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--> Biological Criteria for Water Resource Management Chris O. Yoder and Edward T. Rankin This paper has two goals: to describe a framework for developing biological criteria for water-quality assessment and to suggest what roles biological indicators should have in water resource management and policy. A principal objective of the Clean Water Act is to restore and maintain the physical, chemical, and biological integrity of the nation's surface waters (Clean Water Act Section 101[a]). Although this goal is fundamentally ecological, regulatory agencies have attempted to reach it by measuring chemical and physical, but not biological, variables (Karr et al., 1986). The rationale for this choice is well known; the chemical water-quality criteria developed through laboratory toxicity tests on selected aquatic organisms serve as surrogates for the ecologically based goals of the Clean Water Act. The presumption that improvements in chemical water quality will restore biological integrity has come into question during the past 20 years. The traditional chemical water-quality approach may give an impression of empirical validity and legal defensibility, but it does not directly measure the ecological health and well-being of surface water resources. Nor does it comprehensively address the definition of pollution in Section 502 of the Clean Water Act: ". . . man-made or man-induced alteration of the chemical, physical, biological or radiological integrity of water . . ." (Karr, 1991). State and federal programs have become focused on the objective of controlling point-source discharges of chemicals. However, a growing number of professionals and various stakeholders have become convinced that an attack on point sources of toxins alone will not fully achieve the Clean Water Act's goal related to surface waters. In addition to an overemphasis on point sources and toxins, many state and federal programs also suffer from
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--> an overreliance on prescriptive approaches to management and regulation; a tendency to rely on anecdotal information; and an emphasis on administrative activities that frequently results in a skewed allocation of resources among different programs. Finally, attainment of national goals is also hindered by a lack of consistency in the environmental statistics reported by different states. One result is that in some cases, well-intentioned but basically flawed management strategies have increased environmental degradation as shown, for example, by the Ohio Environmental Protection Agency (1992). Fortunately, the U.S. Environmental Protection Agency (EPA) and others are becoming increasingly aware of these shortcomings and have initiated efforts such as the environmental indicators initiative and the Intergovernmental Task Force on Monitoring Water Quality (1992, 1993, 1995). Major Factors That Determine Water Resource Integrity Beyond chemical contaminants, multiple factors are responsible for the continuing decline of surface water resources in Ohio (Ohio Environmental Protection Agency, 1995) and the United States (Benke, 1990; Judy et al., 1984). These include the modification and destruction of riparian habitat, sedimentation of bottom substrates, and alteration of natural flow regimes. Because biological integrity is affected by many factors, controlling chemicals alone does not assure its protection or restoration (Figure 1). We need a broader focus on the entire water resource if we are to successfully reverse the decline in the overall quality of the nation's waters. Therefore, ecological concepts and biological criteria must be further incorporated into the management of surface water resources. Disparities in the Use of Indicators Although a growing number of states and organizations rely primarily on biological indicators to assess the condition of their water resources, others choose to emphasize chemical and physical indicators. The following examples demonstrate the inherent risks of relying solely on these indicators. Out of 645 stream and river segments analyzed in Ohio, biological indicators revealed impairment in 49.8 percent of the segments where chemical indicators detected none (Ohio Environmental Protection Agency, 1990; Rankin and Yoder, 1990a). The converse was true for only 2.8 percent of stream segments. The remarkable discrepancy between biological and chemical assessments is due to fundamental differences in what they measure. Biological communities respond to a wide variety of chemical, physical, and biological factors. Thus, biological indicators are able to detect a wider range of environmental disturbances than can measures of chemical water-quality alone. Another example is the proportion of waters that various states reported were
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--> Figure 1 Five principal factors that influence and determine the integrity of surface water resources. SOURCE: Modified from Karr et al. (1986). impaired due to habitat degradation (Figure 2). Twenty-five of 58 states and territories that report such statistics claimed that zero miles of rivers and streams had been negatively affected by the modification of habitat. Of the states that did report such impairment, only 15 reported an effect on more than 100 miles. These statistics are difficult to believe given the pervasive nature of well-documented practices that modify habitat, such as flood control, impoundments, agriculture and forestry, resource extraction, and urban development (Benke, 1990; Judy et al., 1984). The wide variation in state statistics is probably due to the use of different indicators and programmatic biases toward the control of toxic chemicals and point-source discharges (Ohio Environmental Protection Agency, 1990; Rankin and Yoder, 1990a).
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--> Figure 2 Miles of habitat-impaired rivers and streams reported by the states to the U.S. Environmental Protection Agency. SOURCE: United States Environmental Protection Agency (1994). NOTE: Twenty-five states reported no miles of aquatic-life use impairment associated with habitat degradation. Ohio's use of quantitative biological criteria had some additional ramifications that affect the statistics related to the Clean Water Act 305[b]. For example, the proportion of stream miles that failed to attain standards increased from 9 percent in 1986 (based on a mix of chemical water-quality and qualitative biological assessments) to 44 percent in 1988, primarily because quantitative biological criteria were included in the assessment process beginning in 1988 (Ohio Environmental Protection Agency, 1988). The nearly fivefold increase in nonattainment illustrates the significant differences that can exist between states that use different assessment methods, especially whether or not biological assessments are included. These examples demonstrate that relying on chemical water-quality information alone is apt to result in underestimates of environmental degradation. Underestimates are especially likely when assessing watershed-level effects. This is because the interaction of aquatic and riparian habitats, land use, and nutrient dynamics is particularly difficult to measure and characterize without using robust biological assessment tools and indicators. Ironically, much of the concern expressed about using biological criteria has been over the risk of failing to detect water-quality impairment. This concern seems misplaced in light of the preceding examples.
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--> The Complementary Roles of Different Indicators The EPA Environmental Monitoring and Assessment Program (United States Environmental Protection Agency, 1991) distinguishes environmental indicators on the basis of whether they best measure stresses, exposures, or responses. Stressor indicators identify activities that have an impact on the environment. These include land-use changes and discharges of pollutants. Exposure indicators identify components of the environment that have been subjected to a substance or activity that could potentially change the environment directly or indirectly. Response indicators detect the status of a particular resource component, usually biological, in relation to external stresses and exposure to those stresses. Particular indicators function most appropriately in one of the three indicator categories, although they may double as a secondary evaluator of another indicator class. For example, chemical measures generally function best as exposure indicators but may indirectly provide insights to response. Biological measures are inherently response oriented and may or may not provide more than qualitative insights to exposure and stressors. These comparisons of chemical and biological indicators illustrate a national problem: the inappropriate use of stressor and exposure indicators as substitutes for response indicators. States that do not have well-developed bioassessment programs still must report on the status of their waters to EPA on a biennial basis. Thus, they are forced to use whatever information is available. Usually, the readily available information is in the form of stressor or exposure indicators. In attempting to resolve the obvious inconsistencies in measuring the condition of aquatic resources, a fundamental step is to recognize and establish appropriate roles for the different chemical, physical, and biological indicators. An accurate portrayal of the condition of the nation's surface waters depends on the use of suites of these indicators, each in their appropriate role as stressor, exposure, or response indicators. Development of Biological Criteria Underlying Theory and Concepts Without a sound theoretical basis, it would be difficult if not impossible to develop biological criteria and meaningful measures of ecological condition. Obvious ecological degradation such as fish kills stimulated the landmark environmental legislation of the past 2 decades, but that biological focus was lost in the quest for easily measurable water-quality indicators (Karr, 1991). The biological integrity provision of the Clean Water Act, which was initially difficult to specify in practice (Ballentine and Guarria, 1975), was eventually defined by Karr and Dudley ( 1981 ) as ". . . the ability of an aquatic ecosystem to support and maintain a balanced, integrated, adaptive community of organisms having a spe-
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--> cies composition, diversity, and functional organization comparable to that of the natural habitats of a region." It was this definition that provided the theoretical underpinnings for developing a framework within which quantitative biological criteria could be derived. The essential concepts of how to measure and define biological performance, natural habitats, and regional variability were each dealt with through a number of key research projects in the early 1980s and together provided the framework and tools needed to derive biological criteria. Given the above definition, the eventual attainment of the Clean Water Act goal of biological integrity requires much more than merely achieving a high level of species diversity, numbers, and/or biomass. In fact, in some situations, increases in any one of these parameters can be a sign of degradation. Managers must also strive for more than the health and well-being of certain target species. The conservation of individual species, although necessary, is not sufficient for ensuring biological integrity. Conservation policy should promote management practices that maintain integrity, prevent endangerment, and enhance the recovery of species and ecosystems (Angermier and Williams, 1993). Water-resource management must include in its goal the objective of achieving self-sustaining, functionally healthy aquatic ecosystems. Achieving this state will foster other ecological goals as well because functionally healthy communities include the elements of biodiversity and rare species inherent to the more narrowly focused management efforts. We believe that biocriteria can play an important role in meeting these challenges. Understanding Biological Integrity: A Prerequisite to Biological Criteria The term biological integrity originated in Section 101[a] of the Federal Water Pollution Control Act amendments of 1972 and has remained a part of the subsequent reauthorizations. Early attempts to define biological integrity in ways that could be used to measure attainment of the legislative goals were inconclusive. One of the better known of these efforts failed to produce a consensus definition or framework for determining biological integrity (Ballentine and Guarria, 1975). Biological integrity was considered relative to conditions that existed prior to European settlement; the protection and propagation of balanced, indigenous populations; and ecosystems that are unperturbed by human activities. These criteria (especially the first and last) could be construed as referring to a pristine condition that exists in only a few, if any, ecosystems in the United States. Subsequently, an EPA-sponsored work group concluded that it is difficult and perhaps impractical to precisely define and assess biological integrity, when it is viewed as a pristine condition (Gakstatter et al., 1981). Rather, the pristine vision of biological integrity became a conceptual goal of pollution abatement efforts, although it may never be fully realized in many parts of the United States given current, past, and future uses of surface waters. More recently, efforts to construct a workable, practical definition of bio-
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--> logical integrity have provided the supporting theory needed to develop standardized measurement techniques and criteria to determine whether efforts are complying with that goal. Biological integrity is now defined as ". . . the ability of an aquatic ecosystem to support and maintain a balanced, integrated, adaptive community of organisms having a species composition, diversity, and functional organization comparable to that of the natural habitats of a region" (Karr and Dudley, 1981). This is a workable definition that directly alludes to the measurable characteristics of biological community structure and function found in the least-impacted habitats of a region. This definition and its underlying ecological theory provide the basis for developing quantitative biological criteria based on conditions at regional reference sites. The EPA adopted a facsimile of this definition in their biological criteria national program guidance (United States Environmental Protection Agency, 1990). The emerging issue of biodiversity should not be equated with biological integrity, even though the two concepts share many attributes (Karr, 1991). They differ in that biodiversity is primarily focused on ecosystem elements (i.e., genetic diversity, populations, bioreserves, etc.), whereas biological integrity includes these elements but also encompasses ecosystem processes (i.e., nutrient cycles, trophic interactions, speciation, etc.). The often-cited ecosystem approach to environmental management (e.g., Great Lakes Water Quality Initiative) can be even more restricted to dealing with elements that are not direct ecological parameters (i.e., chemical water-quality surrogates). Both the biodiversity and the ecosystem approaches would benefit by including the concept of biological integrity to improve the chances that each effort would succeed and assure that environmental problems are addressed from an ecological perspective. New Multimetric Biological-Community Evaluation Mechanisms A variety of quantitative indices for assessing biological data have been developed in the past 20 years. These indices represent significant advances because they use biological information for resource characterizations and for determining the attainment of environmental goals. Examples include the Index of Biotic Integrity, as originally developed by Karr (1981) and modified by many others (Leonard and Orth, 1986; Miller et al., 1988; Ohio Environmental Protection Agency, 1987b; Steedman, 1988); the Index of Well-Being (Gammon, 1976; Gammon et al., 1981); the Invertebrate Community Index (DeShon, 1995; Ohio Environmental Protection Agency, 1987b); the EPA Rapid Bioassessment Protocols for macroinvertebrate assemblages (Plafkin et al., 1989); and the Benthic Index of Biotic Integrity (Kerans and Karr, 1992). Although quantitative biological indices have been criticized for potentially oversimplifying complex ecological processes (Suter, 1993), raw data must be distilled to be interpretable. Multimetric evaluation mechanisms extract ecologically relevant information from complex biological data while preserving the
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--> opportunity to analyze such data on a multivariate basis. Several features of these multimetric indices minimize the problem of data variability. Variability is first controlled by specifying standardized methods and procedures (e.g., Ohio Environmental Protection Agency, 1989b) and providing iterative training exercises and supervision to implement them. Second, variability is in effect "compressed" through the application of multimetric evaluation mechanisms (e.g., Index of Biotic Integrity, Invertebrate Community Index), which reduce raw measurements into discrete, calibrated scoring categories. Last, variability is partitioned according to background factors that determine ecological potential (e.g., ecoregions), a process that results in a graduated set of criteria based on regional potential. The results are evaluation mechanisms, such as the Index of Biotic Integrity and the Invertebrate Community Index, that have acceptably low replicate variability (Davis and Lubin, 1989; Fore et al., 1993; Rankin and Yoder, 1990b; Stevens and Szczytko, 1990). Multimetric indices have been criticized as representing a loss of rich information because the data are reduced to a single index value. However, this presumes that the supporting data are never viewed or examined beyond the calculation of the index itself. These criticisms are without foundation: The need to examine subcomponents of the indices and even the raw data is implicit throughout the biocriteria process. Theoretically sound quantitative measures, as opposed to raw data, are clearly necessary throughout the process of environmental management. Although interpretation of raw data by qualified biologists will always be necessary, it is not realistic to expect that their qualitative judgment alone will be an acceptable substitute for a more empirical process. Fortunately, biological judgment can be incorporated into structured frameworks. These include the frameworks developed by the state of Maine using multivariate techniques (Davies et al., 1993), the state of Florida using a multimetric approach (Barbour et al., 1996), the EPA Rapid Bioassessment Protocols (Plafkin et al., 1989), and the regional reference-site approach (Hughes et al., 1986; Ohio Environmental Protection Agency, 1987b, 1989a; Yoder and Rankin, 1995a). Simply stated, multimetric indices can satisfy the demand for a straightforward numerical evaluation that expresses a relative value of aquatic community health and well-being and allows program managers (who are frequently nonscientists) to "visualize" relative levels of biological integrity. These measures also provide a means to establish quantitative biological criteria. Karr's Index of Biotic Integrity Modified by the Ohio Environmental Protection Agency The Index of Biotic Integrity originally proposed by Karr (1981) and later refined by Karr et al. (1986) and others incorporates 12 metrics (Table 1). These fall within four broad groupings: species richness and composition, trophic composition, environmental tolerance, and fish abundance and condition. Although no single metric consistently functions across an entire environmental gradient
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--> TABLE 1 Index of Biotic Integrity Metrics Used by the Ohio EPA to Evaluate Headwater Sites, Wading Sites, and Boat Sites Metric Headwater Sitesa,b Wading Sitesb Boat Sitesc 1. Number of native speciesd X X X 2. Number of darter species X Number of darter and sculpin species X Percent round-bodied suckerse X 3. Number of sunfish speciesf X X Number of headwater speciesg X 4. Number of sucker species X X Number of minnow species X 5. Number of intolerant species X X Number of sensitive speciesh X 6. Percent green sunfish Percent tolerant species X X X 7. Percent omnivores X X X 8. Percent insectivorous cyprinids Percent insectivores X X X 9. Percent top carnivores X X Percent pioneering speciesi X 10. Number of individuals Number of individuals (less tolerants)j X X X 11. Percent hybrids Percent simple lithophils X X. Number of simple lithophils X 12. Percent diseased individuals Percent DELT anomaliesk X X X a Sites with drainage areas <20 sq. mi. b Sampled with wading electrofishing methods. c Sampled with boat electrofishing methods. d Excludes all exotic and introduced species. e Includes all species of the genera Moxostoma, Hypentelium, Minytrema, and Ericymba, and excludes Catostomus commersoni. f Includes only Lepomis species. g Species designated as permanent residents of headwaters streams. h Includes species designated as intolerant and moderately intolerant (Ohio Environmental Protection Agency, 1987b). i Species designated as frequent and predominant inhabitants of temporal habitats in head- water streams. j Excludes all species designated as tolerant, hybrids, and non-native species. k Includes only animals with deformities (D), eroded (E) fins or barbels, lesions (L), or tumors (T). NOTE: This table lists the original Index of Biotic Integrity metrics of Karr (1981). In cases where the Ohio EPA uses modifications of the original metric, the modifications appear below the original metrics.
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--> and for all types of impacts, their aggregation in the Index of Biotic Integrity provides sufficient overlap and redundancy to yield a consistent and sensitive measure of biological integrity (Angermier and Karr, 1986). The index is a quantitative, ordinal, if not linear, measure that responds in an intuitively correct manner to known environmental gradients (Steedman, 1988). When incorporated with mapping, monitoring, and modeling information, the Index of Biotic Integrity is valuable for determining management and restoration requirements for warm-water streams (Bennet et al., 1993; Steedman, 1988). As an aggregation of community information, the Index of Biotic Integrity and its facsimiles provide a way to organize complex data and reduce it to a form that permits interpretation and comparisons with communities whose condition is known. Although the Index of Biotic Integrity incorporates elements of professional judgment, it also provides the basis for quantitative criteria for determining what constitutes exceptional, good, fair, poor, and very poor conditions. The process of tailoring the Index of Biotic Integrity to regional conditions represents an important example of the use of biological judgment in biological criteria (Miller et al., 1988). Streams and rivers occur in many sizes throughout Ohio. They contain different fish assemblages and must be sampled by different methods. Thus, the Ohio EPA needed to modify Karr's original index to apply it to these different stream sizes and adjust it to account for biases induced by different sampling gear. The three different modified Indices of Biotic Integrity that were developed for Ohio rivers and streams (Table 1; Ohio Environmental Protection Agency, 1987b) are: 1) a headwaters index for application to headwater streams (locations with a drainage area <20 mi2); 2) a wading-site index for application to stream locations with watersheds >20 mi2 sampled with wading methods; and 3) a boat-site index for locations sampled from boats. All modifications follow the guidance on metric modification provided by Karr et al. (1986). Although the Index of Biotic Integrity has worked well in Ohio and in many parts of the United States (Miller et al., 1988), Canada (Steedman, 1988), and Europe (Oberdorff and Hughes, 1992), problems have been encountered in semiarid western U.S. drainages (Bramblett and Fausch, 1991) and cold-water streams in northern states (Lyons, 1992). In both cases, the characteristics of the fauna differ from the presumptions made in the original index. Lyons (1992) was initially confounded by the degradation of cold-water streams in Wisconsin because it resulted in an increased diversity of fish species, a change that is counted as an improvement in water quality by the original Index of Biotic Integrity. He overcame these problems and constructed an index tailored to cold-water streams (Lyons and Wang, 1996). Bramblett and Fausch (1991) encountered difficulties due to a lack of pollution-intolerant species in the harsh and highly variable hydrological conditions of the Arkansas River basin in Colorado. These examples emphasize the need to consider the inherent characteristics of the regional fauna when developing multimetric biological assessment tools. Because surface water resources naturally vary across the nation, nationally
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--> uniform biological criteria are neither feasible nor desirable. However, it is desirable to have a national framework based on the concepts of the regional reference site approach (Hughes et al., 1986), which will promote national consistency between state bioassessments. The key is to describe the framework within a common national goal, such as the maintenance and restoration of biological integrity. Initial Considerations A number of fundamental decisions need to be made prior to adopting a set of biological monitoring methods. This is a critical juncture in the process because bad decisions will reduce the effectiveness of the efforts well into the future. These vital initial choices include which sampling methods to use, when to sample, which organisms to monitor, which parameters to measure, and which level of taxonomic precision to use. If any axiom applies, it is: When in doubt, take more measurements than seem necessary at the time because uncollected information cannot be retrieved later. Parameters that require little or no extra effort to acquire should be included until the evidence shows they are unnecessary. One example in Ohio is external anomalies on fish. We decided to record this information even though it was not apparent how it would be useful. This measure has proved to be one of our most valuable assessment tools. For macroinvertebrates, the decision to identify midges to the genus and species level was also fortuitous given the value of this group in diagnosing impairments. Of course, samples could always be archived for later reinspection, but the logistical burdens that this entails are undesirable. Another important consideration is to assure that qualified and regionally experienced staff do the fieldwork. In ecological assessment, like many other professions, the most skilled and experienced individuals are sought to direct, manage, and supervise. However, biological field assessment requires an equivalent level of expertise in the field because many of the critical pieces of information are recorded and, to a degree, interpreted in the field. There is simply no substitute for the intangibles gained by direct observations in the field. This is not a job to be left to technicians. The staff who perform the fieldwork should also plan that work, process the data, interpret the results, and write reports. Such staff, particularly the more experienced individuals, also contribute to policy development. Water-Quality Standards: Designated Uses and Criteria The Ohio water-quality standards (Ohio Administrative Code 3745-1) consist of designated habitat classifications. Chemical, physical, and biological criteria are specific to each classification and are designed to be consistent with the goals specified by each classification. Protection and restoration requirements are a function of habitat classification. The Ohio water-quality standards de-
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--> Figure 8 Cumulative frequency diagram for Index of Biotic Integrity scores for 1,160 Ohio sites and Invertebrate Community Index scores for 854 Ohio sites measured before and after 1988. NOTE: Measurements from before 1988 are represented by the line labeled ''Earliest Data." Measurements from after 1988 are represented by the line labeled "Latest Data."
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--> Figure 9 Map of sampling sites on Ohio rivers and streams depicting changes in the Index of Biotic Integrity (IBI) at 1,160 sites and Invertebrate Community Index (ICI) at 527 sites. NOTE: Symbols indicate the directions of changes in the two indicators at the various sites based upon a comparison of measurements from the earliest and latest sampling years before and after 1988. apart). The analysis included 1,160 sites for the Index of Biotic Integrity, 845 sites for the Modified Index of Well-Being, and 527 sites for the Invertebrate Community Index. Significant improvements have been observed for each index (Table 3; Figures 8 and 9). The Invertebrate Community Index showed both the largest increase and the largest shift in the frequency of sites entering the good and exceptional performance ranges (i.e., scores meeting the warm-water habitat and exceptional warm-water habitat criteria), but the fewest sites exiting the poor and very poor ranges (scores <14). In contrast, the fish community Index of Biotic Integrity
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--> had the greatest number of sites exiting the poor and very poor ranges (scores >26-28) but the fewest sites entering the good and exceptional ranges. Considerable improvements have been documented in Ohio rivers and streams, as reflected by the biological criteria. The predominant pattern begins with the recovery of the macroinvertebrate community (as measured by the Invertebrate Community Index), followed later by improvements in fish abundance and biomass (as indicated by the Modified Index of Well-Being), and then finally structural and functional improvements (as measured by the Index of Biotic Integrity).
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--> TABLE 3 Summary of Results for Ohio Stream and River Sites with at Least 2 Years of Biological Data Collected between 1979 and 1994, at Least Once Before (earliest) and Once After (latest) 1988 Index of Biotic Integrity Modified Index of Well-Being Invertebrate Community Index Category Earliest Latest Earliest Latest Earliest Latest 10th percentile 16 20 3.8 4.8 6 14 25th percentile 24 28 5.7 6.6 16 26 Median 32 36 7.4 8.1 32 38 75th percentile 42 45 8.6 9.2 42 46 90th percentile 48 50 9.3 9.9 46 52 Mean 32.2 36.2 6.91 7.72 28.9 35.5 Degrees of freedom 1159 844 527 t value 16.89 14.34 12.53 Mean difference 3.95 0.80 6.65 t-test P value P <0.0001 P <0.0001 P <0.0001 Wilcoxon (Z) 24.40 13.50 11.55 Wilcoxon text P value P <0.0001 P <0.0001 P <0.0001 NOTE: Data pairs show descriptive statistics for earliest and latest measurements of the Index of Biotic Integrity, the Modified Index of Well-Being, and the Invertebrate Community Index at the various sites. Paired t-test and Wilcoxon's Z-test statistics compare the means of the earliest and latest measurements of each of the indices. SOURCE: Ohio Environmental Protection Agency (1996). These substantial improvements notwithstanding, a significant proportion of Ohio's rivers and streams remain too polluted and/or physically degraded to meet their biological criteria. Figure 10 illustrates the aggregate changes in impairment that took place between 1988 and 1996 and projects the trend through the year 2002. Figure 10 makes it clear that the proportion of impairment associated primarily with point-source discharges is declining at a more rapid rate than impairment associated with nonpoint sources. Nonpoint sources include habitat modifications, nutrient enrichment, and sedimentation. The state will need to pay more attention to nonpoint sources and watershed-level effects if it is to reach milestones such as the Ohio 2000 goal of 75 percent of full attainment. This will require a significant restructuring of state water-quality management programs that are presently heavily oriented toward point sources of pollutants.
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--> Figure 10 Observed and forecasted reductions, by percent, in the proportion of river and stream miles failing to attain their habitat classification biological criteria. NOTE: Forecasts from 1997 to 2002 are based on the observed rate of restoration during the period 1988-1996. The dashed horizontal line represents the Ohio 2000 goal. Overall Impact of Biological Criteria The use of biological criteria has proved useful to the Ohio Environmental Protection Agency for several reasons: The results use direct measurements of ecological condition, rather than surrogate or symptomatic measures. This helps focus management programs on actual environmental results rather than administrative goals only (i.e., number of permits issued, grant dollars awarded, etc.). The resulting knowledge of aquatic community conditions can more efficiently guide management and regulatory activities that might otherwise be forced to rely on prescriptive approaches to information gathering (e.g., effluent characterization, major discharger lists, 303[d] and 304 lists). The results provide objective measurements with which to assign appropriate habitat classifications to individual rivers and streams. The results provide a means to assess the applicability and effectiveness of the antidegradation policy in the Ohio water-quality standards (i.e., extending antidegradation concerns to nonpoint sources and habitat influences, defining high-quality waters).
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--> Chemical and narrative water-quality criteria and standards can be more appropriately applied and take into account the integrated dynamics of the receiving waters when relevant biological assessment information is available (Yoder, 1991a,b). The biological results provide a legal basis for enforcement against entities discharging chemicals for which there are no existing water-quality standards or effluent guidelines (or at least provide the impetus to designate new chemical criteria or whole-effluent toxicity limits). The results provide a basis for regulating nonchemical environmental degradation (e.g., certifications of dredging permits [Section 401 water-quality certifications], non-point-source management). Remaining Challenges We have demonstrated how biological criteria can be developed and used within a state water-resource-management framework. Nonetheless, some important challenges remain. The cumulative costs associated with environmental mandates, many of which consist of prescription-based regulations, have recently come into question. Both the regulated community and the public desire evidence of real-world results in return for the expenditures made necessary by federally and state-mandated requirements. Biological criteria seem particularly well suited to address these concerns because the underlying science and theory is robust (Karr, 1991), and biocriteria directly assess the biological condition of aquatic habitats. Although no single environmental indicator can do it all, biological criteria have a major role to play. A lack of information from or an overreliance on any single class of indicators can result in environmental regulation that is inaccurate and either under- or overprotective of the resource. Accounting for cost is not only a matter of dollars spent but also of program effectiveness. A credible and genuinely cost-effective approach to water-quality management should include an appropriate mix of chemical, physical, and biological measures, each in their respective roles as stressor, exposure, and response indicators. The public must come to see comprehensive monitoring designs using such cost-effective indicators as a part of the cost of doing business, perhaps at the expense of other programs when new evidence suggests that the resources allocated are disproportionate to the magnitude of the present problems (e.g., point versus nonpoint sources). Based on our experience over the past 17 years, it is evident that including biological criteria in a state's monitoring and assessment effort has multiple benefits: It can foster a more complete integration of important ecological concepts, better focus water resource policy and management, and enhance strategic planning. Some specific examples include:
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--> Watershed approaches to monitoring, assessment, and management—The monitoring and assessment design inherent to biological criteria is fundamentally oriented to yield information on watersheds. Integrated point, nonpoint, and habitat assessment and management—Biological criteria integrate the effects of all stressors over time and space. The attendant use of chemical, toxicological, and physical tools can connect probable causes to observed impairments. This should provide a firm setting for the collaborative use of the same information for the management and regulation of both point and nonpoint sources (including habitat), two realms that have thus far been treated independently. Cumulative effects—Biological communities inhabit the receiving waters all of the time and reflect the integrative, cumulative effect of various stressors. Biodiversity issues—The basic biological data provide information about species, populations, and communities of concern. Interdisciplinary focus—Because biosurvey monitoring and assessment design is inherently interdisciplinary, the biological criteria approach provides the opportunity to bring ecologists, toxicologists, engineers, and other professionals together in planning and conducting assessments, interpreting results, and using information in strategic planning and management actions. Biological criteria are an emerging and increasingly important issue for the EPA, the states, and the regulated community, and their use is growing nationwide. However, much remains to be done, particularly in the area of national and regional leadership. Technical guidance and expertise is needed to ensure a nationally consistent and credible approach and to resolve outstanding technical concerns (Yoder and Rankin, 1995a). Outstanding policy issues, such as the EPA's policy of independent applicability, need to be resolved in a manner that will encourage states to participate. In an era of declining government resources, we must develop ways to do more biological monitoring to support the biocriteria approach. Based on our experience in Ohio, the staffing of state biological assessment programs should include a minimum of one work year equivalent for every 1,200 miles of perennial streams and rivers. This estimate may vary by region and should incorporate lake-surface area in states with a substantial number of lakes (Yoder and Rankin, 1995a). The EPA must consider the potential for bioassessments and biocriteria to modify the present capital- and resource-intensive system of tracking environmental compliance on the basis of specific pollutants. The biological criteria approach should prove a more cost-effective way of managing the nation's water-quality programs. Acknowledgments This paper would not have been possible without the many years of fieldwork, laboratory analysis, and data assessment and interpretation performed by
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--> members (past and present) of the Ohio Environmental Protection Agency, Ecological Assessment Section. Several staff members contributed extensively to the development of the biological assessment program, including Dave Altfater, Randy Sanders, Marc Smith, and Roger Thoma (fish methods, MIwb, and IBI metrics development) and Mike Bolton, Jeff DeShon, Jack Freda, Marty Knapp, and Chuck McKnight (macroinvertebrate methods and ICI metrics development). None of this effort would have been possible without the excellent data management and processing skills contributed by Dennis Mishne. Other staff (past and present) who made important contributions include Paul Albeit, Ray Beaumier, Chuck Boucher, Bernie Counts, Beth Lenoble, and Paul Vandermeer. Dan Dudley and Jim Luey contributed extensively to the early development and review of the then-emerging concepts of biological integrity, ecoregions, reference sites, and biological assessment in general. Charlie Staudt provided many hours of support in the development of the computer programs used for data analysis. Finally, Gary Martin and the late Pat Abrams provided solid management support for the concept of biological criteria and biological assessment at the Ohio Environmental Protection Agency. References Angermier, P. L., and J. R. Karr. 1986. Applying an index of biotic integrity based on stream-fish communities: Considerations in sampling and interpretation. North American Journal of Fisheries Management 6:418-427. Angermier, P. L., and J. E. Williams. 1993. Conservation of imperiled species and reauthorization of the endangered species act of 1973. North American Journal of Fisheries Management 18(7):34-38. Ballentine, R. K., and L. J. Guarria. 1975. Paper published in The Integrity of Water: Proceedings of a Symposium, March 10-12, 1975, Washington, D.C. U.S. Environmental Protection Agency, Office of Water and Hazardous Materials. Washington, D.C.: U.S. Government Printing Office. Barbour, M. T., J. Gerritsen, G. E. Griffin, R. Frydenborg, E. McCarron, J. S. White, and M. L. Bastian. 1996. A framework for biological criteria for Florida streams using benthic macroinvertebrates. Journal of the North American Benthological Society 15(2): 185-211. Benke, A. C. 1990. A perspective on America's vanishing streams. Journal of the North American Benthological Society 9(1):77-88. Bennet, M. R., J. W. Kleene, and V. O. Shanholtz. 1993. Total maximum daily load nonpoint source allocation pilot project. File Report. Blacksburg, Va.: Virginia Department of Agricultural Engineering. Bramblett, R. G., and K. D. Fausch. 1991. Variable fish communities and the index of biotic integrity in a western Great Plains river. Transactions of the American Fisheries Society 120:752. Davies, S. P., L. Tsomides, D. L. Courtemanch, and F. Drummond. 1993. State of Maine Biological Monitoring and Biocriteria Development Program Summary. Augusta: Maine Department of Environmental Protection. Davis, W. S., and A. Lubin. 1989. Statistical validation of Ohio EPA's invertebrate community index. Pp. 23-32 in Proceedings of the 1989 Midwest Pollution Biology Meeting, Chicago, Ill., W. S. Davis and T. P. Simon, eds. EPA 905/9-89/007. Washington, D.C.: U.S. Environmental Protection Agency.
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Representative terms from entire chapter: