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In cooperation with the Wisconsin Department of Natural Resources

Nutrient Concentrations and Their Relations to the Biotic Integrity of Wadeable Streams in Wisconsin

Dale M. Robertson¹, David J. Graczyk¹, Paul J. Garrison², Lizhu Wang³, Gina LaLiberte², and Roger Bannerman⁴

¹ U.S. Geological Survey, Wisconsin Water Science Center, Middleton, Wisconsin
² Wisconsin Department of Natural Resources, Monona, Wisconsin
³ Michigan Department of Natural Resources, Ann Arbor, Michigan
⁴ Wisconsin Department of Natural Resources, Madison, Wisconsin

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Table of Contents

Abstract

Purpose and Scope
Approach

Methods of Data Collection and Analysis

Field Methods
Watershed Boundaries and Environmental Characteristics
Data Summaries
Statistical Methods

Water Quality and Its Relations with Environmental Characteristics in the Watershed

Relations between Water Quality and Environmental Characteristics in the Watershed
Regionalization Schemes for Reference Water Quality and the Response in Water Quality to Changes in Land Use

Physical Habitat

Benthic Chlorophyll a and Periphytic-Diatom Communities and Their Relations with Water-Quality, Environmental, and Physical-Habitat Characteristics

Relations with Individual Characteristics
Effects of Multiple Characteristics on Benthic Chlorophyll a Concentrations and Diatom Indices
Reference Values for Benthic Chlorophyll a Concentrations and Diatom Indices

Relations with Individual Characteristics
Effects of Multiple Characteristics on Macroinvertebrate Indices
Reference Values for the Macroinvertebrate Indices

Relations with Individual Characteristics
Effects of Multiple Characteristics on Fish Indices
Reference Values for the Fish Indices

Multiparameter Biotic Indices to Estimate Nutrient Concentrations in Wadeable Streams

Summary and Conclusions

Final Regionalization Scheme for Wisconsin Streams
Reference Conditions
Responses of Water Quality to Changes in Land Use
Responses of Biotic Indices to Changes in Nutrient Concentrations
Multiparameter Biotic Indices to Estimate Nutrient Concentrations in Wadeable Streams
Nutrient Concentrations Controlling the Biotic Integrity of Streams

Literature Cited

Appendixes

Appendix 1. Stream identification information, location information, and summary statistics for flow and water-quality data collected for each of the 240 studied wadeable streams in Wisconsin.
Appendix 2. Physical-habitat characteristics of each of the 240 studied wadeable streams in Wisconsin.
Appendix 3. Diatom nutrient-tolerance ranking for individual diatom taxa.
Appendix 4. Biological data for each of the 240 studied wadeable streams in Wisconsin.

Figures

Figure 1. National nutrient ecoregions and major land uses in the upper Midwest.

Figure 2. Two regionalization schemes considered for wadeable streams in Wisconsin: A, level III ecoregions with major land-use/land-cover categories and B, environmental phosphorus zones.

Figure 3. Environmental phosphorus zones in the upper Midwest.

Figure 4. Sites on wadeable streams in Wisconsin included in this study.

Figure 5. Distributions of median monthly total phosphorus, total nitrogen, and suspended chlorophyll a concentrations, and Secchi tube depth.

Figure 6. Total phosphorus and total nitrogen concentrations as a function of the percentage of total agriculture in the watersheds of the studied wadeable streams in Wisconsin.

Figure 7. Suspended chlorophyll a concentrations and Secchi tube depths as a function of median total phosphorus and total nitrogen concentrations.

Figure 8. Percentages of explained variance in water quality described by anthropogenic/land-use, basin, soil and surficial-deposit characteristics, and interactions among categories.

Figure 9. Percentages of explained variance in A, Secchi tube depths and B, suspended chlorophyll a concentrations described by nutrients, environmental characteristics, and interactions among categories.

Figure 10. Percentiles of A, total agriculture in the watersheds, B, total phosphorus, and C, total nitrogen in streams in the level III ecoregions and environmental phosphorus zones.

Figure 11. A, suspended chlorophyll a concentrations and B, Secchi tube depths in Reference sites, High sites, and sites with only reference total nitrogen or reference total phosphorus concentrations.

Figure 12. Total phosphorus concentrations as a function of the percentage of agricultural land use in A, environmental phosphorus zones and B, level III ecoregions, and response curves for phosphorus concentrations as a function of the percentage of agriculture in the watershed, in C, EPZs, and D, level III ecoregions.

Figure 13. Total nitrogen concentrations as a function of the percentage of agricultural land use in A, environmental phosphorus zones and B, level III ecoregions, and response curves for nitrogen concentrations as a function of the percentage of agriculture in the watershed, in C, EPZs, and D, level III ecoregions.

Figure 14. Suspended chlorophyll a concentrations as a function of total phosphorus and total nitrogen concentration, by environmental phosphorus zones and by level III ecoregions.

Figure 15. Secchi tube depth as a function of total phosphorus and total nitrogen concentration, by environmental phosphorus zones and by level III ecoregions.

Figure 16. Response curves for A, suspended chlorophyll a concentrations and B, Secchi tube depths in the environmental phosphorus zones as a function of the percentage of agriculture in the watershed.

Figure 17. Distributions of benthic chlorophyll a concentrations, Diatom Nutrient Index values, Diatom Siltation Index values, and Diatom Biotic Index values.

Figure 18. Benthic chlorophyll a concentrations, Diatom Nutrient Index values, and Diatom Siltation Index values as a function of total phosphorus and total nitrogen concentration in the environmental phosphorus zones and level III ecoregions.

Figure 19. Benthic chlorophyll a concentrations, Diatom Nutrient Index values, and Diatom Siltation Index values as a function of dissolved phosphorus, dissolved nitrite plus nitrate, dissolved ammonia, and total Kjeldahl nitrogen concentrations in the environmental phosphorus zones.

Figure 20. Benthic chlorophyll a concentrations and Diatom Nutrient Index values for five nutrient categories.

Figure 21. Percentage of explained variance in benthic chlorophyll a concentrations and diatom index values described by nutrients, other water-quality characteristics, environmental and physical-habitat characteristics, and interactions among categories.

Figure 22. Benthic chlorophyll a concentrations, Diatom Nutrient Index, Diatom Siltation Index, and Diatom Biotic Index values in Reference sites, High sites, and sites with only reference total nitrogen or reference total phosphorus concentrations.

Figure 23. Distributions of macroinvertebrate Hilsenhoff Biotic Index values, the percentages of individuals that were Ephemeroptera, Plecoptera, or Trichoptera, and the percentages of taxa that were Ephemeroptera, Plecoptera, or Trichoptera.

Figure 24. Hilsenhoff Biotic Index values, the percentages of individuals that were Ephemeroptera, Plecoptera, or Trichoptera, and the percentages of taxa that were Ephemeroptera, Plecoptera, or Trichoptera as a function of total phosphorus and total nitrogen concentration in the environmental phosphorus zones and level III ecoregions.

Figure 25. Hilsenhoff Biotic Index values, the percentages of individuals that were Ephemeroptera, Plecoptera, or Trichoptera, and the percentages of taxa that were Ephemeroptera, Plecoptera, or Trichoptera as a function of dissolved phosphorus, dissolved nitrite plus nitrate, dissolved ammonia, and total Kjeldahl nitrogen concentrations in the environmental phosphorus zones.

Figure 26. Percentages of explained variance in six macroinvertebrate index values described by nutrients, other water-quality characteristics, environmental and physical-habitat characteristics, and interactions among categories.

Figure 27. Hilsenhoff Biotic Index values, the percentages of individuals that were Ephemeroptera, Plecoptera, or Trichoptera, and the percentages of taxa that were Ephemeroptera, Plecoptera, or Trichoptera in Reference sites, High sites, and sites with only reference total nitrogen or reference total phosphorus concentrations.

Figure 28. Distributions of fish Index of Biotic Integrity values, the percentages of the fish that are carnivorous, the percentages of fish considered pollution intolerant, and the number of fish caught.

Figure 29. Fish Index of Biotic Integrity values, the percentages of fish considered pollution intolerant, and the percentages of the fish that are carnivorous as a function of total phosphorus and total nitrogen concentration in the environmental phosphorus zones and level III ecoregions.

Figure 30. Fish Index of Biotic Integrity values, the percentages of fish considered pollution intolerant, and the percentages of the fish that are carnivorous as a function of dissolved phosphorus, dissolved nitrite plus nitrate, dissolved ammonia, and total Kjeldahl nitrogen concentrations in the environmental phosphorus zones.

Figure 31. Percentages of explained variance in eight fish index values described by nutrients, other water-quality characteristics, environmental and physical-habitat characteristics, and interactions among categories.

Figure 32. Fish Index of Biotic Integrity values, the percentages of the fish that are carnivorous, and the percentages of fish that are considered pollution intolerant in Reference sites, High sites, and sites with only reference total nitrogen or reference total phosphorus concentrations.

Figure 33. Measured and estimated A, total phosphorus and B, total nitrogen concentrations for the four-parameter regression models, C, measured phosphorus concentrations as a function of Biotic Index of total Phosphorus values, and D, measured nitrogen concentrations as a function of Biotic Index of total Nitrogen values.

Figure 34. Proposed regionalization scheme for defining nutrient criteria for wadeable streams in Wisconsin.

Tables

Table 1. Reference concentrations for total phosphorus, total nitrogen, and suspended chlorophyll a, and turbidity in selected national nutrient and level III ecoregions and environmental phosphorus zones.

Table 2. Summary statistics for median monthly water-quality and environmental characteristics of the watersheds of the sites in the studied wadeable streams in Wisconsin.

Table 3. Median and average monthly concentrations for total and dissolved phosphorus, suspended chlorophyll a, total nitrogen, nitrite plus nitrate, ammonia, and Kjeldahl nitrogen, and Secchi tube depths.

Table 4. Spearman correlation coefficients between median concentrations of total phosphorus, dissolved phosphorus, total nitrogen, nitrite plus nitrate, ammonia, total Kjeldahl nitrogen, median Secchi tube depths, suspended chlorophyll a concentrations, percentages of urban and agricultural areas, point-source loadings of phosphorus, and specific environmental characteristics.

Table 5. Spearman correlation coefficients between residualized logarithmically transformed median concentrations of total phosphorus, dissolved phosphorus, nitrogen, nitrite plus nitrate, total Kjeldahl nitrogen, ammonia and suspended chlorophyll a, median Secchi tube depths, and specific residualized environmental characteristics.

Table 6. Results of forward stepwise-regression analyses used to explain the variance in raw and residualized water-quality concentrations.

Table 7. Results from redundancy analysis between water-quality and environmental characteristics.

Table 8. Reference conditions for total phosphorus, total nitrogen, and suspended chlorophyll a concentrations, and Secchi tube depths in the environmental phosphorus zones and level III ecoregions.

Table 9. Responses in total phosphorus total nitrogen, suspended chlorophyll a, and Secchi tube depth to changes in the percentage of agricultural land use in the watershed.

Table 10. Summary statistics for the physical-habitat characteristics and biotic indices.

Table 11. Spearman rank correlation coefficients between benthic chlorophyll a concentrations and diatom community indices, and median water-quality, environmental, and physical-habitat characteristics.

Table 12. Thresholds or breakpoints in the responses of benthic chlorophyll a concentrations and diatom indices to changes in nutrient concentrations.

Table 13. Percentiles of total phosphorus and total nitrogen concentrations for the studied streams having both benthic chlorophyll a and diatom index data.

Table 14. Results of forward stepwise-regression analysis to explain the variance in benthic chlorophyll a concentrations and the three diatom indices.

Table 15. Spearman rank correlation coefficients between macroinvertebrate-community indices and median water-quality, environmental, and physical-habitat characteristics.

Table 16. Thresholds or breakpoints in the responses in macroinvertebrate indices to changes in nutrient concentrations.

Table 17. Results of forward stepwise-regression analyses to explain variance in macroinvertebrate indices.

Table 18. Spearman rank correlation coefficients (r_s) between fish-community indices and median water-quality, environmental, and physical-habitat characteristics.

Table 19. Thresholds or breakpoints in the responses in fish indices to changes in nutrient concentrations.

Table 20. Results of forward stepwise-regression analyses to explain variance in fish indices.

Table 21. Results of forward stepwise-regression analyses to explain variance in total phosphorus and total nitrogen concentrations with biotic indices.

Table 22. Reference conditions for water quality, chlorophyll a, diatoms, macroinvertebrates, and fish indices.

Table 23. Summary of thresholds or breakpoints in the responses of suspended chlorophyll a concentrations, Secchi tube depth, and various biotic indices to changes in nutrient concentrations.

Conversion Factors

Temperature in degrees Celsius (°C) may be converted to degrees Fahrenheit (°F) as follows:

°F=(1.8x°C)+32

Specific conductance is given in microsiemens per centimeter at 25 degrees Celsius (µS/cm at 25°C).

Concentrations of chemical constituents in water are given either in milligrams per liter (mg/L) or micrograms per liter (µg/L).

Abbreviations

Acknowledgments

Technical Reviewers

Editorial and Graphics

Approving Official

Abstract

Excessive nutrient (phosphorus and nitrogen) loss from watersheds is frequently associated with degraded water quality in streams. To reduce this loss, agricultural performance standards and regulations for croplands and livestock operations are being proposed by various States. In addition, the U.S. Environmental Protection Agency is establishing regionally based nutrient criteria that can be refined by each State to determine whether actions are needed to improve a stream's water quality. More confidence in the environmental benefits of the proposed performance standards and nutrient criteria will be possible with a better understanding of the biotic responses to a range of nutrient concentrations in different environmental settings.

The U.S. Geological Survey and the Wisconsin Department of Natural Resources collected data from 240 wadeable streams throughout Wisconsin to: 1) describe how nutrient concentrations and biotic-community structure vary throughout the State; 2) determine which environmental characteristics are most strongly related to the distribution of nutrient concentrations; 3) determine reference water-quality and biotic conditions for different areas of the State; 4) determine how the biotic community of streams in different areas of the State respond to changes in nutrient concentrations; 5) determine the best regionalization scheme to describe the patterns in reference conditions and the responses in water quality and the biotic community; and 6) develop new indices to estimate nutrient concentrations in streams from a combination of biotic indices. The ultimate goal of this study is to provide the information needed to guide the development of regionally based nutrient criteria for Wisconsin streams.

For total nitrogen (N) and suspended chlorophyll (SCHL) concentrations and water clarity, regional variability in reference conditions and in the responses in water quality to changes in land use are best described by subdividing wadeable streams into two categories: streams in areas with high clay-content soils (Environmental Phosphorus Zone 3, EPZ 3) and streams throughout the rest of the State. The regional variability in the response in total phosphorus (P) concentrations is also best described by subdividing the streams into these two categories; however, little consistent variability was found in reference P concentrations in streams throughout the State.

Reference P concentrations are smilar throughout the State (0.03–0.04 mg/L). Reference N concentrations are divided into two categories: 0.6–0.7 mg/L in all streams except those in areas with high clay-content soils, where 0.4 mg/L is more appropriate. Reference SCHL concentrations are divided into two categories: 1.2–1.7 µg/L in all streams except those in areas with high clay-content soils, where 1.0 µg/L may be more appropriate. Reference water clarity is divided into two categories: streams in areas with high clay-content soils with a lower reference water clarity (Secchi tube depth, SD, of about 110 cm) and streams throughout the rest of the State (SD greater than or equal to about 115 cm). For each category of the biotic community (SCHL and benthic chlorophyll a concentrations (BCHL), periphytic diatoms, macroinvertebrates, and fish), a few biotic indices were more related to differences in nutrient concentrations than were others. For each of the indices more strongly related to nutrient concentrations, reference conditions were obtained by determining values corresponding to the worst 75th percentile value from a subset of minimally impacted streams (streams having reference nutrient concentrations).

By examining the biotic community in streams having either reference P or N concentrations but not both, the relative importance of these two nutrients was determined. For SCHL, P was the more important limiting nutrient; however, for BCHL and all macroinvertebrate indices, it appears that N was the more important nutrient when concentrations were near reference concentrations. For other diatom indices and all fish indices, small additions of P or N appear to have little effect on these communities when nutrient concentrations are near reference conditions.

Concentrations of P and N in streams increase as the percentage of agricultural land increases. Concentrations of P increase more quickly and concentrations of N increase more slowly in response to increasing percentages of agriculture in areas with high clay-content soils than do streams in the rest of the State. The response in water clarity is similar in streams throughout the State; however, the streams in areas with high clay-content soils have poorer reference water clarity, and, therefore, as the percentage of agriculture increases, their clarity remains lower than in streams in areas with other soil types.

As nutrient concentrations increase, many biotic indices change. This result indicates that these nutrients have direct or indirect effects on the composition of the biotic community. Thresholds were identified at which a small change in nutrient concentrations results in a relatively large change in the biotic communities. The thresholds in the response to changes in P concentrations range from about 0.04 mg/L for BCHL, to 0.06–0.07 mg/L for diatom and fish indices, to about 0.09 mg/L for macroinvertebrate indices. The thresholds in the response to changes in N concentrations range from 0.5 mg/L for the fish indices and one macroinvertebrate index to about 0.9–1.2 mg/L for the diatom and other macroinvertebrate indices. Most of the biotic indices had a wedge-shaped response to increases in nutrient concentrations. At relatively low nutrient concentrations, the biotic indices ranged widely, but at relatively high concentrations, the indices generally were poor. The wedge-shaped distribution indicates that at low nutrient concentrations, factors other than nutrients often limit the health of biotic communities, whereas, at high nutrient concentrations, nutrients and factors correlated with high nutrient concentrations are the predominant factors.

The biotic communities that are present in a stream reflect the overall ecological integrity; therefore, they integrate the effects of many different stressors and thus provide a broad measure of their aggregate effect. Nutrient concentrations by themselves explained from about 6 to 13 percent of the total variance in the components of the biotic communities or from about 14 to 23 percent of the explained variance. Nutrient concentrations were most important in affecting SCHL concentrations and macroinvertebrate communities, and least important in affecting BCHL, periphytic diatoms, and fish-community structure. For each component of the biotic community, nutrients by themselves only explained a small part of the overall variance; about half of the variance could not be explained by the variables examined in this study and about one-third of the explained variance could not be assigned to single categories of environmental characteristics.

By use of a combination of four biotic indices, two new multiparameter indices (Biotic Index of total Phosphorus, BIP, and Biotic Index of total Nitrogen, BIN) were developed to estimate P and N concentrations in streams from biotic data collected in streams. These multiparameter models estimated high and low nutrient concentrations equally well. The BIP predicted P concentrations better than the BIN predicted N concentrations. The difference in the accuracy of these indices was consistent with biotic indices being more correlated with P concentrations than with N concentrations. This result suggests that P is more important than N in affecting most biotic communities as nutrient concentrations increase above reference concentrations.

Although specific mechanisms of how nutrients affect the biota in wadeable streams were not examined in this study, the results indicate that nutrients are important in controlling the biotic health of streams. Although the biotic-community structure represents the overall ecological integrity of the stream, nutrients alone explained only a small part of the variance in the biotic community. Therefore, it is difficult to predict the exact result of reducing nutrient concentrations without also modifying the factors typically associated with high nutrient concentrations. Nutrient concentrations in many streams, especially those in agricultural areas, are well above the concentrations where thresholds in the response were found to occur; therefore, small reductions in nutrient concentrations in these streams are not expected to have large effects on the biotic community. Even with these limitations, however, it is expected that reducing nutrient concentrations will improve the biotic community, further the beneficial ecological functioning of most streams, and improve the quality of downstream nutrient-limited receiving waters.

Introduction

Elevated nutrient concentrations above background conditions are one of the most common stressors (contaminants) affecting streams throughout the United States. Problems associated with elevated nutrient concentrations in surface water are not new, but they are among the most persistent. According to the National Water Quality Inventory: 1996 Report to Congress by the U.S. Environmental Protection Agency (USEPA), 50 States, Tribes, and other jurisdictions surveyed water-quality conditions in 19 percent of the Nation's 3.6 million miles of rivers and streams and found overenrichment of nutrients to be the second most common reason for impairment following the combined effects of suspended sediment and siltation (U.S. Environmental Protection Agency, 1996). Excessive nutrients in rivers and streams can result in the overgrowth of benthic algae in shallow areas and in areas with fast current and an overabundance of phytoplankton and macrophytes in deep areas with slow current. High algal and macrophyte biomass can cause severe diurnal fluctuations in dissolved oxygen and pH because of biotic production and respiration, and can generate harmful organic materials when part of the population dies (Welch and others, 1992). These conditions can lead to an increase in the availability of toxic substances, reduction in available aquatic habitat, modifications to the composition of the biotic communities, and a decrease in the overall usefulness of the stream (Miltner and Rankin, 1998; Dodds and Welch, 2000). Excessive transport of nutrients has also been linked to eutrophication of downstream lakes and impoundments, outbreaks of Pfiesteria in bays and estuaries in various Gulf and Mid-Atlantic States, and hypoxia in the Gulf of Mexico (U.S. Environmental Protection Agency, 2000a).

Under recommendations of the Clean Water Action Plan released in 1998, the USEPA has developed a National strategy to develop waterbody-specific nutrient criteria for lakes and reservoirs, rivers and streams, wetlands, and estuaries (U.S. Environmental Protection Agency, 1998); this study is concerned with those for rivers and streams. The intent of this strategy is to get all States and tribes to establish nutrient standards, that, if enforced, will reduce nutrient concentrations and improve the beneficial ecological uses of surface waters. The best way to control nutrient concentrations is to reduce that part contributed by humans, not that part contributed naturally. It has been recognized that various environmental characteristics, such as land use, geology, soils, climate, and hydrology (including human modifications and hydrologic structures) are important in determining water quality (Monteith and Sonzogni, 1981; Clesceri and others, 1986; and Robertson, 1997). Because these characteristics vary greatly across the United States, the establishment of regional nutrient criteria makes scientific sense.

Various frameworks have been used to divide large areas into smaller areas of relatively similar environmental characteristics to minimize the natural variation in water quality within the areas and maximize the differences among the areas. One such framework is the ecoregion delineation developed and refined by Omernik (1987; 1995) and Omernik and others (2000). Ecoregions are a mapped-classification system of regions with assumed relative homogeneity in ecological characteristics. These regions were said to be defined on the basis of relative differences in a suite of environmental characteristics, such as land use/land cover, land-surface form, geology, physiography, climate, soils, potential natural vegetation, and other environmental characteristics (Omernik, 1987 and 1995). The USEPA has taken the initial step in developing regional nutrient criteria based on combining Omernik's 84 level III ecoregions into 14 national nutrient ecoregions for the conterminous United States (U.S. Environmental Protection Agency, 1998; fig. 1). On a subregional basis, such as a specific State, each of these 14 nutrient ecoregions can be further subdivided into the original level III ecoregions. Wisconsin is subdivided into two national nutrient ecoregions (ecoregions 7 and 8; fig. 1), which are further subdivided into four primary level III ecoregions: Northern Lakes and Forests (NLF), North Central Hardwood Forests (NCHF), Southeastern Wisconsin Till Plains (SWTP), and the Driftless Area (DFA)(Omernik and others, 2000; fig. 2A). In addition, there are small pieces of the Western Cornbelt Plains and the Central Cornbelt Plains ecoregions. Because the Cornbelt Plains ecoregions represent only a small part of the State, they will not be discussed in this report. The nutrient ecoregions provide an initial classification scheme for developing nutrient criteria; however, the USEPA expects individual States and tribes to evaluate and possibly develop alternative regionalization schemes (U.S. Environmental Protection Agency, 2000b).

Figure 1. National nutrient ecoregions (U.S. Environmental Protection Agency, 1998) and major land uses in the upper Midwest (U.S. Geological Survey, 2000 and 2004).

Figure 2. Two regionalization schemes considered for wadeable streams in Wisconsin: A, level III ecoregions (Omernik and others, 2000) with major land-use/land-cover categories (Lillesand and others, 1998) and B, environmental phosphorus zones (Robertson and others, 2006).

The nutrient ecoregions proposed by the U.S. Environmental Protection Agency (1998) may define spatial patterns in water quality; however, this regionalization scheme has some inherent problems. Although the boundaries between ecoregions are supposed to represent the differences in a suite of related environmental characteristics (Omernik, 1995), specific boundary lines are often based on differences in a single environmental characteristic and that characteristic may not be the primary one affecting a specific water-quality characteristic. Therefore, greater variations in water quality may occur within an ecoregion than among ecoregions. Second, in defining most ecoregions, the relative importance of each environmental characteristic is often unknown and can vary from one area of the country to another in an unknown way. Therefore, the differences in water quality among ecoregions can be difficult to attribute to any specific environmental characteristic. Finally, for many applications, such as establishing reference conditions for nutrient criteria, the environmental characteristics used to delineate regions of similar water quality should be restricted, as much as possible, to characteristics that are intrinsic, or natural, and are not the result of human activities (U.S. Environmental Protection Agency, 2000a). Here, "reference" water quality refers to background concentrations or the potential water quality that could be achieved in the absence of human activities. By comparing the lines delineating the ecoregions with the land-use patterns in figure 1, it is apparent, however, that land use was the most important characteristic in defining the ecoregions in the upper Midwest. Although the ecoregion delineation is supposed to represent differences in a full suite of environmental characteristics, nutrient ecoregions primarily subdivide the upper Midwest (and Wisconsin) into areas of forest, cropland and pasture, and cropland.

To overcome the problems just described with the USEPA's nutrient ecoregions, SPARTA (SPAtial Regression-Tree Analysis) was developed to define characteristic-specific zones with similar reference water quality for the upper Midwest (Robertson and Saad, 2003) and refined to remove both the direct and indirect effects of land use (Robertson and others, 2006). There are two steps in the SPARTA process to delineate water-quality zones. The first step is to use regression-tree analysis (Breiman and others, 1984) to describe the relations between a single dependent variable (for example, phosphorus (P) concentrations) and various independent variables (for example, clay content of the soil and basin slope) thought to affect the distribution of the dependent variable. The second step of SPARTA is to use the regression-tree results (specific characteristics and breakpoints) to divide the entire study area into zones representing each of the final branches of the analysis.

To refine the SPARTA approach, land-use-adjusted (residualized) water-quality and environmental characteristics were computed and used to remove the direct and indirect effects of land use from the data for each site. Although SPARTA can be applied with only intrinsic or natural characteristics to remove the direct effects of land use as done by Robertson and Saad (2003), these natural characteristics themselves may be strongly correlated with land use. Thus, even if natural characteristics are the only factors included in the analysis, land use can be indirectly incorporated into the results. SPARTA was applied to the land-use-adjusted nutrient data and land-use-adjusted environmental-characteristic data for sites throughout the upper Midwest to develop environmental water-quality zones (Robertson and others, 2006). Because the biota in streams were assumed to be more influenced by P than by total nitrogen (N) concentrations, environmental P zones (EPZs) were examined as an alternative regionalization scheme for the wadeable streams of Wisconsin (fig. 3). The upper Midwest is divided into five EPZs based primarily on the clay content of the soil and secondarily on the slope of the terrain. Wisconsin consists of four major EPZs and one minor EPZ (fig. 2B). Streams in EPZ 1 have basins with the lowest clay content, streams in EPZ 2 have soils with moderate clay content and low-gradient terrain, streams in EPZ 3 have basins with high clay content and low-gradient terrain, and streams in EPZ 4 have basins with moderate clay content and steep terrain. Only a small part of EPZ 5 (basins with high clay content, steep terrain, and high soil erodibility) is in Wisconsin and, therefore, is not examined in detail in this study. Each of these zones contain streams with relatively similar reference P concentrations and with P concentrations that should respond similarly to changes in land use.

Figure 3. Environmental phosphorus zones (EPZs) in the upper Midwest from Robertson and others (2006).

After relatively homogenous geographic areas are chosen, several approaches have been used to define quantitative nutrient criteria. The approach suggested by the USEPA to define possible criteria is based on the reference or potential water quality of each area. In other words, the criteria should be based on the conditions that are attainable in the geographic location of each stream (U.S. Environmental Protection Agency, 2000a). Reference concentrations for P, N, suspended chlorophyll a (SCHL, also referred to as sestonic chlorophyll), and turbidity have been defined from the frequency distribution of all available data (from USEPA's Storage and Retrieval, STORET, database) for each area. It has been suggested that the lower 25th percentile of all concentration data for an area may represent this reference condition (U.S. Environmental Protection Agency, 2000b). In other words, 25 percent of all the sites have water quality at least as good as this reference condition. It has also been suggested that the upper (highest or worst) 75th percentile of the concentration data for a subset of streams thought to be minimally impacted for a defined area may represent this reference condition. In other words, 75 percent of the minimally impacted sites have water quality at least as good as this reference condition. The final criterion should be between these two concentrations. Another approach to estimate reference concentrations for each relatively homogeneous area is to develop a multiple linear-regression model that relates water quality to various anthropogenic factors or characteristics such as the percentages of agriculture and urban area in the watershed (Dodds and Oakes, 2004). With this approach, the estimated concentration of a constituent occurring in the absence of human activities (in other words, with 0-percent agricultural and 0-percent urban areas) represents the reference concentration. These relations or equations can also be used to place confidence intervals on the reference concentrations.

An alternative approach to defining the nutrient criteria is based on thresholds in the response between nutrient concentrations and biotic indices such as algal productivity (chlorophyll a concentration), water clarity, or diatom or fish biotic indices (U.S. Environmental Protection Agency, 2000a). The biotic community that is present in a stream, however, reflects more than just the nutrient concentrations that are or were present in the stream. The biotic community represents the overall ecological integrity of the stream (in other words, physical, chemical, and biological integrity), and thus provides a broad measure of the aggregate effect of all stressors (Barbour and others, 1999). Biotic communities are controlled by many physical, chemical, and biological factors, though they may be directly affected by only a subset of variables. Watershed characteristics (such as geomorphology, geochemistry, and land use/land cover) control the physical/chemical habitat of the stream where the biota live (Frissell and others, 1986; Poff, 1997). Nutrients have been shown to directly affect the productivity and species composition of primary producers, such as macrophytes and benthic and suspended algae, and indirectly affect the primary and secondary consumers in controlled nutrient-enrichment experiments (for example, Mundie and others, 1991; Peterson and others, 1993; Perrin and Richardson, 1997); however, only limited studies have shown observational linkages between nutrients and the health of the biotic communities in natural streams. Among the limited studies in natural environments, Miltner and Rankin (1998) reported that macroinvertebrate- and fish-community indices were negatively correlated with N and P concentrations in wadeable streams in Ohio. Zorn (2003) reported that P was one of the important variables for predicting the presence or absence of specific fish species in Michigan streams. Heiskary and Markus (2003) also reported significant negative correlations between macroinvertebrate- and fish-community characteristics and P and N concentrations in nonwadeable rivers in Minnesota.

If relations between nutrient concentrations and biotic integrity are used to define criteria, the final nutrient criteria should be chosen to minimize degradation in the biotic integrity of the streams. In other words, the criteria should be the concentrations that would not result in high algal concentrations or degradation of other biotic indices. One of the difficulties in defining nutrient criteria is determining the chlorophyll a concentration or other biotic index values for which a stream is considered degraded or impaired. The assumption made with this biotic-response approach is that each of the subregions in a regional framework also delineates an area with streams whose biotic indices respond in a similar manner to changes in nutrients. Whichever approach is used, the final criteria must be stringent enough to protect the specific site and cause no adverse effects in downstream waters.

Reference nutrient concentrations, the responses in nutrient concentrations to changes in land use, and the biotic responses to changes in nutrients may differ in streams throughout Wisconsin. There would be more confidence in the potential environmental benefits of enforcing nutrient criteria and standards for the State, if the criteria and standards were based on the most appropriate regionalization scheme (such as nutrient ecoregions or EPZs), and if the criteria and standards were based on the appropriate regionally defined thresholds to biotic response. Defined nutrient criteria and thresholds for responsive biotic indices would enable the use of monitoring data to identify streams affected by excessive nutrients and to direct rehabilitation efforts.

The two regionalization schemes being considered for the establishment of nutrient criteria for Wisconsin, level III ecoregions and EPZs, are shown in figure 2. The USEPA developed the preliminary criteria based on median concentrations of all the data measured at each site rather than mean concentrations, because a median value represents the concentration most frequently occurring in the stream, and a statistical summary based on median values reduces the effects of outliers and values reported as less than their respective detection limits. The USEPA has provided preliminary criteria for P, N, SCHL, and turbidity for the national nutrient ecoregions and most level III ecoregions (table 1). The proposed criteria by the USEPA for P, based on the 25th-percentile approach, are 0.033 mg/L for national nutrient ecoregion 7 and 0.010 mg/L for national nutrient ecoregion 8 (same as the NLF ecoregion). The USEPA has refined the P criteria for level III ecoregions in national nutrient ecoregion 7: 0.029 mg/L for NCHF, 0.070 mg/L for DFA, and 0.080 for SWTP (U.S. Environmental Protection Agency, 2000b and 2001).

Table 1. Reference concentrations for total phosphorus, total nitrogen, and suspended chlorophyll a, and turbidity in selected national nutrient and level III ecoregions (U.S. Environmental Protection Agency, 2000b and 2001) and environmental phosphorus zones (EPZs) from Robertson and others (2006).

[USEPA, U.S. Environmental Protection Agency; NCHF, North Central Hardwood Forests; DFA, Driftless Area; SWTP, Southeastern Wisconsin Till Plains; --, no data; NTU, nephelometric turbidity units; FTU, formazin turbidity units; mg/L, milligram per liter; µg/L, microgram per liter]

^a USEPA level III ecoregion identification numbers

Robertson and others (2006) estimated median reference P concentrations for the two major national ecoregions and four major EPZs in Wisconsin (fig. 2) by use of the multiple linear-regression approach (previously described). They found that median reference P concentrations for the two national nutrient ecoregions were similar, approximately 0.015–0.016 mg/L. They also found that the four major EPZs could be combined into two zones based on estimated reference P concentrations of 0.012 mg/L for EPZ 1 and approximately 0.021–0.023 mg/L for EPZ 2, EPZ 3, and EPZ 4, and the four major EPZs could be combined into three zones based on the response of P concentrations to changes in land use: EPZ 1 was least responsive, EPZ 2 and EPZ 4 were moderately responsive, and EPZ 3 was most responsive. Streams in EPZ 3, with the highest clay content, had high reference P concentrations (similar to EPZs 2 and 4), but were the most responsive to changes in land use.

Purpose and Scope

In 2001, the U.S. Geological Survey (USGS) and the Wisconsin Department of Natural Resources (WDNR), began a collaborative study to: 1) describe how the nutrient concentrations and biotic-community structure in streams differ throughout Wisconsin; 2) determine which environmental characteristics of watersheds are most strongly related to the distribution of nutrient concentrations in streams; 3) determine reference water-quality and biotic conditions for different areas of the State; 4) determine how the biotic community of streams in different areas of the State respond to changes in nutrient concentrations; 5) evaluate existing regionalization schemes in terms of describing patterns in reference water-quality conditions and patterns in biotic response to changes in nutrient concentrations; and 6) develop new multiparameter biotic indices to predict nutrient concentrations in streams. The ultimate goal of this study is to provide the information needed to guide the development of regionally based nutrient criteria for streams in Wisconsin.

Because the biotic response in streams was expected to vary as a function of stream size, and wadeable streams are sampled with different techniques than nonwadeable streams, the study was divided into two parts. The first part involved sampling 240 wadeable streams in 2001–03, and the second part involved sampling approximately 40 nonwadeable streams in 2003. In this report, the results of the first part of this study are presented: nutrient concentrations and their relations to the biotic integrity of wadeable streams in Wisconsin. The second part of the study will be presented in a separate report.

Approach

Because simultaneously collected hydrological, water-quality, and biological data were not available to determine how the biotic integrity of Wisconsin streams is related to changes in nutrient concentrations, a network of streams was selected to represent the wadeable and nonwadeable streams in the level III ecoregions and EPZs in the State. Although some level III ecoregions were combined into larger-scale national nutrient ecoregions and some of the EPZs were combined because there were no statistical differences in reference concentrations or in the responses to changes in land use among some zones in the upper Midwest study (Robertson and others, 2006), each of the level III ecoregions and EPZs were examined separately in this study. The locations of the 240 wadeable streams are shown in figure 4 and listed in appendix 1. To try to obtain streams that represent the range in environmental conditions in Wisconsin, approximately the same number of sites was chosen in each of the level III ecoregions. To try to obtain streams with a wide range in nutrient concentrations, sites in each ecoregion were chosen to try to represent a full range in the percentage of agricultural land, although this was not always possible. Discharge and water quality of each stream were sampled monthly over a 6-month period (May through October). Benthic (attached) algae and diatoms were sampled once during the period. During 2001, 157 streams with the smallest watersheds were sampled (2.2 to 222 km², but generally less than 90 km²). During 2002, 78 larger streams were sampled (40 to 1,947 km²). In 2003, 42 nonwadeable streams (not discussed in this report) and five additional wadeable streams were sampled (11 to 106 km²). Data on macroinvertebrate and fish populations were not collected as part of this study, but were available from past surveys. A prerequisite for site selection was that the macroinvertebrate and fish populations in the stream had been sampled during the past 5 years.

Figure 4. Sites on wadeable streams in Wisconsin included in this study. Water-quality and biotic data for each site are given by site identification number in the appendixes.

For each site, the drainage basin was digitized and a geographic information system (GIS) was used to describe the environmental characteristics of the watershed. Various multivariate statistical approaches were then used to determine how the environmental characteristics of the watershed were related to water quality and biotic-community structure. The data were used to determine which stratification scheme (level III ecoregions or EPZs) best describes the distributions in reference nutrient concentrations and the responses in nutrient concentrations to changes in land use. Reference concentrations of P, N, and SCHL, and water clarity were estimated by use of the multiple linear-regression approach and the 25th-percentile approach for the best regionalization scheme. Reference values for the biotic indices were estimated by use of the 75th-percentile approach by examining the values of the biotic indices at minimally impacted sites (sites with nutrient concentrations at or below the estimated reference concentrations). Water-quality data were statistically compared with biotic indices describing the suspended and benthic algae, diatoms, macroinvertebrates, and fish to determine how the biotic integrity of streams is related to changes in nutrient concentrations, and whether or not thresholds in P and (or) N concentrations can be defined above which the biology is adversely affected. Two new multiparameter indices were then developed to estimate P and N concentrations in streams on the basis of the biotic-community structure.

Methods of Data Collection and Analysis

Field Methods

Discharge, Water Chemistry, and Suspended Chlorophyll a Concentrations

Streamflow and water quality in each stream were sampled monthly over a 6-month period (May through October). Each site was sampled near the middle of the month regardless of flow conditions. During each visit, discharge and field parameters (specific conductance, water temperature, dissolved oxygen, and pH) were either measured or estimated and a water-quality sample was collected.

Discharge was determined at each site with a current meter (Rantz and others, 1982), with a stage/discharge relation for a continuous-recording gaging station at the site, or estimated from a nearby site. If the site was not wadeable because of high flow and did not have a continuous-recording gage, the discharge was estimated from a nearby streamflow-gaging station by use of relations between previous discharge measurements at the site and at the nearby station.

Specific conductance, water temperature, dissolved oxygen, pH, and, at some sites, turbidity were measured in the field at the time of sampling by use of a multiparameter meter. The meters were calibrated each day before use. Water clarity was measured by use of a 120-cm Secchi tube (also referred to as a transparency tube, U.S. Environmental Protection Agency, 2004). The Secchi tube was held into the flowing stream and filled. The tube was then held perpendicular to the ground and drained until the Secchi disk at the bottom of the tube became visible. The water level in the tube was read to the nearest centimeter and defined as the Secchi tube depth (SD). If the disk was visible when the tube was full, the value was reported as greater than 120 cm.

All water samples were collected by use of the equal-width-increment (EWI) method with a hand-held DH-59 depth-integrating sampler (Edwards and Glysson, 1999), except when stream conditions were not appropriate (stream velocity less than approximately 0.45 m/s, maximum depth less than 0.15 m, or the stream was nonwadeable); in this case, a grab sample was collected with an open bottle at the center of the flow. Samples were then split into appropriate bottles for lab analysis. Samples to be analyzed for dissolved constituents were filtered in the field through 0.45-µm membrane filters. Samples to be analyzed for SCHL were obtained by filtering a known volume of water through a 5-µm membrane filter. The filter was then placed in a labeled petri dish and wrapped in aluminum foil. All samples were chilled until they were delivered to the Wisconsin State Laboratory of Hygiene for analysis, except samples to be analyzed for SCHL, which were frozen, kept in the dark, and delivered to the WDNR Research Laboratory. All samples were analyzed for P, dissolved phosphorus (DP), total Kjeldahl nitrogen (TKN), dissolved nitrite plus nitrate nitrogen (NO₃-N), dissolved ammonia nitrogen (NH₄-N), and SCHL. In July of 2002, samples were also collected for analysis of suspended sediment. All chemical analyses of water samples (except SCHL) were done by the Wisconsin State Laboratory of Hygiene in accordance with standard analytical procedures described in the "Manual of Analytical Methods, Inorganic Chemistry Unit" (Wisconsin State Laboratory of Hygiene, 1993). At the WDNR Research Laboratory, the filters for SCHL analysis were placed in tubes containing 90 percent acetone, stored at least 24 hours, sonicated for 15 minutes, and stored an additional 24 hours in a freezer. The trichromatic chlorophyll a content of the samples was determined by means of a USEPA-approved method (Greenberg and others, 1992). Throughout this report, the water-chemistry, water-clarity, and SCHL data are collectively referred to as "water-quality data." All water-quality data were input into the USGS National Water Information System (NWIS) (U.S. Geological Survey, 1998).

Benthic Chlorophyll a and Diatoms

Samples for benthic chlorophyll a (BCHL) and periphytic-diatom analyses were collected once during August or September. Care was taken to avoid collecting samples within 2 weeks of appreciable rainfall to minimize the potential effect of scouring. Samples were collected by brushing a known area of three to five rocks with a toothbrush. Following collection, the samples were placed on ice and kept in the dark. Within 12 hours of sampling, the sample was diluted to a known volume with distilled water, homogenized in a blender, and a portion was filtered through two 3–5 µm glass-fiber filters. One filter was placed in 90 percent acetone and was analyzed for its acid-corrected chlorophyll a content (BCHL) by means of a USEPA-approved monochromatic method (Greenburg and others, 1992). The other filter was used for the determination of ash-free dry weight. This sample was dried overnight at 105°C and ashed at 550°C for 1 hour. The sample was weighed before and after ashing. Because of either lack of suitable substrate or lack of water at the time of collection, BCHL samples were collected from only 199 sites.

The sample for microscopic analysis of the diatom assemblage was obtained from a portion of the homogenized sample used for BCHL analysis; however, if rocks were not available, samples were collected from sticks. The sample was preserved with Lugols solution and cleaned with hydrogen peroxide and potassium dichromate (van der Werff, 1955). A portion of the diatom suspension was dried on a cover slip and mounted in Naphrax. Specimen were identified and counted under an oil-immersion objective (1,400 or 1,750X). At least 300 diatoms were counted from two slides in each sample. The keys used to identify the species included Patrick and Reimer (1966, 1975), Camburn and others (1984–86), Dodd (1987), and Krammer and Lange-Bertalot (1986, 1988, 1991a,b). Because of either lack of coarse substrate or lack of water at the time of collection, samples were collected from only 214 sites.

Physical Habitat and Fish

Physical-habitat and fish data were collected by the WDNR once at each site between 1997 and 2002. The physical habitat was determined for a stream length equal to 35 times the mean stream width, or a minimum of 100 m. This length is generally sufficient to encompass about three meander sequences (Simonson and others, 1994; Wang and others, 1996). The physical habitat and fish were quantified between late May and late August when low flows facilitated effective sampling and large-scale seasonal fish movement was unlikely to occur (Lyons and Kanehl, 1993). At each site, 30 physical-habitat characteristics, including channel morphology, bottom substrates, cover, bank conditions, riparian vegetation, and land cover were measured or visually estimated along 12 transects by use of standard procedures (Simonson and others, 1994). The entire length of each site was electrofished with either two backpack units in tandem or a single tow-barge unit with three anodes (Lyons and Kanehl, 1993; Simonson and Lyons, 1995). Efforts were made to collect all of the fish greater than or equal to 25 cm in length. All captured fish were identified to species, counted, and weighed in aggregate by species.

Macroinvertebrates

Macroinvertebrate samples were collected by the WDNR once at each site between 1999 and 2002. Two types of macroinvertebrate samples were collected according to procedures described by Hilsenhoff (1988). Samples were generally collected from each site during low flow in early October by use of a 600-µm mesh D-frame kick net. The first sample was collected from riffles or rocky substrates. If rocky substrates were not present, then vegetative snags (areas with overhanging grasses, logs, woody debris, and leaf packs) were sampled to ensure that the sample was from the most representative habitat at each site. The second sample was collected only from snags to ensure that these data were comparable between all sites because rocky substrates were not always present.

Samples from riffle or rocky substrates were collected by placing the net on the stream bottom and kicking an area immediately upstream of the net to dislodge the macroinvertebrates and wash them into the net. In addition, individual rocks were picked up and the attached macroinvertebrates were removed and added to the sample. This process was repeated in at least three locations within the same riffle or different riffles until at least 200–300 organisms were collected. Samples were collected in snags by placing a net in the water column downstream of the snag, where it would collect most of the dislodged debris. The snags were then disturbed by scraping or shaking them with a net, hands, or feet. At each site, all available snag types at multiple locations were sampled, with first consideration given to larger snags in higher water-velocity habitats.

Samples were sorted and identified at the laboratory of Dr. Stanley Szczytko at the University of Wisconsin, Stevens Point. The samples were placed in a glass pan positioned over a 6.5-cm² grid. All of the organisms from randomly chosen grid squares were selected until a minimum of 125 organisms having tolerance values cited in the literature (such as the values in Hilsenhoff, 1988) were picked, or, until the entire sample was sorted. All of the picked organisms were counted and identified to the species or the lowest taxonomic level possible.

Watershed Boundaries and Environmental Characteristics

Watershed boundaries for the sampled streams were manually digitized from 1:24,000-scale USGS topographic quadrangle maps. The environmental characteristics thought to affect the water quality and biology in the streams were compiled for each watershed used in this study: land use/land cover (Lillesand and others, 1998); soil characteristics (from the USSOILS digital coverage of the State Soil Geographic, STATSGO, data base; Schwarz and Alexander, 1995); types of surficial deposits (Fullerton and others, 2003); annual air temperature and precipitation (National Climatic Data Center, 2002); mean land-surface slope (based on 30-m DEM data resampled to 100 m; U.S. Geological Survey, 1999); and average annual runoff (Gebert and others, 1987).

Point-source loadings of phosphorus upstream of each sampling site (PtS) were estimated from monthly mean P concentrations and monthly mean discharge volumes as reported by the dischargers (for example, wastewater-treatment plants and cheese factories) in their Discharge Monitoring Report with the WDNR (James Baumann, Wisconsin Department of Natural Resources, written commun., 2004). The number of concentrations and discharge volumes reported in a month varied with the size and type of discharger and ranged from one sample per month to daily samples. For sites where P concentrations were not required to be measured and, therefore not reported, the P concentrations were estimated based on the size and type of discharger.

All basin characteristics were compiled in digital form by use of a GIS. A digital coverage of each watershed was used to compute the average or percentage value for each environmental characteristic, including the PtS for each of the 240 watersheds. A summary of the environmental characteristics (with the specific metric describing each environmental characteristic) for all of the watersheds used in this study is given in table 2.

Table 2. Summary statistics for median monthly water-quality and environmental (anthropogenic/land-use, basin, soil, and surficial-deposit) characteristics of the watersheds of the sites in the studied wadeable streams in Wisconsin.

[mg/L, milligram per liter; log, logarithm to base 10 transformation; µg/L, microgram per liter; C, Celsius; µS/cm, microSiemen per centimeter; cm, centimeter; (m³/s)/km², cubic meter per second per square kilometer; km², square kilometer; mm, millimeter; mm/year, millimeter per year; --, unitless; %, percent; mm/hr, millimeter per hour; kg/km², kilogram per square kilometer; >, greater than; no PtS, only sites with less than 12 kg/km² of point-source loading of phosphorus are included in this part of the analysis; summary statistics based on monthly values]

^a All values greater than 120 cm were set to 120 cm for computation of summary statistics, which result in the mean values being biased low.

Data Summaries

All of the water-quality data collected in this study were input into the USGS NWIS database (U.S. Geological Survey, 1998) and are summarized in appendix 1. In computing summary statistics, all of the data were used regardless of whether or not flow could be detected. During some samplings, the water in the streams was found to have dried up, and no water-quality data were collected. All data reported as less than the detection limit were set to one-half of the detection limit, and all SD data greater than 120 cm were set to 120 cm prior to any statistical and graphical analyses.

Physical-habitat data were summarized into: mean wetted width, depth, thalweg depth, and stream gradient; the percentage of the stream reach with riffles, runs, or pools; the mean depth of sediment; the percentage of the bottom of the stream reach composed of different substrates, embedded rocky substrate, and covered by algae or macrophytes; the percentage of the stream reach that contains fish cover, is shaded, and that has streambank erosion; and buffer width (Simonson and others, 1994). The physical-habitat data for each site are summarized in appendix 2.

Three metrics were used to summarize the diatom-community data: the Diatom Nutrient Index (DNI), the Diatom Siltation Index (DSI; Bahls, 1992), and the Diatom Biotic Index (DBI). The DNI computations are based on tolerance values assigned to individual taxa. DNI values range from 1 to 6, with 1 representing species typically found with the lowest nutrient concentrations (oligotrophic, good water quality) and 6 typically representing species found with the highest nutrient concentrations (hypereutrophic, poor water quality). The values for Wisconsin diatoms (appendix 3) were generated largely from Van Dam and others (1994), but values were also assigned based upon experience with the diatom communities in Wisconsin. If no autecological data were known, the taxa were not assigned a value and were not included in the DNI calculation. Because the index is based upon relative abundance, rare species have little effect on the final index value. The formula used to calculate DNI value is

(1),

where

n_i = number of individuals of species i;

t_i = tolerance-index value for species i;

j = total number of species in the sample with tolerance-index values; and

N = total number of individuals in the sample having tolerance-index values.

The second metric for the diatom community is the Diatom Siltation Index (DSI). This index is based on the sum of all individuals in the Navicula (including Cavinula, Chamaepinnularia, Craticula, Diadesmis, Fallacia , Fistulifera, Geissleria, Hippodonta, Kobayasiella, Luticola, Mayamaia, Placoneis , and Sellaphora ), Nitzschia (including Psammodictyon and Tryblionella ), and Surirella taxa. These taxa were chosen because they have good motility; therefore, this metric reflects the degree of siltation in a reach (Bahls, 1992). The scale for the index is 0–100 with lower values indicating less silt and thus better water quality.

To assess stream biotic integrity, a multimetric index called the Diatom Biotic Index (DBI) was created. The DBI is based on both diatom indices, DNI and DSI. For computing the DBI, each metric was standardized to the 95th percentile for a number of reference sites and then the two metrics were averaged. For sites with an individual metric exceeding its 95th percentile, the metric was set to 100. The scale of the DBI is 0 to 100, with higher values indicating better biotic integrity. The DBI is intrinsically designed to be sensitive to nutrient enrichment and the effects of sedimentation. The reference sites used to standardize the metric were chosen by combining the northern ecoregions (NLF, NCHF) and the southern ecoregions (DFA, SWTP). Reference sites for the southern ecoregions were those where P concentrations for August were less than or equal to 0.050