques for Wildlife Habitat Management of Wetlands, McGraw-Hill, New York, 1992. Riley, R., Kent, D. M., Salgado, C., and DeBusk, T., The Riley encased methodology for establishing mangroves, Land Water, May/June, 48, 1999. Shisler, J. K., Creation and restoration of coastal wetlands of the northeastern United States, in Wetland Creation and Restoration: The Status of the Science, Kusler, J. A. and Kentula, M. E., Eds., Island Press, Washington, D.C., 1990, 143. Stahre, P. and Urbonas, B., Stormwater Detention, Prentice-Hall, Englewood Cliffs, NJ, 1989. Tilghman, N. G., Characteristics of urban woodlands affecting breeding bird diversity and abundance, Land. Urban Plan., 14, 481, 1987. Tomlinson, P. B., The Botany of Mangroves, Cambridge University Press, Cambridge, U.K., 1986. U.S. Army Corps of Engineers, HEC-1 Flood Hydrograph Package: User’s Manual, Water Resources Support Center, Hydrologic Engineering Center, Davis, CA, 1981. Wegner, J. F. and Merriam, G., Movements by birds and small mammals between wood and adjoining farmland habitats, Appl. Ecol., 16, 349, 1979. Zentner and Zentner, Ball Ranch mitigation program, Report prepared for the Sienna Corporation, Walnut Creek, CA, 1989. Zentner and Zentner, Natoma Station monitoring report, Report prepared for the U.S. Army Corps of Engineers, Sacramento District, Walnut Creek, CA, 1990. Zentner and Zentner, Lower Laguna Creek drainage master plan: mitigation program, Report prepared for Sacramento County, Walnut Creek, CA, 1992.

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Rolband, Michael S. et al “Wetland Mitigation Banking” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

7

Wetland Mitigation Banking Michael S. Rolband, Ann Redmond, and Tom Kelsch

CONTENTS Background Regulatory Context The Banking Process Types of Wetland Mitigation Banks Perspectives on Mitigation Banking Ecological Perspective Regulatory Management Perspective User Perspective Economics Demand for the Product Service Area Regulatory Climate Service Area Size Mitigation Ratios Performance Requirements Monitoring and Maintenance Requirements Permitting Difficulty Attitudes about Mitigation Alternatives Regulatory Stability User Requirements Competitive Supply of the Product and Product Alternatives On-Site Opportunity for Wetland Mitigation Off-Site Opportunities for Wetland Mitigation Other Wetland Banks In Lieu Fee Alternatives

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Risk Assessment and Presale of the Product Revenue Projections Costs of Mitigation Bank Development Land Costs Hard Costs Soft Costs Long-Term Stewardship Economic Projections References

Mitigating the environmental impacts of necessary development actions on wetlands and other aquatic resources is a central premise of wetland regulatory programs. Offsetting losses through the restoration or creation of replacement wetlands has been promoted as a way to achieve a goal of no net loss of remaining wetland resources while still permitting unavoidable impacts to occur. As evidenced by recent studies, however, the effectiveness of on-site compensatory mitigation efforts has produced mixed results. Success rates range from 27 to 50 percent, due in part to 22 to 34 percent of the mitigation projects never being built (Redmond, 1991; Gallihugh, 1998; DeWeese, 1994; Brown and Veneman, 1998). In response to problems associated with individual mitigation efforts, there has been growing interest in the concept of mitigation banking. Mitigation banking refers to the restoration, creation, enhancement, and, in certain circumstances, the preservation of wetlands, for the purpose of compensating for multiple wetland losses in advance of development actions. It typically involves the consolidation of small, fragmented wetland mitigation projects into one large contiguous site. Units of restored, created, enhanced, or preserved wetlands are expressed as credits which may subsequently be withdrawn to offset debits incurred at a project development site. Mitigation banks provide greater flexibility to landowners needing to comply with mitigation requirements and can have several advantages over individual mitigation projects. To the advantage of permit applicants, mitigation banks may reduce permit processing times and provide more cost-effective compensatory mitigation. Most permit applicants do not wish to become wetland experts, but rather they are simply seeking authorization to move forward with their development projects. Through the purchase of credits from an approved mitigation bank, these applicants can transfer the responsibility for providing mitigation to an entity who has the expertise, resources, and incentive to ensure that the mitigation is ultimately successful. Mitigation banking also enhances the effectiveness of wetland protection programs. The environment benefits from consolidation of compensatory mitigation into a single large parcel, or contiguous parcels, that maximize the opportunity to successfully restore important wetland functions. Establishment of a mitigation bank often involves financial resources, planning, and scientific expertise not practicable to many project-specific compensatory mitigation proposals. Consolidation of resources can increase the potential for the establishment and long-term management ©2001 CRC Press LLC

of successful mitigation. Also, mitigation banking typically ensures that compensatory mitigation is implemented and functioning in advance of project impacts. This reduces temporal losses of aquatic functions and uncertainty over whether the mitigation will be successful in offsetting project impacts. Finally, consolidation of compensatory mitigation within a mitigation bank increases the efficiency of limited regulatory agency resources. The review and compliance monitoring of mitigation projects is improved and, thus, agency ability to ensure the success of efforts to restore, create, or enhance wetlands for mitigation purposes is improved.

BACKGROUND The concept of mitigation banking in the United States dates back to the early 1980s when resource agencies, and some in the regulated community, were looking for ways to mitigate wetland impacts more efficiently and effectively. A 1988 U.S. Fish and Wildlife (USFWS) report profiled 13 mitigation banks that were in existence at the time (Short, 1988). Many of these banks were established by enterprising individuals who saw the opportunity to establish joint partnerships to protect and restore priority wetlands using funds from ports, transportation agencies, and others who needed to offset unavoidable impacts. These early efforts were initiated in the absence of any federal or state policies on how to establish mitigation banks. In 1991, in response to a request by Congress, the Corps of Engineers Institute for Water Resources, in collaboration with the Environmental Law Institute and others, initiated a comprehensive study of mitigation banking. The purpose of the study was to determine the potential of mitigation banking for achieving established wetland goals and to determine the applicability of mitigation banking to the U.S. Clean Water Act Section 404 regulatory program. The study included a critical review and evaluation of existing mitigation banks and an analysis of the economic, policy, and other institutional issues affecting banking (Reppert, 1992). The study identified 40 mitigation banks in existence, and another 60 banks that were under development or being considered for approval. An increase in the development of mitigation banks from 1988 to 1992 was the result of state departments of transportation recognizing the ecological, economic, and administrative benefits of consolidating mitigation efforts. The increase in banks was also instigated in part because the 1991 Intermodal Surface Transportation Efficiency Act specifically authorized the use of federal funds for such purposes. Virtually all of the existing banks identified in the Institute for Water Resources study were single-user banks—banks established by a public agency or private company to satisfy their own mitigation needs. Of those banks under development, however, the survey identified several commercial banks whose intent was to offer mitigation credits for sale to the general public. Local agencies, private entrepreneurs, and joint ventures between government agencies and private entities sponsored the commercial bank proposals. The Environmental Law Institute in a 1994 study also noted the trend toward commercial banks.

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Another important finding of the mitigation banking study was the need for a specific policy providing ecological, economic, and legal standards for the establishment and use of banks. A policy would reduce uncertainty, and in so doing, encourage further investment. At the federal level, the Army Corps of Engineers (Corps) and the U.S. Environmental Protection Agency (USEPA) first acknowledged the potential role for mitigation banks in the Section 404 regulatory program in their 1990 Memorandum of Agreement (MOA). In discussing options for providing compensatory mitigation, the memorandum indicates that use of mitigation banks may be acceptable where a bank has been approved by the agencies. In November 1995, the Corps, USEPA, USFWS, National Marine Fisheries Service, and Natural Resources Conservation Service issued the Federal Guidance for the Establishment, Use and Operation of Mitigation Banks. This policy statement details the terms and conditions under which the agencies may approve a mitigation bank for use as compensatory mitigation within the Section 404 regulatory program and the Swampbuster provisions of the Farm Bill. Mitigation banking has been endorsed by both the Bush and Clinton administrations within their comprehensive plans for reforming federal wetland programs. Moreover, Congress has entertained several legislation proposals to promote the use of mitigation banks. In 1998, Congress passed a new transportation bill (TEA-21) that provides further support for the use of mitigation banks to offset wetland impacts that result from transportation projects. In addition to the federal policy, approximately 20 states have established, or are in the process of establishing, policies on mitigation banking. While many of these policies are generally consistent with the federal policy, each is tailored to the unique regulatory requirements of state wetland legislation and is responsive to particular regional conditions. The interest in establishing mitigation banks appears to be increasing, owing in part to the release of federal and state policies. In 1994, the Institute for Wetland Resources identified 46 existing wetland banks in the United States. Only 1 of these 46 was a privately owned bank offering credits to the general public (Environmental Law Institute, 1994). By 1998, the Corps identified over 200 mitigation banks that were either approved or under agency review. Of these 200, approximately 40 existing banks and 75 proposed banks were private commercial banks (unpublished Institute for Wetland Resources survey). Other banking trends include the increased use of mitigation banks as a watershed management tool and the use of mitigation credits for other environmental programs such as endangered species and water quality programs.

REGULATORY CONTEXT In the United States, Section 404 of the Clean Water Act establishes a program to regulate the discharge of dredged or fill material into waters of the United States, including wetlands. Activities typically regulated under this program include fills for development, water resource projects such as dams and levees, infrastructure

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development including highway and airport construction, and conversions of wetlands to uplands for farming or forestry. In 1990, the Bush Administration implemented the Memorandum of Agreement between the Department of the Army and the Environmental Protection Agency Concerning Determination of Mitigation under the Clean Water Act Section 404(b)(1) Guidelines (USEPA, 1990). The Mitigation MOA establishes criteria that must be satisfied before a dredge and fill permit can be obtained from the Corps. First, all practicable steps to avoid wetland impacts must be undertaken by evaluating less damaging project alternatives. Second, applicants must minimize potential impacts to wetlands. And finally, all remaining, unavoidable impacts must be offset through compensatory mitigation. Compensatory mitigation includes the restoration of historic or degraded wetlands, the enhancement of functions of existing wetlands, the creation of new wetlands from uplands, or, in exceptional circumstances, the protection of existing wetlands through acquisition or conservation easement. As clarified in the Mitigation MOA, there is a preference for compensatory mitigation to be located on-site or as close to the impact site as possible. In this way, environmental impacts to local flooding, water quality, fish and wildlife habitat, and other public interests are minimized. In addition, there is a preference that mitigation be in kind, that is, of the same habitat type as the wetlands to be impacted to ensure the mitigation provides similar functions and values. These preferences notwithstanding, the Mitigation MOA also identifies mitigation banking as an option for offsetting unavoidable impacts. On August 24, 1993, the White House Office on Environmental Policy issued a comprehensive plan for reforming federal wetland programs. Regarding mitigation, the plan acknowledges that the aforementioned sequential criteria constitute a logical, predictable, and reasonable framework and that mitigation banking is appropriate in some circumstances. The plan suggests Congress should endorse banking as a compensatory mitigation option under the Section 404 regulatory program. As an outgrowth of the interagency wetland plan, on November 28, 1995, the Corps, the USEPA, the Natural Resource Conservation Service, the USFWS, and the National Marine Fisheries Service issued a Memorandum to the Field titled “Federal Guidance for the Establishment, Use and Operation of Mitigation Banks” (U.S. Army Corps of Engineers et al., 1995). The “Guidance” encourages mitigation banking as an alternative under the Mitigation MOA. The “Guidance” states that mitigation banking is appropriate when compensation for permitted impacts cannot be achieved at the development site or would not be as environmentally beneficial. Under the terms and conditions of the Guidance, applicants for a Section 404 dredge and fill permit may seek approval from the Corps to compensate for unavoidable impacts through the purchase of credits from an operational mitigation bank. In such circumstances, the Corps may approve use of the mitigation credits where onsite mitigation is not practicable (i.e., available and capable of being done) or use of the bank is environmentally preferable to other mitigation options. Moreover, the agencies have established a general preference for using mitigation bank credits to offset minor impacts associated with activities authorized under nationwide and other general permits. Nationwide and general permits are designed to facilitate decision making on relatively small-scale projects and for impacts typically associated with ©2001 CRC Press LLC

linear projects such as road development or utility line installation. Upon authorization by the Corps, the legal responsibility for providing mitigation is transferred to the mitigation bank sponsor through the sale of mitigation credits. In addition to regulation at the federal level, some states have promulgated regulations on mitigation banking (Table 1). Regulation by states has the effect of applying the philosophy and needs of the region to the practice. There are no cases (as of April 1999) where state law alters the existing federal regulations. In all cases, mitigation banking is one tool in the mitigation toolbox. The Banking Process A bank sponsor, who proposes to establish and operate a mitigation bank, initiates mitigation banking. Pertinent regulatory agencies form a mitigation bank review team (MBRT) to work with the bank sponsor. In the United States, the mitigation bank review team typically consists of an interagency group of federal, state, tribal, and/or local regulatory and resource agency representatives. The sponsor will discuss the concept with the MBRT in a pre-application meeting, thereby providing early feedback for the banker as to whether the concept appears viable. The MBRT members may also have knowledge of the project area that will assist the banker in its decision to move forward with the concept. For example, there may be a proposal to build a wellfield proximal to the proposed bank site that would likely affect hydrologic restoration. Alternatively, there may be a recent or pending designation of the area as being of significant conservation interest, to which the bank can then contribute. A prospectus describes the proposed project, and in permitting parlance would constitute a permit application. In the early stages, the prospectus will normally be presented at a conceptual to moderate level of detail, depending on the region or project. Once the proposed bank has been deemed appropriate, then the final details are developed and provided to the MBRT. An accepted prospectus becomes a Mitigation Banking Instrument (MBI). The MBI is developed by the bank sponsor, in consultation with the MBRT, and submitted to the MBRT for review and approval. The role of the instrument is to authorize the mitigation bank project. The instrument includes a preamble describing the project and sections regarding the establishment, operation, maintenance, and monitoring of the project. If a Section 404 permit is also required, the permit is issued as a separate authorization. Because of the importance of these projects, all MBIs must be publicly noticed. The Institute for Water Resources (1996) has developed a model MBI. The MBRT continues to oversee the project once the mitigation bank has been approved and implemented. The operational phase of the project will continue for years, until the project has been declared ecologically successful and transitions to the long-term management phase. During this phase, the sponsor preserves the project site, completes the physical site work, markets, sells, or uses the credits, and monitors and maintains the site. The MBRT reviews the monitoring reports, performs compliance inspections, approves the release of credits, and approves the use of credits for the offset of permitted impacts. The mitigation bank is then debited an ©2001 CRC Press LLC

Table 1

Examples of State Wetland Banking Regulations

State

Wetland Banking

Mitigation Bankers

Florida

Department of Environmental Protection and the 5 regional water management districts (Chapter 62342, Florida Administrative Code)

Private or public

Hackensack Meadowlands District, New Jersey

Interagency Compensatory Wetland Mitigation Agreement (Corps, EPA, NJDEP, Hackensack Meadowlands commission, NMFS and FWS)

Private or public

Louisiana

State’s Coastal Zone

Private or public

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Use of Credits

Mitigation Type

Credits

Management

Land

Some advance credits may be available; credit releases are made with increasing function at the site; some credits are withheld until success Possible to obtain umbrella agreement for establishment and operation of multiple bank sites; some advance use of credits may be approved; preservation credits available once site is preserved Preservation and financial assurances are required for advance credits; preservation is for 20 yr for marshes and 50 yr for forested wetlands

Restoration of native, pre-existing habitats is preferred

Credit assessments are made using a functional assessment methodology

Perpetual management is required and must be endowed prior to the sale of credits

State lands are prohibited for use as mitigation banks

Wetland restoration, creation or enhancement may be used; uplands within the bank site may be assigned credit

Compensation amounts for the offset of impacts is decided on a case by case basis

Endowed perpetual management is required

Public or private lands may be used

Wetland restoration, creation, enhancement, and protection may be used

Credits are assessed using a functional assessment procedure; compensation amounts for the offset of impacts is decided on a case by case basis

Public or private lands may be used

Table 1 (continued) State

Examples of State Wetland Banking Regulations

Wetland Banking

Maryland (nontidal)

Title 26, Subtitle 23, Code of Maryland Regulations (COMAR 26.23.04.06). Nontidal Wetlands Mitigation Banking Act

Maryland (tidal)

Department of the Environment (COMAR 26.24.05.01.B(9) in consultation with local, state, and federal agencies Federal guidance; state provides its assent by signing the Mitigation Banking Instrument

North Carolina

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Mitigation Bankers

Use of Credits

Mitigation Type

Credits

Private or public

Up to 50% of the credits may be available for use in the first two years after construction when the site has been preserved in perpetuity, construction is completed, and the bonding requirements (private banks only) are met; subsequent credit releases are made based on demonstrated increases in function at the site No specific regulations on mitigation banks

Based on rations with a 50% increase in the ratio currently required of the mitigation, uses a wetlands bank; ratios vary by Cowardin Classification and wetlands location

Private

Mitigation via payment to a trust fund for those permit applicants requiring a 401 water quality certification

No specific state provisions governing credit release, perpetual management, financial assurances, service areas, or credit assessments for impacts

Management

Land State lands may be used

The state regulations require “adequate, dedicated financial surety” exists for the perpetual land management

Public or private lands may be used

Table 1 (continued)

Examples of State Wetland Banking Regulations

State

Wetland Banking

Virginia (nontidal)

Does not specifically regulate wetlands banking; regulations establish maximum service area size for banks approved in accordance with applicable federal and state guidance, laws, or regulations Virginia Marine Resource Center (VMRC) and Virginia Institute of Marine Science (VIMS)

Virginia (tidal)

Washington

Federal guidance, state rules expected by December 1999

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Mitigation Bankers

Use of Credits

Mitigation Type

Credits

Department of Environmental Quality (DEQ) authorizes permits to use nontidal wetland banks

Uplands within the bank site may be assigned credit; wetland restoration, creation, enhancement, and protection may be used

Private or public

Sequencing is required prior to authorizing the use of an approved bank as compensation from permitted wetland losses

Generally habitat restoration or creation; also enhancement or in exceptional circumstances, preservation

Private or public

Use of credits prior to meeting all performance standards is allowed

Wetland restoration, creation, enhancement, and preservation may be used, though restoration is preferred

No specific state provisions governing credit release, perpetual management, financial assurances, service areas, or credit assessments for impacts Credit assessments are made using the Function Specific Credit Calculation methodology; performance standards are used to determine credit availability and bank success Long-term management and financial assurances are required

Management

Land Public or private lands may be used

Provisions for longterm management and maintenance are required

The state Department of Ecology and local governments will be signatories to the banking instruments

Public or private lands may be used

amount representing the loss of wetland functions at the impact site. The permit applicant financially compensates the bank sponsor in exchange for allowing its bank to be debited. The bank should be protected and managed over the long term after it has been declared successful. The bank must be legally protected from other future land uses through a conservation easement or similar mechanism. The sponsor is responsible for assuring the financial stability of the bank project over the long term. This necessitates appointment of a long-term manager. This may be the sponsor or another entity, such as a public land manager or environmental organization. Types of Wetland Mitigation Banks The type of sponsor and the operational use can characterize wetland mitigation banks. Sponsors include governmental or quasi-governmental agencies, nonprofit organizations, conservation groups, and private for profit companies. Operational uses include single user and open market sales. The two major types of banks are dedicated banks and commercial banks. Dedicated banks are typically created to compensate for a specific type of activity by a single entity. Dedicated banks include industrial banks—banks created by agreement or permit to mitigate for a specific user impact in a geographic area by a private company. An example of an industrial bank in the United States is the Tenneco LaTerre (Reppert, 1992), a bank sponsored by a private corporation, Tenneco LaTerre, for the purpose of mitigating wetland losses occurring from its oil and gas exploration activities in Louisiana coastal marshes. Another example is the Sunrise Valley Nature Park, a bank that consolidates mitigation for impacts on various separate parcels owned by Mobil Land Development Company in Reston, VA. Another type of dedicated bank is a public works bank. State highway departments, port authorities, or local governments for the purpose of providing mitigation for public works projects sponsored 75 percent of the 46 banks identified by the Corps Institute for Water Resources in 1994. Commercial banks are established by entities whose wetlands credits are available for purchase on the open market by unrelated entities who need wetlands compensation for permitted wetlands impacts. Several types of commercial banks exist. Private entrepreneurial banks create wetland credits and sell them at a profit sufficient to produce an economic return commensurate with the risk undertaken. Nonprofit banks provide mitigation for various activities to achieve a particular ecological and/or economic objective. Credit prices are set so that the sponsor can recover only the costs of the mitigation. For example, the Ohio Wetlands Foundation, a nonprofit group formed by members of the Ohio Homebuilders Association, created a mitigation bank so that builders could have a source of wetland credits and solve an industry problem in a specific geographic area. Governmental banks are banks established by a governmental agency (typically a local government agency) to create mitigation and cover its costs from sales to mitigate for impacts by other governmental or private entities.

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PERSPECTIVES ON MITIGATION BANKING Ecological Perspective Mitigation banking offers several ecological advantages compared to other mitigation vehicles. Many of the regulations governing mitigation banking require that sites be managed over the long term. Therefore, an important evaluation criterion of the mitigation bank planning and approval process is the long-term viability of the site. Long-term viability must be assessed within the context of expected changes in the site's landscape setting. An assessment of this type typically receives less attention in other mitigation planning. Mitigation banking typically consolidates many smaller mitigation obligations into one moderate to large site. Larger sites are more likely to provide significant wetland functions and values than smaller sites. Because banks are typically larger projects than project-specific mitigation, they often include adjacent upland communities. This results in a natural mosaic of upland and wetland in the landscape, increasing the function and value of the wetland and adjacent areas. Also, the relatively large size of mitigation banks facilitates their contribution to watershedbased planning efforts. Mitigation banks typically provide compensation in advance of impacts, meaning that there is little temporal loss in function and value. By contrast, traditional onsite or off-site mitigation is typically implemented or functioning after project impacts occur. This is particularly true of in lieu fee programs, wherein impacts are offset by payment to a management program or conservation organization. As a consequence, temporal losses of functions and values routinely occur. In cases when mitigation fails, function and value losses may be permanent. One disadvantage of mitigation banks is that they do not replace lost functions and values at the point of impact. As such, in cases where on-site mitigation would be viable, that option should be given strong consideration in advance to reduce the local ecosystem impacts. Certain wetland functions, such as flood storage and attenuation and water quality, may not be transferable off site. In some states, such as Florida, nontransferable functions and values are handled separately to ensure their appropriate resolution. Regulatory Management Perspective Mitigation banks can also enhance the effectiveness of wetlands protection programs. U.S. federal guidance mandates a team approach to the assessment of mitigation banking proposals. This approach brings the regulatory and commenting agencies to the same table, with the result that all agencies can contribute their expertise and perspective to the project review. This results in a more meaningful review of the bank proposal and can act to shorten the processing time. The review of one large mitigation proposal is more efficient than the review of numerous, project-specific mitigation proposals. A potential disadvantage of this process is the ability of the team to meet. Limited budgets can limit travel, and coordinating

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multiple schedules can be difficult. Reaching consensus on the terms and conditions of the MBI, as is encouraged by the Federal Guidance, can also be a difficult and time-consuming process. Establishment of a mitigation bank involves financial resources, planning, and scientific expertise at a level not practicable to many project-specific mitigation proposals. Consolidation of resources increases the potential for the successful establishment and long-term management of the mitigation project. Finally, consolidation of mitigation within a bank increases the efficiency of limited agency resources for inspection and compliance monitoring. This improves the agency’s ability to ensure the success of efforts to restore, create, or enhance wetlands for mitigation purposes. User Perspective Most permit applicants do not wish to become wetland experts but are seeking authorization to move forward with their development projects. Mitigation projects may be implemented by people whose main expertise and incentive lies in an unrelated area, such as housing development or highway construction. By contrast, mitigation banks are typically sponsored by organizations whose staff is devoted to implementing successful mitigation projects. A successful banking team provides the permit applicant with the talent and expertise necessary to ensure that the environmental goals set forth in a permit will be achieved. Mitigation banking offers several other advantages to the user. The mitigation plan already has been approved by the agencies, so the time required reviewing and approving a permit is substantially reduced. Purchasing credits from an approved mitigation bank transfers the liability for mitigation success from the user to the banker. The use of credits from a mitigation bank is generally cost competitive with on-site mitigation. By contrast, the costs of implementing successful on-site mitigation may be considerable and open-ended.

ECONOMICS The primary economic issues faced by the sponsor of a mitigation bank resemble those of any business with a product for sale, particularly the real estate development industry. The product a bank sponsor offers is not necessarily a functioning and valuable wetland, but a mechanism that allows a bank user to impact wetlands elsewhere. The goal of the user is more likely to be houses, highways, utility corridors, or some other development or development-related activity. However, mitigation banks are somewhat unique in that the sale of credits produces parks and greenways that provide societal benefits to people other than the bank user. Several factors must be assessed in order to determine the viability of a mitigation bank (Table 2). Each of the factors that contribute to the economic viability of a wetland mitigation bank can be quantified. However, these factors are extremely variable and volatile, which is why the wetland mitigation banking industry, in its current state of development, is extremely risky. One major element of risk is that ©2001 CRC Press LLC

bank demand, product quality, product alternatives, and sometimes price are established or strongly influenced by regulations and policies. These regulations and policies may change frequently and often vary from project to project and user to user. Table 2

Factors That Should Be Assessed in Order to Determine the Viability of a Mitigation Bank

Competitive supply of the product and product alternatives Risk assessment and presale of the product Capital and operating costs of product development Long-term stewardship of the product

Demand for the Product The demand for mitigation credits is driven by three primary factors: service area, regulatory climate, and user requirements (Figure 1). Service Area The service area is the geographic area in which a particular mitigation bank can compensate impacts to wetlands. Ecological concerns (e.g., mitigating in the same watershed where the impact occurred) tend to constrain the appropriate size of a wetland mitigation bank's service area. Economic concerns dictate that the larger the service area, the more likely it is that the bank will experience a sufficient level of demand to be economically practicable. This is because the larger the service area, all other factors being equal, the greater the number of expected wetland impacts. Regulatory Climate Regulatory climate is a term often used to describe how businesses perceive the difficulty of conducting business in a particular location because of local, state, and federal government regulatory agencies. If the regulatory climate is so restrictive that no user in a service area can obtain a permit to impact wetlands, there will be little demand for wetland mitigation credits (Figure 2). Similarly, the regulatory climate may impose mitigation conditions so burdensome (e.g., wetland replacement ratios, performance monitoring, etc.) that impacting wetlands is not cost-effective. Conversely, mitigation requirements may be so minimal that users can satisfy requirements without purchasing credits from a mitigation bank. Thus, the regulatory climate strongly influences the demand for credits generated by economic activity users within a service area. Regulatory decisions in several key areas determine the perception of regulatory climate (Table 3). Service Area Size Free-market advocates suggest that service areas should be unlimited, or at least very large, to encourage maximum levels of competition. This competition could ©2001 CRC Press LLC

Figure 1

Demand for mitigation credits vs. service area size, regulatory climate for obtaining permits to impact wetlands, and the level of prospective user impact to wetlands in the service area.

lead to lower prices, and if the quality of created wetland credits was linked to mitigation ratio requirements, to higher quality wetland compensation. However, the current regulatory climate appears to be focusing on limiting service areas to relatively small areas. If the service areas are too small, the economic practicability of wetland banks will be eliminated. For example, the state of Maryland has recently (1993 to 1998) averaged about 15 hectares (ha) of wetland impacts per year and has established 20 service areas. A mitigation bank is unlikely to be economically viable. Mitigation Ratios Mitigation ratios are established by regulators to compare functions and values of the impacted wetland to the mitigated wetland. The expected likelihood of success

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Figure 2

Regulatory climate factors vs. demand for mitigation credits.

Table 3

Regulatory Decisions in These Key Areas Determine the Perception of Regulatory Climate for Mitigation Banking

Service area size Mitigation ratios Performance requirements Monitoring and maintenance requirements Permitting difficulty Attitudes toward compensatory mitigation alternatives and determinations of the practicability of avoidance of impacts Regulatory stability

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(i.e., a safety factor) and the temporal loss of functions and values (i.e., the time lag between impact and successful mitigation site completion and maturity) are also considered. In other words, the appropriate (mitigation) compensation ratio is based on the level of functional replacement provided, the speed at which functional replacement is achieved, and the risk that the compensation wetland will not perform as expected (King et al., 1993). Some regulators grant lower mitigation ratios to mitigation banks in recognition of their reduced temporal loss, reduced risk of failure, and greater function and value relative to smaller, discontiguous sites. Other regulators grant the same mitigation ratio to mitigation banks as to other types of mitigation. In at least one instance (Maryland), the mitigation ratio is greater for banks than for other types of mitigation. The latter has the effect of discouraging the development of a mitigation banking industry. Another aspect of risk involves the predictability of the mitigation ratios. Some regulators are moving toward specific ratio standards for particular types of impacts. Others are moving toward function and value assessment methodologies that are repeatable, predictable, and useful for all mitigation efforts. This minimizes staff demands, maximizes predictability, and ensures fairness for both users and mitigation providers. However, other regulators seem to be moving in the opposite direction by relying on the best professional judgment of the permit project manager for each specific project. This can lead to unpredictable or biased decisions and consume more staff and banker resources. Regulatory decisions to penalize or reward mitigation banks with higher or lower mitigation ratios are a key indicator of the regulatory climate and are the key determinants of the demand for credits in a service area. The predictability of mitigation ratios by the bank sponsor affects the economic risk and marketability of the bank credits. Once the sponsor of a bank has estimated the projected wetland impacts in a service area over a given time frame, the mitigation ratio must be used to estimate mitigation credits (i.e., mitigation ratio times projected wetland impacts equals the mitigation credits). If the mitigation ratio cannot be predicted, it is difficult to sell the product to a user. This is particularly true if competing products, such as contributions to a trust fund, have a specified payment rate or mitigation ratio. Unpredictability of the mitigation ratio also makes it difficult to develop a sound business plan and economic model for project financing. Performance Requirements The demand and cost of creating mitigation credits are affected by the nature of wetland performance requirements. The risk of obtaining performance requirements is related to the specificity of the requirements. For example, if a forested wetland is required to have 12 5-cm-dbh trees and 160 seedlings per ha by the end of the third growing season, the cost can be projected relatively accurately. However, if performance requirements are less specific (e.g., provide forested wetland vegetation), bank construction costs could reasonably be reduced 75 percent by only planting seedlings. The risk is high, however, that a future compliance inspection will determine that performance requirements are not satisfied, and the bank will

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not be permitted to sell credits. The inability of the bank to guarantee credits, or credits at a fixed price, will lead potential bank users to develop their own mitigation sites rather than purchase bank credits. Experience has shown that providers of sitespecific mitigation rarely suffer hardship due to poor performance. A linkage between performance requirements and mitigation ratios is logical and could be structured so as to provide economic incentives to achieve the desired outcome of mitigation quality. Unfortunately, at this time, many regulators are not supportive of this practice. Historically, a significant shortcoming of traditional wetland mitigation has been the lack of specific performance requirements. This has led to the low success rate of mitigation projects and a significant number of mitigation projects never being initiated (Redmond, 1991; DeWeese, 1994; Brown and Veneman, 1998; Gallihugh, 1998). Monitoring and Maintenance Requirements The demand and cost of creating mitigation credits is affected by the nature, specificity, and duration of monitoring and maintenance requirements. At this time, there are no national or regional standards for these requirements, and extreme variability between projects and locations has been noted. For example, monitoring duration often varies from 5 to 10 years, and the number of vegetation sample plots from 0.4 to 8 or more per ha. The sponsor of a proposed bank must attempt to project these costs in order to complete his business plan. The longer the duration and the more specific the monitoring and maintenance requirements, the greater the demand for credits as most permit applicants prefer to minimize temporal commitments to an individual project. Permitting Difficulty There is no demand for credits unless a permit is issued by the appropriate agencies to allow a wetland area to be impacted. There is also no demand for credits if impacts are allowed to wetlands without compensation or the only compensation required is the preservation of the remaining wetland on a project site. Neither situation allows a wetland bank to be economically viable. Therefore, a careful assessment of historical permitting actions and trends is necessary to predict the regulatory climate characteristic to a specific service area. Attitudes about Mitigation Alternatives A mitigation bank cannot sell its credits unless the regulatory agencies agree that the credits are appropriate compensation and that on-site alternatives are not ecologically preferable or practicable. Therefore, the demand for credits will depend on whether on-site mitigation is practicable or whether or not alternatives such as in lieu fee programs are deemed preferable. Attitudes appear to differ with geographic and political region.

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Regulatory Stability Wetland regulations in the United States change regularly. If users perceive that regulations will be relaxed, demand will be reduced as some users wait for less restrictive regulations. Conversely, a perception of impending greater restrictions will lead to an acceleration of demand as users try to gain entitlement approvals before permitting requirements increase. User Requirements At this time in the evolution of the wetlands banking industry, users simply desire to purchase the quantity of credits required by the regulators for a particular impact. As long as the user, after paying the sponsor, can clearly transfer all wetland performance risk to the bank sponsor, the user has no interest in the quality of the product. In most credit sales, this is the situation. This is not the case in single user banks and in situations where contractually the users share in the risk. User demand for quality could occur where more than one bank exists in the same service area if the quality of wetlands affected the permitting decision. However, as of 1999, there are no reports of quality of wetlands mitigation becoming a permitting factor. Wetland impacts are the result of development activities. As such, the fundamental determinants of the level of mitigation bank user demand are the rate of economic growth in a given service area, local zoning regulations, and the terms and conditions of wetland permits that authorize wetland impacts. A related determinant is the areal threshold of wetland impact that requires wetland mitigation. The lower the threshold, the greater the mitigation acreage created by users of wetland resources. Competitive Supply of the Product and Product Alternatives The economic feasibility of a wetland bank is dependent upon the availability of compensatory options. The quantity, cost, and regulatory favorability of these options must be assessed and compared to the product offered by the proposed bank to estimate what share of the estimated demand the proposed bank can capture. At some time in the future, regulations may exist that will also require the assessment of the quality offered by these alternative options. However, at this time, it is not an issue in supply or demand analysis for credits. Options for compensatory mitigation of wetland impacts should be evaluated to determine the economic feasibility of a wetland bank. These options include on-site and off-site opportunities for wetland mitigation, the availability of suitable wetland restoration and creation landforms, existing and proposed wetland banks, and in lieu fee alternatives. On-Site Opportunity for Wetland Mitigation The nature of the activities in a service area can create significant opportunities for on-site mitigation. For example, if demand is caused by large planned community ©2001 CRC Press LLC

developments in a geographic area that includes broad flat floodplains with substantial areas of prior-converted croplands (PC lands), then it is likely that these projects will have ample opportunity for on-site mitigation. As the land values of these parcels are based upon development yield projections (i.e., how many houses can be built), and houses are not usually allowed in floodplains today, there is no allocated land cost to these potential mitigation areas. Therefore, PC lands can be converted to wetlands with minimal capital costs. In this example, the supply of on-site mitigation opportunities could render a wetland bank economically nonviable. Alternatively, this could be a fine banking opportunity if a developer used these lands to create a bank sufficiently large to handle his needs and sell credits to others or a banker created a joint venture with the developer using his credit needs as an “anchor tenant.” Either way, the availability of on-site opportunities for wetland mitigation greatly influences the economic viability of a prospective bank. Off-Site Opportunities for Wetland Mitigation In many areas of the United States in the mid- to late-1990s, developers, wetland consultants, and entrepreneurs have learned that one way to avoid the rigorous review process, performance requirements, monitoring duration, and sometimes greater mitigation ratios of wetland banks was to simply develop off-site mitigation. The economies of scale associated with bank projects are achieved in this case by pooling the mitigation needs of multiple projects in one location, or simply building mitigation areas in phases with savings primarily occurring in the monitoring and maintenance phases. Areas within the proposed service area that are suitable for easy conversion to wetlands (such as PC lands), and have a low value for other allowable uses, offer a product that could be used instead of mitigation bank credits. If this is also an area where regulators typically impose minimal performance, monitoring, and maintenance requirements, a bank is unlikely to be able to compete against this supply alternative. In the current regulatory climate, off-site mitigation can save considerable costs. A review reveals that many off-site mitigation sites had a cost advantage over banks because they only required 5 years of monitoring (vs. 10 years in most banks), had less restrictive performance requirements, and had no financial assurance requirements or maintenance funding requirements. If this practice is encouraged in the service area of a proposed bank, the product supply opportunity could render the bank economically nonviable. Other Wetland Banks Obviously the quantity and characteristics of credits available from existing wetland banks in the proposed service area, as well as proposed banks, should be determined by contacting the appropriate regulatory agencies. It should be emphasized that wetland banks have developed on an ad hoc basis during the late 1970s to late 1990s. The result is that the agreements that establish these banks are variable, and existing banks could be operating under agreements that give an economic or ©2001 CRC Press LLC

marketing edge over new banks. Therefore, characteristics such as mitigation ratios and performance requirements that could favor other wetland banks over the proposed bank in the same service area should be analyzed carefully. In Lieu Fee Alternatives In lieu fee alternatives emerged in the late 1990s to become one of the most significant supply threats to the economic viability of wetland mitigation banks. Current regulatory structures have permitted compensation for wetland impacts to be in the form of monitoring fees that do not cover the full costs of compensation. Several U.S. examples illustrate this point. The Bracut Marsh public commercial bank developed by the California Coastal Conservancy is forecasted to recover only 54 percent of total costs at sellout (Environmental Law Institute, 1993). The Fairfax Land Trust accepted a US $315,000 payment to purchase and preserve a ±28 ha tract of wetlands to mitigate for 3 ha of forested wetland impact at the Stafford County, VA, Airport. The Trust had no funds remaining after the transaction for taxes and long-term maintenance and monitoring, and relies upon public donations to fund its operation (Hal Wiggins, personal communication). In the Chicago area, in lieu fee alternatives virtually stopped mitigation credit sales in late 1998 by charging significantly less than the credit costs of private banks (John Ryan, personal communication). Another advantage of in lieu fee programs is that many do not face the compensation timing constraints and service area restrictions endorsed by banks. Regulatory pressure usually minimizes the presale of mitigation credits which exposes private banks to the difficulty of raising capital and the risk of exposing capital. Regulatory pressure, environmental groups, some state laws, and federal guidance restrict the service areas of banks. Presale restrictions increase the risk and cost of credit production and increase capital needs, while service area requirements reduce the demand for credits for specific banks. Therefore, in lieu fee programs that do not face these restrictions can gain an economic edge. For example, one Nature Conservancy Trust Fund agreement with the Corps allows the Fund to collect fees from anywhere in Virginia, and expend the funds on preservation, restoration, and creation projects anywhere in the state. By contrast, a state law requires mitigation bank service areas to occur in the hydrologic unit code (HUC) of the bank or on an adjacent HUC within the same rivershed of the bank. The Trust also had the advantage of establishing fees on an impact-by-impact basis and operating without any specific, published monitoring and maintenance requirements. The Norfolk District of the Corps has recognized the potential economic edge for in lieu fees and has established fees at the same (or greater) prices charged by mitigation banks. By contrast, the Chicago Corps District had nine permitted and successful wetland banks by the end of 1998. The recent expansion of the in lieu fee program, in conjunction with a nonprofit organization known as Corlands, has almost eliminated bank credit sales. The Corlands program has several advantages over mitigation banks, including no assumption of risk, no specific performance requirements, lesser service area restrictions, and no timing restrictions. Also, Corlands pricing does not cover all direct project costs, and some monies are directed toward studies rather than restoration and creation of resources. The Savannah Corps District ©2001 CRC Press LLC

has handled these competitive problems by allowing the use of in lieu fees only in those service areas that do not contain an operating wetland bank (Ryan, 1998). At this time, U.S. federal regulators appear to have no consistent policy relating to in lieu fee programs, although Corps, USEPA, and USFWS officials have discussed this issue in public forums. As of May 1999, an interagency policy memorandum on this topic was still being negotiated. Risk Assessment and Presale of the Product Ecological concerns about the success of mitigation bank developments cause some regulators, environmental groups, and policy makers to substantially limit or prohibit presales of credits and, in some cases, call for no sales until after 5 or 10 years of successful monitoring. In lieu fee programs and traditional on-site or offsite mitigation programs typically do not face this requirement. An intriguing issue is the role of the U.S. federal government in determining presale requirements. The development of a wetland bank is a real estate development project, with a relatively unique product. While the agencies involved in mitigation banking limit the amount of presales for mitigation bank credits, other federal agencies involved in the financial banking industry have placed extreme pressure on financial institutions that provide federally guaranteed deposits to minimize lending exposure to speculative real estate ventures. They do this by requiring substantive (i.e., 50 percent or more) presales or preleasing. These agencies have learned that development projects without substantial precommitments from users have a high likelihood of economic failure because expected future demands do not always materialize at the projected time or price. At this time, the regulators of mitigation banks rarely make the connection between economic and ecological success. Most mitigation bank regulators focus on minimizing presales, yet the mitigation option that is stated as being the most ecologically preferred, on-site mitigation, is a 100 percent presale. Every dollar of funds raised by the presale of credits reduces the capital required to develop a mitigation bank. Rarely have traditional commercial lenders financed private banks as bank development activity is considered extremely risky. Venture capitalists and speculation investors, expecting rates of return in the 30 to 45 percent range, have been the primary capital sources to date. The availability of capital from presales can dramatically reduce the amount of capital required to be raised and, thus, reduce the cost of the project by reducing the investment return required. Pre-sales decrease the market risk and reduce exposure to regulatory climate risk. As the projected length of time that a project sell-out period increases, so too does the risk premium necessary to compensate investors for this exposure. For example, the possibility exists that changes to the definition of wetlands areas may cause some isolated wetlands to be nonjurisdictional (Lazarus, 1998). Other recent court rulings may cause certain activities in wetlands to no longer require mitigation (Lee, 1997; McElfish, 1997). These actions could clearly reduce mitigation bank credit demand. From the previous discussion, it is evident that delaying the timing of credit sales increases the risk and the cost of mitigation. This increases the break-even ©2001 CRC Press LLC

sales price and must be factored into any economic analysis of a prospective bank. Figure 3 illustrates the effect on credit prices for a hypothetical conversion of prior converted cropland from credit sale delays on the break even sales price (Shabman et al., 1998). This dramatic example may actually underestimate the effect of postponing sales in today's economic climate. An investment rate of return of 35 to 45 percent is typically required by potential bank investors to account for the cost of money, risk level, and alternative investment opportunities. Alternative Approaches to Ecological Success $14,000 $12,610 $12,000

Price per Credit

$10,000

$8,000 $6,562 $6,000

$3,701

$4,000

$3,966

$3,041

$2,000

$0 No Assurances

Figure 3

Bonding Case 1

Bonding Case 2

Bonding Case 3

Wait to Sell

Wetland credit prices under alternative scenarios for ecological success (given 20 percent target rate of return) for prior converted cropland conversion. No Assurances—30 percent of credits sold during the construction year, balance sold over the next 12 years, no bond. Boding Case 1—Same sales rate as no assurances, but performance guaranteed by a surety bond. Bonding Case 2—30 percent of credits sold in the first year after construction, balance sold over the next 11 years, performance bonded. Bonding Case 3—30 percent of credits sold in the fifth year after construction, balance sold in the next 7 years, all bonded. Wait to Sell Case—no presales, 30 percent of credits sold in the 10th year after construction, balance sold in the next 2 years. (Adapted from Shabman et al., 1998. With permission.)

Revenue Projections Assessing the economic viability of a proposed bank requires a projection of revenues. The prospective banker must project the rate at which credits can be expected to be sold over time and the price at which credits can be sold. The competitive supply of the product and product alternatives, and expected presale restrictions, will affect the projection. The projection typically exhibits an inverse relationship between sales rates (or assumption of the product) and sale price. Local knowledge of mitigation costs, land ©2001 CRC Press LLC

costs, and development pressures is necessary to make such projections. Revenue expectations vary widely throughout the United States (Table 4). By comparison, wetland mitigation costs for highway projects in the mid-Atlantic States have often exceeded US $272,000 per ha (Dennison and Schmid, 1997). As for any product, classical supply/demand relationships will determine the appropriate price of mitigation credits. Figure 4 illustrates the supply and demand relationship relative to credit price (Shabman et al., 1994). Table 4

Reported Mitigation Credit Sale Prices (US$ per ha) throughout the United States

State

Credit Sale Price

Date

Source

3,500–12,400 29,700–39,500

1998 1998

Georgia Florida

60,000–80,000 118,600–148,300

1999 1998

Central Virginia

148,300–197,700

1998

Northern Virginia

197,700–308,900

1998

New Jersey

370,700–494,200

1998

Washington

617,800

1999

Michael Henry, Hydrik Consulting Tom Sutliff, Ohio Wetlands Foundation Art Berger, Wet Inc. Ann Redmond, Florida Wetlands Bank Mike Kelly, Williamsburg Environmental Michael Rolband, Wetland Studies and Solutions, Inc. Bob Sokolove, U.S. Wetland Services Steve Johnson, Paine Field

Louisiana Ohio

COSTS OF MITIGATION BANK DEVELOPMENT Costs must be estimated for the specific project to determine whether it is feasible after assessing the demand for mitigation credits and the likely sales price for these credits. Land development costs, including mitigation bank development costs, usually fall into three categories: land costs, hard costs, and soft costs. Land Costs Although land costs can be a relatively small portion of a development project’s total cost, the costs are always extremely site specific and may vary considerably. Mitigation banking costs are no different. Therefore, the particular land selected for acquisition has an enormous influence over the costs of the project. For example, prior converted cropland has relatively low conversion costs. Sometimes only plugging drains is necessary to restore wetland’s hydrology. Costs can be enormous if the site requires extensive cut and removal of soil and rock, followed by topsoil replacement. Alternatively, the sale of these materials could pay for the entire mitigation project if the material removed is quality sand and aggregate. Often an entire tract must be purchased, but only a portion can be utilized as a wetland bank, causing an increase in the net land cost per wetlands area of creation. As wetland mitigation becomes more common, landowners and realtors are realizing that land easily convertible to wetlands may have a higher value under that ©2001 CRC Press LLC

CASE A

CASE B - High-cost restoration - High-value development pressure

Price

Price

- Low-cost restoration - Low-value development pressure

supply supply

Pa

demand

Pb demand Qa

Quantity

Quantity

Qb

CASE C

CASE D - High-cost restoration - Low-value development pressure

Price

Price

- Low-cost restoration - High-value development pressure

supply

supply

Pd

Pc demand Qc

Figure 4

Quantity

demand Qd

Quantity

Regional economic effects on the potential for mitigation credit markets. (Adapted from Shabman et al.,1994. With permission.)

use than under more traditional uses (e.g., cropland). This is due to the intrinsic low cost of converting such land when there is a limited supply in a service area. Hard Costs Hard costs include earthwork, erosion and sediment controls, planting, amenity and habitat enhancements, monitoring, and maintenance. Earthwork costs typically are the largest cost component, often running from 50 to 80 percent of a project’s cost. The primary cost variables are the volume of material to be moved, the distance the material needs to be moved, the number of moves of material, surface area of final grade lands, and geometry of the site. For example, sites with a high aspect ratio (long and skinny) cost more than low aspect ratio (square) sites to grade. Topography, hydrologic characteristics, and soils establish these variables,

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although design concepts, particularly topsoil treatments, can also have a significant effect. Local regulatory practices typically establish erosion and sediment control requirements. Because many wetland mitigation projects are built adjacent to natural wetlands, erosion and sediment controls can be critical in avoiding unnecessary impacts. It is not uncommon to see redundant (i.e., dual) erosion and sediment control systems required next to sensitive areas such as wetlands and streams. Therefore, these costs are highly variable depending upon site conditions and local regulations. Costs range from less than US $247 per ha for regularly shaped sites surrounded by development, to more than US $24,700 per ha for irregularly shaped creation sites surrounded by wetlands with redundant control requirements. Some local regulations also require cash escrows and bonds to assure compliance with erosion and sediment control regulations which must be accounted for in cash flow projections. The cost of providing wetland plantings for a mitigation project is highly dependent upon regulator opinions. Currently, there are three general schools of thought regarding appropriate plantings for wetland mitigation. The first school recommends providing good soils, wetland hydrology, and an erosion cover crop, and then allowing wetland plants to volunteer. The second school recommends planting a wetland seed mix and seedlings of the target species to assure that a wetland with the desired species composition is obtained. Often nonpioneer species (e.g., Quercus spp.) that are desired in a mature system will be planted as seedlings, while pioneer species (e.g., Acer rubrum, Nyssa sp.) will not be planted because they are expected to volunteer naturally. Finally, the third school of thought recommends planting a wetland seed mix, seedlings, and a selection of larger, mast producing specimens to minimize the temporal loss of habitat, particularly for forested wetlands. For example, sites planted with 2.5- to 5-cm-diameter trees provide a 7- to 15-year headstart over seedling-planted sites. Estimated costs associated with these three schools of thought are illustrated in Table 5. As can be seen from the table, there are significant cost differences between the planting schemes. Amenities such as birdhouses, deadfalls, observation blinds, nature trails, boardwalks, and interpretive stations can be accomplished at all price ranges. Rarely do regulators require such features, but several bankers have provided significant packages of amenity and habitat enhancements to maximize wildlife use and human educational interaction. Costs have been reported from US $250 to US $12,350 per ha. Amenity programs can become very expensive very quickly. For example, boardwalk costs range from US $86 to US $325 per m2. Monitoring and maintenance costs are directly related to the duration of the requirement, the type of wetland system designed, the performance requirements specified in the banking instrument, and natural events. For example, maintenance can be very expensive if a flood washes away newly planted trees. Monitoring and maintenance costs are often difficult to estimate because the maintenance aspect is dependent upon the timing of natural events, and regulatory requirements of the mitigation banking instrument. The latter is usually negotiated after initial economic investment.

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Table 5

Estimated Costs for Wetland Plantings Associated with Three Schools of Thought*

School

E&S Seeding1

Wetlands Seed Mix2

Seedlings @ 2.44 m (8 ft) O.C.3

Container Shrubs @ 3.65 m (12 ft) O.C.4

1 2 3

$3,600.00 $3,600.00 $3,600.00

N/A $7,200.00 $7,200.00

N/A $5,050.00 $5,050.00

N/A N/A $11,250.00

4 cm (1.5 in.) dbh trees @ 12.19 m (40 ft) O.C.5

Total Plant Cost/Ha

N/A N/A $9,050.00

$3,600.00 $15,850.00 $36,150.00

* The table depicts an order of magnitude difference between vegetation options for a forested wetland. 1

$0.36/m2—a typical cost for sites in the 1-ha size range (seed and mulch).

2

$0.72/m2—a typical cost for a diverse wetlands seed mix on a 1-ha site (seed and mulch).

3

Furnished and installed bare root seedlings with fertilizer and mulch (with warranty); 1,680/ha [2.44 m off center (O.C.)] at $3.00 each, with 1 yr 65 percent survival warranty (1,500 to 2,000 seedlings).

4

1 gal container grown plants (furnished and installed); 750/ha (3.65 m O.C.) at $15.00 each (with 1 yr warranty).

5

4 cm dbh trees furnished and installed; 67/ha (12.19 m O.C.) at $135.00 each (with 1 yr warranty).

Soft Costs Soft costs include regulatory approval costs, design costs, review fees, financial assurances, marketing, accounting, administrative costs, and taxes. Regulatory approval costs are dependent upon the trading rules established by regulators to increase the probability of mitigation success (Shabman et al., 1994). Site monitoring, site maintenance, the costs (liability) of project failure, mitigation site design standards, and other factors may be required. Meeting the requirements of the trading rules can increase costs and should be incorporated into the capital and operational costs budgeted for the development of a bank. Entrepreneurial bankers across the country have reported extremely significant direct costs (in the range of several hundreds of thousands to millions of dollars) attributable to gaining regulatory approval of the mitigation banking instrument, and related costs of controlling the selected site(s) during the approval process. Costs include land contract costs (i.e., attorney fees, surveys, title searches, options, and deposit costs), legal fees for developing the mitigation banking instrument, and design development costs such as groundwater monitoring, biological surveys, wetland delineation, and preliminary mitigation plan design. Figure 5 illustrates how regulatory approval costs can have a very large effect upon wetland credit prices. The design of the constructed wetland will have to meet two important constraints. These include the constraint imposed by the site itself, and the constraints imposed by regulators and the local plan approval processes. Before the site is committed, a preliminary mitigation design should be completed to ensure the site has the potential to support wetlands and to ensure that the constructed wetland will be acceptable as mitigation. Once the above constraints have been addressed, ©2001 CRC Press LLC

Figure 5

Wetland credit prices vs. approval costs and demand uncertainty. (Adapted from Shabman et al., 1998. With permission.)

final site design will include site hydrology analysis, water budget analysis, grading plans, erosion and sediment controls, soil requirements, and vegetation selection and distribution. Design costs can vary with the site. A relatively simple site, such as a flat floodplain area where the hydrologic linkage with a contiguous stream can be relatively easily restored, may have low design costs. Conversely, the design costs of a complex site that includes a number of wetland cells with differing hydrologic and soil requirements may be high. Some mitigation projects require permits from federal and state agencies, though usually these fees are relatively small. However, many local government fees, required to process grading and erosion and sediment control plans, can be quite expensive. Costs may be hundreds to thousands of dollars per ha in some localities. Financial assurances are methods used to provide some guarantee that a mitigation bank will succeed, and that there are funds available for planned or unplanned contingencies. Contingencies include site monitoring, maintenance, erosion repair, and vegetation replanting due to storm effects. Financial assurances will result in economic costs to the mitigation banker if the mitigation bank does not succeed. ©2001 CRC Press LLC

Surety bonds can be used to ensure that a mitigation bank meets specified criteria. The bonds are only released after the specified criteria are met. This may include completion and approval of the entire project, or it may address certain milestone criteria at which, upon completion, a part of the bond is released. For example, milestone criteria could include a certain vegetation density or percent cover over a specified part of the mitigation site. Escrow accounts are another way to ensure that funds are available to conduct maintenance and other activities (e.g., monitoring) that are necessary for a successful mitigation bank. Funds are paid into the escrow account when the mitigation bank receives payment for credits. The amount of deposits into the escrow account can vary depending upon the factors being considered by regulators. If regulators accept the presale of wetland credits, then the payment into the escrow account could reflect the risk of failure of the mitigation site, as well as the requirement to have funds available for monitoring, maintenance, and catastrophic events. If regulators approve the wetland mitigation (e.g., based upon site design standards), then the escrow payment may be lower (or a portion of funds collected can be released) and reflect only the future costs of maintenance and other activities. Marketing, accounting, and administrative costs can become a very significant cost component of wetland bank development over time. Prospective sponsors must budget for the staff time, office overhead, and direct expenses related to their activities. If the product is successfully marketed to users, costs carefully accounted for, and the entire product managed well, a wetland bank can be ecologically successful but still be a financial failure. Real estate taxes are typically assessed upon private landowners. Thus, until the site is transferred to a long-term steward that is tax exempt, this is a cost that must be budgeted for. Some localities will provide significant tax reductions on land held in open space for conservation, some may tax the land at the projected value of the mitigation credits, and others at its market value based on traditional highest and best use valuations. These practices vary by state and locality and can range in costs by an order of magnitude. Therefore, this cost must be evaluated on a site by site basis.

LONG-TERM STEWARDSHIP The purposes of long-term stewardship are to provide long-term maintenance and to assure that no inappropriate land use occurs. Although the ideal wetland would be self-sustaining and require no maintenance, such wetlands are very rare. This is particularly true in the short term. In reality, a lack of maintenance can result in the modification or loss of wetland functions. One common result of a lack of maintenance is the invasion of unwanted vegetation. Brazilian pepper (Schinus terebinthifolius), common reed (Phragmites australis), purple loosestrife (Lythrum salicaria), and, in some regions, cattails (Typha sp.) are undesirable in the United States. Other results can occur also, including topographical changes due to sedimentation or scouring, and water level changes due to problems with input or output systems. Each of these occurrences can result in ecological changes that may need to be addressed in order to achieve the goals of the wetland mitigation effort. For ©2001 CRC Press LLC

example, scouring can result in the formation of new channels in the wetland, causing water to exit the wetland more quickly, and decreasing the residence time available for water column/sediment/vegetation interactions. The best way to assure long-term protection of any mitigation area from intentional disturbance by man is to record a deed restriction, easement, or conservation covenant among the land records of the local courthouse that has jurisdiction to the lands in the mitigation area. This document should identify the legal description of the property to be protected and the activities allowed on the property. Once this protection is in place, the land can be safely transferred to the long-term steward. In order to guarantee that appropriate long-term stewardship occurs, some entity must be responsible for the site, and there must be sufficient funds available to accomplish the necessary tasks. Entities available for long-term stewardship include federal, state, and local resource agencies, nonprofit organizations, and private entities engaged in land conservation. Government entities require the funds and the manpower to conduct long-term stewardship of a site. Funding of a government agency, or even the existence of the government agency, is dependent upon a legislative body, and the priorities of such bodies may change in the future. Some of this uncertainty can be alleviated by the presence of long-term funding from a financial assurance program (e.g., trust fund established by the banker), although manpower problems may still occur owing to budgetary constraints. Government entities may also have trouble protecting the wetland site from other government entities or programs. Revenue generating activities such as forestry and agriculture are allowed in wetlands and may appear attractive to government entities suffering budgetary constraints. In addition, large wetland areas may be the most economically attractive routes for new roads, particularly if no other alternatives are available. Political pressure on the government entity could result in impacts to the constructed wetland site. Nevertheless, government entities can, and do, protect and maintain an extremely diverse array of natural areas very well. Nonprofit organizations are less vulnerable to political pressure, but they are vulnerable to financial problems. These organizations have been known to sell, for commercial purposes, properties that they consider less important relative to other properties. Some nonprofits, such as the Nature Conservancy, relinquish some of their acquisitions to government entities. However, most nonprofit organizations operate on low budgets and are able to utilize low paid or volunteer personnel. Thus, nonprofit organizations may be able to efficiently utilize trust fund or other available money for long-term maintenance and protection. Certain nonprofit organizations may reap additional benefits from a mitigation site that tends to provide additional incentives to protect the site. For example, educational institutions could utilize such a site for long-term research or study with the assurance that the site will remain available for a very long time. The mitigation banker does not usually wish to retain ownership of a site after all of the mitigation credits have been sold. There is no more profit to be made either from the site or from the long-term stewardship of the site. Thus the mitigation banker, or other private entity that owns the land, will probably donate the site to a governmental entity or a nonprofit organization. However, occasionally there are ©2001 CRC Press LLC

private entities that simply want to conserve the land. The deed restriction discussed earlier ensures long-term protection regardless of land ownership changes. Economic Projections The economic aspects of developing a wetland mitigation bank have been discussed in this chapter. To gain the financial resources needed to actually build and operate a bank from investors, venture capitalists, or financial institutions, these elements must be quantified and analyzed. Supply, demand, and regulatory policies should be assessed by the prospective banker to assess the price that wetland mitigation credits can be sold at in a specific area. Figure 6 illustrates the relationships of the factors discussed throughout this chapter. The fundamental elements needed to economically justify creation of a bank include capital and operating cost budgets, cash flow requirement projections (i.e., how the costs budgeted are expected to be expended over time), and sales rate and sale price projections (which are typically inversely related). These elements should be combined into one cash flow spreadsheet to model the economics of the proposed bank. Development of the model then allows sensitivity analyses to determine the effects of the more variable elements of the project. These elements include sales rates and credit prices, presale requirements, and phasing of capital expenditures. The model can then be used to estimate potential returns on the capital needed to develop the proposed wetland bank. The returns will be adjusted based upon the perceived level of risk by potential funding sources and compared to alternative investment options by such sources. A mitigation banker must find capital sources that recognize the proposed bank to be a superior investment alternative based upon its risk tolerance and investment interests. The appropriate capital budgeting techniques used in this analysis are identical to those involved in any capital intensive industry, and thus are not described in detail herein. For those inexperienced in such techniques, there are a number of excellent textbooks that address this topic (e.g., Bierman and Smidt, 1993). The fundamental economic test that a wetland bank must meet is the ability to sell credits at a price that exceeds expected costs and investor return requirements. This seemingly simple concept is very difficult to predict at this stage in the development of the mitigation banking industry. The market at this time is thin, and it is dependent upon regulatory practices and policies that often appear to change faster than wetland banks can be approved, constructed, and grown. A successful wetland bank is one that satisfies both economic and ecological criteria. Several banks have achieved that goal to date. Whether or not these dual goals can be achieved consistently throughout the country by this nascent industry, and create wetlands ecologically superior to traditional on- and off-site approaches, remains to be seen. The USEPA has proposed to engage the U.S. National Academy of Science to study this question over the 1999–2000 time period.

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Figure 6

Regulatory policies influence wetland mitigation credit markets. (Adapted from Shabman et al., 1994. With permission.)

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REFERENCES Apogee Research, An examination of wetlands programs: opportunities for compensatory mitigation, National Wetland Mitigation Banking Study, IWR Report 94-WMB-5, U.S. Army Corps of Engineers, Institute for Water Resources, 1994. Bierman, Jr., H. and Smidt, S., The Capital Budgeting System: Economic Analysis of Investment Projects, 8th ed., Prentice-Hall, Upper Saddle River, NJ, 1993. Brown, S. and Veneman, P., Compensatory Wetland Mitigation in Massachusetts, Research Bulletin Number 746, Massachusetts Agricultural Experiment Station, University of Massachusetts, Amherst, 1998. Brumbaugh, R. and Reppert, R., First Phase Report National Wetland Mitigation Banking Study, IWR Report 94-WMB-4, U.S. Army Corps of Engineers, Institute for Water Resources, 1994. Dennison, M. S. and Schmid, J. A., Wetlands Mitigation: Mitigation Banking and Other Strategies for Development and Compliance, Government Institutes, Rockville, MD, 1997. DeWeese, J., An Evaluation of Selected Wetland Creation Projects Authorized through the Corps of Engineers Section 404 Program, U.S. Fish and Wildlife Service, Sacramento, CA, 1994. Eggers, S. D., Compensatory Wetland Mitigation: Some Problems and Suggestions for Corrective Measures, U.S. Army Corps of Engineers, St. Paul District, 1992. Environmental Law Institute and Institute for Water Resources, Wetland mitigation banking report: resource document, National Wetland Mitigation Banking Study, IWR Report 94-WMB-2, U.S. Army Corps of Engineers, Institute for Water Resources, 1994. Environmental Law Institute, Wetland Mitigation Banking, Environmental Law Institute, Washington, D.C., 1993. Environmental Law Institute, Wetland Mitigation Banking, Environmental Law Institute, IWR Report 94-WMB-6, U.S. Army Corps of Engineers, Institute for Water Resources, Washington, D.C., 1994. Environmental Law Reporter and Environmental Law Institute, Wetlands Deskbook, Washington, D.C., 1993. Gallihugh, J. L., Wetland Mitigation and 404 Permit Compliance Study, Vol. 1, Report and Appendices A, B, D, E, U.S. Fish and Wildlife Service, Barrington, IL, 1998. Institute for Water Resources, National Wetland Mitigation Banking Study, Model Banking Instrument, Water Resources Support Center IWR Technical Paper WMB-TP-1, U.S. Army Corps of Engineers, May 1996. King, C., Bohlen, C., and Adler, K. J., Watershed Management and Wetland Mitigation: A Framework for Determining Compensation Ratios, University of Maryland System Draft Report #UMCEES, CBL-93-098, 1993. Lazarus, R., U.S. v. Wilson imposes limits on the reach of Section 404, Natl. Wetlands Newsl., 20, 2, 1998. Lee, G., Schlanger, P., and Murray, C., A decision well reasoned, Natl. Wetlands Newsl., 19, 2, 1997. Liebesman, L. R., Maryland adopts landmark wetlands mitigation banking legislation, Md. Builder, July/August, 1993. McElfish, J., The Tulloch Rule is overturned, Natl. Wetlands Newsl., 19(2), 1997. Redmond, A., Report on the Effectiveness of Permitted Mitigation, Florida Department of Environmental Regulation, 1991. Reppert, R., Wetlands Mitigation Banking Concepts, National Wetland Mitigation Banking Study, IWR Report 92-WMB-1, 1992. ©2001 CRC Press LLC

Salvesen, D., Wetlands: Mitigation and Regulating Development Impacts, 2nd ed., Urban Land Institute, Washington, D.C., 1994. Scodari, P. and Brumbaugh, R., Commercial Wetland Mitigation Credit Ventures: 1995 National Survey, IWR Report 96-WMB-9, U.S. Army Corps of Engineers, Institute for Water Resources, Alexandria, VA, 1996. Scodari, P., Shabman, L., and White, D., Wetlands Credit Markets: Theory and Practice, IWR Report 95-WMB-7, U.S. Army Corps of Engineers, Institute for Water Resources, Fort Belvoir, VA, 1996. Shabman, L., Stephenson, K., and Scodari, P., Wetland credit sales as a strategy for achieving no net loss: the limitations of regulatory conditions, Wetlands, 18, 471, 1998. Shabman, L., Scodari, P., and King, D., National Wetland Mitigation Banking Study, Expanding Opportunities for Successful Mitigation: The Private Credit Market Alternative, Institute for Water Resources, Water Resources Support Center, IWR Report 94-WMB-3, U.S. Army Corps of Engineers, Alexandria, VA, 1994. Short, C., Mitigation Banking, Biological Report 88(41), U.S. Department of the Interior, Fish and Wildlife Service, Research and Development, Washington, D.C., 1988. Want, W. L., Law of Wetlands Regulation, The Clark Boardman Callaghan Environmental Law Series, West Group, New York, 1998. U.S. Army Corps of Engineers, U.S. Environmental Protection Agency, National Marine Fisheries Service, and Natural Resources Conservation Service, Federal guidance for the establishment, use and operation of mitigation banks, Fed. Regist., 60, 58605, 1995. U.S. Environmental Protection Agency, Memorandum of Agreement between the Department of the Army and the Environmental Protection Agency Concerning the Determination of Mitigation under the Clean Water Act Section 404(b)(1) Guidelines, 1990. Wilkey, P. L., Sundell, R. C., Bailey, K. A., and Hayes, D. C., Wetland Mitigation Banking for the Oil and Gas Industry: Assessment, Conclusions, and Recommendations, Argonne National Laboratory, Argonne, IL, 1994.

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Kent, Donald M. “Monitoring Wetlands” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

8

Monitoring Wetlands Donald M. Kent

CONTENTS Reasons for Monitoring Measures Properties of Individual Plants Properties of Vegetation Communities Landform Properties Properties of Soil Hydrologic and Hydraulic Properties Aquatic Physical and Chemical Properties Organismal Properties Properties of Individual Wildlife and Fish Species Properties of Wildlife and Fish Communities Approaches to Monitoring Selecting a Monitoring Approach Investment and Return Investment, Return, and Area Investment, Return, and Time Measures and Monitoring Approaches Investment, Measures, and Area Investment, Measures, and Time Monitoring Design and Analysis Design Analysis References

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Wetlands monitoring is the checking, watching, or tracking of wetlands for the purpose of collecting and interpreting data, which is then used to record or control the wetland or processes affecting the wetland. Not to be confused with wetlands assessment or evaluation which is the valuation of wetlands, monitoring of wetlands is a component of mitigation efforts (Kusler and Kentula, 1990; U.S. Army Corps of Engineers, 1989), the Environmental Protection Agency’s Environmental Monitoring and Assessment Program (Paul et al., 1990; Leibowitz et al., 1991), and other programs designed to protect, conserve, and understand wetland resources (New Hampshire Water Pollution Control Commission, 1989; Haddad, 1990; Walker, 1991). Monitoring efforts are conducted for several reasons using a variety of techniques to measure and assess an array of structural and functional parameters. The process of developing and implementing a monitoring program can be reduced to four basic steps (Figure 1). First and foremost, the reason for monitoring must be identified and clearly stated. Second, a determination of the measures appropriate for achieving the stated objective(s) must be made. Third, an approach commensurate with the level of investment and the required return must be selected. The size of the area to be monitored, as well as the length of time the area will be monitored, will affect selection of an approach. Finally, the information gathered from the monitoring effort must be analyzed and interpreted.

REASONS FOR MONITORING For the most part, wetland monitoring is conducted for a relatively few, discrete reasons. Habitat mapping and trend analysis monitoring are conducted to identify wetlands resources and to detect changes in these resources over time. Examples of mapping and trend analysis monitoring include efforts in coastal and seaway Canada (Rump, 1987), coastal India (Nayak et al., 1989), migratory bird habitat in central California (Peters, 1989), and the National Wetlands Inventory project (Dahl and Pywell, 1989). Perhaps the largest monitoring effort of this type is the Environmental Monitoring and Assessment Program (Paul et al., 1990; Liebowitz, 1991). The program, designed to monitor the condition of wetlands, has stimulated mapping and trend analysis monitoring throughout the United States (Haddad, 1990; Johnston and Handley, 1990; Orth et al., 1990). Initial aspects of the wetland ecosystems component of the Environmental Monitoring and Assessment Program focus on determining the sensitivity of various metrics for detecting known levels of stress and determining the spatial and temporal variability of proposed wetland indicators of condition (U.S. Environmental Protection Agency, 1990). Wildlife and fisheries management monitoring is also a type of habitat mapping and trend analysis monitoring. It is conducted to provide information about species richness and species abundance over time and to assess the effects of management strategies. The wildlife or fisheries population (Henny et al., 1972; Neilson and Green, 1981; Hink and Ohmart, 1984; Young, 1987; Molini, 1989), habitat indicators of wildlife richness and abundance (Weller and Fredrickson, 1974; Koeln et al., 1988), or both (Weller, 1979; Weller and Voigts, 1983) are monitored. ©2001 CRC Press LLC

Figure 1

Steps for developing and implementing a wetland monitoring program.

A second reason for monitoring is to determine the effectiveness of enhancement, restoration, and creation efforts. Examples include evaluation of habitat created using dredge spoil (Newling and Landin, 1985; Landin et al., 1989) and restoration of degraded habitats (Pacific Estuarine Research Laboratory, 1990). There are numerous monitoring efforts associated with Section 404, state, and local wetland fill permits (Kusler and Kentula, 1989; U.S. Army Corps of Engineers, 1989; Erwin, 1991) as well. Impact analysis constitutes a third reason for monitoring. Monitoring is conducted to determine the response of wetlands to identified direct and indirect impacts. Examples include monitoring of impacts to wetlands on and adjacent to hazardous waste sites (Watson et al., 1985; Hebert et al., 1990), as well as impacts from discrete and continuous chemical contamination events (McFarlane and Watson, 1977; Woodward et al., 1988). Other examples of impact analysis monitoring include studies of the effects of highway construction (Cramer and Hopkins, 1981), siting impacts from generating station construction and operation (Wynn and Kiefer, 1977), ©2001 CRC Press LLC

effects on wetland flora from exposure to electromagnetic fields (Guntenspergen et al., 1989), and impacts from agricultural practices (Hawkins and Stewart, 1990; Walker, 1991). Finally, wetlands may be monitored to determine the potential for, or effectiveness of, wetlands for treating point source or nonpoint source discharges. Treatment monitoring has been applied to studies of the effectiveness of constructed wetlands for domestic wastewater treatment (Hardy, 1988; Choate et al., 1990; Tennessee Valley Authority, 1990), mine drainage (Eger and Kapakko, 1988; Stark et al., 1988; Stillings et al., 1988), stormwater runoff (Meiorin, 1991), and agricultural runoff (Costello, 1991).

MEASURES A large number of measures have been applied, or potentially can be applied, to monitoring of wetland structure and function (Table 1). Commonly used measures include measures of the properties of individual plants and animals, measures of the properties of vegetation and wildlife communities, measures of aquatic physical and chemical properties, and measures of soil properties. Less commonly used are measures of hydrologic and hydraulic properties such as flood frequency and groundwater depth. Generally unused are potentially useful measures of landform properties such as heterogeneity and patch characteristics (Forman and Godron, 1986). The latter properties are particularly important in the preservation and creation of wetlands for wildlife and are likely to be useful for other aspects of habitat mapping and trend analysis monitoring. Measures of organismal properties are typical of impact analysis monitoring programs. Properties of Individual Plants Measures of the properties of individual plants are used to assess the condition of natural plants and propagules. In theory, the properties of a plant are affected by any factor that alters the growth and maintenance of the plant. Factors that affect plant growth and maintenance include soil nutrients, soil moisture, disease, pest infestations, and anthropogenic and other disturbances. Information obtained from measurements of the properties of individual plants can be applied to trend analysis monitoring, enhanced, restored, and created wetlands monitoring, impact analysis monitoring, and treatment monitoring. The simplest measure of an individual plant is survival, that is, whether the plant is dead or alive. For living plants, measures include basal area, which is the area of exposed stem if the plant were cut horizontally, and stem diameter, which is the maximum width of the area of exposed stem if the plant were cut horizontally. Basal area and stem diameter are usually measured in centimeters (2.5 cm equals 1 in.) above the ground by ecologists and range managers, and 1.4 m (4.5 ft) above the ground by foresters. Plant height is the mean vertical distance from the ground at the base of a plant to the uppermost level of a plant. Cover, including ground cover (herbaceous plants and low growing shrubs) and canopy cover (other shrubs and ©2001 CRC Press LLC

Table 1

Measures of Wetland Structure and Function

Properties of individual plants Basal area Biomass Canopy diameter Cover Properties of vegetation communities Basal cover Biomass Cover Cover type Density Landform properties Accessibility Dispersion Heterogeneity Isolation Properties of soil Classification Moisture Hydrologic and hydraulic properties Flood storage volume Frequency of flooding Groundwater depth Groundwater recharge volume Aquatic physical/chemical properties Biological oxygen demand Chlorophyll Turbidity Dissolved solids Nutrients Organismal properties Behavior Bioaccumulation Growth and development Properties of individual wildlife and fish species Abundance Association Age structure Properties of wildlife communities Abundance Biomass Density

Growth rate Productivity Stem diameter Survival Evenness Richness Survival Stratification

Interaction Shape Size

Organic content Texture Surface Surface Surface Surface

water water water water

depth area velocity width

pH Salinity Temperature Toxicants

Metabolism Reproduction Tissue health Density Mortality Presence/absence Evenness Niche overlap Richness

trees), is that part of the ground area covered by the vertical projection downward of the aerial part of the plant. Typically, the vertical projection downward is viewed as a polygon drawn around the plant’s perimeter and ignores small gaps between branches. Canopy diameter is the average maximum width of the polygon used for canopy cover. Basal area, stem diameter, plant height, cover, and canopy diameter, if repeatedly measured over time, can be used as indicators of plant growth rate. Plants allocate net production to leaves, twigs, stem, bark, roots, flowers, and seeds. The accumulated living organic matter is the biomass and is usually expressed ©2001 CRC Press LLC

as the dry weight per unit of area. Determining the allocation to each part is generally invasive in that the parts must be removed from the plant and either weighed or analyzed for energy or nutrient content. Nevertheless, individual plant productivity can be estimated by sampling leaves, flowers, or seeds (Figure 2).

Figure 2

Monitoring of individual plants during the appropriate season will indicate if reproduction is occurring. Productivity can be estimated by sampling leaves, flowers, or seeds.

Properties of Vegetation Communities Just as factors that affect plant growth and maintenance are reflected in measurements of the properties of individual plants, factors which affect more than one individual plant will be reflected in measurements of the properties of vegetation communities. Therefore, measures of the properties of vegetation communities are of use in assessing the condition of natural and mitigated vegetation communities. Measures of the properties of vegetation communities include extensions of the measures applied to individual plants as well as measures which are unique to the characterization of communities. Measures of community survival, basal

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cover, cover, and biomass require the accumulation of measures of individual plants. The cumulative expression of these measures, relative to the number of individuals assessed in the case of survival, or relative to the size of the area assessed in the case of basal cover and cover, provides for the description of the vegetation community. Properties unique to vegetation communities include cover type, which is the assignment of the community or parts of the community, to predetermined categories (Figure 3). “Classification of wetlands and deepwater habitats of the United States” (Cowardin et al., 1979) is the most commonly used system for describing cover type and its widespread use provides for comparison among disparate monitoring efforts. Nevertheless, the development of other descriptive systems is sometimes required in order to maximize information return. Other measures unique to the community level are density, which describes the number of individuals per unit of area, and richness, which is the list of plant species identified in the community of interest. If each individual plant within the sampling area is identified, then evenness can be determined. Evenness describes how the species abundances are distributed among the species. Another widely used measure of community structure, diversity, combines richness and evenness. Because diversity measures combine richness and evenness, they confound the number of species, the relative abundances of the species, and the homogeneity and size of the area sampled, and are, therefore, less useful than measures of richness and evenness. Finally, measures of stratification, a diversity index reflecting the amount of foliage at various levels above the ground, describe the vertical structure of the vegetation community.

Figure 3

Wetlands can be monitored for cover type, which is the assignment of the plant community, in this case emergent macrophytes, to predetermined categories.

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Landform Properties Measures of landform properties are used by landscape ecologists to identify and describe individual communities and the relationships among communities. The measures can be valuable to wetland scientists interested in local and regional planning issues, particularly because these issues relate to wildlife and trend analysis. However, the measures have been infrequently used and, therefore, require precise definition and identification of limitations, when applied. Some measures of landform properties, such as shape and size, can be applied to studies of single wetlands. Shape is typically described as a ratio of wetland perimeter to wetland area (Bowen and Burgess, 1981). Size is described as the area of the wetland or by some linear dimension such as length, width, or the ratio of length to width. Other measures of landform properties require consideration of more than a single wetland. Accessibility describes the distance along a corridor of suitable habitat from one wetland to another and reflects the perceived ease of species movement (Bowen and Burgess, 1981). Dispersion describes the pattern (e.g., clumped, uniform, random) of spatial arrangement among wetlands (Pielou, 1977). Isolation describes the distance of a wetland from other wetlands (Bowen and Burgess, 1981) and interaction describes the perceived influence of a wetland on another wetland through consideration of the distance between wetlands (MacClintock et al., 1977). Properties of Soil Measures of the properties of soils are useful in describing wetland structure and provide clues to wetland function. As part of mapping and trend analysis monitoring efforts, measures of soil properties help to distinguish between wetland and nonwetland areas and provide information as to changes to these areas. If monitored as part of a wetland enhancement, restoration, or creation effort, including efforts associated with the establishment of treatment wetlands, measures of the properties of soil indicate the development of hydric conditions. Soil is typically classified according to such characters as color, texture, and size and shape of aggregates. The Department of Agriculture Soil Conservation Service system is the commonly used taxonomic classification system in the United States. Based upon the kind and character of soil properties and the arrangement of horizons within the profile, the system also provides information about the use and management of the soil. Soil texture is based on the relative proportions of the various soil separates in a sample and is estimated from its plasticity when extruded and by feeling its grittiness (Hays et al., 1981). Soil moisture is the percent of a given amount of soil consisting of water and is estimated by the loss of weight on drying. Soil organic content is the percent of a given amount of soil consisting of organic matter and is estimated by loss of weight upon ignition.

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Hydrologic and Hydraulic Properties Simply stated, a wetland is a wetland because it is wet. Hydrologic and hydraulic measures provide useful descriptors of wetland structure and also provide valuable information as to wetland function. Measures of hydrologic and hydraulic properties provide information about the extent of wetlands as well as the effect of intrinsic and extrinsic changes to wetlands. Treatment monitoring benefits from measures of hydrologic and hydraulic properties in determining maximum treatment capacity. Measures of hydrologic and hydraulic properties as part of enhanced, restored, and created wetland efforts are integral to an assessment of project success. Velocity describes the speed at which water travels and reflects not only the depth and width of the water body, but also the topographical gradient and the extent and type of vegetation. Water depth, width, and area are descriptors of wetland structure. Monitored over time, and in relation to extreme events, these measures provide an empirical estimate of the frequency of flooding and of flood storage volume. Flood storage volume can also be estimated using one of a number of computerized hydrological models (U.S. Army Corps of Engineers, 1981; Soil Conservation Service Hydrology Units 1982 and 1986; Huber and Dickinson, 1988). Model inputs include wetland and watershed slope, vegetative cover, soil type, and surface type (i.e., pervious or impervious). Groundwater depth, the distance below the ground surface at which water occurs, can be determined empirically through the installation and monitoring of wells (Figure 4). Groundwater recharge volume, the volume of surface water moving down through the soil to an underlying groundwater system or aquifer, can be estimated using the aforementioned hydrological models. Aquatic Physical and Chemical Properties The quality of water affects the growth, maintenance, and reproduction of wetland flora and fauna. Wetland water quality is revealed by measures of aquatic physical and chemical properties. Water quality reflects the condition of the surrounding environment and is affected by human activities such as watershed erosion and point and nonpoint source discharges. Wetland water quality also reflects the condition of the wetland itself. Measures of aquatic physical and chemical properties are particularly applicable to monitoring of the effects of impacts to wetlands and monitoring the effectiveness of treatment wetlands. Water temperature influences the rate of metabolic reactions, the reactivity of enzymes, and the amount of oxygen that can be dissolved in water. The pH of water affects organismal physiological reactions and membrane characteristics. Dissolved oxygen concentrations must be sufficient to enable diffusion from the water into an animal’s blood. Salinity affects water quality through its effect on the ability of species to maintain osmotic balance. Turbidity restricts the depth to which solar radiation can penetrate the water column. Dissolved solids, such as carbonates, bicarbonates, chlorides, phosphates, nitrates, and salts of calcium, magnesium, sodium, and potassium, affect organismal ionic balance and other physiological processes. Biological oxygen demand, the amount of oxygen required by bacteria ©2001 CRC Press LLC

Figure 4

Groundwater depth, the distance below the ground surface at which water occurs, can be monitored empirically using wells.

while stabilizing decomposable organic matter under aerobic conditions, reflects the trophic status of the aquatic body and possibly the extent and type of inputs to the aquatic body. Trophic status is also revealed by measures of chlorophyll and of nutrients such as nitrates, nitrites, and phosphates. Measures of toxicants such as heavy metals, volatile organic compounds, and petroleum hydrocarbons provide a direct measure of contaminants. Organismal Properties As with plants, factors that alter the growth, reproduction, and maintenance of an individual organism will affect the properties of that organism. Therefore, those properties will be of use in assessing the condition of that organism. Factors that affect organismal growth, reproduction, and maintenance include water quality, including the presence or absence of environmental toxins, and the availability of food and cover. Measures of the properties of individual organisms are of particular use in monitoring impacted wetlands or in assessing the affects to a natural wetland used for treatment of a discharge. Organismal behavior, such as predator avoidance, foraging effectiveness, and intraspecific social interactions, is modified by factors that affect the wetland. So, too, the rate or age of the onset of reproduction and the rate of growth and development are similarly affected. Factors may also affect organismal metabolism, such as oxygen consumption, photosynthesis, nutrient uptake, or enzymatic reactivity. In a more direct sense, organisms express a response to unfavorable environmental factors by the bioaccumulation of chemical constituents. In some cases, bioaccumulated ©2001 CRC Press LLC

chemical constituents are evidenced by changes in tissue health, such as lesions and tumors. Properties of Individual Wildlife and Fish Species Factors that grossly affect organisms, particularly those factors that affect reproduction and growth and development, will be reflected in properties of individual wildlife and fish populations. Therefore, measures of the properties of individual wildlife and fish species are of use in monitoring impacted wetlands. These measures are also useful for trend analysis monitoring efforts in that they reflect the condition of the wetland relative to the focal species. Finally, measures of the properties of individual wildlife and fish species provide important information as to the value of enhanced, restored, and created wetlands as wildlife or fisheries habitat. The simplest measures of individual wildlife and fish species are presence/absence and abundance (Figure 5). Requiring relatively more effort are measures of population density, the number of individuals per unit of area. Other measures are useful in assessing the potential persistence of the species. Mortality can be expressed as either the probability of dying or as the death rate. The complement of mortality is survival, the probability of living. Natality is the production of new individuals in the population and can be described as the maximum or physiological natality or as the realized mortality. Changes in mortality, survival, and natality are reflected in the age structure of the wildlife or fish population. Declining or stabilized populations are characterized by relatively fewer young in the reproductive age classes and a relatively larger proportion of individuals in older age classes. Conversely, growing populations are characterized by a relatively larger proportion of the younger age classes. Monitoring efforts interested in determining how a wildlife or fish species is distributed throughout the wetland will use measures of association. Properties of Wildlife and Fish Communities Analogous to the situation with vegetation communities, factors which affect the reproduction, growth, and development of more than one wildlife or fish species will be reflected in measurements of the properties of wildlife and fish communities. As with measures of the properties of individual wildlife and fish species, measures of the properties of wildlife and fish communities are applicable to impact monitoring, trend analysis monitoring, and enhanced, restored, and created wetlands monitoring efforts. Measures of wildlife or fish community abundance and density provide gross estimates of wetland condition and suitability (Figure 6). The number of species occurring in the community is the richness. Evenness refers to how the species abundances are distributed among the species. The richness and evenness measures are frequently combined to form a single measure of diversity. Again, the major criticism of diversity measures is that they confound a number of variables that characterize community structure (Ludwig and Reynolds, 1988). Alternatively, biomass can be used to quantitatively describe wildlife and fish communities. More ©2001 CRC Press LLC

Figure 5

Wildlife presence/absence and abundance can reasonably be monitored with a minimum of effort.

commonly applied to measures of vegetation communities, biomass is of use in describing community structure, particularly energy flow. Monitoring efforts interested in determining how coexisting species use common wetland resources will use measures of niche overlap.

APPROACHES TO MONITORING Two broad approaches are available for monitoring wetlands: remote and contact. Remote monitoring is the acquisition of information about a wetland from a distance, without physical contact. Conversely, contact monitoring is the acquisition of information about a wetland from near at hand, with physical contact. Remote monitoring of wetlands provides a level of spatial and temporal sampling that is impractical with contact techniques. Because data are available at large and synoptic scales, large-scale patterns can be discerned and large-scale processes can be measured. Space-based remote sensing instruments measure electromagnetic radiation reflected and emitted by the earth’s surface. Visible and thermal satellite data provide location information about broad vegetation cover types and extent of inundation (Carter et al., 1976; Roughgarden et al., 1991; Wickland, 1991). Comparison of images along a temporal gradient provides information about land use and vegetation successional changes (Mackey and Jensen, 1989; Nayak et al., 1989; Byrne and Dabrowska-Zielinska, 1981). Information is provided at mesoscale, macroscale,

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Figure 6

Fish community species richness and abundance can be monitored using fish traps such as this fyke net.

and megascale levels (see Table 2 adapted from Delcourt and Delcourt, 1988). Aerial photography provides similar information at the microscale and mesoscale levels (Cartmill, 1973; Haddad, 1990; Jean and Bouchard, 1991). Contact monitoring cannot reasonably provide information about large-scale processes, but it does provide for the acquisition of more detailed structural and functional information on a site-specific basis. Contact monitoring includes measures of habitat structure, measures of water quality and hydrology, measures of animal populations, and measures of contaminant levels in wildlife and fish (Cramer and Hopkins, 1981; Bosserman and Hill, 1985; Watson et al., 1985; Eger and Kapakko, 1988; Hardy, 1988; Pritchett, 1988; Shortelle et al., 1989; Clausen and Johnson, 1990; Conner and Toliver, 1990; Hebert et al., 1990; Oberts and Osgood, 1991; Walker, 1991). Specific measurement parameters used in contact monitoring are discussed in Hays et al. (1981), Cooperrider et al. (1986), Graves and Dittberner (1986), and Adamus and Brandt (1990). Historically, many contact monitoring programs have been based upon structural parameters related to vegetation (Larson, 1987; Carothers et al., 1989; Landin et al., 1989). The recent trend is toward the measurement of appropriate indicators which reflect wetland functional condition (Brooks et al., 1989; Paul et al., 1990; Kent et al., 1992).

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Table 2

Spatial Hierarchy of Delcourt and Delcourt (1988)

Hierarchical Domains

Sublevels

Area (m2) 1.5 × 10

14

Map Scale Smaller scale

Global Megascale

1 × 1014

1:20,000,000

1 × 1012

1:2,000,000

1 × 1010

1:200,000

1 × 108

1:20,000

1 × 106

1:2,000

1 × 104

1 1:1 200

1 × 102

1 1:1 200

1 × 10

1 1:1 200

Continent Macroscale

Macroregion Mesoregion

Mesoscale Microregion Macrosite Microscale

Mesosite Microsite

SELECTING A MONITORING APPROACH Ideally, all monitoring programs would have adequate numbers of trained personnel using replicated, quantitative techniques, and large sample sizes. Monitoring would be conducted over a sufficient number of years to provide for identification of stochastic variation. Finally, the data would be subjected to parametric statistical analysis and the results would have global application. Realistically, the vast majority of monitoring programs, if not all monitoring programs, are constrained by the availability of resources. Government monitoring programs are often understaffed, and the programs are subject to periodic reassessment of priorities. Monitoring programs conducted by academics are sensitive to available funding. Mitigation monitoring is conducted for the minimum time necessary, using minimum funding, because of the understandable disinterest of developers in long-term commitment of resources to what may be perceived as an ancillary activity. Investment and Return Given the enormous amount of resources required to effectively and accurately monitor a single wetland, notwithstanding the resources required to monitor wetlands at the megascale, it is clear that each monitoring opportunity requires consideration of investment and return. That is, what information about the wetland is required and what resources can be applied in pursuit of this information? This consideration must bear in mind that a commitment of resources to one monitoring opportunity necessarily diminishes the resources available for subsequent monitoring opportunities.

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Furthermore, the relationship between investment and return can be expected to be logarithmic: as investment increases, the rate of increase in return decreases (Figure 7). To illustrate this point, consider the monitoring of a 32-ha created wetland for wildlife species richness. For a single sampling technique, a pedestrian transect, 1 h of sampling resulted in the identification of 8 species, 4 h of sampling identified 15 total species, 7 h of sampling identified 18 total species, and 10 h of sampling identified 22 total species. A decline in return relative to the level of investment is also evident when multiple sampling techniques are used. After 84 h of pedestrian transect sampling, fyke net sampling, scan sampling, and auditory sampling identified 32 species. An additional 84 h (for a total of 168 h) of sampling identified 12 more species, and a third 84-h sampling event (total of 252 h) yielded 7 more species.

Figure 7

The relationship between investment and return can be expected to be logarithmic. That is, as investment increases the rate of increase in return decreases.

Clearly, in this example an investment of 1 h was inadequate to reasonably describe species richness on the site. However, if the question is, “Are wildlife species present on the site?” then 1 h is adequate. Conversely, if it is important to identify nearly every species using the site, then an investment of 250 or more hours is more realistic. In most cases, information needs, whether for determination of wildlife species richness or any other wetland variable, fall somewhere in-between these two extremes. Continuing with the example, an investment of more than 250 h only results in the identification of 7 more species than an investment of 168 h. This is a 16 percent increase return for a 50 percent increase in investment. Again, in the real world of budget constraints and additional commitments, selecting a level of

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investment commensurate with the level of information required is essential. Nevertheless, the effect of a less than comprehensive monitoring effort on project output must be recognized prior to initiation of the monitoring effort. As discussed above, information returned from a given monitoring effort is a function of the amount of resources invested in that effort. The rate of increase in return decreases as investment increases. Selection of an appropriate level of investment, and consequently development of a pragmatic monitoring program, is expedited by considerations of space and time. Investment, Return, and Area For a given level of investment, return decreases with increasing area to be monitored. Conversely, as the area to be monitored increases, the level of investment must be correspondingly increased to maintain a given level of information return. The relationship between area and return for a given level of effort, as it was between investment and return, is logarithmic because sampling parameters are typically area specific (Figure 8). Returning to the example of monitoring wildlife species richness, a monitoring technique which comprehensively samples a 1-ha site will only sample one half of a 2-ha site, one fourth of a 4-ha site, and so on. As such, in order to monitor larger areas, either the investment must be increased accordingly or a lesser return accepted.

Figure 8

When sampling parameters are area specific, the relationship between area and return is inversely logarithmic. A relatively greater level of investment is required to maintain a given level of return as the size of the area to be monitored increases.

The scale at which an investment is made, and at which a return is realized, is a function of the reason for monitoring. Habitat mapping and trend analysis monitoring occurs on the microscale to the megascale, with monitoring for wildlife and fisheries management occurring primarily at the microscale and mesoscale. Enhanced, restored, and created wetlands monitoring occurs at the microscale.

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Impact analysis monitoring occurs at the microscale and mesoscale, and treatment monitoring occurs at the microscale. Investment, Return, and Time The effect of time on monitoring investment and return has historically been restricted to consideration of the length of the monitoring program. Habitat mapping and trend analysis monitoring occur for undefined but generally prolonged periods. Monitoring of fisheries and wildlife management occurs for extended periods corresponding with management goals, although, short-term efforts do occur (Shortelle et al., 1989). Monitoring of enhanced, restored, and created wetlands generally occurs for finite periods of time (typically 2 to 7 years) and frequently corresponds with permit conditions. Impact analysis monitoring continues until such time as the impact source is remedied or the nature and extent of the impact is understood. Treatment monitoring occurs for extended periods corresponding with treatment needs. Characteristic of most programs, regardless of the reason for monitoring, is establishment of a monitoring scheme that proceeds without variation until program completion. This type of effort, which can be described as continuous investment monitoring, is represented by a straight line when time is plotted against monitoring investment. In some instances, continuous investment monitoring can be reasonably modified over time to achieve an efficient balance between investment and return. For example, monitoring of an enhanced, restored, and created wetland may include tracking of structural aspects related to construction (e.g., propagule survival, surface or groundwater elevation), as well as determination of wetland function (e.g., species richness, sediment, and toxicant retention). As an alternative to the maintenance of this relatively large investment throughout the monitoring period, structural and functional monitoring could be confined to an initial critical period, perhaps as little as 2 to 3 years, and subsequent years committed to monitoring of functional components. The level of investment could be reduced to levels commensurate with habitat mapping and trend analysis after a suitable period (Figure 9a). Similarly, impact analysis monitoring could be reduced to levels commensurate with habitat mapping and trend analysis after an initial investment intensive period (Figure 9b). In each case, gaining an understanding of the relationship between site specific structure and function is essential. Measures and Monitoring Approaches A number of the measures listed in Table 1 cannot be accomplished through remote sensing and, therefore, only lend themselves to contact monitoring. Properties of soil, individual organisms, and aquatic physical and chemical properties cannot reasonably or effectively be measured without observations at the site. Measures of the properties of individual plants and vegetation communities also largely lend themselves to contact monitoring. However, measurement of vegetation community cover type and canopy cover can reasonably be accomplished through the use of ©2001 CRC Press LLC

Figure 9

Most monitoring programs use a continuous level of investment throughout. Reasonably, monitoring investment for enhanced, restored, and created wetlands (a) and impact analysis (b) could consist of an initial investment intensive period followed by a less intensive period(s) designed to balance investment and return.

aerial photography or satellite imagery, and individual tree canopy cover can be accomplished through the use of aerial photography when vegetation is sparse and relatively isolated. Similarly, properties of individual wildlife species and wildlife communities can in certain situations be remotely monitored. For example, presence and absence, abundance, and density can be accomplished through aerial photographs or aerial flyovers of large species living in open habitats. Measures of landform properties only lend themselves to remote monitoring, either through the use of aerial photography or satellite imagery. Measures of hydrologic and hydraulic properties are accomplished by contact monitoring, remote monitoring, or a combination of the two. Measures of groundwater and surface water depth can only be accomplished through contact monitoring. Measures of surface water area are best accomplished through remote monitoring, whereas monitoring of the frequency of flooding can reasonably be accomplished through either contact or remote monitoring. Flood storage volume and groundwater recharge volume which require knowledge of wetland and watershed slope, vegetative cover, soil type, and surface type (impervious or pervious) require both contact and remote monitoring. Investment, Measures, and Area Any measure can be applied to any size monitoring area given a sufficient investment. Nevertheless, for many measures there is a point where either the investment required is too large to be practical or the return is diminished to the point where the information in nonrepresentative of actual conditions. Measures of the properties of individual plants, vegetation communities, soil, organisms, and aquatic physical and chemical properties are generally most effective at the microscale level, although vegetation community cover type is effectively applied to the mesoscale and macroscale and in rare cases to the megascale (Haddad, 1990). Measures of landform properties are most effectively applied to the mesoscale and ©2001 CRC Press LLC

macroscale, although these measures are applicable to the microscale. Hydrologic and hydraulic measures are most effectively applied at the microscale and mesoscale, but measures of flood frequency and surface water area are effective at the macroscale. Properties of individual wildlife species and wildlife communities are also most effectively measured at the microscale, although measures of abundance, density, and richness are sometimes applicable at the mesoscale. Investment, Measures, and Time Analogous to the relationship between measure and area, any measure can be applied over any length of time given a sufficient investment. Generally, contact monitoring measures require a greater investment in time, money, and skill than remote monitoring measures and are, therefore, more difficult to apply over long periods of time because of inevitable staff turnover and shifting priorities. Properties of individual plants, soil, organisms, and aquatic physical and chemical properties, some hydrologic and hydraulic measures, and individual and community wildlife properties are more effectively measured over relatively short periods of time. Monitoring programs to be conducted for relatively long periods of time are more effective if they use remote measures such as properties of landforms, vegetation community cover type, or in some cases wildlife abundance, density, or richness measures. Contact monitoring measures are applicable over long periods, even given a reasonable level of investment, if sample size and frequency are reduced. However, a reduction in investment inevitably reduces the monitoring return through reduction in result representativeness. No attempt should be made to minimize sample size and frequency without first assessing the impact to the return, and then determining if the reduction in return is acceptable. For example, a reduction in sample size or sample frequency will increase the variance in the data. This variance must be evaluated to determine if it obscures the information sought, that is, the data are too imprecise to be of practical use. Frequently, this reduction in monitoring investment results in the production of an index rather than a direct measure. An alternative approach to reducing sample size and sample frequency of contact measures is to establish a relationship between the contact measure of interest and a remote measure. In this way, the remote measure becomes an indicator of a wetland condition generally determined through a contact measure.

MONITORING DESIGN AND ANALYSIS Design Monitoring data is the product of either an experimental or an observational approach (Ludwig and Reynolds, 1988). An experimental approach, analogous to a true experiment (Hicks, 1982), presumes that the wetland is subject to experimental manipulation. That is, one or more independent variables are manipulated and their effect on one or more dependent variables is determined. Any differences among the dependent variables are attributed to the independent variables. By contrast, an ©2001 CRC Press LLC

observational approach uses measurements on the wetland over a range of natural conditions. The observer can either study and compare separate wetlands or different parts of the same wetland, subject to differing conditions at the same point in time. Alternatively, the same wetland or part of a wetland can be studied or compared at two separate points in time. Regardless of the approach, effective monitoring of wetlands requires an initial understanding of monitoring goals. This statement of the problem should include one or more criteria for assessing the results of the monitoring effort. The criteria should be measurable, and the achievable accuracy of the measure understood. If an experimental approach is used, then the independent variables must be defined and levels established. Consideration must be given as to how the data are to be collected. The extent of the difference to be detected and the degree of variability in the data will determine the number of observations required. In the absence of this information, the sample size should be as large as resource constraints will allow. Because there are always a number of variables that cannot be controlled, the order of data collection should be randomized. Randomization averages out the effects of time, and because randomization allows the monitoring effort to proceed as if the errors of measurement were independent, a key assumption of most statistical tests is satisfied (Hicks, 1982). Analysis Information obtained from a monitoring effort must be examined and interpreted. Statistics facilitate this examination and interpretation and are generally descriptive or inferential in nature. Descriptive statistics simply describe a sample, whereas inferential statistics allow inference, or the development of conclusions, based on a sample. Descriptive statistics include measures of central tendency such as mean, mode, and median. Measures of dispersion, such as standard deviation and variance, indicate how the data are distributed around the central tendency. Spatial pattern can be an important descriptor of wetland plant and animal communities. Three basic spatial patterns, random, clumped, and uniform, are recognized. Random patterns suggest environmental homogeneity or nonselective behavioral patterns, clumping suggests that individuals are aggregated in more favorable parts of the habitat, and uniform patterns suggest negative interactions between individuals (Pemberton and Frey, 1984). Analysis of spatial patterns is useful in that it suggests hypotheses that might be tested to explain underlying causal factors. Species-abundance relationships reveal information about the structure of the wetland community and lead to theories about such issues as community stability, resource partitioning, and species-area relationships (Hutchinson, 1959; McGuinness, 1984). Species abundances can be based upon number of individuals per species or, alternatively, on other variables such as biomass or percent cover. The simplest analysis of species-abundance relationships is a frequency distribution. Other characterizations of species-abundance relationships are indicated by indices of species richness, species evenness, and diversity. ©2001 CRC Press LLC

Patterns in species interactions are revealed by analyses of species affinity. Niche overlap indices indicate how coexisting species use a common resource. Measures of interspecific association indicate whether or not two species select or avoid the same habitat. When a sample contains quantitative measures of species abundances, the covariation in abundance between species can be assessed. Inferential statistics allow for testing of hypotheses about the data. If the population from which the sample was drawn can be assumed to have a known distribution and the sample items are independent of each other and normally distributed, then parametric statistics can be used. The t-test is used to compare the means of two samples (or the mean of a sample to a standard), and an analysis of variance (ANOVA) test is used to compare more than two means. In experimental situations where independent and dependent variables are defined, regression analysis is used to predict the effect of changing the independent variable on the dependent variable. Correlation analysis determines the degree of association between two factors of unknown relationship. When no assumption about the distribution of the population or the sample can be made, nonparametric statistics are appropriate. The MannWhitney U test and the Kruskal-Wallis k-sample test are analogous to the t-test and ANOVA, respectively, and a contingency table can be used to test for associations between variables. Community classification is the grouping or clustering of objects based upon their resemblance (Ludwig and Reynolds, 1988). Its use in monitoring of wetlands is in comparing a wetland with a reference wetland or in comparing the same wetland at two different points in time. Community classification is accomplished through the use of resemblance functions, association analysis, and cluster analysis. Community ordination is a term used to describe a set of techniques in which samples are arranged in relation to one or more coordinate axes, thereby facilitating identification of similar and dissimilar samples (Ludwig and Reynolds, 1988). Ordination techniques are intended to simplify and condense large data sets, and to identify key environmental factors. Techniques include polar ordination, principle component analysis, correspondence analysis, and various nonlinear ordinations. Community interpretation techniques provide for the evaluation of the effects of environmental factors on the patterns revealed by community classification and ordination. Interpretation of community classification data is accomplished with discriminate analysis, whereas interpretation of ordination data is accomplished through the use of regression and correlation statistics (Ludwig and Reynolds, 1988).

REFERENCES Adamus, P. and Brandt, K., Impacts on Quality of Inland Wetlands of the United States: A Survey of Indicators, Techniques, and Applications of Community Level Biomonitoring Data, U.S. Environmental Protection Agency Report EPA/600/3-90/073, 1990. Bosserman, R. W. and Hill, P. L., Community ecology of three wetlands, in Proceedings of the Pennsylvania State University Wetlands Water Management Mined Lands Conference, 1985.

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Bowen, G. W. and Burgess, R. L., A Quantitative Analysis of Forest Island Pattern in Selected Ohio Landscapes, Publication ORNL-TM-7759, Oak Ridge National Laboratory, 1981. Brooks, R. P., Arnold, D. E., Bellis, E. D., Keener, C. S., and Croonquist, M. J., A methodology for biological monitoring of cumulative impacts on wetland, stream and riparian components of watersheds, Draft presentation at International Symposium: Wetlands and River Corridor Management, 1989. Byrne, G. F. and Dabrowska-Zielinska, K., Use of visible and thermal satellite data to monitor an intermittently flooding marshland, Remote Sens. Environ., 11, 393, 1981. Carothers, S. W., Mills, G. S., and Johnson, R. R., The creation and restoration of riparian habitat in southeastern arid and semiarid regions, in Wetland Creation and Restoration: The Status of the Science, Report No. 600/3-89/038, Kusler, J. A. and Kentula, M. E., Eds., U. S. Environmental Protection Agency, Corvallis, OR, 1989, 359. Carter, V., Alsid, L., and Anderson, R. R., Man’s impact upon wetlands, in ERTS-1: A New Window on Our Planet, Geological Survey Professional Paper 929, Williams, R. S. and Carter, W. D., Eds., 1976, 293. Cartmill, R. H., Evaluation of remote sensing and automatic data techniques for characterization of wetlands, in Proceedings of the ERTS-1 Symposium, Washington, D.C., 1973, 1257. Choate, K. D., Watson, J. T., and Steiner, G. R., Demonstration of Constructed Wetlands for Treatment of Municipal Wastewaters: Monitoring Report for the Period March 1988–October 1989, Report No. TVA/WR/WQ-90/11, Tennessee Valley Authority, Knoxville, TN, 1990. Clausen, J. C. and Johnson, G. D., Lake level influences on sediment and nutrient retention in a lakeside wetland, J. Environ. Qual., 19(1), 83, 1990. Conner, W. H. and Toliver, J. R., Observations on the regeneration of bald cypress (Taxodium distichum) in Louisiana swamps, South. J. Appl. For., 14(3), 115, 1990. Cooperrider, A. Y., Boyd, R. J., and Stuart, H. R., Inventory and Monitoring of Wildlife Habitat, U.S. Department of the Interior, Bureau of Land Management, Service Center, Denver, CO, 1986. Costello, C. J., Wetlands treatment of dairy animal wastes in Irish drumlin landscape, in Constructed Wetlands for Wastewater Treatment, Hammer, D. A., Ed., Lewis Publishers, Chelsea, MI, 1991, 702. Cowardin, L. M., Carter, V., Golet, F. C., and LaRoe, E. T., Classification of Wetlands and Deepwater Habitats of the United States, Fish Wildlife Service Biological Report FWS/OBS-79/31, 1979. Cramer, G. H. and Hopkins, W. C., The Effects of Elevated Highway Construction on Water Quality in Louisiana Wetlands, Federal Highway Administration Report No. LA-75-4G-F, 1981. Dahl, T. E. and Pywell, H. R., National status and trends study: estimating wetland resources in the 1980’s, in AWRA Wetlands: Concerns and Successes Symposium, American Water Resources Association, 1989, 25. Delcourt, H. R. and Delcourt, P. A., Quaternary landscape ecology: relevant scales in space and time, Land. Ecol., 2, 23, 1988. Eger, P. and Kapakko, K., Use of wetlands to remove nickel and copper from mine drainage, in Proceedings of the Tennessee Valley Authority First Annual Conference on Constructed Wetlands for Wastewater Treatment, Tennessee Valley Authority, 1988, 780. Erwin, K. L., An evaluation of wetland mitigation in the South Florida Water Management District, Report to the South Florida Water Management District, Contract No. C89-0082-A1, 1991. Forman, R. T. T. and Godron, M., Landscape Ecology, John Wiley & Sons, New York, 1986. ©2001 CRC Press LLC

Graves, B. M. and Dittberner, P. L., Variables for Monitoring Aquatic and Terrestrial Environments, U.S. Fish and Wildlife Service Biological Report 86(5), 1986. Guntspergen, G., Keough, J., Stearns, F., and Wikum, D., ELF Communications System Ecological Monitoring Program: Wetland Studies—Final Report, Prepared for Submarine Communications Project Office, Technical Report E06620–2, Contract No. N00039-88-C-0065, Washington, D.C., 1989. Haddad, K., Marine wetland mapping and monitoring in Florida, Fish Wildlife Service Biological Report 90(18), 1990, 145. Hardy, J. W., Land Treatment of municipal wastewater on Mississippi Sandhill Crane National Refuge for wetlands/crane habitat management: a status report, in Proceedings of the Tennessee Valley Authority First Annual Conference on Constructed Wetlands for Wastewater Treatment, Tennessee Valley Authority, 1988, 186. Hawkins, A. S. and Stewart, J. L., Environmental Management Program: Long Term Resource Monitoring Program for the Upper Mississippi River System, Prepared for the Winona County Soil and Water conservation District, Lewiston, MN and the Fish and Wildlife Service, Onalaska, WI, Report No. EMTC-90/05, 1990. Hays, R. L., Summers, C., and Seitz, W., Estimating Wildlife Habitat Variables, Fish Wildlife Service Biological Report FWS/OBS-81/47, 1981. Hebert, C. E., Haffner, G. D., Weis, I. M., Lazar, R., and Montour, L., Organochlorine contaminants in duck populations of Walpole Island, J. Great Lakes Res., 16(1), 21, 1990. Henny, C. J., Anderson, D. R., and Pospahala, R. S., Aerial Surveys of Waterfowl Production in North America, 1955–71, U.S. Department of the Interior, Fish Wildlife Service Special Scientific Report—Wildlife 160, 1972. Hicks, C. R., Fundamental Concepts in the Design of Experiments, Holt, Rinehart and Winston, New York, 1982. Hink, V. and Ohmart, R. D., Middle Rio Grande Biological Survey Final Report, U.S. Army Corps of Engineers, Albuquerque, NM, 1984. Huber, W. C. and Dickinson, R. E., Stormwater Management Model, Version 4: User’s Manual, Cooperative Agreement CR-811607, U.S. Environmental Protection Agency, Athens, GA, 1988. Hutchinson, G. E., Homage to Santa Roasalia, or why are there so many kinds of animals? Am. Nat., 93, 145, 1959. Jean, M. and Bouchard, A., Temporal changes in wetland landscapes of a section of the St. Lawrence River, Canada, Environ. Manage., 15, 241, 1991. Johnston, J. B. and Handley, L. R., Coastal Mapping Programs at the U.S. Fish and Wildlife Service’s National Wetlands Research Center, Fish Wildlife Service Biological Report 90(18), Federal Coastal Wetland Mapping Programs, 1990, 105. Kent, D. M., Reimold, R. J., Kelly, J., and Tammi, C., Coupling wetlands structure and function: developing a condition index for wetlands monitoring, in Ecological Indicators, Vol. 1, McKenzie, D. H., Hyatt, D. E., and McDonald, J. V., Eds., Elsevier Science, Essex, England, 1992, 559. Koeln, G. T., Jacobson, J. E., Wesley, D. E., and Rempel, R. S., Wetland inventories derived from Landsat data for waterfowl management planning, in Proceedings of the 53rd Wildlife Management Institute of North American Wildlife and Natural Resources Conference, Wildlife Management Institute, 1988, 303. Kusler, J. A. and Kentula, M. E., Wetland Creation and Restoration: The Status of the Science, Island Press, Washington, D.C., 1990. Landin, M. C., Clairain, E. J., and Newling, C. J., Wetland habitat development and longterm monitoring at Windmill Point, Virginia, Wetlands, 9, 13, 1989.

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Larson, J. S., Wetland mitigation in the glaciated northeast: risks and uncertainties, in Mitigating Freshwater Wetland Alterations in the Glaciated Northeastern United States: An Assessment of the Science Base, Larson, J. S. and Neill, C. S., Eds., University of Massachusetts Environmental Institute Publication No. 87–1, 1987. Leibowitz, N. C., Squires, L., and Baker, J. P., Environmental Monitoring and Assessment Program: Research Plan for Monitoring Wetland Ecosystems, EPA Report No. EPA/600/3–91/010, 1991. Ludwig, J. A. and Reynolds, J. F., Statistical Ecology, John Wiley & Sons, New York, 1988. MacClintock, L., Whitcomb, R. F., and Whitcomb, B. L., Evidence for the value of corridors and minimization of isolation in preservation of biotic diversity, Am. Birds, 31, 6, 1977. Mackey, H. E. and Jensen, J. R., Wetlands mapping with spot multispectral scanner data, in Proceedings of the 9th Annual Convention on Surveying and Mapping Auto–Cartography, Baltimore, MD, 1989. McFarlane, C. and Watson, R. D., The detection and mapping of oil on a marshy area by a remote luminescent sensor, in Proceedings of the American Petroleum Institute Oil Spill Conference, American Petroleum Institute, New Orleans, LA, 1977, 197. McGuinness, K. A., Equations and explanations in the study of species-area curves, Biol. Rev., 59, 423, 1984. Meiorin, E. C., Urban runoff treatment in a fresh/brackish water marsh in Fremont, California, in Constructed Wetlands for Wastewater Treatment, Hammer, D. A., Ed., Lewis Publishers, Chelsea, MI, 1991, 677. Molini, W. A., Pacific flyway perspectives and expectations, in Proceedings of the 54th Conference of the Wildlife Management Institute for North American Wildlife and Natural Resources, Wildlife Management Institute, 1989, 529. Nayak, S., Pandeya, A., Gupta, M. C., Trivedi, C. R., Prasad, K. N., and Kadri, S. A., Application of satellite data for monitoring degradation of tidal wetlands of the Gulf of Kachchh Western India, Acta Astron., 20, 171, 1989. Neilson, J. D. and Green, G. H., Enumeration of spawning salmon from spawner residence time and aerial counts, Trans. Am. Fish. Soc., 110(4), 554, 1981. New Hampshire Water Pollution Control Commission, Assessment of Wetlands Management and Sediment Phosphorous Inactivation, Kezar Lake, New Hampshire: Phase 2 Implementation and Monitoring, Concord, NH, 1989. Newling, C. J. and Landin, M. C., Long-Term Monitoring of Habitat Development at Upland and Wetland Dredged Material Disposal Sites 1974–1982, Dredging Operations Technical Support Program Report No. WES/TR/D-85-5, 1985. Oberts, G. L. and Osgood, R. A., Water quality effectiveness of a detention/wetland treatment system and its effect on an urban lake, Environ. Manage., 15(1), 131, 1991. Orth, R. J., Moore, K. A., and Nowak, J. F., Monitoring seagrass distribution and abundance patterns: a case study from the Chesapeake Bay, in Fish Wildlife Service Biological Report 90(18): Federal Coastal Wetland Mapping Programs, U.S. Fish and Wildlife Service, 1990, 111. Pacific Estuarine Research Laboratory, A Manual for Assessing Restored and Natural Coastal Wetlands with Examples from Southern California, California Sea Grant Report No. T-CSGCP-021, La Jolla, CA, 1990. Paul, J. F., Holland, A. F., Schimmel, S. C., Summers, J. K., and Scott, K. J., USA EPA Environmental Monitoring and Assessment Program: an ecological status and trends program, U.S. Fish and Wildlife Service Biological Report, 90(18), 71, 1990. Pemberton, S. G. and Frey, R. W., Quantitative methods in ichnology: spatial distribution among populations, Lethaia, 17, 33, 1984.

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Peters, D. D., Status and trends of wetlands in the California Central Valley, in AWRA Wetlands: Concerns and Successes, American Water Resources Association, 1989, 33. Pielou, E. C., Mathematical Ecology, John Wiley & Sons, New York, 1977. Pritchett, D. A., Creation and monitoring of vernal pools at Santa Barbara, California, in Proceedings of Environmental Restoration: Science and Strategies for Restoring the Earth, University of California Press, Berkeley, CA, 1988, 282. Roughgarden, J., Running, S. W., and Matson, P. l. A., What does remote sensing do for ecology? Ecology, 72(6), 1918, 1991. Rump, P. C., The state of Canada’s wetlands, in Proceedings of the Peatlands Symposium, Association of State Wetland Managers, Edmonton, Canada, 1987, 259. Shortelle, A. B., Dudley, J. L., Prynoski, B., and Boyajian, M., Vernal pool wetlands: wildlife values, acidification and a need for management, in AWRA Wetlands: Concerns and Successes Symposium, American Water Resources Association, Tampa, FL, 1989, 463. Soil Conservation Service, Technical release no. 20 (TR-20), National Technical Information Service, 1982. Soil Conservation Service, Technical release no. 55 (TR-55), National Technical Information Service, 1986. Stark, L. R., Kolbash, R. L., Webster, H. I., Stevens, S. E., Dionis, K. A., and Murphy, E. R., The Simco #4 Wetland: biological patterns and performance of a wetland receiving mine drainage, in Mine Drainage and Surface Mine Reclamation, Vol. I, Information Circular No. 9183, U.S. Bureau of Mines, 1988, 332. Stillings, L. L., Gryta, J. J., and Ronning, T. A., Iron and manganese removal in a Typha dominated wetland during ten months following its construction, in Mine Drainage and Surface Mine Reclamation, Vol. I. Information Circular No. 9183, U.S. Bureau of Mines, 1988, 317. Tennessee Valley Authority, Design and Performance of the Constructed Wetland Wastewater Treatment System at Phillips High School, Bear Creek, Alabama, Report No. TVA/WR/WQ-90/5, Tennessee Valley Authority, Chattanooga, TN, 1990. U.S. Army Corps of Engineers, HEC-1 Flood Hydorgraph Package: User’s Manual, Water Resources Support Center, The Hydrologic Engineering Center, Davis, CA, 1981. U.S. Army Corps of Engineers, Evaluation of Freshwater Wetland Replacement Projects in Massachusetts, New England Division, Waltham, MA, 1989. U.S. Environmental Protection Agency, Environmental Monitoring and Assessment Program, EPA/600/3–90/060, 1990. Walker, W. W., Water quality trends at inflow to Everglades National Park, Water Res. Bull., 27, 59, 1991. Watson, M. R., Stone, W. B., Okoniewski, J. C., and Smith, L. M., Wildlife as monitors of the movement of polychlorinated biphenyls and other organochlorine compounds from a hazardous waste site, in Proceedings of the Northeast Fish and Wildlife Conference, 1985, 91. Weller, M. W., Birds of some Iowa wetlands in relation to concepts of faunal preservation, Proc. IA Acad. Sci., 86, 81, 1979. Weller, M. W. and Fredrickson, L. H., Avian ecology of a managed glacial marsh, Living Bird, 12, 269, 1974. Weller, M. W. and Voigts, D. K., Changes in the vegetation and wildlife use of a small prairie wetland following a drought, Proc. IA Acad. Sci., 90, 50, 1983. Wickland, D. E., Mission to planet earth: the ecological perspective, Ecology, 72(6), 1923, 1991.

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Woodward, D. F., Snyder-Conn, E., Riley, R. G., and Garland, T. R., Drilling fluids and the Arctic tundra of Alaska, U.S.A.: assessing contamination of wetlands habitat and the toxicity to aquatic invertebrates and fish, Arch. Environ. Contam. Toxicol., 17(5), 683, 1988. Wynn, S. L. and Kiefer, R. W., Monitoring vegetation changes in a large impacted wetland using quantitative field data and quantitative remote sensing data, in Proceedings of the Sensing of Environmental Pollutants 4th Joint Conference, New Orleans, LA, 1977, 178. Young, D. A., Petroleum extraction and waterfowl utilization within a major wetland complex: are they compatible? in Proceedings of the Peatlands Symposium, Edmonton, Canada, 1987, 165.

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DeBusk, Thomas A. et al “Wetlands for Water Treatment” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

9

Wetlands for Water Treatment Thomas A. DeBusk and William F. DeBusk

CONTENTS General Features of Wetlands that Contribute to Contaminant Removal Types of Treatment Wetlands Free Water Surface Wetlands Subsurface Flow Wetlands Hybrid Treatment Wetlands Treatment Wetland Components Treatment Wetland Vegetation Hydroperiod and Hydraulics Treatment Wetland Soils Treatment Wetland Contaminant Removal Processes Physical Removal Processes Biological Removal Processes Chemical Removal Processes Planning and Design Treatment Wetlands as a Unit Process Regulatory Issues Preliminary Feasibility and Alternatives Analyses Design Considerations Construction and Management Construction Management Performance Suspended Solids and Organic Carbon Removal Removal of Organic Carbon and Suspended Solids from Waste Stabilization Pond Effluents

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Nitrogen Removal Nitrogen Removal from Food Processing Wastewaters Phosphorus Removal Removal of Trace Metals, Toxic Organic Compounds, and Complex Mixtures of Contaminants Wetland Treatment of Urban Runoff Wetland Treatment of Landfill Leachates Pathogen Removal Conclusions References

Wetlands are widely regarded as biological filters, providing protection for water resources such as streams, lakes, estuaries, and groundwater. Although naturally occurring wetlands have always served as ecological buffers, research and development of wetland treatment technology is a relatively recent phenomenon. Studies of the feasibility of using wetlands for wastewater treatment were initiated during the early 1950s in Germany. In the United States, wastewater to wetlands research began in the late 1960s and increased dramatically in scope during the 1970s. As a result, the use of wetlands for water and wastewater treatment has gained considerable popularity worldwide. Currently, an estimated 1000 wetland treatment systems, both natural and constructed, are in use in North America (Cole, 1998). The goal of water and wastewater treatment is the removal of aqueous contaminants in order to decrease the possibility of detrimental impacts on humans and the rest of the ecosystem. The term contaminant is used in this context to refer to any constituent in the water or wastewater that may adversely affect human and environmental health. Many contaminants, including a wide variety of organic compounds and metals, are toxic to humans and other organisms. Other types of contaminants may not be hazardous in the conventional sense but nevertheless pose an indirect threat to our well being. For example, loading of nutrients (e.g., nitrogen and phosphorus) to waterways can result in excessive growth of algae and unwanted vegetation. This growth diminishes the recreational, economic, and aesthetic values of lakes, bays, and streams. Constructed wetlands have been successfully used as treatment systems for domestic wastewater effluent, from single-residence wetlands to large municipal wastewater treatment facilities. Similarly, wetlands may be used effectively for treatment of animal and aquaculture wastes. The use of wetland retention basins for treatment of stormwater runoff has become relatively commonplace. The composition of stormwater varies greatly, depending on the surrounding land use. For example, urban runoff may contain soil particles, dissolved nutrients, heavy metals, oil, and grease. Residential and agricultural runoff may also contain organic matter and pesticides. A variety of industrial wastes, including pulp and paper, food processing, slaughtering and rendering, chemical manufacturing, petroleum refining, and landfill leachates are amenable to wetland treatment.

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While wetlands are remarkable in their ability to treat diverse types of contaminated waters, there are limitations that should be carefully considered prior to the implementation of any wetland treatment system. Pretreatment, for example, primary sedimentation or anaerobic or aerobic stabilization, is often required prior to feeding domestic and industrial effluents into a wetland. Selected water contaminants removed within a treatment wetland ultimately may become available for assimilation and potentially be toxic to wetland biota. A sound understanding of wetland contaminant removal processes (Reddy and D’Angelo, 1997), the long-term fate of these contaminants, and contaminant removal effectiveness of various wetland types is critical in the proper design and operation of treatment wetlands.

GENERAL FEATURES OF WETLANDS THAT CONTRIBUTE TO CONTAMINANT REMOVAL The unique combination of structural and functional attributes sets wetlands apart from terrestrial and aquatic ecosystems in their ability to remove or sequester nutrients and toxic environmental contaminants. For example, shallow water, low current velocity, and the physical filtering action of plant stems and leaves provide favorable conditions for settling of particulate matter. Wetlands also provide substrates for a multitude of chemical and microbiological processes, promoting nutrient removal and storage within the complex maze of microsites in the soil and vegetation cover. The total surface area available for microbial activity in the soil and the overlying dead plant material (litter or detritus) is extremely high in wetlands. Physical, chemical, and microbiological processes are further enhanced in wetlands by retention of water for extended periods within this biologically active zone. Another important characteristic of wetlands is the presence of anaerobic (oxygendepleted) soils during periods of flooding which gives rise to an aerobic–anaerobic interface, or boundary, near the soil surface. This juxtaposition of aerobic and anaerobic conditions provides an environment for unique chemical and microbiological reactions that greatly enhance the removal of nutrients from inflowing water. Wastewater-borne labile organic carbon compounds, expressed as biochemical oxygen demand (BOD), are readily removed in anaerobic and aerobic microenvironments of treatment wetlands. Reduced nitrogen (N) compounds (e.g., ammonium) are nitrified in aerobic regions, from which the products can migrate (either by bulk transfer or diffusion) to anaerobic regions. Denitrification of the produced NOx species occurs rapidly due to the preponderance of anaerobic conditions and ready availability of labile carbon compounds from decaying vegetation and organic soils. In this sequential N removal process, nitrification typically is the rate-limiting step due to low oxygen availability in many parts of the system. Plant uptake and adsorption to soil surfaces contribute to short-term phosphorus (P) removal in wetlands. However, the only prominent, long-term P sink is thought to occur through soil accumulation. Large wetland areas are therefore required to achieve substantial P removal. Treatment wetlands differ in two fundamental ways from more conventional wastewater treatment unit processes. First, wetlands sacrifice consistently high ©2001 CRC Press LLC

microorganism densities and strict process control for reduced construction costs and operator attention. Second, solids processing occurs internally in wetlands, so no biological sludge management is required, at least on the short-term. For the above reasons, treatment wetlands have moderate to high land requirements. In summary, a number of physical, chemical, and biological processes operate concurrently in constructed and natural wetlands to provide contaminant removal (Figure 1). Removal of contaminants may be accomplished through storage in the wetland soil and vegetation or through losses to the atmosphere. Knowledge of the basic contaminant removal concepts is extremely helpful for assessing the potential applications, benefits, and limitations of wetland treatment systems. These processes are described in more detail in a later section.

TYPES OF TREATMENT WETLANDS Treatment wetlands are generally classified as either free water surface (FWS) or subsurface flow (SSF) systems (Figure 2). Subsurface flow wetlands are the common system design implemented in Europe for domestic wastewater treatment which has greater than 500 treatment wetlands. In North America, with around 600 treatment wetlands, the FWS type is more common (Cole, 1998). In the United States, FWS wetlands for domestic wastewater treatment commonly occur in communities with 1000 or fewer people, although some large FWS wetlands exist in cities with populations greater than 1 million. As of late 1998, South Dakota was the state with the greatest number of operational (nonpilot) FWS wetlands (42), followed by Florida (24) and California (11) (U.S. Environmental Protection Agency, 1999). The widespread use of treatment wetlands in South Dakota, a state with harsh winter conditions, provides a good indication of the versatility of treatment wetlands, as well as the circumstances under which they are an appropriate and competitive technology. Because of low capital (owing to inexpensive land) and operating costs and the ability to provide winter water storage, waste stabilization ponds (WSPs) have been widely implemented for domestic wastewater treatment during the past four decades. As of 1991, 246 communities in South Dakota were using WSPs. A drawback of WSPs is that they do not consistently provide low effluent suspended solids, ammonium, and total phosphorus concentrations. A strong interest in additional protection of the quality of water resources, as well as in creating new wildlife habitat, has led to the upgrading of many WSPs with treatment wetlands during the past decade (Dornbush, 1993). The South Dakota state regulatory agency has encouraged the use of constructed wetlands by providing design guidelines and economic assistance with treatment wetland construction. Free Water Surface Wetlands FWS design typically incorporates a shallow layer of surface water, flowing over mineral (sandy) or organic (peat) soils. Vegetation often consists of marsh plants, such as Typha (cattails) and Scirpus (bulrush), but may also include floating and submerged aquatic vegetation and wetland shrubs and trees (Figure 3). Natural ©2001 CRC Press LLC

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Volatilization NH3 Volatile organics

CO2, CH4

N2

Plant storage

Decomposition

Denitrification

Organic C

NO3-

Sedimentation

Inflow

SURFACE WATER DETRITUS (LITTER)

Plant uptake Burial

Soil storage (peat)

Figure 1

Adsorption - NH4+, metals, P, organics (to clays, Fe/Al hydroxides, organic matter)

SOIL

Precipitation - P (with Fe, Al, Ca) - Metals (with sulfides)

Contaminant removal mechanisms and transformations in free water surface (FWS) wetlands.

Outflow

Inflow

SURFACE WATER DETRITUS (LITTER) SOIL (MINERAL OR PEAT)

FREE WATER SURFACE WETLAND

WATER LEVEL DETRITUS (LITTER) Inflow

Outflow

GRAVEL OR SOIL

SUBSURFACE-FLOW WETLAND Figure 2

Schematic of free water surface (FWS) and subsurface flow (SSF) wetlands.

wetlands, both forested and herbaceous, have also been effectively used as FWS treatment wetlands. For some treatment applications, FWS wetlands are designed and managed to encourage dominance by either floating or submerged macrophytes. Water depth is one parameter that can be controlled to discourage emergent macrophytes, thereby allowing development of either a floating aquatic macrophyte (FAM) or submerged aquatic vegetation (SAV) system. Free water surface wetlands vary dramatically in size, from less than 1 ha to greater than 1000 ha. Large FWS wetlands are even being used as a nutrient control technology to treat runoff from entire regional watersheds. For example, over 16,000 ha of FWS wetlands are being constructed in South Florida to remove P from agricultural drainage water before it enters the Everglades (Moustafa et al., 1999). Free water surface wetlands offer ecological and engineering benefits beyond water treatment. Free water surface wetlands used for treating agricultural and urban ©2001 CRC Press LLC

Figure 3

Free water surface (FWS) wetlands incorporate a shallow layer of surface water and vegetation such as these emergent macrophytes.

runoff also reduce hydraulic runoff peaks from storm events. Effectiveness depends on the wetland size (volume), location in the watershed, and configuration of inlet and outlet structures (Schueler, 1996). Many FWS treatment wetlands provide both a recreational amenity and wildlife habitat. Iron Bridge wetland in Orlando, FL and the Arcata marsh in Humbolt, CA have each provided water treatment and other ecological and aesthetic benefits for more than a decade (Jackson, 1989; Gearhart and Higley, 1993). Subsurface Flow Wetlands Subsurface flow wetlands differ from FWS wetlands in that they incorporate a rock or gravel matrix that the wastewater is passed through in a horizontal or vertical fashion (see Figure 2). Unless the matrix clogs, the top layer of the bed in horizontal flow systems will remain dry. The SSF configuration offers several advantages, including a decreased likelihood of odor production and no insect proliferation within the wetland as long as surface ponding is avoided. Unlike FWS wetlands, SSF systems provide no aesthetic or recreational benefits and few, if any, benefits to wildlife. Subsurface flow wetlands continue to provide effective treatment of most wastewater constituents through the winter in temperate climates. The subsurface microbial treatment processes still function, albeit at a reduced rate, even when the surface vegetation has senesced or died, and the matrix surface is covered with snow and ice. Subsurface flow wetlands also can be operated in a vertical flow fashion which ©2001 CRC Press LLC

can reduce matrix clogging problems and enhance certain contaminant removal processes such as nitrification. Because of the high cost of the gravel or rock matrix, SSF wetlands never attain the large spatial footprint of the large FWS wetlands. Concerns over matrix clogging and the potential high cost of renovation also limit the deployment of extremely large SSF wetlands. However, SSF are finding increased use for small applications, such as for small communities or single family homes. The limitations of septic systems for nutrient control have become more apparent in the past two decades (Hagedorn et al., 1981), and SSF wetlands are one technology that is being deployed to improve nutrient removal performance (Mitchell et al., 1990). Subsurface flow systems are the only wetland configuration suitable for this purpose, because they create no standing water, thereby limiting the likelihood of human exposure to wastewater pathogens (House et al., 1999). Hybrid Treatment Wetlands A number of treatment wetlands have been constructed that combine different wetland types. Some, such as the 134-ha Eastern Service Area wetland in Orlando, FL, consist of constructed FWS wetlands that are followed by natural forested wetlands. This particular configuration was based on regulatory needs, with the natural parcel receiving water only after pretreatment by the constructed wetland (Schwartz et al., 1994). Other hybrid systems are based on specific contaminant removal needs. For example, the key to enhanced nitrogen removal in SSF wetlands is to create an intermediate step in the process train with an oxygenated environment that enhances nitrification in the rock or gravel matrix. Many subsurface wetlands accomplish this by sequencing horizontal flow beds with vertical flow beds or by recirculating partially treated effluent onto an inflow region rock or sand filter to enhance nitrification (Cooper et al., 1997; Reed and Brown, 1995). Other investigators have recommended operating vertical flow beds with various draw and fill cycles to enhance chemical oxygen demand (COD) and nitrogen removal, but performance using this technique has been mixed (Burgoon et al., 1995; Boutin et al., 1997). Regardless of the approach used to stimulate nitrification, the SSF wetland is configured so that the nitrate-rich effluent subsequently is introduced into an anoxic section of the bed where denitrification readily occurs.

TREATMENT WETLAND COMPONENTS Treatment Wetland Vegetation Macroscopic vegetation is the most prominent feature of treatment wetlands. Free water surface wetlands can develop as simple monocultures of weedy or competitive species such as Typha (cattail) or Phragmites (reed), but more often they contain a diversity of other emergent and floating plants within genera such as Pontederia, Sagittaria, Eleocharis, Utricularia, and Lemna. Treatment wetlands typically are planted just prior to initial flooding to ensure rapid vegetative cover ©2001 CRC Press LLC

development and to facilitate initiation of water treatment. Under extreme operating conditions, such as high organic, nutrient, or hydraulic loading rates, the system may remain a monoculture or near monoculture for the life of the treatment system. Similarly, subsurface flow wetlands usually remain dominated by the species planted prior to startup, typically Phragmites (reed), Scirpus (bulrush), or Typha (cattail). This is due both to the difficulty of seeds and other propagules in becoming established on the bed’s surface and the often high organic loading provided to SSF systems. Under less rigorous environmental conditions, the vegetative community that develops over time in FWS wetlands may bear little resemblance to the species originally planted. At the Eastern Service Area Treatment Wetland (Orlando, FL), one constructed wetland is quite shallow and experiences periodic drydown, and the second, while also shallow, is continuously inundated. This system is used for further polishing of domestic wastewater that has received conventional, advanced treatment to levels below 5 mg BOD/l, 5 mg total suspended solids (TSS)/l, 3 mg N/l, and 1 mg P/l (Schwartz et al., 1994). Upon wetland startup, 13 species were planted into the mineral soils at a density of 336 plants per ha. The vegetative communities in the two constructed wetlands were sampled at 1 and 4 years after planting. Only one of the most abundant species occurring at year four in each wetland was a species that was originally planted (Table 1). In a wetland ecosystem self-organization experiment, Mitsch et al. (1998) established two individual 1-ha wetlands for treating Olentangy River water in Ohio. One wetland was planted with 2400 propagules (rootstock and rhizomes), representing 13 plant species at an overall density of 0.24 plants/m2. Species planted included Nelumbo lutea, Nymphaea odorata, and Potamogeton pectinatus in the deepest (0.6 m depth) region, Scirpus validus and Scirpus fluviatilis at moderate (0.3 m) depth, and Spartina pectinata, Sparganium eurycarpum, Acorus calamus, Sagittaria latifolia, Pontedereria cordata, Juncus effusus, Saururus cernuus, and Cephalanthus occidentalis in the shallow, littoral (0 to 0.3 m) region. The second wetland, adjacent to the first, was left unplanted. The wetlands were evaluated each year after startup for water treatment aspects and flora and fauna characteristics. As of year three, 9 of the 13 original stocked species in the planted wetland were still present, although the total number of macrophyte species had increased to 65 (Mitsch et al., 1998). The unplanted wetland had similar vegetation at year three, and had 54 macrophyte species. However, only 1 of the 13 species originally planted in the adjacent wetland were present. Treatment performance and fauna of the two systems were remarkably similar at year three. Whether or not to plant a FWS treatment wetland upon start-up, as well as what density to plant, is dictated by the urgency to achieve an operational system. If the system is built a year or two in advance of water or wastewater treatment needs, and the design does not call for any specific vegetation components, then existing studies clearly show that natural recruitment can be relied upon for development of a diverse plant community. Depending on the depth and nutrient regime of the wetland, mats of filamentous algae, phytoplankton, or submerged macrophytes likely will dominate in the wetland water column prior to development of a dense emergent macrophyte community. ©2001 CRC Press LLC

Table 1

Characteristics of Herbaceous Vegetation Communities in Two Shallow Eastern Service Area Wetlands, Orlando, FL P/R

Frequency Rank Year 1 Year 4

Wetland #1 Mikania scandens Baccharis halimifolia Myrica cerifera Hydrocotyle umbellata Ludwigia repens Galium tinctorium Setaria geniculata Ptilimnium capillaceum Juncus effusus Andropogon golmeratus

R R R R R R R R P R

NF* 28 35 19 26 55 43 27 39 86

1 2 3 4 5 6 7 8 9 10

R R R R R R R R P R

3 118 10 13 NF 19 41 16 33 1

1 2 3 4 5 6 7 8 9 10

Wetland #2 Hydrocotyle umbellata Mikania scandens Panicum repens Typha domingensis Cyperus compressus Ludwigia repens Fuirena scirpoidea Juncus scirpoides Sagittaria lancifolia Rhynchospora microcephala

Notes: Wetland #1 experiences periodic drydown and wetland #2 is continuously inundated. Values represent frequency of occurrence rankings in May 1988 and December 1992, 1 and 4 years after planting. The first column denotes whether the species was originally planted (P) or naturally recruited (R). * NF means not found. Source: Wallace, Ecosystem Research Corporation, Gainesville, FL.

From a water treatment perspective, macrophyte vegetation provides a number of functions (Brix, 1997). In FWS wetlands, dense macrophyte stands shade the water column reducing or eliminating phytoplankton populations. Conversely, in sparse macrophyte stands, the emergent stems may serve as attachment sites for periphytic algae. In either case, the submerged plant portions also provide surface area for colonization by bacteria that contribute to processing of carbon, nitrogen, and other wastewater constituents. Emergent and floating macrophytes shield the water from direct sunlight and, therefore, moderate the temperature of the shallow water column. These plants also tend to dissipate or block wind and wave energy and, therefore, help maintain quiescent conditions in the water column. This promotes settling of wastewaterborne solids and inhibits resuspension of flocculant sediments from the bottom. ©2001 CRC Press LLC

Some floating species, such as the duckweeds, provide a dense floating mat on the water’s surface that can inhibit oviposition by mosquitoes and even act as a barrier to prevent escape of odors produced in the bottom sediments or water column. The contribution of belowground plant tissues, rhizomes, and roots to wastewater treatment depends on the system configuration. In FWS wetlands, the roots of emergent plants assimilate nutrients from the soil porewater and also provide an anchoring function that can reduce erosion. In a SSF wetland or a FWS wetland dominated by floating plants, the roots are in intimate contact with the wastewater and O2 leakage from the roots can enhance microbial processes (e.g., nitrification) that require oxic conditions. Oxygen is transported internally to the plant roots either by passive molecular diffusion or convective (bulk) flow of air through the internal lacunae. The convective air flow can be driven by a number of physical processes, including wind velocity gradients in the plant canopy and temperature or humidity differences between the interior and exterior of the plant (Brix, 1993, 1994). In SSF and FAMs systems, the plant roots are in intimate contact with the wastewater. In these systems, the diffuse root mats harbor bacteria which may also benefit from oxygen transported from the foliage to the rhizosphere. For these systems, a number of investigations have been conducted to determine the effectiveness of various vegetation types on treatment performance. Gersberg (1985) conducted one of the first studies to compare effectiveness of cattail (Typha) and bulrush (Scirpus) for ammonia removal in a SSF treatment wetland. Using large gravel bed systems receiving domestic wastewater, these investigators found that bulrush (Scirpus) beds provided greater ammonia removal than either unvegetated beds or beds containing cattail (Typha). A nitrogen mass balance revealed that macrophyte uptake accounted for only a small percentage of the N removed from the wastewater, so the higher nitrification rate for the Scirpus bed was attributed to this specie’s greater oxygen transport capacity. Results from numerous small-scale studies have demonstrated that the species of plants used in treatment wetlands can affect system contaminant removal performance, particularly for the nutrients N and P. Both emergent and floating macrophytes have been rigorously characterized with respect to their N and P uptake capability. For most nutrient-laden waters in moderate climates, water hyacinth (Eichhornia crassipes) and water lettuce (Pistia stratiotes) provide the highest rates of N and P uptake among floating species (DeBusk et al., 1996a; Reddy and DeBusk, 1985). Among emergent macrophytes, cattail (Typha), bulrush (Scirpus), and reed (Phragmites) provide some of the highest N and P removal rates. Short-term P uptake rates in excess of 37 g P/m2 per year have been reported for these productive floating and emergent species (Reddy and DeBusk, 1985; Tanner, 1996). Despite obvious between-species differences in nutrient uptake, assessments of full-scale treatment wetlands reveal few differences in contaminant removal performance between wetlands dominated by different plant species. This is because in the long-term, most of the nutrients assimilated by the plant standing crop are recycled by plant senescence, detritus production, and decomposition back into the water column and sediment compartments. What does seem to affect contaminant removal performance on a large scale is overall plant habit, that is, whether emergent, submerged, or floating species dominates the plant community. These differences ©2001 CRC Press LLC

in habit influence water column and sediment-water interface environmental conditions, such as dissolved oxygen concentrations and solar radiation inputs, that are the critical master factors in controlling element cycling and contaminant removal in wetlands (Reddy and D’Angelo, 1997; Reddy et al., 1999). Finally, where prominent plant morphological differences exist (or differences in survival) in a given waste stream, performance differences for a particular plant habit (e.g., floating or emergent) will occur. For example, floating aquatic macrophyte systems dominated by large-leafed water hyacinth and pennywort provide superior BOD removal performance compared to small-leafed duckweeds. This is likely because of the greater surface area of underwater rhizomes and roots and greater foliage to rhizosphere O2 transport by the larger-leafed floating species (Clough et al., 1987). Hydroperiod and Hydraulics In the natural landscape, hydroperiod (frequency and duration of flooding) is a prominent factor that dictates wetland occurrence and characteristics. Factors that influence natural wetland hydroperiod include surface water and groundwater inputs and losses. Additionally, the total annual volume of rainfall and evapotranspiration, as well as the seasonal distribution of these atmospheric water fluxes, will influence the type of wetland that occurs. Hydroperiod characteristics are less of a concern for wetlands that receive a relatively constant hydraulic loading, such as from a domestic wastewater source. Such wetlands often are isolated from groundwaters. Moreover, evapotranspiration and rainfall generally need to be accounted for only in assessing their influence on the effluent quality and mass contaminant removal budget and to ensure storage for large rainfall events (e.g., 25- or 100-year storm events). Atmospheric water fluxes also must be carefully understood and addressed where unusual climatic (extremely low annual rainfall) or site-specific conditions (highly permeable soils) exist. Treatment wetland hydraulics relate to the ability of the wetland to physically accommodate water inputs, as well as internal reactor design characteristics that contribute to contaminant removal. Because FWS wetlands are shallow basins, typically 0.5 to 1.5 m deep with the water column occupied in part by macrophytes, water passing through the wetland is subject to a certain amount of friction-induced headloss. Shallow areas and areas with dense vegetation provide the most resistance to flow. Headloss in FWS wetlands will not be a major design concern unless the hydraulic loading rate is unusually high, the wetland aspect ratio (length to width ratio) is high, or the flow path extremely long. By contrast, careful hydraulic design is paramount in SSF wetlands where all of the flow is being routed through a gravel or rock matrix. Common parameters related to the hydraulic design of treatment wetlands include hydraulic loading rate (HLR, usually expressed in cm/day) and hydraulic retention time (HRT, units usually in days). The former is obtained by dividing the flow (Q ) by the wetland area (A); the latter is calculated by dividing the flow by the water volume (V ) of the wetland. Performance forecast modeling of treatment wetlands is based on the concept that these systems behave as plug-flow reactors, with flow moving in lock step ©2001 CRC Press LLC

through the treatment wetland. However, tracer studies conducted with emergent, floating, and submerged macrophyte-dominated treatment wetlands (DeBusk et al., 1990; Kadlec and Knight, 1996) reveal that flow patterns depart widely from ideal plug-flow characteristics. Temperature-related water column density gradients, the heterogeneous and clumped nature of vegetation, and uneven microtopographical features result in the development of rapid flow paths and internal dispersion and mixing (Kadlec, 1990). The net outcome is that some of the influent water reaches the effluent end of the system long before the calculated hydraulic retention time (HRT), and a considerable amount is held longer than the calculated HRT. From a performance-forecasting standpoint, these deviations from plug-flow have been addressed by using different hydraulic reactor models. For example, HRT has been modeled using several continuously stirred tank reactors (CSTRs) in series, or a plug-flow reactor followed by multiple CSTRs (Kadlec, 1997; King et al., 1997). Recognition of nonideal flow characteristics, as documented by full-scale tracer studies, has led to most treatment wetlands being designed with a means of evenly distributing the influent across the entire width of the wetland. Once water enters the wetland, however, flows coalesce into small rills that then combine to create large short-circuiting channels. These flow channels typically remain intact until the water is redistributed by structural means. Both deep channels and earthen berms perpendicular to flow have been used to redistribute water in wetlands (Kadlec and Knight, 1996). However, neither rational design parameters nor performance benefits for these structural modifications have been rigorously characterized. Nevertheless, large treatment wetlands are often compartmentalized for purposes of improving flow distribution, as well as to facilitate dry-down and maintenance of selected portions of the treatment system (Figure 4). Treatment Wetland Soils Surface flow treatment wetlands can be constructed on both mineral and organic soils. Organic soils are generally categorized as having greater than 12 to 20 percent organic matter content, a pH less than 6.0, low bulk density, and high water holding and cation exchange capacities (Faulkner and Richardson, 1989). Mineral soils have a low organic matter content, high bulk density, and often provide greater nutrient availability than organic soils. Soils in FWS treatment wetlands serve several functions. Soils must provide appropriate physical and chemical support for the emergent macrophytes. The soil should have physical properties that facilitate planting and recruitment of the aquatic vegetation and that physically can support the plants under flooded conditions. The soil also must provide adequate nutrition to support macrophyte growth, particularly when treating waters of unusual chemical composition (e.g., industrial wastewaters) that may be deficient in certain plant macro- and/or micronutrients. Treatment wetland subsoils must exhibit low permeability. In most instances, unless treating exceptionally clean waters or effluents, the treatment wetland should be isolated from groundwaters (aquifers). Therefore, the soil profile should contain an impermeable layer (hydraulic conductivity of less than 10–6 cm/s) that inhibits vertical water movement. If the soil profile is overly permeable and the internal ©2001 CRC Press LLC

Figure 4

The Iron Bridge wastewater polishing wetland in Orlando, FL has compartments to improve flow distribution.

wastewaters must be isolated from groundwaters, then a clay, bentonite, or vinyl liner can be incorporated into the wetland. Permeability of the site soils will also dictate their suitability for use in constructing berms. Because soil is a prominent storage component (for nutrients etc.) in a FWS wetland, a soil type must be selected that does not add undesirable contaminants to the overlying water column. Extreme examples of soil nutrient export have actually resulted in treatment wetland use being curtailed. Operation of a forested wetland in central Florida that received advanced secondary domestic effluent was halted due to an export of P and organic N. This hydrologically altered wetland had been dry for more than 10 years prior to rehydration with wastewater effluent. Water quality transects within the wetland and laboratory soil column incubations revealed that the exported compounds resulted from flooding the highly oxidized soils. Almost 1 year of water exchange was required to reduce export of soil constituents to the water column (Figure 5). Some of the large wetlands in South Florida constructed on muck soils, previously used for agricultural crops, have also exhibited an increase in water column nutrient levels upon flooding. The Everglades Nutrient Removal Project, a 1370-ha wetland designed to reduce P in agricultural drainage waters from the range of 150 to 200 µg/l to 50 µg/l, exhibited water column P levels up to 370 µg/l in the first 2 months after flooding (Koch, 1991). Within 10 months, water column P concentrations had declined to 46 µg/l. Experience with these wetlands demonstrates that historical land use and soil management practices (e.g., fertilization of agricultural fields) can influence wetland water quality during the start-up phase. ©2001 CRC Press LLC

Figure 5

Profiles of total organic carbon (TOC) with distance through a forested treatment wetland 3 and 6 months after flooding with advanced secondary domestic wastewater effluent. Prior to rehydration, the wetland had been hydrologically altered (dry) for more than 10 years.

Regardless of the system design approach and type of vegetation that is encouraged, large FWS wetlands usually provide some habitat for fauna. Previous land use practices, therefore, should be assessed not only based on nutrient control, but also from a wildlife health aspect. For example, residual pesticides in previously farmed vegetable farms soils may be released into the surface water upon flooding, creating a hazard to waterfowl and other wildlife using the wetland.

TREATMENT WETLAND CONTAMINANT REMOVAL PROCESSES An understanding of wetland contaminant removal processes can facilitate wetland design, and aid dramatically in system troubleshooting should contaminant removal performance not be as expected. Wetlands provide effective transformation and storage of many water-borne constituents. Contaminants removed from the inflow are either re-exported in an aqueous, but more innocuous form (e.g., chlorine to chloride), are stored in the sediments (e.g., P, metals), or are lost from the system in a gaseous form (e.g., methane, carbon dioxide, nitrogen gas). A number of physical, biological, and chemical processes are responsible for contaminant removal in wetlands. Physical Removal Processes Wetlands are capable of providing highly efficient physical removal of contaminants associated with particulate matter in the water or waste stream. Surface water typically moves very slowly through wetlands due to the characteristic broad sheet flow and the resistance provided by rooted and floating plants. Sedimentation of ©2001 CRC Press LLC

suspended solids is promoted by the low flow velocity and by the fact that the flow is often laminar (not turbulent) in wetlands. Mats of floating plants in wetlands may serve, to a limited extent, as sediment traps, but their primary role in suspended solids removal is to limit resuspension of settled particulate matter. Efficiency of suspended solids removal is proportional to the particle settling velocity and the length of the wetland. For practical purposes, sedimentation is usually considered an irreversible process, resulting in accumulation of solids and associated contaminants on the wetland soil surface. However, resuspension of sediment may result in the export of suspended solids and yield a somewhat lower removal efficiency. Some resuspension may occur during periods of high flow velocity in the wetland. More commonly, resuspension results from wind-driven turbulence, bioturbation (disturbance by animals and humans), and gas lift. Gas lift results from production of gases such as oxygen from photosynthesis, and methane and carbon dioxide produced by microorganisms in the sediment during decomposition of organic matter. For some wetlands, build-up of sediment to detrimental levels can occur, necessitating dry-down and sediment consolidation. Biological Removal Processes Biological removal processes represent a prominent pathway of contaminant removal in wetlands. Probably the most widely recognized biological process for contaminant removal in wetlands is plant uptake. Wetland plants readily take up contaminants that are also essential nutrients, such as nitrate, ammonium, and phosphate. However, certain wetland plant species are also capable of uptake and even significant accumulation of certain toxic metals such as cadmium and lead. The rate of contaminant removal by plants varies widely, depending on the plant growth rate and the concentration of the contaminant in the plant tissue. Woody plants, that is, trees and shrubs, provide relatively long-term storage of contaminants compared with herbaceous plants. However, contaminant uptake rate per unit area of land is often much higher for herbaceous macrophytes such as Typha. Algae may also provide a significant amount of nutrient uptake but are relatively susceptible to the toxic effects of heavy metals. Storage of nutrients in algae is relatively short-term, due to the rapid turnover (life cycle) of the algae. Bacteria and other microorganisms also provide uptake and short-term storage of nutrients and some other contaminants in the soil. As plants age and eventually die, dead plant material, known as detritus or litter, accumulates at the soil surface. Some of the nutrients, metals, or other elements previously removed from the water by plant uptake are lost from the plant detritus and recycled back into the water. Leaching of water-soluble contaminants may occur rapidly upon the death of the plant or plant tissue, while a more gradual loss of contaminants occurs during decomposition of detritus by bacteria and other organisms. Recycled contaminants may be flushed from the wetland in the surface water or may be removed again from the water by biological uptake or other means. In most wetlands, there is a net accumulation of plant detritus because the rate of decomposition is substantially reduced relative to upland ecosystems by the low availability of oxygen for the decomposers. Anoxic and anaerobic conditions ©2001 CRC Press LLC

generally prevail in wetland soils because the extremely low diffusion rate of oxygen in water (approximately 10,000 times slower than in air) is not sufficient to replenish the oxygen consumed by the microbial decomposers. Therefore, decomposition of the detritus is not complete, resulting in accumulation and burial of partially decomposed organic matter. In this manner, some of the contaminants originally taken up by plants can be trapped and stored as peat. Peat may accumulate to great depths in wetlands and can provide long-term storage for contaminants. However, peat is also susceptible to decomposition if the wetland is drained or soils otherwise exposed, in which case the contaminants incorporated in the peat may be released and either recycled or flushed from the wetland. Although microorganisms may provide a measurable amount of contaminant uptake and storage, their metabolic processes play a much more significant role in the removal of organic compounds. Microbial decomposers, primarily soil bacteria associated with the native organic matter in wetlands, use organic carbon (C) as a source of energy converting it to carbon dioxide (CO2) or methane (CH4) gases. This affords a biological mechanism for removal of a wide variety of organic C compounds including those found in municipal wastewater, food processing wastewater, pesticides, and petroleum products. The efficiency and rate of organic C degradation by microorganisms are highly variable among different types of organic compounds. Microbial metabolism also affords removal of inorganic nitrogen, that is, nitrate and ammonium, in wetland soils. Certain bacteria (e.g., Pseudomonas spp.) metabolically transform nitrate into nitrogen gas (N2), a process known as denitrification. The N2 is subsequently lost to the atmosphere, thus denitrification represents a means for permanent removal, rather than storage, of nitrogen by the wetland. Removal of ammonium in wetlands can occur as a result of the sequential processes of nitrification and denitrification. Nitrification, the microbial (Nitrosomonas and Nitrobacter) transformation of ammonium to nitrate, takes place in aerobic regions of the soil and surface water. The newly formed nitrate can then undergo denitrification when it diffuses into the deeper, anaerobic regions of the soil. The coupled processes of nitrification and denitrification are universally important in the cycling and bioavailability of nitrogen in wetland and upland soils. Chemical Removal Processes In addition to physical and biological processes, a wide range of chemical processes are involved in the removal of contaminants in wetlands. The most important chemical removal process in wetland soils is sorption, which results in short-term retention or long-term immobilization of several classes of contaminants. Sorption is a broadly defined term for the transfer of ions (or molecules with positive or negative charges) from the solution phase (water) to the solid phase (soil). Sorption actually describes a group of processes that include adsorption and precipitation reactions. Adsorption refers to the attachment of ions to soil particles, either by cation exchange or chemisorption. Cation exchange involves the physical attachment of cations, or positively charged ions, to the surfaces of clay and organic matter particles in the soil. Cations are bonded to the soil by electrostatic attraction, a much weaker ©2001 CRC Press LLC

force than chemical bonding; therefore, the cations are not permanently immobilized. Many contaminants in wastewater and runoff exist as cations, including ammonium (NH4+, a plant nutrient) and trace metals such as copper (Cu2+). The capacity of soils for retention of cations, expressed as cation exchange capacity (CEC), generally increases with increasing clay and organic matter content. Chemisorption represents a stronger and more permanent form of bonding than cation exchange. A number of metals and organic compounds can be immobilized in the soil via chemisorption with clays, iron (Fe), aluminum (Al) oxides, and organic matter. Phosphate can also bind with clays and Fe and Al oxides through chemisorption. Phosphate can also precipitate with iron and aluminum oxides to form new mineral compounds (Fe- and Al-phosphates) that are potentially very stable in the soil and afford long-term storage of phosphorus. In the Everglades and other wetlands that contain high concentrations of calcium (Ca), phosphate can precipitate to form Ca-phosphate minerals which are stable over a long period of time. Another important precipitation reaction that occurs in wetland soils is the formation of metal sulfides which are highly insoluble and are, therefore, an effective means for immobilizing many toxic metals in wetlands. Volatilization, which involves diffusion of a dissolved compound from the water into the atmosphere, is another potential means of contaminant removal in wetlands. Ammonia (NH3) volatilization can result in significant removal of nitrogen if the pH of the water is high (greater than about 8.5). However, at neutral or low pH, ammonia nitrogen exists almost exclusively in the ionized form (ammonium, NH4+ ) which is not volatile. Many types of organic compounds are volatile and are readily lost to the atmosphere from wetlands and other surface waters. Although volatilization can effectively remove certain contaminants from the water, it may prove to be undesirable in some instances, due to the potential for polluting the air with the same contaminants.

PLANNING AND DESIGN Constructed and natural wetlands have been used extensively to treat many types of wastewaters and other contaminated waters such as urban and agricultural runoff. High levels of removal can be achieved for a number of contaminants, including suspended solids, nutrients, metals, and organic compounds, in treatment wetlands. However, there are inherent limitations to the effectiveness of wetlands for wastewater treatment. In some cases, it may not be possible to achieve the desired level of concentration reduction due to natural background levels. Also, there is a relatively high degree of time-dependent variability in treatment efficiency associated with wetlands, especially when compared with conventional treatment technologies (Kadlec, 1997). Because the contaminant removal interactions among vegetation, soils, and hydrologic wetland components are complex, contaminant removal efficiency varies widely among the types of treatment wetlands. Moreover, the actual pollutant loading that the treatment wetland can accommodate also varies. Treatment performance criteria for contaminant removal in wetlands may be based on the contaminant concentration in the wetland outflow or on the total or ©2001 CRC Press LLC

percent mass removal of the contaminant. As a case in point, the efficiency of nutrient removal decreases significantly as inflow concentration approaches the natural background concentration of the nutrient in the wetland, even though the outflow concentration may be well within the desired range. Conversely, nutrient removal efficiency, in terms of mass removal, may increase substantially as the loading rate is increased to moderate levels, yet the outflow concentration may exceed the desired level. It is important that the selected criteria accurately reflect the actual performance of the wetland relative to the objectives and intended uses of the wetland treatment system. The actual performance of treatment wetlands is generally dependent on a multitude of factors, including inflow concentration, mass loading rates, wetland design, and climate.

TREATMENT WETLANDS AS A UNIT PROCESS For proper design and operation of treatment wetlands, it is important that they be considered part of an overall water treatment train. Because of this, the contaminant removal effectiveness of the treatment wetland will be influenced by the performance of the upstream unit processes. Examples of upstream unit processes for wetlands include sedimentation ponds or deep forebays for wetlands treating urban runoff and primary clarifiers and/or secondary treatment processes (e.g., activated sludge) for wetland treatment of domestic wastewaters. Similarly, treatment wetlands are not always the final unit process in a treatment train. Wetlands can be followed by filtration or disinfection processes. The treatment train concept is critical to consider for wetland treatment for two reasons. First, an overall assessment of the strengths and weaknesses of the unit processes allows for a financially as well as technically optimized system. Second, the performance of the treatment wetland will be dictated in part by the quality (e.g., average, variability) of the water discharged by the upstream unit processes. For example, the treatment train might include a conventional, advanced secondary domestic wastewater treatment plant (WWTP) to nitrify ammonium, a treatment wetland designed to provide further N removal through denitrification, and disinfection by chlorine. If during the lifetime of the treatment system the WWTP becomes overloaded and fails to nitrify the wastewater ammonium, then the wetland’s effluent quality will dramatically decline. In turn, ammonium discharged from the wetland would reduce the effectiveness (or increase the cost) of the subsequent disinfection unit process. Regulatory Issues Federal, state, and local regulations must be carefully reviewed before using a constructed or natural wetland for water treatment in the United States. Almost all point discharges into wetlands, whether municipal, agricultural, or industrial wastewaters, will require a National Pollutant Discharge Elimination System (NPDES) permit. Federal regulations on the use of constructed wetlands for wastewater

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treatment are not overly restrictive, and the wetland essentially can be considered to be a unit process in the wastewater treatment train. By contrast, federal regulations discourage the use of natural wetlands for wastewater treatment because wetlands are considered waters of the United States and should not be used for waste assimilation or transport (U.S. Environmental Protection Agency, 1990). However, natural wetlands have been, and still are, used for wastewater treatment, and the U.S. Environmental Protection Agency recognizes this by noting that natural wetlands can be used for this purpose under limited conditions. The promoter of a natural wetland-based treatment system, therefore, will have to successfully argue that the wastewater introduction will not interfere with designated and existing uses of the wetland or downstream waters. In many cases, this means that prior to discharge into the natural wetlands, concentrations of oxygen-demanding substances, solids, nutrients, and heavy metals of a water source must be extremely low and that water introduction will not significantly alter the wetland hydroperiod. These restrictions, of course, can limit the effectiveness of the natural wetland as a unit treatment process. Florida, a state with a long history of regulating the use of wetlands for water treatment, has developed a Wetland Application Rule that recognizes four distinct types of wetlands used for domestic wastewater treatment: natural receiving wetlands, natural treatment wetlands, hydrologically altered treatment wetlands, and constructed wetlands. Depending on the classification, the rule places limits on the quality of the effluent introduced into the wetland, the hydraulic loading rate, the N and P loading rate, and the ongoing biological monitoring requirements to ensure no disruption of system integrity. Natural receiving wetlands, the most protected category, can only receive tertiary treated effluent at loadings not to exceed 5 cm (2 in.) of water per week and 25 g N and 3 g P per m2 per year. Constructed wetlands, the least restrictive category, have no fixed hydraulic or nutrient loading limits, but the wetland designer must provide reasonable assurance that the system will provide treatment and will function as a viable wetland habitat. Preliminary Feasibility and Alternatives Analyses A number of site-specific issues must be addressed siting a treatment wetland. Information needs include characteristics (mean, and ranges of flow and composition) of the water or wastewater source, treatment goals, regulatory requirements, and site topography, hydrology, soil, and climate characteristics. One of the most fundamental decisions relates to goals. Is the wetland simply to provide water treatment, or also to provide wildlife habitat and/or an aesthetic or recreational asset to the community? It should be noted that wildlife activity (e.g., waterfowl excreta) could elevate nutrient and fecal coliform levels both within a wetland and at the final discharge. Similarly, constituents in the waste stream, for example, pesticides in agricultural runoff or metals in industrial wastewaters or urban runoff, can accumulate in the wetland sediments and plants, and ultimately impact wetland fauna. In general, the safest region in a treatment wetland to encourage wildlife habitat is the mid-region, away from high contaminant loadings in the influent and some

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distance from the effluent, where the wastes cannot impair the quality of the discharged water. As part of an analysis, the investigator should assess the effectiveness and economic costs of alternative technologies, alternative sites, and alternative process trains. For municipal wastewater treatment, the treatment wetland typically provides secondary and/or advanced treatment. However, there may exist a more cost-effective combination of unit processes (with or without a wetland component) that provides for lower life-cycle costs.

DESIGN CONSIDERATIONS Treatment wetlands can be inexpensive to construct and simple to operate compared to many water treatment technologies. This has led to their widespread use during the past two decades. There are several approaches to designing wetlands for water treatment, and there are good references available to support these designs. Several fundamental questions should be addressed in the design process. Will a treatment wetland provide the desired effluent contaminant concentrations? What is the most reliable, cost-effective treatment wetland configuration? How much land will be required? The bulk of the costs of treatment wetlands, such as land purchase and berm construction, is related to area requirements. How robust is the system (i.e., what is the risk of compliance problems)? A good starting point for design is to access one of the available pollutant reduction models that has been verified against the North American Wetland Treatment System Database (NAWTSD). The NAWTSD, compiled with funding from U.S. Environmental Protection Agency, lists the performance of several hundred treatment wetlands. Several authors and agencies have provided overviews of the NAWTSD, as well as discussions of the benefits and limitations of these data for design purposes (Kadlec and Knight, 1996; U.S. Environmental Protection Agency, 1999). Both area-specific and volume-specific models are available for treatment wetland design purposes. Both types of models allow the designer to derive the acreage of wetland from the inflow pollutant concentration, desired outflow concentration, and hydraulic loading rates (Reed et al., 1995; Kadlec and Knight, 1996). Kadlec and Knight (1996) have reported a suite of pollutant settling rate coefficients for the areal model that is useful for initial design purposes. The areal model is as follows: ln((Ce – C*)/(Ci – C*)) = – (k/q) where Ce equals the target effluent concentration (mg/l). Ci equals the influent concentration (mg/l). C * equals the background concentration (mg/l). k equals the first-order areal rate constant, m/yr. q equals the hydraulic loading rate, m/yr. ©2001 CRC Press LLC

This equation can be rearranged in the form A = (0.0365Q/k) ln((Ci – C*)/(Ce – C*)) where A equals the wetland area in ha. Q equals the influent flow, m3/day. One key factor to note with respect to wetland design is that treatment wetlands are extraordinarily effective at removing or transforming some constituents but are quite inefficient at removing others. As the wetland area must be based on worstcase conditions, it is these latter constituents that will dictate the overall area requirement. If just one constituent in a waste stream is markedly expanding the wetland area requirement (and therefore cost), then as part of an alternative analysis the designer should assess techniques for accelerating removal of this contaminant, either in the wetland or in an upstream or downstream unit process. An equally important concept for design is that treatment wetlands exhibit a background constituent concentration below which the wetland is unlikely to provide treatment. This is the C* in the areal model described above. A designer can obtain some perspective for which contaminants are recalcitrant (difficult to remove) in a wetland environment, and what the likely minimum contaminant background levels are, by inspecting k and C* values for the areal model for FWS wetlands derived from the NAWTSD (Table 2). The values in Table 2 also provide an indication of the relative performance of FWS and SSF treatment wetlands. On a unit area basis, SSF systems are more effective than FWS at removing a number of constituents including BOD and nitrate–nitrogen. Table 2

Preliminary Model Parameters for FWS and SSF Wetlands Developed from the North American Wetlands Treatment System Database*

Parameter

k20 (m/yr) FWS SSF

BOD TSS Organic N NH4-N NOx-N TN TP

34 1000 17 18 35 22 12

180 1000 35 34 50 27 12

C * (mg/L) FWS

SSF

3.5 + 0.053Ci 5.1 + 0.16Ci 1.5 0 0 1.5 0.02

3.5 + 0.053Ci 7.8 + 0.063Ci 1.5 0 0 1.5 0.02

* The k20 represents the first-order areal rate constant at 20°C and C * represents the likely wetland background concentration. Ci represents the constituent inflow concentration. Higher k values correspond to higher mass contaminant removal rates. (From Kadlec and Knight, 1996.)

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Most predictive models used for sizing treatment wetlands are empirical and not intimately linked to actual contaminant removal processes. The k values depicted above, for example, are lumped parameters that simply represent the net result of all the internal processes in a wetland that contribute to the removal of a constituent. For phosphorus, the k value is low because burial of P in the sediment is thought to be the only long-term P removal process (Richardson and Craft, 1993). Phosphorus is actually quite dynamic in wetlands, owing to settling of influent particulate P, assimilation and short-term or long-term storage of ortho- or soluble reactive P by algae, microorganisms, and macrophytes, sorption of dissolved organic or organic P forms, and decomposition of the organic P forms. While some of the above processes can be rapid and sequester large amounts of P, slow burial is the ultimate removal mechanism. The k value for the areal model, therefore, integrates all of these processes. The areal model described above, or one of the volumetric models, can be used for preliminary sizing. At a more detailed design level, the engineer must account for site-specific factors as well as expected seasonal variations in performance in the final area requirement assessment (Kadlec and Knight, 1996). When is more site-specific information needed beyond published design models? First, intimate process knowledge becomes more critical for treating unusual waste streams, for challenging environmental conditions (e.g., cold climates), or for conditions under which stringent permit limits will be imposed. Further, if the wetland does not seem appropriate for the site (e.g., exceeds available area), more site-specific information may be required. The use of smaller wetland footprints beyond what is predicted by the design model may be realistic, but this may only be achieved by a better understanding of the contaminant removal processes with the particular wastewater and, perhaps, stricter operator monitoring and control once the system is operational. Outdoor mesocosm-scale studies can be performed to assess site-specific removal rates for the limiting contaminants. Hydraulic analyses can be performed to determine if the addition of more cells or internal berms can improve internal flow patterns and, therefore, water treatment performance.

CONSTRUCTION AND MANAGEMENT Construction Construction plans and specifications must be prepared once design parameters are established. While the prominent cost components of a treatment wetland are often related to earthmoving (land grubbing and grading and berm construction), allowance must be made for other features including access roads, water distribution (inlet and outlet), structures and plumbing, and rock or gravel placement (SSF wetlands). Erosion control measures also must be incorporated for the berms. A suitable soil substrate for planting must be provided for FWS wetlands, such as a loamy soil that the roots and rhizomes can readily penetrate. If initial vegetation planting is performed, consideration must be given to the type of propagules (e.g.,

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seeds, bare-root seedlings, or field-harvested plants), planting density (1 m spacing is often used), and appropriate elevations and planting depths. Fertilizer also may be required to initially establish the plant standing crop depending upon the water source to be treated. Care must be taken not to compact the soil matrix prior to and during planting. Control of water elevations during the wetland start-up period can help ensure the success of the planted vegetation and to discourage unwanted, weedy species. Construction costs, on an areal basis, are substantially greater for SSF wetlands than for FWS wetlands due to the purchase and transport of the rock or gravel matrix. In some cases, this is compensated for by the fact that the matrix reduces the area required to treat a given flow. The median cost of FWS wetlands in North America as of 1995 was US $44,600 per ha, and for subsurface flow wetlands, US $358,000 per ha (Kadlec and Knight, 1996). Management In general, the operational risks associated with using a wetland for water treatment purposes are probably no greater than those for more conventional treatment processes. As with any other water treatment technology, the wetland treatment system must be properly designed and operated. Moreover, in preparing and negotiating a permit for a treatment wetland, it is critical to establish compliance limits that account for the likely temporal variations in wetland performance. Operators of treatment wetlands have little short-term control over the diverse microenvironments that provide actual water treatment within the wetland. Poor effluent quality caused by a problem internal to the treatment wetland can be difficult to correct quickly. A phenomenon that has caused compliance problems in several FWS wetlands is die-off of emergent vegetation in the immediate vicinity of the effluent discharge. This creates an area of open water where phytoplankton can thrive. These unicellular algae can, in turn, be exported from the system, creating high effluent concentrations of TSS and associated constituents. Establishment of long-term permit limits (annual average or long-term running averages) for compliance will help minimize the risks associated with such unpredictable events. This is particularly critical for treatment wetlands used for nonpoint source treatment, such as for agricultural and urban runoff, where temporal variations in hydraulic and contaminant loading are pronounced and unpredictable. If short-term guidelines, such as daily or weekly permit levels, are required, they should be substantially more liberal than the longterm guidelines. Most wetlands are sampled on an inflow–outflow basis, for both water quantity and quality characteristics (Figure 6). For performance upset situations, it can be helpful to collect samples with distance from the inflow to outflow to determine whether an undesirable export situation is simply the passage of an inflow parameter through the system without treatment or whether the contaminant is being generated internal to the wetland. If it is being generated internally, then transect sampling can be a useful diagnostic tool to pinpoint the location and cause.

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Figure 6

Automatic recording device used for measuring water stage and flows in FWS wetlands.

Some management of treatment wetlands may be required to ensure that nuisance or public health problems are avoided (Kent and DeBusk, 1997). A common nuisance situation is odors. Suites of inorganic and organic sulfur-containing compounds are found in many types of wastewaters and can contribute to odors in treatment wetlands. A good rule of thumb for FWS systems is to reduce or eliminate anaerobic zones in the inflow region of treatment wetlands, for example, by pretreating high organic strength wastewaters. Food processing wastewaters, which are high in COD, have been successfully treated in FWS systems by incorporating step-feed influent structures to spread the high strength wastewaters over a greater area. Dense mats of floating macrophytes, such as duckweeds, have been shown to successfully reduce odors. Subsurface flow wetlands generally appear to be more successful at containing odors, although effluent structures (e.g., weirs) that cause agitation can disperse odors. Treatment wetlands can also be breeding grounds for mosquitoes which are both a nuisance and vectors for a host of human and animal diseases (Russell, 1999). If a treatment wetland is located in an area where such mosquito-borne diseases occur, it is critically important that its siting, design, and operation mitigate potential health risks. Situating the wetland away from human habitation (further than the typical flight line of mosquitoes) is important as is the incorporation of vegetation types and densities that facilitate biological (e.g., fish) and, if necessary, chemical control of the mosquito larvae.

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PERFORMANCE The NAWTSD, with approximately 200 wetland sites (Knight et al., 1993), provides a general basis for assessing contaminant removal performance of treatment wetlands. The database was developed from wetland treatment systems that collectively represent a wide range of contaminant types and loading and design characteristics. Therefore, it should be recognized that these data provide only a rough approximation of specific contaminant removal efficiency in treatment wetlands. There are many wetland configurations, dedicated to various water treatment goals, for which performance data exist. To clearly understand the potential contaminant removal performance of a treatment wetland, it is helpful to review longterm performance of existing systems, focusing on those contaminants that originally dictated wetland area requirements and contaminant loading rates. For example, FWS and SSF wetlands have been designed for BOD and TSS removal from domestic wastewaters (i.e., secondary treatment), as well as for N and P removal (i.e., advanced treatment). Although in both circumstances the wetland is used for domestic wastewater treatment, the design loading rates for the different contaminants will vary widely (e.g., the area requirement for P removal is markedly higher than that for BOD removal) as will the contaminant removal performance. A brief overview of treatment wetland performance for specific types of contaminants using different configurations (e.g., FWS, SSF) is presented below. Suspended Solids and Organic Carbon Removal Wetlands are capable of achieving a high efficiency for suspended solids (TSS) removal from the water column. Suspended matter in the water may contain a number of contaminants, such as nutrients, heavy metals, and organic compounds. These contaminants may themselves be in particulate form or they may be physically or chemically bound to the particulate matter. Thus, in cases where the bulk of the contaminant load is associated with particulate matter, physical settling of suspended solids can result in efficient removal of the contaminants from the water or wastewater stream. Among the FWS wetlands included in the NAWTSD, the average outflow concentration of TSS was 14 mg/l, compared with an average inflow concentration of 46 mg/l. This represents a mass removal rate of 7.0 kg/ha per day, or 68 percent removal on a percentage basis. For SSF systems, average inflow TSS levels of 48 mg/l were reduced to 10 mg/l, providing a mass removal rate of 35 kg/ha per day, 74 percent removal (Kadlec and Knight, 1996). Although wetlands generally provide effective removal of suspended solids, the removal efficiency decreases substantially when the concentration of TSS approaches the natural background level of about 10 mg/l. Most wetland treatment systems are overdesigned for TSS removal (having been designed for removal of other contaminants). Therefore, background concentrations of TSS are often attained a short distance downstream from the wetland inflow. Wetlands contain vast numbers of organic C-using microorganisms adapted to the aerobic (O2-rich) surface waters and anaerobic (O2-depleted) soils and, thus, are ©2001 CRC Press LLC

capable of highly effective removal of organic compounds from a variety of wastewaters. Organic matter contains approximately 45 to 50 percent carbon (C) which is used by these microorganisms as a source of energy. Many microorganisms consume oxygen (O2) to break down organic C to carbon dioxide (CO2) and methane (CH4), both of which are lost to the atmosphere. Therefore, the release of excessive amounts of organic C to surface waters can result in a significant depletion of O2, and the subsequent mortality of fish and other O2-dependent aquatic or marine organisms. Wetlands also store and recycle copious amounts of organic C contained in plants and animals, dead plant material (litter), microorganisms, and peat. Therefore, wetlands tend to be natural exporters of organic C as a result of decomposition of organic matter into fine particulate material and dissolved compounds. The more readily degradable organic C compounds typically found in domestic wastewaters are rapidly removed in wetlands. Biological removal of a variety of recalcitrant (not readily decomposed) organic C compounds, including lignin-based compounds and petroleum products, can also be achieved in wetlands, although removal rates may be substantially lower. A commonly used parameter for biologically available C is biochemical oxygen demand (BOD), which is actually a measure of the rate of O2 consumption by microorganisms using the available organic C in the water or soil. The normal procedure for determining BOD in water samples measures the amount of O2 depletion occurring over a 5-day period (BOD5). Chemical oxygen demand (COD) is a second common indicator of oxygen-demanding substances, particularly for industrial wastewaters and runoff streams. Wetlands generally provide substantial removal of BOD5, despite a naturally occurring background level of approximately 1 to 6 mg/l. The average outflow concentration of BOD5 for FWS wetlands in the NAWTSD was 8.0 mg/l, compared with an average inflow concentration of 30.3 mg/l, representing a mass removal rate of 5.1 kg/ha per day and 71 percent removal efficiency. For SSF systems, average inflow BOD5 levels of 27.5 were reduced to 8.6 mg/l, at a mass removal rate of 18.4 kg/ha per day and 63 percent removal efficiency (Kadlec and Knight, 1996). Removal of Organic Carbon and Suspended Solids from Waste Stabilization Pond Effluents Waste stabilization ponds (WSPs) are a domestic wastewater treatment technology common to both developed and developing countries. Treatment ponds differ in their characteristics based mainly on organic matter (BOD5) loading. Highly loaded (225 to 560 kg BOD5/ha per day) ponds usually are anaerobic, moderately loaded (55 to 200 kg BOD5/ha per day) ponds typically are facultative, and lightly loaded (65 to 135 kg BOD5/ha per day) ponds are often aerobic (Water Environment Federation, 1996). Aerobic conditions also are promoted in some WSPs by using floating, mechanical aerators. Ponds frequently are connected in sequence to improve hydraulics and enhance treatment performance of the overall WSP system. While providing effective pathogen and BOD removal, WSPs typically produce an effluent with high TSS concentrations, as well as a high (alkaline) daytime pH. The high TSS usually is not related to the influent TSS loading, but rather is caused by the high microalgae densities that develop in the ponds. Because microalgae are ©2001 CRC Press LLC

organic particles, the exported TSS can exert a long-term oxygen demand as well as release nutrients in receiving waters. The particulate matter also can clog equipment (e.g., fine emitters) used for irrigating agricultural crops with the pond effluent. During the daytime, high water column pH conditions created by microalgae photosynthesis in the WSP may result in effluent unionized ammonia (NH3-N) concentrations that are toxic to aquatic fauna in the receiving waters. Polishing of WSP effluents has become one of the prominent uses for treatment wetlands. In the early 1990s, WSPs were the most common type of front end or preliminary treatment of wastewaters prior to introduction into a FWS or SSF wetland (Reed and Brown, 1995). The original work on wetland treatment of WSP effluents dealt with simply allowing floating aquatic plants, typically water hyacinths, to grow over a portion of the effluent region of the ponds (Orth and Sapkota, 1988; Wolverton and McDonald, 1979). Later studies and commercial applications involved the use of small-leaved floating plants (duckweeds) for this same purpose. Results from outdoor tank studies with the floating plant duckweed demonstrate that while this small-leaved plant is slightly less effective than WSPs for BOD5 removal, it does provide an effluent sharply lower in TSS than the nonvegetated ponds (Table 3). Floating plants, therefore, can be effective, when deployed in the final cell (for WSPs in series) or effluent region (for a single pond) for reducing particle export from WSP systems. Table 3

Influent and Effluent BOD5 and TSS Concentrations*

Influent WSP effluent Duckweed pond effluent

BOD5 mg/l

TSS mg/l

193 84 99

59 133 36

* BOD5 and TSS concentrations in microcosm waste stabilization ponds (WSP) and floating plant-dominated (duckweed) ponds; Each system received primary domestic effluent at a hydraulic retention time of six days. Values represent means of weekly measurements from duplicate tanks over a seven month (November–June) period.

While partial floating macrophyte cover can improve WSP treatment, the periodic plant harvest requirement for large ponds often is undesirable. Free water surface and SSF wetlands now are more commonly used than floating plant systems for upgrading WSP effluents. Wetlands have proved effective for this purpose as long as realistic TSS loadings are used (typically 10 to 48 kg/ha per day) and adequate emergent vegetation cover is maintained, particularly in the effluent region of the wetland (Gearhart and Higley, 1993). Nitrogen Removal Nitrogen (N) is a major component of municipal wastewater, stormwater runoff from urban and agricultural lands, and wastewater from various types of industrial ©2001 CRC Press LLC

processes. Environmental and health problems associated with excessive amounts of N in the environment have been well documented. For example, high concentrations of nitrate in drinking water supplies can cause methemoglobinemia, or blue baby syndrome, in infants. Unionized ammonia (NH3), found in certain types of wastewater effluent, is potentially toxic to many aquatic and marine organisms. In addition, eutrophication of surface waters frequently is linked with elevated N concentrations, especially in coastal and estuarine environments. Nitrogen exists in many forms in the environment and transformations among different forms may occur rapidly and frequently. Municipal and industrial wastewater may contain significant amounts of both organic and inorganic forms of N. Inorganic N, which includes nitrate, nitrite, and ammonium, may also be present at high concentrations in agricultural and urban runoff. In the environment, nitrite and, especially, nitrate are usually found in well-aerated waters, while ammonium is the more persistent form of inorganic N in anaerobic wetland soils. Wetlands provide a relatively high level of N removal, although the natural background level of N in wetland outflows is typically greater than 1 mg/l owing primarily to decomposition and export of the native organic matter (Figure 7). The average outflow concentration of ammonium (expressed as N) in North American FWS wetlands was 2.2 mg/l, compared with an average inflow concentration of 4.9 mg/l. The average ammonium removal on a mass and percentage basis was 0.35 kg/ha-day and 38 percent, respectively. Mean inflow and outflow values for ammonium for SSF wetlands were 6.0 and 4.5 mg/l, with mass and percentage removal rates of 0.62 kg/ha per day and 9 percent, respectively (Kadlec and Knight, 1996). Biological conversion of ammonium to nitrate (nitrification) accounts for much of the ammonium removal in wetlands. The use of higher strength wastewaters and the prevalence of anaerobic conditions in the subsurface bed explain the generally poorer ammonium removal performance of SSF wetlands.

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3

Figure 7

Profiles of nitrogen species (ammonium, nitrate, and organic nitrogen) with distance through a newly flooded forested treatment wetland. Prior to rehydration, the wetland had been hydrologically altered (dry) for more than 10 years. This caused a short-term export of organic N and ammonium.

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Nitrate removal efficiency is generally high in wetlands. The average reduction in nitrate (including nitrite) concentrations for the NAWTSD was 5.6 to 2.2 mg/l for FWS wetlands and 4.4 to 1.6 mg/l for SSF wetlands. Mass nitrate plus nitrite removal rates (and percentage removal efficiencies) for these respective systems were 0.40 kg/ha per day (51 percent) and 1.89 kg/ha per day (61 percent) (Kadlec and Knight, 1996). The biological process of denitrification, that is, conversion of nitrate to nitrogen gas, provides a means for complete removal of inorganic N from wetlands as opposed to storage within the vegetation or soil. Accordingly, the removal efficiency for nitrate is frequently very high for wetlands with high inflow concentrations. The average reduction in total N (inorganic plus organic forms) concentrations in North American wetlands was 9.03 to 4.27 mg/l for FWS wetlands and 18.92 to 8.41 mg/l for SSF wetlands. Mean total N mass removal rates averaged 1.06 and 5.85 kg/ha per day for FWS and SSF wetlands, respectively (Kadlec and Knight, 1996). Note that the higher mass removal rates for SSF wetlands than FWS systems for N, BOD5 , and TSS is related principally to their higher mass loading rates (typically four to six times higher than FWS wetlands) on a unit area basis. Nitrogen Removal from Food Processing Wastewaters Recently, investigators have begun sequencing wetland unit processes to enhance overall system nitrogen removal performance. A pilot-scale combined wetland system in the state of Washington was tested in 1995 for its effectiveness in treating potato processing wastewaters (Kadlec et al., 1997). Potato processing produces a high organic strength wastewater, with a chemical oxygen demand (COD) of approximately 3000 mg/l. The goal of the pilot-scale wetland treatment system was to produce an effluent suitable for land irrigation. Nitrogen reduction by the wetland was a particularly difficult challenge because TN levels needed to be reduced from above 150 mg/l to less than 50 mg/l. Project scientists selected a combination system comprised of two horizontal FWS wetlands (HSF1 and HSF2), followed by a downward, vertical flow SSF wetland (VFW3), and then a final horizontal flow FWS wetland (HSF4) to test. Wastewater flowed by gravity through the two initial horizontal flow wetlands which were designed to provide organic matter (COD) and TSS removal. Effluent from HSF2 was then spray-irrigated onto the surface of the vertical flow wetland (HSF3), designed for nitrification. Effluent from VFW3 was then gravity-fed to HSF4, where the principal goal was denitrification of the nitrate produced in VFW3. Performance data for the pilot-scale combined system demonstrated dramatic changes in effluent quality in the various system unit processes (Kadlec et al., 1997). The initial wetlands (HSF1, HSF2) provided good COD and TSS removal as well as mineralization of organic N to NH4-N (Table 4). In the vertical flow wetland, the bulk of the NH4-N was nitrified to NO3-N. Some denitrification also appeared to occur in this wetland. The final wetland (HSF4), in turn, removed much of the remaining nitrate by denitrification. Many wastewater constituents, such as N, require sequential exposure to different environments in order to be transformed and removed from the water. The ©2001 CRC Press LLC

Table 4

Summer Performance (mg/l) of a Pilot-Scale Combined Wetland for Treating Potato Processing Wastewaters*

COD TSS Organic N NH4-N NO3-N

Influent

HSF1

HSF2

2986 607 91 73 1

1056 85 10 129 1

601 72 3 116 1

Effluent VFW3 209 48 13 26 43

HSF4 161 37 12 29 13

* The initial two horizontal flow wetlands, HSF1 and HSF2, were designed for chemical oxygen demand (COD) and suspended solids (TSS) reduction. The third wetland, VFW3, a vertical flow wetland, was designed to enhance nitrification. The final wetland, HSF4, was designed for denitrification. From Kadlec et al. (1997). With permission.

implementation of combined systems that link separate unit processes in the appropriate sequence to effect contaminant removal is a useful strategy for removing complex compounds. Data from the pilot-study with potato processing wastewater were used to design a full-scale system that was recently placed into operation. Phosphorus Removal Phosphorus (P), like N, is a major plant nutrient. Hence, addition of P to the environment often contributes to eutrophication of lakes and coastal waters. In many cases, wetlands do not provide the high level of efficient long-term removal for P that they provide for N. This is, in part, due to the lack of a gaseous sink, analogous to denitrification, for P removal. Nevertheless, most wetlands can provide significant P removal from water and wastewater through a combination of physical, chemical, and biological processes. Orthophosphate (HPO42– and H2PO4– ) is the predominant inorganic form of P in surface waters. This form of P readily accumulates in wetland vegetation and soils as a result of biological uptake and chemical bonding. Based on data from the NAWTSD, the average orthophosphate concentration reduction was 1.75 to 1.11 mg/l for FWS wetlands with mass and percentage removal rates of 0.12 kg/ha per day and 41 percent (Kadlec and Knight, 1996). The NAWTSD does not provide orthophosphate data for SSF wetlands. Organic forms of P are much less biologically and chemically reactive in wetlands than orthophosphate. Particulate organic P may be removed by settling from the water column. Both dissolved and particulate organic P ultimately may be biologically broken down to inorganic P (mineralization) and subsequently removed through biological and chemical processes. The average total P concentration reduction for FWS wetlands was 3.78 to 1.62 mg/l, and that for SSF wetlands, 4.41 to 2.97 mg/l. Respective mass (and percentage) removal rates for FWS and SSF systems were 0.17 kg/ha per day (34 percent) and 1.14 kg/ha per day (22 percent) (Kadlec and Knight, 1996).

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It is notable that of the contaminants discussed above, P is the element that has the lowest background concentration in treatment wetlands. Several large emergent marshes in Florida have provided long-term effluent P concentrations below 0.05 mg/l (Jackson, 1989; Moustafa et al., 1999). Removal of Trace Metals, Toxic Organic Compounds, and Complex Mixtures of Contaminants A number of metals are required in trace amounts for plant or animal growth. Some of these micronutrients such as copper, selenium, and zinc, are toxic at higher concentrations and may be found in certain types of wastewaters. Other metals, such as cadmium, mercury, and lead, found in industrial or other types of wastewater, have no known biological benefit and are toxic even at relatively low concentrations. Furthermore, certain metals have a tendency to become concentrated at higher levels of the food chain. This biomagnification effect can lead to serious health hazards to higher consumers, including humans. Removal of metals in wetlands may occur through a number of processes including plant uptake, soil adsorption, and precipitation. Plant uptake rates and tolerance of metals vary considerably among plant species. Some upland plant species have demonstrated the ability to store high concentrations of metals in roots and other tissues. Metals may also tend to accumulate on the root surfaces of plants, rather than being absorbed into the plant. Wetlands can be effective sinks for metals due to the relative immobility of most metals in wetland soils. A number of metals, including cadmium, copper, nickel, lead, and zinc, form nearly insoluble compounds with sulfides under anaerobic conditions in wetland soils. In addition, some metals, such as chromium, copper, lead, and zinc, form strong chemical complexes (chemisorption) with organic matter in the soil or water. Metals, for example chromium and copper, also may be bound through sorption to clays and oxides of manganese, aluminum, and iron. Nickel also binds with organic matter and iron/manganese but may become mobilized under certain conditions. In addition to heavy metals, there are multitudes of degradation-resistant and toxic natural and man-made organic compounds that may be present in wastewater effluent or runoff. Quantitative data on removal of organics, such as pesticides and petrochemicals, are limited. However, measurable removal of a wide variety of organic compounds has been documented for wetland treatment systems. Both mineral and organic soils may adsorb organic compounds via chemisorption (strong interaction) or physical adsorption (weak interaction). Microbes are capable of degrading most classes of organic pollutants, but the rate of degradation varies considerably, depending on chemical and structural properties of the organic compound, and the chemical and physical environment in the soil. For example, highly halogenated hydrocarbons such as polychlorinated biphenyls (PCBs) are extremely resistant to decomposition due to their low solubility in water and the lack of a structural site for enzyme attachment for degradation. Other possible mechanisms for removal of organics in wetlands are volatilization and photochemical (sunlight) degradation. ©2001 CRC Press LLC

Studies have documented successful wetland treatment of PCBs, lindane, pentachlorophenol, and atrazine. In most cases, actual removal processes, for example sediment retention or microbial degradation, were not determined. However, there is substantial evidence that a number of toxic organics, such as pentachlorophenol, break down readily under the alternating aerobic and anaerobic conditions found in wetland soils. Wetland Treatment of Urban Runoff There currently exists considerable interest in using wetlands for treating urban runoff which can contain a mixture of suspended solids, oxygen-demanding substances, heavy metals, hydrocarbons, and other potentially toxic organics. Treatment of urban runoff is particularly challenging because factors such as intervals between storm events can influence the contaminant loading to the wetland, and the volume (size) of the rainfall event can control the hydraulic residence time of contaminants in the treatment wetland. In short, wetlands used for urban runoff treatment are exposed to very dynamic hydraulic and contaminant loadings, and this requires the deployment of unique designs. Several features are common to wetlands used for treating urban runoff (Livingston, 1989; Schueler, 1996). Deep forebays (e.g., 1 m deep) focus sediment accumulation at the inflow points to the wetland. Energy dissipation devices reduce velocities that could scour sediments or vegetation. Variable water depths encourage a diversity of vegetation, and a permanent water pool helps to maintain wetland characteristics during droughts. Also, most runoff treatment wetlands have a defined sediment cleanout schedule (usually every 5 to 10 years) and a specific wetland-towatershed area ratio (often 0.01 to 0.02). Performance of wetlands for urban runoff treatment is often difficult to quantify. Inflow–outflow sampling must be performed on a storm-event basis, and much of the inflow water (and associated contaminants) can be lost to vertical seepage. As a consequence, contaminant loadings to groundwaters often are ignored in documenting performance of stormwater detention facilities. Because of the strong affinity many metals have for particles, the sediments and vegetation of several urban runoff detention facilities have been evaluated for heavy metal retention (Shutes et al., 1997). A marsh system used for urban runoff detention in California provided the following heavy metal removal efficiencies: lead (83 percent), chromium (53 percent), zinc (51 percent), copper (32 percent), and nickel (12 percent). Analysis of ecosystem compartments in the marsh demonstrated highest metal concentrations in the surficial sediments followed by plant roots (Meiorin, 1989). This accumulation of metals in the sediments is common for urban runoff detention facilities (Shutes et al., 1997). Wetland Treatment of Landfill Leachates Like urban runoff, landfill leachates contain a mixture of contaminants whose concentrations can vary markedly over time. Landfill leachate composition is controlled by the type of solid waste disposed of in the landfill, the amount of leachate ©2001 CRC Press LLC

dilution from rainfall, and the age of the landfill. Leachate composition can change over time, with the fresh leachate often anoxic with a high COD, and the more aged leachate containing a reduced oxygen demand but higher concentrations of heavy metals and potentially toxic organics. Constructed wetlands have been scrutinized as a technology for treating landfill leachates for only the past decade (Kadlec, 1998), with most of the existing treatment wetland performance data pertaining to oxygen demanding substances and nutrients (DeBusk, 1998). Mass balances of leachate-borne heavy metals have been obtained but only at the mesocosm scale. One 14 month study investigated metal removal from landfill leachate in wetlands dominated by either floating (Lemna minor) or emergent macrophytes (Typha latifolia) (DeBusk et al., 1996b). Landfill leachate amended with lead (Pb) and cadmium (Cd) to levels of 100 and 400 µg/L, respectively, was fed continuously to outdoor mesocosm tanks situated adjacent to a landfill. Despite the high hydraulic loading rate (14.6 cm/day) and short hydraulic retention times (2.4 days), both wetland types removed approximately 50 percent of the influent metals (Table 5). On a mass basis, the surficial sediments were the principal site of metal sequestration, containing over 2000 mg/m2 of both Pb and Cd. Nearly all sediment Pb and Cd were present as metal sulfides which suggests that these metals are neither bioavailable nor toxic to biota (Di Toro, 1992). Cattail roots were the second most prominent site of metal immobilization, containing between 250 to 400 mg/m2 of Pb and Cd. No overt toxic affects to flora or fauna (fish) were observed in this investigation. Table 5

Influent and Effluent Lead and Cadmium Concentrations (µg/L) for Landfill Leachate*

Influent Cattail effluent Duckweed effluent

Lead

Cadmium

396 196 219

105 52 52

* Lead and cadmium concentrations fed to microcosm wetlands dominated by cattail (Typha domingensis ) and duckweed (Lemna minor ). Values represent weekly means from duplicate microcosms that received a continuous flow of spiked leachate for 14 months.

Pilot-scale wetlands dominated by Typha and Lemna were evaluated for their effectiveness at removing hydrocarbons from industrial wastewater (Salmon et al., 1998). Wetlands containing these two macrophytes performed comparably, reducing total hydrocarbon (THC) levels from 60 mg/l to below 8 mg/l at a 1.5 day hydraulic retention time. A mass balance showed that of the 1431 g THC/m2 loaded to the wetlands over a 1 year period, around 25 percent were volatilized and 63 percent were degraded by microbial activity. Only a small fraction (4 percent) was recovered on and in the wetland sediments. Treatment wetlands can effectively remove many metals and synthetic organic compounds from wastewaters. Because there is little performance information ©2001 CRC Press LLC

available upon which to design these systems, pilot-scale studies typically should be conducted prior to implementation of a full-scale system. The pilot should be used to develop design information as well as to obtain a clear understanding of the fate and bioavailability of potentially toxic compounds both within the unit processes and in the system effluent. Pathogen Removal Pathogens in wastewaters and surface waters that are fed into treatment wetlands include bacteria, viruses, protozoa, and helminths (parasitic worms). Because treatment wetlands were first studied and deployed in industrialized countries of Europe and North America where post-treatment disinfection is common, the removal of pathogens internal to wetlands was not of critical concern. Indeed, the biggest consideration was the occasional increase in pathogens observed as wastewater was passed through FWS wetlands caused typically by avian activity. More recently, with interest in using wetlands globally for wastewater treatment purposes, the pathogen removal performance of wetlands is being scrutinized. Most enteric bacteria and viruses are not thought to survive long outside the host organisms, so natural die-off is one process by which these organisms are removed in treatment wetlands. Other factors that contribute to bacteria and virus removal in natural systems include competition, predation, sedimentation, filtration, adsorption, pH extremes, and photolysis. Routine microbiological assessments of surface waters use groups of organisms such as total and fecal coliforms as indicators of fecal contamination. From an international public health perspective, multiple cell waste stabilization ponds are recommended as a cost-effective domestic wastewater treatment technology. The WSP water column conditions (high UV and pH) are effective for bacteria and virus destruction, and a long hydraulic retention time (HRT) contributes to helminth ova removal by sedimentation (Bartone and Arlosoroff, 1987). Welldesigned and operated WSPs usually meet the microbiological standards of less than 1000 fecal coliforms/100 ml recommended for the agricultural reuse of domestic effluents (World Health Organization, 1989). Studies with various treatment wetland configurations have demonstrated a 1-log (90 percent) to 2-log (99 percent) reduction for indicator bacteria, depending on the system HRT. Studies with pilot-scale SSF wetlands planted with Phragmites demonstrated 1.1- to 1.9-log removals of Escherichia coli and total coliforms at HRTs in as short as 6 hours (Green et al., 1997). These removals increased to as high as 3.1 log after 48-hours detention in the gravel bed. Plant root exudates, or the microorganisms associated with the plant rhizosphere, may also contribute to pathogen destruction. Gravel filled SSF wetlands containing bulrush (Scirpus) achieved greater total coliform reduction (from 6.7 × 107/100 ml to 5.77 × 105/100 ml) than a nonvegetated gravel bed (Gersberg et al., 1989). Studies with indicator organisms reveal that the bacterial and viral reduction provided by treatment wetlands is slightly superior to conventional wastewater treatment processes such as activated sludge. The latter typically provides a 1-log reduction. Tests in a Typha-dominated FWS wetland that received secondary effluent ©2001 CRC Press LLC

at a nominal 3.3 day HRT demonstrated 90 to 99 percent reductions of both bacteria and virus indicators (MS-2 bacteriophage, Gersberg et al., 1989). These investigators cautioned, however, that post disinfection would be needed to satisfy local requirements (1000 TC/100 ml) for discharge. While wetlands can be an effective technology for pathogen removal, there remain some concerns for using treatment wetlands, particularly FWS configurations, in tropical regions. Many human parasites have complex life cycles that rely on intermediate animal hosts. Wetland environments, unlike WSPs, can support many of these avian and mollusk hosts and, therefore, provide an environment for proliferation rather than the destruction of the pathogenic organisms.

CONCLUSIONS Treatment wetlands offer great promise for the low-cost treatment of contaminated waters and wastewaters. Successful wetland treatment system designs require careful consideration of the actual contaminant removal processes and how the rates and effectiveness of these processes may change over time as the system ages. A sound understanding of treatment goals and contaminant removal processes will lead to the most appropriate and cost-effective unit process (e.g., wetland, waste stabilization pond, conventional WWTP) or sequence of unit processes.

REFERENCES Bartone, C. R. and Arlosoroff, S., Irrigation reuse of pond effluents in developing countries, Water Sci. Technol., 19, 289, 1987. Boutin, C., Lienard, A., and Esser, D., Development of a new generation of reed-bed filters in France: first results, Water Sci. Technol., 35, 315, 1997. Brix, H., Macrophyte-mediated oxygen transfer in wetlands: transport mechanisms and rates, in Constructed Wetlands for Water Quality Improvement, Moshiri, G. A., Ed., Lewis Publishers, Boca Raton, FL, 1993. Brix, H., Functions of macrophytes in constructed wetlands, Water Sci. Technol., 29, 71, 1994. Brix, H., Do macrophytes play a role in treatment wetlands? Water Sci. Technol., 35, 11, 1997. Burgoon, P. S., Reddy, K. R., and DeBusk, T. A., Performance of subsurface flow wetlands with batch-load and continuous-flow conditions, Water Environ. Res., 67, 855, 1995. Clough, K. S., DeBusk, T. A., and Reddy, K. R., Model water hyacinth and pennywort systems for secondary treatment of domestic wastewater, in Aquatic Plants for Water Treatment and Resource Recovery, Reddy, K. R. and Smith, W. H., Eds., Magnolia Publishing, Orlando, FL, 1987. Cole, S., The emergence of treatment wetlands, Environ. Sci. Technol., 33, 1998. Cooper, P., Smith, M., and Maynard, H., The design and performance of a nitrifying verticalflow reed bed treatment system, Water Sci. Technol., 35, 215, 1997. DeBusk, T. A., Langston, M. A., Burgoon, P. S., and Reddy, K. R., A performance comparison of vegetated submerged beds and floating macrophyte systems for wastewater treatment, in Constructed Wetlands in Water Pollution Control, Cooper, P. F. and Findlater, B. C., Eds., Pergamon Press, Oxford, England, 1990.

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DeBusk, T. A., Peterson, J. E., and Reddy, K. R., Use of aquatic and terrestrial plants for removing phosphorus from dairy wastewaters, Ecol. Eng., 5, 371, 1996a. DeBusk, T. A., Laughlin, R. B., and Schwartz, L. N., Retention and compartmentalization of lead and cadmium in wetland microcosms, Water Res., 30, 2707, 1996b. DeBusk, W. F., Evaluation of a constructed wetland for treatment of leachate at a municipal landfill in northwest Florida, in Constructed Wetlands for the Treatment of Landfill Leachates, Mulamoottil, G., McBean, E., and Rovers, F. A., Eds., Lewis Publishers, Boca Raton, FL, 1998. Di Toro, D. M., Mahony, J. D., Hansen, D. J., Scott, K. J., Carlson, A. R., and Ankley, G. T., Acid volatile sulfide predicts the acute toxicity of cadmium and nickel in sediments, Environ. Sci. Technol., 26, 96, 1992. Dornbush, J. N., Constructed wastewater wetlands: the answer in South Dakota’s challenging environment, in Constructed Wetlands for Water Quality Improvement, Moshiri, G. A., Ed., Lewis Publishers, Boca Raton, FL, 1993. Faulkner, S. P. and Richardson, C. J., Physical and chemical characteristics of freshwater wetland soils, in Constructed Wetlands for Wastewater Treatment: Municipal, Industrial and Agricultural, Hammer, D. A., Ed., Lewis Publishers, Boca Raton, FL, 1989. Gearhart, R. A. and Higley, M., Constructed open surface wetlands: the water quality benefits and wildlife benefits, in Constructed Wetlands for Water Quality Improvement, Moshiri, G. A., Ed., Lewis Publishers, Boca Raton, FL, 1993. Gersberg, R. M., Gearhart, R. A., and Ives, M., Pathogen removal in constructed wetlands, in Constructed Wetlands for Wastewater Treatment: Municipal, Industrial and Agricultural, Hammer, D. A., Ed., Lewis Publishers, Boca Raton, FL, 1989. Gersberg, R. M., Role of aquatic plants in wastewater treatment by artificial wetlands, Water Res., 20, 363, 1985. Green, M. B., Griffen, P., Seabridge, J. K., and Dhobie, D., Removal of bacteria in subsurface flow wetlands, Water Sci. Technol., 35, 109, 1997. Hagedorn, C., McCoy, E. L., and Rahe, T. M., The potential for ground water contamination from septic effluents, J. Environ. Qual., 10, 1, 1981. House, C. H., Bergmann, B. A., Stomp, A. M., and Frederick, D. J., Combining constructed wetlands and aquatic and soil filters for reclamation and reuse of water, Ecol. Eng., 12, 27, 1999. Jackson, J., Man-made wetlands for wastewater treatment: two case studies, in Constructed Wetlands for Wastewater Treatment: Municipal, Industrial and Agricultural, Hammer, D. A., Ed., Lewis Publishers, Boca Raton, FL, 1989. Kadlec, R. H. and Knight, R. L., Treatment Wetlands, Lewis Publishers, Boca Raton, FL, 1996. Kadlec, R. H., Burgoon, P. S., and Henderson, M. E., Integrated natural systems for treating potato processing wastewater, Water Sci. Technol., 35, 263, 1997. Kadlec, R. H., Overland flow in wetlands: vegetation resistance, J. Hydraulic Eng., 116(5), 691, 1990. Kadlec, R. H., Deterministic and stochastic aspects of constructed wetlands performance and design, Water Sci. Technol., 35, 149, 1997. Kadlec, R. H., Constructed wetlands for treating landfill leachate, in Constructed Wetlands for the Treatment of Landfill Leachates, Mulamoottil, E., McBean, G., and Rovers, F. A., Eds., Lewis Publishers, Boca Raton, FL, 1998. Kent and DeBusk, Managing treatment wetlands, Land and Water, November/December, p. 52, 1997. King, A. C., Mitchell, C. A., and Howes, T., Hydraulic tracer studies in a pilot scale subsurface flow constructed wetland, Water Sci. Technol., 35, 189, 1997.

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Koch, M. S., Soil and Surface Water Nutrients in the Everglades Nutrient Removal Project, Technical Publication #9104. South Florida Water Management District, West Palm Beach, FL, 1991. Knight, R. L., Ruble, R. W., Kadlec, R. H., and Reed, S., Wetlands for wastewater treatment: Performance database, in Constructed Wetlands for Water Quality Improvement, Moshiri, G. A., Ed., Lewis Publishers, Boca Raton, FL, 1993. Livingston, E. H., Use of wetlands for urban stormwater management, in Constructed Wetlands for Wastewater Treatment: Municipal, Industrial and Agricultural, Hammer, D. A., Ed., Lewis Publishers, Boca Raton, FL, 1989. Meiorin, E. C., Urban runoff treatment in a fresh/brackish water marsh in Fremont, California, in Constructed Wetlands for Wastewater Treatment: Municipal, Industrial and Agricultural, Hammer, D. A., Ed., Lewis Publishers, Boca Raton, FL, 1989. Mitchell, D. S., Breen, P. F., and Chick, A. J., Artificial wetlands for treating wastewaters from single households and small communities, in Constructed Wetlands in Water Pollution Control, Cooper, P. F. and Findlater, B. C., Eds., Pergamon Press, Oxford, England, 1990. Mitsch, W. J., Wu, X., Nairn, R. W., Weihe, P. E., Wang, N., Deal, R., and Boucher, C. E., Creating and restoring wetlands—a whole-ecosystem experiment in self-design, BioScience, 48, 1019, 1998. Moustafa, M. Z., Newman, S., Fontaine, T. D., Chimney, M. J., and Kosier, T. C., Phosphorus retention by the Everglades Nutrient Removal Project: An everglades stormwater treatment area, in Phosphorus Biogeochemistry in Subtropical Ecosystems, Reddy, K. R., O’Conner, G. A., and Schelske, C. L., Eds., Lewis Publishers, Boca Raton, FL, 1999. Orth, H. M. and Sapkota, D. P., Upgrading a facultative pond by implanting water hyacinth, Water Res., 22, 1503, 1988. Reddy, K. R. and DeBusk, W. F., Nutrient removal potential of selected aquatic macrophytes, J. Environ. Qual., 14, 459, 1985. Reddy, K. R. and D’Angelo, E. M., Biogeochemical indicators to evaluate pollutant removal efficiency in constructed wetlands, Water Sci. Technol., 35, 1, 1997. Reddy, K. R., White, J. R., Wright, A., and Chua, T., Influence of phosphorus loading on microbial processes in the soil and water column of wetlands, in Phosphorus Biogeochemistry in Subtropical Ecosystems, Reddy, K. R., O’Conner, G. A., and Schelske, C. L., Eds., Lewis Publishers, Boca Raton, FL, 1999. Reed, S. C. and Brown, D., Subsurface flow wetlands—a performance evaluation, Water Environ. Res., 67, 244, 1995. Richardson, C. J. and Craft, C. B., Efficient phosphorus retention in wetlands: fact or fiction? in Constructed Wetlands for Water Quality Improvement, Moshiri, G. A., Ed., Lewis Publishers, Boca Raton, FL, 1993. Russell, R. C., Constructed wetlands and mosquitoes: health hazards and management options—an Australian perspective, Ecol. Eng., 12, 107, 1999. Salmon, C., Crabos, J. L., Sambuco, J. P., Bessiere, J. M., Basseres, A., Caumette, P., and Baccou, J.C., Artificial wetland performances in the purification efficiency of hydrocarbon wastewater, Water, Air, Soil Pollut., 104, 313, 1998. Schueler, T., Unpublished data, presented at Surface-Flow Constructed Treatment Wetland Technology Assessment Workshop, sponsored by U.S. Environmental Protection Agency, U.S. Bureau Mines, and City of Phoenix, 1996. Schwartz, L. N., Wallace, P. M., Gale, P. M., Smith, W. F., Wittig, J. T., and McCarty, S. L., Orange County, Florida Eastern Service Area reclaimed water wetlands reuse system, Water Sci. Technol., 29, 273, 1994.

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Shutes, R. B. E., Revitt, D. M., Mungur, A. S., and Scholes, L. N. L., The design of wetland systems for the treatment of urban runoff, Water Sci. Technol., 35, 19, 1997. Tanner, C. C., Plants for constructed wetland treatment systems—a comparison of the growth and nutrient uptake of eight emergent species, Ecol. Eng., 7, 59, 1996. U.S. Environmental Protection Agency, Water Quality Standards for Wetlands: National Guidance, U.S. Environmental Protection Agency, Office of Water Regulations and Standards, EPA 440/S-90–011, Washington, D.C., 1990. U.S. Environmental Protection Agency, Free Water Surface Wetlands for Wastewater Treatment: A Technology Assessment, Draft Document, March 1999. Wallace, P. M., Ecology of Created Wetland Communities in the Orange County Eastern Service Area Reclaimed Water Wetlands System, Fourth Operational Year Monitoring Report, March 1991–August 1992, Ecosystem Research Corporation, Gainesville, FL, 1992. Water Environment Federation, Operation of Municipal Wastewater Treatment Plants, Manual of Practice, Vol. II, Water Environment Federation, Alexandria, VA, 1996. World Health Organization, Health Guidelines for the Use of Wastewater in Agriculture and Aquaculture, World Health Organization Technical Report Series No. 778, WHO, Geneva, Switzerland, 1989. Wolverton, B. C. and McDonald, R. C., Upgrading facultative wastewater lagoons with vascular plants, J. Water Pollut. Contr. Fed., 51, 305, 1979.

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Kent, Donald M. “Design and Management of Wetlands for Wildlife” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

10

Design and Management of Wetlands for Wildlife Donald M. Kent

CONTENTS Design Size Relationship to Other Wetlands Disturbance Design Guidelines Management Management Approaches Management Techniques Vegetation Management Burning Grazing Herbicide Application Mechanical Management Blasting Bulldozing, Draglining, and Dredging Crushing Cutting Disking Propagation Water-Level Manipulation Artificial Nesting and Loafing Sites Fisheries References

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Wildlife management had been concerned primarily with the administration and regulation of waterfowl and furbearer harvests prior to the 1930s. It was about this time that wildlife managers, as well as the public, recognized that wildlife resources were not limitless. Leopold crystallized this emerging perspective in his book Game Management (1933) that gave birth to the scientific management of wildlife populations and wildlife habitats. Wetlands are especially critical habitats for wildlife and exceed all other land types in wildlife productivity (Vaught and Bowmaster, 1983; Cowardin and Goforth, 1985; Payne, 1992). Wildlife species use wetlands on either a permanent or transitory basis for breeding, food, and shelter (Pandit and Fotedar, 1982; Rakstad and Probst, 1985). In the United States, wetlands provide critical habitat for 80 of 276 threatened and endangered species. Approximately 64 percent of the wildlife in the Great Lakes region of the United States inhabit or are attracted to wetlands, including 62 percent of the birds, 69 percent of the mammals, and 71 percent of the amphibians and reptiles (Rakstad and Probst, 1985). From 67 to 90 percent of commercial fish and shellfish species are either directly or indirectly dependent upon wetlands (Peters et al., 1979; Vaught and Bowmaster, 1983; Radtke, 1985). Wetlands are also the principal habitat for furbearers and waterfowl. Approximately 10 to 12 million ducks breed in the contiguous United States and 45 million ducks depend on wetlands throughout the United States and Canada for their existence (Vaught and Bowmaster, 1983; Radtke, 1985). Wetland wildlife has a quantifiable economic value. Hundreds of millions of dollars are spent annually on birdwatching and other wildlife observations. Freshwater fisherman spent $7.8 billion dollars in 1980 and waterfowl hunters spent $950 million in 1975 (Radtke, 1985). In 1975–1976, more than 8.5 million furbearer pelts with a value in excess of $35.5 million were harvested (Chabreck, 1979). Valuable wetland habitats are being lost and degraded at an alarming rate; more than 200,000 ha of wetlands are lost per year (Low in Payne, 1992). Annual losses to agriculture range from 1 to 4 percent (Weller, 1981). Prairie potholes in the United States and Canada are lost at a rate of 1 to 2 percent per year, and 75 percent of northern central United States wetlands were lost between 1850 and 1977 (U.S. Department of Agriculture, 1980; Radtke, 1985; Melinchuk and Mackay, 1986). Bottomland hardwood forests were cleared at a rate of 66,800 ha per year between 1940 and 1980, reducing forested wetlands in some states by 96 percent (Korte and Fredrickson, 1977; Radtke, 1985). Coastal wetlands have also suffered dramatic losses, with more than 10,000 ha per year being lost from Gulf Coast wetlands (Chabreck, 1976; Gagliano, 1981). Many of the wetlands that remain are degraded from channelization, damming, and agricultural and urban surface runoff. As well, these remaining wetlands are typically fragmented or isolated and occur on private land. Coincident with the loss and degradation of wetlands is a decline in continental waterfowl populations. Breeding mallard populations have declined at a rate of up to 19 percent since 1970 (Melinchuk and MacKay, 1986). Weller (1981) estimates that 90 million waterfowl nests were lost in the north central United States between 1850 and 1977, and wintering waterfowl populations declined by 70 percent between

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the mid-1950s and 1983 (Whitman and Meredith, 1987). The effect of wetland loss and degradation on other wildlife remains largely undetermined. The exceptional value of wetlands as wildlife habitat, and the continued loss and degradation of wetlands, necessitate the careful design of new habitats and the management of existing habitat. The majority of high quality wetland habitats, those uninfluenced by extrinsic disturbances, are already preserved in parks and refuges. Opportunities for designing new habitats are few. Many other wetland habitats offer less than optimal habitat. For the latter, application of management techniques can increase productivity.

DESIGN A wetland designed for wildlife is the combination of details and features that, when implemented, results in the provision of habitat for wildlife that use wetlands to satisfy all or part of their life requisites. The design should be a preliminary sketch or plan for work to be executed, conceived in the mind, and fashioned skillfully. In practice, designs for wetlands can take several forms. The simplest and earliest efforts at designing wetlands for wildlife were characterized by the preservation of wildlife habitat. The most prominent effort among these in the United States was the establishment of the National Wildlife Refuge system, which protects uplands as well as wetlands. Florida's Pelican Island was the first refuge, established in 1903 by President Theodore Roosevelt to protect egrets, herons, and other birds that were being killed for their feathers. There are presently over 450 National Wildlife Refuges, comprising a network that encompasses over 90 million acres of lands and waters. Southern bayous, bottomland hardwood forests, swamps, prairie potholes, estuaries, and coastal wetlands are represented. Preservation of wetland wildlife habitat continues, although at a slower pace, through the efforts of government initiatives and private organizations. The restoration and enhancement of historic wetlands and the creation of new wetlands characterize more complex approaches to the design of wetlands for wildlife. Illustrative of large-scale restoration efforts in the United States, the 1985 Food Security Act has provided for wetland restoration on Farmers Home Administration and Conservation Reserve Program lands. Almost 55,000 acres of agricultural lands were restored to wetlands between 1987 and 1989, and another 90,000 acres were targeted for restoration in 1990 and 1991 (Mitchell in Kusler and Kentula, 1990). The North American Wetlands Conservation Act enacted in 1989 will provide 25 million dollars annually in federal matching funds over the next 15 years for restoration of wetlands vital to waterfowl and other migratory birds. Wildlife managers (Weller, 1987, 1990; Weller et al., 1991) have effected other restoration and enhancement efforts for years, in some cases to counteract the effects of previous management efforts (Talbot et al., 1986; Newling, 1990; Rey et al., 1990). At a generally smaller scale, restoration designs occur as mitigation requirements for regulated wetland fills (Kusler and Kentula, 1990). Designing wetlands for wildlife through creation of wetlands is undoubtedly a greater challenge than preservation or restoration. Whereas design through ©2001 CRC Press LLC

preservation is accomplished through observation of current wildlife use, and design through restoration is accomplished through historical knowledge of wildlife use, creation requires the attraction of wildlife to a new resource. Prominent among creation efforts in the United States is the Dredged Material Research Program of the U.S. Army Corps of Engineers (1976). Authorized by the River and Harbor Act of 1970, the USACOE Waterways Experiment Station (WES) initiated research in 1973 that included testing and evaluation of concepts for wetland and upland habitat development (Garbisch, 1977; Lunz et al., 1978a). WES has since designed and constructed thousands of acres of freshwater and coastal wetlands from dredged material and demonstrated the value of these wetlands to wildlife (Cole, 1978; Crawford and Edwards, 1978; Lunz et al., 1978b; Webb et al., 1988; Landin et al., 1989). As is the case with restoration, the wetland regulatory process has resulted in a large number of small-scale wetlands designed at least in part for wildlife (Michael and Smith, 1985; USCOE, 1989; Kusler and Kentula, 1990). The fundamental principles for effective design are the same regardless of whether a design for wetland wildlife is accomplished through preservation of existing wildlife habitat, restoration or enhancement of historic wetland habitat, or the creation of new wetland habitat. The principles are related to minimum habitat area, minimum viable population, and tolerance of the wildlife species for disturbance. Therefore, the objective of this chapter is to discuss the effects of wetland size, the relationship of the wetland to other wetlands, and anthropogenic disturbance on wetland effectiveness for providing wildlife habitat. Size Size is generally the first factor considered in designing a wetland for wildlife. Ideally, the objectives of the design, for example provision of all life requisites for the species of interest, determine wetland size. More often, land or financial constraints, or even mitigation requirements, pre-ordain the size of the wetland. In these instances, an assessment should be made of what wildlife species can reasonably be supported. Grinnell and Swarth (1913) were the first to note the relationship between the number of species and the size of the habitat in their study of montane peaks. Following attempts to quantify the relationship for terrestrial habitats (Cain, 1938), it was the application of the concept to true islands which led to its widespread recognition (MacArthur and Wilson, 1963, 1967). In what has become known as the theory of island biogeography, the greater the size of the island the greater the species richness. This relationship is described by S = CAz where S is the number of species, A is the area, and C and z are dimensionless constants that need to be fitted for each set of species-area data (Figure 1, MacArthur and Wilson, 1967). The relationship is thought to occur primarily because larger islands have more habitat and greater habitat diversity.

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z

s

Figure 1

The theory of island biogeography suggests that the greater the size of the island, the greater the species richness (MacArthur and Wilson, 1967). The relationship is thought to occur because larger islands have more habitat and greater habitat diversity.

Although the theory has its origins in terrestrial ecology, there are reservations about the applicability of island biogeography to terrestrial reserves (Kushlan, 1979; Harris, 1984; Forman and Godron, 1986). Certainly there are inherent differences between the two systems because the nature of surrounding habitat is far more distinct for oceanic islands than terrestrial islands. This should result in differences in true island and terrestrial island immigration rates. Nevertheless, the relationship between terrestrial island size and species richness has been demonstrated to hold for birds and large mammal species and for habitat types such as forests, urban parkland, caves, and mountains (Culver, 1970; Vuilleumier, 1970, 1973; Brown, 1971; Peterken, 1974, 1977; Moore and Hooper, 1975; Galli et al., 1976; Gavareski, 1976; Whitcomb, 1977; Thompson, 1978; Fritz, 1979; Gottfried, 1979; Robbins, 1979; Bekele, 1980; Whitcomb et al., 1981; Ambuel and Temple, 1983; Lynch and Whigham, 1984). Therefore, despite inherent differences between oceanic and terrestrial islands, there is evidence that the same isolating mechanisms are operating. The degree to which these mechanisms operate is, of course, dependent upon the degree of habitat insularity, which in turn depends on species-specific habitat specificity, tolerance, and vagility. Harris (1984) has suggested that the isolating mechanisms operate most strongly on amphibians and reptiles, followed by mammals, resident birds, and then migratory birds. The degree to which the latter group is susceptible depends on breeding site fidelity and the extent to which reproduction is restricted to the breeding site. ©2001 CRC Press LLC

An estimated 47 million ha of wetlands were lost in the contiguous United States between 1780 and 1980 (Dahl, 1990). This loss has fragmented and insularized remaining wetlands, producing in many cases relatively small terrestrial islands. Wildlife populations are increasingly isolated and reduced in size, leading inevitably to extinction (Senner, 1980). Extinction occurs for several reasons. First, small, closed populations are more susceptible to extrinsic factors such as predation, disease, and parasitism, and to changes in the physical environment. Second, demographic stochasticity, the random variation in sex ratio and birth and death rates, contributes to fluctuations in population size (Steinhart, 1986), increasing the susceptibility of small, closed populations to random extinction events. Finally, small, closed populations suffer genetic deterioration, primarily due to inbreeding, leading to a decrease in population fitness (Soulé, 1980; Allendorf and Leary, 1986; Ralls et al., 1986; Soulé and Simberloff, 1986). The effects of inbreeding depression can be illustrated by considering the fate of a small, closed population (Senner, 1980). For an effective population size (Ne , number of breeding individuals) of 4, constrained by extrinsic factors to a maximum of 10 individuals, genetic heterozygosity declines with each successive generation (Figure 2). As heterozygosity declines, the average survival of offspring declines due to inbreeding depression. Inbreeding depression includes viability depression, which is the failure of offspring to survive to maturity, and fecundity depression, which is the tendency for inbred animals to be sterile. Mammals, in which the male X chromosome is always hemizygous, also suffer sex ratio depression by way of a relative increase in males. As fecundity decreases, the population size can no longer be maintained at its limit. The probability of survival, while initially very high, drops very sharply after approximately 15 generations. The population approaches extinction after approximately 25 generations. The rate of loss of heterozygosity per generation (f ) for inbreeding populations is equal to 1/2Ne , and animal breeders note an obvious effect on fecundity as f approaches 0.5 or 0.6 (Soulé, 1980). Because ∆ f = 1 – (1 – 1/2Ne ,)t substituting 0.6 for ∆ f and solving for t (number of generations) indicates that the number of generations to the extinction threshold is approximately 1.5 times Ne (Soulé, 1980). Smaller populations and those with shorter generation times become extinct in less time than larger populations and those with longer generation times (Figure 3). Domestic animal breeders have determined that an inbreeding rate of 2 or 3 percent per generation is sufficient for selection to eliminate deleterious genes (Stephenson et al., 1953; Dickerson et al., 1954). Citing differences between domestic and natural populations, Soulé (1980) has recommended that a 1 percent inbreeding rate be adapted as the threshold for natural populations. Because f = 1/2Ne ,

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Figure 2

The fate of a small (Ne = 4), closed population constrained by extrinsic factors to few individuals (10 in this example) is a decline in genetic heterozygosity, a decline in offspring survival, and a decline in population size (Senner, 1980).

the minimum effective population size is 50 if the inbreeding rate is to be maintained at 1 percent. However, even at this rate a population of Ne = 50 will lose about 1/4 of its genetic variation in 20 to 30 generations (Soulé, 1980). Setting the inbreeding rate at 0.1 percent, Franklin (1980) has recommended a minimum effective population size of 500 individuals for long-term survival. Depending upon generation length, number of young, and percent survival, the minimum viable population size may be somewhat more or less than 500 (Shaffer, 1981). Of course, genetic risks are not the only threat to long-term population survival. Demographic risks such as disease, meteorological catastrophes, and populations too dispersed to effect breeding can also be important contributors to the determination of minimum viable population size when populations are very small (Goodman, 1987). However, with few exceptions, genetic deterioration should occur well in advance of demographic extinction, and demographic risks will be seen to exacerbate genetic risks. Also, many demographic risks are largely unpredictable, therefore negating the development of effective design criteria discrete from those derived from genetic considerations. Therefore, it seems reasonable to emphasize genetic risks when estimating minimum viable population size. For reproductively isolated populations, minimum refuge size is a product of home range size and minimum effective population size. The home range sizes for many wetland wildlife species are poorly understood (as is the degree of reproductive ©2001 CRC Press LLC

e

Figure 3

Based on the rate of loss of heterozygosity per generation for inbreeding populations, the number of generations to the extinction threshold is 1.5 times Ne (Soulé, 1980). Smaller populations and those with shorter generation times become extinct in less time than larger populations and those with longer generation times.

isolation). Nevertheless, for purposes of illustration, consider three distinct wetland species: the bullfrog (Rana catesbeiana), the Pacific water shrew (Sorex bendirei), and the mink (Mustela vison). The bullfrog is territorial only during breeding and has been observed living and breeding in permanent ponds as small as 1.5 m diameter (Graves and Anderson, 1987). Pacific water shrew are territorial and have a home range of approximately 1 ha (Harris, 1984). Male and female mink have generally distinct home ranges of approximately 12 ha (Allen, 1986). Minimum refuge sizes for a minimum effective population size of 50 are 1.9 × 10–2, 54.5, and 600 ha, respectively. For a minimum effective population size of 500 individuals, refuge sizes are 0.19, 545, and 6000 ha. These estimates are likely conservative because they assume all individuals in the populations are contributing to the gene pool. As

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noted above, minimum refuge size is modified by generation length, number of young, and percent survival. Some existing preserves appear to be large enough to support minimum viable populations of at least some species. National Wildlife Refuges range in size from 0.4 ha in Mille Lacs, AK, to 8 million ha in Yukon Delta, AK, and average approximately 80,000 ha. Although many refuges consist of uplands as well as wetlands, it is clear that at least some National Wildlife Refuges are large enough to support minimum viable populations of some species. However, few opportunities remain for the preservation of such large tracts, and wetlands outside the refuge system may not be large enough to support minimum viable populations. For example, 28 percent of wetland habitats in the east and central Florida region are less than 2 ha in size, and 40 to 60 percent are less than 20 ha in size (Gilbrook, 1989). Restoration, enhancement, and creation of riverine wetlands have sometimes resulted in relatively large contiguous habitats (Baskett, 1987; Weller et al., 1991). More frequently however, these efforts result in wetlands of hundreds of ha, tens of ha, and even areas of less than 1 ha (Michael and Smith, 1988; Ray and Woodroof, 1988; Reimold and Thompson, 1988; U.S. Army Corps of Engineers, 1989; Landin and Webb, 1989). Relationship to Other Wetlands Given the paucity of large wetlands available for preservation, and the very real possibility that smaller wetlands will not support minimum viable populations of many species, it is necessary to provide mechanisms for interpopulation movement. Interpopulation movement increases effective habitat size and creates a metapopulation (Gilpin, 1987). The metapopulation is composed of an interacting system of local populations that suffer extinction and are recolonized from within the region. The metapopulation will be sustained if the source(s) of colonists are proximally located, and the immigration rate is greater than the reciprocal of the time to extinction (Brown and Kodric-Brown, 1977). The metapopulation has a decreased danger of accidental extinction compared to individual populations and an ability to counter genetic drift through occasional migration. In the absence of a metapopulation, the wetland internal disturbance regime becomes the critical design feature, and the minimum dynamic area must be large enough to support internal colonization sources (Pickett and Thompson, 1978). As discussed above, this minimum dynamic area is likely to be unattainable in many instances. Metapopulations can reasonably be established and maintained if interpopulation movement can be effected at a minimum rate of every few generations (Wright, 1969; Nei et al., 1975; Kiester et al., 1982; Allendorf, 1986). The rate of movement is a function of the distance between populations and the quality of the intervening habitat, and will vary among species based upon dispersal ability, habitat specificity, and habitat tolerance. The rate of movement between populations varies inversely with the distance between habitats (McArthur and Wilson, 1967; Diamond, 1975; Gilpin, 1987; Soulé, 1991). Animals find it more difficult to disperse from one habitat to another as the distance between habitats increases, and habitats are more likely to be recolonized following extinction events if habitats occur in close proximity (Figure 4, Wilson ©2001 CRC Press LLC

and Willis, 1975; Brown and Kodric-Brown, 1977; Soulé, 1991). For example, unoccupied spruce grouse (Dendragapus canadensis) habitat is significantly further from occupied habitat than other occupied habitat (Fritz, 1979). Rodents, rabbits, and hares appear to benefit from proximally located habitats (Soulé, 1991), and mammals typically disperse less than 5 home range diameters (Chepko-Sade and Halpin, 1987). Generally, small, sedentary, cursorial species with narrow habitat tolerances will be more greatly affected by the distance between habitats than will large pedestrian species, volant species, migratory species, and species with broad habitat tolerances.

Figure 4

Animals find it easier to disperse from one habitat to another if those habitats are closely juxtaposed than if habitats are widely separated. For example, in this photograph two wetlands are separated by about 20 m of upland. Close juxtaposition of habitats also facilitates recolonization following local extinction events.

Ultimately, interpopulation movement is determined by the quality of the intervening habitat, with quality a function of species-specific habitat tolerance (Harris, 1984; Forman and Godron, 1986; Noss, 1987). At its simplest, intervening habitat can be viewed as either unsuitable or suitable, with unsuitable habitat constituting barriers to movement and suitable habitat constituting corridors. Roadways are classic barriers that inhibit the movement of mammals (Oxley et al., 1974; Madsen, 1990), and are responsible for the death of an estimated 100 million amphibians, reptiles, birds, and mammals per year (Arnold, 1990; Anon., 1992; Lopez, 1992). Some bird species have an intrinsic aversion to abandoning cover (Diamond, 1973; Soulé, 1991), and rodents tend to stay within suitable habitat (Holekamp, 1984; Garrett and Franklin, 1988; Wiggett and Boag, 1989). At the mesoscale and ©2001 CRC Press LLC

macroscale, lakes, mountains, and valleys affect mammal dispersal (Shirer and Downhower, 1968; Seidensticker et al., 1973; Storm et al., 1976). For species such as these with poor dispersal abilities or narrow habitat tolerances, individual populations comprising the metapopulation must be connected by suitable habitat. Alternatively, discrete habitats could be closely juxtaposed. Roadway crossings should be avoided, although movement can be effected for some species by the use of underpasses (Arnold, 1990; Madsen, 1990; Soulé, 1991). Unsuitable habitat is more likely to be used as a corridor if the transit time between populations is short enough that forage and cover are not required. Other species have greater dispersal abilities and broader habitat tolerances. For example, a Florida black bear (Ursus americanus) traversed eight major highways, a dozen other roadways, a river, several canals, fences, farmland, and skirted suburban areas in an 11 week journey (Arnold, 1990). As another example, McIntyre and Barrett (1992) noted that Australian bird species preferred to move within forested areas but traversed open areas when necessary. For species such as these, intervening habitat is comprised of various degrees of suitability, each of which can be tolerated for various lengths of time. Few, if any, specific guidelines exist for determining the effective distance between populations, or the quality of intervening habitat. In the absence of specific information demonstrating the broader abilities of dispersers, wetland design efforts should focus on establishing corridors of habitat similar to that being used by existing individual populations. Individual populations should be located as closely together as possible so as to minimize transit time, and corridor width should approach home range diameter as transit time increases. Noss (1993) recommends that corridors be wide enough to minimize edge effects, meet the needs of the most sensitive species, and accommodate a variety of successional stages. This conservative approach is more likely to ensure long-term survival of species with poor dispersal abilities such as amphibians, reptiles, some bird species, and small mammals than other designs. For large mammals and some bird species, corridors may consist of dissimilar habitat if species use can be demonstrated. For migratory bird species, corridors may be largely unnecessary, and design efforts should focus on providing habitat requirements at breeding and wintering sites, and stopping points in-between. Disturbance Disturbance is a change in structure caused by factors external to the hierarchical level of the system of interest (Pickett et al., 1989). As it pertains to wildlife, disturbance alters birth and death rates by directly killing individuals or by affecting resources upon which those individuals rely (Petraitis et al., 1989). Generally, small areas are more susceptible to disturbance than larger areas (Norse et al., 1986). Disturbance can be either natural, in that it occurs as part of normal community dynamics, or anthropogenic. Natural disturbance provides for the continued existence of species which use temporary habitats (Soulé and Simberloff, 1986), and species richness is generally maximized at moderate frequencies or intensities of disturbance (Connell, 1978; Pickett and White, 1985). Natural disturbance regimes

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can be accommodated within a single large reserve (alpha diversity) or by a series of reserves (beta diversity) serving a metapopulation. By contrast, anthropogenic disturbance is generally detrimental to overall, longterm community health. Anthropogenic disturbance can be intrusive, in which people or domesticated animals enter wetland habitat and normal community processes are disrupted. Intrusive disturbance has been institutionalized at many local, state, and federal reserves with the advent of nature trails and viewing areas. Other examples of intrusive disturbance include boat and vehicle incursions, hunting, and lumbering. The primary effect of intrusive disturbance is to disrupt waterfowl, wading bird, and raptor nesting and foraging (Pough, 1951; Kushlan, 1976; Palmer, 1976; Kale, 1978; U.S. Fish and Wildlife Service, 1984; Short and Cooper, 1985; Peterson, 1986; Ambruster, 1987). The simplest way to counter intrusive disturbance is to prevent access to habitats. However, denying public access to natural areas is difficult to justify and even more difficult to enforce. Moreover, denied access diminishes the educational value of natural areas. Therefore, more realistic efforts will focus on minimizing the disruptive effects of the intrusion by avoiding sensitive areas, restricting access at critical times of the year, and limiting the number of people accessing the area at any one time. Secondary efforts will include avoiding damage to vegetative and water resources by constructing low impact trails and establishing viewing areas at the periphery of the habitat. Anthropogenic activities external to the wetland also cause disturbances. Classically, the wetland edge was viewed as an area of increased vertebrate biomass and productivity (Leopold, 1933; Lay, 1938; McAtee, 1945; Giles, 1978). Certain species of wildlife, notably deer (Odocoileus spp.), rabbits (Sylvilagus spp.), gamebirds, and some raptors, appear to benefit from the juxtaposition of wetlands and uplands (Bider, 1968; Gates and Hale, 1974; Gates and Gysel, 1978; Petersen, 1979; Wilcove et al., 1986). More recently, there has been recognition of the detrimental effects of this juxtaposition (Figure 5). The edge is subject to an increased frequency and severity of fire, hunting and poaching, nest predation, nest parasitism, and a replacement of the native mammal community by exotic species (Stalmaster and Newman, 1978; Robbins, 1979; Tremblay and Ellison, 1979; Rodgers and Burger, 1981; Whitcomb et al., 1981; Brittingham and Temple, 1983; Wilcove, 1985a, b; Klein in Brown et al., 1989; Soulé, 1991). Differences in microclimate, browsing, and other disturbances favor weedy plant species and a vegetation community that differs markedly from the interior (Wales, 1972; Ranney, 1977; Forman and Godron, 1986; Lovejoy et al., 1986). The extent of the edge effect on interior species has only been reasonably quantified for forest bird species. The effect, if any, to emergent and other open wetland systems remains to be addressed. Nevertheless, Wales (1972) and Ranney (1977) determined that major vegetation changes occur from 10 to 30 m into the forest, with the greatest effects occurring along the southerly edge. Lovejoy et al. (1986), working in tropical rainforests, found microclimate varied up to 100 m into the interior, and that interior birds did not occur within 50 m of the edge. In temperate forest, Whitcomb et al. (1981) found a negative impact from surrounding altered habitats on interior bird species occurring up to 100 m from the forest edge, whereas ©2001 CRC Press LLC

Figure 5

Although the edge was historically viewed as an area of increased productivity, recent evidence has illustrated detrimental effects to interior species from exposure to anthropogenic disturbance.

Wilcove (1985a) found edge-related nest parasitism and predation to songbirds up to 600 m from the edge. Whether the edge is viewed as beneficial or detrimental depends upon the objectives of the design. Regardless of objectives, a decrease in the ratio of interior to edge will increase the relative number of edge adapted species and decrease the relative number of interior adapted species. Isodiametric (round) wetlands will maximize the interior-to-edge ratio, whereas elongated wetlands will minimize the interior-to-edge ratio. Isodiametric wetlands also have a secondary benefit of minimizing internal dispersal distance (Diamond, 1975; Wilson and Willis, 1975). Maintaining interior adapted species with a decreasing interior-to-edge ratio will likely require increased protection and management. Clearly, detrimental edge effects on interior species will be reduced as refuge size increases. If the edge effect extends to 100 m (Lovejoy et al., 1986), then wetland refuges must be larger than 10 ha to accommodate interior species. Wetland refuges must be greater than 100 ha to accommodate interior species if the effect extends 600 m as Wilcove (1985a) has suggested. Wetlands smaller than these minimum sizes will, in theory, accommodate only edge adapted or disturbance tolerant species. However, caution must be exercised in extrapolating effects on temperate and tropical forest bird species to wetland species in general. Another way in which to minimize disturbance associated with detrimental edge effects is to establish upland buffers to the wetland refuge. Wetland-related wildlife use surrounding uplands to fulfill part of their life requisites. In the United States, ©2001 CRC Press LLC

Errington (1957) noted rabbits, woodchucks, foxes, ducks, herons, small birds, and mammals, skunks, minks, and muskrats using the upland areas adjacent to Iowa and South Dakota marshes. Deer (Odocoileus virginianus) and pheasants (Phasianus colchicus) seek cover in dense upland vegetation surrounding wetlands (Gates and Hale, 1974; Linder and Schitosky, 1979). Red-tailed hawk (Buteo jamaicensis), pheasant, northern harrier (Circus cyaneus), and leopard frog (Rana pipiens) forage in upland borders (Errington and Breckenridge, 1936; Dole, 1965; Gates and Hale, 1974; Petersen, 1979). Salamanders, and some frogs and toads, spend the majority of their adult lives in fields and forests (Behler and Find, 1979). Waterfowl, turtle, and mammal breed along the upland border of wetlands (Allen and Shapton, 1942; Errington, 1957; Jahn and Hunt, 1964; Pils and Martin, 1978; Weller, 1978; DeGraaf and Rudis, 1986; Kirby, 1988). The upland buffer also serves as a travel corridor, a refuge during periods of high water, and a shield for wetland species from anthropogenic activity (Meanly, 1972; Porter, 1981) In large part, buffers proposed to protect wetland wildlife are based upon best professional judgment or knowledge of species spatial requirements. Leedy et al. (1978) determined that a buffer of up to 92 m is necessary on either side of a stream to provide required wildlife habitat elements. This opinion has surfaced in several efforts to establish effective wildlife buffers in New Jersey (Roman and Good, 1985; Diamond and Nilson, 1988; New Jersey Department of Environmental Regulation, 1988). Adopting a more rigorous approach, Brown et al. (1989) considered spatial requirements, and using guilds and indicator species, determined that buffers should be 98 to 224 m wide, with a minimum of 15 m of upland included in the buffer. Buffers in the latter study are measured from the waterward edge of forested areas, whereas marsh buffers are measured from the landward edge of the wetland. There are few studies which note the distance at which wildlife are disturbed by human activity, and much of this information is anecdotal (Short and Cooper, 1985; Brady and Buchsbaum, 1989; Brown et al., 1989). Disturbance distances for 23 species of birds range from 6 to 459 m, and averaged 74 m (Table 1). No strong patterns are apparent that would suggest a relationship between disturbance distance and taxonomic group, body size, or ecological niche. In a less direct manner, buffers protect wetland wildlife by removing sediment, nutrients, salt, bacteria, virus, and chemical pollutants from agricultural and urban surface runoff (Karr and Schlosser, 1977; Sullivan, 1986; U.S. Department of Agriculture Soil Conservation Service, 1986; Potts and Bai, 1989). Berger (1989) suggested that one reason for a general decrease in the number of amphibians is excessive agricultural chemical pollution. Vegetated buffers can be effective in preventing or minimizing environmental degradation of wetlands (Dillaha et al., 1989). Determining appropriate buffer widths adequate to reduce the level of disturbance to wetlands from surface runoff has received more rigorous examination than attempts to establish buffers based upon direct disturbance to wildlife. Recommendations vary considerably and reflect regional and local differences in the aforementioned factors (Table 2). Empirical studies have identified buffer needs of 15 m to 61 m to prevent natural debris and sediment accumulation in streams (Trimble and

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Table 1

Disturbance Tolerance Distances for 23 Species of Birds (Short and Cooper, 1985; Brady and Buchsbaum, 1989; and Brown et al., 1985)

Scientific Name

Common Name

Disturbance Distance (m)

Anas americana Anas clypeata Anas discors Anas fulvigula Anhinga anhinga Ardea horodias Buteo jamaicensis Calidra alba Calidris alpina Calidris mauri Calidris minutilla Casmerodius albus Catoptrophorus semipalmatus Charadrius vociferus Egretta caerulea Egretta thula Eudocimus albus Fulica americana Haliaetus leucocephalus Pandion haliaetus Pelicanus occidentalis Phalacrocorax auritus Podilymbus podiceps

American widgeon Northern shoveler Blue-winged teal Mottled duck Anhinga Great blue heron Red-tailed hawk Sanderling Dunlin Western sandpiper Least sandpiper Great egret Willet Killdeer Little blue heron Snowy egret White ibis American coot Bald eagle Osprey Brown pelican Double-crested cormorant Pied-billed grebe

92 92 92 37 6 100 31 73 92 73 73 18 73 55 55 73 73 37 459 6 6 6 73

Average

74

Sartz, 1957; Broderson, 1973). Theoretical treatments have identified buffer needs of 15 to 224 m (New Jersey Department of Environmental Regulation, 1988; Brady and Buchsbaum, 1989; Brown et al., 1989; Dillaha et al., 1989; Potts and Bai, 1989; East Central Florida Regional Planning Council, 1991). Generally, wetland refuges surrounded by natural vegetation, particularly trees and shrubs that serve as visual and auditory screens from anthropogenic activity, will likely withstand smaller buffers than wetlands surrounded by disturbed habitats. Wetlands surrounded by passive recreational areas, such as golf courses and ball fields, will likely withstand smaller buffers than wetlands surrounded by residential, commercial, and industrial development. Wetland refuges surrounded by gentle slopes, natural vegetation communities, and course-grained, well-drained soils with a relatively high organic content will likely withstand smaller buffers than refuges surrounded by steeper slopes, disturbed vegetation communities, and fine-grained, poorly drained soils with a relatively low organic content. Buffer size should generally increase with an increase in the quality of the wetland, an increase in the size and intensity of surrounding development, and a decrease in surface runoff particle size.

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Table 2

Suggested Buffer Widths to Reduce the Level of Disturbance to Wetlands

Suggested Buffer (m) Up to 43

15–61

15–92

15–86+

92–122

23–224

20–30 75+

0–159

Basis Maximum distance sediment transported from logging road to streams Natural control of debris and sediment accumulation in streams Vegetation interspersion, wetland size, quality of surrounding habitat, potential for impacts Slope, vegetation, soil characteristics, value of wetland, intensity of development Water quality maintenance, public health protection, wildlife protection Groundwater drawdown, sediment and turbidity control, wildlife Vegetation Vegetation and soil characteristics, development intensity, and type Wetland quality, soil type

Study Trimble and Sartz (1957)

Broderson (1973)

Roman and Good 1985

New Jersey Department of Environmental Regulation (1988) Brady and Buchsbaum (1989)

Brown et al. (1989)

Dillaha et al. (1989) Potts and Bai (1989)

East Central Florida Regional Planning Council (1991)

Design Guidelines Wildlife responds to largely physical characteristics when selecting a habitat. Nevertheless, three factors, 1. The size of the wetland 2. The relationship of the wetland to other wetlands 3. The level and type of disturbance

will largely determine the effectiveness of a wetland for long-term wildlife use. Ideally, a wetland designed for wildlife, regardless of whether the design uses preservation, enhancement, restoration, or creation to achieve its objectives, should be large enough to support an estimated minimum viable population of the species of interest. In many cases, establishment of a wildlife community is the objective, and an indicator species with significant areal requirements should be identified. The wetland should be large enough to support at least 50, and ideally 500, breeding individuals, and area requirements can be estimated from knowledge about species home range sizes. In most instances, wetlands being designed will be large enough to support minimum viable populations of only those species with small areal requirements. For other species, the design should focus on establishment of a metapopulation through ensuring interpopulation movement. Interpopulation movement can be ©2001 CRC Press LLC

effected for all species by close juxtaposition of individual populations or provision of habitat corridors. Species with broad habitat tolerances will use a variety of landscape elements as long as barriers such as roads, waterways, and waterbodies are avoided. Wetlands for migratory bird species will not require corridors. However, designers of wetlands for migratory bird species should recognize that sustainability of a seasonal population is dependent upon one or more wetlands hundreds or even thousands of kilometers distant. Wetlands designed to support disturbance intolerant species should limit intrusions or protect critical areas during sensitive times of the year. Interior species are unlikely to thrive in wetlands less than 100 ha in size or to persist in wetlands less than 10 ha. The establishment of an upland buffer can increase the effective size of a wetland. Buffers should be established on a case by case basis through consideration of soil type, vegetation type in the buffer, adjacent land use, slope, runoff particle size, wetland quality, and indigenous wildlife. Nevertheless, a buffer of approximately 200 m width appears to be adequate in most instances to minimize direct disturbance to wildlife and to reduce water quality impacts from contaminated surface runoff. An upland buffer also provides life requisites to many wetland wildlife species.

MANAGEMENT Management Approaches Management of wildlife ranges from passive approaches exemplified by preservation of self-regulating habitat, to semi-active approaches such as the installation of nest boxes, to active approaches such as impoundments that require periodic water, soil, and vegetation manipulation. As management schemes become more active, and intrinsically more complex, monetary and labor costs increase, and the chances for sustainability and success decrease. To many, purchase of existing habitat is the only feasible way of protecting unique areas for bird nesting or migration stopovers (Weller, 1987). In the United States, this approach is illustrated by the actions of the National Audubon Society, Ducks Unlimited, The Nature Conservancy, and the U.S. Fish and Wildlife Service. The latter group has been purchasing Waterfowl Production Areas since the 1960s in an effort to maintain continental waterfowl populations. It is vainglorious to expect that managers can improve on the complex dynamic processes of natural undisturbed wetlands. Active management will, by necessity, enhance habitat for some species while degrading habitat for other species. Management may fail because of inadequate or inaccurate information, imprecise water control, colonization, and modification by nuisance species, or even political or public pressure to terminate or modify management techniques or goals (Fredrickson, 1985). Therefore, it seems reasonable to reserve active management for wetlands known to be degraded and created wetlands. Historically, wildlife management overwhelmingly emphasized waterfowl, and other species were managed incidentally, if at all (Figure 6, Payne, 1992). In large ©2001 CRC Press LLC

part, management was applied to game species (Graul and Miller, 1984). Single species (or in some cases guild) management typically included prioritization of wildlife species, determining the requirements of these species, obtaining information on local environmental conditions, and determining the wildlife value and growth requirements of local plants (Chabreck, 1976).

Figure 6

Until recently, wildlife management efforts were focused largely on preservation and creation of waterfowl production areas.

The 1970s gave rise to regulations in the United States that required the management of wildlife for diversity, as well as to an increased public interest in managing species other than waterfowl (Rundle and Fredrickson, 1981). Both consumptive and nonconsumptive species were to be preserved (Odom, 1977; Martin, 1979). The Colorado Nongame Act of 1973 required that all native species be perpetuated, and the 1976 National Forest Management Act required the maintenance of animal diversity. Managers recognized early that the single species approach was inadequate for ensuring the maintenance of diversity, and yet it was impossible to manage for all species (Wagner, 1977). Graul and Miller (1984) reviewed several approaches to managing for diversity. The management indicator approach is intended to benefit a featured species. Typically, this is a game species, but sometimes a threatened or endangered species or a species of public interest is selected (Gould, 1977). The relationship between the featured species and other species must be understood to ensure maintenance of diversity. The ecological indicator approach manages for stenotopic species, species having a narrow range of adaptability to changes in environmental conditions. The approach assumes that eurytopic species, species tolerant of wide variation in their environment, will have their requirements satisfied indirectly (Graul et al., 1976). ©2001 CRC Press LLC

The habitat diversity approach manages vegetation stand type and age class rather than individual wildlife species (Siderits and Radtke, 1977). The success of the approach is sensitive to the size of habitat blocks set aside. Finally, the special features approach emphasizes habitat features such as snags, edges, perches, etc. (Graul, 1980). None of these approaches have been tested in any rigorous manner that would permit determination of their effectiveness. Another approach applicable to active management, and also untested, is to mimic the soil, hydrology, or vegetation of a natural, undisturbed wetland that has the desired species. Management Techniques There are a number of techniques used to manage wetlands for wildlife (e.g., Weller, 1987; Whitman and Meredith, 1987; Payne, 1992). The majority of these techniques are directed at managing vegetation. Other techniques are directed at providing nonvegetative structural requirements such as feeding opportunities and breeding sites. Selected vegetation management techniques, including burning, grazing, herbicide application, mechanical manipulation, propagation, and water level manipulation, are discussed herein. Artificial breeding and loafing sites are also discussed. Vegetation Management Burning Fires were a naturally occurring event in many palustrine emergent wetlands prior to mankind's intervention. Fire functioned to eliminate accumulated plant material and to return nutrients to the soil. Burning has been widely used for marsh management, particularly in United States Gulf Coast marshes (Hoffpauer, 1968; Chabreck, 1976; Wright and Bailey, 1982). Wetland wildlife managers use fire to promote the growth of green shoots, roots, and rhizomes of grasses and sedges that are then available to foraging geese. Fallen seeds are exposed to ducks, and dead plant material is eliminated which increases the value of the habitat to ducks, muskrats (Ondatra zibethica), and nutria (Myocastor coypus). Burning creates deep pools and edge for nesting and feeding waterfowl and controls or eliminates undesirable vegetation. There are three types of burns: cover, root, and peat (Payne, 1992). Cover (surface, wet) burns are conducted when the water level is at or above the root horizon and are used to convert monotypic stands of reed (Phragmites communis), cattail (Typha spp.), or unproductive sedge to plants that provide food and cover to nutria, muskrat, duck, and geese. This is accomplished by releasing plants with an earlier growing season than the undesirable plant species. Root burns are used to control or eliminate climax vegetation or other plants of low wildlife value. Hotter than cover burns, root burns are conducted when the soil is dry to a depth of 8 to 15 cm. The roots of undesirable plants are burned, and more desirable plants, which have roots extending to greater depths, are spared. Peat burns are conducted during droughts in an effort to convert marsh into aquatic habitat. The fire burns a hole in ©2001 CRC Press LLC

the peat which then fills with water. Peat burns are more common in freshwater marshes where there is sufficient organic material in the soil to support the fire than in coastal marshes. Timing of the burn depends upon the type of burn (cover, root, or peat) and the intended objective. Patchy late summer cover burns expose insects to migrating shorebirds (Bradbury, 1938). Root and peat burns at this time of year can be used to eliminate reed and cattail, and are most effective if reflooding can be accomplished (Mallik and Wein, 1986). Late summer or early fall cover burning will decrease muskrat populations by decreasing the availability of den building material (Daiber, 1986). Elimination of undesirable woody vegetation is also accomplished in late summer or early fall through a root burn (Linde, 1985). Patchy winter cover burns increase edge and access for waterfowl nesting the following spring and provide for control of reed and cattail (Ward, 1942; Beule, 1979). Olney threesquare (Scirpus olneyi), American bulrush (Scirpus americanus), and saltmarsh bulrush (Scirpus robustus) benefit by late winter cover burning and reflooding of saltmarsh cordgrass (Spartina patens). Spring cover burns will increase muskrat populations by stimulating the production of succulent shoots (Daiber, 1986). Fires are difficult to direct and extinguish. This is particularly true in bog wetlands where fires may burn for days or weeks (Payne, 1992). A means for extinguishing the fire, either an auxiliary water supply that can be sprayed on the marsh or a means for reflooding, should be provided. Burns should not be conducted during the waterfowl breeding season because ducklings are particularly susceptible to fast fires through dead vegetation. Nor should burns be conducted in areas with a high erosion potential or in drought years unless a root or peat burn is intended. Burning has some short-term adverse impacts on wetland wildlife. Inevitably, cover is reduced, forcing ducks to concentrate in unburned areas which increases their susceptibility to predation. Winter cover is reduced, which has an ancillary effect of reducing the wetland's ability to trap and retain snowfall. This latter effect can be significant in precipitation deficit regions (Ward, 1968). Burning also results in a short-term reduction in the insect prey base (Opler, 1981). These short-term impacts are overshadowed by the long-term benefits to wildlife. Grazing Weller (1987) has suggested that bison grazing on northern prairies may have benefited certain wildlife species by opening up dense stands of vegetation. Grazing in wetlands arrests plant community succession and tends to reduce undesirable perennials and increase annuals (Chabreck et al., 1989). Grazing animals may create openings in dense vegetation bordering riparian areas (Krueger and Anderson, 1985). In uplands bordering wetlands, grazing reduces cover for predators and fuel for fires and inhibits grassland invasion by brush. Today, cattle are the primary agent for habitat management through grazing, although sheep, horses, and even muskrat can be effective. Sheep are more easily controlled than cattle and tend to be more effective at removing undesirable plants through their close grazing (Ermacoff, 1968). Horses are better at controlling woody

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vegetation (Pederson et al., 1989). Muskrats are very effective at reducing persistent emergent cover but are very difficult to control (Weller, 1987). Waterfowl are the primary beneficiaries of controlled grazing, although shorebirds and furbearers may occasionally benefit. Openings in otherwise impenetrable cover allow access by nesting ducks. Migrating and wintering waterfowl, especially snow geese (Chen caerulescens), forage on exposed seeds, sprouts stimulated by grazing, and roots and rhizomes exposed by hoof impacts (Chabreck, 1968; Daiber, 1986). Snipe (Gallinago gallinago) concentrate in overgrazed marshes, and upland sandpiper (Bartramia longicauda) are attracted by shorter cover (Chabreck, 1976; Weller, 1987). O’Neil (1949) has suggested that marshes managed for furbearers be subjected to grazing. Grazing intended to improve waterfowl nesting habitat should occur in the winter or early spring (Rutherford and Snyder, 1983), and no more than 50 percent of the forage plants should be removed annually (Payne, 1992). To increase the value of marsh habitat for migratory and wintering waterfowl, grazing may continue through spring and early summer. Cattle (or other grazing agent) should be excluded July through September to promote the growth and seed production of annual grasses and sedges (Chabreck, 1976; Chabreck et al., 1989). The marsh can be flooded in fall and winter to attract ducks or left unflooded to attract geese. There are drawbacks to the management of wetlands through grazing in that management for one guild, waterfowl, inevitably reduces habitat value for other species such as small birds and mammals. Cattle trample dens and cause underground tunnels to collapse. Domestic grazers also compete with wildlife herbivores for food. Overgrazing has more serious consequences in that it negatively impacts habitat for all wildlife species. Ducks and other waterfowl will not nest or feed in exposed areas, and overgrazing of riparian areas can destroy the streambank, increase erosion, and decrease plant vigor. Herbicide Application Application of herbicides should, in most instances, be used as a last resort. Herbicides exhibit lethal and sublethal effects on plants and animals, a problem exacerbated by the difficulty needed to apply and control their application. Therefore, application is likely to affect nontarget species, including desired species. Furthermore, some herbicides will persist in the soil and potentially could contaminate surface water and ground water. In aquatic environments, treatment of large areas can result in oxygen depletion from decaying plants. To minimize the potential negative impacts of herbicide application, herbicides should only be used where other methods are ineffective, treat the smallest area possible, apply the herbicide when hazards to wildlife are least, and follow the manufacturer’s instructions. The safest herbicides are organophosphates and carbamates because they persist in the environment for relatively short periods compared with organochlorine herbicides. Despite the drawbacks of herbicide use, they can be effective in eliminating or reducing dense emergent vegetation. Herbicides have successfully been used to create open water areas in dense marsh vegetation, thus increasing habitat value for ©2001 CRC Press LLC

waterfowl (Weller, 1987; Payne, 1992). Water hyacinth (Eichornia crassipes) has, for the most part, been impossible to control without the use of herbicides. Extensive efforts have been undertaken along the eastern seaboard of the United States to rid coastal marshes of common reed and to encourage revegetation with native species (Jones and Lehman, 1987). Application of herbicide to reed is generally required in two successive years, with burning off of dead reed following treatment. Annuals, biennials, and perennials are most easily controlled at the seedling stage. Summer annuals are best treated in spring, and winter annuals in fall, during germination. Biennials are most effectively treated as rosettes in the fall. The vegetative stage of annuals and perennials is moderately to poorly controlled with herbicides, and control during seed set is largely ineffective (Hansen et al., 1984). Aquatic vegetation is best treated when the foliage is above the water and is rain or dew free (Gangstad, 1986). Herbicides injected into the ground are most effective when applied in the spring. Cattail and common reed are treated somewhat differently than other plants. Existing growth should be cut so that the herbicide can be applied to seedlings. If cutting is not feasible, a herbicide that will be translocated from the leaves to the rhizome should be applied to the foliage in the fall immediately prior to senescence. For the most part, woody perennials can be effectively treated at any time of the year. Herbicides can be applied to cuts or notches in the trunk or to stumps after cutting. Herbicides applied to the canopy should occur after the leaves have fully formed but prior to the development of a heavy cuticle. Mechanical Management Mechanical methods of management include blasting, bulldozing, crushing, cutting, disking, draglining, and dredging. These methods are applied to decrease the density of existing vegetation or to ensure the maintenance of open areas in newly created or restored habitat. This is achieved by scraping vegetation, deepening basins, breaking up sod-forming grasses or organic deposits, and retarding woody growth. The result is the creation of open water areas for feeding and breeding waterfowl in deeper marshes and for ground-dwelling wildlife in shallow marsh and wet meadow habitat (Linde, 1985; Weller, 1987; Payne, 1992). Also, mechanical management is used to prepare newly constructed wetlands or dewatered wetlands for planting. Prior to management through mechanical methods, an evaluation should be conducted that compares long-term expected benefits to short-term impacts to plant and animal communities. Blasting Blasting is a relatively inexpensive method for creating openings in dense, emergent vegetation. It is best applied in mineral soils, when the water level is at the soil surface to 20 cm below the surface (Hopper, 1971). Historically, dynamite was used to blast openings, but beginning in the 1960s ammonium nitrate fertilizer mixed with fuel oil has been used. The latter is approximately 1/10th the cost of dynamite and much safer to handle. ©2001 CRC Press LLC

Basins created by blasting are generally bowl or cone-shaped with steep sides. Although blasting is the least expensive way to create open water areas and to reduce the density of vegetation, there is very little control over the shape and size of the areas created. Sloughing of the sides will likely occur over the first year or two following blasting, and basins in organic soils may refill requiring a repeat of the process. Bulldozing, Draglining, and Dredging Bulldozing, draglining, and dredging are used to open up dense marsh by scraping or basin deepening (Figure 7). To prevent regrowth, basins must be excavated below the photic zone. The methods can be applied to dewatered impoundments, or to newly constructed wetlands that have not been flooded, to the benefit of waterfowl and muskrat (Linde, 1985; Weller, 1987). The abrupt edges and absence of vegetation which characterize newly constructed basins cause them to appear unnatural for a few years and will likely be less attractive to water birds (Provost, 1948; Strohmeyer and Fredrickson, 1967; Weller, 1987).

Figure 7

Bulldozers can be used to open up dense vegetation or create new basins.

Bulldozing is the more economical of the methods and allows the most accurate contouring. It can be effective in dry to moist mineral soils. Draglines are used in wetter situations and are more difficult to control than a bulldozer. The dragline can be operated from upland, on mats placed on the wetland, or on a frozen surface during winter in temperate regions. Dredging is the most costly of the methods and is generally impractical except for very large projects.

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Crushing Crushing is an effective method for creating openings in dense, persistent emergent vegetation such as cattail and common reed. It is most effective when treatment takes advantage of the plant’s natural carbohydrate translocation cycle (Beule, 1979). Carbohydrate reserves stored in the rhizome are used in the spring to produce new leaves. When fully formed, the new leaves will replace the drawn-down reserves by translocating the products of photosynthesis. Crushing of the new leaves when they are nearly fully formed interrupts the cycle and rhizome reserves are depleted. Regrowth will be stunted or precluded. Successive annual treatments are generally sufficient to permit establishment of more desirable plant species from the seed bank or from planting or seeding. Cutting Plants vary in their response to cutting. For example, cattail abundance is generally decreased by cutting, especially successive annual cutting. By contrast, marsh smartweed (Polygonum coccineum) appears to tolerate and, perhaps, even to be stimulated by cutting (Pederson et al., 1989). Cutting is effective because rhizomes are dependent upon oxygen which is supplied to the rhizomes by way of aerenchyma in the leaves. If the leaves are cut, and the substrate remains inundated, oxygen flow to the rhizomes is cut off and the plant will die (Laing, 1941). An effective strategy is to cut the leaves as close to the ground surface as possible in the fall or winter. In colder climates, the leaves can be on ice. Cutting success can be improved by inundating the stubble to a depth of 10 to 20 cm and repeating the operation for 2 or 3 consecutive years. Disking Disking is used to break up sod-forming grasses too dense to be used by grounddwelling wildlife. Also, disking helps to prevent growth of woody vegetation, retards filling in of wetlands by breaking up organic deposits, prepares new marsh for planting by breaking up sod and peat, and aerates dewatered soils thereby stimulating plant growth upon reflooding (Griffith, 1948; Burger, 1973). The technique is most effective if repeated every 3 to 4 years. Disking is largely ineffective for control of cattail and other plants with rhizomes due to resprouting (Payne, 1992). Managers of marshes in the southeastern United States use disking to stimulate the growth of Asiatic dayflower (Aneilema keisak) which is a valuable duck food. Disking occurs in February or March. Fall disking creates habitat for common snipe (Gallinago gallinago). Propagation Propagation of wetland plants was relatively common in the 1940s, and was used primarily to establish emergents and marsh edge species (Linduska, 1964). The technique has experienced a revival of sorts with the advent of wetlands mitigation. ©2001 CRC Press LLC

However, propagation is a relatively expensive operation, and failures are still common. In most instances, the natural seed bank will be more than adequate provided water levels are appropriate (Ermacoff, 1968; Weller, 1987). Nevertheless, propagation can be an effective strategy for many wetlands (Figure 8). For example, three-cornered rush (Scirpus olneyi) can be successfully established in the southeastern United States salt marsh, where it enhances habitat for snow geese and muskrat (Ross and Chabreck, 1972). It can also be an effective strategy for small areas and for newly created areas if invasion by undesirables species is likely (Weller, 1990; Payne, 1992). Payne and Copes (1986) suggest that propagation is a reasonable strategy if less than 15 percent of an existing wetland is vegetated with desirable perennial species.

Figure 8

Although the natural seed bank will generally be more than adequate under appropriate conditions, propagation can be effective for small or newly created wetlands.

Propagules come in many forms including seed, rootstock, rhizome, tuber, cutting, seedling, and transplant. Seeding constitutes the least expensive propagation method and is especially appropriate for large areas. It is also the most convenient method, in that seeds can be stored until needed. Note, however, that the cost and logistical difficulty of seeding increase dramatically if seeds are collected rather than purchased. Mulch can increase the success of most seeding efforts. The mulch shelters seedlings, helps to retain soil moisture, and reduces soil erosion until such time as the propagules establish soil-holding root systems. Tubers can also be broadcast and, in some instances, can be harvested mechanically. Tubers, roots, and rhizomes make maximum use of the plant material but are susceptible to washout if flooding occurs after propagation (Payne, 1992). Cuttings are a relatively low cost option and can be collected fairly quickly. As with tubers, ©2001 CRC Press LLC

roots, and rhizomes, cuttings are susceptible to washout. Survival is generally lower than that of rooted propagules (Payne, 1992). Seedlings and transplants are the most expensive propagules but exhibit generally greater survival than other propagules (Payne, 1992). Seedlings generally permit greater flexibility because they can be stored and planted over longer periods of time than transplants. Transplanting should ideally occur during dormancy. Propagules should be selected on a site-specific basis. Undoubtedly, the wildlife manager will select plants that maximize use of the wetland by target wildlife, for example, summer annuals for ducks or annual and perennial grasses and legumes for geese. However, care should be taken to ensure that the selected species are appropriate for the intended soil and hydrological conditions. Water quality, especially turbidity and pH, can also be important factors contributing to propagation success. Whenever possible, propagules should originate from a nearby site, or the site itself, as native, locally adapted species are most successful (Coastal Zone Resource Division, 1978). Other factors, such as the ability to withstand waves or ice, can be important in certain situations. Annuals are propagated by seed. Broadcasting is less reliable than row cropping and requires 50 to 55 percent more seed (Crawford and Bjugstad, 1967). Perennials are propagated by seed or vegetative propagule. The latter should consist of vigorous stock, and care should be taken not to damage roots and the stem during removal, transport, and planting. The propagule should be pruned after planting to decrease transpiration, and a support should be provided to prevent wind damage. Large, dense plantings are more resistant to herbivory than small, sparse plantings (Payne, 1992). To ensure that adequate nutrients are available, the propagule can be fertilized as planted, or the surrounding soil can be fertilized just prior to planting. Slow release fertilizer is generally best and can be placed in or near the planting hole. If placed in the planting hole, the fertilizer should be well below the root ball so as not to burn the roots. As a maintenance measure, fertilizer can be applied during or just before the growing season. Warm season species are best fertilized in the spring or early summer, whereas cool season species are best fertilized at the time of seeding and again at midwinter (Carpenter and Williams, 1972). Water-Level Manipulation Water depth and hydroperiod determine wetland plant community composition. Therefore, water-level manipulation encourages or discourages particular types and species of plants. Wildlife is attracted to the type of vegetation community that develops in response to water-level manipulation and is directly affected through the creation or elimination of aquatic habitat. Manipulating water levels in wetlands can be an effective management technique for increasing wildlife productivity (Wilson, 1968; Chabreck, 1976; Rundle and Fredrickson, 1981; Fredrickson and Taylor, 1982; Knighton, 1985). The degree of substrate drainage, surface relief, and substrate composition are equally important in determining vegetation interspersion (Knighton, 1985). Water-level manipulation can be accomplished relatively inexpensively if a small, simple water control ©2001 CRC Press LLC

structure can be constructed at the downstream end of a naturally occurring basin. The water control structure should be capable of effecting a complete drawdown and should be able to manipulate water levels with a precision of 5 to 7 cm. Costs increase dramatically if levee construction is required. Water level manipulation can also be affected by pumping water in or out of an impoundment, although this option is more costly than periodic adjustment of a water control structure. Deep organic soils (greater than 15 cm) generally preclude management through water level manipulation because reflooding results in floating mats (Knighton, 1985). Manipulation includes both drawdown and flooding. Drawdowns reduce or eliminate undesirable plant species, facilitate decomposition of vegetation and the return of nutrients to the soil, allow desirable plant species to germinate or recover from flood stress, concentrate prey for wildlife, and reduce or eliminate nuisance fish and wildlife (Kadlec, 1960; Linde, 1969; Weller, 1987). Flooding decreases the density of emergent vegetation and increases the production of invertebrates, thereby enhancing the suitability of the wetland for waterfowl breeding and brood rearing. The response of plants to the manipulation of water levels depends on the timing and extent of drawdown and flooding and the plant community successional stage. Consequently, wildlife community composition and productivity are dependent on these factors. For example, late summer flooding will freeze-proof emergent marshes, thereby increasing muskrat numbers. Conversely, a partial late fall drawdown will expose invertebrates and minnows to migrant ducks. A fall drawdown conducted over several years will reduce or eliminate muskrat, carp (Cyprinus carpio), and bullhead (Ictalurus spp.). Management techniques using water level manipulation have been developed largely to benefit waterfowl. Generally, achievement of a 1:1 ratio of emergent vegetation to open water will maximize waterfowl use (Weller and Spatcher, 1965; Weller, 1975; Bookhout et al., 1989; Pederson et al., 1989). Maintenance of pool level will encourage perennial emergent vegetation suitable for dabbler duck nesting and brood rearing. A partial drawdown in spring will encourage regrowth of perennials if more cover is needed (Weller, 1987; Bookhout et al., 1989). Management for migratory and wintering waterfowl requires periodic drawdown and reflooding. Typically, drawdown occurs in spring or early summer so that soils can drain, thereby stimulating germination and growth of grasses, sedges, and other seed-producing plants from the seed bank (Kadlec, 1960). The seed bank may be inadequate in saline wetlands and impounded bays (Pederson and Smith, 1988). Floating-leaved and submergent vegetation density will be reduced by drawdown, and perennial emergent vegetation growth will be suppressed. Complete drawdown exposes mudflats and encourages dense stands of moist soil plants, primarily nonpersistent annual and biennial emergents which are prolific seed producers. Seed viability declines if mudflat emergents are continuously flooded for several years (Knighton, 1985). Complete drawdown also encourages undesirables such as willow (Salix spp.) and purple loosestrife (Lythrum salicaria), therefore, periodic inspections will be required to prevent invasion by these species. The wetland is reflooded in fall to make moist soil plant seeds available to waterfowl. Reflooding should be gradual to avoid flotation of emergents, scouring, and mortality from turbidity (Weller, 1987). Reflooding to a depth of 10 to 25 cm benefits dabbling ducks, ©2001 CRC Press LLC

whereas diving ducks and common moorhen (Gallinula chloropus) benefit from deeper depths (Payne, 1992). Water levels can also be manipulated to manage wetlands for rails and shorebirds (Neely, 1959; Griese et al., 1980; Rundle and Fredrickson, 1981; Payne, 1992). Rails are attracted to wetlands with robust emergent vegetation and water depths less than 50 cm deep, and preferably less than 15 cm deep. When spring rail use is desired, wetlands vegetated with annual grasses and smartweeds must be dewatered over the winter to protect vegetation from ice and waterfowl. Fall use can be encouraged by reflooding drawn down wetlands in late summer (Johnson and Dinsmore, 1986). Shorebirds are attracted to gradual drawdowns creating extremely shallow water (0 to 5 cm) interspersed with exposed, saturated soil. Spring flooding or disking can be used to suppress growth of aquatic plants. Rail management will, to some extent, also benefit dabbling ducks, whereas shorebird management will provide some benefit to geese. Manipulation of water levels in hardwood forests can encourage oaks and other mast producing species to the benefit of ducks, beaver (Castor canadensis), woodcock (Scolopax minor), mink (Mustela vison), squirrel (Sciurus spp.), raccoon (Procyon lotor), and muskrat. These so-called green tree reservoirs are flooded to a mean depth of approximately 40 cm in the fall and drawn down in late winter or early spring prior to the start of the growing season. Complete drawdown is essential to preclude development of undesirable vegetation (Hunter, 1978; Vaught and Bowmaster, 1983). Waterfowl are especially benefited by the described manipulations because flooding makes mast available to wintering waterfowl, and the drawdown concentrates invertebrate prey items for migrating waterfowl. Water-level manipulation is a long-term management strategy, and some type of managed disturbance will be required at intervals of 5 years or less (Payne, 1992). Drawdowns for waterfowl should reasonably be accomplished every 2 to 4 years in marshes. Moreover, the cyclic nature of water-level manipulation management, variation inherent in the seed bank, and seed bank response to edaphic conditions will likely result in differing marsh vegetation communities from year to year. In green tree reservoirs, the soil should be allowed to dry out every few years so as to discourage the establishment of vegetation characteristic of wetter habitats. Flooding should be withheld for a 2 to 3 year period after acorn production to encourage seedling establishment. Annual differences in wildlife community composition and use can be expected. Artificial Nesting and Loafing Sites Development of upland areas adjacent to wetlands and overharvesting of trees have contributed to a decline in waterfowl populations. Artificial nesting and loafing sites can be an effective means for increasing the local habitat carrying capacity for waterfowl when upland nesting sites are limited (Johnson et al., 1978; Lokemoen et al., 1984). Artificial areas increase the shoreline to wetland surface area ratio, thereby increasing the amount of sites available to breeding pairs. Islands provide an increased measure of security from predators as well as reducing the level of anthropogenic disturbance. Nevertheless, artificial areas are less likely to be accepted ©2001 CRC Press LLC

by wildlife and are more likely to be used by upland nesting waterfowl such as Canada goose (Branta canadensis) than by marsh edge birds such as American coot (Fulica americana) (Weller, 1987). Artificial nesting and loafing sites should be located in areas frequented by waterfowl, but where natural nesting and loafing sites are limited. Depending upon local predators and the nature of proximal disturbance, islands should be located 9 to 170 m from the mainland and closer to leeward than to the windward side of the mainland (Jones, 1975; Giroux, 1981; Ohlsson et al., 1982). Water depth around the island should be 0.5 to 0.75 m (Hammond and Mann, 1956). Islands should be 0.5 to 5 ha in size because smaller areas are too small to support a breeding pair and larger areas might support predators (Duebbert, 1982; Higgins, 1986). Long, narrow, rectangular islands will maximize the number of breeding pairs. A low profile will be less attractive to predators, although the area should be high enough to prevent flooding of ground nesters (Hoffman, 1988). Groundcover should be fairly dense (greater than 50 percent) and consist of grass, legumes, and forbs. Coots and grebes like vegetation extending into the water (Swift, 1982). Dense grass will encourage nesting by gulls, whereas woody shrubs will provide habitat for herons (Soots and Parnell, 1975). Trees should be avoided because they provide perches for raptors and crows. Fisheries Managing wetlands for wildlife while at the same time maintaining a fisheries is difficult, and in many instances the two are mutually exclusive. The typical freshwater wetland managed for wildlife is subject to variations in water level, whereas management for fisheries requires a relatively stable water level. Water level fluctuation increases turbidity, and periodically increases water temperature and decreases dissolved oxygen levels. Freezing to the bottom is also likely to occur in shallow impoundments. As such, only those species tolerant of fluctuating and relatively extreme conditions, such as carp and bullhead, will persist, and then only if deeper areas are provided which can serve as refugia. Carp and bullhead typically are discouraged in managed freshwater wetlands because as they dislodge vegetation and increase turbidity. Wetlands connected to deepwater habitats provide a better opportunity for reconciling the conflict between wildlife and fisheries. Diked wetlands on Lake Erie that are managed for waterfowl allow for fish movement through water control pipes (Petering and Johnson, 1991). Relatively many species have been recorded in the wetlands (Johnson, 1989), but for the most part these species are tolerant of extreme conditions and do not use the diked wetlands for spawning. The effectiveness of the diked wetlands for fisheries is limited by the number of access points and the provision of sufficiently deep water only during the fall waterfowl migration period. Limited access and lowered water levels also limit the fisheries’ value of impounded coastal wetlands. Waterfowl management requires the closure of impoundments during peak fish recruitment periods. As such, marsh nursery use by estuarine-dependent saltwater species decreases, whereas use by freshwater species increases (Herke et al., 1987). Nevertheless, use of impounded coastal ©2001 CRC Press LLC

wetlands by marine transient species can be increased by providing deepwater refugia such as perimeter ditches, and by reducing impediments to movement between the impoundment and deepwater areas. The latter can be accomplished by increasing the number of water control structures and by using water control structures such as variable crest weirs or slotted weirs which permit fish movement.

REFERENCES Allen, A. W., Habitat Suitability Index Models: Mink, Revised, U.S. Fish and Wildlife Service Biological Report 82(10.127), 1986. Allen, D. L. and Shapton, W. W., An ecological study of winter dens, with special reference to the eastern skunk, Ecology, 23(1), 59, 1942. Allendorf, F. W. and Leary, R. F., Heterozygosity and fitness in natural populations of animals, in Conservation Biology: Science of Scarcity and Diversity, Soulé, M. E., Ed., Sinauer Associates, Sunderland, MA, 1986. Allendorf, R. W., Genetic drift and the loss of alleles versus heterozygosity, Zoo Biol., 5, 181, 1986. Ambruster, M. J., Habitat Suitability Index Models: Greater Sandhill Crane, U.S. Fish and Wildlife Service Biological Report 82(10.140), 26, 1987. Ambuel, B. and Temple, S. A., Area dependent changes in the bird communities and vegetation of southern Wisconsin forests, Ecology, 64, 1057, 1983. Anon., Animal road kill toll tops 2,600 at 64 of Florida’s state parks, Florida Environ., June 1992, 6(6), 1992. Arnold, C., Wildlife corridors, CA Coast Ocean, Summer 1990, 10, 1990. Baskett, R. K., Grand Pass Wildlife Area, Missouri: modern wetland restoration strategy at work, Incr. Our Wetland Res., 220, 1987. Behler, J. L. and Find, F. W., The Audubon Society Field Guide to North American Reptiles and Amphibians, Alfred A. Knopf, New York, 1979. Bekele, E., Island Biogeography and Guidelines for the Selection of Conservation Units for Large Mammal, Ph.D. dissertation, University of Michigan, Ann Arbor, 1980. Berger, L., Disappearance of amphibian larvae in the agricultural landscape, Ecol. Int. Bull., 17, 65, 1989. Beule, J. D., Control and Management of Cattails in Southeastern Wisconsin Wetlands, Wisconsin Department Natural Resources, Technical Bulletin 112, 40, 1979. Bider, J. R., Animal activity in uncontrolled terrestrial communities as determined by a sand transect technique, Ecol. Monogr., 38, 269, 1968. Bookhout, T. A., Bednarik, K. E., and Kroll, R. W., The Great Lakes marshes, in Habitat Management for Migrating and Wintering Waterfowl in North America, Smith, L. M., Pederson, R. L., and Kaminski, R. M., Eds., Texas Technical University Press, Lubbock, 1989, 131. Bradbury, H. M., Mosquito control operations on tide marshes in Massachusetts and their effect on shore birds and waterfowl, J. Wild. Manage., 2, 49, 1938. Brady, P. and Buchsbaum, R., Buffer Zones: The Environment’s Last Defense, Report submitted to the City of Gloucester, MA by Massachusetts Audubon: North Shore, 1989. Brittingham, M. C. and Temple, S. A., Have cowbirds caused forest songbirds to decline? BioScience, 33, 31, 1983.

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Broderson, J. M., Sizing Buffer Strips to Maintain Water Quality, M. S. thesis, University of Washington, 1973. Brown, J. H., Mammals on Mountaintops: Nonequilibrium Insular Biogeography, Am. Natural., 105, 467, 1971. Brown, J. H. and Kodric-Brown, A., Turnover rates in insular biogeography: effect of immigration on extinction, Ecology, 58, 445, 1977. Brown, M., Schaefer, J., and Brandt, K., Buffer Zones for Water, Wetlands, and Wildlife in the East Central Florida Region, Center for Wetlands Publication #89–07, University of Florida, Gainesville, FL, 1989. Burger, G. V., Practical Wildlife Management, Winchester Press, New York, 1973. Cain, S., The species-area curve, Am. Midl. Natural., 19, 573, 1938. Carpenter, L. H. and Williams, G. L., A Literature Review on the Role of Mineral Fertilizers in Big Game Range Management, Colorado Game, Fish and Parks Department Special Report 28, 1972. Chabreck, R. H., Weirs, plugs and artificial potholes for the management of wildlife in coastal marshes, Proc. Marsh Estu. Manage. Symp., 1, 178, 1968. Chabreck, R. H., Management of wetlands for wildlife habitat improvement, in Estuarine Processes, Vol. 1, Wiley, M., Ed., Academic Press, New York, 1976, 226. Chabreck, R. H., Wildlife harvest in wetlands of the United States, in Wetland Functions and Values: The State of Our Understanding, Greeson, P. E., Clark, J. R., and Clark, J. E., Eds., American Water Resources Association, Minneapolis, MN., 1979, 618. Chabreck, R. H., Joanen, T., and Paulus, S. L., Southern coastal marshes and lakes, in Habitat Management for Migrating and Wintering Waterfowl in North America, Smith, L. M., Pederson, R. L., and Kaminski, R. M., Eds., Texas Tech University Press, Lubbock, 1989, 249. Chepko-Sade, B. D. and Halpin, Z. T., Mammalian Dispersal Patterns, University of Chicago Press, Chicago, IL, 1987. Coastal Zone Resources Division, Handbook for Terrestrial Wildlife Habitat Development on Dredged Material, U.S. Army Corps of Engineers Technical Report D-78–37, 1978. Cole, R. A., Habitat Development Field Investigations, Buttermilk Sound Marsh Development Site, Atlantic Intracoastal Waterway, Georgia: Summary Report, Technical Report D78–26, U.S. Army Corpos of Engineers Waterways Experiment Station, Vicksburg, MS, 1978. Connell, J. H., Diversity in tropical rain forests and coral reefs, Science, 199, 1302, 1978. Cowardin, L. M. and Goforth, W. R., Summary and research needs, in Water Impoundments for Wildlife: A Habitat Management Workshop, General Technical Report NC-100, U.S. Department of Agriculture, Forest Service, North Central Forest Experimental Station, 1985, 135. Crawford, H. S., Jr. and Bjugstad, A. J., Establishing Grass Range in the Southwest Missouri Ozarks, U.S. Forest Service Research Note NC-22. 4, 1967. Crawford, J. A. and Edwards, D. K., Habitat Development Field Investigations, Miller Sands Marsh and Upland Habitat Development Site, Columbia River, Oregon, Appendix F: Postpropagation Assessment of Wildlife Resources on Dredged Material, Technical Report D-77-38, U.S. Army Corps of Engineers Waterways Experiment Station, Vicksburg, MS, 1978. Culver, D. C., Analysis of simple cave communities. I. Caves as islands, Evolution, 24, 463, 1970. Dahl, T. E., Wetland Losses in the United States 1780’s to 1980’s, U.S. Department of the Interior, Fish and Wildlife Service, Washington, D.C., 1990.

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Daiber, F. C., Conservation of Tidal Marshes, Van Nostrand Reinhold, New York, 1986. DeGraaf, R. M. and Rudis, D. D., New England Wildlife: Habitat, Natural History and Distribution, U.S. Department of Agriculture Forest Service, Northeast Forest Experiment Station General Technical Report NE-108, 1986. Diamond, J. M., Distributional ecology of New Guinea birds, Science, 179, 759, 1973. Diamond, J. M., The island dilemma: lessons of modern biogeographic studies for the design of natural preserves, Biol. Conserv., 7, 129, 1975. Diamond, R. S. and Nilson, D. J., Buffer delineation method for coastal wetlands in New Jersey, in Symposium on Coastal Water Resources, Proceedings of the American Water Resources Association, 1988. Dickerson, G. E., Blunn, C. T., Chapman, A. B., Kottman, R. M., Krider, J. L., Warwick, E. J., and Whatley, J. A., Jr., in collaboration with Baker, M. L., Lush, J. L., and Winters, L. M., Evaluation of Selection in Developing Inbred Lines of Swine, University of Missouri College of Agriculture Research Bulletin 551, 1954. Dillaha, T. A., Sherrard, J. H., and Lee, D., Long-term effectiveness of vegetative filter strips, Water Environ. Technol., November 1989. Dole, J. W., Summer movements of adult leopard frogs, Rana pipiens, Ecology, 46(3), 236, 1965. Duebbert, H. F., Nesting of waterfowl on islands in Lake Audubon, North Dakota, Wild. Soc. Bull., 10, 232, 1982. East Central Florida Regional Planning Council, ECFRPC Wetland Buffer Criteria and Procedures Manual, Draft No. 7, East Central Florida Regional Planning Council, Winter Park, FL, 1991. Ermacoff, N., Marsh and Habitat Management Practices at the Mendota Wildlife Area, California Department Fish Game Leaflet 12, 1968. Errington, P. L., Of Men and Marshes, Macmillan, New York, 1957. Errington, P. L. and Breckenridge, W. J., Food habits of marsh hawks in the glaciated prairie region of north-central United States, Am. Midl. Natural., 17, 831, 1936. Forman, R. T. T. and Godron, M., Landscape Ecology, John Wiley & Sons, New York, 1986. Franklin, I. R., Evolutionary change in small populations, in Conservation Biology: an Evolutionary—Ecological Perspective, Soulé, M. E. and Wilcox, M. E., Eds., Sinauer Associates, Sunderland, MA, 1980, 135. Fredrickson, L. H., Managed wetland habitats for wildlife: why are they important? in Water Impoundments for Wildlife: a Habitat Management Workshop, General Technical Report NC-100, U.S. Department of Agriculture, Forest Service, North Central Forest Experimental Station, 1985, 1. Fredrickson, L. H. and Taylor, T. S., Management of Seasonally Flooded Impoundments for Wildlife, U.S. Fish and Wildlife Service Resource Publication 148, 1982. Fritz, R. S., Consequences of insular population structure: distribution and extinction of spruce grouse populations, Oecologia, 42, 57, 1979. Gagliano, S. M., Special report on Marsh Deterioration and Land Loss in the Deltaic Plain of Coastal Louisiana, Coastal Environments, Baton Rouge, LA, 1981. Galli, A. E., Leck, E. C. F., and Forman, R. T. T., Avian distribution patterns within different sized forest islands in central New Jersey, Auk, 93, 356, 1976. Gangstad, E. O., Freshwater Vegetation Management, Thomas Publishers, Fresno, CA, 1986. Garbisch, E. W., Jr., Recent and Planned Marsh Establishment Work Throughout the Contiguous United States: A Survey and Basic Guidelines, Technical Report D-77-3, U.S. Army Corps of Engineers, Waterways Experiment Station, Vicksburg, MS, 1977.

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Buchsbaum, Robert “Coastal Marsh Management” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

11

Coastal Marsh Management Robert Buchsbaum

CONTENTS Historical Coastal Marsh Management Coastal Wetland Destruction Mosquito Control Biology of Salt Marsh Mosquitoes Habitat Alteration by Grid Ditching Pesticides and Bacterium Exploitation of Coastal Wetlands Marsh Diking Contemporary Marsh Management Recent Trends in Coastal Wetland Loss Mosquito Control by Open Marsh Water Management OMWM vs. Grid Ditching Effect of OMWM on Mosquitoes Effect of OMWM on Marsh Processes Other Potential Management Uses of OMWM Recommendations for Mosquito Control Impacts of Docks and Piers Buffer Zones and Coastal Wetlands Water Quality Aspects of Buffers Pathogenic Microorganisms Nitrogen Wildlife Habitat Aspects of Buffers Examples of Buffer Protection Programs Restoration of Degraded Wetlands with Particular Emphasis on Introduced Species Future Considerations References

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HISTORICAL COASTAL MARSH MANAGEMENT When European settlers first arrived in the northeast United States, they often settled around salt marshes (Nixon, 1982). Marshes were valued as a source of food for livestock because there was little open grazing land. Native Americans of the northeastern United States, unlike their counterparts in other parts of North America, did not regularly maintain open lands. Marshes had traditionally been used for grazing sheep and cattle in Europe (Jensen, 1985), thus it was not surprising that they would be similarly valued in the New World. As more and more farmland was cleared for pasture, attitudes toward coastal wetlands changed for the worse. Marshes were at best ignored and at worst were perceived as worthless land that bred mosquitoes and other pestilence. The best use of the coastal wetlands was in being reclaimed and put to some useful purpose. Up until about the 1970s, the two most widespread management activities in coastal wetlands were outright destruction and mosquito abatement.

COASTAL WETLAND DESTRUCTION Coastal wetlands have been filled and degraded to create more land area for homes, industry, and agriculture. Estimates of wetland lost since colonial times have not always distinguished coastal from inland wetlands, so we must rely to some extent on estimates of all wetland to estimate coastal wetland losses. Dahl (1990) estimated that the United States has lost 30 percent of its original wetlands acreage (53 percent if Alaska and Hawaii are excluded). An estimated 46 percent of the original wetlands area of Florida and Louisiana, the two states with the largest acreage of coastal wetlands (almost seven million ha combined), have been lost (Watzin and Gosselink, 1992). About 90 percent of California’s original area of wetlands have been destroyed (Figure 1, Watzin and Gosselink, 1992). Evaluations of coastal wetland loss suggest that over one half of the original U.S. salt marshes and mangrove forests have been destroyed, much of it between 1950 and the mid-1970s (Watzin and Gosselink, 1992). Between the mid-1950s and mid-1970s, the coterminous United States lost an estimated 373,300 acres of vegetated estuarine wetlands, a 7.6 percent loss (Frayer et al., 1983). Such losses and modifications have been particularly acute in San Francisco Bay. Most of the bay’s tidal marshes have been filled by the activities of gold miners, agriculture, and salt production. Hydrologic changes caused by dams, reservoirs, and canals have reduced the freshwater flow to only about 60 percent of its original volume. Similar activities have occurred in other urban areas. Major airports were built on filled tide lands in New York City, Boston, and New Orleans. The upscale Back Bay section of Boston was once a shallow embayment fringed with salt marshes. Old maps of the city indicate extensive areas of water that are now dry land. Similarly, the original shoreline of Manhattan was irregular with bays and inlets, a far cry from the present almost linear expanse of piers and highways. Marshland, with its rich, peaty soil, was often reclaimed for agriculture in Europe. Both mangrove swamps and salt marshes in Florida have also been destroyed ©2001 CRC Press LLC

Figure 1

Salt marsh dominated by pickleweek (Salicornia virginica) near Stinson Beach, CA; over 90 percent of California’s wetlands, including most of its original coastal marshes, have been destroyed.

to create waterfront homes and marinas and for the construction of the Intracoastal Waterway (Florida Department of Natural Resources, 1992a, b). Over 40 percent of the salt marshes and mangroves in Tampa Bay have been lost since 1940 (Florida Department of Natural Resources, 1992a, b). Lake Worth in Palm Beach County has lost 87 percent of its mangroves and 51 percent of its salt marshes.

MOSQUITO CONTROL Mosquito control activities in coastal wetlands have involved both physical alteration of the habitat to make it less suitable for mosquito breeding (source reduction), and the use of chemical and/or biological agents to directly kill adult and larval mosquitoes. Although the use of pesticides often receives the most public attention, habitat alteration is ultimately of more concern because of its potential to irreversibly alter coastal wetlands. Biology of Salt Marsh Mosquitoes Mosquito breeding areas on salt marshes and mangrove forests typically occur at the irregularly flooded upper edges of these habitats (Figure 2). Sites may include spring tides associated with the new and full moons. Mosquitoes may also breed among sporadically inundated tufts of high marsh plants, such as salt marsh hay ©2001 CRC Press LLC

(Spartina patens) in East Coast marshes. Eggs of most species such as Aedes solicitans, the most common nuisance mosquito in the northeastern United States, are laid on the surface of a marsh typically in shallow depressions or along the edges of drying salt pannes at least several days after the last spring tide. The eggs incubate in the air and hatch only after the subsequent spring tide or rain refills depressions on the marsh surface. The larvae, known as wrigglers because of their corkscrewlike movements, undergo four feeding stages (instars) and a nonfeeding but active pupal stage. Adults emerge in anywhere from several days to several weeks after the eggs hatch depending on the temperature.

Figure 2

Typical habitat of salt marsh mosquito larvae during a spring tide; the pools are within a short form smooth cordgrass (Spartina alterniflora) marsh and will usually dry up prior to the next spring tide precluding a permanent fish population.

Salt marsh mosquitoes typically produce several broods per year and are said to be multivoltine. Because they are tied to the lunar tidal cycle, the emergence of adults from marshes tends to be synchronized. Coastal residents experience this as periodic waves of mosquitoes, which may occur every 2 or 4 weeks depending on the height of the spring tide and weather conditions. The success of mosquito breeding on a salt marsh depends on a number of factors. If the pool dries out before the larvae can complete all stages and emerge as adults, the larvae will die. Similarly, permanent pools that support predatory fish such as Fundulus spp. and Gasterosteus spp. will not support mosquito larvae and are not a suitable habitat for eggs. Low marsh areas that are flooded daily by tides are not sites of mosquito breeding because they do not provide the prolonged period of air incubation the eggs require, and they are accessible to predatory fish.

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Habitat Alteration by Grid Ditching Although the most radical habitat alteration for mosquito control is filling the marsh, most mosquito control activities have involved water management of some kind. Habitat alteration for mosquito control in coastal wetlands reached the zenith of activity in the United States during the Depression (Provost, 1977). Both the Civilian Conservation Corps and the Works Progress Administration had programs to reclaim marshes by digging ditches at regular intervals on the marsh surface. Although these ditches were ostensibly intended to remove standing water from the marsh surface and to lower the water table, they really were built without regard for where pannes existed or mosquitoes actually bred. As a result, many marshes or sections of marshes that did not breed mosquitoes were ditched. At the time such considerations were not considered significant because a major purpose of the ditching projects was to put people to work. The grid ditching pattern, estimated to have occurred in over 95 percent of northeast marshes, is evident from an airplane. The effectiveness of controlling mosquitoes by grid ditching marshes, and its impacts on marsh processes, has been debated for the last 40 years (Bourne and Cottam, 1950; Lesser et al., 1976; Provost, 1977). The debate was largely initiated by the publication of observations that waterfowl use of a marsh in Kent County, DE, had declined after the marsh was subjected to grid ditching (Bourne and Cottam, 1950). Bourne and Cottam noted declines in invertebrate populations in the ditched portion of this marsh compared to an unditched section. They also noted the dominance of high marsh shrubs, groundsel tree (Baccharis halmifolia), and salt marsh elder (Iva frutescens) along the edge of ditches. Bourne and Cottam predicted that these high marsh shrubs would continue to spread onto the ditched marsh at the expense of the previously existing smooth cordgrass, Spartina alterniflora, as long as the ditches remained functional. This initiated a long standing debate about grid ditching between wildlife managers, whose goal was to manage salt marshes for waterfowl, and mosquito control agencies, whose goal was to reduce mosquito populations. In retrospect, there really is very little evidence on either side about the harmful effects of grid ditching on marsh wildlife (Provost, 1977). The marsh ditching debate centered on the purported lowering of water tables and gradual drying out of marshes. Clearly, a ditch that drains a panne will negatively affect wildlife that depends on that panne. But because marsh peat has such a strong affinity for water, the water table itself may only be lowered in the immediate vicinity (ca. 1 m) of the ditch (Balling and Resh, 1982). Thus, ditches are not likely to cause an overall lowering of the marsh water table. Lesser et al. (1976) reexamined the Kent County, DE, marsh in the 1970s and found that, contrary to the prediction of Bourne and Cottam, smooth cordgrass still dominated much of the ditched marsh even though the ditches were maintained in good working order. After the cessation of navigational dredging in the channel, which had caused a general lowering of the water table in the marsh, the area of high marsh shrubs had actually declined, and smooth cordgrass had increased (Provost, 1977). Dredging of navigable waters adjacent to marshes (Lesser et al., 1976) often complicates studies of the effect of ditching.

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The most intensive studies of the effects of ditching on marsh vegetation and marsh organisms have been carried out in San Francisco Bay, New Jersey, and Delaware marshes. Putting aesthetic considerations aside, ditching a marsh obviously increases the amount of tidal water flowing into the high marsh, creating narrow bands where low marsh habitats penetrate into high marsh. Strips of smooth cordgrass penetrate salt marsh hay habitat along ditches in East Coast marshes. Ditching allows the tall ecophenotype of smooth cordgrass (Valiela et al., 1978), which dominates the lower part of the intertidal zone along the edges of tidal creeks, to extend into the high marsh. Increased productivity of marsh vegetation and invertebrates can result from this change (Shisler et al., 1975; Lesser et al., 1976; Balling and Resh, 1983). The improper placement of dredge spoils and other structural alterations of the habitat, however, compromise such factors. Ditching increases the heterogeneity of the marsh, both in terms of physical characteristics and the biota. The banks of the mosquito ditches are characterized by lowered salinities compared to the adjacent high marsh because regular tidal flushing prevents the build up of hypersaline conditions (Balling and Resh, 1982). In addition, the substratum along the edge of ditches is likely to be better oxygenated than areas further back because of the lowered water table at low tide (Mendelssohn et al., 1981; Howes et al., 1981; Balling and Resh, 1982). In San Francisco Bay, pickleweed (Salicornia virginica), a low marsh species, tends to have higher productivity along ditches than elsewhere on the marsh (Balling and Resh, 1983). Balling and Resh attribute this higher productivity to the tendency of near-ditch areas to have lower salinities than the surrounding marsh. In less saline marshes, the tendency of pickleweed to be outcompeted by baltic rush (Juncus balticus), a brackish water species, is also attributed to lower average salinities along ditches. The response of invertebrates to ditching in San Francisco Bay varies seasonally. The diversity of arthropods decreased away from ditches during the dry season in San Francisco Bay salt marshes (Balling and Resh, 1982). The reverse was true during the wet season except in a natural channel and an old ditch that had relatively greater biomass of vegetation and more complex structure than most of the ditches present (Balling and Resh, 1982). Balling and Resh conclude that the arthropod community adjacent to mosquito ditches will eventually resemble that adjacent to natural channels. Along the east coast, a number of studies indicate that ditching has no marked effect on invertebrate populations of salt marshes (Shisler and Jobbins, 1975; Lesser et al., 1976; Clarke et al., 1984). Lesser et al., for example, found an increase in populations of fiddler crabs (Uca spp.) and the salt marsh snail (Melampus bidentatus) in ditched marshes compared to controls. Ditching may very likely enhance fish populations of salt marshes. Fish density and diversity increased in ponds when these were connected to a ditching system ((Resh and Balling, 1983). As long as ponds are not drained, ditching increases the amount of available marsh habitat to fish by increasing the amount of open water at high tide. It also allows the fish access to parts of the marsh that are normally not available to them. The ditches serve as corridors by which fish may enter the vegetated surface of the marsh at high tides (Rozas et al., 1988). This movement of fish, particularly the mummichog (Fundulus heteroclitus), is important to the productivity of marsh fish in that it allows ©2001 CRC Press LLC

the fish to feed on invertebrates of the marsh surface, resulting in more rapid growth rates (Weissberg and Lotrich, 1982). Ecologically, it is a mechanism by which the productivity of the vegetated surface of the marsh is transported into the surrounding estuarine habitats as these fish become prey for larger fish or birds. Using flume nets, more than 3 times as many individual fish and 14 times the fish biomass per area were caught in intertidal rivulets of tidal freshwater marshes than in larger creek banks (Rozas et al., 1988). These intertidal rivulets are structurally similar to mosquito ditches. Ditching of salt marshes has historically been considered harmful to populations of salt marsh birds (Urner, 1935; Bourne and Cottam, 1950). Clarke et al. (1984) found lower numbers of shorebirds, waders, terns, and swallows on ditched marshes compared to adjacent control marshes that had substantial areas of pannes. Because there were no differences in invertebrate populations, they attributed this observation to difficulty of foraging along ditches, possibly because of their steep sides. Other than swallows, the number of passerines (songbirds) was unaffected. Perhaps the most destructive aspect of ditching to salt marsh ecosystems has been related to the placement of dredge spoils. In many cases, spoils have simply been left along the side of the excavated creek bank where they form levees that are rarely, if ever, inundated by the tides. These levees are typically colonized by species of plants normally found at the upland edge of the marsh, such as the salt marsh elder in east coast marshes. If the levees are high enough, the normal flow of high tides over the surface of the marsh is impeded. The negative impact of dredge spoil dispersal can be avoided by proper management procedures designed to ensure that the spoils do not form levees along the border of mosquito ditches. A rotary ditcher, for example, spreads dredge spoils thinly over the marsh surface and has a temporary fertilizing effect (Burger and Shisler, 1983). Using the dredge spoils from ditches to create small islands that do not impede the general sheet flow of water over a marsh during a high tide may actually be beneficial to wildlife that require a mixture of upland and wetland habitats. Shisler et al. (1978) found that clapper rails (Rallus longirostris) frequently nested on spoil islands in New Jersey marshes. Pesticides and Bacterium Pesticides are still used to control salt marsh and mangrove mosquitoes. Broadspectrum pesticides, such as organophosphates (e.g., malathion) or pyrethroids (e.g., resmethrin), are sprayed on marshes in an attempt to kill emerging adults as they fly off the marsh. In Essex County, Massachusetts, malathion use has been timed to coincide with the emergence of adults from the marsh before they have had a chance to disperse to upland habitats (personal communication, W. Montgomery, Essex County Mosquito Control Project). These pesticides break down relatively quickly in the environment compared to those in wide use 20 years ago, such as organochlorines (e.g., DDT, dieldrin). However, organophosphates are toxic to nontarget organisms, particularly aquatic invertebrates and fish. Bacillus thuringiensis israelensis (Bti) is a bacterium that produces a protein toxin that affects mosquito larvae. Bti may be spread by hand or aerially over salt ©2001 CRC Press LLC

pannes that contain mosquito larvae. Although more specific than pesticides, Bti may still have some impact on nontarget dipterans that may occur in marshes, particularly chironomids (Lacey and Undeen, 1986). Chironomid larvae are an important item in the diet of sticklebacks (Ward and Fitzgerald, 1983). Bti treatment of salt marsh pools may potentially impact the food sources of these fish that are essential in the trophic structure of salt marshes because they are consumed by other fish, birds, and mammalian predators. Bti is less toxic to chironomid larvae than to mosquito larvae (Lacey and Undeen, 1986), thus avoiding nontarget effects on chironomids requires judicious measurement of final concentrations.

EXPLOITATION OF COASTAL WETLANDS When coastal wetlands were not being destroyed outright, or ditched for mosquito control, they were sometimes managed to provide useful products. Humans have used the vegetation itself. Salt marsh hay is still cut from northeast marshes. Although not the most ideal fodder for livestock, it has the advantage of containing virtually no weed seeds; thus, it is much sought after by gardeners for mulch. Nixon (1982) cited a 19th century survey that showed that farmers in Rhode Island cut 1557 metric tons of salt marsh hay from more than 1015 ha of marsh in 1875. The salt marsh hayers benefited from the creation of mosquito ditches that drained pannes and created a regular grid pattern on the marsh, making it easier to move equipment around on the marsh. As the hayer’s were primarily interested in the salt marsh hay, a high marsh species, they would sometimes build dikes or other barriers to restrict regular tidal inundation. Marshes have also been managed to provide wildlife for hunting. Typically, impoundments have been created on salt marshes to provide open water habitat for waterfowl. Impoundments often create a new set of problems, most notably invasion by aggressive, alien plant species such as common reed (Phragmites australis) and purple loosestrife (Lythrum salicaria) that are more tolerant of brackish conditions. Impoundments may reduce the exchange of tidal water into the marsh and, thus, reduce the ability of coastal wetlands to export organic matter into surrounding coastal waters (Montague et al., 1987). They also act as barriers to the movement of marsh fish, as well as anadromous fish, that may be passing through marshes. In tropical regions, tannins are extracted from mangrove bark, and the wood is used for charcoal. Mangrove swamps, however, have not historically been managed to the extent that salt marshes have.

MARSH DIKING Diking of marshes has been carried out to create impoundments for wildlife, for flood control, to create pleasure boating and swimming areas, and for the construction of causeways for roads and railroads. Often this causes habitat degradation behind the dike because tidal flushing is reduced and the water stagnates.

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Diking can have drastic effects on marsh vegetation and, by extension, seriously alter populations of marsh fauna. If salinities behind the dike are diminished due to reduced tidal flushing, aggressive brackish water species such as the common reed and cattails (Typha spp.) will replace the natural salt marsh vegetation (Figure 3; Niering and Warren, 1980; Roman et al., 1984; Beare and Zedler, 1987). Overall productivity of the vegetation may increase in response to lowered salinities or decrease if the tidally restricted area becomes hypersaline (Zedler et al. 1980).

Figure 3

Common reed (Phragmites australis) encroaching on salt marsh cordgrass; such scenes are common along the upland edge of East Coast marshes, particularly where tidal flow has been restricted.

Often, marsh creeks behind dikes have lower water quality than those seaward. Portnoy (1991) observed lower dissolved oxygen and higher than normal levels of sulfides behind a dike on the Herring River in Wellfleet, MA. This area is plagued by periodic fish kills and high numbers of mosquitoes, both consequences of stagnation. In the past, road construction on fill over marshes did not plan for maintenance of adequate tidal flushing in their design. Roads block sheet flow of tidal water over the marsh surface, and culverts for tidal creeks are often too small to maintain the normal tidal range and flushing. Flood and ebb tides behind a road across a marsh may be delayed several hours by an inadequately sized culvert compared to that seaward of the road, and the tidal range may be reduced by 25 per cent or more. Restoring the normal tidal circulation to a formerly diked area can reverse these negative effects. Slavin and Shisler (1983) noted substantial increases in wading birds, waterfowl, shorebirds, and gulls in a marsh when the dike of a tidally restricted salt marsh hay farm was breached. Conversely, the number of passerines declined. They also observed increases in smooth cordgrass and declines in salt marsh hay. ©2001 CRC Press LLC

Recent studies of Connecticut salt marshes have documented a striking decline in brackish species and expansion of the natural salt marsh with removal of dikes (Sinicrope et al., 1990). Simply removing a dike, however, does not always lead to the return of the natural salt marsh vegetation. If the peat has been oxidized or eroded behind the dike, the wetland surface may be lowered and the area may remain unvegetated and flooded (personal communication, J. Portnoy, S. Warren). Marshes are dynamic systems that may move up or down the shoreline in response to changes in sea level. Diking along the upland edge of marshes, a common flood control measure in urban areas, prevents the normal migration of the marsh. The future of such marshes is dubious if rising sea levels occur as a result of increases in atmospheric carbon dioxide and other greenhouse gases.

CONTEMPORARY MARSH MANAGEMENT U.S. federal and state laws and regulations reflect a new appreciation by the general public for the function and value of coastal wetlands. The outright legal destruction of large areas of coastal marshes and mangrove swamps is, hopefully, a thing of the past. Nonetheless, several significant management issues still remain. These isssues include recent wetland losses caused by direct or indirect human impacts, the effects of activities that are still permitted by federal and state wetlands regulations such as mosquito control procedures, and the construction of docks and piers over marshes. Other issues include the cumulative impact of activities in watersheds surrounding the coastal wetland including activities in wetland buffers, and restoration of degraded coastal wetlands. Recent Trends in Coastal Wetland Loss Losses of coastal wetlands still occur, albeit at a slower rate than prior to 1970. The U.S. Fish and Wildlife Service’s National Wetlands Inventory Project estimated that a net loss of 28,665 ha of vegetated estuarine wetlands occurred in the coterminous United States between 1974 and 1983 (Tiner, 1991). This is about 1.5 percent of the total existing wetland area in 1973 and represents a decline in the rate of loss from the mid-1950s through the mid-1970s. An increase of 4670 ha occurred in nonvegetated estuarine wetlands, such as tidal flats. Recent losses have been subtler than those of the past, consisting primarily of a transformation of estuarine vegetated wetlands to deepwater habitat rather than conversion to urban or agricultural land (Tiner, 1991). The majority of the recent losses have occurred in the Mississippi delta and the Florida Everglades (Field et al., 1991). Studies along the northern Gulf of Mexico have implicated rapid shoreline subsidence, and the inability of marshes to keep up with this subsidence due to relatively low accretion rates as major factors (DeLaune et al., 1989; Turner and Rao, 1990). Localized alteration of hydrology caused by the building of canals and levees for flood control has increased surface water levels on marshes, stressing and killing the vegetation (DeLaune et al., 1989; Mendelssohn and McKee, 1988). The

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break up of a vegetated wetland into smaller, and then larger, ponds can occur several kilometers from a canal (Turner and Rao, 1990). Louisiana lost 2.9 percent (23,887 ha) of its wetlands from 1974 to 1983, largely through these processes. During this same period, Texas experienced a loss of about 4049 ha and New Jersey and South Carolina lost over 405 ha (Tiner 1991). For the entire United States, 18,200 ha were lost to urban development, about 1000 ha of which were mangrove swamps in Florida. Another 1620 ha were converted to agricultural use. Mosquito Control by Open Marsh Water Management In response to the concern over grid ditching, a technique called Open Marsh Water Management (OMWM) was developed in the mid-Atlantic states in the 1950s (Ferrigno et al., 1975). OMWM consists of a system of reservoirs and canals in mosquito breeding areas that allow predatory fish, generally Fundulus sp., access to waterlogged areas of high marsh where mosquito larvae develop. Often, the reservoirs are hectare-sized champagne ponds and are at least 1 m deep to provide an adequate refuge for the fish during the several weeks of neap tide. In the northeast, old mosquito ditches are converted into reservoirs by deepening them and plugging up their junction with natural tidal creeks (Hruby and Montgomery, 1985). Canals are dug from reservoirs to mosquito breeding areas to allow passage of fish (Figure 4). The success of OMWM in controlling salt marsh mosquitoes has been documented (Ferrigno et al., 1975; Hruby et al., 1985).

Figure 4

This small reservoir pool and two shallow radial canals in a salt marsh in Gloucester, MA, are part of an OMWM system. The reservoir is 1 m deep, and the canals are 0.3 m deep. The reservoirs of OMWM systems of mid-Atlantic and southern salt marshes may be as large as 1 ha or more.

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OMWM vs. Grid Ditching The chief advantage of OMWM compared to grid ditching is that it does not drain the pannes and pools on the marsh surface. An OMWM system, unlike grid ditches, is not connected to tidal creeks. Seawater enters reservoirs and canals by sheet flow directly over the marsh surface during spring high tides, in the same manner as it floods nearby mosquito breeding habitats. The water is then trapped in the reservoirs and canals and does not drain out at low tide because the connections with the tidal creeks have been eliminated. Where water flows through old mosquito ditches, a sill at the junction of the ditch with a tidal creek, or at a relatively high point in marsh topography, achieves the same goal. OMWM systems are site specific. This is more a function of the recent overall enlightened management of salt marshes than of OMWM in particular, as grid ditching could also have been site specific. For an OMWM system to function properly, managers must identify the mosquito breeding sites through a monitoring program and then integrate the OMWM design into the hydrology of the area. Another advantage of OMWM systems compared to grid ditching is that they are easier to maintain. Grid ditches periodically have to be cleaned out or redug because the steep banks become scoured by tidal action and often collapse. This is particularly true in the northeastern United States where large tidal ranges occur. Ironically, this often creates more of a mosquito problem than was initially present because a clogged ditch that no longer allows passage to fish is an excellent mosquito breeding habitat. Portnoy (1982) found higher numbers of mosquitoes (Aedes cantator) in mosquito ditches, even those treated with a larvicide, than in natural surface pools untreated with larvicides in a diked river basin on Cape Cod. As OMWM systems are not subject to the scouring action of tidal water rushing through creeks, they should require less maintenance, although no one has actually compared the two types of systems for any length of time. Finally, the development of rotary ditchers has allowed dredge spoils to be placed over marshes in a thin layer, thus reducing impacts to vegetation. In OMWM systems in Massachusetts, vegetation visibly recovers the second growing season after deposition of spoils by rotary ditchers (personal communication, W. Montgomery, T. Hruby). In New Jersey, thin deposition of dredge had little visible effect on vegetation even in the initial year of deposition (Burger and Shisler, 1982). Effect of OMWM on Mosquitoes Monitoring the success of OMWM efforts in terms of its ability to control mosquitoes is complex because mosquito numbers and the numbers of broods produced per year vary substantially both temporally and spatially on the marsh depending on the timing of rainfall, spring tide events, and temperature. Nevertheless, Hruby et al. (1985) determined that the number of mosquito larvae declined by 75 to 99 percent in the OMWM marsh compared to the numbers in the same site the year before it was altered, while larval numbers in adjacent control areas remained roughly the same.

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In addition to allowing fish predation on the mosquito larvae, OMWM systems are likely to interfere with the hatching cycle of mosquito eggs which need to incubate for a period in air. The number of larval mosquitoes that survive to pupate as adults on the marsh surface are negatively correlated with both tidal inundation and with fish numbers (Figures 5 and 6). In a comparison of mosquito emergence in an unaltered marsh, a grid ditched marsh, and an OMWM system, significantly fewer mosquitoes were observed emerging from the OMWM system than the unaltered marsh (unpublished data). No mosquitoes emerged from the ditched marsh. Fish were present in the ditches of the grid ditched marsh and on the unaltered and OMWM system marsh surfaces during the spring tide.

Figure 5

Relationship between tide height above the marsh surface during a spring high tide and the number of mosquitoes that survive to emerge as adults; sampling occurred in a ditched marsh, an OMWM marsh, and an ulaltered control marsh.

Effect of OMWM on Marsh Processes There have been few studies of the long-term impacts of OMWM. Brush et al. (1986) concluded that OMWM had little impact on bird numbers on a Massachusetts salt marsh. In the first year after a ditched marsh was converted to an OMWM ©2001 CRC Press LLC

Figure 6

Relationship between fish numbers on the salt marsh and emerging mosquito numbers; sampling occurred in a ditched marsh, an OMWM marsh, and an unaltered control marsh.

system, shorebirds (e.g., sandpipers, plovers, and others) numbers increased, presumably because the spoils provided accessible foraging for invertebrates. Their numbers declined in subsequent years as the vegetation grew up through the spoils. Marsh passerines declined at first and then increased to pre-OMWM levels, and other groups of birds were unaffected. Brush et al. suggest that bird numbers were more closely related to the number of pannes on a marsh than to whether it was altered by OMWM, ditched, or remained unaltered. Peat cores from a northern Massachusetts salt marsh were examined for invertebrates using Berlese funnels. No significant difference in types of organisms was detected between unaltered marsh and an OMWM marsh (unpublished data). Occasionally, the reservoir pools of an OMWM system become stagnant during periods of hot weather and neap tides. Mosquito ditches that have been appropriated for OMWM seem especially prone to this because they tend to be long, narrow, and relatively deep. A thick layer of algae may form on the surface with hypoxic or anoxic water beneath. In Massachusetts’ marshes, fish kills have not been observed, but in some Florida counties stagnation and declining water quality of pools have occurred leading to declines in fish populations. Constructing channels to increase tidal exchange has mitigated the latter. Other Potential Management Uses of OMWM In the future, OMWM techniques may be useful as part of an integrated approach to restoring degraded marshes. A decline in the invasive brackish water common reed is evident in OMWM marshes that include a perimeter ditch at the border of the upland. The perimeter ditch prevents freshwater surface and groundwater flow from moving out over the natural salt marsh, and salinity levels are maintained at ©2001 CRC Press LLC

those appropriate for salt marsh vegetation. Dredge spoils created by OMWM may be potentially useful for enhancing nesting areas for marsh birds. Dredge spoils may be colonized by wildlife such as rails that forage on marshes but which require drier nest sites (Shisler et al., 1978). Judicious placement of dredge spoils into small islands rather than levees may enhance wildlife habitat of the marsh without impeding circulation. Another way to enhance wildlife habitat value is to design the reservoir pools so that they function as foraging areas for shorebirds and waders. This can be accomplished by creating a gently sloping edge to reservoir pools. Recommendations for Mosquito Control OMWM is an alteration of a salt marsh with ecological consequences and should only be used if there is a documented, compelling reason. Ideally, OMWM would comprise one facet of an integrated approach to mosquito control. At a minimum, an integrated pest management (IPM) approach to controlling nuisance mosquitoes on salt marshes is a useful theoretical framework for determining when and how to intervene. An IPM approach should begin by determining a threshold level at which control measures are necessary. Although many residents near marshes might assert that even one mosquito bite per hour is too many, a threshold level of annoyance that is specifically related to salt marsh mosquitoes needs to be determined before proceeding with any action. If the predetermined annoyance threshold has been exceeded, then trapping of adult mosquitoes is carried out to determine if mosquitoes from salt marshes are a significant cause of the nuisance. If disease is an issue, the threshold will be some indicator of the presence of the disease. If salt marsh mosquitoes are shown to be the cause of the nuisance, then monitoring of levels of larval mosquitoes in salt marshes should be initiated. Standards developed by the Massachusetts Audubon Society and the Essex County (Massachusetts) Mosquito Control Project include one full year of mosquito larvae monitoring before any alterations are considered (Hruby and Montgomery, 1986). Monitoring includes weekly sampling with a standard mosquito larvae dipper throughout the growing season at stations where mosquito larvae are suspected of occurring. Control is considered appropriate if three broods of mosquitoes containing at least five larvae per dip occur during the growing season. If two broods are observed with greater than five mosquitoes per site, then the site is monitored for another season before a decision is made. Potential impacts of alterations on other marsh organisms, such as breeding and foraging birds, are also taken into consideration before proceeding further. When control is justified, the most ecologically benign procedures or combination of procedures should be selected. A proactive approach includes public education to reduce the risk of public exposure and review of drainage plans for developments, subdivisions, and roads by appropriate local and state agencies to ensure that they do not contribute to reduced tidal flushing in coastal wetlands. A source reduction approach for disturbed habitats that supports large populations of larval mosquitoes includes OMWM (1 ha or larger) areas, or selective ditching in smaller areas. Although commonly used, pesticide application for nuisance control is less cost effective than source reduction in the long term. Pesticide use requires repeated ©2001 CRC Press LLC

application and will likely require increasing application frequencies and doses as mosquitoes develop resistance (Ofiara and Allison, 1986). Any control mechanism should be evaluated on a regular basis and modified as necessary. Both mosquitoes and nontarget organisms should be monitored. Impacts of Docks and Piers Outright destruction of coastal marshes by private landowners is no longer allowed in many parts of the country. However, building docks and piers over salt marshes is still a permitted activity. Although there are few data on this subject, there is concern on the part of resource managers about the impacts of these structures on marshes. Managers are particularly concerned about the effects of shading on vegetation, and the potential for scouring around support posts. The Massachusetts Office of Coastal Zone Management has recently sponsored research on the effects of small docks and piers on tidal flats and eelgrass beds adjacent to salt marshes. The U.S. Army Corps of Engineers, which is responsible for permitting these structures under Section 404b(1) of the Federal Water Pollution Control Act of 1972, requires that marsh vegetation beneath a dock not be impacted by the dock. No design standards are provided to ensure that light does reach plants beneath the dock. The Massachusetts Executive Office of Environmental Affairs has adopted pier guidelines (Table 1). The docks themselves may be less of a problem than the human and boat traffic they encourage. Marshes are very susceptible to trampling, as scientists who have repeatedly sampled in the same area of marsh are well aware. A catwalk over a marsh is probably a less harmful way to reach a boat than walking directly on the marsh. Footsteps gradually erode a path into the marsh surface leading to loss of vegetation and subsidence of peat. The boats themselves, maneuvering in the shallow water of marsh creeks, resuspend sediment that may affect submerged vegetation and shellfish beds. Boats also contribute to the erosion and slumping of peat along the banks of marshes. Community piers are an effective alternative to individual piers. Buffer Zones and Coastal Wetlands Wetland laws typically regulate activities within a wetland, and for a small buffer strip around a wetland. The rationale for regulatory authority over a fringe area around the wetland is that activities within that fringe potentially impact the wetland. A number of states have adopted wetland buffers ranging from 45 to 300 m from the edge of wetlands in critical areas (Table 2). Buffer distances such as these represent a compromise between protecting the private property interest of the landowner and the public interest in the wetland. In the ideal world, each buffer distance would be determined after a case by case analysis of soils, hydrology, slope, vegetational cover, the characteristics of the wetland resources being protected, and the nature of the development being proposed. Some states have adopted such a case by case approach using a ranking system based on the above criteria (Diamond and

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Table 1

Summary of Guidelines for the Construction of Piers in Critical Areas

T-docks and floats at the end of a pier are preferred in that they allow a vessel to be parallel to the shoreline and in deeper water; in any case, the stern of a vessel at a pier must be facing toward deeper water Seasonal docks are preferred over permanent ones Pile driving is preferred over jetting Docks should be as small in scope as possible with as few pilings as possible; the goal is to allow unimpeded transport of sediment Spacing of pilings should be no closer than 20× the diameter of the pilings when the dock is located in a salt marsh and never any closer together than 3 m Piers in salt marshes should be of wood construction; if treated lumber is used, only nonleaching types No stabilization structures should be proposed even if the pier is adjacent to an eroding bank Piers located in an anadromous fish run should be designed so that there is no change in the rate of flow within the run; construction should not be allowed at times when the fish are running The pier and the uses of the pier should not encroach upon navigational channels and mooring areas A dock should not be designed to end in water that is too shallow to float the vessel at all tides; however, very long elevated walkways are likely to have environmental impacts No pier should extend beyond the length of existing piers used for similar purposes and in no case should the length extend more than 1/3 of the way across a water body A higher level of environmental review is required for proposed walkways wider than 4 ft, proposed floats greater than 300 ft2, and any structure proposed to be less than 4 ft above a salt marsh; the review should clearly demonstrate that the structure will have no adverse impact on the marsh, adjacent mudflats, or submerged aquatic vegetation The dock should be designed so that the approach path of vessels is at least 50 ft from the edge of any salt marsh The primary factor in determining the location of the dock on a lot is the avoidance of sensitive resources The necessity for mitigation should be avoided to the maximum extent possible Adapted from the Massachusetts Office of Coastal Zone Management (1988).

Table 2

Buffer Distances around Critical Areas Adopted by Various States

State Maine Maryland New Jersey Rhode Island Washington Wisconsin

Minimum Buffer Zone Width (m) 45–90 90 90 60 60 300

Adapted from Brady and Buchsbaum (1989).

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Nilson, 1988). Unfortunately, political realities, the expense of evaluating individual properties, and scientific uncertainty about the relationship of the buffer to wetland functions often make a case by case approach unrealistic. The rationale for protecting coastal and inland wetlands is that they perform certain functions that are valuable to the public. Coastal wetlands enhance the water quality of coastal waters by removing potentially harmful constituents before they reach open water. Coastal wetlands are also rich habitats for fish, shellfish, and wildlife. Water quality and habitat values are related because poor water may result in increased incidence of disease in marine organisms as well as rendering some organisms, such as shellfish, unharvestable. In addition to these two values, buffers also help to maintain the aesthetics of a wetland by preserving scenic qualities. Water Quality Aspects of Buffers Pathogenic Microorganisms Buffers can play a key role in a management strategy to reduce pollution before it reaches a coastal wetland (Abernathy et al., 1985; Lawrence et al., 1985, 1986). Particulate pollutants, such as microorganisms, suspended solids, and pollutants associated with sediments, are slowed down and entrained by physical processes within the soil. Soluble nutrients may be removed by biological activity within the buffer. Pollution by bacteria and viruses from domestic wastes is a particular concern in coastal wetlands. Bacteria and viruses render shellfish unfit for human consumption and create a public health threat at swimming beaches. The percentage of potentially harmful microorganisms in sewage and stormwater that reaches a wetland area is largely a function of time. The longer it takes for these organisms to be transported from their source to a resource area, the less their numbers will be at that resource area. Bacteria and viruses are adsorbed to soil particles in the buffer, downgradient movement is slowed, and they eventually die (Hemond and Benoit, 1988). The velocity of water in surface runoff is much faster than that of groundwater. Therefore, stormwater runoff will potentially pose a greater risk of microbial contamination than domestic sewage. Stormwater runoff was the major source of high fecal coliform counts in Buttermilk Bay and many other water bodies in Wareham, MA (Heufelder, 1988). Fecal coliform bacteria were almost completely attenuated within 7 m of leaching fields. Characteristics of the buffer determine in part the extent that microorganisms are detained. Important characteristics include the permeability of the soil, slope gradient, extent of vegetation, and degree of channelization. Occasionally, water percolating through the soil will reach an impermeable layer, such as bedrock or clay, and then be channelized rapidly into coastal wetlands. The distance a microorganism will travel from its source and still remain viable depends not only on characteristics inherent to the buffer but also on the type of organism. Survival times of up to 27 weeks in soils were reported for some microorganisms, and the distances these organisms traveled from their source ranged from 2 to 837 m (Hagedorn, 1984). Viruses may travel as far as 400 m in groundwater from sewage infiltration basins (Keswick and Gerba, 1980). In seawater and freshwater, ©2001 CRC Press LLC

average times for a 90 percent decline in concentrations of coliform bacteria were 2.2 and 57 hr, respectively (Mitchell and Chamberlain, 1978). The lateral distance microorganisms will travel from a leaching field can be estimated based on these die-off rates and the groundwater flow rate. Under certain conditions, viruses and bacteria which are potentially harmful to humans may travel further from their source than buffer distances of even 100 m, particularly in poorly functioning or antiquated wastewater treatment systems or in stormwater runoff. Buffer distance is a difficult parameter to isolate in the field because it usually correlates with overall density of development. In one study, there was a significant correlation between concentrations of fecal coliform bacteria in a salt marsh creek and the number of houses within 30 m of the edge of the marsh (Bochman, 1991). The number of houses within 30 m, however, correlated with the total number of houses within the watershed. Nitrogen Nitrogen is of particular concern to coastal waters because it stimulates blooms of phytoplankton and seaweed. These blooms harm the aesthetics of coastal waters and threaten other marine plant and animal life due to the lowering of light and dissolved oxygen levels. The efficacy of removing nitrogen with a large buffer depends on whether the nitrogen enters the wetland through surface runoff or in groundwater. Nitrate in surface runoff will slowly percolate through the root zone of plants where it will be taken up by vegetation both in the buffer and the wetland (Woodwell, 1977; Ehrenfeld, 1987; Knight et al., 1987). However, if the surface runoff is very rapid, vegetation in the buffer may be ineffective in taking up nitrogen before it reaches a wetland. Nitrogen in groundwater will only be removed when it nears the surface of a wetland or a water body either through plant and algal uptake or through denitrification. When in the groundwater, nitrogen in groundwater is thought to be conservative and not subject to attenuation by biological or physical processes (Valiela and Costa, 1988). If this is the case, a large buffer will not protect a coastal embayment or salt pond from potential eutrophication unless it provides sufficient area for dilution. Instead, nitrogen can be managed by land use controls that limit the amount of nitrogen entering the watershed or by wastewater treatment procedures (denitrification systems, tertiary treatment) that remove nitrogen before it can enter groundwater. Wildlife Habitat Aspects of Buffers There are several reasons why a buffer is thought to enhance the wildlife habitat value of a coastal wetland. Many wildlife species that live in wetlands depend on the surrounding upland for cover, nesting, foraging, and migration. According to the classic ecological notion of the ecotone, wildlife will be more abundant at the wetland-buffer boundary because two habitat types coexist in close proximity. Many birds and mammals forage on the abundant invertebrates and fish of a salt marsh but require the surrounding upland for nesting or as a refuge during high tides. In ©2001 CRC Press LLC

addition to providing an alternative source of essential resources, a buffer also insulates the animals of the wetland from disturbance from developments located around its periphery, a characteristic that is particularly valuable for smaller wetlands. Human activity brings not only noise but also domestic animals and those native and alien wildlife (e.g., starlings, cowbirds) that are well adapted to human habitation and often compete with native wildlife in transitional areas. A 90 m distance has been recommended to provide a buffer against disturbance around state and federal wildlife refuges and conservation areas (Diamond and Nilson, 1988). Determining the actual impact of different sized buffers on wetland wildlife in the field is difficult, primarily because wildlife respond to a number of different aspects of the landscape that cannot be controlled. A reasonable approach is to infer the role of the wetland buffer by examining the life history of different animals. In habitat evaluation procedures, the U.S. Fish and Wildlife Service typically considers buffer size as one component in calculating habitat suitability indices. Rogers et al. (1975) also evaluate the adequacy of aquatic buffers in ranking of biotic natural areas of the Eastern Shore of Maryland. Several examples illustrate the importance of buffers to wildlife species that occur in coastal salt and freshwater marshes. The area immediately surrounding coastal wetlands is important for providing the seclusion a number of species of waterfowl need to nest free from predation and disturbance. Kirby (1988) states that, “It has long been recognized that lands adjacent to areas managed for waterfowl play a major role in the entire management scheme.” Black ducks (Anas rubripes) nest either on islands in wetlands or in areas immediately adjacent to wetlands up to 1.2 km from the wetland (Kirby, 1988). Ideal nesting habitat for black duck is heavily vegetated on at least one side to provide concealment from predators. Gadwalls (Anas strepera) typically nest in drier shoreline areas within 30.5 m of water (DeGraaf and Rudis, 1986). Canada geese (Branta canadensis) also nest near the water’s edge in fresh and salt marshes (DeGraaf and Rudis, 1986). Short and Cooper (1985) suggest that human disturbance and the resulting loss of nesting sites have been the most important factors contributing to declines in some great blue heron (Ardea herodias) populations in New York and British Columbia. Nesting sites for great blue herons may be many miles from their foraging areas (DeGraaf and Rudis, 1986), and creeks and shallow ponds on marshes are typical feeding sites. A suitable foraging area should be at least 100 m from human activities and habitation (Short and Cooper, 1985). In Essex Bay, MA, there is a significant inverse correlation between the extent of development in the buffer along the marsh edge and the number of foraging wading birds (Figure 7). The U.S. Fish and Wildlife Service’s Habitat Suitability Index (HSI) for mink (Mustela vison) assumes that a minimum strip of 100 m along the edge of a wetland enhances the value of that wetland to mink (Allen, 1986). In a study in Alaska, 68 percent of all observations of mink were either in a wetland itself or within 100 m of the wetland. Rapid declines of mink populations along the shores of Lake Ontario were associated with small increases in the human population along the shoreline. River otter (Lutra canadensis) also occur along the edges of coastal ponds, salt marshes, and estuaries (DeGraaf and Rudis, 1986).

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Figure 7

Relationship between the number of houses along the marsh edge and number of foraging wading birds in a salt marsh in Essex Bay, MA.

Examples of Buffer Protection Programs A number of states and regions have investigated the question of setting buffer distances around coastal and inland wetlands. These include Maryland (Rogers et al., 1975), Rhode Island (Rhode Island Department of Environmental Management, 1984), New Jersey Pinelands (Roman and Good, 1985), central Florida (Brown and Shaeffer, 1990), and Washington (Castelle et al., 1992a, b). In addition, Phillips and Phillips (1988) have independently proposed a method for the estimation of shoreline buffer zones. Two approaches are illustrated. Phillips and Phillips (1988) proposed a simple physical model to enable managers to predict the size of a buffer necessary to attenuate pollutants from stormwater. The delivery of pollutants to a wetland or water body is related to the energy of surface flow generated during a storm event. This energy can be predicted from 1. 2. 3. 4.

The The The The

permeability of water into the soil within the buffer slope gradient down which the stormwater runs width of the area over which the runoff flows (i.e., the buffer size) surface roughness of the buffer (Phillips and Phillips, 1988)

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The first is measured as hydraulic conductivity or infiltration rate and is generally available from the U.S. Soil Conservation Service for the various soil types. The second is obtained during the planning stages of any project. Vegetation is a major part of surface roughness which is factored into the model as a roughness coefficient. The use of this model requires field measurement of a nearby reference buffer to which the buffer under consideration is compared. Because this model is based on transport of suspended particles, it is valid for bacteria and pollutants associated with particles, but not for soluble pollutants. The New Jersey Pinelands Commission, which has regulatory authority over development in the globally significant pine barrens of southeastern New Jersey, has set up a systematic decision-making flow chart for establishing buffer distances from wetlands resources (Roman and Good, 1985). Major variables taken into account are the perceived value of the wetland resource and the nature of the activity being proposed near the wetland. The New Jersey Pinelands model includes several special cases in which the buffer is set at 90 m. These include proposed activity in a designated preservation district (the core area of the Pinelands), within 90 m of an Atlantic white cedar (Chamaecyparis thyoides) swamp, activities in which resources are extracted, and activities that include on-site wastewater treatment systems. Atlantic white cedar swamps in the Pine Barrens are afforded particularly strong protection because they contain a large assemblage of rare species, and they have historically been severely impacted by human activity. Activities proposed within existing densely developed areas are required to have a minimum buffer of 15 m. If none of the special cases apply, then the proposed activity is subject to a threestep procedure to establish the buffer distance. The evaluator of the project develops a numerical index of wetland value based on comparing the site-specific wetland to a standard evaluation scheme. Criteria comprising the value index include presence or absence of endangered species, vegetation quality, surface water quality, water quality maintenance, wildlife habitat, and socio-cultural values. Potential localized, cumulative, and watershed-wide impacts, and the slope, are all considered in deriving this index. Last, the value and impacts indices are averaged to derive a buffer delineation index. Wetlands that have a high value index and a high potential for impacts are assigned the largest buffer distance, 90 m. Restoration of Degraded Wetlands with Particular Emphasis on Introduced Species One aspect of degradation of salt marshes that is often associated with human activities is the invasion of salt marshes by introduced species, many of which are exotics (nonnative). The danger is that these species, lacking natural controls in their adopted habitats, will outcompete the native vegetation and cause overall changes in the flora and fauna of the marshes. Introduced low marsh species tend to colonize bare substrate at the lower margins of marshes (Frenkel and Boss, 1988) and do not generally compete with the native salt marsh species. A number of species of low marsh plants have been introduced into Pacific Coast marshes. These include the cordgrasses Spartina alterniflora and

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S. anglica, and the eelgrass Zostera japonica (Frenkel and Boss, 1988; Calloway and Josselyn, 1992). The cordgrass S. densiflora, a native of Chile, has been accidentally introduced into Humboldt Bay, CA. Modes of introduction of these species include transport of seeds in oyster culture and in ballast from ships. An example of an introduced low marsh plant competing with a native species is S. x townsendi, a vigorous hybrid of the native cordgrass, S. anglica, and the introduced North American species, S. alterniflora. This hybrid is outcompeting S. anglica in British salt marshes. The mid and high marshes are less susceptible to colonization by introduced species than the bare substrate of the low marsh (Frenkel and Boss, 1988). In most salt marsh systems, these sections tend to be densely vegetated already, requiring a strong competitive ability if an introduced species is to be successful. In addition, the suspected sources of marsh plant introductions, oyster culture and ballast, are generally in direct contact only with low marsh. Nonetheless, the changing mosaic nature of salt marsh vegetation (Miller and Egler, 1950) insures that bare areas are always present for opportunistic species. Salt marsh hay, a native of East Coast marshes, has colonized a number of marshes in the Pacific Northwest. In Siuslaw Estuary, OR, this species is limited to the relatively open Scirpus maritimus–Deschampsia caespitosa community of the mid marsh, where its dense growth habit has crowded out native marsh plants. The plants were likely introduced accidentally in packing material (Frenkel and Boss, 1988). Management of mid and high marsh invasive species requires mechanical removal of plants, an increasingly difficult process as the plant population expands. The vegetation of the upper edge of coastal wetlands, where saltwater grades into freshwater, is particularly prone to alteration through changes in hydrology. As mentioned earlier, invasions of common reed along the east coast and cattails in southern California have been associated with reduction in salinities (Niering and Warren, 1980; Roman et al., 1984; Beare and Zedler, 1987). Mitigation requires restoration of the historical tidal exchange by removing barriers and, where fill has been deposited, regrading the marsh to elevations suitable for salt marsh species. Because common reed tolerates a wide range of salinities up to about 25 ppt, mechanical removal of this species may be required to allow reestablishment of salt marsh species. Although salt marsh managers often view common reed-dominated habitats as less valuable to wildlife than the native salt marsh (Roman et al., 1984), it is important to consider the context in which this plant occurs within a mixed landscape. In urban marshes, dense stands of common reed adjacent to salt marshes may provide the only cover for wildlife species in an otherwise exposed environment. The common reed-dominated zone of Belle Isle Marsh in metropolitan Boston, MA was the part of the marsh where muskrats (Ondatra zibethica) and red-winged blackbirds (Agelaius phoenicus) nested and black-crowned night herons (Nycticorax nycticorax) roosted (Buchsbaum and Hall 1990). The American bittern (Botaurus lentiginosus), an uncommon shy species, occurs primarily within the common reed zone at this marsh.

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FUTURE CONSIDERATIONS In the last 20 years, the amount of legal protection given coastal wetlands by the federal government and many coastal states has increased substantially. This, combined with an increase in our scientific understanding of the ecology of these habitats, offers hope for future protection and management. Nevertheless, the largescale destruction of marshes for development is being replaced by other potential threats. Mosquito control will continue to be a major factor in the management of salt marshes and mangrove swamps. Therefore, evaluation of the effects of OMWM and related procedures on various marsh processes is still required. The projected expanding human population along the coast (Culliton et al., 1990) will continue to place pressure on coastal wetlands even when the actual wetland is protected. More people means development of buffer areas and surrounding watersheds, incremental losses of small pieces of coastal habitats, and increases in recreational-related structures and activities that impact coastal wetlands. Management and mismanagement of coastal wetlands in the past and present have occurred on a local scale in which decisions have focused on protecting (or not protecting) individual wetlands. The most significant future management issues will be related to global and regional changes that occur on a scale beyond which there is any precedent for management. The transformation of marshes to deepwater habitats in Louisiana due to shoreline subsidence, the largest single cause of recent vegetated wetland loss in the United States, is not a problem amenable to traditional localized solutions. Similarly, addressing the threat to coastal wetlands caused by projected rising sea levels will require global cooperation.

REFERENCES Abernathy, A. R., Zirschy, J., and Borup, M. P., Overland flow wastewater treatment at Easley, S.C., J. Water Poll. Control Fed., 57, 291, 1985. Allen, A. W., Habitat Suitability Models: Mink, Revised, U.S. Fish and Wildlife Service Biological Report 82(10.127), 1986. Allen, A. W. and Hoffman, R. D., Habitat Suitability Index Models: Muskrat, U.S. Department of the Interior Fish and Wildlife Service, FWS/OBS-82/10.46, 1984. Balling, S. S. and Resh, V. H., Arthropod community response to mosquito control recirculation ditches in San Francisco Bay salt marshes, Environ. Entomol., 11, 801, 1982. Balling, S. S. and Resh, V. H., The influence of mosquito control recirculation ditches on plant biomass, production, and composition in two San Francisco Bay salt marshes, Estuar. Coastal Shelf Sci., 16, 151, 1983. Beare, P. A. and Zedler, J. B., Cattail invasion and persistence in a coastal salt marsh: the role of salinity reduction, Estuaries, 10, 165, 1987. Bochman, A., Buffer Zones and Water Quality in the Parker River/ Essex Bay Area of Critical Environmental Concern: A Correlational Study, Unpublished M.S. thesis, Harvard University Extension School, 1991. Bourne, W. S. and Cottam, C., Some Biological Effects of Ditching Tidewater Marshes, Research Report 19, U.S. Fish and Wildlife Service, 1950.

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Brady, P. and Buchsbaum, R., Buffer Zones: The Environment’s Last Defense, Massachusetts Audubon Society, Wenham, MA, 1989. Brown, M. and Shaeffer, J., Buffer Zones for Wetlands and Wildlife in East Central Florida, University of Florida Center for Wetlands, Publication 89–07, Gainesville, 1990. Brush, T., Lent, R. A., Hruby, T., Harrington, B. A., Marshall, R. M., and Montgomery, W. G., Habitat use by salt marsh birds and response to open marsh water management, Colonial Waterbirds, 9, 189, 1986. Buchsbaum, R. and Hall, M., Baseline Assessment of Belle Isle Marsh in Anticipation of Restoration, Report to the Massachusetts Environmental Trust, 1992. Burger, J. and Shisler, J., Succession and productivity on perturbed and natural Spartina salt marsh areas in New Jersey, Estuaries, 6, 50, 1983. Calloway, J. C. and Josselyn, M., Introduction and spread of Spartina alterniflora in south San Francisco Bay, Estuaries, 15, 218, 1992. Castelle, A. J., Conolly, C., Emers, M., Metz, E. D., Meyer, S., and Witter, M., Wetlands Buffers: An Annotated Bibliography, Publication No. 92–10, Adolfson Associates, prepared for Shorelands and Coastal Zone Management Program Washington State Department of Ecology Olympia, 1992a. Castelle, A. J., Conolly, C., Emers, M., Metz, E. D., Meyer, S., and Witter, M., Mauermann, S., Erickson, T., and Cooke, S. S., Wetlands Buffers: Uses and Effectiveness, Publication No. 92–10, Adolfson Associates, prepared for Shorelands and Coastal Zone Management Program Washington State Department of Ecology Olympia, 1992b. Clarke, J. A., Harrington, B. A., Hruby, T., and Wasserman, F., The effect of ditching for mosquito control on salt marsh use by birds in Rowley, MA, J. Field Ornithol., 55, 160, 1984. Culliton, T. J., Warren, M. A., Goodspeed, T. R., Remer, D. G., Blackwell, C. M., and McDonough, III, J. J., Second Report of Coastal Trends: 50 Years of Population Change along the Nation’s Coast, 1960–2010, U.S. Department of Commerce, National Oceanic and Atmospheric Administration, NOAA Strategic Assessment Branch, Rockville, MD, 1990. Dahl, T. E., Wetlands Losses in the United States 1780’s to 1980’s, U.S. Department of the Interior, Fish and Wildlife Service, Washington, D.C., 1990. DeGraaf, R. M. and Rudis, D. D., New England Wildlife: Habitat, Natural History, and Distribution. NE Forest Experiment Station General Technical Report NE-108, U.S. Department of Agriculture and Forest Service, Broomall, PA, 1986. DeLaune, R. D., Whitcomb, J. H., Patrick, W. H., Perdue, J. H., and Pezeshki, S. R., Accretion and canal impacts in a rapidly subsiding wetland. I. 137Cs and 210Pb techniques, Estuaries, 12, 247, 1989. Diamond, R. S. and Nilson, D. J., Buffer delineation method for coastal wetlands of New Jersey, Symp. Coast. Water Res. Amer. Water Res. Assoc., May 1988. Dzierzeski, M., Vegetation Changes in a Hydrologically Altered Salt Marsh Ecosystem, Unpublished Masters thesis, University of New Hampshire, Durhham, 1991. Ehrenfeld, J. G., The role of woody vegetation in preventing groundwater pollution by nitrogen from septic tank leachate, Water Res., 21, 605, 1987. Executive Office of Environmental Affairs, Report of the Living Resources Committee of the Technical Advisory Group on Marine Affairs, Commonwealth of Massachusetts, Executive Office of Environmental Affairs, 1991. Ferrigno, F., Slavin, P., and Jobbins, D. M., Salt marsh water management for mosquito control, in Proceedings of the 62nd Annual Meeting of the New Jersey Mosquito Control Association, 1975, 32.

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Field, D. W., Reyer, A. J., Genovese, P. V., and Shearer, B. D., Coastal Wetlands of the United States: Accounting of a Valuable National Resource, Strategic Assessment Branch, National Oceanic and Atmospheric Administration and U.S. Fish and Wildlife Service, 1991. Florida Department of Natural Resources, Florida Salt Marshes, 1992a. Florida Department of Natural Resources, Florida’s Mangroves, 1992b. Frayer, W. E., Monahan, T. J., Bowden, D. C., and Graybill, F. A., Status and Trends of Wetlands and Deepwater Habitats in the Coterminous United States, Department of Forest and Wood Science, Colorado State University, Ft. Collins, 1983. Frenkel, R. E. and Boss, T. R., Introduction, establishment and spread of Spartina patens on Cox Island, Siuslaw Estuary, Oregon, Wetlands, 8, 33, 1988. Hagedorn, C., Microbiological aspects of groundwater pollution due to septic tanks, in Groundwater Pollution Microbiology, Bitton, G. and Gerba, C. P., Eds., John Wiley & Sons, New York, 1984, 181. Hemond, H. F. and Benoit, J., Cumulative impacts on water quality functions of wetlands, Environ. Manage., 12, 639, 1988. Heufelder, G., Bacterial monitoring in Buzzards Bay, EPA 503/4-88-001, U.S. Environmental Protection Agency, Region 1, Boston, MA, 1988. Howes, B., Howarth, R., Teal, J. M., and Valiela, I., Oxidation-reduction potentials in a salt marsh. Spatial patterns and interactions with primary production, Limnol. Oceanogr., 26, 350, 1981. Hruby, T., Montgomery, W. G., Lent, R. A., and Dobson, N., Open marsh water management in Massachusetts: adapting the technique to local conditions and its impact on mosquito larvae during the first season, J. Am. Mosquito Contr. Assoc., 1, 85, 1985. Hruby, T. and Montgomery, W. G., Open Marsh Water Management Manual for Open Tidal Marshes in the Northeast, Massachusetts Audubon Society, Ecosystems Management Series, 1986. Jensen, A., The effect of cattle and sheep grazing on salt marsh vegetation at Skallingen, Denmark, Vegetatio, 60, 37, 1985. Jordan, T. E., Correll, D. L., Peterjohn, W. T., and Weller, D. E., Nutrient flux in a landscape: the Rhodes River watershed and receiving waters, in Watershed Research Perspectives, Correll, D., Ed., Smithsonian Environmental Research Center, Edgewater, MD, 1986. Keswick, B. and Gerba, C., Viruses in groundwater, Environ. Sci. Technol., 14, 1290, 1980. Kirby, R. E., American Black Duck Breeding Habitat Enhancement in the Northeastern United States: A Review and Synthesis, U.S. Fish and Wildlife Service Biological Report 88(4), 1988. Knight, R. L., McKim, T. W., and Kohl, H. R., Performance of a natural wetland system for wastewater management, J. Water Poll. Control. Fed., 59, 746, 1987. Lacey, L. A. and Undeen, A. H., Microbial control of mosquitoes and blackflies, Ann. Rev. Entomol., 11, 265, 1986. Lawrence, R. R., Leonard, R., and Sheridan, J. M., Managing riparian ecosystems to control nonpoint pollution, J. Soil Water Conserv., 40, 871, 1985. Lawrence, R. R., Sharpe, J. K., and Sheridan, J. M., Long term sediment deposition in the riparian zone of a coastal plain watershed, J. Soil Water Conserv., 41, 266, 1985. Lesser, C. R., Murphy, F. J., and Lake, R. W., Some effects of grid system mosquito control ditching on salt marsh biota in Delaware, Mosquito News, 36, 69, 1976. Massachusetts Office of Coastal Zone Management, Guidelines for Dock and Pier Construction in ACECs and Ocean Sanctuaries, Massachusetts Coastal Zone Management, Executive Office of Environmental Affairs, Boston, MA, 1988.

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Mendellsohn, I. A., McKee, K. L., and Patrick, W. H., Oxygen deficiency in Spartina alterniflora roots: metabolic adaptations to anoxia, Science, 214, 439, 1981. Mendelssohn, I. A. and McKee, K. L., Spartina alterniflora die-back in Louisiana: time course investigation of soil waterlogging effects, J. Ecol., 76, 509, 1988. Miller, W. B. and Egler, F. E., Vegetation of the Weqetequock-Pawcatuck tidal-marshes, Conn. Ecol. Monogr., 20, 143, 1950. Mitchell, R. and Chamberlain, C., Survival of indicator organisms, in Indicators of Viruses in Water and Food, Berg, G., Ed., Ann Arbor Science, Ann Arbor, MI, 1978, 15. Montague, C. L., Zale, A. V., and Percival, H. F., Ecological effects of coastal marsh impoundments: a review, Environ. Manage., 11, 743, 1987. Niering, W. A. and Scott Warren, R., Vegetation patterns and processes in New England salt marshes, BioScience, 30, 301, 1980. Nixon, S., The Ecology of New England High Salt Marshes, FWS/OBS-81–55, U.S. Fish and Wildlife Service, Office of Biological Service, Washington, D.C., 1982. Ofiara, D. D. and Allison, J. R., A comparison of alternative mosquito abatement methods using benefit-cost analysis, J. Am. Mosquito Contr. Assoc., 2, 522, 1986. Phillips, J. D. and Phillips, L. R., Delineation of shoreline buffer zones for stormwater pollution control, Symp. on Coastal Water Res. Amer. Water Res. Assoc., 1988. Portnoy, J. W., Report on the Mosquitoes of the Herring River Basin, Cape Cod National Seashore, 1982, Report No. 12, U.S. Department of Interior, National Park Service, North Atlantic Region Water Resource Program, 1982. Portnoy, J. W., Summer oxygen depletion in a diked New England estuary, Estuaries, 14, 122, 1991. Provost, M. W., Source reduction in salt marsh mosquito control: past and future, Mosquito News, 37, 689, 1977. Resh, V. H. and Balling, S. S., Tidal circulation alteration for salt marsh mosquito control, Environ. Manage., 7, 77, 1983. Rhode Island Department of Environmental Management Coastal Resource Management Plan, Office of Environmental Coordination, Providence, RI, 1984. Rogers, J., Syz, S., and Golden, F., Maryland Upland Natural Areas Study, Vol. 1, Eastern Shore, Prepared for Maryland Coastal Zone Management Program, Annapolis, MD, 1975. Roman, C. T. and Good, R., Buffer delineation model for New Jersey pinelands wetlands, Division of Pinelands Research Center for Coastal and Environmental Studies, Rutgers University, New Brunswick, NJ, 1985. Roman, C. T., Niering, W. A., and Warren, R. S., Salt marsh vegetation change in response to tidal restriction, Environ. Manage., 8, 141, 1984. Rozas, L. P., McIvor, C. C., and Odum, W. E., Intertidal rivulets and creekbanks: corridors between tidal creeks and marshes, Mar. Ecol. Prog. Ser., 47, 303, 1988. Shisler, J. and Jobbins, D. M., Aspects of biological productivity in mosquito ditched salt marshes, Proc. NJ Mosquito Exterm. Assoc., 62, 48, 1975. Shisler, J., Lesser, F. H., and Schultze, T. L., Re-evaluation of some effects of water management on the Mispillion Marsh, Kent County, Delaware, Proc. NJ Mosquito Exterm. Assoc., 66, 276, 1975. Shisler, J., Schultze, T. L., and Howes, B. I., The effect of the marsh elder (Iva frutescens) on the standing crop biomass of Spartina patens and associated wildlife, Biol. Conserv., 14, 159, 1978. Short, H. J. and Cooper, R. J., Habitat suitability index models: great blue heron, U.S. Fish Wildlife Service Biological Report, 82(10.99), 1985.

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Sinicrope, T. L., Hine, P. G., Warren, R. S., and Niering, W. A., Restoration of an impounded salt marsh in New England, Estuaries, 13, 25, 1990. Slavin, P. and Shisler, J. K., Avian utilization of a tidally restored salt hay farm, Biol. Conserv., 26, 271, 1983. Tiner, R. W., Recent changes in estuarine wetlands of the coterminous United States, Coastal Wetlands Coastal Zone 1991 Conference—ASCE, 1991, 100. Turner, R. E. and Rao, Y. S., Relationships between wetlands fragmentation and recent hydrologic changes in a deltaic coast, Estuaries, 13, 272, 1990. Urner, C. A., Relation of mosquito control in New Jersey to bird life of the salt marshes, Proc. 22nd Annu. Meet. New Jersey Mosquito Exterm. Assoc., 130, 1935. Valiela, I. and Costa, J., Eutrophication of Buttermilk Bay, a Cape Cod coastal embayment: concentrations of nutrients and watershed nutrient budgets, Environ. Manage., 12, 539, 1988. Valiela, I., Teal, J. M., and Deuser, W. G., The nature of growth forms of the salt marsh grass, Spartina alterniflora, Am. Nat., 112, 461, 1978. Ward, G. and Fitzgerald, G. J., Fish predation on the macrobenthos of tidal salt marsh pools, Can. J. Zool., 61, 1358, 1983. Watzin, M. C. and Gosselink, J. G., The Fragile Fringe: Coastal Wetlands of the Continental United States, Louisiana Sea Grant College Program, Louisiana State University, Baton Rouge, LA, U.S. Fish and Wildlife Service, Washington, D.C., and National Oceanic and Atmospheric Administration, Rockville, MD, 1992. Weissberg, S. B. and Lotrich, V. A., Importance of an infrequently flooded intertidal marsh surface as an energy source for the mummichog, Fundulus heteroclitus: an experimental approach, Mar. Biol., 66, 307, 1982. Woodwell, G. M., Recycling sewage through plant communities, Am. Sci., 65, 556, 1977. Zedler, J. B., Winfield, T., and Williams, P., Salt marsh productivity with natural and altered tidal circulation, Oecologia, 44, 236, 1980.

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Kent, Donald M. “Watershed Management” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

12

Watershed Management Donald M. Kent

CONTENTS Managing Watersheds Elements of Management Definition and Delineation Watershed Characterization Prioritization Developing and Implementing a Watershed Program Monitor and Adjust Source Control Municipal Wastewater Best Management Practices (BMPs) Agricultural BMPs Urban Stormwater Runoff BMPs Innovative Solutions Watershed-Based Trading Case Study—the Chesapeake Bay Watershed Anacostia Watershed Restoration References

Wetlands tend to occupy topographic low points in the landscape and are thus recipient of water and eroded materials from higher in the landscape. The influx of water and other materials gives each wetland its character, supports its internal processes, and in part determines wetland function and value. In meager or excess amounts, water and other materials may alter or hinder wetland processes and

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diminish functions and values. Therefore, effective wetland management requires management of parts of the landscape contributing water and other materials to wetlands. The contributing areas of the landscape constitute the watershed. The concept of managing at the scale of watersheds has been evolving in the United States for about 100 years. In the 1890s, the U.S. Inland Waterways Commission recommended to Congress that each river system be treated as an integrated system. Throughout the first half of the 20th century, the focus of watershed management was on the use of water resources for energy, navigation, flood control, irrigation, and drinking water (U.S. Environmental Protection Agency, 1995b). In 1944, the Pick-Sloan Plan proposed to reduce flood damage by constructing large dams. The plan was opposed by some that believed that more effective flood control could be accomplished by managing rural, upstream watersheds than by constructing large dams (Peterson, 1998). During the 1950s and 1960s the management emphasis shifted to protecting drinking water (U.S. Environmental Protection Agency, 1995b). The Federal Water Pollution Control Act of 1956 funded publicly owned treatment works, and the Water Quality Act of 1965 required states to develop standards for interstate waters. The Clean Water Act and Safe Drinking Water Act of the 1970s and 1980s further emphasized large-scale protection of water resources. The Clean Water Act established a permitting program for point source polluters, provided additional funding for wastewater treatment and state water quality programs, and authorized programs to reduce, prevent, and eliminate pollution to surface and ground waters. The Safe Drinking Water Act established the basis for protecting surface and ground water supplies with an emphasis on preventing contamination. In recent years, the focus of water quality management has shifted to include nonpoint sources of pollution. Watershed management provides a necessary framework for managing nonpoint pollution. As a result, the U.S. Environmental Protection Agency developed the Watershed Protection Approach (1995a). Through focus on hydrologically defined resource areas, rather than jurisdictional boundaries, the Watershed Protection Approach is designed to more effectively protect and restore aquatic resources and protect human health than the historical approaches. The Approach targets priority problems, involves stakeholders, seeks integrated solutions, and measures success.

MANAGING WATERSHEDS A watershed is technically a divide separating one drainage area from another (Chow, 1964). More commonly, and as applied to watershed management, watersheds are areas that drain to surface water bodies. Watersheds come in all shapes, and range in size from a few to several million km2. Depending upon the type and extent of water quality problems, administrative boundaries, and technical constraints, watershed management may be applied to local watersheds, major watersheds, river basins, aquifers, or composites of surface watersheds and aquifers. From a water quality standpoint, watersheds have two elements. Terrestrial habitats, including urban, suburban, and rural areas, are the sources of particulate ©2001 CRC Press LLC

and dissolved materials. Particulate and dissolved materials derive from wastewater discharges, stormwater runoff, and erosion. The other element, surface water bodies including streams, rivers, ponds, lakes, estuaries, and coastal habitats, are the receptacles for particulate and dissolved materials. Materials may become trapped in the receiving water body or be transported downstream. Watershed management attempts to sustain and improve water quality by focusing on hydrologically defined resource areas. This is in contrast to historical efforts to regulate individual point sources of pollution. Watershed management also integrates various efforts to manage nonpoint sources of pollution. A fundamental premise of watershed management is that water quality and ecosystem issues can be more effectively addressed at the watershed level than at the level of the individual waterbody or polluter (U.S. Environmental Protection Agency, 1995a). Because watershed management addresses both point and nonpoint sources of pollution, it is an effective mechanism for protecting water and habitat quality. Several benefits, all of which save time and money, derive from watershed management’s holistic approach (U.S. Environmental Protection Agency, 1995a, b, 1996a). Regulatory efficiency is enhanced by coordinated monitoring, shared responsibility for assessment, and consolidated permitting. Decision making is improved by consideration of all stressors affecting water quality, systematic review of watershed basins, an increase in the availability and level of detail of watershed information, and a pooling of resources. An enlarged information base, systematic review, and enhanced coordination improve targeting of resources; and resources are focused on environmental results rather than programmatic activities such as permitting and reporting. Finally, innovative solutions are encouraged by watershed management, including ecological restoration, protection of critical areas, wetland mitigation banking, and watershed-based trading. Inherent to a successful watershed management program is stakeholder involvement. Stakeholders are individuals and organizations that are affected by water quality management decisions. This includes state and federal agencies charged with protecting water quality, businesses that rely on water or discharge waste, and citizens that use waterbodies and waterways for drinking water or recreation. Stakeholders share responsibility for monitoring, setting priorities, and developing and implementing management strategies. Elements of Management Watershed management has five elements (see Figure 1): 1. 2. 3. 4. 5.

Definition and delineation Characterization Prioritization Program development and implementation Monitoring and adjustment

Each of these elements will now be discussed briefly.

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Figure 1

Elements of a watershed management program.

In addition, watershed management requires development of a project team and public support. The former may include local, state, regional, and federal regulating agencies, research scientists, policymakers, trade associations representative of pollution sources, and nongovernmental organizations. The composition of the project team will vary with geographic scope and institutional infrastructure. Public support is important for developing applicable management goals, encouraging ©2001 CRC Press LLC

cooperation among disparate project team members, implementing management actions, and monitoring success. Definition and Delineation Definition and delineation are the selection of management boundaries. Management boundaries may encompass local watersheds, groups of local watersheds, river basins, aquifers, or some combination of watersheds, basins, and aquifers. Ideally, management boundaries should be large enough to benefit from an economy of scale, take advantage of government and technical expertise, and yet be manageable for the long term (U.S. Environmental Protection Agency, 1995a). As mentioned above, the boundary will in practice reflect the type and extent of the water quality issues and administrative boundaries. Nested watersheds, where small watersheds are subsets of larger watersheds, facilitate management at multiple scales (U.S. Environmental Protection Agency, 1995b). For example, local stakeholders can manage local watersheds, while state or regional entities can manage river basins. Watershed Characterization The watershed should be characterized after the management boundary has been defined and delineated. The purpose of characterization is to describe the physical characteristics of the watershed, to determine the water quality status and trends of watershed waters, and to identify potential water quality stressors and their sources. The physical description of the watershed should include geology, topography, soils, land use, hydrology, and significant biological resources. The latter may include threatened and endangered species and critical habitat. Surface water bodies should be described with respect to their designated uses and physicochemical and biological water quality. A baseline water quality monitoring program will need to be established if existing information is inadequate. Ideally, the baseline program will include physical, chemical, and biological indicators of water condition (see Chapter 8). Potential point (e.g., wastewater treatment facilities, industrial discharges) and nonpoint (e.g., urban stormwater, agricultural runoff) sources of pollution should be described by location, type, and absolute and relative loadings to the receiving body (Table 1). Rarely does one source or one type of pollution cause a problem. Existing control measures should also be described. Projecting expected watershed demographics and land use as they relate to potential sources of pollutants is also helpful at this stage. Prioritization Watershed characterization may identify few issues, and available resources may be sufficient to effect comprehensive management. More likely, the extent and degree of watershed issues will exceed the resources expected to be available for management. In such instances, watershed goals, targets, and action items must be prioritized. Prioritization may be logically directed at individual waterbodies or ©2001 CRC Press LLC

Table 1

Water Quality Stressors Typically Associated with Land Uses and Land Use Activities

Land Use or Activity Agriculture

Construction Forestry Golf courses Impoundments Industrial discharge Mining Septic systems Urban runoff

Wastewater treatment facility

Stressor Sediment Nutrients Bacteria Pesticides Sediment Sediment Nutrients Pesticides Altered hydrology Inorganic and organic chemicals Metals Sediments Metals Nutrients Bacteria Sediment Nutrients Bacteria Pesticides Altered hydrology Metals Nutrients Bacteria

waterways within the watershed (Table 2). Alternatively, specific pollutants or pollutant sources could be prioritized. Water quality impairments that pose a risk to public health should receive top priority and be addressed as quickly as possible. Other policy-related criteria include water quality goals, designated water uses, and waterbody or waterway value. These criteria are related when the waterbody or waterway is used for drinking water, commercial fishing, or recreation. Waters with more stringent water quality goals, greater designated uses, and higher value might reasonably receive high priority. Table 2

Criteria for Prioritizing Watershed Management Efforts Directed at Improving Waterbody and Waterway Water Quality

Degree of waterbody/waterway impairment Designated use of the waterbody/waterway Knowledge about water quality, stressors, and sources Probability of success Resources available for management Risk to ecosystem health Risk to public health Stakeholder support Type of waterbody/waterway impairment Value of the waterbody/waterway Water quality goals for the waterbody/waterway

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Programmatic criteria, including knowledge about watershed waters, resources available for management, stakeholder support, and probability of success, also impact the implementation of management actions. Insufficient knowledge about the watershed will require a return to the characterization stage. Alternatively, insufficient knowledge about individual waters may eliminate their consideration from the management process. In the absence of sufficient resources, some goals, targets, and action items may have to be eliminated. Lack of stakeholder support may necessitate initiation of an education program and postponement of actions. Conversely, projects with stakeholder support will be easier to implement. Goals, targets, and action items with a high probability of success are important at the beginning of a watershed management program to demonstrate program effectiveness to stakeholders. The type and degree of water quality impairment and ecosystem health relate directly to the physical, chemical, and biological character of the waterbody or waterway. Waterbody and waterway water quality can be compared against regulatory or designated use standards or against the minimum requirements of aquatic organisms such as fish. Reviewing plant and animal richness and diversity can assess ecosystem health. Systems with impaired water quality or poor ecosystem health may be priorities. As noted above, the threat to public health will be the superceding criteria for prioritization. In the absence of a public health risk, other criteria may become superceding based upon local or regional policy concerns, programmatic constraints, or stakeholder interest. Nevertheless, particularly in the early stages of a watershed management program, formalized evaluation of assorted criteria facilitates consideration of multiple perspectives, flexible problem solving, and stakeholder support. This approach also provides a basis for reevaluating a goal, target, or action item if circumstances change. A matrix analogous to the site selection criteria matrix illustrated in Chapter 5 could be used to ensure careful consideration of all issues. Developing and Implementing a Watershed Program Developing and implementing a watershed management program requires knowledge of the type and degree of water quality problems, the source of the problems, and the available and achievable solutions. This was achieved in the characterization stage. The prioritization stage helped determine the sequence of management actions. This stage has two components: program development and program implementation. Program development focuses on defining a strategy for improving watershed water quality. This is accomplished by setting management goals, targets, and action items. Goals are long-term visions of the watershed and may be programmatic, activity-based, centered on best management practices (BMPs) installation, water quality-oriented, or biological. An example goal might be stating that all surface waters will support commercial and recreational fisheries by the year 2010. Setting of additional near-term or interim goals may facilitate continued stakeholder support and a sense that progress is being made toward long-term goals. Goals are supported by targets, which are specific, quantifiable objectives. For example, reducing nutrient ©2001 CRC Press LLC

loads by 50 percent and restoring historical riparian vegetation will restore commercial and recreational fisheries. Finally, action items ensure that goals and targets will be achieved. Action items are specific projects with assigned roles and responsibilities and a scheduled completion date. For example, the local chapter of the ecological restoration society will restore bank vegetation along a 1 km stretch of the headwater stream extending from point A to point B, beginning May 1, 2000 and completing the restoration by June 30, 2000. Together, the goals, targets, and action items will be a mix of local and watershedwide regulations, management practices, economic incentives, and education and training programs. Again, one of the benefits of watershed management is the opportunity for innovative solutions, such as pollution trading (discussed later in this chapter), ecological restoration (see Chapter 6), and mitigation banking (see Chapter 7). Installation of controls should be site specific and tailored to hydrology, topography, geology, the resource to be protected, and politics. Documentation in the form of a watershed management plan is fundamental to program development. The plan should describe the watershed, characterize water quality and pollutant sources, list priorities, and describe the process leading to setting of goals, targets, and action items. In addition, the plan should define roles and responsibilities, identify funding sources and mechanisms, establish a schedule, and describe how program effectiveness will be assessed. Documenting development of the watershed management program facilitates reevaluation, clarifies intent and the decision-making process, and serves as a reference for future management. The plan should be periodically updated. Program implementation requires reaching consensus on goals, targets, and action items, developing an organizational infrastructure for effecting controls, and establishing procedures. Consensus is facilitated by stakeholder involvement in watershed definition and delineation, characterization, prioritization, and program development. An organizational infrastructure must carry out management actions, account for funds, maintain the schedule, and communicate to stakeholders. Controls must be properly installed and subject to periodic inspection and maintenance. Effective actions should be documented as procedures and become part of the watershed plan. Successful watershed management programs will secure commitments for funding and installation and management of controls. Commitments should come from both those implementing and administering actions and from those installing controls. Commitments may be formal or rely on public accountability (U.S. Environmental Protection Agency, 1995a). The former are written and detail expectations for all parties. The latter provide for public review through meetings or publications. Funding may derive from the operating budgets of participating organizations, businesses, municipal bonds, taxes, grants from nonparticipating organizations, donations, or fees. Additional support may come from in-kind contributions. Large or complex watershed management programs may benefit from a funding schedule. The schedule would reflect potential funding sources, application dates, dates funding is required, and tasks to obtain funding (U.S. Environmental Protection Agency, 1995a). Ultimately, successful programs have multiple incentives for stakeholder participation (Table 3, U.S. Environmental Protection Agency, 1995a). Stakeholders ©2001 CRC Press LLC

should be thoroughly educated about the reasons, goals, and progress of the watershed management program. Individuals responsible for implementing, installing, and maintaining pollution controls should receive adequate training and technical assistance. Individuals and businesses should be compensated for control costs that benefit society as a whole. Table 3

Incentives for Participating in a Watershed Management Program (U.S. Environmental Protection Agency, 1995a)

Incentive Cost–Share Education Purchase

Regulation Tax advantage Technical assistance

Description Payment to polluters for the installation of controls Including function and value of waterbodies and waterways; goals, targets, and action items; benefits of controls; and progress Purchase of critical areas including source water protection areas, riparian areas, critical habitat, lands from owners unwilling to institute controls Environmental laws and regulations, zoning ordinances, use restrictions, performance standards Conservation easements, credits for installation of controls Installation of controls, training of on-site managers, provision of procedural documents

Monitor and Adjust Ideally, monitoring will have been effected prior to the implementation of any management actions to characterize the watershed and provide a baseline for comparison, and after the implementation of management actions, monitoring documents the effectiveness, or ineffectiveness, of the watershed management program. Documented monitoring results also provide the basis for communicating with stakeholders and facilitate long-term maintenance of pollutant controls. Perhaps most importantly, monitoring provides a basis for making adjustments to the watershed management program. Adjustments will be necessary if management actions are partly or wholly ineffective at achieving program goals or targets. Program adjustments will also be necessary if management actions are effective; goals and targets must be reprioritized. Finally, monitoring provides a basis for making program adjustments in response to significant land-use changes. Monitoring plans should derive directly from program goals, targets, and action items. Continuing with the earlier example, monitoring of native fisheries might include direct counts of fish, preferably by age class. Depending upon program goals, monitoring may encompass biological, chemical, physical, and programmatic parameters (see Chapter 8). Table 4 lists parameters commonly monitored as part of a watershed management program. Chemical and physical parameters should be monitored routinely, as well as during storm events, to characterize the initial flush of pollutants. Biological parameters effectively may be monitored seasonally or annually. Voluntary citizen monitoring programs have become increasingly common in the United States. The success of these programs is dependent upon effective training and a good quality assurance/quality control program. ©2001 CRC Press LLC

Table 4

Parameters Likely To Be Monitored in a Watershed Management Program

Type Parameter Biological

Chemical

Physical

Programmatic

Benthic macroinvertebrate richness Biotic index Fish and wildlife abundance Fish and wildlife richness Vegetation cover or density Vegetation richness Biological oxygen demand Dissolved oxygen Nutrient concentration pH Toxicants Suspended solids Temperature Turbidity Enforcement actions Funds received and disbursed Meetings Permit issuance Reports

SOURCE CONTROL Municipal Wastewater Municipal wastewater contains suspended solids, biodegradable organics (e.g., proteins, carbohydrates, fats), pathogens, and nutrients such as nitrogen and phosphorus. Depending upon the service area, wastewater may also contain organic and inorganic carcinogens, mutagens, teratogens, acutely toxic compounds, pesticides, heavy metals, and dissolved organics. In the absence of high concentrations of the latter constituents, nutrients are the primary constituents of concern. Excessive nutrients discharged to aquatic environments increase the growth of undesirable plants and algae, decrease dissolved oxygen levels, and in some instances promote ammonia toxicity. In the early 20th century in the United States, wastewater was discharged directly to streams and rivers via storm sewers. The accumulation of sludge, odors, and other unsightly conditions led to the separation of storm drains and sewers, and the construction of wastewater treatment facilities. Initially, most treatment facilities provided only primary treatment, which consisted of screening and sedimentation to remove floating and settleable solids. Later, the U.S. Environmental Protection Agency mandated secondary treatment as the minimum standard for facilities. Secondary treatment involves biological and chemical processes to remove most of the organic matter. Treated wastewater was historically disposed of by the easiest method possible. For coastal communities, this may have included ocean discharge, a practice that is ©2001 CRC Press LLC

increasingly discouraged (Metcalf and Eddy, 1991). Away from the coast, discharge to inland surface waters is the most common method for disposing of treated wastewater. Surface discharge relies on the assimilative capacity of the receiving water, a capacity that has been increasingly exceeded for many waterways in the latter part of the 20th century. In response, many wastewater facilities are being required to provide advanced treatment. Advanced wastewater treatment removes additional suspended and dissolved substances, especially nitrogen and phosphorus. At conventional treatment facilities, advanced processes remove nitrogen by biological nitrification and denitrification, separate stage biological denitrification, airstripping, breakpoint chlorination, and ion exchange (Metcalf and Eddy, 1991). Phosphorus is removed by chemical precipitation with metal salts or lime, and filtration. Microorganisms can also be stressed to force additional phosphorus uptake. Large wastewater facilities exceed the treatment needs and financial resources of small communities. Clustered homes may use a package treatment facility. More typically, rural homes will use on-site treatment consisting of a septic tank and disposal field. BOD, SS, N, P, bacteria, and viruses are the primary constituents of concern with on-site disposal. Onsite systems should be set back from surface and ground waters, the distance of the setback contingent upon system capacity and soil permeability (Metcalf and Eddy, 1991). Schueler (1995) has noted that more than one on-site septic system per 2.8 ha can result in shellfish bed closures (Figure 2).

Figure 2

On-site septic systems located too close to coastal waters can result in shellfish bed closures.

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Natural treatment systems have many of the same treatment processes as conventional facilities (e.g., sedimentation, filtration) and have additional, unique treatment processes (e.g., photosynthesis, plant uptake). Land-based and wetland systems effect treatment of municipal wastewater. Both types of systems are preceded by mechanical pretreatment including fine screening and primary sedimentation. The three fundamental types of land treatment are slow rate, rapid infiltration, and overland flow (Metcalf and Eddy, 1991). Slow rate systems are potentially the most effective land treatment and entail the application of wastewater to vegetated land to provide treatment and irrigation. Wastewater is consumed by plants and is evapotranspirated. Treatment is effected in large part by wastewater percolation through the soil. Rapid infiltration systems entail the intermittent application of wastewater to shallow, unvegetated infiltration or spreading basins. As with slow rate systems, treatment occurs as wastewater percolates through the soil. Overland flow systems are relatively less effective than slow rate and rapid infiltration systems and are used in areas with relatively impermeable soils. Wastewater is distributed across the upper part of a graded, vegetated slope, and runoff is collected in ditches at the toe of the slope. Treatment is effected primarily by evapotranspiration. Wetland systems are inundated areas supporting aquatic vegetation (Kadlec and Knight, 1996). Filtration, sedimentation, precipitation, plant uptake, and other processes effect significant reduction and removal of wastewater constituents. Chapter 9 discusses wetland treatment systems at length. Best Management Practices (BMPs) Best management practices (BMPs) are operational procedures designed to reduce pollutant discharge to surface water or groundwater and to minimize changes to hydrology and hydraulics. BMPs reduce the pollutant load by reducing the volume of discharge water, reducing the concentration of pollutants in discharged water, or both. A watershed management program may include agricultural and urban BMPs (Tables 5 and 6). Agricultural BMPs Modern agricultural practices rely on fertilizers and pesticides to increase crop yield. Excess or misapplied fertilizer can cause algal blooms, stimulate growth of noxious plants, and decrease available oxygen for fish and other aquatic organisms. High concentrations of nitrogen may cause methemoglobinemia (see Chapter 5). Pesticides can be chronically or acutely toxic to humans and aquatic organisms. Agricultural practices may also be accompanied by excessive erosion. Sediment erosion increases surface water turbidity and may smother benthic organisms. Nutrients, pesticides, and heavy metals occur in particulate form or can be attached to dirt, sediment, and detritus. Sediment accumulation may also alter waterway hydrology and hydraulics by increasing flow velocity and decreasing flow capacity. Fertilizer BMPs operate by reducing the amount of fertilizer used and retaining unused fertilizer on-site (Bottcher et al., 1995; South Florida Water Management District, 1999). BMPs include soil chemistry management and calibrated soil ©2001 CRC Press LLC

Table 5

Best Management Practices (BMPs) for Agriculture

Fertilizer control Banding fertilizer Calibrated soil testing Cover crop On-farm retention of drainage water Soil chemistry management Split application Pesticide control Buffer zone Spill management Integrated pest management Mixing, loading, and washdown location Precise application Sediment control Bank contouring Bank stabilization Sediment traps and settling basins

Table 6

Best Management Practices (BMPs) for Urban Stormwater Runoff

Buffer zones Fertilizer management Green parking lots Land-use restrictions Limit soil disturbance Minimize impervious surface Stormwater retention and treatment

testing, banding and split application, on-site retention of drainage water, and buffer zones (Table 5, Figure 3). Soil chemistry management maintains the soil pH to maximize the availability of nutrients to plants, minimize nutrient leaching, and immobilize metals. Calibrated soil testing bases fertilizer recommendations on yield–response curves developed by correlating soil nutrient levels with crop yields. Banding places fertilizer in strips adjacent to plant roots and is most effective for crops without a continuous root mat. Split application is the practice of applying half the total amount of fertilizer semi-annually rather than all at one time. Buffer zones between crops and surface waters help prevent the misapplication of fertilizer to adjacent surface waters. Pesticide BMPs rely heavily on educating and training applicators (Florida Department of Agriculture and Consumer Services and the Florida Department of Environmental Protection, 1998). Mixing, loading, and equipment washdown locations should be permanent, consisting of an impermeable surface located close to the storage building. Impermanent mixing, loading, and washdown locations should be relocated frequently to prevent the accumulation of pesticides to toxic levels. Both permanent and impermanent facilities should be located away from surface ©2001 CRC Press LLC

Figure 3

Buffer zones between fields and surface waters minimize misapplication and filter runoff.

waters. As with fertilizers, establishment of buffer zones between crops and surface water will minimize the inadvertent application of pesticides to water resources. Establishing formal practices and procedures effects spill management. For example, containers are stored upright with tight, closed covers in sealed bottom, covered facilities. Any spills are contained with barriers and absorbent material. Precise application reduces the quantity of pesticide used on the crop. Techniques include controlled droplet technology, canopy-dimension spray machine discharge towers, drift control agents, spray calibration and maintenance, and minimized bandwidth. Pesticide quantities can be further reduced through integrated pest management (IPM). IPM encompasses a broad spectrum of practices that cumulatively minimize pests (Leslie, 1994; Florida Department of Agriculture and Consumer Services and the Florida Department of Environmental Protection, 1998). Key pests and beneficial organisms are identified, and cultural practices are implemented to minimize pests and enhance biological controls. Practices include soil preparation, crop rotation, use of resistant crop varieties, variable planting dates, modified irrigation, and cover crops. Beneficial organisms may also be augmented. Chemicals are applied only when pests are present. Sediment erosion can be minimized by increasing ditch sideslopes to reduce erosion potential, contouring the top of the bank away from surface waters, and stabilizing the bank (Bottcher et al., 1995; South Florida Water Management District, 1999). The latter can be achieved by using rock gabions at the water line, rip-rap, or establishing rooted plants. Sediment that escapes the site can be captured in settling traps and basins. Sediment traps are barriers placed in widened sections of ©2001 CRC Press LLC

ditches or canals. The traps slow water velocity, inducing the settling of particles. Settling basins are sumps in the bottom of ditches or canals that collect suspended particles. Accumulated sediments must routinely be removed from sediment traps and settling basins. Urban Stormwater Runoff BMPs In urban and suburban areas, runoff from impervious substrates is an important contributor to watershed degradation. Impervious cover increases stream peak discharge, velocity, and volume. The increase in flows widens streambanks and downcuts the streambed. In addition, impervious surfaces collect pollutants from the atmosphere, vehicles, and other sources which are, in turn, transferred to surface waters during storm events (Figure 4). Stormwater runoff constituents include nutrients, metals, hydrocarbons, bacteria, and viruses (Bingham, 1994). Stream degradation, including disruption of benthic communities and fisheries, occurs at 10 to 25 percent impervious cover (Hollis, 1975; Garie and McIntosh, 1986; Luchetti and Fuersteburg, 1993; Schueler, 1995). Minimizing lot setbacks, decreasing road widths, and narrowing or eliminating sidewalks can reduce the amount of impervious cover. Clustering development can reduce the amount of impervious surface by 10 to 50 percent, primarily by reducing roadways (Schueler, 1995).

Figure 4

Pollutants from the air, vehicles, and other sources are transferred from parking lots to surface waters during storm events.

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Runoff from impervious surfaces can be retained and treated using stormwater BMPs. Stormwater BMPs include in-line treatment methods like inlets, catchbasins, sump pits, and oil and grit separators (Schueler, 1987; England, 1997). However, in-line BMPs have limited storage capacity, and large storm events tend to resuspend trapped particles. Frequent maintenance is required. End of pipe BMPs include detention basins and retention ponds, vegetated filters, and filtration and infiltration devices. These BMPs are discussed at length in Chapter 5. The principles of impervious surface reduction and stormwater retention and treatment can also be extended to parking lots (Schueler, 1995). To reduce surface area, spaces can be allotted for compact cars. The use of porous pavement and interlocking pavers instead of asphalt and concrete increases infiltration and decreases runoff. Stormwater retention and treatment can be effected by directing surface runoff to specially designed areas within or adjacent to the parking lot (Bitter and Bowers, 1994). These areas filter or infiltrate runoff. Other methods for reducing watershed degradation from urban runoff include land use restrictions, buffer zones, limiting soil disturbance, and fertilizer management. Land use restrictions can prevent development in, or in close proximity to, critical areas such as streams, floodplains, riparian zones, shorelines, wetlands, and steep slopes. Buffer zones between critical areas and development offer additional protection by preventing inadvertent or indirect impacts. Buffer zones also provide filtering and infiltration of runoff and reduce disturbance to wildlife. Limiting soil disturbance on construction sites to the immediate work area, and implementing sedimentation and erosion controls, will minimize transport of soil and other particles to aquatic resources (see Chapter 5). Lawn and garden fertilizers can be managed in much the same way as described for agricultural areas to prevent eutrophication and methemoglobinemia.

INNOVATIVE SOLUTIONS Watershed management provides opportunities for innovative problem solving. Chief among these are habitat restoration, mitigation banking, and watershed-based trading. Wetland enhancement, restoration, and creation are the subject of Chapter 6, and mitigation banking is the subject of Chapter 7. Watershed-based trading is discussed next. Watershed-Based Trading Watershed-based trading is an exchange of effluent control responsibility between pollutant dischargers to achieve water quality objectives (U.S. Environmental Protection Agency, 1996b; Commonwealth of Virginia, 1996). A market-based approach, watershed-based trading enables a pollutant source with a high cost of control to purchase allowances from dischargers elsewhere in the watershed with a lower cost of control. Allowances are a quantity of effluent the discharger is allowed to release. The exchange of allowances does not increase overall effluent discharge in the watershed. ©2001 CRC Press LLC

Several potential benefits accrue from watershed-based trading to regulators, trading partners, and the community at large (U.S. Environmental Protection Agency, 1996b). Economically, the cost of pollutant control is reduced for individual dischargers by purchasing the least expensive option and taking advantage of economies of scale. As a direct corollary, the overall cost of managing water quality in the watershed is reduced. Dischargers that sell allowances reap direct financial benefits. Environmentally, watershed-based trading may achieve equal or greater water quality for the same or less cost, provide an incentive to polluters to go beyond the minimum pollutant reduction required, and encourage innovation. Innovation may address broader goals like conservation and preservation, ecological restoration, and endangered species protection. Socially, watershed-based trading encourages dialogue among stakeholders. Programmatically, watershed-based trading provides managers with a flexible approach for accelerating watershed-wide water quality improvement programs for point and nonpoint sources of pollution. Participants in a watershed-based trading program could include point source dischargers, indirect dischargers (i.e., industrial or commercial operations that discharge to a treatment facility), and nonpoint sources. Trading might occur intrafacility, between point source dischargers, between indirect dischargers, between nonpoint sources, or between point and nonpoint sources (Table 7). Fundamentally, a trading program requires a buyer to compensate the seller to reduce pollutant loads sufficiently to bring both facilities or land uses into compliance with discharge or water quality standards. The basis for the trade may be total maximum daily load or another expression of total effluent per unit of time that establishes a loading capacity for the defined area. Alternatively, the basis for the trade may be a point source permit. Public or private banks that buy and sell pollutant allowances may also effect trading. Table 7

Types of Watershed-Based Trading (U.S. Environmental Protection Agency, 1996b)

Intra-facility trading Point source to point source trading Pretreatment trading Nonpoint to nonpoint source trading Point to nonpoint source trading

A facility cost-effectively allocates pollutant discharges among outfalls A point source purchases an allowance from a second point source rather than reduce its own pollutant discharge An indirect discharger purchases an allowance from a second indirect discharger rather than increase its own pretreatment A nonpoint source polluter purchases an allowance from a second nonpoint source polluter rather than enhance its own control practices A point source purchases an allowance from a nonpoint source rather than reduce its own pollutant discharge

Trading systems may take one of three forms (Commonwealth of Virginia, 1996). Open trading systems represent a slight departure from the typical permit process by allowing regulated sources to modify permits to reflect an exchange of pollution control requirements. Allowances are only created when a source discharges less than the amount allowed under the permit. A closed trading system establishes an effluent discharge limit for a specified group of dischargers within a geographical ©2001 CRC Press LLC

area. Responsibility for effluent control is delegated to individual group members, and trading can only occur if total effluent discharge does not exceed the prescribed limit. A full closed system extends the closed system concept to all effluent discharge sources within the watershed. All point and nonpoint sources are assigned an initial allocation of allowances. As with closed systems, new pollutant sources are only permitted by acquiring existing allowances. Development of an effective watershed-based trading program requires consideration of several issues. Trading demand is created when discharge limits are constrained. When trading demand is created, there should be a clear transfer of financial and legal obligations. Trading is most effective when partners are close, as effluent distribution will shift with increasing distance. Nonpoint sources and best management practices are more difficult to quantify than point sources. Trades between point sources and nonpoint sources should include a trading ratio that favors the point source to compensate for this uncertainty. Monitoring of receiving waters, best management practices, and finances is essential.

CASE STUDY—THE CHESAPEAKE BAY WATERSHED Management of the Chesapeake Bay watershed illustrates many of the principles and practices discussed throughout this chapter. Chesapeake Bay is the largest estuary in the United States (Figure 5). The Bay is home to more than 2700 species of fish and wildlife, and is surrounded by 15 million people. A total of 48 major tributaries drain over 25,910 ha in Maryland, New York, Pennsylvania, Virginia, and West Virginia. The Bay is shallow, averaging only 8 m, and has a ratio of land area to water volume of 10 to 1. Exacerbated by shallow water and the large land-to-water ratio, the Chesapeake Bay’s decline was evident in the 1950s, but it was not until the 1970s that scientists attributed the decline in Bay water quality to three factors: excess nutrients, sediment, and toxic chemicals. Nutrients originated from domestic wastewater discharges and agricultural and stormwater runoff. Agricultural areas, construction sites, and erosion were the source of sediments. Toxic chemicals originated with businesses within the watershed. Initial efforts to reverse the decline in Chesapeake Bay water quality focused on upgrading wastewater treatment facilities. However, these efforts were insufficient to accomplish Bay restoration, leading to comprehensive efforts to control nonpoint sources of pollution. In 1983, the governors of Virginia, Maryland, Pennsylvania, the mayor of the District of Columbia, and the U.S. Environmental Protection Agency agreed to cooperate toward solving Chesapeake Bay water quality problems. In 1987, the Chesapeake Bay Executive Council, as it became known, established a goal of reducing nutrient input to the Bay by 40 percent from 1985 levels by the year 2000. Several interrelated programs, including the Chesapeake Bay Preservation Act and the Anacostia Restoration Agreement, have been developed to accomplish this goal. The Virginia General Assembly enacted the Chesapeake Bay Preservation Act in 1988 to establish a cooperative program between state and local governments to ©2001 CRC Press LLC

New York

Pennsylvania

West Virginia

Virginia

Figure 5

The Chesapeake Bay and Anacostia River (shaded area) watersheds.

reduce nonpoint source pollution. Inherent to the Act is an effort to balance economic interests and water quality concerns by requiring the use of resource management practices for environmentally sensitive lands. The Act establishes a relationship between local land use decisions and water quality protection by granting local governments the authority to manage water quality. With the exception of towns that drain directly to the Atlantic Ocean, all cities and counties bordering on tidal waters (i.e., Tidewater, VA) are required to comply with the Act. ©2001 CRC Press LLC

The Act established the Chesapeake Bay Local Assistance Board. The Board is comprised of nine individuals representing various locales and interests such as agriculture, environmental management, nonagricultural businesses, and government. According to the Act, the Board is charged with promulgating and maintaining regulations, providing technical and financial assistance to Tidewater governments, providing technical assistance and advice to regional and state agencies, and ensuring that local government plans and ordinances are in compliance with Act regulations. The Board is assisted by the Chesapeake Bay Local Assistance Department (1995), a state agency in the Secretariat of Natural Resources. The Department provides technical assistance and advice to local governments. Assistance includes administering a grants program, interpreting Act regulations, compliance reviews of comprehensive plans and ordinances, and review of private development plans. The Department also provides training for local planners and engineers. The Board promulgated Chesapeake Bay Preservation Area Designation and Management Regulations in 1989. The Regulations establish a framework for compliance and require local Tidewater governments to adopt a water quality program. Listed in the Regulations are 11 performance criteria (Table 8). Local programs, which tend to differ among localities, adopt or amend local land use plans and ordinances to incorporate water quality protection measures consistent with the Act. Program compliance has three phases. Table 8

Chesapeake Bay Preservation Area Designation and Management Regulations Performance Criteria (Chesapeake Bay Local Assistance Department, 1995)

Minimize impervious cover Minimize disturbed land Preserve existing vegetation Pump out septic tanks every 5 years and require 100 percent reserve drainfields for new development Erosion and sediment control for disturbances greater than 2500 ft2 No net increase in stormwater pollutant loadings for new development and 10 percent reduction in loadings for redevelopment Plan review for developmemt exceeding 2500 ft2 Agricultural conservation plans Forestry best management practices Evidence of wetland permits prior to clearing or grading Regular and periodic best management practice maintenance

Phase I objectives include determining the geographic and ecological extent of environmentally sensitive lands, mapping said lands, designating Chesapeake Bay Preservation Areas, and implementing water quality performance criteria. Chesapeake Bay Preservation Areas are those lands that have the potential to most directly impact water quality. These are lands that protect water quality, Resource Protection Areas (RPAs), and lands that could potentially damage water quality, Resource Management Areas (RMAs). RPAs are presumed to filter pollutants from runoff and include a 30 m landward buffer. Development within RPAs is restricted to water dependent projects, redevelopment, water wells, passive recreation, and historic and archeological activities. ©2001 CRC Press LLC

RMAs are contiguous with the inland boundary of RPAs. If improperly used or developed, RMAs have the potential to degrade water quality or otherwise damage RPAs. RMAs include lands with highly erodible soils or steep slopes, highly permeable soils, 100-year floodplains, and nontidal wetlands not included in RPAs. Development is permitted in RMAs in accordance with performance standards. Local governments can designate parts of RPAs and RMAs as Intensely Developed Areas under certain conditions. More than 50 percent of the land area must be covered by impervious surface, the land area must have public water and sewer, or the existing housing density must be 10 or more units per ha. The designation is intended to encourage redevelopment and infill activity rather than new development. Phase II objectives require local governments to adopt a Comprehensive Plan or Plan Amendment that incorporates water quality protection measures consistent with the Act. The Comprehensive Plan provides a policy framework for community development. Act regulations require that the comprehensive plan address physical constraints to development, protection of potable water supplies, shoreline erosion, access to waterfront areas, and redevelopment. Local governments may include other elements in the Comprehensive Plan. For example, Table 9 lists policies of the Fairfax County, Virginia Comprehensive Plan, which emphasize prevention of pollution from nonpoint sources (Fairfax County, 1990). Table 9

Policies of the Fairfax County, Virginia Comprehensive Plan Designed to Prevent and Reduce Pollution of Surface Waters

Implement a best management practice (BMP) program Update BMP requirements as more effective strategies become available Minimize impervious surfaces Minimize the application of fertilizers, pesticides, and herbicides to lawns and landscaped areas Preserve stream valleys when locating and designing stormwater dentention and BMP facilities Update erosion and sediment regulations; minimize grading Retrofit stormwater management ponds to become BMPs Monitor BMP performance Maintain high standards for discharges from point sources

Phase III objectives require local governments to adopt or revise a zoning ordinance (e.g., erosion and sediment control) that protects water quality consistent with the Act. Many local governments amend existing ordinances to encompass the 11 performance criteria listed in the Act. Phase III provides local governments with an opportunity to revisit the criteria and incorporate language specific to their local land use management program. An evaluation of the compatibility of Act regulations with local development standards also occurs in Phase III. Nutrient pollution in the watershed is declining, but additional efforts are required (Chesapeake Executive Council, 1996; Chesapeake Bay Program, 1999). Phosphorus loads to the bay were reduced by 2,721,500 kg per year between 1985 and 1997, and the 40 percent reduction goal is likely to be achieved. Nitrogen loads declined by 14,515,000 kg per year, but additional reductions are needed to achieve the 40 percent reduction goal. Gains in nonpoint source nitrogen reduction were offset by increases in point source nitrogen. Additional reductions to achieve the 40 percent goal may come from upgrades to wastewater treatment facilities, an option that was ©2001 CRC Press LLC

earlier rejected because of cost. Nutrient reductions have not markedly improved water clarity in the Bay. Other indicators suggest the program is having an impact. Many waters of the watershed that had been closed to fishing because of kepone contamination have been reopened. Industries within the watershed reduced chemical releases by 67 percent between 1988 and 1997. More than 700,000 ha of farmland were placed under nutrient management (i.e., comprehensive plans for efficient nutrient use) between 1985 and 1997. Restoration efforts have reforested 350 km of riparian zone, and fish passage construction and barrier removal have reopened 1000 km of spawning habitat. Anacostia Watershed Restoration The Anacostia River watershed is a critical area within the Chesapeake Bay Program and illustrates management approaches at a local level. The Anacostia River has a 70 ha watershed in the state of Maryland and the District of Columbia. Water quality problems in the Anacostia are largely attributed to combined sewer overflows, urban runoff, and erosion from construction activities and surface mining operations (Metropolitan Washington Council of Governments, 1990; Anacostia Restoration Team, 1991). The situation has been exacerbated by a 75 percent reduction in watershed forest cover. The State of Maryland, Montgomery and Prince George (Maryland) Counties, and the District of Columbia initiated the Anacostia Restoration Agreement in 1987. The Anacostia River Restoration Committee is the primary oversight group and is comprised of representatives from the aforementioned entities, County and District of Columbia departments, state and federal agencies, and nongovernmental organizations. Various policy and technical committees coordinate the participation of more than 60 different agencies. The Restoration Committee established broad water quality, biological, land use, and outreach goals for the Anacostia watershed (Table 10). The primary programmatic mechanism for accomplishing these goals is the development of Subwatershed Action Plans (SWAPs). SWAPs detail the schedule and location of watershed projects and are intended to streamline the approval of individual projects and define roles and responsibilities. Each SWAP will assess water quality and the aquatic community, define goals and targets, identify management opportunities, prioritize projects, and monitor results. In addition, each SWAP will develop plans to increase wetlands and forest cover within the subwatershed. Management actions are focused on implementation of basin-wide controls, stream restoration, and communicating with stakeholders (Metropolitan Washington Council of Governments, 1990). Basin-wide controls include abatement of combined sewer overflows, retrofitting of urban stormwater controls, new discharge restrictions on point sources of pollution, enhanced stormwater and sediment control regulations for development, and surface mine reclamation. Stream restoration efforts include the establishment of stream buffers, riparian restoration, streambank stabilization, and fish habitat enhancement. The progress of the basin-wide controls and stream

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Table 10

Goals of the Anacostia River Restoration Committee

Reduce pollutant loads in the tidal estuary by the turn of the century Enhance aquatic diversity and provide for an urban fishery Restore the spawning range of anadromous fish Increase the acreage of tidal and nontidal wetlands Expand the range of forest cover and create a contiguous corridor Make the public aware and increase volunteer participation Adapted from Metropolitan Washington Council of Governments, 1992. With permission.

restoration efforts is being assessed through baseline, performance, and storm event water quality sampling, and biological and habitat surveys. Considerable effort is being devoted to communicating watershed issues, project goals, and results to stakeholders (Metropolitan Washington Council of Governments, 1998). An annual report details the implementation of controls, restoration efforts, and monitoring results. A quarterly newsletter is devoted to citizen accomplishments and restoration activities. Subbasin educational documents have also been developed. In addition, subbasin coordinators promote public participation through slide presentations, stream walks, and clean-up efforts. The Anacostia Watershed Restoration Program continues to make progress toward its goals (Metropolitan Washington Council of Governments, 1998). Installation of a swirl concentrator facility has reduced floatable material and total phosphorus discharges from the largest combined sewer overflow by 25 to 30 percent. No fish kills have been reported in the river since 1992, and submerged aquatic vegetation has begun to reestablish itself in lower sections of the river. Stream restoration projects have been initiated and completed, and native fish reintroduced to part of the watershed have survived. Anadromous fish spawning habitat has been increased by 30 km by removal and modification of barriers to fish movement. Tidal and nontidal wetlands have been created, and amphibians have been restored to vernal pool habitats. More than 25,000 trees have been planted on 20 ha in support of riparian forest restoration. The Interstate Commission on the Potomac River Basin public outreach program has communicated to more than 60,000 people, and the Anacostia River Education Center was established by the District of Columbia and the Potomac Electric Power Company. Efforts continue to control stormwater runoff and high sediment loads and to expand recreational opportunities in the watershed.

REFERENCES Anacostia Restoration Team, A Commitment to Restore Our Home River: A Six Point Plan to Restore the Anacostia River, Metropolitan Washington Council of Governments, Washington, D.C., 1991. Bingham, D. R., Wetlands for stormwater treatment, in Applied Wetlands Science and Technology, Kent, D. M., Ed., Lewis Publishers, Boca Raton, FL, 1994, 243. Bitter, S. and Bowers, J., Bioretention as a water quality best management practice, Water Prot. Tech., 1(3), 114, 1994. ©2001 CRC Press LLC

Bottcher, A. B., Izuno, F. T., and Hanlon, E. A., Procedural Guide for the Development of Farm Level Best Management Practice Plans for Phosphorous Control in the Everglades Agricultural Area, Version 1.1, Circular 1777, University of Florida Cooperative Extension Service, 1995. Chesapeake Bay Local Assistance Department, A Guide to the Bay Act, 1995. Chesapeake Bay Program, The State of the Chesapeake Bay: A Report to the Citizens of the Bay Program, EPA 903-R99-013, CBP/TRS 222/108, Annapolis, MD, 1999. Chesapeake Executive Council, Commonwealth of Virginia Shenandoah and Potomac River Basins Tributary Nutrient Reduction Strategy, final comment draft, Virginia Secretary of Natural Resources, Chesapeake Bay Local Assistance Department, Department of Conservation and Recreation, Department of Environmental Quality, 1996. Chow, V. T., Handbook of Applied Hydrolog, McGraw-Hill, New York, 1964. Commonwealth of Virginia, Commonwealth of Virginia Shenandoah and Potomac River Basins Tributary Nutrient Reduction Strategy, Virginia Secretary of Natural Resources, Virginia Chesapeake Bay Local Assistance Department, Virginia Department of Conservation and Recreation, and Virginia Department of Environmental Quality, 1996. England, G., Stormwater sediment control using baffle boxes and inlet devices, in Proceedings of the Fifth Biennial Stormwater Research Conference, South Florida Water Management District, 1997, 142. Fairfax County, Policy Plan: The Countywide Policy Element of the Comprehensive Plan for Fairfax County, Virginia, 1990. Florida Department of Agriculture and Consumer Services and the Florida Department of Environmental Protection, Best Management Practices for Agrichemical Handling and Farm Equipment Maintenance, 1998. Garie, H. and McIntosh, A., Distribution of benthic macroinvertebrates, Water Res. Bull., 22, 447, 1986. Hollis, G., The effect of urbanization on floods of different recurrence intervals, Water Res. Res., 11(3), 431, 1975. Kadlec, R. H. and Knight, R. L., Treatment Wetlands, Lewis Publishers, Boca Raton, FL, 1996. Leslie, A. R., Ed., Integrated Pest Management for Turf and Ornamentals, Lewis Publishers, Boca Raton, FL, 1994. Luchetti, G. and Fuersteburg, R., Relative fish use in urban and nonurban streams, Proceedings of the Conference on Wild Salmon, Vancouver, Canada, 1993. Metcalf and Eddy, Wastewater Engineering: Treatment, Disposal, and Reuse, 3rd ed., McGraw-Hill, New York, 1991. Metropolitan Washington Council of Governments, The state of the Anacostia: 1989 status report, prepared for the Anacostia Watershed Team, Washington, D.C., 1990. Metropolitan Washington Council of Governments, Anacostia Watershed Restoration Progress and Conditions Report, 1990–1997, prepared for the Anacostia Watershed Restoration Committee, 1998. Peterson, J. W., Meet the National Watershed Coalition, Land Water, January/February, 10, 1998. Schueler, T., Controlling Urban Runoff—Practical Manual for Planning and Designing Urban Best Management Practices, Metropolitan Washington Council of Governments, Washington, D.C., 1987. Schueler, T., Site Planning for Urban Stream Protection, prepared for the Metropolitan Washington Council of Governments, Washington, D.C., 1995. South Florida Water Management District, Guidebook to Develop a BMP Environmental Protection Plan, draft, 1999.

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U.S. Environmental Protection Agency, Watershed Protection: A Project Focus, Environmental Protection Agency 841-R-95-003, Office of Water, Washington, D.C., 1995a. U.S. Environmental Protection Agency, Watershed Protection: A Statewide Approach, Environmental Protection Agency 841-R-95-004, Office of Water, Washington, D.C., 1995b. U.S. Environmental Protection Agency, Why watersheds, Environmental Protection Agency 800-F-96-001, Office of Water, Washington, D.C., 1996a. U.S. Environmental Protection Agency, Draft Framework for Watershed-Based Trading, Environmental Protection Agency 800-R-96-001, Office of Water, Washington, D.C., 1996b.

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Kent, Donald M. “Managing Global Wetlands” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

13

MANAGING GLOBAL WETLANDS Annette M. Paulin and Donald M. Kent

CONTENTS The Ramsar Convention Membership Wetland Definition and Classification Management Additional Support Case Studies Florida Everglades Functions and Values Threats and Impacts Conservation Effort The Mekong Delta Functions and Values Threats and Impacts Conservation Efforts The Pantanal Functions and Values Threats and Impacts Conservation Efforts The Wadden Sea Functions and Values Threats and Values Conservation Efforts References

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Historically, a significant amount of wetland management effort has been focused on individual wetlands. Commonly, individual wetlands are encompassed and regulated by a single authoritative unit, such as a town or city. With the advent of watershed management, the focus has broadened to include simultaneous consideration of multiple wetlands. Watershed management typically requires cooperation and coordination among several authoritative units, such as municipalities, counties, and states. Managing wetlands that cross international boundaries or wetlands within a single country, but that are of international importance, poses additional challenges. Managing wetlands that cross international boundaries requires cooperation and coordination among countries. This may be accomplished informally by communication between respective government environmental agencies, or formally by establishment of joint proclamations or management plans. Understanding that an individual country’s best interests are served by ensuring protection of shared wetland resources is fundamental to effective management. In some instances, wetland function and value, even those contained with a single country, may be of global importance. For example, a wetland may function as flood storage for downstream, transboundary communities. Continued wetland function would ensure that downstream communities have adequate irrigation for agriculture and are protected from catastrophic floods. Alternatively, a wetland may be a critical breeding or wintering area for migratory fish and wildlife. Managing wetlands at the global level could be self-regulating. That is, resource users, whether governments or private entities, would recognize the value of wetlands and protect their investment. The World Trade Organization might be a model or a vehicle for this type of management. Success would require sustainable use of resources, an accurate valuation of all wetland values, and a mediation process. Lending institutions might also effect management. Funding of significant development projects would require conduct of a cost–benefit based environmental impact assessment. The perspectives of both the applicant country and the international community would need to be considered. Some international lending institutions require an impact assessment, but the process does not appear to adequately value wetland resources. Presently, the United Nations through the Ramsar Convention on Wetlands effects management of global wetlands. Participation in the Ramsar Convention is voluntary, and there is no enforcement authority. Wetland protection is effected through education and management assistance. The first part of this chapter describes the approach and operation of the Ramsar Convention. The balance of the chapter describes four wetlands of international significance. The wetlands illustrate a range of challenges and the conservation efforts being enacted to protect wetland functions and values.

THE RAMSAR CONVENTION The primary instrument for the protection and management of wetlands on a global scale is the Ramsar Convention on Wetlands of International Importance (Ramsar Convention, 1999). As of 1999, there were 114 contracting nations and 975 ©2001 CRC Press LLC

wetlands totaling 70.7 million ha. The United Nations Educational, Scientific, and Cultural Organization serves as the depository for information, and funds the Convention. The Ramsar Convention recognized that wetlands have great economic, ecological, and cultural value, and that encroachment and loss of wetlands must be reduced. Because water resources for wetlands often cross political boundaries, the Ramsar Convention now provides a framework for intergovernmental cooperation in the conservation and wise use of wetlands. Membership A nation must have one listed wetland to become a participating member of the Ramsar Convention. A wetland can be listed if it satisfies at least one of several criteria. Geographical and ecological criteria include being representative of a natural or near natural wetland, common to more than one biogeographical region, representative of a wetland that plays an important role in the natural connection of a major river basin or coastal system (especially where located in a transborder position), or unique as a rare or unusual type of wetland in the biogeographical region. A wetland may also satisfy listing criteria by playing a role in plant or animal species integrity. For example, the wetland may support rare, vulnerable, or endangered species by maintaining genetic and ecological diversity of the flora or fauna of a region, by providing habitat for plants or animals at a critical stage of their biological cycle, or by containing one or more endemic plant or animal species or communities. The final criteria for listing are related to biological value and are based on supporting regulatory waterfowl communities of 20,000 or more individuals, a substantial number of a particular group of waterfowl, or 1 percent of the individuals in a population of a species or subspecies of waterfowl. If a wetland satisfies at least one criterion, it will be listed as a wetland of international importance. If a wetland fails to satisfy the criteria, measures may be taken to restore or enhance its values and function in order to meet one of the criteria. If these measures fail, the site will be not be listed. Dues are paid based on a sliding scale determined by Gross National Product of the member nation. Members of the Ramsar Convention have certain obligations. These include considering wetland conservation within the framework of land-use planning, promoting conservation of wetlands throughout their region, and establishing wetland reserves. Members must also provide in-country training in the fields of wetland research and management, and exchange information and data with other members of the Convention. Finally, members must consult with the Ramsar Convention regarding management implementation, especially when it involves transboundary wetlands, shared water resources, or shared development aid for projects. Wetland Definition and Classification The Ramsar Convention has a definition and classification system for identifying wetlands of international importance. Wetlands are defined as “areas of marsh, fen, peatland, or water, whether natural or artificial, permanent or temporary, with water that is static or flowing, fresh, brackish or salt, including areas of marine water the ©2001 CRC Press LLC

depth of which at low tide does not exceed six meters … and may incorporate riparian and coastal zones adjacent to the wetland, and islands or bodies of marine water deeper than six meters at low tide lying within the wetlands” (Davis, 1994). The definition incorporates ecosystems that are an integral part of a major water system and is broader than many other operational wetland definitions (see Chapter 1). The Ramsar Convention (Davis, 1994) recognizes four major types of wetlands. Marine wetlands are coastal wetlands including rocky shores and coral reefs. Estuarine wetlands are located between salt water and fresh water bodies, or dry land including deltas, tidal marshes, and mangrove swamps. Lacustrine wetlands are wetlands associated with lakes. Palustrine wetlands are isolated marshes, swamps, or bogs. Wetlands are classified when listed as wetlands of international importance by the Ramsar Convention. Management The Convention assists members with plans of action for management of their wetlands by sharing information and through the activities of a Scientific and Technical Review Panel. The Panel is comprised of experts in the field of wetland management. They provide members with expert opinions and assist with the design of management plans. The management planning process has three steps: description of the site, forming evaluations and objectives, and designing an action plan or prescription. Description of the site includes identifying the wetland type(s) and creating an inventory of the flora and fauna. This first step is used to establish criteria for listing the site. The evaluation provides a detailed report of the site, including information on biological diversity, integrity, rarity, fragility, history, cultural and aesthetic value, social and economic value, education and research opportunities, and potential uses for recreation. In evaluating the site, concise objectives are formed for best management practices. These objectives are based strictly on the evaluation of the site, and are intended to fully protect its characteristics. After the objectives are outlined, any factors that may hinder their achievement, including both natural and humaninduced, are identified. Considering the objectives and mitigating factors, operational objectives are developed. Management strategies are established based on the best possible alternatives under the given circumstances. A limit of acceptable change is established to meet protection obligations. Finally, a plan of action is outlined which may include zoning, habitat management, species management, contextual uses, education, and research initiatives. Specific projects and work programs are designed to implement these actions. Over time, reviews of progress at the site are presented to the Ramsar Convention to ensure operational success, or to facilitate changes in objectives and action plans. Additional Support The Ramsar Convention provides additional assistance to member nations through three documents: the Montreux Record and Monitoring Procedure (Davis, ©2001 CRC Press LLC

1994), Towards the Wise Use of Wetlands (Davis, 1993), and the Economic Valuation of Wetlands (Barbier et al., 1997). The Montreux Record lists priority sites that are undergoing ecological changes owing to human activities. The Scientific and Technical Review Panel supports a procedure for monitoring site changes and implementing management strategies. National management strategies are encouraged to comply with guidelines described in Towards the Wise Use of Wetlands (Davis, 1993). The Ramsar Convention defines the wise use of wetlands as “sustainable utilization for the benefit of mankind in a way compatible with the maintenance of the natural properties of the ecosystem” (Davis, 1993). Guidelines include establishing an integrated approach to policy making using coordinated efforts of national, regional, and local institutions, and providing policies that promote wetland protection in land-use planning, environmental audits, financial incentives, and permit processes. Wise Use provides examples of management strategies by describing wetland inventories, monitoring techniques, research on identifying values, wetland use, and landscape function. The management strategies include establishment of training programs and promotion of public awareness (Davis, 1993). Actions outlined by Wise Use include maintaining ecological integrity, sustainable use, balancing restrictions with cultural uses, and integrating wetland management with development plans. The latter seeks to achieve a balance between conservation and the use of wetland resources. Wise Use also provides 17 case studies to illustrate management issues and lessons learned from the implementation of the Wise Use guidelines. The Ramsar Convention developed a guide, Economic Valuation of Wetlands (Barbier et al. 1997), to facilitate the economic valuation of wetland resources. Developed in conjunction with the Department of Environmental Economics and Environmental Management at the University of York, the Institute of Hydrology, and the World Conservation Union (IUCN), the document outlines several approaches to valuate the economic value of wetlands and weigh the benefit of development strategies with the degradation it may cause to wetland resources. One somewhat unique aspect of the document is the emphasis on social, cultural, and political values for decision-making. This is accomplished by addressing not only the valuation of direct economic benefits from wetlands (e.g., timber and food resources), but also indirect economic benefits (e.g., biological functions such as flood attenuation and future uses and benefits) and nonuse values (e.g., biodiversity and cultural heritage). The framework provided in Economic Valuation of Wetlands includes seven steps (Table 1). In practice, the valuation requires an interdisciplinary approach and the cooperative effort of specialists, including economists, hydrologists, fishery and wetland biologists, and sociologists. The first step in the valuation is selecting the appropriate approach. There are three assessment approaches, impact analysis, partial valuation, and total valuation, available based on the type of development and degree of impact to the wetland’s integrity. An impact analysis assesses the external costs of off-site development or activities such as discharges from industries or mining activities. This analysis would compare the benefit of the activity to the losses in specific wetland resources from off-site impacts. A partial valuation is used to evaluate changes in the allocation or ©2001 CRC Press LLC

Table 1

Framework for the Economic Valuation of Wetlands (Barbier et al., 1997)

1. Select an assessment approach 2. Define the wetland boundary and system boundary 3. Identify and rank wetland components, functions, and attributes 4. Relate components, functions, and attributes to type of use value 5. Identify information required to assess uses 6. Quantify economic values 7. Implement the appraisal method

the alternative uses of wetland resources. An example is the diversion of floodwater for irrigation. While not all of the wetland resources may be impacted, valuation of the benefits lost is compared to benefits gained from the irrigation project. In addition, evaluation of the alternative uses of the floodwater would be considered. Such alternatives may be the use of floodplains for fish farming or agricultural benefits from nutrient loading of floodwaters. The third approach, total valuation, is used to evaluate a wetland’s contribution to society as a whole. This valuation may be applied in wetland preservation strategies and regional natural resource assessments. The second step in the framework is to define the wetland area, time scale, and analytical boundaries of the assessment. The defined parameters will differ based on the assessment approach. An impact assessment may only require a short time scale such as the time discharge flows from an industry and small analytical boundaries such as the area of the water resources impacted. A total valuation may require consideration of all wetland resources, a large analytical boundary, and an extended time scale. Step three identifies the corresponding functions and attributes of the wetland. This step requires review of previous research and may require additional research to ensure thorough identification of the functions and attributes to be considered in the valuation. This critical stage will also require the collaborative teamwork of specialists in differing fields. Step four in the valuation defines and prioritizes the values of the identified wetland functions and attributes (see Chapter 3). Functions and values may be use (e.g., direct, indirect, and option or quasi-option) or nonuse. Direct uses are values most often used in economic valuation studies because these activities are directly marketed. Examples of direct use include agricultural resources, fuelwood, recreation, harvesting, and transportation uses. In many instances, these activities are used for subsistence purposes and appropriate evaluation techniques must be applied. However, either as marketed or subsistence resources direct use can be quantified based on a marketed economic value. Indirect use values are primarily ecological functions that protect or support direct uses. These values include flood attenuation, erosion control, nutrient retention, and groundwater recharge. Option and quasi-option values are potential future uses (both direct and indirect) and future information use. Value accrues by delaying development or exploitation. Option and quasi-option values may change with changes in economic, social, and scientific circumstances. Nonuse value is the wetland’s intrinsic existence value. The most difficult to quantify, nonuse value includes biodiversity, cultural heritage, and preservation for ©2001 CRC Press LLC

future generations. Contributions to conservation campaigns are one indicator of the value society places on these uses. Step five in the valuation process is to obtain detailed information about the identified values. This information includes scientific data, statistical data on human uses, economic inputs and outputs of activities, and survey results, and is critical to the valuation process. For example, scientific data will provide details on indirect use values such as flood retention capacity, degree of erosion protection, populations of harvested species, and growth rates of forests. Statistical data on human uses and economic inputs—outputs of activities, including agricultural and fishery yields and tourism revenues—provide detail on direct marketable uses. Information from surveys may provide valuable information on option based or intrinsic values that are otherwise difficult to quantify. Once an appropriate assessment approach is chosen, and values are identified, classified, and defined statistically, the actual valuation is conducted. Most critical in this sixth step is choosing the proper technique for valuating resources and their use. Economic Valuation for Wetlands does not detail the methods for each technique, although it provides a list of advantages and disadvantages. For example, market prices may be applied to direct market uses, while surrogate market price may be applied to a wetland resource that is not marketed but is closely related to a marketed good or service. Another direct use method is indirect substitute, where the cost of an alternative source of resources is applied to the wetland resource such as water imported from the outside vs. water used from the wetland. For indirect uses, approaches such as the value in changes in productivity and damage costs avoided can be used to determine the impact of ecological degradation. To determine option values and nonuse values, the contingent valuation method is most widely used. Because of its context in nonuse valuation, this method is controversial. However, it does provide an assessment of an individual’s or society’s willingness to pay for the value, or how much compensation they would require upon loss of the use. Another approach to determining option and nonuse values is to determine the sustainable yield of current activities and alternative or compensating projects that could be offered. If current activities are not sustainable, alternative scenarios that offer higher social returns are offered. Compensating projects offer mitigation for environmental degradation, while maintaining long-term sustainability in the overall natural system and ensuring nonuse values. The final step in the valuation framework is to implement the appropriate appraisal method. Again, Economic Valuation of Wetlands does not detail methods but lists advantages, disadvantages, and most appropriate cases for implementation. These methods include cost–benefit analysis, multiple criteria analysis, land suitability/classification models, environmental impact assessments, and cost-effectiveness analysis. Upon implementation of appraisal methods, Economic Valuation of Wetlands stresses the importance of applying the economic valuation methods within social, political, and cultural contexts. In addition, this valuation should be conducted with interdisciplinary collaboration and provide effective institutional capacity building upon decision making. This capacity is based on the training and information provided by those conducting the valuation to those involved in decision making. Thus, the valuation should be founded on thorough information and appropriate ©2001 CRC Press LLC

evaluation techniques. As stated by Delmar Blasco, Secretary General for the Ramsar Convention Bureau, “It is important to stress that economic valuation is not a panacea for all decisions, that it represents just one input into the decision-making process, along with important political, social, cultural and other considerations. The goal of this text is to assist planners and decision-makers in increasing the input from economic valuation in order to take the best possible road towards a sustainable future” (Barbier et al., 1997).

CASE STUDIES Four case studies illustrate the types of management issues confronting countries with wetlands of international importance (Table 2). Both freshwater and coastal wetland systems in developed and developing countries are represented. The types of threats and impacts include excessive resource extraction, altered hydrology, and pollution. Conservation efforts include restoration, international agreements, protection of critical areas, education, and use restrictions. The Everglades is contained wholly within the United States. The Everglades supports many threatened and endangered species and migratory bird populations. Groundwater recharge and surface water flow are critical to the social and economic well-being of burgeoning South Florida. The other case studies illustrate issues associated with transboundary wetlands. The Mekong Delta is a major coastal system. The Delta has recovered from wartime impacts but is now threatened by unsustainable subsistence use and upstream hydropower projects. The Pantanal is a significant contributor to regional biodiversity. In some ways, the situation in the Pantanal mirrors that of the Everglades in the early 20th century—a large, relatively pristine and productive wetland is threatened by a proposal to alter regional hydrology to accommodate economic growth. The Wadden Sea is a major coastal and shallow marine system in northern Europe. Located in a developed region of the world, the Wadden Sea is subjected to coastal armament and pollution. Florida Everglades Known as the River of Grass, the Florida Everglades is one of the largest freshwater marshes in the world, historically encompassing about 3.5 million ha (Figure 1). It extends from Lake Okeechobee in south central Florida to the southern tip of Florida where its wide mouth empties into Florida Bay and other smaller bays and estuaries. The headwaters of the Everglades begin as small streams and lakes in central Florida and flow through the Kissimmee Lakes and River system. This system leads to Lake Okeechobee, a 300 ha natural reservoir formed in the center of the state when sea levels fell during the last ice age. Water historically flowed over the southern edge of Lake Okeechobee and into the Everglades. From here the water flowed to the Gulf of Mexico via the Caloosahatchee River and to the Atlantic Ocean via the St. Lucie River (Robinson et al., 1996).

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Table 2

Characteristics of Case Study Wetlands Function and Value

Threats and Impacts

Fish and wildlife habitat, recreation, tourism, agriculture, drinking water, cultural heritage Fisheries, agriculture, forest resources, transportation

Agriculture, altered hydrology, pollution, exotic vegetation

Protected areas, restoration plan, stormwater treatment areas, agricultural BMPs

Deforestation, pollution, hydropower

Protected areas, education, Commission toward Sustainable Development Protected areas, education, Intergovernmental Committee on Hydrovia Joint Declaration on the Protection, protected areas, prohibited shoreline armoring

Location

Size (ha)

Habitat

Florida Everglades

United States

3,500,000

Freshwater marsh, mangrove, swamp, slough, estuary, shallow bay

Mekong Delta

Laos, Myanmar, Thailand, Cambodia, Vietnam

3,900,000

Melaleuca forest, mangrove, tidal mudflat

Pantanal

Brazil, Bolivia, Paraguay

11,000,000

Swamp, forest, savannah, lake margin scrub, gallery forest

Flood control, fish and wildlife habitat, fisheries, cattle ranching, tourism

Wadden Sea

Denmark, Germany, Netherlands

1,350,500

Tidal channel, mud flat, salt marsh, beach, dune

Primary productivity, fish and wildlife habitat, tourism, recreation

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Cattle ranching, agriculture, overfishing, hunting, Hydrovia project Infrastructure development, pollution, shellfish harvesting

Conservation Efforts

Lake Okeechobee

Florida EAA

WCA-1 WCA-2

WCA-3

Miami Everglades National Park

Florida Bay

Figure 1

The Florida Everglades; the shaded area represents the remnant Everglades–Florida Bay ecosystem; WCAs are water conservation areas. (Modified from South Florida Water Management District, 1999.)

Functions and Values Small elevation differences within the Everglades contribute to the formation of many distinct ecosystems including swampy cypress domes, hydric hardwood hammocks, pinewoods, freshwater sloughs, brackish estuaries, shallow shorelines and embayments, and deeper gulf waters (Frazier, 1996; World Conservation Monitoring Centre, 1990). These diverse ecosystems are habitat for 25 terrestrial and 2 aquatic mammals including the endangered Florida black bear (Ursus americanus), Florida panther (Felis concolor), and West Indian manatee (Trichechus manatus). Over 300 species of birds have been observed, many of which use the area as a stop on their migration to and from Central and South America. Many of North America’s waterfowl and shore birds can be found using the Everglades as a seasonal home. Some, like the wood stork (Mycteria americana), sandhill crane (Grus canadensis), glossy ©2001 CRC Press LLC

ibis (Plegadis falcinellus), and roseate spoonbill (Ajaia ajaja), are threatened or endangered. The threatened or endangered American crocodile (Crocodylus acutus), American alligator (Alligator mississippiensis), indigo snake (Drymarchon corais), loggerhead turtle (Caretta caretta), hawksbill turtle (Eretmochelys imbricata), and green sea turtle (Chelonia mydas) are among the more than 50 species of reptiles inhabiting the Everglades. The flora of the Everglades is as varied as its ecosystems, with 10,000 species of seed-bearing plants and 120 species of trees. Plant types range from the characteristic sawgrass (Cladium jamaicensis), to bromeliads and epiphytic orchids, to tropical (e.g., gumbo limbo, Bursera simaruba) and temperate (oaks, Quercus spp.) trees. Over 60 plant species found in the Everglades are endemic to South Florida. Periphyton is an important component of the Everglades ecosystem. High in calcite and dominated by filamentous blue-green algae, the periphyton community helps build marl, a soil overlying the limestone foundation of the Everglades, and supports an intricate food web. In the transition from fresh to salt water, the Everglades estuarine systems provide a nursery for the Florida Bay’s fishing industry. In 1989, harvesting of lobster (Panulirus argus), stone crab (Menippe mercenaria), and pink shrimp (Penaeus duorarum) brought in revenue of US $61 million (Redfield et al., 1999). One county alone made over $35 million in 1992, making the Everglades and Florida Bay an important economic resource (Robinson et al., 1996). In addition, Everglades National Park is an economic center for tourism with over 250 thousand visitors a year (World Conservation Monitoring Centre, 1990). Another economic interest operating within the Everglades historic boundary is agriculture. Based mainly on sugarcane crops, this multibillion dollar industry relies on the rich soils of drained portions of the Everglades. Unfortunately, the development of this industry has had profound direct and indirect effects on the integrity of the Everglades ecosystem. The Everglades also has important social and cultural values. Demand for freshwater resources increased markedly with the development of South Florida. The porous nature of the limestone baserock recharges an underlying aquifer. A total of 90 percent of the regional population receives its potable water from this aquifer. Freshwater recharge into the southern part of the Everglades prevents saltwater intrusion into the aquifer. The Everglades also protects South Florida residents from heavy surf associated with hurricanes and high tides. Mangrove forests and barrier islands along the coastal edges serve as protective barriers. The Everglades cultural value derives from historic and continued habitation by indigenous peoples (Robinson et al., 1996). The Calusa and Tequesta arrived in Florida 11,000 years ago and thrived until the early 1800s when they succumbed to diseases and war accompanying European settlers. Remnants of the Calusa and Tequesta remain in the form of shell mounds and artifacts. Other Native Americans, the Seminoles, fled to Florida in the early 18th century when tribes were pressured west and south by colonial expansion. A small, non-Christian faction of the Seminoles, the Miccosukee, lives in the Everglades today.

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Threats and Impacts Impacts to the Everglades are related to a long history of cultural and economic development. The earliest records of the first ecological changes to the system are from the late 19th century (Robinson et al., 1996). In 1881, one of the first developers, Hamilton Disston, began draining wetlands around Lake Okeechobee to create farmland. In addition, canals were built within the northern part of the watershed. While these drainage efforts had little effect on Lake Okeechobee and its overflows into the Everglades, it opened the door to reclaiming the Everglades for civilization (Derr, 1993). With mild year-round weather and fertile soils, the area quickly became an agricultural center. To support development, the State of Florida encouraged drainage. The Everglades Drainage District excavated 700 km of canals by 1927. The Hoover Dike was built along the southern shore of Lake Okeechobee in the late 1920s in response to flood disasters caused by several hurricanes. In large part, the dike separated the lake from the Everglades. During this same time period, the Tamiami Trail was built, linking Miami with the Gulf Coast. This roadway, built through the middle of the Everglades, separated the northern and southern Everglades disrupting the natural hydroperiod. Construction of the canals, Hoover Dike, and the Tamiami Trail had a quick and obvious impact on the hydrology of the Everglades. Models indicate that the first four canals removed 1.5 million acre-feet of water per year from the system. Impacts to wildlife, especially waterfowl, were evident. In addition, a 1.2 m decrease in groundwater levels caused soils to oxidize and subside. By 1940, 1.8 to 2.1 m of soil was lost and the slope of the land became concave, reversing the direction of flow in the northern part of the Everglades toward Lake Okeechobee (Robinson et al., 1996; Redfield et al., 1999). In 1948, in response to concerns about droughts and floods in the northern part of the Everglades, Congress authorized an expansion of the canal system (McLean and Bush, 1999). The US $208 million Central and South Florida Project consisted of 1250 km of flood control canals. The Central and South Florida Flood Control District was established to manage the project and to allocate water resources through the system of canals, levees, locks, and dams. Additional levees were built south of Lake Okeechobee to block sheet flow to the increasingly populated Atlantic Coast. The levees were also used to create Water Conservation Areas that would retain the waters of the Everglades. During this same time period, the southernmost 5 percent of the historic Everglades was designated as a National Park. However, this designation would not protect the Everglade’s fragile and valuable ecosystem. By the 1960s, degradation of the Everglades was a sensitive issue throughout the United States. Congress enacted the Water Resource Development Act in 1970, mandating a minimum water delivery to the Everglades. In 1972, the Water Resources Act was passed, mandating the protection of the Everglades through improvements in water quantity and quality. The Central and South Florida Flood Control District was renamed the South Florida Water Management District (one of five newly formed districts in the state), with a new mission of protecting water quality and preserving environmental values ©2001 CRC Press LLC

(Robinson et al., 1996). The new District was required to balance the water resource needs of urban, agricultural, and natural areas. Areas in the northern and central part of the Everglades, between the Everglades Agricultural Areas and Everglades National Park, were placed under protection and designated as the Everglades Protection Area. During the next 20 years, studies would indicate that the Everglades continued to be impacted. Despite new water discharge requirements, the quality and quantity of water resources delivered to the park were inadequate to maintain ecosystem health. For decades, the most obvious change to the ecosystem was alteration of the hydrologic flow. Conversion of more than 50 percent of the historic Everglades to agricultural and urban areas resulted in a decrease in regional water storage. The construction of canals, drainage ditches, and dikes to control floodwaters dramatically changed natural sheet flow and hydroperiod. Sudden releases from the water control structures inundated areas that that had been abnormally dry for extended periods, disrupting foraging and nesting habits of wading birds and herpetofauna. In addition, changes in water depth altered macrophyte and algal community composition, disrupting primary production processes. The disruption in natural sheet flow to the Everglades has impacted areas as far south as the Florida Bay, and possibly the coral reefs of the Florida Keys (Robinson et al., 1996). Because freshwater flow has decreased, salinity levels have risen stressing aquatic organisms adapted to narrower salinity ranges. In addition, waters of Florida Bay have warmed, altering current exchange patterns with the Atlantic Ocean. Reduction in freshwater flows to the Everglades has also impacted mangrove and upland communities of the Atlantic Coast as groundwater recharge decreased and saltwater intrusion increased. The loss in groundwater recharge has also impacted the supply of potable water to the region (Robinson et al., 1996). Groundwater resources, the primary source of South Florida’s water supply, are used at a rate exceeding recharge rates. Freshwater has become scarcer, and potable water must be imported to large parts of the region. Research indicates that the water reaching the Everglades is of poor quality (McCormack et al., 1999). Stormwater runoff diverted by canals to the Everglades Protection Area contains high levels of nutrients from agricultural areas. Once an oligotrophic system, the Everglades now exhibits signs of eutrophication. Natural surface water total phosphorus concentrations are between 4 to 10 ppb, while agricultural runoff concentrations range between 50 to 200 ppb total phosphorus. Natural soil total phosphorus concentrations are between 200 to 500 ppb, whereas soil concentrations downstream of agricultural areas reach 1000 ppb total phosphorus. Elevated concentrations of total phosphorus are linked to shifts in algae species better adapted to high nutrient loading. Sudden algal blooms cause low water column dissolved oxygen levels threatening invertebrate organisms. In Lake Okeechobee, algal blooms cause fish kills. Florida Bay has experienced low dissolved oxygen levels with subsequent impacts to seagrass communities vital to fisheries. High nutrient loading has also contributed to a shift from diverse wetland plant species to a monoculture of cattails in some parts of the Everglades. Under high nutrient conditions, cattails are able to out compete sawgrass and other macrophytes, especially when accompanied by hydrologic changes. ©2001 CRC Press LLC

Hydrologic and other disturbances have created opportunities for the establishment of exotic vegetation (Ferriter et al., 1999). Brazilian pepper (Schinus terebinthifolius), Australian pine (Casuarina litorea), and especially melaluca (Melaleuca quinquenervia), are of primary concern. Melaluca is the dominant exotic in some parts of the Everglades. It thrives in lower water levels and produces thousands of seeds per plant. Melaleuca was introduced to the Everglades in the belief that its high evapotranspiration rate would dry out wet areas (Derr, 1993; Robinson et al., 1996). The latest threat to the Everglades is the accumulation of mercury in sediments (Fink et al., 1999). Mercury is deposited from the atmosphere and converted to highly toxic methylmercury by sulfate bacteria. Methylmercury bioaccumulates and has the potential for affecting wildlife at all levels of the food web. Top predators may accumulate up to 10 million times the concentration in water. Mercury also presents a human health risk, and fish consumption in the Everglades, eastern Florida Bay, and the Big Cypress Conservation Area is either prohibited or restricted. Although the accumulation of methylmercury has not been observed to affect reproductive rates in waterbirds, it may affect eating and foraging habits (Frederick et al., 1999; Bouton et al., 1999). Mercury poisoning is the suggested cause of several Florida panther deaths and reduced litter sizes (Florida Panther Interagency Committee, 1989). Conservation Efforts In 1988, the United States government sued the Florida Department of Environmental Regulation and the South Florida Water Management District for failing to maintain high water quality in waters flowing to the Everglades National Park and the Loxahatchee National Wildlife Refuge. In 1991, a settlement was reached and the defendants agreed to guarantee water quality and water quantity needed to preserve and restore the unique flora and fauna of the Park and the Refuge (Robinson et al., 1996). With a July 2002 deadline to meet these agreements, the state of Florida passed the Everglades Forever Act in 1994. Strategies outlined in the Act are designed to achieve water quality standards by December 31, 2006. The Everglades Forever Act mandated a fair share of restoration costs be borne by agricultural interests. In addition, the Act recognized that urban development encroached upon the natural system and authorized the South Florida Water Management District to impose a homeowner’s tax of an average of US $10 a year (Robinson et al., 1996). Initial management efforts focused on reducing nutrient loading from agricultural area stormwater runoff to the Everglades Protection Area (Chimney et al., 1999). The Everglades Nutrient Removal Project purchased agricultural land and converted it to wetland stormwater treatment areas (STAs). The STAs are macrophyte-based systems designed to remove phosphorus from agricultural stormwater runoff. The original goal of the STAs was to reduce total phosphorus in agricultural runoff to 50 ppb or less and to discharge the treated water to the Everglades Protection Area. To date, the STAs have reduced total phosphorus concentrations to an average of 22 ppb, an 82 percent load reduction (Redfield et al., 1999). A total of 63 metric

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tons of phosphorus have been removed that would have been discharged to the Loxahatchee National Wildlife Refuge. Best Management Practices (BMPs) are being implemented in the Everglades Agricultural Area (Whalen et al., 1999). A total of 16 BMPs have been identified and implemented through regulatory and educational initiatives. The BMPs include increased water detention for settling of nutrients, fertilizer application controls (e.g., direct root application by banding or split application), sediment controls (e.g., ditch improvements and sump installation), pasture management, xeriscaping, and vegetative filtering. During the first 3 years of BMP implementation (1995 to 1998), the total phosphorus load was reduced by 55 percent (Redfield et al., 1999). Additional management initiatives are focused on mercury, supplemental treatment technologies, and a Central and South Florida restudy. Mercury studies will provide information about the complex interactions between phosphorus, sulfate, and other water quality parameters and methylmercury. Other studies and modeling will estimate the no observable adverse effect level for Everglades fish and wildlife. Other research is underway to identify supplemental technologies for reducing nutrient loads to the EPA. Recent studies have suggested that optimal total phosphorus concentrations should not exceed 10 ppb. STA performance must be supplemented to achieve this level. Currently, nine technologies, ranging from chemical treatment to improved treatment wetland design, are being evaluated. Submerged aquatic vegetation and periphyton-based wetland treatment systems appear promising. Successful technologies may be integrated into the existing STAs or implemented as an additional treatment system. However, certain guidelines must be met, including technical feasibility, local acceptability, and large-scale applicability. The most effective technology(s) will be implemented by December 2003. The largest conservation effort is a restudy of the Central and South Florida Project (McLean and Bush, 1999). The main objective of the study is to create additional storage and increase water flowing to the Everglades Protection Area (Redfield et al., 1999). The study will include intensive modeling of hydrologic regimes, ecosystem dynamics, and water storage. Implementation will cost an estimated US $7.8 billion and may include modification of existing flood control structures, optimization of hydrologic regimes, and waterway restoration. Plans to return the Kissimmee River to its natural state are being developed and implemented. This project alone will be one of the largest restoration projects in the world. The Mekong Delta The Mekong River flows from the Himalayan Mountains of China 4200 km to the South China Sea. During its course, the river travels through the countries of Laos, Myanmar (Burma), Thailand, Cambodia, and Vietnam (Figure 2). The river drops its sediment load upon reaching the calm waters of the China Sea, creating shallow plateaus where vegetation grows. This is the Mekong Delta. The Mekong Delta encompasses 3.9 million ha of mangrove and melaleuca forests, tidal mudflats, and shrimp and fish ponds (Frazier, 1996).

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Functions and Values The many wetlands of the Mekong Delta are habitat for breeding colonies of native waterfowl (Frazier, 1996). Over 50 species of migratory birds use the wetlands during the migratory period. The coastal wetlands also provide protection to the coastline, reducing erosion from wave action. The majority of the coastal wetlands are mangrove forests that act as a sediment filter, reducing turbidity and increasing water quality. The inland wetlands are comprised mainly of melaleuca forests. These seasonally flooded wetlands reduce the acidity and sulfur content of underlying soils.

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The Mekong Delta plays a critical role in the economy of Vietnam (Khoa and Roth-Nelson, 1994; Frazier, 1996). Vietnamese rely heavily on the Delta for food, fuelwood, timber, and water resources. Much of the land area is used for agriculture, especially rice production, and aquaculture including fish, crab, and shrimp nurseries. Timber from both mangrove and melaleuca forests are used for firewood, furniture, home construction, and tannin. Other goods from the Delta include paper, glue, synthetic fiber, sugar wine, and salt. In addition, the Delta's waterways are transportation routes between economic centers. Because of its economic significance, the Mekong Delta has become a cultural center of over 14 million inhabitants (20 percent of Vietnam’s population), with a density of 300 to 500 people per km2. Threats and Impacts The Vietnam War contributed heavily to the initial degradation of the Mekong Delta (Khoa and Roth-Nelson, 1994; Frazier, 1996). Napalm bombs and defoliants destroyed much of the vegetation, including nearly half of the mangrove forests. A large section of the Delta, the Plain of Reeds, was drained for combat. Many people were left homeless when the war ended. In an effort to survive and to rebuild their nation, the people of Vietnam exploited what little resources remained. The Mekong Delta has recovered significantly from the war, although the area still suffers from unsustainable uses of resources (Duc, 1993; Khoa and Roth-Nelson, 1994). Inland, melaleuca forests are used as a fuelwood resource and are areas of agricultural and aquacultural activities. Due to the highly acidic soils, production is low, and the farms are often abandoned. Thus, there is a continuous cycle of forest clearing as subsistence farmers move to new plots of forest. As the melaleuca timber resources are depleted, wood from coastal mangroves is exported inland. Mangrove deforestation increases coastal erosion; up to 70 m of land is lost to the sea per year (Bentham et al., 1997). The loss of coastal mangroves also leads to salt water intrusion, disrupting the delicate balance of fresh and salt water to which many organisms have become adapted. Another impact to water quality is pollution from human activities both within the Delta and the upper Mekong River Basin (Mekong River Commission, 1995). Sewage is regularly disposed of directly into the water. Upstream, agricultural activities contribute both nutrients and chemicals into the Delta's water source. These pollutants decrease water quality and compromise the integrity of the wetland ecosystems. Additional threats to the Delta’s integrity arise from the development of the upper Mekong River Basin (Lohmann, 1990). Planned at high water velocity points along the Mekong River are 6 to 50 hydropower projects. The diversion of fresh water may alter the Delta’s hydrologic regime and salinity, reduce organic matter inputs, increase pollution, and decrease fish populations (Lanza, 1996; Wegner, 1997). The hydropower projects are of international importance. In an area of great economic instability, countries have much to gain from reliable and relatively inexpensive energy (Tu, 1996; Osborne, 1996; Chape and Inthavong, 1996). For example, much of the Mekong lies along Cambodia’s border. Presently, Cambodia's

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development needs are minimal because its population is small and its economy is restrained by a socialist political system. However, dams constructed in Cambodia's part of the Mekong River could harness large amounts of energy and be sold to more developed countries, such as Thailand, where populations are booming and economic opportunities are growing due to capitalistic industrialization and a democratic political system. Conflicts in water resource allocation could arise as Mekong River Basin nations achieve economic development at the expense of the economic and cultural stability of the Vietnamese living in the Delta. In a region of historical conflict and instability, cooperation between countries of the Mekong River basin is necessary to maintain current peaceful conditions. Conservation Efforts Although the Mekong Delta is not a site listed under the Ramsar Convention, Vietnam is a participating member (with one listed site). As a member, it must meet obligations outlined by the Convention. These include promoting wetland conservation throughout the region, establishing wetland reserves, considering wetland conservation within the context of land use planning, and promoting training in the management of wetland resources. Through the Ministry of Water Resources and the Ministry of Agriculture, the Vietnam Government has established six protected wetland regions within the Mekong Delta (Khoa and Roth-Nelson, 1994; Beilfuss and Barzen, 1994). They are the Tram Chin Crane Reserve at Dong Thap Muoi, the Nam Can Mangrove Reserve, Vo Doi Melaleuca Protected Forest, and three waterfowl breeding colonies in Bac Lieu, Cai Nuoc, and Dam Doi. Management within these sites includes protection of waterfowl, the reestablishment of mangrove and melaleuca forests, and the restoration of hydrologic regimes. The State Program on the Rational Utilization of Natural Resources and Environmental Protection designs management plans, research studies, and training programs to be implemented in the reserves. The Vietnam government is also promoting community awareness and cooperation (Duc, 1993; Bentham et al., 1997). In promoting sustainable use of mangrove and melaleuca forests, the government is issuing long-term leases of land resources (owned by a socialist political system) to farmers and fishermen. Under the rental agreement, farmers and fisherman gain use rights while promising to restore and conserve resources. In mangrove forests, 5 to 10 ha are leased and 20 to 30 percent of the area may be used for aquaculture. The remainder of the area must be reforested or preserved. At melaleuca forest sites, 10 ha plots are leased; 7.5 ha must be preserved or replanted and 2.5 ha will be used for permanent agriculture. In addition to the leasing plan, community education is provided to facilitate local participation in the plan. The Soil Science Department of Hanoi State University has provided a plan for sustainable agriculture. The plan outlines a system of rotating crops and intercropping using key native flora species. This system reduces the acids and sulfates in the soils that make cultivation unproductive. Also outlined is a proposed land-use plan that designates areas of protection and human activities (Khoa and Roth-Nelson, 1994).

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To address the potential conflicts with the implementation of hydropower projects, the five countries in the Mekong River basin have formed the Mekong River Commission toward Sustainable Development (Tu, 1996). The Commission has conducted a detailed diagnostic study to assess the development projects in context with socioeconomic structures, political frameworks, and ecological resources. From this study, the Commission has outlined a basic plan of action (Mekong River Commission, 1997). The plan addresses the need to build institutional frameworks for environmental management and policy making, establishing basinwide environmental standards, and strengthening environmental regulatory and enforcement mechanisms. To provide critical support to environmental agencies, training programs will be developed to diversify the workforce. Also, information systems will be improved to provide data on basin ecosystems and resources, and pollution and degradation sources. With this additional support, the development of project plans will incorporate environmental concerns and protection. In an initiative to promote community participation, environmental protection will include programs for land tenure and poverty alleviation. The Commission hopes to approach problems with an adaptive and responsive attitude. The Pantanal The Pantanal encompasses 11 million ha and is the world’s largest freshwater wetland. Much of the Pantanal is in the states of Matto Grosso and Matto Grosse de Sol, Brazil, with lesser parts in the countries of Bolivia and Paraguay (Figure 3). The Pantanal is 100 m above sea level, and situated between the Plateau of Matto Grosso (to the east) and the savannas of Bolivia (to the west). Erosional material is deposited at the bottom of the plateau, making the Pantanal an inland alluvial fan. As a broad, flat basin, the Pantanal acts as a reservoir for water flowing from the east to west and is the headwaters for the Paraguay River. Further downstream, the Parana River meets the Paraguay forming a large waterway that flows through five countries. The final destination of this waterway is the Rio De La Plata at Buenos Aires, Argentina. Annual rainfall in the region averages between 100 to 130 cm, with most rainfall occurring between December and June. The Pantanal helps to regulate water flowing to the Paraguay by acting as a sponge for floodwaters. Over a 6-month period, flood waters are slowly released, maintaining dry season flow. The slow release of floodwaters also creates a lag between the high peaks of both the Paraguay and Parana Rivers, preventing catastrophic floods. Functions and Values As noted above, the Pantanal provides flood control for the Paraguay-Parana river system. In addition, the Pantanal enhances water quality because suspended sediments settle in its calm waters before entering the Paraguay. The Pantanal also has high ecological value in its diverse ecosystems and wildlife (Mittermeier et al., 1990; Gottgens et al., 1998). Habitats include swamp, deciduous forest, savannah, lake margin scrub, and gallery forest. The fauna of the Pantanal includes 230 species ©2001 CRC Press LLC

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of fish, 80 species of mammals, and 50 species of reptiles. Residing in small populations are threatened and endangered species such as the giant otter (Pteronura brasiliensis), giant anteater (Myrmecophaga tridactyla), maned wolf (Chrysocyon brachyurus), and the jaguar (Panthera onca). Over 650 species of birds have been ©2001 CRC Press LLC

observed, including the hyacinth macaw (Anodorhynchus hyancinthinus), the rhea (Rhea americana), and the jabiru stork (Jabiru mycteria). The Pantanal is also a significant staging area for three bird migration routes. Economic activities in the Pantanal include fishing, cattle ranching, and tourism. Cattle ranching is the significant activity, with more than 10,000 ha of grazing land and 10 million cattle. The region has its own cultural heritage based around this industry, with unique music, clothing, language, and food. In part because of its unique culture and diverse ecosystems, the Pantanal is quickly becoming a tourist destination. Up to 10,000 people visit the area annually, and the potential for expanded tourism is great. Another cultural value of the Pantanal is that it is home to many indigenous peoples (Bucher et al., 1993). There are 19 reservations in the area and as many as 12 distinct groups of people. Some of these groups are known for their handicrafts and music, while others are close to extirpation. Threats and Impacts Unsustainable economic activities impact the Pantanal ecosystem (Mittermeier et al., 1990; Bucher et al., 1993). Cattle ranching requires large tracts of land for grazing. Wetlands are drained and forests cut down to create pastures for grazing and other agricultural activities. In addition, cattle trample vegetation, and their selective grazing changes vegetative composition. Native fauna such as puma (Felis concolor), jaguar, and armadillo (Dasypus novemcinctus) may compete for food with cattle, or prey on cattle, and thus are hunted and killed. Another impact is overfishing. Large catches of pacu (Colossoma metrei), catfish (Surubum sp.), cacharra (Pseudoplatystoma fasciatum), and other species are taken from area lakes. Much of the overfishing is attributed to the commercial fishing industry which exports fish to large cities in the region. Some local overfishing occurs as well. Hunting has impacted the populations of many indigenous species. In addition to hunting animals to protect livestock, ranchers hunt to supplement their income. Supplemental hunting increases as the value of beef decreases. Interestingly, subsistence hunting rarely occurs, because most residents rely on fish and beef for food. Sport hunting and hunting for skins have made the greatest impacts on native fauna. Otter, ocelot (Felis pardalis), jaguar, and caiman (Caiman crocodilus) are hunted for their skins. Tourists often participate in sport hunting, shooting birds, caiman, and various mammals from roadsides. Apparently passive recreation is also accompanied by impacts (Mittermeier et al., 1990; Junk, 1993). Visitors often disturb wildlife by throwing stones at caiman, or scaring birds into flight to create a photographic opportunity. Little infrastructure or organization exists to effect public education. Other impacts to the Pantanal include siltation of wetlands and waterways as the result of deforestation, and pollution from agricultural activities and mining. Also, wildlife losses occur through the international skin and animal trade (Bucher et al., 1993).

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The greatest threat to the Pantanal is the proposed waterway project known as Hidrovia (Bucher et al., 1993; Gottgens et al., 1998). The Hidrovia project is a system of channels designed to enhance navigation of the Paraguay and Parana Rivers. This enhancement would increase the transport of goods to ports throughout the region, especially in Argentina, Paraguay, and Brazil. The project engenders significant changes to waterway morphology, including bottom dredging, channel straightening, and waterflow regulation. The potential impacts to the Pantanal are not completely understood and could be devastating. The most significant impact would be a reduction in flood storage capacity. With the straightening of river channels, water will flow at a higher velocity, thereby causing the Pantanal to drain. Ironically, this may make the rivers difficult to navigate. Dredging activities will increase turbidity, decreasing water quality and impacting aquatic insects and fish populations. Other likely impacts include the loss of ecological and biological diversity because energy and food web relationships will be altered. Destruction of the wetlands could result in enormous social costs because many regional economic and cultural activities rely on the Pantanal’s ecological resources. Conservation Efforts The Pantanal is listed as a Wetland of International Importance. However, 95 percent of the wetland is privately owned (Bucher et al., 1993). Only three areas are designated as reserves: Pantanal National Park (135,000 ha), the Federal Reserve of Cara-Cara (61,126 ha), and the Taima Ecological Station (12,000 ha). Concerned landowners have formed the Pantanal Defense Society to encourage sustainable practices in the region (Mittermeier et al., 1990). Issues being addressed include wise land use, erosion control, pollution reduction, and improved ranching practices. The Rural Workers’ Union, state governments, scientific research organizations, and environmental agencies have met to discuss sustainable use of key wildlife species. Strategies include caiman farming to reduce poaching and fish farming. Ranching practices are being evaluated to reduce hunting. To implement the Hidrovia project, the basin countries formed the Intergovernmental Committee on the Hidrovia (CIH, Gottgens et al., 1998). As part of development plans, the CIH conducted an economic feasibility study. However, the study may have inaccurately evaluated the environmental costs of the Hidrovia project. A technical panel of reviewers concluded that the study failed to fully estimate significant environmental losses to such industries as fishing and cattle ranching, and that it overestimated the project benefits such as soybean exports and iron ore values (Gottgens et al., 1998). In addition, the panel stated that the study failed to provide information on existing transportation routes and plans that might be more environmentally benign. Finally, the panel criticized the initial study for the lack of input from indigenous cultures and the lack of information on the long-term environmental effects of the project. In response, the InterAmerican Development Bank and the United Nations Development Program contributed US $10 million in technical assistance to conduct assessments of the engineering, economic, and environmental costs of the project. ©2001 CRC Press LLC

Brazil and Paraguay have, or are considering, abandoning the Hidrovia Project because of questionable economic benefits (Bucher et al., 1993; Gottgens et al., 1998). In 1998, Brazil’s Federal Environmental Agency abandoned the proposed construction of the waterway along its border. The national courts terminated all ongoing studies for the construction of Hidrovia, instead focusing on smaller nonstructural improvements. Paraguay is also considering abandoning its portion of Hidrovia. Consultation with the U.S. Army Corps of Engineers has provided insight into past mistakes associated with large channelization projects in the United States, such as those on the Mississippi River and in South Florida. However, Bolivia and Argentina are continuing to dredge new channels to create waterways for commercial use. The Wadden Sea The Wadden Sea stretches from the north coast of The Netherlands, along Germany’s coastline to the Skallingen Peninsula in Denmark (Figure 4). Encompassing 1,350,570 ha, the shallow sea has a wide diversity of ecosystems. Large bays exist where the rivers meet the sea. Tidal channels, mud flats, salt marshes, beaches, and dunes exist where the sea meets the land. Numerous barrier islands and sand bars protect the coastline from the harsh North Sea. Tidal flats comprise two thirds of the Wadden Sea (Frazier, 1996; Common Wadden Sea Secretariat, 1997). Diurnal tides bring 1500 ha of water from the North Sea into the Wadden, doubling its size. Sand and silt carried with the tide settle in the calmer waters of the Wadden creating tidal flats that are exposed at low tide. As the largest stretch of tidal flats in the world, the Wadden Sea accounts for 60 percent of all tidal areas in Europe and North Africa. Functions and Values The productivity of the Wadden Sea can be compared with that of tropical rain forests (Common Wadden Sea Secretariat, 1997). Shallow water and high nutrient levels result in high primary productivity. Phytoplankton and algae support a great abundance and diversity of bird, mammal, and fish species. The environment of the Wadden Sea is highly variable, with extreme temperatures, salinities, and water levels. Indigenous species have adapted to these extremes, and there are more than 250 endemic species and ecotypes. The Wadden Sea is a site of international importance for waterfowl species. Documented are 52 distinct populations of 41 species. Of the individuals from 20 populations, one half use the area during a stage of their annual life cycle and 10 species are endemic. Approximately 10 to 12 million birds stop in the area to rest and feed during migration (Davis, 1993; Dugan, 1993; Frazier, 1996). The Wadden Sea is also habitat for marine mammals such as the harbor seal (Phoca vitulina), the gray seal (Halichoerus grypus), and the bottlenose dolphin (Tursiops truncatus). In addition, the Wadden Sea is a spawning and migration ground for 102 species of fish, with 34 designated as rare or extremely rare. Many of these species are found only in the sea during certain times of the year. Large ©2001 CRC Press LLC

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percentages of major commercial species such as plaice (Pleuronectes platessa), sole (Solea solea), and North Sea herring (Clupea harengus) mature in the Wadden Sea, making the sea economically important to the region. Other economic values of the Wadden Sea are tourism and recreation. Currently, 30 to 40 million people a year visit the Wadden Sea (Davis, 1993). The tourism industry has become one of the most important economic inputs into the area, with some barrier islands completely dependent upon tourism for economic stability. Threats and Impacts The most obvious impact to the Wadden Sea is the construction of embankments, dikes, and shipping ports (Enemark, 1993; Common Wadden Sea Secretariat, 1997). Construction of embankments have reduced the number of bays and impeded the natural migration of barrier islands and sandbars. Collectively, these infrastructure projects have resulted in the loss of 16,000 ha of land area, almost half of what now remains. This loss reduces natural habitat and accentuates the difference between low and high tides, thereby creating the potential for the loss of coastline from erosion and submersion. ©2001 CRC Press LLC

Pollution from rivers that discharge into the Wadden Sea have impacted four of five estuaries in Denmark (Davis, 1993; Dugan, 1993). Only the Verde Ao in Denmark exists in a fairly natural state, while the Ems, Weser, Elbe, and Elder introduce high concentrations of nutrients and contaminants into the Wadden Sea. Contaminants include polychlorinated biphenyls (PCBs) and pesticides that enter food chains and weaken the immune and productive systems of marine mammals. High concentrations of nutrients such as nitrogen and phosphorus contribute to large algae blooms and the production of toxic algae such as those associated with red tides. Contaminants are also transported in tidal waters from the North Sea (Dugan, 1993). Dredge material containing heavy metals are dumped offshore, carried by currents, and deposited into the Wadden Sea. Atmospheric processes deposit nitrogen and cadmium in the Wadden Sea. Other impacts are attributed to commercial activities such as shellfish farming, hunting, and tourism. Shellfish harvest and culture beds impact the ecosystem by changing species composition, food competition between organisms, nutrient balances, and species abundance (Frazier, 1996). The historic hunting of seals has stressed already low populations trying to recover from viral epidemics, although a ban on seal hunting has increased the population from 3,600 to 10,000 (Common Wadden Sea Secretariat, 1997). Increased tourism has led to the construction of additional embankments and disturbance to wildlife. Peak tourism time, May to September, is critical to seal reproduction and pup growth. Disturbance by sightseers on sandy beaches during nursing time reduces fat reserves in seal pups and, hence, their chance for survival. Conservation Efforts To protect the Wadden Sea and its resources, initiatives must encompass the policies of several nations. As an important regional resource which is impacted by activities of several parties, cooperation on a international level is addressed through the Ramsar Convention using guidelines provided in Towards the Wise Use of Wetlands program. From 1978 to 1982, protective measures were hindered by intercountry differences in policies and administration (Davis, 1993; Common Wadden Sea Secretariat, 1997). In 1982, cooperation for protection was established through the Joint Declaration on the Protection of the Wadden Sea. In 1991, the wise use concept was applied to the framework for this trilateral cooperation, and the protection of the Wadden Sea has become the foremost example of the Ramsar Convention’s Wise Use program. Fundamental to the program and the Joint Declaration is a goal of achieving, as much as possible, a natural and sustainable ecosystem in which natural processes proceed in an undisturbed way. The framework describes the actual situation of the Wadden Sea including ecological values, human uses, and potential and existing threats. A reference situation is provided evaluating the full potential values of the ecosystem. The target of these values includes an inventory of the diverse habitat types, the greatest potential integrity of these habitats, protection of organisms, and high water quality. To establish these values, scientific, political, social,

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and cultural parameters are addressed and incorporated into management principles and monitoring and research programs. The management principles mediate between actual policies and human uses. Outlined are seven management principles. They include informed decision-making, best technological and environmental practices, zoning based on ecosystem fragility, and mitigation or restoration where activity cannot be avoided. In applying the management principles, international objectives have been established. Embankments and any other sea defense that would encroach on tidal flats are prohibited. In addition, protection for saltmarsh and dunes, with emphasis on flora and fauna, will be increased. This includes a ban on pesticides, fertilizers, and other toxins. Large areas, including inter- and subtidal zones, will be closed to mussel fishing. Zones of recreation will be established based on the sensitivity of the location and type of activity. For example, hovercrafts and jet-scooters are completely prohibited, while recreational boats and ships are restricted to designated zones. Hunting of migratory birds will be phased out, and hunting of nonmigratory species will be allowed based on the stability of the population. To reduce pollutant discharges, a 50 percent reduction in fertilizer, and up to a 70 percent reduction of some metals is proposed. Direct dumping and offshore discharge will be prohibited, and ports will be upgraded to minimize ship wastes (Enemark, 1993). Monitoring and research programs will assess the progress of these management initiatives, while broadening the knowledge on which to base new strategies and principles. In accordance with the wise use guidelines, each nation has designated national parks, wildlife refuges, and other protected areas within the Wadden Sea. The Netherlands established the Wadden Sea Memorandum, a national planning document that incorporates national, state, and regional administration of management strategies. Areas covered by the memorandum are nature preserves, where any activity that may damage flora, fauna, or scenic views is prohibited. In Germany, states are given authority to establish national parks through the Federal Nature Conservation Act. The parks are part of a zoning system. Zone I areas are heavily regulated, and may be totally restricted from public access. Zone II areas only allow activities which will not degrade ecosystem integrity. In Denmark, the Wadden Sea in its entirety is designated as a nature and wildlife reserve by the Statutory Order of 1982. It prohibits any activities that change or destroy the natural environment. Public access to important seal and bird areas is prohibited. In other areas, boating and recreational activities are regulated. Shell fishing is restricted in major parts of the tidal zone and permitted in shipping routes and waters offshore of the islands. Under the Ramsar Convention, the preserved and restricted areas of the Wadden Sea are listed as sites of international importance. Through the wise use program, Denmark, the Netherlands, and Germany are guided by common objectives to protect the integrity of the Wadden Sea. Every 3 years the Trilateral Governmental Wadden Sea Conference is held and attended by responsible ministries from each nation. The ministries include the Minister of Agriculture, Nature Management in Fisheries, The Netherlands; the Federal Minister for the Environment, Nature Conservation and Nuclear Safety, Germany; and the Minister for the Environment and Energy,

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Denmark. At the conferences, ministers discuss implementation and progress, create additional protective measures, and assess the current quality of the Wadden Sea.

REFERENCES Barbier, E. B., Acreman, M., and Knowler, D., Economic Valuation of Wetlands: A Guide for Policy Makers and Planners, Ramsar Convention Bureau, Gland, Switzerland, 1997. Beilfuss, R. D. and Barzen, J. A., Hydrological wetland restoration in the Mekong Delta, Vietnam, in Global Wetlands: Old World and New, Mitsch, W. J., Ed., Elseveir Science, Amsterdam, 1994, 453. Bentham, W., Chuyen, N. D., van Lavieren, L. B., and Verheught, W. J. M., Rehabitating the Mangrove Forests of the Mekong Delta, Ecoconsult, The Netherlands, 1997. Bouton, S. N., Frederick, P. C., Spalding, M. G., and McGill, H., Effects of chronic, low concentrations of dietary methlymercury on the behavior of juvenile Great Egrets (Ardea albus), Environ. Toxicol. Chem., 18(9), 1934, 1999. Bucher, E. H., Bonetto, A., Boyle, T., Canevari, P., Castro, G., Huszar, P., and Stone, T., Hidrovia, an Intitial Environmetnal Examination of the Paraguay-Parana Waterway, Wetlands for the Americas, Manomet, MA, and Buenos Aires, Argentina, 1993. Chape, S. and Inthavong, C., Protected Areas, Biodiviersity Conservation and the Development Imperative in Lao PDR: Forging the Links, draft report, IUCN World Conservation Unit, 1996. Chimney, M. J., Nungesser, M., Newman, J., Pietro, K., Germain, G., Lynch, T., and Moustafa, M. Z., Stormwater treatment areas—status of research and monitoring to optimize effectiveness of nutrient removal and annual report on operational compliance, in 2000 Everglades Consolidated Report, Draft, South Florida Water Management District, 1999, 6-1. Common Wadden Sea Secretariat, The Trilateral Cooperation on the Protection of the Wadden Sea, Ramsar Convention Bureau, 1997. Davis, T. J., Ed., Towards the Wise Use of Wetlands: Report of the Ramsar Convention Wise Use Project, Ramsar Convention Bureau, Gland, Switzerland, 1993. Davis, T. J., Ed., The Ramsar Convention Manual: A Guide to the Ramsar Converntion on Wetlands of International Importance, Ramsar Convention Bureau, Gland, Switzerland, 1994. Derr, M., Redeeming the Everglades, Audubon, September/October, 48 and 128, 1993. Duc, L. D., Rehabilitation of the melaleuca floodplain forests in the Mekong Delta, Vietnam, in Towards the Wise Use of Wetlands: Report of the Ramsar Convention Wise Use Project, Davis, T. J., Ed., Ramsar Convention Bureau, Gland, Switzerland, 1993, 140. Dugan, P., Ed., Wetlands in Danger: A World Conservation Atlas, Oxford University Press, New York, 1993. Enemark, J., Wise use of the Wadden Sea, in Towards the Wise Use of Wetlands: Report of the Ramsar Convention Wise Use Project, Davis, T. J., Ed., Ramsar Convention Bureau, Gland, Switzerland, 1993, 25. Ferriter, A., Thayer, D., Laroche, F., Bodle, M., and Davis, S., Exotic plants in the Everglades, in 2000 Everglades Consolidated Report, Draft, South Florida Water Management District, 1999, 14-1. Fink, L., Rumbold, D., and Rawlik, P., The Everglades mercury problem, in 2000 Everglades Consolidated Report, Draft, South Florida Water Management District, 1999, 7-1.

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Florida Panther Interagency Committee, Status report. Mercury Contamination in Florida Panthers, Technical Subcommittee, Florida Game and Fresh Water Fish Commission, U.S. Fish and Wildlife Service and Everglades National Park, December 1989. Frazier, S., compiler, Directory of Wetlands of International Importance—an Update, Ramsar Convetion Bureau, Gland, Switzerland, 1996. Frederick, P. C., Spalding, M. G., Sepulveda, G. E., Williams, M. S., Jr., Lorazel, S. M., and Samuelson, D. A., Exposure of great egret (Ardea albus) nestlings to mercury through diet in the Everglades ecosystem, Environ. Toxicol. Chem., 18(9), 1940, 1999. Gottgens, J. F., Fortney, R. H., Meyer, J., Perry, J. E., and Rood, B. E., The case of the Paraguay-Parana Waterway (Hidrovia) and its impact on the Pantanal of Brazil: a summary report to the Society of Wetlands Scientists, Wetlands Bulletin, The Society of Wetlands Scientists, September, 1998, 12. Junk, W. J., Wetlands of tropical South America, in Wetlands of the World: Inventory, Ecology, and Management, Whigham, D. F. et al., Eds., Kluwer Academic Publishers, The Netherlands, 1993, 679. Khoa, L.V. and Roth-Nelson, W., Sustainable wetland use for agriculture in the Mekong River delta of Vietnam, in Global Wetlands: Old World and New, Mitsch, W. J., Ed., Elsevier Science, Amsterdam, 1994, 737. Lanza, G. R., A Review of Nam Leuk Hydropower Development Project Environmental Impact Assessment Final Report, International Rivers Network, Berkeley, CA, 1996. Lohmann, L., Remaking the Mekong, Ecologist, 20(2), 61, 1990. McCormack, P. V., Newman, S., Miao, S., and Fontaine, T. D., Ecological needs of the Everglades, in 2000 Everglades Consolidated Report, Draft, South Florida Water Management District, 1999, 3-1. McLean, A. R. and Bush, E. L., Central and south Florida restudy, in 2000 Everglades Consolidated Report, Draft, South Florida Water Management District, 1999, 10-1. Mekong River Commission towards Sustainable Development, Annual Report, Mekong River Commission Secretariat, Bangkok, Thailand, 1995. Mekong River Commission towards Sustainable Development, Mekong River Basin Diagnostic Study: Final Report, Mekong River Commission, Bangkok, Thailand, 1997. Mittermeier, R. A., De Gusmao Camara, I., Padua, M. T. J., and Blanck, J., Conservation in the Pantanal of Brazil, Oryx, 24 (2), 103, 1990. Osborne, M., Exploring the Mekong’s past and present, Asia Pac. Mag., 2, 26, 1996. Ramsar Convention Bureau, About the Ramsar Convention; Website: http://ramsar, 1999. Redfield, G., Goforth, G. F., and Rizzardi, K. W., Major findings and preliminary implications of the 2000 Everglades Consolidated Report, in 2000 Everglades Consolidated Report, Draft, South Florida Water Management District, 1999, 1. Robinson, G. B., Robinson, S. C., Lane, J., Nelson, D., Hveem, L., and Carrasco, N., Discover a Watershed: The Everglades, produced by the Watercourse for the South Florida Water Management District, Bozeman, MT, 1996. Tu, T. D., Sustainable development in the Mekong River Basin, Land Lines, 8(3), 1996; Website: www.lincolninst.edu/landline/1996/may/mekong2. Wegner, D. L., Review comments on Nam Theun 2 hydrolelectric project environmental assessment and management plan, International Rivers Network, Berkeley, CA, 1997. Whalen, B., P. Whalen, T., and Kosier, T., Effectiveness of best management practices., in 2000 Everglades Consolidated Report, Draft, South Florida Water Management District, 1999, 5-1.

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White, W. C., Review of Greater Mekong Task Force Strategies, International Rivers Network; Website: htp:www.irn.org, 1997. World Conservation Monitoring Centre, Directory of Wetlands of International Importance: Sites Designated for the List of Wetlands of International Importance, Ramsar Convention Bureau, Gland, Switzerland, 1990.

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Ripple, Karen L. “Wetland Education” Applied Wetlands Science and Technology Editor Donald M. Kent Boca Raton: CRC Press LLC,2001

CHAPTER

14

Wetland Education Karen L. Ripple

CONTENTS School System Requirements Science Curriculum Science Content Standards Multidisciplinary Units Student Action Projects Service Learning Credits Wetland Program Features Important to Educators Lesson Plan Format Instructional Objectives Multiple Intelligences Hands-On Learning Cooperative Learning Performance-Based Instruction Evaluating Wetland Programs and Activities Wetland Educational Programs for Grades K–12 National Education Programs WOW!: The Wonders of Wetlands Project WILD Aquatic Education Activity Guide Children’s Groundwater Festival Discover Wetlands A World in Our Backyard: A Wetlands Education and Stewardship Program Wading into Wetlands Student Wetlands Action Projects Schoolyard Habitats Program

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Wicked Big Puddles: A Guide to the Study and Certification of Vernal Pools Handbook for Wetlands Conservation and Sustainability WILD School Sites and Taking Action POW!: The Planning of Wetlands Grant Funding Membership Organizations Wetland Education for Professionals Certification University Continuing Education Independent Professional Training Environmental Concern Inc Wetland Training Institute Richard Chinn Environmental Training Institute for Wetland and Environmental Education and Research Membership Organizations References

In the United States, public perception of wetlands as wastelands is slowly evolving into recognition of wetlands as productive, valuable natural resources (Tiner, 1998). Laws and regulations passed in the last three decades are beginning to curb wetland destruction previously encouraged through Congressional Swamp Land Acts (Kusler and Opheim, 1996). Wetland education is slowly evolving in response to changing attitudes. During the 1960s, college level ecology courses included wetland studies. Wetland education materials for students in grades kindergarten through twelve (K–12), however, were developed much later. Wetland activities within the newly developed environmental study units, and isolated wetland activities connected to regional issues, gradually appeared in classrooms. As the value of wetlands to society increased, the need to include wetland material within our educational programs was finally recognized and acted upon in the late 1980s. Now the field of wetland education is rapidly expanding, and resources of all kinds are readily available to educators. This chapter discusses educational approaches and programs applicable to K–12 wetland educators and identifies key elements of their success. Educational options for wetland professionals are also considered. The chapter reviews national education programs containing activities that can easily be integrated into current K–12 curricula. The programs require minimal teacher preparation. Current educational emphasis on developing student problem-solving skills leads naturally to programs encouraging the creation, restoration, enhancement, and monitoring of schoolyard wetlands. These newer programs are also reviewed. Model programs for student participation in wetland certification are examined next, as students learn to take action to protect our vanishing wetlands. Some funding sources for wetland education projects and ©2001 CRC Press LLC

teacher training are discussed as well as organizations that effectively disseminate information on wetlands. Finally, professional courses for those seeking to improve their techniques, skills, and knowledge in relation to wetlands are examined. Ways of staying abreast of issues, methods, and information in the field of wetlands are also suggested. Undergraduate education and graduate research programs on wetlands, while important to increasing our wetland knowledge base, are not discussed herein. Colleges and universities can more readily supply current information on available courses and programs if contacted directly. Methods of disseminating knowledge are constantly evolving. While research generates knowledge through experimentation and exploration, applied science refines the techniques and tools. The fruits of both research and applied science are transmitted to college students through a combination of lecture and laboratory work. While the lecture/laboratory approach to teaching has been traditionally linked to college courses and college bound student programs, this approach is not always effective. Younger students, and those not college bound, tend to respond to more active educational approaches that have practical applications to their own lives. Primary and secondary school students may not understand the relationship between wetlands and their lives. Alternatively, they may feel wetland losses are too large a problem with which to cope. Consequently, teaching techniques must be continually developed and refined to better fit the capabilities, interests, and needs of younger students as well as the broader community. Those outside the education field may not realize the importance of meeting the educational requirements of school systems as well as the needs of individual educators. Both are important to gain acceptance for a wetland program or activity. Providing something that young students like is not enough to gain access to the classroom.

SCHOOL SYSTEM REQUIREMENTS School systems establish the science curriculum based upon national science content standards or more stringent local standards. Increasingly, curricula include multidisciplinary units and student action projects. In addition, some school systems require students to earn service learning credits as a requirement for graduation. A closer look at these requirements and how they affect new wetland programs is warranted. Science Curriculum A science curriculum includes the science courses that will be available to students in a particular grade and what unit topics will be covered within those courses. The curricula of a school system are set by the superintendent, the curriculum specialists (e.g., science supervisor etc.), and the school board. Some systems allow teacher input. If a school system decides that wetland units will be taught in the second and eighth grades, then wetland materials targeting sixth graders will not likely be considered for use by that school system. ©2001 CRC Press LLC

State, county, district, city, or some combination of these bodies organizes school systems. To understand the science curriculum of a particular school system, begin at the state level by determining what is mandated statewide, then contact appropriate local school boards to learn how state mandates are applied. If wetland programs or materials meet the state level curriculum criteria, they are more likely to be acceptable at the local level. Table 1 provides one example of a school science curriculum and the units covered in some courses. Table 1 Grade

An Example of Part of a District Science Curriculum Course or Class

1

Science

2

Science

3

Science

4

Science

5

Science

6

Science

7

Science

8

Science

9

Environmental Science

9

Biology

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Units or Topics Seeds and plants Earth, dinosaurs, and space Magnets Schoolyard habitat Animal classification Weather Sink or float Wetlands Insects Earth Machines Forestry Ecosystems Atmosphere Electricity Recycling Soils and plants Rocks and minerals Estuaries Physiology Dangerous storms Light Ecology Zoology Astronomy Force and motion Environmental issues Genetics Dynamic earth Acids and bases Global energy Environmental concepts Biomes of North America Terrestrial ecosystems Wetland ecosystems Classification Human anatomy and physiology Plant structures and processes Ecosystems, populations, and communities

Science Content Standards Science content includes scientific knowledge, understanding, and abilities—the essential material contained within the units of a science course. Science content standards dictate the knowledge students should acquire through their science studies. In 1995, the National Research Council of the National Academy of Sciences developed national science standards. The standards are grouped into three grade levels and eight categories, all of which might include some aspect of wetland science (Table 2). The three grade levels are K–4, 5–8, and 9–12. Standards in the first two categories, Unifying Concepts and Processes and Science as Inquiry, are consistent across all grade levels because these lifelong processes are basic for an understanding of the natural world. In the remaining six categories, the middle school standards build on those of the primary grades, and the high school standards build on those of the middle school level. The categories of Physical Science, Life Science, and Earth and Space Science focus on the facts, concepts, principles, theories, and models within each subject area. The Science and Technology category links the natural world and the designed world, with many parallels to the Science as Inquiry category. The category of Science in Personal and Social Perspectives concentrates on decision-making skills in personal and social issues. The final category, History and Nature of Science, reflects the change of science through time and the influence of science on world cultures. These national standards have been adopted unchanged by many educational systems. State and local school systems set their own standards, which usually are more stringent and more detailed than the national standards, and often reflect local needs and issues. Designers of science programs, including those about wetlands, should consider cross-referencing the national standards with their activities, thereby relieving educators of that necessity. Multidisciplinary Units Multidisciplinary programs, projects, or study units incorporate more than one subject or content area. Currently in favor with school systems, this approach blends boundaries between subjects and promotes teamwork among teachers. Multidisciplinary units also remind students that subject areas do not divide the world beyond the school walls. One intriguing aspect of wetlands is that it is not just a science topic, although many treat it as such. Wetlands and water can form a multidisciplinary theme for an entire school or link activities within an entire grade level. Thanks to innovative teachers and a supportive principal at Thomas Jefferson High School for Science and Technology in Fairfax, VA, students there have been studying natural wetlands during integrated freshman biology, language arts, and technology courses. Recently, students began creating research wetlands within a school courtyard and soon will determine effective ways to enhance nearby natural wetlands. Additional courses are expected to shift focus to the wetland and water theme in the future, eventually involving the entire school body in wetland activities.

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Table 2

National Science Content Standards Grades K–4

Grades 5–8

Grades 9–12

Unifying Concepts and Processes Systems, order, and organization Evidence, models, and explanation Change, constancy, and measurement Evolution and equilibrium Form and function

Systems, order, and organization Evidence, models, and explanation Change, constancy, and measurement Evolution and equilibrium Form and function

Systems, order, and organization Evidence, models, and explanation Change, constancy, and measurement Evolution and equilibrium Form and function

Science as Inquiry Abilities necessary to do scientific inquiry Understandings about scientific inquiry

Abilities necessary to do scientific inquiry Understandings about scientific inquiry

Abilities necessary to do scientific inquiry Understandings about scientific inquiry

Physical Science Properties of objects and materials Position and motion of objects Light, heat, electricity, and magnetism

Properties and changes of properties in matter Motions and forces Transfer of energy

Structure of atoms Structure and properties of matter Chemical reactions Motions and forces Conservation of energy and increase in disorder Interactions of energy and matter

Life Science Characteristics of organisms Life cycles of organisms Organisms and environments

Structure and function in living systems Reproduction and heredity Regulation and behavior Populations and ecosystems Diversity and adaptations of organisms

The cell Molecular basis of heredity Biological evolution Interdependence of organisms Matter, energy, and organization in living systems Behavior of organisms

Earth and Space Science Properties of earth materials Objects in the sky Changes in earth and sky

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Structure of the earth system Earth’s history Earth in the solar system

Energy in the earth system Geochemical cycles Origin and evolution of the earth system Origin and evolution of the universe

Table 2 (continued) National Science Content Standards Grades K–4

Grades 5–8

Grades 9–12

Science and Technology Abilities of technological design Understandings about science and technology Abilities to distinguish between natural objects and objects made by humans

Abilities of technological design Understandings about science and technology

Abilities of technological design Understandings about science and technology

Science in Personal and Social Perspectives Personal health

Personal health

Characteristics and changes in populations Types of resources Changes in environments Science and technology in local challenges

Populations, resources, and environments Natural hazards Risks and benefits Science and technology in society

Personal and community health Population growth Natural resources Environmental quality Natural and human-induced hazards Science and technology in local, national, and global challenges

History and Nature of Science Science as a human endeavor

Science as a human endeavor Nature of science History of science

Science as a human endeavor Nature of scientific knowledge Historical perspectives

Adapted from the National Research Council, 1995. With permission.

Student Action Projects In some study units, students are encouraged, and sometimes required, to take action in a way that will make a positive difference in their school, community, or environment. These are often called student action projects. Wetland monitoring, protection, and creation are often the focus of these student environmental action projects. Removing trash from a wetland, raising funds to purchase wetland plants for a restoration project, planting a degraded wetland, and stenciling storm drains to indicate that they empty into a wetland are examples of student action projects (Figure 1). Service Learning Credits In an increasing number of states, students must participate in a minimum number of hours of community service as a requirement for high school graduation. Called service learning credits, students also reflect and communicate what was ©2001 CRC Press LLC

Figure 1

In student action projects, students make a positive difference in their school, community, or environment. Third grade students planting a small constructed wetland at Horsehead Wetland Center is but one example of the many possibilities for positive change that can empower our youth.

learned in providing the service. A coordinator within the school system suggests existing community service activities available to students and coordinates new community service projects. Wetland restoration, creation, monitoring, and protection projects usually provide opportunities for student service learning credits.

WETLAND PROGRAM FEATURES IMPORTANT TO EDUCATORS What features do educators look for when they evaluate new programs? Educators save precious time if activities are presented in a lesson plan format with clearly stated instructional objectives. Activities that encourage use of multiple intelligences and a variety of learning styles, such as hands-on learning, cooperative learning, or performance-based instructional techniques, are favored over the lecture/laboratory technique often used by colleges. A closer look at the needs of educators is necessary to understand how effective wetland programs are designed. The best programs will have many of the features described next. Lesson Plan Format A lesson plan indicates what an educator intends to accomplish with a lesson and how he or she intends to accomplish it. Most school systems require teachers to write detailed daily lesson plans before teaching a class and to have those plans ©2001 CRC Press LLC

at hand during the lesson. Many teachers are required to submit daily lesson plans to a supervisor for approval before the lesson is taught. Lesson plan formats vary from system to system, but minimally contain an objective, an activity and/or assignment, and some type of student assessment to determine if the objective was accomplished. Programs that are organized with lesson plan formats are easier for teachers to incorporate into their plans. Instructional Objectives Instructional objectives are statements expressing what the student is expected to accomplish with an assignment and are an important part of the lesson plan. Statements of objectives, or learning outcomes, have two essential parts: the action verb and the content. The action verb indicates the skill to be achieved, such as “measure.” The content indicates the knowledge to be gained, such as “water temperature.” Simple, specific wording of the objective allows students to clearly understand what is expected of them and eases teacher determination of whether the objective has been accomplished. For example, “students will appreciate wetlands” is a vague and immeasurable objective. “Imitate the sounds or motions of your favorite wetland creature” is more specific, more measurable, and therefore more useful as a guide for both teachers and students. Bloom’s taxonomy is a hierarchy of instructional objectives that build upon each other, extending from simple to complex thinking processes, and from concrete to abstract. The basic levels of the hierarchy from simple to complex are knowledge, comprehension, application, analysis, synthesis, and evaluation (Bloom, 1956). Table 3 summarizes the objectives of each level, the type of thinking skill required to achieve the objective, and action verbs consistent with those thinking skills. Table 3

Verbs for Instructional Objectives (Bloom, 1956)

Thinking Skill Level

Objective

1. Knowledge

Recall information

2. Comprehension

Understand and interpret material

3. Application

Use material in a new situation

4. Analysis

Examine parts and relationships

5. Synthesis

Rearrange parts to form a new idea, plan, or relationship Judge material based on evidence

6. Evaluation

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Action Verbs Consistent with Objective Circle, define, designate, determine, identify, label, list, mark, match, name, select, specify, state, underline Condense, describe, explain, interpret, outline, restate, rewrite, summarize, trace, translate Build, construct, demonstrate, draw, illustrate, make, measure, model, operate, show, solve, use Analyze, classify, compare, contrast, debate, diagram, differentiate, explore, graph, organize, monitor, specify, test Compose, construct, create, design, develop, establish, invent, plan, predict, produce, suggest, write Assess, choose, compare, conclude, decide, evaluate, grade, judge, justify, rank, select, support, value

Within a study unit, action verbs used in the objective should reflect increasingly higher thinking levels as student skills develop. For instance, when a topic is first introduced a student might be expected to “name the three parameters utilized in wetland delineation,” a level 1 skill. After some study, the same student might be able to “describe six characteristics of hydric plants,” which is a level 2 skill. Later the student could be expected to demonstrate a level 3 skill, for example, “illustrate or list the hydric characteristics of a cattail.” A student who could “classify plants as hydric or nonhydric” is using level 4 skills. “Design a wetland planting plan” utilizes thinking skills of level 5, and to “assess the success of a restored wetland” would involve level 6 skills. Multiple Intelligences Traditional IQ tests primarily measure language and logic. Students weak in these areas may excel in other aspects and still be creditable students. Now recognized by educators are other types of intelligence (Lazear, 1999). Originally proposed as a theory by Howard Gardner (1983), these multiple intelligences and their characteristics now include the eight intelligences summarized in Table 4. Table 4

Multiple Intelligences (Gardner, 1983)

Intelligence Verbal/linguistic Logical/mathematical Visual/spatial Musical/rhythmic Bodily/kinesthetic Interpersonal Intrapersonal Naturalist

Description Use of words and language Reasoning, use of numbers, relationship and pattern recognition, analysis, problem solving Use of art and imagination, creation of mental pictures Recognition of rhythmic and tonal patterns, sensitivity to sounds Use of the body in physical motion Communication person-to-person Use of self-knowledge Recognition of the parts of the natural environment

Recognizing these multiple intelligences, educators frequently provide for a variety of learning styles in programs and activities, thereby helping students more readily achieve their objectives (Figure 2). Educators and their students readily accept wetland education programs that both allow, and encourage, students to use all of their talents. Hands-On Learning Hands-on learning activities actively involve students in the learning process. Effective at all ages, this teaching technique works especially well with younger students, hyperactive students, and those with short attention spans (Figure 3). Students handle and manipulate objects other than papers and pencils. Usually students are out of their seats for at least part of the activity, which often occurs outdoors. This is an important way for many students to learn. It has long been known in education that some knowledge is retained if it is simply heard, more is retained if ©2001 CRC Press LLC

Figure 2

Educators use multiple intelligences to role-play salt marsh organisms during a WOW!: The Wonders of Wetlands teacher training workshop. Notice the rising tide, waving sea grasses, swimming fish, and a tubeworm in her black trash bag “tube.”

it is heard and seen, but when a student can manipulate the material in some fashion, then they usually own it. For example, hearing someone talk about fishing can be informative. Watching a fisherman fish is more helpful, but actually fishing is handson learning. Each person has a unique learning style. Hands-on activities fit with many personal learning styles, and are effective with most students. Wetlands activities, such as assessing habitat utilization by birds or amphibians, are often well suited to hands-on teaching techniques. Cooperative Learning Students working together to complete an assignment in small learning groups of mixed ability are practicing cooperative learning. In this teaching technique, students assume a greater responsibility for learning and for helping each other to accomplish the objectives of the activity (Kagan, 1997). The teacher allocates students to each group in a manner that will achieve a high level of heterogeneity. Each member of the group is assigned a job, such as facilitator, reporter, or recorder (Figure 4). The number of jobs required to complete the task as a team determines the number of students per group. The teacher either assigns jobs within a group or, if students select their job, it must be different than during the last activity. Teachers structure the activity so that the objectives, and the steps to achieve the objective, are clear to students. Cooperative procedures and skills are taught to group members ©2001 CRC Press LLC

Figure 3

During hands-on learning, educators determine ground elevations along a transect using meter sticks, string, and a line level. The data gathered will be converted into a topographic map of the wetland site. Use of both manual and intellectual skills in this learning technique can be much more effective than simply examining a prepared topographic map.

prior to the activity and then monitored while groups work on an assignment. Each member of a group receives the same grade on an activity, so all benefit from working together to achieve more than any one person could alone. During a cooperative learning assignment, a four-member group might have a facilitator, a supply manager, a recorder, and a reporter. The facilitator ensures that the group follows each step correctly and in sequence, stays on task, and meets deadlines. The supply manager obtains the needed materials at the appropriate time and returns them when and where appropriate. The recorder writes data, observations, or information collected by the group on the appropriate forms or in an appropriate format. The reporter coordinates writing the group report. This teaching technique requires much more preplanning by teachers but can be extremely effective. When the technique is successful, teachers may not appear to be busy during an assignment, because in essence, they become consultants to their students. Team skills learned by students are directly applicable to the real job world and are skills valued by many employers. Specifically, students focus on assignments and skills needed to successfully work together. They take responsibility for themselves and the success of their group, with students encouraging each other to the benefit of all. Ideally, with no group leader, leadership responsibilities are shared or alternated. In practice, this is probably the most difficult aspect of cooperative learning to achieve and the most important for team success. Many wetland activities effectively use cooperative learning techniques. ©2001 CRC Press LLC

Figure 4

With cooperative learning each group member has a role that contributes toward group accomplishment of the task. During this site survey by educators, there is one person holding the rod, one recorder writing the elevations measured on a data sheet, one instrument person sighting on the rod and stating the elevation observed, and two flaggers who have completed marking the area to be surveyed and are now learning the roles of recorder and instrument person. Through cooperation, the group accomplishes more than one person could working alone.

Performance-Based Instruction Performance-based instruction is a teaching technique in which students must solve problems and think critically by using basic knowledge and skills in real life situations (McTighe, 1996). This type of instruction cannot be assessed using typical multiple choice, true–false, fill in the blank, or short answer tests because there are many possible correct answers. Instead, a group of students solves a problem, such as “design a one acre wetland suitable for frogs” (Figure 5). Some states, such as Maryland, grade schools through performance assessment tests of students. Wetland activities can be a basis for performance-based instruction, especially many aspects of creating and monitoring wetlands.

EVALUATING WETLAND PROGRAMS AND ACTIVITIES There are several potential questions to be asked in determining whether a wetland program fulfills the needs of a school system or if a wetland activity is appropriate for a particular group of students. Is the program or activity suitable for the course and grade level needs of the science curriculum? Does it satisfy the science content ©2001 CRC Press LLC

Figure 5

With performance-based instruction, students solve problems and think critically using basic knowledge and skills in real life situations. These educators are determining how much rainwater runoff from the roof will be available to support a small constructed wetland.

standards (national, state, or local—depending on the target audience)? Is the program or activity multidisciplinary? Are student action projects included from which students could potentially earn service learning credits? Are the activities in a lesson plan format? Are instructional objectives clear, and do they contain appropriate action verbs? Do the activities allow use of multiple intelligences? Are teaching techniques such as hands-on learning, cooperative learning, or performance-based instruction utilized? Programs generating the greatest number of yes answers to these questions likely will apply to a broader audience. Many excellent, but specialized, programs, such as those focusing on action projects, may generate fewer yes answers because they are intended to fit within a larger program. In the activity Water We Have Here? from WOW!: The Wonders of Wetlands (Appendix 15-1 in Slattery and Kesselheim, 1995), students conduct a variety of tests on wetlands water to quantify physical characteristics, compare them to standards, then form some conclusions about the water quality. Using this activity as an example, how would each of the aforementioned school system requirements, and features important to educators, apply? Appropriate subject areas and grade levels are listed in the shaded area of the first page, so fit within the curriculum is easily determined. The summary and objective sections suggest which science content standards are supported by the activity, but the connection could be more explicit. The activity is multidisciplinary. No service learning credit is identified. However, the wrap-up and action section of the procedure section, and the extensions section, suggest student action projects that might provide opportunities for service learning ©2001 CRC Press LLC

credit. The activity is in a lesson plan format, including objectives, procedures, and assessment. The objectives are clear, contain the action verbs measure and monitor, and draw conclusions that are from several thinking skill levels. Many intelligences are utilized in this activity (e.g., verbal/linguistic, logical/mathematical, bodily/kinesthetic, interpersonal, and naturalist), but this can be ascertained only by reading through the activity. The WOW! book indicates that all of the activities involve handson learning. Performance-based and cooperative learning techniques could also be utilized, but this is not stated. Overall, this sample activity fulfills most school system requirements and contains many features important to educators. Presumably, this is the reason why curriculum guides containing this type of activity have been so successful.

WETLAND EDUCATIONAL PROGRAMS FOR GRADES K–12 Within every school there are unsung heroes and heroines, teachers who have pulled together bits and pieces of traditional programs in their own creative ways to challenge their unique group of students to understand and appreciate wetlands. Teachers and a supportive PTA president in Newark, DE, are guiding the entire second grade at Brader Elementary School, not just in the study of wetlands, but also in the enhancement of a schoolyard wetland by enlarging it and planting a greater variety of wetland plants. All involved were delighted when a pair of mallard ducks took up residence and raised a family in the wetland before it was dedicated! Success stories such as this often result from the efforts of small groups of enthusiastic, dedicated educators motivated by the desire to share the knowledge and problem-solving skills that students need to live in harmony with each other and our environment. How do success stories like this develop? What wetland programs do these teachers use? What wetland educational resources are available? Described below are a number of wetland educational programs. Contacts are provided in Table 5. National Education Programs Many programs are available that focus on local wetland issues. The following programs have been successful on a national scale. This is not intended to be an inclusive list, but these popular programs are representative of those that are readily available to educators. Most were produced by nonprofit organizations and predate the national science content standards discussed earlier. WOW!: The Wonders of Wetlands WOW!: The Wonders of Wetlands (Slattery and Kesselheim, 1995) is a unique educator’s guide that focuses entirely on wetlands. WOW! provides hands-on activities designed to excite and educate students, then illustrates that action can be taken. The original WOW! was written by Britt Eckhardt Slattery of Environmental Concern Inc. in 1991. It contains original drawings and 44 hands-on wetland activities for ©2001 CRC Press LLC

Table 5

Contact Information for Educational Programs Described in This Chapter Program

WOW!: The Wonders of Wetlands, POW! The Planning of Wetlands Project WILD Aquatic, WILD School Sites, Taking Action

Contact Environmental Concern Inc. Project WILD

Address P.O. Box P St. Michaels, MD 21663 707 Conservation Lane Suite 305 Gaithersburg, MD 20878 P.O. Box 22558 Lincoln, NE 68542-2558 P.O. Box 47600 Olympia, WA 98504-7600 P.O. Box 1016 Chapel Hill, NC 27514 8925 Leesburg Pike Vienna, VA 22184 8925 Leesburg Pike Vienna, VA 22184-0001

Children’s Groundwater Festival

The Groundwater Foundation

Discover Wetlands A World in Our Backyard

Washington Department of Ecology, Publications Distribution Environmental Media

Wading into Wetlands

National Wildlife Federation

Schoolyard Habitats Program

National Wildlife Federation

Wicked Big Puddles

Vernal Pool Association, Reading Memorial High School

62 Oakland Road Reading, MA 01867

Handbook for Wetlands Conservation and Sustainability

Izaak Walton League of America, Save Our Streams Program

707 Conservation Lane Gaithersburg, MD 20878-2983

©2001 CRC Press LLC

Phone/Web Site 410-745-9620 www.wetland.org 301-527-8900 www.projectwild.org 800-858-4844 www.groundwater.org 360-407-7472 www.wa.gov/ecology/pubs 800-363-3382 800-588-1650 www.nwf.org 800-822-9919 www.nwf.org/nwf/habitats/ schoolyard/index.html 781-944-8200 earth.simmons.edu/vernal/ pool/store.htm 800-BUG-IWLA www.iwla.org/sos

grades K–12 grouped into five chapters: introducing wetlands, plants and animals, water, soil, and issues. In 1995, Environmental Concern partnered with The Watercourse of Bozeman, MT, to revise WOW!. Background information was expanded, the lesson plan format and arrangement were improved, and some activities were revamped. The Watercourse and Western Regional Environmental Education Council produced project WET, an activity guide that focuses on water resources, in 1995. At that time WOW! became the wetlands module for Project WET, and facilitator training was initiated. WOW! can be used as a stand-alone wetlands unit, or selected activities can be integrated into current study units. To assist school systems in determining curriculum fit, a summary of activities by grade level and a unit planning guide are provided. Suggested grade levels are again listed in a shaded block within each activity along with subject areas. Most activities are multidisciplinary. Conserving Wetlands in the original WOW! has been separated into two activities stressing student action. Additionally, an appendix on Planning and Developing a Schoolyard Wetland Habitat has been included. A table cross-referencing the national science content standards and the activities (as has been done separately with the New Jersey State standards) would be a useful addition. Presented in an educator-friendly, lesson plan format, the activities have clearly stated objectives, procedures, assessments, and extensions. A variety of learning styles and teaching techniques are utilized. WOW! is available through Project WET workshops, through workshops provided by Environmental Concern, or directly from Environmental Concern or The Watercourse. While not required for purchase of WOW!, attendance at a workshop is strongly encouraged. Project WILD Aquatic Education Activity Guide Beginning in the early 1970s, Project Learning Tree (American Forest Foundation, 1995) was one of the first hands-on, interdisciplinary, supplemental education programs that contained activities for teachers to use with students K–12. Project Learning Tree is co-produced by the Council for Environmental Education (formerly the Western Regional Environmental Education Council, Inc.) and the American Forest Foundation (formerly the American Forest Institute). It contains a few wetland activities, but forests, not wetlands, are the primary focus. Project WILD was developed as a joint project by the Council for Environmental Education and the Western Association of Fish and Wildlife Agencies using a similar format. The Project WILD Activity Guide (Council for Environmental Education, 1992) became available in 1983. As in Project Learning Tree, a few wetland activities are included, but the primary focus of the project is wildlife, not wetlands. In 1987, the Project WILD Aquatic Education Activity Guide (Council for Environmental Education, 1992) made its national debut as a part of Project WILD. Containing 40 hands-on water and wildlife activities for grades K–12, this guide is designed to either supplement existing courses or stand alone as an aquatic education course. The material does not focus exclusively on wetlands but covers aquatic awareness and appreciation, wildlife values, ecological principles, management and

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conservation, culture and wildlife, trends, issues and consequences, and responsible human actions. Information to assist school systems in determining curriculum fit is provided in a box at the beginning of each activity. Within the box are listed the appropriate grade levels and subjects (many are multidisciplinary). In the appendices, grade level, subject, skills, and topics (one of which is actions) index all activities. A section on Taking Action deals more directly with student action projects, many of which could result in service learning credits. As with most programs, the science content standards are not yet referenced. For educators, each activity is formatted like a lesson plan. Objectives are clearly stated, and methods for attaining those objectives are described. Background information to assist the teacher, materials needed for the activity, a step-by-step procedure for conducting the activity, possible variations and extensions of the activity, and a means to evaluate student performance are also provided. Skills are listed in a box with other information and indicate that multiple learning styles are used in the activities. Of the activities, six make use of problem-solving skills. Group size listings indicate which activities might use cooperative learning techniques. The Project WILD Aquatic Education Activity Guide cannot be purchased. It is distributed free, along with the Project WILD Activity Guide, to participants attending a 6-hour workshop conducted by trained state coordinators and facilitators. To locate your state coordinator, check the listing on the Project WILD Web site. Children’s Groundwater Festival This program is decidedly different, and wildly successful, with twice as many applications as spaces are available. The Groundwater Foundation has been sponsoring the Children’s Groundwater Festival in Nebraska annually since 1989 and the idea has spread to many other states and Washington, D.C. Fifth grade students from across Nebraska travel to Grand Island and for 4 hours become immersed in a wide variety of hands-on activities dealing with groundwater and closely related themes such as wetlands. Presenters are volunteers from government, private industry, and education. Sponsors and grants cover expenses. Teachers prepare students with a groundwater unit before attending the Festival and follow-up with review activities. Toward that end, a Festival Outreach Packet (Groundwater Foundation, 1996) containing hands-on activities is available to educators. These activities fit the Nebraska fifth grade curriculum and content standards are multidisciplinary, have a lesson plan format with objectives, provide for multiple learning styles, and suit a variety of instructional techniques. National standards do not appear to be addressed. In another publication, Bringing the Festival Home (Groundwater Foundation, 1996), students are encouraged to perform a community service project connected with groundwater after attending the festival. For those interested in organizing their own festival, Making Waves: How to Put on a Water Festival (Killham, 1996) is available from The Groundwater Foundation as are a number of other publications and resources.

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Discover Wetlands The curriculum guide Discover Wetlands (Usher et al., 1995) was developed for Washington State educators in 1988, and substantially revised in 1995. Although the guide is state specific, it is so well prepared for classroom use that it is frequently cited and imitated. The guide contains four units: Washington’s wetlands, functions and values, people and wetlands, and field studies. Each unit is subdivided into four or five topics containing hands-on activities. The original target group was grades 4–8, but this was expanded to include grades K–12. Curriculum guidelines are provided in an appendix. The activities are cross-referenced with the Washington State curriculum goals and are multidisciplinary. Student action projects are suggested in the “Extensions” section of some activities. To assist educators, all activities are presented in a lesson plan format. However, the objectives often lack action verbs. The need for multiple learning techniques is satisfied and all activities are hands-on. Performance-based instruction and cooperative learning techniques are not addressed but could be used. To order copies of Discover Wetlands (Publication #88-16) contact the Washington Department of Ecology. A World in Our Backyard: A Wetlands Education and Stewardship Program A World in Our Backyard (Madison and Paly, 1994), produced by the New England Interstate Water Pollution Control Commission, contains both a curriculum guide and a video for use by classroom educators in New England. Wetland information is provided in eight chapters with hands-on activities for students. Topics include wetland science, types, functions, threats, locations, field studies, protection, and adoption. Classes that perform stewardship activities to protect local wetlands receive ‘Adopt A Wetland’ certificates from the U.S. Environmental Protection Agency, Region I, upon application. The guide is intended for use by middle school teachers as a supplement to existing curricula. Suggested uses are as a short wetland unit, for long-term multidisciplinary study, and with extracurricular organizations. Science standards are not addressed, but adopting a wetland would qualify as an action project. For educators, the activities are presented in lesson plan format, but lack an assessment and often an objective. Activities are hands-on and encourage use of multiple intelligences. Cooperative learning and performance-based instruction are not addressed. The video provides information about wetlands in an entertaining format. To order copies of the video and curriculum guide, contact Environmental Media. Wading into Wetlands Wading into Wetlands (National Wildlife Federation, 1997) is part of the very popular Ranger Rick’s NatureScope magazine series published by the National Wildlife Federation for elementary school aged children. The original version,

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produced in 1989, was updated in 1997. Information and hands-on activities are provided for use in the classroom, in community group activities, or individually, to help children understand and appreciate wetlands. The age group targeted is grades K–8, but no attempt is made to correlate the material with curricula or content standards. Many activities are multidisciplinary. Student action projects are not included. For educators, a limited lesson plan format is used that includes objectives. Use of multiple intelligences is promoted in the hands-on activities, but performance-based instruction and cooperative learning techniques are not included. This book is available in bookstores and through the National Wildlife Federation.

STUDENT WETLANDS ACTION PROJECTS Some programs are not intended to function as a curriculum, but rather spur students into taking some environmental action. Toward that end the following programs are successfully empowering students to take action on wetlands. Schoolyard Habitats Program Introduced in 1995, the Schoolyard Habitats Program of the National Wildlife Federation promotes the establishment of habitat learning sites (including wetlands) on school grounds. To accomplish this, a prepaid habitat certification kit may be ordered. The Application for Certification, which is also available at their Web site, may be printed, completed, and returned with the appropriate fee. The application contains five sections that require thought and planning to achieve. The categories include project goals and description, key project participants, components of the habitat (e.g., availability of food, water, cover, and places to raise young), a site diagram, and use of the habitat in the curriculum. The program does not include activities or study units but recognizes the extra effort that is required of students and educators to establish habitat areas. By design, the application focuses attention on the essential elements of habitat and the steps necessary for a successful habitat project. With habitat certification comes the newsletter Habitats and an opportunity to order and post a sign indicating that the National Wildlife Federation certifies the habitat. For further information contact the National Wildlife Federation. Wicked Big Puddles: A Guide to the Study and Certification of Vernal Pools The Wicked Big Puddles (Kenney, 1995) guide resulted from student action projects at Reading Memorial High School in Massachusetts. Individuals, groups, and high school students produced it as an aide in the identification, study, and certification of vernal pools. A vernal pool is a type of depression wetland in which snow melts and spring rains collect. Vernal pools exist long enough to be used for amphibian reproduction but usually are dry in summer. Unless certified, vernal pools ©2001 CRC Press LLC

are presumed not to exist and, therefore, have no protection from destruction. While the certification forms and procedures are specific to Massachusetts, wetlands (and vernal pool) certifications are being conducted in other jurisdictions as well. Check with your state department of natural resources or department of the environment for information about programs in your state. The guide is designed for use by high school students and adults as a project and does not have a lesson plan format. Two publications of the Massachusetts Audubon Society may provide additional help in this endeavor, Vernal Pool Lessons and Activities (Childs and Colburn, undated) and Certified: a Citizen’s Step-by-Step Guide to Protecting Vernal Pools (Colburn, 1993). To order copies of Wicked Big Puddles, contact the Vernal Pool Association. Handbook for Wetlands Conservation and Sustainability The target audiences for the Handbook for Wetlands Conservation and Sustainability (Firehock et al., 1998) are community organizations, such as local chapters of the Izaak Walton League. The focus is on establishing a wetland stewardship program, monitoring the wetland, and taking action to protect or enhance the wetland. Secondary school students working with adults can use the material presented, but again, schools are not the target audience. The needs of school systems and educators are, therefore, not addressed. The Handbook was published by the Save Our Streams Program of the Izaak Walton League of America in 1996. After 2 years of field testing with chapters across the nation, the second edition was printed in 1998. The first three chapters provide basic wetland information in a readable format, while the next five chapters supply information and directions on how to establish a stewardship program, monitor the many aspects of a wetland, and conserve wetlands by taking action. The second half of the Handbook contains references, resources, contacts, and 12 appendices giving detailed directions on some aspect of wetlands conservation. This useful manual is available directly from the Izaak Walton League. Also available are 2-day workshops to provide assistance in the many hands-on aspects of conserving wetlands. WILD School Sites and Taking Action WILD School Sites (Charles, 1993) and Taking Action (Stoner, 1995) are both produced by the Council for Environmental Education in conjunction with Project WILD. Both supply directions for developing schoolyard habitats, including wetlands, through a variety of student actions. WILD School Sites presents the rationale for providing habitat, basic wildlife habitat needs, suggestions for projects, steps for creating a plan, and putting the plan into action. The site may also be certified through the National Wildlife Federation for a small fee. Taking Action, “an educator’s guide to involving students in environmental action projects,” may or may not involve a school site, yet provides ideas of what actions to take and how to accomplish those actions.

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By design, both publications focus on action projects. Neither provides lesson plans nor instructional objectives. Both are multidisciplinary, allow use of multiple intelligences, and could be implemented using a variety of teaching techniques. These two projects thoughtfully guide educators through the difficult process of gaining acceptance for a project and getting it started. For additional information, contact Project WILD. POW!: The Planning of Wetlands Newly available in 1999, POW!: The Planning of Wetlands (Ripple et al., 1999) provides background information for educators on the enhancement, restoration, creation, and monitoring of wetlands. Student activities to accomplish that goal are also provided. Using the same format as WOW!: The Wonders of Wetlands, this action project guide is designed for use with students in grades 5–12. Modified activities are included for use with younger students (grades K–4), but they might also participate with older student partners in the main activities. The national science content standards and activities are cross-referenced in a table to assist school systems in determining curriculum fit. Most activities are multidisciplinary. As a student action project, planning a wetland could provide service learning credits. To guide educators, each activity is presented in a lesson plan format with instructional objectives, procedures, assessments, and extensions. Multiple intelligences are used in most activities, with many opportunities for use of performance-based instruction and cooperative learning. The manual may be ordered from Environmental Concern Inc., which also provides a 3-day course to guide educators in planning wetlands. Package programs are available that include manuals, a 3-day course, and professional guidance on wetland siting and design. Grant Funding Funding is available for obtaining programs such as those discussed above, or developing new wetland educational programs. The larger government grants, available at the national level, require much planning and preparation to produce a grant proposal. They often require matching funds and are highly competitive. State and regional government grants generally offer lower funding levels, require less paperwork, may or may not require matching funds, and are somewhat less competitive. Most large corporations set aside money for grants, some of which may be available for wetland education. While some advertise the funds available in competitive, established programs, others award money if solicited, and if the grant proposal matches their interests. Foundations of all sizes also offer grants. Small local companies should not be overlooked if the grant funding request is modest and will in some way benefit the local area. To save time and effort, match funding request levels and the area to be affected to appropriate organizations at the national, state, and local levels. It is also acceptable to request funds from more than one source. For those unfamiliar with writing grant proposals, it must be stressed—read and follow all directions and guidelines. Those reviewing the proposals may know very ©2001 CRC Press LLC

little about what is being proposed, but they can evaluate whether the required information has been supplied, and whether it is presented in an appealing manner. Incomplete grant proposals typically are rejected. Before the merits of a proposal can be considered, it must be complete. If a proposal is rejected, inquire as to why if reasons or copies of reviewers’ comments are not supplied with the rejection letter. Quite possibly there was no flaw in the proposal, just that more funds were requested than were available. Heed suggestions for improvement, and then do not hesitate to reapply to the same organization or to another. Be sure the program fits the needs and skill levels of the intended audience if requesting grant funds for a program. A wetland program designed for fifth graders may not be utilized at all in areas that have second and ninth grade wetland units in the curriculum. A grant proposal is stronger and more likely to be funded if it can be shown that a need for the program exists. One of many references that can help in writing a strong grant proposal is Grant Funding for Your Environmental Education Program: Strategies and Options prepared by the North American Association for Environmental Education (NAAEE, 1993). Courses are also available offering tips and techniques for writing grant proposals through local college continuing education and workforce training programs. There are several ways of locating appropriate granting agencies, including mailing lists for grant information from appropriate federal agencies, publications with summary lists of available grants, Internet search, and Internet alert groups. Each has advantages. For those who prefer hard copy information booklets and forms, check the local library for directories that provide information on agencies, corporations, and foundations that offer grants. Most directories are updated yearly. Contact the organizations that appear most appropriate for your needs, then request grant information, applications, and placement on their mailing list. Be aware that within each federal agency there are many programs offering grants, most on an annual cycle. In addition, many agencies have regional offices that offer grants within specific areas. If Internet is your information tool of choice, try an Internet “search” for grants or, if you prefer to have a search done for you, try an organization such as U.S. Opportunity Alert (www.usalert.com/public/register.asp) that will deliver targeted funding information automatically via e-mail for a fee. Whatever funding source you choose to pursue, follow their guidelines carefully, fully complete the application and/or proposal, and submit paperwork on time. Membership Organizations A host of pertinent organizations are available that network people and information. Many focus on environmental education, but an increasing number are stressing wetland education. Most have publications, workshops, and/or conferences to help educators stay abreast of new ideas, new materials, and training that is available. Table 6 provides a listing of some national organizations that provide a variety of benefits for wetland educators.

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Table 6

Organizations Which May Benefit Wetland Educators Organization

American Water Resources Association American Water Works Association

Ducks Unlimited

Global Rivers Environmental Education Network (GREEN)

Izaak Walton League of America, Save Our Streams Program National Audubon Society

National Marine Educators Association (NMEA) National Science Teachers Association (NSTA) National Wildlife Federation Nature Conservancy

©2001 CRC Press LLC

Benefits

Address

Phone/Web Site

Publications Conferences Local sections Publications Drinking water week Blue Thumb Project Publications Local chapters Conservation Publications Workshops Network Kits Publications Workshops Local chapters Publications Workshops/camps Local chapters Publications Annual conference Local chapters Publications Convention Programs Publications Workshops Publications Local chapters Conservation

950 Herndon Parkway, Suite 300 Herndon, VA20170-5531

703-904-1225 www.uwin.siu.edu/~awra

6666 West Quincy Ave. Denver, CO 80235

303-794-7711 www.awwa.org

One Waterfowl Way Memphis, TN 38120-2351

800-45ducks www.ducks.org

206 South Fifth Ave., Suite 150 Ann Arbor, MI 48104

734-761-8142 www.igc.org/green www.earthforce.org

707 Conservation Lane Gaithersburg, MD 20878-2983

800-BUG-IWLA www.iwla.org

Education Division 700 Broadway New York, NY 10003 P.O. Box 1470 Ocean Springs, MS 39566-147

212-979-3000 www.audubon.org

1840 Wilson Blvd. Arlington, VA 22201-3000

703-243-7100 www.nsta.org

8925 Leesburg Pike Vienna, VA 22184 4245 North Fairfax Dr., Suite 100 Arlington, VA 22203-1606

800-588-1650 www.nwf.org 703-841-5300 www.tnc.org

228-0-374-7557 www.marine-ed.org

Table 6

Organizations Which May Benefit Wetland Educators Organization

Benefits

Address

North American Association for Environmental Education (NAAEE)

Publications Annual conference Sections/activities Member directory Publications Activities Wetlands projects Publications Conferences Wetlands month

410 Tarvin Road Rock Spring, GA 30739

706-764-2926 www.naaee.org

730 Polk Street San Francisco, CA 94109

317-231-1908 www.sierra.org

4 Herbert Street Alexandria, VA 22305

703-548-5473 www.terrene.org

Sierra Club

Terrene Institute

©2001 CRC Press LLC

Phone/Web Site

WETLAND EDUCATION FOR PROFESSIONALS To be effective and competitive, professionals in any field must stay current with information, techniques, and regulations. Both continuing education courses offered by universities and training workshops offered by independent professional groups are available to wetland professionals. Membership organizations encourage networking of people and sharing of information. They also help members stay abreast of conferences and training opportunities. Certification The U.S. Army Corps of Engineers certification program is specific for wetland delineation. The program was designed to ease the burden of wetland delineation under Section 404 of the U.S. Clean Water Act. Currently that program is not funded. The Society of Wetland Scientists (SWS) has created a certification program to satisfy the need to identify individuals who are qualified to assess and manage wetlands. The certification program has no legal or official standing but signifies that a Professional Wetland Scientist meets the standards established by his/her peers. It requires both education and experience and is designed to meet the needs of a broad range of wetland professionals. For more information on the SWS Professional Certification Program, call 800-627-0629 or check the Web site: www.wetland cert.org. University Continuing Education Universities with postgraduate continuing education programs often have short courses for wetland professionals who wish to improve their knowledge and skills, need to refresh them, or need to catch up on the latest developments in the field. One outstanding example is the Cook College Continuing Professional Education Program of Rutgers, The State University of New Jersey. Rutgers offers programs in wetland delineation, protecting watersheds, stabilization and restoration of disturbed sites, planning hydrology for wetland construction, and more. The Society of Wetland Scientists lists colleges offering training at www.sws.org/colleges. Discover what else is available in your region by contacting local colleges and universities directly. In coastal states, the National Sea Grant College Program provides resources for workshops, research, and grants. Sea Grant is a partnership program consisting of government, industry, and universities and is designed to provide research, education, and outreach on the seas and their land borders. The National Sea Grant Depository of publications is located at the University of Rhode Island and can be accessed by telephone (401-874-6114) or Web site (www.nsgd.gso.uri.edu). The Web site also provides links to all Sea Grant Colleges. If living or working in a coastal state, check the nearest Sea Grant College for a wealth of resources and possible training opportunities.

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Independent Professional Training A wide variety of professional training opportunities are available from many excellent independent commercial and nonprofit organizations. The reputation of the instructor is often key to the quality of the course, but even unknowns are worth some risk if the program has a good reputation. The organizations that follow are representative of what is available, but the list is not comprehensive. For a more complete listing of training opportunities worldwide, check the Ramsar Web site (www.ramsar.org/wurc_training_directory.htm) or the Society of Wetland Scientists Web site (www.sws.org/training). Environmental Concern Inc. Environmental Concern Inc. is a small nonprofit organization with a singular focus on wetlands. They have a long-term commitment to wetland education, search and development, and application of technology in wetland construction, restoration, and enhancement. Quality professional courses are offered in the areas of wetland delineation, wetland hydrology, wetland mitigation, evaluation for planned wetlands, wetlands botany, wetlands horticulture, and more. Wetlands courses for educators are now offered as well. Scholarships are available for college juniors and seniors, government regulators, and educators. Occasionally, conferences on selected wetland topics are convened. Environmental Concern has published a variety of wetland-related books and the Wetland Journal. The Wetland Journal, published four times per year, encourages contributions and comments from professionals in the field. For a professional course schedule, a subscription to the Wetland Journal, or information about other publications, contact Environmental Concern Inc. at P.O. Box P, St. Michaels, MD 21663 or on the Web at www.wetland.org. Wetland Training Institute The Wetland Training Institute, Inc. is a commercial association of instructors from government, private industry, and academia that was formed to provide professional training in water resource conservation, management, and regulation. Courses offered at a variety of locations include wetland delineation, wetland soils and hydrology, wetland construction and restoration, hydrology of constructed wetlands, federal wetland policy, and more. Some publications are also available. Contact the Wetland Training Institute for more information and schedules at P.O. Box 31, Glenwood, NM 88039 or on the Web at www.wetlandtraining.com. Richard Chinn Environmental Training Richard Chinn Environmental Training is a small but well-publicized commercial venture offering courses in Florida and many locations throughout the country. Wetland Delineation and Management Training (plus other courses unrelated to

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wetlands) are scheduled annually. For additional information and schedules contact CET Richard Chinn Environmental Training, Inc., P.O. Box 10776, Pompano Beach, FL 33061-6776 or on the Web at www.richardchinn.com Institute for Wetland and Environmental Education and Research The Institute for Wetland and Environmental Education and Research (IWEER), a private educational organization, offers short training courses for environmental professionals. Courses focus on wetland-related topics which are presented by a faculty of recognized experts, all of whom have experience teaching. Course topics include wetland delineation, wetland hydrology, wetland classification, hydrogeomorphic concepts, plant identification, and more. Courses are offered at a variety of locations and are mostly field oriented. For additional information and a schedule of courses, contact the Institute for Wetland and Environment Education and Research, P.O. Box 288, Leverett, MA 01054 or on the Web at members.aol.com/iweer. Membership Organizations Professional organizations offer a variety of training and networking opportunities that help to maintain and increase levels of competency for wetland specialists. Table 7 lists organizations that may help fill that need.

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Table 7

Organizations of Interest to Wetlands Professionals

Organization American Association for the Advancement of Science (AAAS) Association of State Wetland Managers

Ecological Society of America (ESA)

Environmental Law Institute (ELI) National Association for Conservation Districts (NACD) National Science Foundation (NSF) Society for Ecological Restoration (SER)

Society of Wetland Scientists (SWS)

Soil Science Society of America

Water Environment Federation

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Benefits Publications, meetings Education section Publications, meetings Workshops Registry Publications, meetings Workshops Education section Training Seminars Publications, meetings Conventions Publications Grants Publications, meetings Conferences Training Publications, meetings Training Certification Publications, meetings Conferences Certification Publications Conferences Training

Address

Phone/Web Site

1200 New York Ave., NW Washington, DC 20005 P.O. Box 269 Berne, NY 12023-9746

202-326-6400 www.aaas.org 518-872-1804 www.aswm.org

2010 Mass. Ave., NW, Suite 400 Washington, DC 20036

202-833-8773 www.sdsc.edu/ ~ESA/esa.htm 202-939-3800 www.eli.org 202-547-6223

1616 P St., NW, Suite 200 Washington, DC 20036 P.O. Box 855 League City, TX 77574-0855 4201 Wilson Blvd. Arlington, VA 22230 1207 Seminole Hwy., Suite B Madison, WI 53711

703-306-1234 www.nsf.gov 608-262-9547 www.ser.org

P.O. Box 1897 Lawrence, KS 66044-8897

913-843-1221 www.sws.org

677 South Segoe Rd. Madison, WI 53711-1086

608-273-8095 www.soils.org

601 Wythe Street Alexandria, VA 22314-1994

800-666-0206 www.wef.org

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Slattery, B. E. and Kesselheim, A. S., WOW!: The Wonders of Wetlands, Environmental Concern Inc., St. Michaels, MD, and The Watercourse, Bozeman, MT, 1995. Stoner, D. K., Taking Action, Council for Environmental Education, Bethesda, MD, 1995. Tiner, R. W., In Search of Swampland, Rutgers University Press, New Brunswick, NJ, 1998. Usher, L., Lippy, K., and Pauls, S., Eds., Discover Wetlands, Publication #88-16, Washington Department of Ecology, Olympia, WA, revised July 1995. The Watercourse and Western Regional Environmental Education Council, 1995. Project WET, The Watercourse, Bozeman, MT and WREEC, Houston, TX, 1995.

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