BIOTIC INDICES AND STREAM ECOSYSTEM PROCESSES: RESULTS FROM AN EXPERIMENTAL STUDY 1

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1 Ecological Applications, 6(1), 1996, pp by the Ecological Society of America BIOTIC INDICES AND STREAM ECOSYSTEM PROCESSES: RESULTS FROM AN EXPERIMENTAL STUDY 1 J. BRUCE WALLACE Department of Entomology and Institute of Ecology, University of Georgia, Athens, Georgia USA JACK W. GRUBAUGH, 2 AND MATT R. WHILES 3 Institute of Ecology, University of Georgia, Athens, Georgia USA Abstract. We investigated the ability of the North Carolina Biotic Index (NCBI) and the Ephemeroptera + Plecoptera + Trichoptera (EPT) index to track an experimental manipulation of the invertebrate community and resultant alteration of several ecosystem-level processes in a headwater stream at the Coweeta Hydrologic Laboratory in western North Carolina. Indices were calculated from quantitative monthly or bimonthly benthic samples of moss-covered rockface and mixed substrate habitats, as well as habitat-weighted values based on the proportion of each habitat in the two streams. One stream (C 55) served as a reference stream over the 6-yr period of late 1984 through 1990, whereas the other (C 54) received seasonal treatments with an insecticide for 3 yr ( ). Throughout pretreatment, treatment, and recovery, both the NCBI and EPT indices tracked the disturbance regime of the treatment stream. Indices for the reference stream varied little during the 6-yr period. Both the NCBI and EPT suggested strong changes in the treatment stream during treatment relative to both pretreatment and the reference stream. Following cessation of insecticide treatments, both indices reflected improved biotic conditions during first and second years of recovery in C 54. Compared with fauna of mixed substrates, rockface fauna had lower (better) NCBI values during pretreatment, and exhibited a greater proportional increase in tolerant taxa during treatment than mixed substrates, emphasizing the importance of including rockface communities in environmental monitoring programs. Changes in both the EPT and NCBI indices closely corresponded to changes in ecosystemlevel processes observed in C 54 from pretreatment to treatment, and recovery periods. These processes include: leaf litter processing rates, organic matter storage, fine paniculate organic matter generation and export, and secondary production. With the exception of organic matter storage, all of these processes declined during treatment of C 54, and subsequently increased during recovery. Our results demonstrate the potential of such indices to detect and monitor stream ecosystem changes during and following disturbance. The EPT index was by far the easiest to use from both the standpoint of time required for sample processing and ease of application. Compared with the labor-intensive sample processing, specimen identification and measurement, and data entry required for secondary production calculations, the EPT index was relatively simple and displayed a remarkable ability to track secondary production of invertebrates in the treatment stream. Our data strongly support the inclusion of the EPT and NCBI indices in these southern Appalachian headwater streams as indicators of both degradation and recovery of stream ecosystem processes from chemical-induced disturbance. Key words: biomonitoring; biotic indices; ecosystem processes; EPT index; macroinvertebrates; manipulation; NCBI index; pesticides; recovery; secondary production; streams. INTRODUCTION Biological assessment of aquatic environments has been practiced since the early 1900s (see reviews of Hynes 1960, 1994, Cairns and Pratt 1993). As evidenced by a number of recent books devoted entirely to the subject (e.g., Abel 1989, Plafkin et al. 1989, Rosenberg and Resh 1993, Loeb and Spacie 1994), 1 Manuscript received 15 September 1994; revised 26 January 1995; accepted 16 February Present address: Department of Biology, Ellington Hall, University of Memphis, Memphis, Tennessee USA. 3 Present address: Department of Biology, Division of Mathematics and Sciences, Wayne State College, Wayne, Nebraska USA. 140 interest in this area has grown tremendously. As demands on water resources increase, ambient biological monitoring is becoming a rapid and accurate means of assessing quality of lotic systems, although specific methods remain limited (Cummins 1994). Many methods have assessed stream quality using invertebrates, ranging from assessing physiological and morphological changes of individuals to various measures of community structure (e.g., Rosenberg and Resh 1993). Biotic indices, which evaluate macroinvertebrate community structure, are widely used. These indices often follow the approach described by Chutter (1972), in which tolerance values (TV) ranging from 0 (very intolerant) to 10 (very tolerant) are based on the ability

2 February 1996 BIOTIC INDICES AND ECOSYSTEM PROCESSES 141 of a taxon to inhabit stream or river systems differing in water quality (e.g., Hilsenhoff 1977, Plafkin et al. 1989, Lenat 1993). Other organisms such as fish also have been used extensively in biological monitoring programs (e.g., Plafkin et al. 1989, Loeb and Spacie 1994). Construction of biological indices requires considerable effort (i.e., Hilsenhoff 1987, Karr 1991, Lenat 1993, Kerans and Karr 1994). However, once derived, such indices, like most ecological studies, may have problems such as unclear identification of habitats and habitat characteristics sampled, differences in sampling intensity, lack of both pre- and post-impact data, and reliance on only one "unimpaired" reference site (Lenat and Barbour 1994). Years of sampling may be required to collect sufficient data to detect long-term trends, and determining whether an apparent trend is due to anthropogenic causes or natural variation may be impossible (Charles et al. 1994). Alternatively, biological monitoring offers a relatively affordable means of environmental measurement, compared to chemical data, for assessing degradation of aquatic habitats and loss of biological diversity induced by anthropogenic disturbances (Karr 1991, Hynes 1994). Biological evaluation also must take place against a firm knowledge of local fauna and flora (Karr 1991). Karr (1991, 1993) argues that biological indices, which incorporate concepts such as biological diversity and integrity, provide important measures of ecosystem health. Accordingly, the use of such indices, which encompass individual to landscape-level perspectives, should provide an ecologically robust means of assessing ecosystem health (Karr 1993), although other points of view have been expressed (Suter 1993). Further, experimental approaches, which incorporate aspects of biological monitoring into ecosystem-level manipulations, are lacking. A "top-down" ecosystem-level manipulation of a headwater stream at Coweeta Hydrologic Laboratory (North Carolina, USA) was designed to assess the role of macroinvertebrates (primarily insects) in ecosystem processes. Insecticide treatments reduced insect abundance (Wallace et al. 1989, 1991b), biomass, and secondary production (Lugthart and Wallace 1992), as well as leaf litter processing rates (Cuffney et al. 1990, Chung et al. 1993), without altering microbial respiration or abundance (Cuffney et al. 1990, Suberkropp and Wallace 1992). Treatments also caused an accumulation of in-stream leaf litter (Wallace et al. 1995) and reduced fine paniculate organic matter (FPOM) (Cuffney et al. 1990, Wallace et al. 199la) and paniculate inorganic matter (ash) (Wallace et al. 1993) concentrations and export, which subsequently increased during recovery. Thus, this manipulation demonstrated the importance of invertebrates in a series of ecosystem-level processes ranging from organic matter storage to detritus processing and export. We used these ecosystem-level observations to test if biotic indices corresponded to changes in ecosystem processes. Lenat (1993) has presented a biotic index for southeastern U.S. streams, called the North Carolina Biotic Index (NCBI). We applied this index to the 6-yr record of pretreatment, treatment, and recovery data available from the manipulation study described above to examine changes in stream biotic integrity. Additionally, we applied the EPT index (number of taxa belonging to the Ephemeroptera, Plecoptera, and Trichoptera [Crawford and Lenat 1989, Plafkin et al. 1989, Kerans and Karr 1994]) as a second measure of biotic integrity. We compared NCBI and EPT values for the treated stream to those calculated for a nearby reference stream during the same 6-yr period. Specifically, we address the following questions: Where impairments of ecosystem-level processes have been demonstrated, do the NCBI and EPT indices vary significantly from those of a nearby reference stream? Do NCBIs based on abundance and biomass of taxa differ? Do communities inhabiting different substrates differ in their response? Do indices track ecosystem recovery and how do recovery values differ from those of pretreatment and treatment? STUDY SITES Our two study streams were first-order and drain Catchments (C) 54 and 55 at the Coweeta Hydrologic Laboratory (U.S. Forest Service) in western North Carolina. Catchment vegetation was mixed hardwoods. A dense riparian growth of rhododendron resulted in heavy year-round shading of both streams. Elevation, catchment area, thermal regime, discharge, and aspect (southern) were similar for both streams (see Wallace et al. 1991a). Concentrations of most ions were low (<1 mg/l) and phs of 6.6 to 6.8 were similar to those of other streams in the Coweeta Basin (Swank and Waide 1987). Detailed descriptions of the Coweeta Basin can be found in Swank and Crossley (1987). Based on substratum measurements at 1-m intervals, 65-87% of the substrate in the treatment and reference streams, respectively, consisted of a heterogeneous mixture of cobble, pebble, gravel, sand, and silt, referred to here as mixed substrates (see Lugthart and Wallace 1992). The remaining 35% (treatment) and 13% (reference) of substrate were moss-covered rockface. Additional information on these streams has been reported by Wallace et al. (1989, 1991a), Cuffney et al. (1990), and Lugthart and Wallace (1992). Precipitation during our study was 91.8% of average (180 cm/yr) and encompassed extremes of the 60-yr record at Coweeta. During 1986, precipitation was the lowest on record (123.9 cm = 68.7% of average), whereas 1989 was the wettest year of record (234.1 cm = 129.8% of average). METHODS Benthic sampling Invertebrates in the mixed substrate habitats were sampled with a 400-cm 2 stovepipe corer. All material

3 142 J. BRUCE WALLACE ET AL. Ecological Applications Vol. 6, No. 1 within the corer was scooped out to a depth of =10 cm or until bedrock was encountered. Rockface habitats were sampled by scraping and brushing moss and associated material from a 15 X 15 cm area into a plastic bag pressed to the rock siii'face. Seven benthic samples, four from mixed substrates and three from rockface, were collected from each stream on each sampling date. Sample locations were selected randomly. If the appropriate habitat type was not present at the preselected location, the sample was collected at the closest upstream site. During pretreatment (October 1984 to November 1985), samples were collected monthly from both streams. Sampling started on 12 October During the first year of treatment (January 1986 to November 1986), samples were collected during alternate months from each stream. During December 1988 (end of treatment for C 54) through December 1990, samples were collected monthly in the treatment stream and during alternate months in the reference stream. Invertebrates were elutriated from inorganic substrates, passed through nested 1-mm and 250-jjum sieves, and preserved in a 6-8% formalin solution containing Phloxine B dye. All animals were removed from the 1-mm sieve by hand-picking under 15 x magnification and preserved in 6-8% formalin solution. Material in the <l-mm to 250-u,m size fraction was subsampled (Vs to V M of whole sample) using a sample splitter (Waters 1969). Animals were then removed by hand using a stereomicroscope (15X magnification). Invertebrates were identified (species or genus whenever possible) and counted. Larval Chironomidae (Diptera) were identified as either Tanypodinae or non-tanypodinae; however, a list of chironomid taxa in the treatment stream during initial and subsequent treatments has been published elsewhere (Wallace et al. 1991i>). Most non-insect invertebrates were identified to the ordinal level or higher. Total body length of each individual was measured to the nearest millimetre using 12x magnification with a graduated stage microscope or ocular micrometer. For crayfish, Cambarus bartonii (Cambaridae: Decapoda), carapace lengths were measured. Biomass (ash-free dry mass [AFDM]) estimates for most taxa were obtained using length-mass regressions derived from animals in the study streams or nearby Coweeta streams (Huryn 1986; J. B. Wallace, G. J. Lugthart, and J. O'Hop, unpublished data). Biomass of small, non-insect taxa was obtained from mean mass of >50 individuals in subsamples of representative size classes. Abundances and biomass were estimated separately for mixed substrate and rockface substrates. Habitatweighted abundances and biomass for each stream were then calculated according to the proportion of rockface and mixed substrate present in each stream. Information on methods used to estimate secondary production of invertebrates during pre-treatment and the initial treatment year (1986) can be found in Lugthart and Wallace (1992). Biotic indices Two separate indices were used for assessment of stream biotic integrity in C 54 and C 55: the North Carolina Biotic Index (NCBI) and taxonomic richness of intolerant taxa (Ephemeroptera + Plecoptera + Trichoptera, EPT). EPT is a sensitive indicator of stream perturbations (Crawford and Lenat 1989, Eaton and Lenat 1991) and is widely used by various agencies as part of their environmental monitoring programs (Lenat 1988, Plafkin et al. 1989). The NCBI is based on an extensive data set of benthic stream samples collected throughout North Carolina and is designed to be specific to mountain, Piedmont, or coastal ecoregions of the southeastern United States (Lenat 1993). The NCBI is calculated as: where S is the number of taxa, TV, is the tolerance value of the i* 1 taxon, N, is density of the /' lh taxon as either abundance (numbers per square metre) or biomass (milligrams ash-free dry mass per square metre), and N, is total abundance (or biomass) of macroinvertebrates in the sample. Tolerance values range from 0 (highly intolerant taxa) to 10 (highly tolerant taxa). Taxa from our streams were assigned TVs provided by Lenat (1993) with several modifications. Taxa classified to genus (i.e., Serratella spp. [Ephemeroptera] and Rhyacophila spp. [Trichoptera]) were assigned mean tolerance values of species known to occur or likely to occur in the study streams. Chironomids (Diptera) were assigned a tolerance value of 5.7, based on a mean value provided by Lenat (1993), which may be conservative for our streams. Based on data from Wallace et al. (1991 ft), mean tolerance values of drifting chironomid taxa during insecticide applications ranged from 4.2 to 5.6. For crayfish, the tolerance value of 8.1 provided for collective Cambarus spp. is probably too high for our insecticide treatment, because C. bartonii was quickly eradicated from C 54 during treatment and did not return during recovery (Lugthart and Wallace 1992, Whiles and Wallace 1992). Thus, crayfish were excluded from NCBI calculations. Sampling and sample processing equipment was fitted with meshes of 250 u,m. Larger mesh sizes (i.e., 1 mm) are generally used in bioassessment and biomonitoring protocols (see Plafkin et al. 1989), and collections made with smaller meshes will bias results toward abundant smaller taxa such as Nematoda and Chironomidae. Further, any TVs assigned to nematodes are at best tenuous because they are rarely, if ever, included in freshwater biotic indices (D. Lenat, personal communication). Lugthart et al. (1990) found =55% of non-tanypodinae chironomids collected in C 54 and C 55 were in the 1-mm size class (<1 mm). To make our

4 February 1996 BIOTIC INDICES AND ECOSYSTEM PROCESSES 143 results comparable to other studies, all nematodes were excluded from abundance-based NCBI calculations and non-tanypodinae chironomid abundances were reduced by 55%. Nematodes also were excluded from biomass-based NCBI calculations; 55% of non-tanypodinae abundance was multiplied by mg (average individual AFDM of a 1-mm chironomid) and subtracted from non-tanypodinae biomass to correct biomass-based calculations. Benthic densities from replicate monthly samples were composited by individual habitat type on each sampling date and used to calculate NCBIs. This produced 44 separate NCBIs for both mixed substrate and rockface habitats in the treatment stream, and 32 NCBIs for mixed substrate and rockface habitats in the reference stream. The total number of monthly sampling dates by treatment period was as follows: pretreatment (C 54 = 13 and C 55 = 13); treatment (C 54 = 7 and C 55 = 7); first-year recovery (C 54 = 12 and C 55 = 6); and, second-year recovery (C 54 = 12 and C 55 = 6). To correct for seasonal variation, 0.4 was added to NCBIs derived from autumn data (September-November) and 0.5 was added to indices derived from winter/spring data (December-May) (Lenat 1993). Habitat-weighted NCBI values were calculated for both streams by multiplying habitat-specific NCBIs by the proportion of mixed substrate and rockface habitat in each stream and summing products for individual sampling dates. Insecticide treatment The insecticide methoxychlor (l,l,l-trichloro-2,2- bis [p-methoxyphenyl] ethane; CAS Number ; Southern Agricultural Chemical Company, Kingstree, South Carolina) was applied seasonally to C 54 during a 3-yr period at a rate of 10 mg/l based on discharge at the flume (Wallace et al. 1989, 1991b). The entire stream channel from flume to spring seeps was treated, using two hand-sprayers to apply the insecticide for 4 h (December 1985), followed by 2-h seasonal treatments (every 3 mo) during March 1986 through October Additional details of treatment, including methoxychlor concentrations in stream water, can be found in Wallace et al. (1989). Information about recolonization by invertebrates between subsequent treatments is given by Wallace et al. (1991i>). The reference stream received no insecticide throughout the study. Data analyses Several within- and among-stream comparisons did not meet the assumption of equal variance; therefore, nonparametric methods were used for most analyses (Elliott 1977). To test for differences among multiple comparisons for specific periods (i.e., pretreatment, treatment, recovery year 1, and recovery year 2) and between streams, we used a Kruskal-Wallis ANOVA on ranks, and all pairwise comparisons were tested US- TABLE 1. Average NCBIs (North Carolina Biotic Index) for invertebrate abundances and biomass on rockface and mixed substrates in treatment and reference streams with and without crayfish, nematodes, and chironomids <1 mm in length during October 1984 through December 1990.t Abundances Rockface Mixed substrates Biomass Rockface Mixed substrates Treatment Reference With Without With Without ** ** ** t Asterisks indicate significantly lower biotic index (higher score) for a given comparison within each stream by Mann- Whitney rank sum test, ** P < ing Dunn's test (Zar 1984). Pairwise comparisons were made using the Mann-Whitney rank sum test. Despite readily apparent results from observation, as well as the above statistical tests, we are aware that the study is pseudoreplicated, employing one treatment and one reference stream. To circumvent this problem, we also used Randomized Intervention Analysis (RIA, Carpenter et al. 1989) for between-stream comparisons of NCBIs during specific time periods. Time periods included: pretreatment (October 1984 through November 1984); treatment (January 1986 through December 1988); Ist-yr recovery (January 1989 to December 1989); and 2nd-yr recovery (January 1990 to December 1990). As pointed out by Carpenter et al. (1989), RIA tests the null hypothesis that no change occurred in the experimental stream relative to the reference stream following a manipulation. Rejection of the null hypothesis does not demonstrate that our pesticide manipulation caused the change, only that a change occurred. RESULTS NCBIs Abundance and biomass NCBIs, both with and without excluded taxa (i.e., crayfish, nematodes, and non- Tanypodinae chironomids <1 mm in length), were compared for similar substrates in both streams. Exclusion of these groups did not significantly influence NCBIs for abundances or biomass on rockface substrates in either stream. However, exclusion did result in significantly lower NCBIs for abundances in mixed substrates in both streams and biomass in the reference stream (Table 1). The difference in NCBIs for biomass in mixed substrates among streams was primarily attributable to crayfish with high TVs (8.1), which were eliminated from the treatment stream following initial treatment and, in contrast to insects, did not recolonize during the remainder of the study. Throughout the remainder of this paper we will restrict all NCBI comparisons of abundances and biomass to samples without chironomids <1 mm, nematodes, and crayfish.

5 144 J. BRUCE WALLACE ET AL. Ecological Applications Vol. 6, No. 1 TABLE 2. Average NCBIs (North Carolina Biotic Index) for invertebrate abundances and biomass on rockface and mixed substrates during different time periods in the treatment and reference streams. Values are presented without crayfish, nematodes, and chironomids <1 mm in length. Periods refer to insecticide treatment of C 54 and are as follows: pretreatment = October 1984 to November 1985; treatment = December 1985 to December 1988; recovery year 1 = January to December 1989; and recovery year 2 = January to December 1990.t Abundance Pretreatment Treatment Recovery year 1 Recovery year 2 Biomass Pretreatment Treatment Recovery year 1 Recovery year 2 Treatment Rockface 5.00*** V *** 2.11*** * 3.21*** stream Mixed * Reference stream Rockface Mixed 5.24* * * *** *** ** ** 5.21 t Asterisks indicate significantly lower biotic index (higher score) for a given substrate comparison within each stream and period by Mann-Whitney rank sum test, *P < 0.05, **P < 0.01, ***P < Substrate, abundance, and biomass comparisons NCBIs based on macroinvertebrate abundances were consistently lower (higher biotic integrity) for rockface than mixed substrate habitats, with the exception of C 54 during treatment (Table 2), when rockface abundances generally displayed higher NCBIs than mixed substrates. Hence, rockface communities in the treatment stream showed greater departure from pretreatment conditions than mixed substrates during treatment. However, recovery of abundance-based NCBIs was more rapid for rockface than mixed substrates following cessation of insecticide treatments. During 2ndyr recovery, abundance-based NCBIs on rockface substrates in the treatment stream were lower than pretreatment values (Table 2). Biomass-based NCBIs were lower than those for abundance during pretreatment and recovery (Table 2), particularly for rockface habitats where large filterfeeding taxa such as Parapsyche cardis, which have low TVs, tend to dominate. In contrast, chironomids, which have higher TVs, constitute a large portion of abundances and a much smaller proportion of biomass on such habitats. Recovery of rockface faunal biomass following treatment of C 54 was somewhat slower than that of abundance, as biomass-based NCBIs remained above that of pretreatment well into the second year of recovery (Table 2). Temporal patterns of NCBIs Abundance-based NCBIs on mixed substrates and rockface, along with habitat-weighted values, did not differ significantly in the reference stream throughout the study (P > 0.18 to P > 0.60), nor did those for biomass (P > 0.51 to P > 0.60, Kruskal-Wallis test, ANOVA on ranks). However, substantial changes were evident in the treated stream (Fig. 1), where abundancebased NCBIs displayed large increases during treatments. This effect persisted well into the first year of recovery (Fig. 1). Significant differences existed for NCBIs within the treatment stream between pretreatment vs. treatment and Ist-yr recovery (Table 3). During the 2nd-yr recovery period, NCBIs based on both OJ (0 58 9l si 8 ' i 7 ' > C *-Treatment * A A -o- Reference Years ' \ rj,., '. ',.' i Elapsed Days FIG. 1. NCBIs based on invertebrate abundances on (top to bottom) mixed substrates, rockface, and habitat-weighted values for C 54 (treatment) and C 55 (reference), plotted against elapsed days from initial sampling (12 October 1984). S = start of seasonal insecticide treatments of C 54 and E = end of treatments. Elevated index score indicates decreased biological integrity.

6 February 1996 BIOTIC INDICES AND ECOSYSTEM PROCESSES 145 TABLE 3. Probability values for ANOVA based on ranks of NCBIs (North Carolina Biotic Index) for within-stream comparisons of both habitat-weighted abundances and biomassf for the treatment stream and the reference stream. Comparison periods Pretreatment vs. treatment Pretreatment vs. recovery year 1 Pretreatment vs. recovery year 2 Treatment vs. recovery year 1 Treatment vs. recovery year 2 Recovery year 1 vs. recovery year 2 Treatment Habitatweighted abundance <0.001* <0.05* >0.50 >0.20 <0.001* >0.10 stream Habitatweighted biomass <0.001* <0.01* >0.50 >0.20 <0.01* >0.10 Reference Habitatweighted abundance >0.25 >0.25 >0.25 >0.25 >0.25 >0.25 stream Habitatweighted biomass >0.55 >0.55 >0.55 >0.55 >0.55 >0.55 t Kruskal-Wallis one-way ANOVA on ranks, differences among periods tested using Dunn's multiple comparison procedure. t Asterisks indicate significant differences between comparison periods (P < 0.05). habitat-weighted abundances and biomass did not differ significantly from either Ist-yr recovery or pretreatment. Average habitat-weighted NCBIs during 2nd-yr recovery declined and differed significantly from treatment (Table 3). One of six between-stream comparisons for abundances and biomass NCBIs differed during the pretreatment period. Mixed substrate abundance NCBIs in the treatment stream were significantly greater than those of the reference stream during the pretreatment year, whereas the other five comparisons had probability values ranging from 0.18 to 0.64 (Table 4). During treatment, all NCBIs in the treatment stream were significantly greater (P < ) than those of reference stream. During Ist-yr recovery only indices for mixed substrate abundances, rockface abundances, and rockface biomass in the treatment stream remained significantly higher than those of the reference (Table 4). In 1990 (2nd-yr recovery) there were no differences in NCBIs among streams for either abundances or biomass (P > , Mann-Whitney rank sum test, Table 4). Randomized intervention analysis (RIA) rejected the null hypothesis that no significant changes in NCBIs occurred in the treatment stream relative to the reference for pretreatment and treatment periods, and treatment and 1st- or 2nd-yr recovery for either habitatweighted abundances or biomass (Table 5). In contrast, between-stream RIA analysis did not reject the null hypothesis that no significant differences existed between streams during the pretreatment, or among recovery years, and indicated a strong treatment effect (Table 5). EPT index Before treatment, average number of EPT taxa did not differ among streams (Table 6). Throughout the study, EPT indices remained relatively consistent in the reference stream, ranging from 15 to 24 taxa on each collection date. In contrast, dramatic changes occurred in the treated stream (Fig. 2). Number of EPT taxa ranged from 13 to 21 taxa in the treatment stream prior to treatment and declined dramatically to 2-8 taxa following initial treatment (=day 430, Fig. 2). EPT taxa remained low throughout the treatment period and initial 6 mo of recovery in the treatment stream (summer 1989, =day 1750, Fig. 2). In midsummer 1989, EPT taxa increased sharply as recolonization occurred. Following the sharp decrease with initial treatment, the average EPT index for the treatment stream was only 5.4 during the treatment period, 10.8 during Ist-yr recovery, and increased to 16.5 during 2nd-yr recovery (Table 6). In contrast, EPT of the reference stream averaged 19 during treatment, 20.5 during Ist-yr recovery, and 18.3 during 2nd-yr recovery of the treatment stream. Treatment stream EPTs during treatment and Ist-yr recovery were significantly lower than those of pretreatment (Table 6). Second-yr recovery EPTs in the treatment stream did not differ significantly from pretreatment. Between-stream comparisons show that EPTs did not differ between treatment and reference streams during pretreatment and 2nd-yr recovery pe- TABLE 4. Between-stream (treatment vs. reference) probability values for comparisons of NCBIs (North Carolina Biotic Index) based on abundances and biomass during different time periods of the study. Periods refer to insecticide treatment of the treatment stream and are as follows: pretreatment = October 1984 to November 1985; treatment = December 1985 to December 1988; recovery year 1 = January to December 1989; and recovery year 2 = January to December Comparison! Habitat-weighted abundances Habitat-weighted biomass Mixed substrate abundances Mixed substrate biomass Rockface abundances Rockface biomass Pretreatment 0.57 NS 0.51 NS 0.02* 0.18 NS 0.33 NS 0.64 NS Treatment * * * * * * t Comparisons made by Mann-Whitney rank sum test, *P < 0.05, NS = P > Recovery year NS 0.61 NS 0.02* 0.24 NS 0.044* 0.006* Recovery year NS 0.61 NS 0.11 NS 0.96 NS 0.96 NS 0.13 NS

7 146 J. BRUCE WALLACE ET AL. Ecological Applications Vol. 6, No. 1 TABLE 5. Results of Randomized Intervention Analysis (RIA) performed on NCBIs (North Carolina Biotic Index) based on habitat-weighted abundances and biomass between the treatment stream (C 54) and reference stream (C 55), testing the null hypothesis that no change occurred in the treatment stream relative to the reference stream. Comparison periods Pretreatment vs. treatment Pretreatment vs. recovery year 1 Pretreatment vs. recovery year 2 Treatment vs. recovery year 1 Treatment vs. recovery year 2 Recovery year 1 vs. recovery year 2 * = strong effects of treatment. Habitat-weighted abundances, P < * * < * Habitat- weighted biomass, P < * * 0.006* riods. In contrast, treatment and Ist-yr recovery EPTs differed significantly from those of the reference stream (Table 6). Biotic indices and ecosystem processes Previous studies on these streams have documented several ecosystem-level consequences resulting from this top-down manipulation of the invertebrate community. For example, these heavily shaded, heterotrophic streams primarily depend on allochthonous inputs of leaf litter, which serve as the energy base of the system. Following invertebrate reduction, leaf litter decomposition within the treated stream was greatly reduced compared with pretreatment and untreated reference streams. The time required for 95% processing of leaf litter was about twice that of untreated streams (Cuffney et al. 1990, Chung et al. 1993). Decomposition rates observed for leaf litter corresponded with the EPT index patterns observed during pretreatment, treatment, and recovery in C 54. During treatment, when the EPT index was lowest, decomposition rates (days to 5% of original litter mass remaining in litter bags) decreased. Decomposition rates subsequently increased, along with EPT values during recovery (Fig. 3). Previous work in these streams has also demonstrated strong linkages between invertebrate shredders, leaf litter decomposition, and export of FPOM. Leaf shredding invertebrates generally have low assimilation efficiencies and increase the rate at which coarse particulate organic matter (CPOM) is converted to FPOM, which is more easily transported than CPOM. Treatment greatly reduced shredder biomass and production (Lugthart and Wallace 1992, Chung et al. 1993) in the treatment stream, and resulted in lower rates of conversion of CPOM to FPOM. Hence, concentration and export of suspended FPOM (= organic seston) were much lower during treatment than either pretreatment or reference streams (Wallace et al. 1991a). As observed above for the EPT index and leaf litter decomposition rates, the pattern of FPOM concentrations observed at the flume at the base of the stream and the EPT index were quite similar (Fig. 4). FPOM concentrations in stream water decreased with initial treatment, remained low throughout the 3-yr treatment period, and subsequently increased during recovery. This pattern was accurately reflected by changes in the EPT index. Trends similar to those observed for the EPT index were evident for NCBIs based on both habitat-weighted TABLE 6. Average number of EPT taxa (± 1 so) collected during various periods in the treatment and reference streams. Periods refer to insecticide treatment of the treatment stream and are as follows: pretreatment = October 1984 to November 1985; treatment = December 1985 to December 1988; recovery year 1 = January to December 1989; and recovery year 2 = January to December Period Pretreatment Treatment Recovery year 1 Recovery year 2 Treatment stream! Reference stream:): 18.0 ±2.35" 19.0 ± ± 2.64 k 19.0 ± 1.53* 10.8 ± 5.50»- c 20.5 ± 2.66* 16.5 ±1.83" 18.3 ±1.86 t Same superscript letters for the treatment stream indicate no significant difference among periods for the treatment stream; Kruskal-Wallis ANOVA on ranks and Dunn's test, P > $ Asterisks indicate significantly higher EPT abundances for a given period among streams; Mann-Whitney rank sum test, P < Years r-*- \. 1, 1.,i, i, ^-, Elapsed Days FIG. 2. Number of EPT taxa collected on various sampling dates in C 54 (treatment stream) and C 55 (reference stream). S = start of seasonal insecticide treatments of C 54 and E = end of treatments.

8 February 1996 BIOTIC INDICES AND ECOSYSTEM PROCESSES , -*- EPT taxa Dred maple r Elapsed Days FIG. 3. Number of EPT taxa present on various sampling dates superimposed on leaf litter decomposition rates for red maple and rhododendron in the treatment stream (C 54). Note much longer time periods required'for litter decomposition during the treatment period (=days 430 to 1600). Litter decomposition rates are from Chung (1992) and Chung et al. (1993) and annual values are plotted on the midpoint of the day during each year of measurement. abundance and biomass. For example, EPTs and NCBIs were correlated in the treatment stream for both the inverse (1/NCBI value) of habitat-weighted biomass (r = 0.78, P < 0.001, df = 42) and abundance (r = 0.69, P < 0.001, df = 42) vs. EPTs. DISCUSSION Throughout the study period, both NCBIs and EPT values tracked the disturbance regime of the treated stream, whereas these indices differed little during the 6-yr period in the reference stream. At least for pulsetype disturbances such as toxicants (sensu Bender et al. 1984), which are limited in temporal scale and do not result in chronic, long-term, physical alteration of the system, both indices were effective mechanisms for assessing changes in community structure. Both the NCBIs and EPT reflected strong changes in the treatment stream during treatment relative to both pretreatment or the reference stream. Both indices also were effective indicators of improved biotic conditions during the first and second years of recovery of the treatment stream (Tables 3, 4, and 5). Further, these indices agreed closely with observed patterns of secondary production in the treatment stream, which decreased by 62% from the pretreatment year to the initial treatment year, and subsequently increased during the two recovery years. EPT taxa contribution to total secondary production in the treatment stream decreased from 40.8% in the pretreatment year to 2.5% during treatment, then increased to 26 and 29% during 1st- and 2nd-yr recovery, respectively (Lugthart and Wallace 1992, Whiles and Wallace 1995). These results underscore the potential for such indices in effectively detecting and monitoring stream ecosystem changes during and following our chemical-induced disturbance. Our study demonstrates that invertebrate communities inhabiting moss-covered rockface substrates appear to be excellent bellwethers of toxicant-induced environmental changes in these high-gradient southern Appalachian streams. In general, rockface substrates had lower (better) NCBIs than mixed substrates during pretreatment and recovery of C 54. However, rockface habitats exhibited greater proportional increases in NCBIs during treatment than mixed substrates (Table 2). These results indicate that communities inhabiting rockface habitats have considerable sensitivity to toxicants, demonstrating the value of incorporating these habitats, when present, into environmental monitoring programs. However, these effects may vary with the nature of the disturbance. Drought also appears to have a greater negative impact on rockface than mixed substrates (Lugthart and Wallace 1992). In contrast, rockface habitats were least affected of all habitat types by increased sediments resulting from clear-cut logging and road building in another Coweeta stream (Gurtz and Wallace 1984). Others have also noted differences among pool and riffle habitats of streams (Kerans et al. 1992). Thus, different conclusions may be generated if only one habitat is sampled. Lenat and Barbour (1994) reported EPT taxa richness to be the single most reliable measurement used by North Carolina biologists, and that this index is very sensitive to changes in water quality, while less variable than other measures among years of contrasting discharges. Our results agree, as the reference stream (C 55) had an average EPT (±95% ci) of 19.1 (±0.69) over the 32 sampling dates, despite record dry and wet years. Thus, our data indicate that the EPT, while sensitive to chemical-induced disturbances, is relatively insensitive to natural disturbances such as extreme discharges in these headwater streams. Further, the EPT index declined sharply with initial treatment of C 54, remained low throughout treatment, and displayed a strong increase during the last 6 mo of Ist-yr recovery (Fig. 2). The EPT values, as well as the NCBIs, also C 54 seston --* -- C 54 EPT Elapsed Day FIG. 4. Number of EPT taxa present on various sampling dates superimposed on continuously measured seston concentrations (mg ash-free dry mass/l) in the treatment stream (C 54). Note generally lower seston concentrations during the treatment period (=days 430 to 1600). Seston data are from Wallace et al. (1991a) and J. B. Wallace, unpublished data.

9 148 J. BRUCE WALLACE ET AL. Ecological Applications Vol. 6, No. I closely corresponded with several ecosystem-level changes observed during treatment and recovery of C 54. These included declines and subsequent increases in leaf litter processing rates, FPOM concentrations and export, and macroinvertebrate secondary production. At least for southern Appalachian headwater streams, we cannot agree with the conclusions of Suter (1993) that such indices have no meaning, no diagnostic power, and no regulatory applications. In our streams, they had real value for assessing a wide array of potential ecosystem-level phenonoma. The high conformity between EPT indices and leaf decomposition rates is undoubtedly related to the predominant animals feeding directly on leaf litter (shredders). A number of Plecoptera, Trichoptera, and Tipula spp. (Diptera) constitute the majority of the shredder functional group in these streams. Biomass and production of Tipula spp., as well as plecopteran and trichopteran shredders (Chung et al. 1993), corresponded closely with the EPT index during treatment and recovery. Shredders constitute =30% of total secondary production in these streams (Lugthart and Wallace 1992). However, leaf-shredding trichopterans and plecopterans accounted for >74% of shredder secondary production in litterbags during 2nd-yr recovery of C 54, and Tipula another 25.5%. Two species of Lepidostoma (Trichoptera: Lepidostomatidae) dominated shredder production during recovery (Chung et al. 1993) and they have been shown to play an important role in restoration of leaf litter processing in this stream (Whiles et al. 1993). Thus, the high inverse relationship between the EPT index and leaf litter decomposition rates is not surprising. The relationship between the EPT index and FPOM concentrations and export is also undoubtedly linked through the shredder functional group. A strong linkage between leaf-shredding invertebrates and FPOM concentrations and export in stream water has previously been demonstrated for these streams. Leaf-shredding invertebrates increase the rate at which CPOM is converted to FPOM (Wallace et al. 1982, Cuffney et al. 1990, Wallace et al. 1991a). Other indirect linkages between shredders and FPOM concentrations and export likely exist as well. During treatment, unprocessed leaf litter increased in the treatment stream, compared to the reference (Wallace et al. 1995). Accumulated leaf litter retained FPOM and altered seasonal patterns of FPOM export in the treatment stream during treatment compared to untreated years or that observed for nearby reference streams (Wallace et al. 199 la). Our results demonstrate a relationship between degraded biotic integrity and ecosystem-level processes in headwater streams of forested regions. In our view, similar phenomena likely exist in other regions and biomes. For example, invertebrate grazers decrease algal standing crops in more autotrophic streams (e.g., Lamberti and Resh 1983, Lamberti and Moore 1984, Feminella et al. 1989). Treatment of a Japanese stream with an insecticide reduced macroinvertebrate populations, including grazers, and their removal was subsequently accompanied by a large increase in algal biomass (Yasuno et al. 1982). Feeding activities of grazers also influence FPOM export to downstream reaches in laboratory streams (Mulholland et al. 1983, Lamberti et al. 1989) as well as nutrient dynamics (Mulholland et al. 1991). Detritivores can increase the rate of both nutrient (Grimm 1988) and dissolved organic matter release (Meyer and O'Hop 1983). Microfiltering collectors can increase the rate at which FPOM is removed from stream seston (Morin et al. 1988) and macrofiltering insects can suppress drifting invertebrates in the water column (Georgian and Thorp 1992). Invertebrate predators can influence shredder abundances and leaflitter processing rates (Oberndorfer et al. 1984, Malmqvist 1993). The above examples, as well as many others, suggest important linkages between consumers and ecosystem level processes in streams. Hence, destruction of biotic integrity potentially influences many processes in addition to those documented by us in this study. In summary, both the NCBI and EPT indices tracked several system-level parameters, including secondary production. The EPT index is by far the easiest to use, from both the standpoint of time required for sample processing and ease of application. Expense and time involved with sampling, processing, and data analyses are important considerations in any biomonitoring program (see Karr 1991). The relatively simple and inexpensive EPT index displayed a remarkable ability to track secondary production, a far more laborious and expensive measure. Furthermore, most secondary production measurements are expressed on an annual basis, requiring measurements over a full year for a single value. Thus, biotic indices allow for much greater temporal and spatial replication. Likewise, the use of NCBIs based on biomass showed greater proportional changes from pretreatment to treatment and recovery periods than NCBIs based on abundances (Table 2). However, the usefulness of biomass NCBIs may remain limited because of the much greater effort and expense involved with measuring and weighing specimens. In addition to the NCBI, our data strongly support the inclusion of the EPT index in these southern Appalachian headwater streams for detection of both degradation and recovery of stream ecosystem processes following chemical-induced disturbance. In regions where taxa belonging to the three EPT orders are less abundant, or among streams of different sizes and community structure (e.g., Crunkilton and Duchrow 1991), the EPT index may be more difficult to apply. Others also have noted that the EPT index is often lower in headwater streams than larger downstream reaches, which may compromise its use for biological monitoring (Lenat and Barbour 1994, Stewart and Loar 1994). This suggests that between-stream EPT comparisons may need to be restricted to streams of

10 February 1996 BIOTIC INDICES AND ECOSYSTEM PROCESSES 149 similar size and elevation within a regional setting. In such situations, multiple approaches incorporating stream size, distributional patterns, and biological attributes of fish have been used (Miller et al. 1988). Even when streams of similar size and elevation are compared, the close agreement between EPT and the NCBI, as observed in our study, may not apply to all disturbances. For example, clear-cutting of a headwater catchment at Coweeta in 1977 resulted in increased insolation, which increased stream temperature (Swift 1983) and shifted, in the short term, the stream energy base from allochthonous sources to primary production (Webster et al. 1983). Immediately after clear-cutting, increases in EPT taxa occurred in the stream draining the clear-cut catchment, compared with that of a nearby reference stream draining a forested catchment. Several taxa normally confined to larger downstream reaches colonized the stream draining the clear-cut catchment without significant loss of headwater EPT taxa. With forest regrowth, the number of EPT taxa exhibited successive declines in 1983 and 1993 compared to that of the reference stream. In contrast, compared to the reference stream, the NCBI indicated lower biotic integrity immediately following clear-cutting, which subsequently improved during forest regrowth (M. Stone and J. B. Wallace, unpublished data). Thus, biotic indices may vary with the nature of disturbance, i.e., pulse vs. press (sensu Bender et al. 1984), demonstrating the pitfalls of relying solely on one index such as the EPT for all disturbances (see also Karr 1994). The extent to which biotic indices are linked with system-level processes remains to be shown for either (a) other types of anthropogenic disturbances, or (b) larger downstream reaches, where physical forces such as greater discharge may overwhelm the capacity of investigators to detect changes in processes. However, even in very large rivers, in view of some of the roles of consumers in studies cited above, as well as numerous other studies, these indices may indicate the potential for a pervasive influence of disturbance on community structure, secondary production, food webs, and food resource availability to higher aquatic organisms. This requires a blending of basic and applied research, an important future direction for lotic scientists. ACKNOWLEDGMENTS This research was supported by grants from the National Science Foundation and U.S. Forest Service. Dr. Wayne T. Swank, G. B. Cunningham, and other personnel of the Coweeta Hydrologic Laboratory of the U.S. Forest Service provided discharge data and assistance with various aspects of the project. We thank Drs. G. J. Lugthart, K. Chung, T. F. Cuffney, A. D. Huryn, F. Smith-Cuffney, and D. Imm for assistance with various aspects of the project. Field and laboratory assistance was provided by A. Lingle, S. Stroud, T. Pressley, G. Schurr, L. Houston, P. Egens, P. Vila, and B. Goldowitz. Constructive comments of referees and Roy Stein on the manuscript were valuable. This work would not have been possible without the efforts of the North Carolina Division of Environmental Management and the long-term efforts of D. Lenat and D. Penrose and others in constructing the North Carolina Biotic Index. Dr. S. R. Carpenter developed the program used for Randomized Intervention Analysis. LITERATURE CITED Abel, P. D Water pollution biology. Ellis Horwood, Chichester, UK. Bender, E. A., T. J. Case, and M. E. Gilpin Perturbation experiments in community ecology: theory and practice. Ecology 65:1-13. Cairns, J., and J. R. Pratt A history of biological monitoring using benthic macroinvertebrates. Pages in-d. M. Rosenberg and V. H. Resh, editors. Freshwater biomonitoring and benthic macroinvertebrates. Chapman and Hall, New York, New York, USA. Carpenter, S. R., T. M. Frost, D. Heisey, and T. Kratz Randomized intervention analyses and the interpretation of whole ecosystem experiments. Ecology 70: Charles, D. F., J. P. Smol, and D. R. Engstrom Paleolimnological approaches to biological monitoring. Pages in S. L. Loeb and A. Spacie, editors. Biological monitoring of aquatic systems. Lewis, Boca Raton, Florida, USA. Chung, K Recovery of macroinvertebrate fauna from insecticide-induced disturbance in small headwater streams. Dissertation. University of Georgia, Athens, Georgia, USA. Chung, K., J. B. Wallace, and J. W. Grubaugh The impact of insecticide treatment on abundance, biomass, and production of litterbag fauna in a headwater stream: a study of pretreatment, treatment, and recovery. Limnologica 28: Chutter, F. M An empirical biotic index of the quality of water in South African streams and rivers. Water Research 6: Crawford, J. K., and D. R. Lenat Effects of land use on the water quality and biota of three streams in the Piedmont province of North Carolina. Water-Resources Investigations Report , U.S. Geological Survey, Raleigh, North Carolina, USA. Crunkilton, R. L., and R. M. Duchrow Use of stream order and biological indices to assess water quality in the Osage and Black river basins of Missouri. Hydrobiologia 223: Cuffney, T. E, J. B. Wallace, and G. J. Lugthart Experimental evidence quantifying the role of benthic invertebrates in organic matter dynamics of headwater streams. Freshwater Biology 23: Cummins, K. W Bioassessment and analysis of functional organization of running water ecosystems. Pages in S. L. Loeb and A. Spacie, editors. Biological monitoring of aquatic systems. Lewis, Boca Raton, Florida, USA. Eaton, L. E., and D. R. Lenat Comparison of a rapid bioassessment method with North Carolina's qualitative macroinvertebrate collection method. Journal of the North American Benthological Society 10: Elliott, J. M Some methods for the statistical analysis of samples of benthic invertebrates. Second edition. Freshwater Biological Association Scientific Publication Number 25. Feminella, J. W, M. E. Power, and V. H. Resh Periphyton response to invertebrate grazing and riparian canopy in three northern California coastal streams. Freshwater Biology 22: Georgian, T., and J. H. Thorp Effects of microhabitat selection on feeding rates of net-spinning caddisfly larvae. Ecology 73: Grimm, N. B Role of macroinvertebrates in nitrogen dynamics of a desert stream. Ecology 69:

11 150 J. BRUCE WALLACE ET AL. Ecological Applications Vol. 6, No. 1 Gurtz, M. E., and J. B. Wallace Substrate-mediated response of stream invertebrates to disturbance. Ecology 65: Hilsenhoff, W. L Use of arthropods to evaluate water quality of streams. Technical Bulletin Number 100, Department of Natural Resources, Madison, Wisconsin, USA An improved biotic index of organic stream pollution. Great Lakes Entomologist 20: Huryn, A. D Secondary production of the macroinvertebrate community of a high-elevation stream in the southern Appalachian Mountains. Dissertation. University of Georgia, Athens, Georgia, USA. Hynes, H. B. N The biology of polluted waters. Liverpool University Press, Liverpool, UK Historical perspective and future direction of biological monitoring of aquatic systems. Pages in S. L. Loeb and A. Spacie, editors. Biological monitoring of aquatic systems. Lewis, Boca Raton, Florida, USA. Karr, J.R Biological integrity: a long-neglected aspect of water resource management. Ecological Applications 1: Defining and assessing ecological integrity: beyond water quality. Environmental Toxicology and Chemistry 12: Biological monitoring: challenges for the future. Pages in S. L. Loeb and A. Spacie, editors. Biological monitoring of aquatic systems. Lewis, Boca Raton, Florida, USA. Kerans, B. L., and J. R. Karr A benthic index of biotic integrity (B-IBI) for rivers of the Tennesee Valley. Ecological Applications 4: Kerans, B. L., J. R. Karr, and S. A. Ahlstedt Aquatic invertebrate assemblages: spatial and temporal differences among sampling protocols. Journal of the North American Benthological Society 11: Lamberti, G. A., S. V. Gregory, L. R. Ashkenas, A. D. Steinman, and C. D. Mclntire Productive capacity of periphyton as a determinant of plant-herbivore interactions in streams. Ecology 70: Lamberti, G. A., and J. W. Moore Aquatic insects as primary consumers. Pages in V. H. Resh and D. M. Rosenberg, editors. The ecology of aquatic insects. Praeger, New York, New York, USA. Lamberti, G. A., and V. H. Resh Stream periphyton and insect herbivores: an experimental study of grazing by a caddisfly population. Ecology 64: Lenat, D.R Water quality assessment of streams using a qualitative collection method for benthic macroinvertebrates. Journal of the North American Benthological Society 7: A biotic index for the southeastern United States: derivation and list of tolerance values, with criteria for assigning water-quality ratings. Journal of the North American Benthological Society 12: Lenat, D. R., and M. T. Barbour Using benthic macroinvertebrate community structure for rapid, cost-effective, water quality monitoring: rapid bioassessment. Pages in S. L. Loeb and A. Spacie, editors. Biological monitoring of aquatic systems. Lewis, Boca Raton, Florida, USA. Loeb, S. L., and A. Spacie, editors Biological monitoring of aquatic systems. Lewis, Boca Raton, Florida, USA. Lugthart, G. J., and J. B. Wallace Effects of disturbance on benthic functional structure and production in mountain streams. Journal of the North American Benthological Society 11: Lugthart, G. J., J. B. Wallace, and A. D. Huryn Secondary production of chironomid communities in insecticide-treated and untreated headwater streams. Freshwater Biology 24:417^127. Malmqvist, B Interactions in stream leaf packs: effects of a stonefly predator on detritivores and organic matter processing. Oikos 66: Meyer, J. L., and J. O'Hop Leaf-shredding insects as a source of dissolved organic carbon in headwater streams. American Midland Naturalist 109: Miller, D. L., P. M. Leonard, R. M. Hughes, J. R. Karr, P. B. Moyle, L. H. Schrader, B. A. Thompson, R. A. Daniels, K. D. Fausch, G. A. Fitshugh, J. R. Gammon, D. B. Haliwell, P. L. Angermeier, and D. J. Orth Regional applications of an index for biotic integrity for use in water resource management. Fisheries (Bethesda) 13: Morin, A., C. Back, A. Chalifour, J. Boisvert, and R. H. Peters Effect of black fly ingestion and assimilation on seston transport in a Quebec lake outlet. Canadian Journal of Fisheries and Aquatic Sciences 45: Mulholland, P. J., J. D. Newbold, J. W. Elwood, and C. L. Horn The effect of grazing intensity on phosphorus spiralling in autotrophic streams. Oecologia (Berlin) 53: Mulholland, P. J., A. D. Steinman, A. V. Palumbo, J. W. Elwood, and D. B. Kirschtel Role of nutrient cycling and herbivory in regulating periphyton communities in laboratory streams. Ecology 72: Oberndorfer, R. Y., J. V. McArthur, and J. R. Barnes The effect of invertebrate predators on leaf litter processing in an alpine stream. Ecology 65: Plafkin, J. L., M. T. Barbour, K. D. Porter, S. K. Gross, and R. M. Hughes Rapid bioassessment protocols for use in streams and rivers: benthic macroinvertebrates and fish. United States Environmental Protection Agency EPA/ 444/ Rosenberg, D. M., and V. H. Resh, editors Freshwater biomonitoring and benthic macroinvertebrates. Chapman and Hall, New York, New York, USA. Stewart, A. J., and J. M. Loar Spatial and temporal variation in biomonitoring data. Pages in S. L. Loeb and A. Spacie, editors. Biological monitoring of aquatic systems. Lewis, Boca Raton, Florida, USA. Suberkropp, K., and J. B. Wallace Aquatic hyphomycetes in insecticide-treated and untreated streams. Journal of the North American Benthological Society 11: Suter, G. W A critique of ecosystem health concepts and indexes. Environmental Toxicology and Chemistry 12: Swank, W. T, and D. A. Crossley, editors Forest hydrology and ecology at Coweeta. Ecological Studies Series Number 66. Springer-Verlag, New York, New York, USA. Swank, W. T, and J. B. Waide Characterization of baseline precipitation and stream chemistry and nutrient budgets for control watersheds. Pages in W. T. Swank and D. A. Crossley, editors. Forest hydrology and ecology at Coweeta. Ecological Studies Series Number 66. Springer-Verlag, New York, New York, USA. Swift, L. W., Jr Duration of stream temperature increases following forest cutting in the southern Appalachian Mountains. Pages in Proceedings of the International Symposium on Hydrometeorology, Denver, Colorado, USA, June American Water Resources Association, Bethesda, Maryland, USA. Wallace, J. B., T. F. Cuffney, J. R. Webster, G. J. Lugthart, K. Chung, and B. S. Goldowitz. 1991a. Export of fine organic particles from headwater streams: effects of season, extreme discharges, and invertebrate manipulation. Limnology and Oceanography 36: Wallace, J. B., A. D. Huryn, and G. J. Lugthart. 1991k Col-