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Freshwater Ecosystems: Revitalizing Educational Programs in Limnology (1996)

Chapter: Bringing Biology Back into Water Quality Assessments

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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Page 300
Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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Suggested Citation:"Bringing Biology Back into Water Quality Assessments." National Research Council. 1996. Freshwater Ecosystems: Revitalizing Educational Programs in Limnology. Washington, DC: The National Academies Press. doi: 10.17226/5146.
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BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 289 Bringing Biology Back into Water Quality Assessments G. Wayne Minshall Department of Biological Sciences Idaho State University Pocatello, Idaho SUMMARY For some time now, the quality of the nation's inland waters has been evaluated largely on the basis of chemical and toxicological criteria. However, more recent theories reflect the idea that the native biota, exposed to the full suite of environmental conditions in nature, more accurately reflect the suitability of that environment for survival and long-term persistence. This paper examines the reasons for the resurgence of interest in the biological assessment of water quality and highlights some important considerations in the application of this approach to inland aquatic ecosystems. Biological integrity is a concept central to successful bioassessment because it identifies the essential factors to be measured and provides a reference against which the degree of environmental disturbance or stress, either natural or anthropogenic, can be evaluated. Major anthropogenic stresses on the integrity of inland aquatic ecosystems include livestock grazing; forestry; agriculture; mining and smelting; urban usage; manufacturing; impoundment and diversion; and lake-, marsh-, and stream-bottom alteration. Measurement of the biological health of aquatic ecosystems is a complex issue involving multiple spatial and temporal scales and methodological and logistical considerations. Nevertheless, direct assessment of the status or ecological health of aquatic organisms and communities is essential for proper resource management of inland waters and their sustained diversity and productivity. INTRODUCTION Over the past few decades, water quality has been defined primarily in chemical terms. More recently, however, water management agencies

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 290 have been increasingly aware of the need to bring biology back into the water quality equation. In some cases, chemical monitoring has actually exceeded the ability to detect biological impacts of chemical contaminants, so that large sums of money have been spent to remove contaminants that do not even affect aquatic organisms. On the other hand, reliance on chemical criteria or laboratory-derived toxicological information taken out of the environmental context has often allowed levels of toxicants or other materials that are harmful to aquatic populations. The return of biology to environmental assessments has brought with it a need for knowledge about whole-organism biology, the study of which has become increasingly neglected in academic institutions over the past several decades. At the same time, many exciting developments, in fields ranging from molecular biology to landscape ecology, have potential application to the study and management of inland aquatic resources. This paper reviews the historical basis for the application of biological methods to water quality assessment and discusses factors that need to be considered in evaluating the biological integrity of inland aquatic ecosystems. HISTORICAL BACKGROUND Modern bioassessment of inland aquatic ecosystems has given rise to several terms and concepts regarding protection or restoration of aquatic environments (Steedman, 1994). Foremost among these are the ideas of integrity, which relates to whether biological systems are intact or restorable, and health, management, and sustainability, which relate to modification of sites by human activity. The idea of biological, and subsequently ecological, integrity is traceable at least as far back as the writings of Aldo Leopold (1949), but its emergence as a formal ecosystem concept did not occur until the mid-1970s (e.g., Cairns, 1977a,b). The Water Quality Act Amendments of 1972 (P.L. 92-500) formalized the term ''biological integrity" under the directive to restore and maintain the "chemical, physical, and biological integrity of the nation's waters." Initially, the primary focus was on chemical and physical aspects of the environment and on toxicity tests performed in the laboratory on both individual contaminants and complex mixtures of waste effluents from industry and other sources. The idea of biological integrity gradually evolved to include naturalness, sustainability, and ecosystem balance, structure, and function (Jackson and Davis, 1994). Karr and Dudley (1981) defined biological integrity as the "ability of an aquatic ecosystem to support and maintain a balanced, adaptive community of organisms having a species composition, diversity, and functional organization comparable to that of natural habitats within a region." Others have refined,

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 291 clarified, and extended this concept to specific applications in a series of articles (e.g., Karr, 1991, 1993; Kay, 1991; King, 1993; Steedman and Haider, 1993; Polls, 1994; Steedman, 1994). Still others have addressed the choice of indicators (Keddy et al., 1993), monitoring considerations (Munn, 1993), and descriptions of national programs to measure ecological integrity (EPA, 1990; Marshall et al., 1993; Woodley, 1993; Jackson and Davis, 1994). Because of increasing public awareness of environmental problems, beginning in the mid-1960s and continuing to the present, key ecological issues have been codified as catch phrases, such as ecological or ecosystem health, management, and sustainability. • Ecological health generally is regarded as a condition in which natural ecosystem properties are not severely restricted, the ability for progressive self-organization is present, the capacity for self-repair when stressed is preserved, and minimal external support for management is needed (Steedman and Regier, 1990; Karr, 1993). Sites modified by human activity may be considered ecologically healthy "when their management neither degrades the sites for future use, nor results in degradation beyond their borders" (Steedman, 1994). Problems associated with deterioration of ecosystem health must be addressed at a landscape scale of resolution since significant cumulative and interactive effects otherwise might be overlooked. • Ecosystem management may be defined as "the skillful, integrated use of ecological knowledge at various scales to produce desired resource values, products, services, and conditions in ways that also sustain the diversity and productivity of ecosystems" over the long term (Avers, 1992). In practice, it means blending the needs of people and environmental values in such a way as to achieve healthy, productive, and sustainable ecosystems. The extent to which these goals are attained can be determined only through biological assessment and monitoring of resource conditions. • Ecosystem sustainability is "the ability to sustain diversity, productivity, resilience to stress, health, renewability, and/or yields of desired values, resource uses, products, or services from an ecosystem while maintaining the integrity of the ecosystem over time" (Overbay, 1992). Its return to the resource management equation has come about through the National Environmental Protection Act, the Endangered Species Act, and numerous other federal laws passed during the 1960s and 1970s (e.g., Overbay, 1992) and sustained by federal court decisions. Ecosystem management and sustainability are likely to have a major influence on research and management of many inland aquatic ecosystems in the United States for years to come. These concepts are especially relevant to federal agencies with large land holdings and broad responsibilities for the terrestrial and aquatic resources occupying them, such as

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 292 the Bureau of Land Management, the National Forest Service, the National Park Service, and the Refuge Management branch of the Fish and Wildlife Service. Both the concepts and the recognition of their responsibility to implement them are new to these agencies; it is still unclear how (and at what rate) they will move toward instituting formal ecosystem management policies. However, it is clear that biological assessments and continued monitoring will be important to ecosystem management in inventorying aquatic biological resources and their status, in assessing the effects of various management practices on them, and in determining suitable management strategies through research and adaptive management practices to ensure ecosystem sustainability. Federal Legislative Initiatives Historically, three major federal legislative actions are responsible for current efforts to increase the use of biological measures and a more meaningful ecological perspective: (1) the Water Quality Act, (2) the National Environmental Policy Act, and (3) the Endangered Species Act. The Water Quality Act of 1965 (P.L. 89-234) and the related Clean Water Act (Federal Water Pollution Control Act Amendments of 1972, Clean Water Act of 1977, Water Quality Act of 1987) were enacted in response to widespread surface water degradation and a growing public environmental awareness and concern. The implications of this legislation for inland aquatic science range from classroom to courtroom, and its implementation provides substantial opportunities for involvement in all aspects of water science. The National Environmental Policy Act of 1969 (NEPA; P.L. 91-190) was responsible for interjecting an ecological perspective into subsequent federal legislation and actions, particularly as they relate to natural resource-oriented projects. NEPA set forth a national policy to protect and improve the national environment by requiring detailed consideration of proposals for federal legislation, construction (e.g., dam construction, channel alteration, draining of wetlands), or resource extraction (e.g., water diversion, logging, or livestock grazing) likely to significantly affect the quality of the air, land, and water environments. Among other things, the law required the identification of (1) any adverse environmental effects that cannot be avoided should the proposal be implemented; (2) alternatives to the proposed action, including total abandonment and mitigation of damages; (3) the relationship between local short-term human uses of the environment and maintenance and enhancement of long-term productivity (i.e., sustainability); and (4) any irreversible and irretrievable commitments of resources that would be involved in the proposed action. This act consistently has been upheld and expanded by federal court decisions. Increasingly, NEPA and its legal interpretations have had far-reaching

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 293 implications for the management of inland aquatic resources at the ecosystem and landscape scales. NEPA has resulted in a more holistic and long-range view of past, present, and future management actions on natural resources in an ecosystem context and has called for a greater and more thorough knowledge of resource states under different management treatments. The Endangered Species Act of 1973 (ESA; P.L. 93-205) protects all species (except pests) of plants and animals in danger of extinction. Twelve percent of all animal species live in inland waters, and many species are restricted to limited geographic ranges. As freshwater habitats have been destroyed, altered, or polluted, biodiversity and ecosystem integrity have declined. The listing of federally recognized threatened or endangered freshwater species is an important means of tracking total biological integrity (Covich, 1993). The Endangered Species Act has served to emphasize the importance of identifying and preserving the diversity of inland aquatic organisms and their habitats, and of assessing long-term trends in their conditions. Several recent developments stemming from these legislative acts have brought the biological aspects of water quality to the forefront: (1) the initiation of several large federal monitoring and assessment programs that emphasize the measurement of water quality in biological rather than solely chemical or physical terms; (2) legal mandates to institute biological criteria into state water quality standards in the next few years; and (3) comprehensive assessment of the status of resources throughout the Columbia River Basin and how to manage these resources. These directives and comprehensive programs at the state and national levels will severely overload existing resource management personnel, a situation that is unlikely to be alleviated at the current rate of qualified graduates entering the work force. Federal Monitoring and Assessment Specific federal programs of monitoring and assessment have been instituted by the Environmental Protection Agency (EPA) and the U.S. Geological Survey (USGS). Presumably, the newly instituted National Biological Service (NBS) also will emphasize biological assessments through wetland surveys, inventories of biological resources, and the like, unless these responsibilities are abrogated by the new Congress. At the moment, the premier U.S. federal program involving bioassessment is the National Water Quality Assessment (NAWQA) Program of the USGS (Gurtz, 1994). This program is designed to integrate chemical, physical, and biological data to assess the status of, and trends in, national water quality. It consists of 60 study units (major river basins and large aquifers) located throughout the country (Gurtz, 1994) that represent

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 294 major natural and human-impacted conditions that influence water quality. Data collection began in 1991. Biological data include (1) analysis of aquatic organism tissues for a wide array of chemical contaminants; (2) characterizations of algal, macroinvertebrate, and fish communities; and (3) characterizations of vegetation growing in and along streams. Physical data include streamflow and characterizations of in-stream, bank, and floodplain habitats (Meador and Gurtz, 1994). NAWQA has developed nationally consistent biological sampling methods so that results are comparable across different river basins and geographic regions. The Environmental Monitoring and Assessment Program (EMAP) of the EPA has comparable goals, but it uses a more extensive set of sampling sites, including lakes and wetlands (Hunsaker et al., 1990; Paulsen et al., 1991). This program has spent much time developing its philosophical and conceptual underpinnings and has lagged behind NAWQA in making available a standard set of methods and in initiating a full-fledged data collection program. However, a series of pilot studies focusing on lakes was conducted in 1991 (Larsen and Christie, 1993). An Intergovernmental Task Force on Monitoring Water Quality led by representatives of NAWQA and EMAP (Gurtz and Muir, 1994) is working to develop a national water quality survey that would demonstrate effective collaboration among federal agencies. This group has chosen to focus on biological aspects of water quality to better understand the condition of the nation's stream communities and to identify opportunities and barriers to cooperative partnerships (M. Gurtz, USGS, personal communication, 1994). The national survey will initially aim to characterize reference conditions, but the long-term goal is to include all streams regardless of their condition. The Clean Water Act mandates state development of criteria to measure water quality conditions based on biological assessments of natural ecosystems. The general authority for biological criteria comes from Section 101, which establishes as the objective of the act the restoration and maintenance of the chemical, physical, and biological integrity of the nation's waters. This section also includes an interim water quality goal for the protection and propagation of fish, shellfish, and wildlife. Propagation includes the full range of biological conditions necessary to support reproducing populations of all forms of aquatic life and other life that depend on aquatic systems (EPA, 1990). Sections 303 and 304 provide specific directives for the development of biological criteria. Section 303 requires states to adopt protective water quality standards that consist of uses, criteria, and antidegradation measures. Section 303(c)(2)(B), enacted in 1987, requires states to adopt numerical criteria for toxic pollutants specified by EPA. The section further requires that states adopt criteria based on biological assessment and monitoring methods, consistent with information published by EPA under Section 304(a)(B). Section 304 directs

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 295 EPA to develop and publish water quality criteria and information on methods for measuring water quality, including biological monitoring and assessment methods to determine (1) the effects of pollutants on aquatic community components (e.g., plants, plankton, fish) and community attributes (e.g., diversity, productivity, stability) in any body of water, and (2) the factors necessary to restore and maintain the ecological integrity of all navigable waters (EPA, 1990). Development and use of biological criteria also will help states to meet the intent of several other legislative acts that require an assessment of risk to the environment (including resident aquatic communities) to determine the need for regulatory action (EPA, 1990). Some examples of the latter are the Comprehensive Environmental Response, Compensation, and Liability Act of 1980; the Federal Land Policy and Management Act of 1976; the Fish and Wildlife Conservation Act of 1980; NEPA; the Resource Conservation and Recovery Act first enacted in 1976; and the Wild and Scenic Rivers Act passed in 1968. Under the Clean Water Act, states were required to begin instituting narrative biological criteria into state water quality standards during 1991-1993; numeric criteria and full implementation are scheduled to occur within a few years (EPA, 1990). These requirements also apply to federal agencies responsible for the management of large tracts of public land (e.g., the U.S. Forest Service and Bureau of Land Management), especially in the western United States. Narrative biological criteria are general definable statements of conditions or attainable goals of biological integrity and water quality for a given use designation; numeric criteria establish specific values based on measures such as species richness, presence or absence of indicator taxa (taxonomically related groups), and trophic composition. Need for Cooperation Among Federal Agencies A good example of cooperation among federal agencies in addressing these aspects using biological assessment is the cooperative survey of the Apalachicola-Chattachoochee-Flint River Basin recently initiated by NAWQA and the NBS (NAWQA Information Sheet, April 6, 1994). This river basin, one of the largest in the eastern Gulf Coast Plain, was known for its rich diversity of at least 45 species of unionid mussels, but these populations have either declined or died out. Mussels are sensitive indicators because they are sessile and are dependent on good water quality, physical habitat conditions, and populations of host fish. The life cycle of unionid mussels is closely linked to fish because mussel larvae are obligate parasites on fish before becoming free- living adults. Conservation efforts to protect or restore declining mussel populations require information on both mussel and fish populations in watersheds with differing

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 296 land uses and water quality conditions. The NAWQA program, with its interest in monitoring water quality, and the NBS program, with its interest in identifying and protecting populations of aquatic organisms, both require information on the distribution and abundance of aquatic organisms in different environmental settings. Recent awareness of the rapidly declining status of 76 anadromous fish stocks in the Columbia River Basin (Nehlsen et al., 1992), together with documentation of declining freshwater habitat conditions (Sedell and Everest, 1991), has resulted in intensive efforts by several federal agencies to head off potential extensive curtailment of their resource extraction activities throughout the entire river catchment. Of immediate concern is the fact that protections offered for threatened and endangered fish under the ESA could result in severe curtailment or alteration of U.S. Forest Service and Bureau of Land Management activities. Ideally, improved management of aquatic and riparian ecosystems on lands administered by these two agencies, combined with improvements in hydropower operations, hatchery practices, and fish harvest management, can prevent additional stocks from becoming extinct and preclude the need to extend the protections of the Endangered Species Act to other at- risk anadromous fish stocks (U.S. Forest Service-U.S. Bureau of Land Management, 1994). In addition, both agencies are required by the Clean Water Act of 1976 (33 USC 1251, 1329) to ensure that activities occurring on lands they administer comply with requirements concerning the discharge or runoff of pollutants. A reasoned response to this new information on serious declines in anadromous fish stocks and aquatic habitat conditions is crucial to the two agencies' success in meeting the "continuing compliance" obligations of NEPA, ESA, the National Forest Management Act of 1976 (NFMA), the Federal Land Policy and Management Act of 1976 (FLPMA), and other environmental laws. By using the latest scientific information on chemical, physical, and biological integrity, the agencies will be better able to ensure the long-term viability of anadromous fish species and the continuing production of goods and services from public lands. Interim and longer-term management strategies are being examined in several geographically specific environmental impact statements as required under NEPA; also under development is a comprehensive ecosystem management plan for the interior Columbia River Basin (Science Integration Team, 1994). STRESSES ON BIOLOGICAL HEALTH OF INLAND WATERS The biological integrity of inland aquatic ecosystems is being assaulted in many ways (Power et al., 1988; Resh et al., 1988; Covich, 1993). Numerous

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 297 anthropogenic disturbances affect inland waters and their associated riparian ecosystems: • Livestock grazing contributes to increased inorganic sediments, nutrients, and organic matter; breakdown of stream banks; and removal of riparian vegetation. • Forestry and logging practices, including extensive road building, introduce sediments and logging slash; remove large woody debris; increase runoff and streambed scour; and erode stream banks. • Agricultural practices add sediments, nutrients, and toxicants; deplete streams through irrigation withdrawal; channelize streams; drain wetlands; and destroy riparian habitats. Pesticides applied to forest and agriculture lands often reach waterways. • Mining and smelting operations release heavy metals and other poisonous substances to water bodies via surface, subsurface, and aerial pathways. • Urban usage removes water for domestic consumption; adds sewage and many complex household and other chemicals; converts stream channels into concrete-lined gutters; and contributes fertilizers, herbicides, and pesticides. • Manufacturing and processing operations release chemicals and heated water and, along with motorized vehicles, contribute airborne pollutants that reach waterways. • Fish management practices use poisons to remove unwanted species and introduce exotic species. • Impoundment for flood control, electric power generation, navigation, and recreation drowns rivers, changes flow patterns, alters nutrient and sediment loads and temperatures, and thereby destroys the habitat and impedes or blocks the movement of native aquatic fauna. • Diking, channelization, and removal of woody debris for navigation, flow "enhancement," flood control, or fish passage all speed up the flow of water; destroy habitat; disrupt in-stream processing of organic matter and nutrients; and prevent interchange of nutrients, organic matter, and sediments within the riparian environment. • Production of electricity by coal-fired or nuclear reactor steam plants depletes water by evaporation and by diversion from natural water bodies and may increase temperature, trace elements, and other chemicals. Nutrients from many of the above activities, particularly nitrogen and phosphorus, cause the accelerated enrichment (cultural eutrophication) of lakes and streams. This can result in large-scale fish kills and the elimination of desirable fish species, production of foul odors, uncontrolled growth of algae and toxic bacteria, and obnoxious accumulations of filamentous algae and vascular plants. Not only do these activities affect the ecological integrity of inland

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 298 aquatic ecosystems, but the effects of each type of disturbance may be synergistic among types and cumulative in space and/or time (Sidle, 1990). Although viewed as relatively local, they often have large-scale, far-reaching effects. Some large-scale stresses affecting aquatic ecosystems, whether natural or human induced, are rapid and dramatic. Examples include certain recent cases of massive deforestation, urbanization, development of crop- and pasturelands, forest fires, plant disease outbreaks, and insect infestations. Other disturbances occur over extended periods of time and, hence, often are not recognized as such until the situation becomes extremely difficult or impossible to reverse. These include acidification, some types of logging and mining, livestock grazing, fire suppression, irrigation, and potentially, global climate change (Minshall, 1992, 1993; Covich, 1993). Global climate change could profoundly alter riparian ecosystems through its effect on terrestrial vegetation, thermal and hydrologic regimes, nutrient cycles, and so on (Firth and Fisher, 1992). Fast or slow, disturbances of riparian ecosystems may result in changes in water temperature or runoff, channel straightening, scouring or sedimentation, loss of physical habitat, alteration of food base, and waterlogging or drying of riparian soils. Challenges of Assessing Biological Integrity Although legislation calls for maintaining biological integrity, measuring the biological health of inland waters is extremely complex; this complexity results not only from the need to account for natural variations in time and space, but also from the need to consider individual species as well as interactions among organisms in a particular aquatic community. Importance of Scale (Space and Time) There is no single correct scale for the study, assessment, or management of aquatic-riparian ecosystems (O'Neill et al., 1986; Levin, 1992; Johnson et al., 1993); rather, the appropriate scale depends on the scientific question or management problem being addressed. The importance of various environmental factors and the interpretation of measurements taken on aquatic ecosystems vary with scale (O'Neill et al., 1986; Minshall, 1988). Further, since ecosystem boundaries vary with scale, the spatial boundaries also must be correlated with the temporal framework appropriate for a particular disturbance (O'Neill et al., 1986). Most ecosystems extend over comparatively large areas and persist for long periods of time. It is thus difficult to devise large-scale, single-value measurements of ecosystem integrity. However, the hierarchical structure of ecosystems results in a series of scaled interactions that can act as natural integrators of local processes. For example, measurement of community metabolism of a river segment or lake can serve to integrate the status of

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 299 conditions from a myriad of spatial patches and compartments within these ecosystems. This natural integration is especially evident at the scale of entire water catchments (King, 1993) where, for example, the ecological health of an entire forest may be reflected in the condition of the stream flowing through it. Use of scales of assessment close to the scale of the entire ecosystem will increase the likelihood that observed changes will be of consequence for the entire ecosystem. These larger-scale integrated measures are invaluable for detecting changes in loss of ecosystem integrity, but they may have to be supplemented with finer-scale measurements to determine cause and effect (King, 1993). For example, measurements of primary production at the level of patches or compartments within river segments or lakes are necessary to determine the relative importance of each and to isolate the specific factors responsible for any differences. In lakes, measurements performed on the plankton are better for lakewide assessments of phosphorus availability, whereas measurements utilizing attached algae such as Cladophora permit more localized assessments (Cairns et al., 1993). As noted above, there is a range of biological information that can be used to evaluate water quality. Studies at the population and community levels of organization emphasize species populations and interactions within and among them, such as competition. In this approach, the physical environment is seen as external to the system of organisms and biotic interactions (King, 1993). Population and community studies emphasize biotic interactions, whereas ecosystem studies focus on the processing and transfer of matter and energy in which the environment is an integral (as opposed to external) part of the system (O'Neill et al., 1986; King, 1993). Study of landscapes commonly addresses patterns of distribution within and among ecosystems, thus generally implying spatial scales of relatively broad extent. Geology, topography, and climate all influence the characteristics of a river basin or watershed ecosystem (e.g., Minshall et al., 1985) and thus act at the scale of the landscape (Omernik, 1987; Hughes and Larsen, 1988). Landscape patterns (such as regional or river basin) influence many ecological phenomena in inland aquatic ecosystems (Hughes et al., 1986; Karr, 1991). For example, streamflow characteristics vary with the type of soils and underlying geology; the topographic relief; and the form, amount, and timing of precipitation. The resulting flow regime in turn influences a variety of ecologically relevant features, including channel form, substratum size composition and stability, woody debris, and the nature of the food base. Patterns of stream discharge and disturbance regimes show a strong geographic separation (Minshall, 1988; Poff and Ward, 1989), implying the operation of landscape-level phenomena. Differences in flow regimes, coupled with climatically mediated thermal characteristics and geologically determined substratum characteristics,

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 300 are expected to contribute to correspondingly different geographic and other landscape-level patterns in biotic structure and function (Vannote et al., 1980; Minshall et al., 1985; Poff and Ward, 1990). Patterns of disturbances, both natural and human induced, resulting in a mosaic of patches of different ages and composition also may enter into the natural pattern of variability (White and Pickett, 1985). For example, the River Continuum Concept proposes that there are a number of features of even pristine stream-riparian ecosystems that change progressively throughout a river basin, therefore requiring a landscape perspective for proper interpretation (Vannote et al., 1980). The influence of riparian vegetation, the annual amount of terrestrial leaf litter in the channel, the availability of dissolved organic matter, and the modal size of particulate organic matter all generally decrease with distance from the headwaters of a stream system. The relative contributions of photosynthesis and community metabolism and the composition of macroinvertebrate functional feeding groups also change gradually and in a predictable fashion along the so-called river continuum. In addition, the effects of disturbances vary along a river system; some (especially if widely dispersed) become dissipated with increasing stream size, whereas others may act cumulatively. For example, the effects of moderate amounts of logging or farming within a basin or the entrance of low levels of sewage or nutrients from a point source may be dissipated through biological means and physical dilution as the water proceeds downstream. The effect of an isolated incident, such as the building of a single low-head dam, may be imperceptible, whereas the erection of many such dams can cumulatively have numerous adverse effects on many aspects of the downstream riverine ecosystem. Landscape is the scale of many forestwide and forestwide-regionwide land uses (e.g., logging, mining, livestock grazing) and their associated management practices that affect aquatic-riparian ecosystems. It also is the focus of many larger-scale problems for resource managers (e.g., drought, acid rain, forest and range fires, disease, and pest outbreaks). The catchment is an appropriate landscape unit for examining streamriparian ecosystem responses to disturbances on the order of years to decades (Minshall, 1993). However, for events occurring at intervals of 102 to 104 years, such as wildfire, the focus becomes the entire forest, which itself may cover numerous catchments. Landscape-scale events may affect aquatic ecosystems at various lower levels of resolution because of the hierarchical nature of these systems (Harris, 1980; Frissell et al., 1986; Pringle et al., 1988). The scale of an ecological system refers to its spatial and temporal dimensions (Allen and Hoekstra, 1992). The concept of ecological integrity is scale dependent (King, 1993). Maintenance of ecological integrity

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 301 implies perpetuation of some normal state or norm of operation within some prescribed range of variation. Thus, measuring or observing ecosystem integrity, or its loss, requires observations over sufficient space and time to identify the range of variation (King, 1993). Aquatic ecosystems require certain spatial and temporal bounds for the maintenance of their structure and function. For example, many stream ecosystems are dependent on terrestrial leaves and other forms of allochthonous detritus for food and habitat. Thus, severing the connection between the stream and riparian vegetation, because of road building, logging, or livestock grazing, can disrupt the integrity of the stream ecosystem. On a larger scale, attempts to manage or rehabilitate a river without regard to actions in the upstream portions of the river and its tributaries (such as land-use practices that alter the amount, kind, and timing of allochthonous detritus inputs) will be largely ineffectual. A minimum extent may be required for some processes to operate or interaction to take place. For example, nutrient cycling in streams occurs in a spiraling fashion along the water course, with important excursions into the hyporheic and floodplain zones (Newbold et al., 1982a,b; Green and Kauffman, 1989; Triska et al., 1989). Efforts to assess, protect, or restore biological integrity that do not use the appropriate scale are destined to fail. Failure to observe the system at the appropriate spatial-temporal scale can make inferences about ecological integrity of ecosystems difficult or impossible (e.g., Frost et al., 1988). Furthermore, setting spatial or temporal boundaries of a system smaller than the minimum required for persistence and interaction can affect system function and actually may lead to a loss of ecosystem integrity (King, 1993). For example, it is now clear that survival of native anadromous fish populations over much of the Pacific Northwest is dependent on resource management decisions made throughout the entire Columbia River Basin rather than on a stream-by-stream basis, as done in the past. Ecosystems in particular must be defined simultaneously in terms of space and time, since ecological dynamics occur over a broad spectrum of space-time scales (O'Neill et al., 1986; Allen and Hoekstra, 1992). For example, lake and stream ecosystem responses occur at scales ranging from millimeters and seconds to hundreds of kilometers and millions of years (Frost et al., 1988; Minshall, 1988, 1993). Small-scale events, such as the rolling of a rock or a 5°C change in temperature, recur with relatively high frequency; larger-scale events, such as a major flood, forest fire, or volcanic eruption, are progressively more rare. Phosphorus released by feeding zooplankton is taken up immediately by algae (Lehman and Scavia, 1982), whereas acid additions to a whole lake take several years to show effects on adult lake trout (Schindler et al., 1985). Recovery of species abundance on individually disturbed rocks in an otherwise intact streambed occurs within a few days to weeks (Robinson and Minshall,

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 302 1986), whereas recovery from the effects of a catchment-wide fire or a massive dam failure takes tens to hundreds of years (Minshall et al., 1983, 1989; Richards and Minshall, 1992). The demise of salmonid populations in the Pacific Northwest, referred to earlier, is symptomatic of a region- or basinwide loss of ecological integrity within the present century. Importance of Seasonal Variations in Time Variations in activity, condition, distribution, and abundance (hence, recruitment and/or mortality) of aquatic organisms across seasons are common. This is to be expected in strongly seasonal (temperate) environments, but such variations are found even in the tropics (Covich, 1988). Temporal variation in biotic responses should be accounted for in biological assessments of environmental conditions or determinations of change, but frequently it is not. Consideration of seasonal differences is especially important when comparing data from different locations or for the same area over time. For assessment of long-term trends, samples must be collected at the same general time within a season. However, in comparative studies among sites or years, sampling times should be determined on the basis of cumulative temperatures (e.g., number of degree-days) rather than specific calendar date. Temperature often is the primary factor responsible for seasonal variations in the temperate parts of the world. Even when light is the primary factor responsible, temperature is a reasonable surrogate because both are strongly influenced by the amount of solar radiation reaching the surface of a water body. This implies the need for continuous long-term records of temperature from the waters in question and/or the development of reliable regressions between air temperature or solar radiation and water temperature for specific localities and habitats of interest. Nonequilibrium Nature of Aquatic Systems and the Role of Disturbance Recently, there has been a paradigm shift from a belief in the dominance of equilibrium processes in ecology to one that emphasizes the importance of nonequilibrium processes (e.g., Harris, 1986; Botkin, 1990; Reice, 1994). Previously, the dynamics within ecological levels of organization, from populations through ecosystems, were viewed as being controlled primarily by processes that were density dependent and tended toward equilibrium conditions. The present view, whose implications have yet to be fully appreciated by most ecologists and resource managers, is that these same dynamics are controlled largely by processes that are density independent and of a nonequilibrium type. Consequently, they are believed to be heavily dependent on random (stochastic) forces and hence to be disturbance driven and variable rather than constant. Though reality probably lies somewhere between these two extremes, the latter currently

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 303 dominates ecological thinking. Nonequilibrium theory suggests that the development of ecological systems is nonlinear, heavily influenced by discontinuous catastrophic disturbances, and hence largely unpredictable and with multiple developmental pathways (Kay, 1991). Whether aquatic ecosystems are perceived as equilibrium or nonequilibrium actually may depend on the spatiotemporal scale being considered (O'Neill et al., 1986) and on the magnitude and time since disturbance. For example, in a year-to-year and section-by-section context, most natural stream ecosystems may be perceived to be open and nonequilibrium in character. They receive substantial inputs (environmental influences) from outside their boundaries and exert comparable influences (outputs) on adjacent ecosystems. They also are dynamic and continually changing. Many lakes possess similar features but usually to a lesser extent (e.g., Wetzel, 1983). However, when viewed in broader contexts, aquatic systems may exhibit several levels of stable behavior and show substantial spatial homogeneity (i.e., attributes associated with equilibrium conditions) (e.g., Harris, 1986; Frost et al., 1988; Minshall, 1988, 1993). Disturbance and the resultant change in conditions have long been recognized as an important factor affecting the structure and dynamics of ecological systems at various levels of organization. More recently, emphasis has shifted from viewing disturbance as rare and unpredictable to treating it as a natural process that occurs at different spatial and temporal scales with varying degrees of predictability (e.g., Pickett and White, 1985; Resh et al., 1988; Townsend, 1989; Fisher, 1990). Postdisturbance ecosystem development should be expected to exhibit the characteristics of self-organizing, nonequilibrium systems (Kay, 1991). The development of such systems is expected to proceed in irregular spurts from one steady state to another. Each spurt results in the system moving further from equilibrium and becoming more organized. For example, in ecosystem succession, each of the stages (seres) corresponds to a transient steady state, and the displacement of a previous seral stage by the next is a spurt that results in increased organization (Kay, 1991). Because change (both natural and human induced) is implicit in the modern, nonequilibrium view of ecosystems, its consideration is important in developing and applying the concept of ecosystem integrity to inland waters. Also, since ecosystem integrity is a scale-dependent concept, measuring or observing integrity or its loss in inland aquatic ecosystems requires observations over sufficient temporal extent to identify and characterize their patterns (King, 1993). Tools for Measuring Biological Integrity Tools for measuring biological integrity can be divided into two main groups: (1) those that measure ''structural" integrity, meaning those that

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 304 analyze the building blocks of aquatic communities; and (2) those that measure "functional" integrity, meaning tools that assess processes occurring in aquatic communities (Cairns, 1977a,b; Karr, 1991, 1993). Collectively, these structural and functional measures help restore a biological focus in water quality measurements. Structural Attributes An unusual change in one or more structural characteristics is interpreted as evidence of ecological stress. Impairment of biological structure of aquatic communities may be indicated by the absence of pollution-sensitive taxa, dominance by any particular taxon and low overall taxa richness, or changes in community composition relative to the reference condition (Plafkin et al., 1989). Fundamental measurements of ecosystem structure are (1) the number of species or other taxonomic units present, (2) the number (or mass) of individuals per species, and (3) the particular kinds of species present. Historically, aquatic ecologists have done a fairly good job of measuring structural aspects, although they are hampered by the inability to identify all of the species in a biotic community because of the lack of accurate, up-to-date taxonomic keys and comprehensive systematic treatises at the species level, particularly those of a regional nature. A still greater problem is the limited availability of qualified biologists capable of using even the existing taxonomic keys. The particular structural attributes analyzed to determine biological integrity vary with the level of developmental complexity (plants, invertebrates, fish) and ecological organization (population, community, ecosystem) of the focal organism or group (e.g., Johnson et al., 1993). The most commonly used groups for determining structural responses vary with the type of aquatic habitat under investigation and the scale appropriate to the question being addressed. In lentic waters, microscopic algae and invertebrates (particularly rotifers, cladocerans, and copepods) are most commonly emphasized. These may be separated further into those suspended in the water (i.e., plankton) and those growing attached to natural or introduced substrata (i.e., aufwuchs on sticks or microscope slides, respectively). In lotic waters, the most commonly studied groups are attached algae (periphyton), macroscopic invertebrates living on the stream bottom (benthos), and fish. The suitability of a particular group is determined by the specific environmental stressor, the generation time of the organisms, and their mobility and dispersal capabilities. For example, fish may be better indicators of physical habitat conditions in streams and benthic macroinvertebrates more indicative of water quality conditions (C. Yoder, personal communication, 1994). Algae have generation times of hours or days, macroinvertebrates

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 305 of months or a year or more, and fish of several years. Consequently, the algae may have already recovered from an environmental impact before an investigator arrives on the scene, whereas fish populations may not recover quickly enough to show the effects of multiple episodes within a season or year. Fish are more mobile than attached algae or benthic invertebrates and thus may avoid the effects of a pollutant or other disturbance entirely. High mobility also makes it difficult to obtain an accurate population estimate for fish. However, small relatively immobile organisms, such as algae or snails, may have higher dispersal capabilities than larger more mobile forms because they are often carried long distances by air currents or birds. Frequently, only a single group is used in a bioassessment but the use of several groups is preferred because they provide a better coverage of trophic levels and functional groups, sensitivity to a broader range of stresses, and a greater array of response times. At the population level, the presence and abundance of one or more key species or so-called indicator organisms may be used to determine the condition of the aquatic environment. However, this approach has been effective only in the limited cases where the monitored species responded clearly to specific types of water quality. To be useful as an indicator organism, an individual species must have a narrow range of tolerance for suitable environmental conditions that are known and related to some attribute of interest to humans. Few species satisfy these requirements because tolerance of a narrow range of conditions means that the organism may not be found widely in space or time and therefore will be of limited general utility as an indicator of environmental stress (Warren, 1971). Individual species often are not satisfactory for detecting gradual increases or decreases in pollution and are not sensitive to low levels of pollution. In addition, individual species may flourish or languish for reasons (such as competitors or predators, substratum or food conditions, or current velocity) that have nothing to do with pollution. Other measures at the population level may be more responsive to ecological dynamics than is species abundance. These are built on the properties of individuals (size, growth rate, content of particular components such as fats or certain enzymes) or populations (birth and death rates, population growth rate, age-frequency distribution). For example, a decline in birth rate may signal a significant change in environmental conditions and therefore provide a possible assessment approach for detecting human-induced impacts (W. T. Edmondson, personal communication, 1994). Similar methods could be applied to insects and other macroinvertebrates in streams. The effects of toxic substances on individual species have been studied widely under laboratory conditions, but rarely are toxicity tests applied to intact or partially isolated systems in nature (Rand and Petrocelli, 1985). Mortality generally is the criterion in such tests, but a sublethal condition

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 306 is just as important in controlling populations (Edmondson, 1993). Properties other than mortality (such as those given above) also vary with toxicity. Neoplastic lesions in fish have been correlated with pollutants (Russell and Kotin, 1957; Tyler et al., 1991). Chironomid head capsule morphology (Warwick, 1988) and diatom shell structure (W. T. Edmondson, personal communication, 1994) can be distorted when the organisms live in conditions containing sublethal toxic material. In Lake Orta in north Italy, copper-rich waste discharged from a rayon factory for many years caused elimination of the entire biota. The sedimentary record for the lake shows an initially normal diatom community; at progressively shallower levels, there are increasing proportions of distorted diatoms, gradual elimination of the various species, and finally no biota. Whether other metals would have the same effect as copper or whether different species of diatoms respond differently to different metals is not known. Population guilds and communities generally have proven more satisfactory than individual indicator organisms for the assessment of water quality. For example, paleolimnologists working on acid rain have been able to determine the pH of lakes on the basis of groups of diatoms found in bottom sediments. In a slightly different approach, planktonic-diatom biomass estimates were examined for the sediments of three lakes with contrasting types of lake disturbance: acidification (Gaffeln, southwestern Sweden), point-source eutrophication (Lough Augher, northern Ireland), and catchment agriculture (Akassjon, Northern Sweden) (Anderson, 1994). At Gaffeln, biomass declined steadily with acidification until extinction of the planktonic diatoms occurred. In Akassjon, maximum biomass coincided with the maximum areal extent of arable land. At Augher, nutrient input from a creamery resulted in steady biomass increases. Multivariate statistical analyses of species groupings have proven to be more useful than individual indicator organisms in a number of studies, particularly those involving pollution (e.g., Clarke, 1993). Various methods are employed to obtain samples of the biota for use in water quality assessment. These vary with the type of habitat (e.g., lotic or lentic), location (e.g., open water or benthic), size (microscopic or macroscopic), type of substratum (e.g., mud or rock), and degree of mobility of the organisms being sampled. Application of these methods requires skill and training. Planktonic forms may be trapped, pumped, enclosed, or netted from specific or composite locations within the water column (Lind, 1985; Wetzel and Likens, 1991). Attached microscopic forms can be scraped or brushed to free them from the substratum and then removed by suction for placement in a sample container or collection onto a filter (Lind, 1985; Porter et al., 1993). Benthic macroinvertebrates are collected by a variety of dredges, nets, corers, baskets, and other devices depending on water depth, substratum type, and current velocity (Lind, 1985; Wetzel and Likens, 1991; Cuffney et al., 1993a,b). Fish commonly are collected

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 307 by electrofishing, seining, or gill netting, or are observed directly by snorkeling or scuba diving (Meador et al., 1993b). Selection of methods is best made based on experience and reference to reputable published sources (e.g., APHA, 1989; MacDonald et al., 1991; Meador et al., 1993a, in addition to those cited above). In addition, important aspects of statistical design and data analysis must be incorporated into the study plan, but these are beyond the scope of this paper (e.g., Green, 1979; MacDonald et al., 1991; Norris and Georges, 1993; Resh and McElravy, 1993). Bioassessment of pollution in inland waters has been going on for a long time (e.g., Hynes, 1960; Karr, 1991; Cairns and Pratt, 1993), and a number of different measures and indices have been proposed. However, the recent renewed interest in biological integrity and its assessment has resulted in a reevaluation of those metrics that seem most effective, in an attempt to come up with standardized protocols (e.g., Karr et al., 1986; Plafkin et al., 1989; Karr, 1991; Resh and Jackson, 1993; Kerans and Karr, 1994). After representative samples of the biota are obtained, the organisms are identified and enumerated, their biomass and condition are determined, and relevant metrics are calculated. A biological "metric" is an absolute or derived measure that is sensitive to ecological conditions. The number of species present in a community is one of the simplest and most reliable metrics. A variety of other metrics have been examined for use in bioassessment, although additional testing and refinement still are needed. Some are sensitive to certain types of pollution, such as organic matter or inorganic sediments, whereas others are more general in their response. Some of the most popular metrics for periphyton, macroinvertebrates, and fish are listed in Table 1. Adjustments in the list are necessary to accommodate regional differences in distribution and assemblage structure and function (e.g., Miller et al., 1988). Richness is the number of different kinds of individuals (usually species or genera) in the total community or a specified assemblage. The Pinkham and Pearson community similarity index incorporates abundance and species composition (Plafkin et al., 1989). The quantitative similarity index for taxa compares two communities in terms of presence or absence of taxa and relative abundance (Shackleford, 1988). The Hilsenhoff biotic index summarizes the tolerances of macroinvertebrates to organic pollution (Hilsenhoff, 1987, 1988; Plafkin et al., 1989). Dominance measures assume that a highly skewed species- abundance distribution reflects an impaired community. The Hydropsychidae: Trichoptera ratio includes the mildly pollution-tolerant hydropsychids but excludes the pollution-intolerant arctopsychids from the family total. Functional feeding group designations are based on the manner in which food is obtained (see Merritt and Cummins, 1984, for details). Scrapers rasp or chew food growing attached to a surface such as rock, wood, or living aquatic vascular plants. Filterers employ self-constructed nets or specialized anatomical

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 308 structures to remove food particles from the water. Shredders feed on leaves, needles, sticks, and other coarse-sized particles (particularly of terrestrial origin). TABLE 1 Commonly Used Biological Metrics in the Bioassessment of Water Quality and Physical Habitat Conditions Attached algae (periphyton)a Soft-bodied forms Taxa richness (number of genera) Dominant phylum Specific indicator taxa of different levels and causes of pollution Diatoms Shannon diversity index Pollution tolerance index (after Lange-Bertalot, 1979) Siltation index (relative abundance of motile forms) Similarity index (Whittaker and Fairbanks, 1958) Macroinvertebratesb Structure Taxa richness Richness of taxa within the generally pollution-sensitive orders Ephemeroptera- Plecoptera-Trichoptera) (= EPT index) Pinkham-Pearson index Quantitative Similarity index for taxa Community balance Hilsenhoff biotic index Proportion of individuals in the dominant taxon Dominants-in-common at each site for five most abundant taxa Proportion of Hydropsychidae individuals of the total number of Trichoptera Functional feeding group components Proportion of scraper abundance to sum of scrapers plus filterers Proportion of shredder abundance to total individuals Quantitative similarity index for functional feeding groups Fishc Species composition Total number of fish species Species richness and composition of selected groups (e.g., darters, suckers, sunfish) Presence of pollution-intolerant species Trophic composition Proportion of individuals as omnivores Proportion of individuals as insectivores Proportion of individuals as top carnivores Abundance and health Number of individuals in sample Proportion of individuals with disease or anomalies a From Bahls (1993). b From Barbour et al. (1992). c From Karr et al. (1986) and Leonard and Orth (1986). Table 2 compares values of macroinvertebrates from five unaffected

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 309 reference streams (i.e., located in relatively undisturbed catchments) and five affected streams (catchments heavily grazed by livestock) in two different ecoregions in southern Idaho (Robinson and Minshall, 1995). Values are presented for each of the individual metrics examined in the study. Total scores (i.e., summed index values) based on all 18 metrics and on the 8 to 10 most responsive metrics identified through principal components analysis (PCA) are given at the bottom of the table. Some metrics are more responsive (and consistent) than others to the effects of livestock grazing. For example, in the Snake River Plain (SRP) ecoregion, the metrics Ephemeroptera-Plecoptera- Trichoptera (EPT) richness, Hilsenhoff biotic index (HBI), and biotic condition index (BCI) show consistent

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 310 responses to grazing; in the Northern Basin and Range (NBR) region, EPT/ Chironomidae, taxa richness, EPT richness, HBI, BCI, percent dominance, Shannon-Weiner diversity (H'), Simpson's index, and percent EPT show the most consistent responses. Nevertheless, most metrics clearly show lower values for the affected sites than for the unaffected ones. Also, the metrics differ in their effectiveness between ecoregions. For example, of the 18 metrics, 9 showed a large and consistent difference between affected and unaffected streams in the NBR but only 4 did so in the SRP. Often a variety of different scores are seen for a given metric, even within the unaffected streams, presumably reflecting differences in environmental effects. Examination of the total metric scores (Figure 1) showed substantially lower values for the affected streams in both ecoregions, thereby quantifying the effects of livestock grazing. In using the full 18 metrics, the condition of both the grazed and the ungrazed treatment types appears comparable between the two ecoregions. However, in using the subset of metrics selected by PCA, the unaffected streams in NBR are shown to be in better condition and the affected streams in worse condition than those in SRP, FIGURE 1 Health of macroinvertebrate communities in streams affected and unaffected by livestock grazing in two Idaho ecoregions: the Snake River Plain (SRP) and Northern Basin and Range (NBR). SOURCE: Based on data reported in Robinson and Minshall (1995). © 1995 by CRC Press.

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 311 a fact that supports on-site impressions of stream conditions in the two regions. Functional Attributes Functional integrity involves processes such as photosynthesis and community respiration, nutrient transfer, energy flow (secondary production), and decomposition. Abnormal rates of activity and accumulation or depletion of materials are indications of disruptions of an ecosystem's functional integrity. For example, during the course of a whole-lake acidification experiment, the entire nitrification process was halted at a pH of less than 5.7, possibly due to the loss of all nitrifying bacteria (Rudd et al., 1988). High rates of ecosystem metabolism commonly are associated with eutrophication; lowered or negative rates of metabolism may indicate active decomposition of excess supplies of organic matter trapped in the sediments. To distinguish abnormal conditions from normal ones, the natural range of variability must be known for each particular location (e.g., Frost et al., 1988). Functional measures of ecosystem integrity are often considered less sensitive to environmental factors, including pollutants, than measures of structure. Also, system function may give a very different impression of the effects of environmental stress than structural responses, particularly species composition (Frost et al., 1995). However, these apparent anomalies simply may occur because functional measures operate at longer temporal or larger spatial scales than structural measures or because their responses are less well known. It often is assumed that ecosystems are functionally resilient to alteration of structure due to compensatory responses, but this assumption has not been tested adequately for aquatic ecosystems. A compensatory functional response occurs when rates and amounts of ecosystem processes remain unchanged in the face of changes in structural composition, such as alteration of species dominance or loss of species richness. Studies of compensation and complementarity in ecosystem function are few and limited largely to lakes (Schindler, 1987; Howarth, 1991; Frost et al., 1995). The loss of species from ecosystems produces inconsistent results. In some cases, such losses are accompanied by no apparent compensation (e.g., primary production remains unchanged); in others, considerable alteration of ecosystem processes occurs (e.g., Kitchell and Carpenter, 1993). Understanding the reasons for such apparent inconsistency is an important area for future research. Measurement of ecosystem function has been avoided because methods dealing with it have been lacking or are thought to be more difficult and time- consuming to employ. Although it often is easier to evaluate aspects of system structure, this is not always the case. For example, it generally is easier to make a measurement of open-water community metabolism

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 312 in a lake than to collect, identify, count or weigh, and determine the appropriate structural metrics for even one group of organisms. In addition, freshwater ecologists now have the fundamental tools needed to begin assessing some functional aspects of ecosystem integrity (e.g., metabolism chambers, nutrient uptake techniques, leaft pack and litter bag decomposition). Although methods are now available to begin adding functional measures to the assessment of ecological integrity in inland waters, many are still crude and simplistic. Considerable refinement in methodology and the development of additional approaches and techniques are sorely needed. In addition, further efforts are needed to develop practical, cost-effective techniques for routine bioassessment and to train field technicians in their use. All aquatic life is dependent, either directly or indirectly, on the photosynthetic fixation of radiant and chemical energy into primary organic compounds. Part of this fixation occurs directly in the aquatic environment through primary production (i.e, "self-produced," or autochthonous); the remainder occurs on land, with the organic matter entering the aquatic environment in the form of leaves and other terrestrial plant products (i.e., "produced outside," or allochthonous). Thus, primary production is an important measure of both the autotrophic capacity of a water body and the availability of food resources for the various consumer levels in the food web. In addition, when used in conjunction with measurement of community respiration, it provides useful insights into the relative importance of autochthonous versus allochthonous carbon resources to energy flow in an ecosystem. Primary production in inland waters is measured by determining the amount of carbon dioxide taken up, the amount of higher-energy carbon compounds produced, or the amount of oxygen evolved in the presence of light. Measurement techniques all involve determination of changes in the amount of reactant or of one or more products in either "open systems" (i.e., the water in all or a portion of a basin or channel) or "closed systems" (i.e., sealed bottles, spheres, recirculating chambers, and the like containing isolated portions of water and components of the ecosystem). Although the measurement of primary production is a time-honored procedure in aquatic ecology, there is still considerable room for improvement. Difficulties include accurate correction for diffusion losses and gains across the air-water interface in open system measurements and nutrient depletion and the lack of adequate circulation (or flow) in closed systems. The potential for other artifacts due to enclosure in containers of various sizes and shapes is likely but has yet to be examined. Additional problems arise in interpreting the meaning of whole-system values when the contributions of various compartments or subsystems remain undetermined and in extrapolating from closed-system measurement of ecosystem fragments to realistic intact-system estimates.

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 313 Community respiration is the combined sum of the respiration by autotrophs and heterotrophs (e.g., bacteria, fungi, invertebrates, vertebrates). It measures the total consumptive process of the ecosystem of fixed carbon, whether autochthonous or allochthonous in origin. The procedure involves measurement of the amount of oxygen consumed, organic carbon compounds used, or carbon dioxide produced using techniques identical to those for primary production. Measurements are made in the dark in order to eliminate the confounding and counteracting effect of photosynthesis. Respiration also occurs during the daytime and is believed be higher than at night, but techniques for routine measurement of daytime respiration outside the artificial conditions of the laboratory are not available. A special case of measuring utilization of organic carbon compounds involves determination of the rate of mineralization of dead plant matter. This material may be exposed directly in the water in the form of loosely aggregated packs of tree leaves or other materials, or enclosed in mesh bags (Petersen and Cummins, 1974; Allard and Moreau, 1986; Cummins et al., 1989). Control of mesh size has been used to exclude or enclose selected decomposer groups (Minshall and Minshall, 1978), but in many cases, inhibition of decomposition due to the creation of anaerobic conditions may occur (Cummins et al., 1989). The usual measure is the amount of weight lost (after correction for leaching and handling) per time interval, adjusted for temperature. Stressed aquatic environments generally show rates of decomposition reduced from those of unstressed ones. In particular, increased levels of acid, chlorine, chlorine plus ammonia, and salt have been shown to lower litter decomposition (Reice and Wohlenberg, 1993). Another important process in aquatic ecosystems is biogeochemical cycling. In lakes and wetlands, cycling and uptake rates commonly are measured in vertical tubes or other types of containers (Harris, 1986; Wetzel and Likens, 1991). In streams, where the cycling does not occur in place but extends longitudinally downstream in a sort of corkscrew fashion, this process is termed ''spiraling." In flowing waters, several aspects can be measured, such as rate of downstream movement, rate of cycling, turnover length, and turnover time (Newbold et al., 1982a,b; Minshall et al., 1992), which may provide insights into the effect of various human and anthropogenic impacts. In addition, nutrient uptake rates are measured relatively easily by releasing known amounts of a nutrient (such as nitrogen or phosphorus) at an upstream location, allowing the nutrient-enriched water to flow over a representative stretch of stream, and determining the amount removed at some downstream point (Munn and Meyer, 1990; Stream Solute Workshop, 1990). Nutrient uptake rates can provide insights into an important aspect of nutrient cycling or spiraling and enable the measurement of various environmental stresses on this aspect of ecosystem functioning.

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 314 In addition to measures of ecosystem processes such as energy flow and nutrient cycling, measurement of functional integrity should include genetic and evolutionary aspects of the biota (Regier, 1993). Biological systems are in continual states of adjustment (adaptation) to their environment and evolve over time. Thus, the scale of their response and the patterns subsequently produced can be expected to be the product of selection by long-term evolution. Failure to address evolutionary aspects adequately has led to major misconceptions regarding ecosystem properties and processes such as succession (Hagen, 1992; Colinvaux, 1993). The dilution, isolation, and extinction of genetic pools are bound to be major problems in inland waters both now and in the future (Noss and Cooper-rider, 1994). Awareness of this problem is just becoming widespread and is restricted mainly to fish and mollusks (Williams and Miller, 1990; Nehlsen et al., 1992; Bogan, 1993), but effects on other aquatic groups are expected to be equally severe (e.g., Zwick, 1992). Methods for measuring these long-term fitness features of ecosystem functional integrity have only recently begun to be developed for freshwater organisms. They are often more difficult to apply to field conditions and require larger numbers of samples than other types of processes because of the sensitive, tedious, and specialized laboratory analyses involved (Funk and Sweeney, 1990; Robinson et al., 1992) or the detailed life history, functional, and behavioral information required (e.g., Resh et al., 1994; Usseglio-Polatera, 1994). Nonetheless, it is important to address genetic and evolutionary components in the assessment of the ecological integrity of inland waters. Two general types of tools are available to do this: (1) those that permit determination of the genetic makeup, particularly the extent of allelic heterozygosity and gene polymorphism, of populations within the ecosystem; and (2) those that assess the occurrence of various measurable traits expected to evolve in particular environments or be selected for under different types and frequencies of environmental stress. Most natural environments are predominantly nonequilibrium, populated by organisms whose populations have relatively high levels of heterozygosity (Hedrick, 1986). Genetic diversity will decline as populations near extinction. Species traits that are likely to contribute to fitness, and hence are likely to be selected for or against under particular environmental conditions, include (1) physiological adaptations to generally unfavorable physical conditions, (2) adaptations for defense, (3) food harvesting and somatic development, (4) reproduction, and (5) tactics for escape in space or time (Southwood, 1988). These traits are postulated to vary in a predictable manner in relation to the degree of stress (adversity) in the environment and the frequency of disturbance or extent of temporal heterogeneity (Southwood, 1977, 1988; Hildrew and Townsend, 1987). Environmental disturbances to which a population has not adapted over evolutionary

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 315 time may adversely affect its genetic diversity and threaten its long-term survival. Tools that permit determination of the degree of genetic heterozygosity rely on biochemical measures of gene frequencies and polymorphisms. One common biochemical approach is to use gel electrophoresis to identify loci associated with various enzymes (Shaw and Prasad, 1970; Harris and Hopkinson, 1976). The loci are then scored, based on relative distance of the bands from the origin and determinations of allelic frequencies and heterozygosities. Finally, population polymorphisms and population heterozygosities are calculated from the scored loci (Ayala, 1982). Measurement of species traits important to long-term fitness uses relevant morphological, physiological, and behavioral features of a population (e.g., size, body form, reproductive capacity, mode of respiration, dispersal ability, feeding method). This procedure involves the selection of traits, quantification of the extent to which the different life history stages of individual populations possess those traits, derivation of species-trait and species-habitat-type matrices, and evaluation of relationships between the two matrices using appropriate statistical analyses (Doledec and Statzner, 1994; Usseglio-Polatera, 1994). Bioassessment procedures that incorporate multiple measures (metrics) of the responses of population aggregates ("communities") are recommended because different measures are sensitive to different types of water quality impairment. Often individual metric scores are summed in the belief that a collective "signal" is easier to discern than individual ones (Plafkin et al., 1989; Karr, 1991, 1993). However, some of the metrics respond in opposite ways, many are biased toward a particular type of pollution (e.g., organic wastes), and not all types of pollution are represented or adequately determined. Therefore, the common practice of summing the results of individual metrics to obtain a single total score tends to conceal valuable information and to produce equivocal results. Additional work is needed to remove uninformative redundancy and develop metrics specific to different types of degradation. IMPLICATIONS FOR THE FUTURE Modern water science encompasses a broad array of skills and areas of expertise. Future scientists, teachers, and resource managers will need to be broadly trained in these areas. The complexity and magnitude of the questions facing researchers and resource managers will increasingly require an interdisciplinary approach and the ability to work cooperatively. The ecological integrity of inland waters is being assailed on many fronts. Direct assessment of the biota is crucial for the protection and management of aquatic resources. Sound understanding of basic biological

BRINGING BIOLOGY BACK INTO WATER QUALITY ASSESSMENTS 316 (ecological) relationships is prerequisite to sound management (Jumars, 1990; Edmondson, 1993). Several large federal monitoring and assessment programs under development emphasize the measurement of water quality in biological, rather than solely chemical-physical, terms. Consequently, the need for training in systematics, basic biology, and ecology of key groups of inland aquatic flora and fauna (e.g., diatom algae, macroinvertebrates, fish) will increase in the future. Considerations of scale are increasingly a part of the process by which aquatic ecologists approach a variety of ecological issues and problems (King, 1993). The effects of natural and human-caused factors on inland aquatic ecosystems require consideration at multiple spatiotemporal scales that include adequate heterogeneity across landscapes (Covich, 1993). Hierarchy theory commonly is used to address these questions of scale (Allen and Hoekstra, 1992). Thus, ecological integrity will have to address questions of scale and hierarchy; the approach will vary with the particular research question or management problem. However, for the immediate future, the ecosystem and landscape perspectives will be especially important if sustainable biological aquatic resources are to be protected adequately in the face of pressures from the burgeoning human population. Numerous pressing challenges face the future of inland waters, their study, and their management for sustained benefits and uses. Ecosystem, landscape, and global perspectives will be necessary to provide adequate quantity and high- quality water for human use and natural habitats (Covich, 1993). The remaining natural freshwater habitats that have high biodiversity or endemic species should be protected (e.g., Boon, 1992), and degraded waterways should be restored and their natural linkages reestablished. Basinwide planning and management are needed to protect and restore riverine ecosystems and avoid cumulative effects. Agency personnel and the public must be educated on the new strategies and techniques in aquatic ecosystem management and restoration at the catchment and basin levels. Improved funding is necessary for research and education to enhance information, improve skills, and increase the number of personnel to ensure proper management of sustainable inland aquatic ecosystems. Most of the methods being applied to the evaluation of structural and functional attributes of inland aquatic ecosystems have been around for a long time. Although there is much to be said for the use of widely accepted, time- tested approaches, there is also the danger that complacency will lead to lack of needed improvement and innovation. Continued refinement of existing methods is necessary. At the same time, many exciting new developments in biology— including genetic markers; molecular, morphological, and behavioral indicators of exposure to toxic substances; and molecular measures of function—are emerging as fertile

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Freshwater Ecosystems: Revitalizing Educational Programs in Limnology Get This Book
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To fulfill its commitment to clean water, the United States depends on limnology, a multidisciplinary science that seeks to understand the behavior of freshwater bodies by integrating aspects of all basic sciences—from chemistry and fluid mechanics to botany, ichthyology, and microbiology. Now, prominent limnologists are concerned about this important field, citing the lack of adequate educational programs and other issues.

Freshwater Ecosystems responds with recommendations for strengthening the field and ensuring the readiness of the next generation of practitioners. Highlighted with case studies, this book explores limnology's place in the university structure and the need for curriculum reform, with concrete suggestions for curricula and field research at the undergraduate, graduate, and postdoctoral levels. The volume examines the wide-ranging career opportunities for limnologists and recommends strategies for integrating limnology more fully into water resource decision management.

Freshwater Ecosystems tells the story of limnology and its most prominent practitioners and examines the current strengths and weaknesses of the field. The committee discusses how limnology can contribute to appropriate policies for industrial waste, wetlands destruction, the release of greenhouse gases, extensive damming of rivers, the zebra mussel and other "invasions" of species—the broad spectrum of problems that threaten the nation's freshwater supply. Freshwater Ecosystems provides the foundation for improving a field whose importance will continue to increase as human populations grow and place even greater demands on freshwater resources. This volume will be of value to administrators of university and government science programs, faculty and students in aquatic science, aquatic resource managers, and clean-water advocates—and it is readily accessible to the concerned individual.

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