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Fish Assemblages in the Upper Esopus Creek, NY: Current Status, Variability, and Controlling Factors
Barry P. Baldigo, Scott D. George, and Walter T. Keller

Northeastern Naturalist, Volume 22, Issue 2 (2015): 345–371

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Northeastern Naturalist Vol. 22, No. 2 B.P. Baldigo, S.D. George, and W.T. Keller 2015 345 2015 NORTHEASTERN NATURALIST 22(2):345–371 Fish Assemblages in the Upper Esopus Creek, NY: Current Status, Variability, and Controlling Factors Barry P. Baldigo1,*, Scott D. George1, and Walter T. Keller2 Abstract - The Upper Esopus Creek receives water diversions from a neighboring basin through the Shandaken Tunnel (the portal) from the Schoharie Reservoir. Although the portal is closed during floods, mean flows and turbidity of portal waters are generally greater than in Esopus Creek above their confluence. These conditions could potentially affect local fish assemblages, yet such effects have not been assessed in this highly regulated stream. We studied water quality, hydrology, temperature, and fish assemblages at 18 sites in the Upper Esopus Creek during 2009–2011 to characterize the effects of the portal input on resident-fish assemblages and to document the status of the fishery resource. In general, fish-community richness increased by 2–3 species at mainstem sites near the portal, and median density and biomass of fish communities at sites downstream of the portal were significantly lower than they were at sites upstream of the portal. Median densities of Salmo trutta (Brown Trout) and all trout species were significantly lower than at mainstem sites downstream from the portal—25.1 fish/0.1 ha and 148.9 fish/0.1 ha, respectively—than at mainstem sites upstream from the portal—68.8 fish/0.1 ha and 357.7 fish/0.1 ha, respectively— yet median biomass for Brown Trout and all trout did not differ between sites from both reaches. The median density of young-of-year Brown Trout at downstream sites (9.3 fish/0.1 ha) was significantly lower than at upstream sites (33.9 fish/0.1 ha). Waters from the portal appeared to adversely affect the density and biomass of young-of-year Brown Trout, but lower temperatures and increased flows also improved habitat quality for mature trout at downstream sites during summer. These findings, and those from companion studies, indicate that moderately turbid waters from the portal had few if any adverse impacts on trout populations and overall fish communities in the Upper Esopus Creek during this study. Introduction The Upper Esopus Creek is a historic trout fishery and recreational stream in the Catskill Mountains of southeastern New York state. The Shandaken Tunnel (the portal) delivers water that is usually cool during the summer from the Schoharie Reservoir to the Upper Esopus Creek (hereafter, the Esopus), and maintains a minimum flow in the Esopus as stipulated by state regulations (CCE 2007). On average, these releases are more turbid than the waters of the Esopus and have been the focus of controversy because of the perception that the turbidity of the portal’s inflow waters negatively affects water quality and trout populations in the upper basin (CCE 2007). Any stressors that adversely affect Salmo trutta L. (Brown Trout) and Oncorhynchus mykiss (Walbaum) (Rainbow Trout) populations might negatively 1US Geological Survey, New York Water Science Center, 425 Jordan Road, Troy, NY 12180. 2New York State Department of Environmental Conservation (Retired), 65561 State Highway 10, Stamford, NY 12167. *Corresponding author - bbaldigo@usgs.gov. Manuscript Editor: Rudolf G. Arndt Northeastern Naturalist 346 B.P. Baldigo, S.D. George, and W.T. Keller 2015 Vol. 22, No. 2 affect the local economy because these fisheries are important resources. Both the concentration of suspended sediment and the duration of turbid episodes can adversely affect the health of lotic species and thus aquatic communities (Newcombe and Macdonald 1991, Shaw and Richardson 2001). Turbidity has been linked with decreased visual acuity and growth rates in trout under laboratory conditions (Shaw and Richardson 2001; Sigler et al. 1984; Sweka and Hartman 2001a, b), reduced density and biomass of macroinvertebrates in the wild (Wagener and LaPerriere 1985), reduced biomass and primary productivity in periphyton (Bilotta and Brazier 2008, Quinn et al. 1992), and siltation-related habitat degradation in all 3 communities (Henley et al. 2000). Field and mesocosm studies, moreover, have found a range of neutral to negative impacts of turbidity on feeding, growth, density, and biomass of resident trout or their populations (Redding 1987, Sweka and Hartman 2001a, White and Harvey 2007). Because there are few data on fish communities and turbidity in the Esopus, the potential effects of turbidity on local fish communities and trout populations in this basin are largely speculative and subject to debate. Discharge of relatively cool water from the portal could also have beneficial effects on cold-water species. Water temperatures in parts of the Esopus frequently exceed upper thermal limits for growth (~19 °C) and survival (~25 °C) of Brown Trout (Hasnain et al. 2013, Wehrly et al. 2007) for long low-flow periods that occur during most summers (Ross 2012). The portal provides supplemental flows that can greatly exceed natural flows in the mainstem Esopus (at their ju ncture) by as much as 1 order of magnitude. The temperature of portal water often is below the upper lethal limits for Brown Trout during warm months, and may allow the species to survive, if not grow, in nearby reaches immediately downstream from the portal that otherwise might be uninhabitable. Several factors other than the portal input may affect water quality and impact stream ecosystems in the Upper Esopus Basin including (1) effluent discharges from sewage-treatment plants within the watershed; (2) runoff from a ski area; (3) withdrawals from streams for various purposes, including snow-making, water supply, and bottling; and (4) areas with high concentrations of septic-systems adjacent to tributaries. These factors, combined with water releases from the portal, subject local fish populations and communities to a range of thermal, hydrologic, and waste-water stresses in the upper basin. Unbiased contemporary data are needed to quantify the status of water quality, fish assemblages, and potential impairments in the Esopus. In 2009, the US Geological Survey (USGS), in collaboration with the New York State Department of Environmental Conservation (NYSDEC), New York City Department of Environmental Protection (NYCDEP), and Cornell Cooperative Extension of Ulster County (CCEUC), began a comprehensive assessment of the biological condition (fish, invertebrates, and diatoms) and water quality (turbidity and nutrients) in the Esopus. Primary objectives of this study were to assemble contemporary data to define the effects of discharges from the portal on resident fish assemblages and to foster better decision-making regarding the management of the fishery, water quality, and water quantity (allocated water from the Schoharie Northeastern Naturalist Vol. 22, No. 2 B.P. Baldigo, S.D. George, and W.T. Keller 2015 347 Reservoir to the Ashokan Reservoir) in the Esopus. Specific goals of this effort were to characterize the natural variability and its relationship with water quality for the biological communities, and to assess the potential effects of point sources of turbidity on fish assemblages. We gave special attention to trout populations because of their economic value and location at the top of the aquatic food chain, as well as the perception that the local fishery and other natural resources within the basin have declined due to the turbid waters delivered to the upper basin via the portal. In this paper, we describe the status of resident fish assemblages and trout populations at 18 sites in the Esopus that we sampled annually, with a few exceptions at selected sites, from June through August during 2009–2011. Study Area The Upper Esopus Creek Basin is located in the south-central Catskill Mountains of southeastern New York and follows a 67.3-km semi-circular course from its headwaters at Winnisook Lake, to its terminus at the Ashokan Reservoir (Fig. 1). Its 497-km2 watershed is contained entirely within the Catskill Park and drains some of the most rugged and mountainous terrain in the Park. Forested land comprises over 95% of the watershed, and commercial and residential development occupies the rest (CCE 2007). The Shandaken Portal and the Esopus confluence occur in the village of Shandaken just west of Phoenicia (Fig. 1). As of about 2005, water discharged from the portal had the highest median turbidity (8.8 Nephelometric Turbidity Units [NTU]) of any tributary previously assessed in the upper basin; however, its contribution to the total annual sediment load of the Esopus was not fully defined (CCE Figure 1. Location of 20 sites on the Upper Esopus Creek surveyed 2009–2011. Northeastern Naturalist 348 B.P. Baldigo, S.D. George, and W.T. Keller 2015 Vol. 22, No. 2 2007). Channels for several Esopus tributaries cut into clay-rich till and also contribute a significant portion of the total annual sediment load during rain and snow-melt events (CCE 2007). The portal can discharge more than 2 million m3 of water per day into the Esopus when at maximum output. Since 1977, water releases have been regulated by Part 670 of Title 6 of New York Compilation of Codes, Rules and Regulations (NYCRR; CCE 2007). In 2006, a NY State Pollutant Discharge Elimination System (SPDES) permit placed additional limits on flow, turbidity, and temperature in order to protect the health of aquatic biota in the Esopus (CCE 2007). Methods We collected fisheries and water-quality data at 18 sites in 2009 and 2010 and at 16 sites in 2011 (Fig. 1). Of the 18 sites, 10 were on tributaries with drainage areas of 10.3–83.9 km2, 1 (esop0) is a small headwater mainstem site of 30.3 km2, and 7 have drainages larger than 100 km2 and are located on the mainstem of the Upper Esopus (Table 1). Elevations of sites range from 189 m to 455 m (Table 1). At the time, there were no natural or man-made barriers to fish passage in the study area. We distributed mainstem sites evenly above and below the portal in an attempt to distinguish differences in fish communities caused by the portal from those that occurred naturally due to normal shifts in stream conditions between successively larger stream sites. In general, we chose study sites (reaches) so that they encompassed 2 or more complete geomorphic channel-unit sequences (Fitzpatrick et al. 1998, Meador et al. 2003, Simonson et al. 1994). Most study reaches were 20–35 mean-channel widths long (maximum of 100 m). At each reach, we collected fish by electrofishing up to a 100-m-long section using a backpack shocker (Smith-Root Model 12B) and 3 or 4 netters. After blocking each reach with seines, quantitative surveys consisted of 3 electrofishing passes through each reach; sample-area dimensions were measured on site. At sites with narrow channels (less than 15 m), we placed the blocking seines completely across the channel at the upstream and downstream ends of the study section. At sites with wider channels, we completed 3 replicate surveys in relatively small near-shore subreaches. At each subreach, we affixed 1 blocking seine to the bank, stretched it perpendicular to the bank, and attached it to a rock or rebar 6–8 m from shore. We placed a second 25-m seine oriented downstream and parallel to shore, attached it to a second rock or rebar, and placed a third seine between the second rebar and shore. Hence, each subreach was 6–8 m wide x 25 m long and was blocked by seines on 3 sides and by shoreline on the fourth. We processed fish collected from each sub-reach and pass separately. We identified each fish to species and recorded the lengths and weights of all individuals longer than 150 mm. For very abundant species (some minnows), which were usually shorter than 150 mm, we measured subsamples and obtained lengths and weights from 40–50 individuals; thereafter, we recorded total weights and counts by species in batches of 10–50. We returned all fish to the stream after processing. We characterized several key habitat features within each study reach and subreach (see Table A in Supplemental File 1, available online at http://www.eaglehill. Northeastern Naturalist Vol. 22, No. 2 B.P. Baldigo, S.D. George, and W.T. Keller 2015 349 us/NENAonline/suppl-files/n22-2-N1280-Baldigo-s1, and for BioOne subscribers, at http://dx.doi.org/10.1656/N1280.s1). We recorded total reach length and 10 measurements of reach width, and employed a modified point and transect method (Fitzpatrick et al. 1998) to measure depth and velocity and to estimate dominant substrate-size categories at 3 points—center, 25%, and 75% of crossection—along each of 10 evenly spaced transects. We used the total length and mean width to calculate sample areas, and included all measurements in mean depth and velocity calculations. We determined the dominant and subdominant particle sizes from the frequency of each category. Data for mean and median water-year (1 October–30 September) temperature, discharge, suspended-sediment concentration, and turbidity, along with annual suspended-sediment loads at all study sites, were from McHale and Siemion (2014) and USGS (2014). We determined standard metrics for entire fish communities and selected species populations for each survey (each site and year that they were sampled). We used the number of fish captured during each pass to estimate density and biomass (and Table 1. Site names, codes, and USGS site numbers for sites surveyed in the Upper Esopus Creek, 2009–2011. Elev. = elevation. Drainage Site USGS site Latitude Longitude area Elev. Stream and site name code number (°N) (°W) (km2) (m) Tributary sites Fox Hollow fox 01362199 42.1161111 74.380556 10.3 309 Peck Hollow peck 01362215 42.1255556 74.376389 12.3 351 Broadstreet Hollow broad 01362232 42.1125556 74.358694 23.7 296 Bushnellsville Creek bush 01362197 42.1247222 74.401139 29.5 336 Esopus Creek at OlivereaA esop0 0136219203 42.0525000 74.456222 30.3 455 Birch Creek birch 013621955 42.1089787 74.451818 32.4 377 Little Beaverkill lbeav 01362497 42.0195364 74.266258 42.7 205 Woodland Valley Creek wood 0136230002 42.0797222 74.334583 53.4 268 Beaverkill beav 01362487 42.0467580 74.276814 64.7 213 Stony Clove Creek at Chichester stoc0 01362370 42.1020278 74.310889 80.0 292 Stony Clove Creek at Phoenicia stoc1 01362398 42.0830556 74.315833 83.9 245 Upstream mainstem sites Esopus Creek at Big Indian esop2 0136219565 42.1041667 74.435833 111.9 355 Esopus Creek at Shandaken esop3 0136219710 42.1194444 74.397500 152.0 317 Esopus Creek at Allaben esop3a 01362200 42.1170341 74.380149 165.0 305 Downstream mainstem sites Esopus Creek downstream esop3b 0136223005 42.1133333 74.361889 181.0 287 of portal Esopus Creek upstream of esop4 01362250 42.0925000 74.335972 215.7 268 Phoenicia Esopus Creek at Phoenicia esop4a 01362405 42.0819444 74.312028 357.4 238 Esopus Creek downstream esop4b 01362420 42.0636111 74.306389 365.2 225 of Phoenicia Esopus Creek at Mt Tremper esop5 01362430 42.0468889 74.280000 373.0 207 Esopus Creek at Boiceville esop6 01362500 42.0142588 74.270425 497.3 189 ASite esop0 was also treated as an upstream mainstem site for selected analyses. Northeastern Naturalist 350 B.P. Baldigo, S.D. George, and W.T. Keller 2015 Vol. 22, No. 2 95% confidence intervals [CI]) for the entire fish community and for each species population with the Moran-Zippin method of proportional reduction (Van Deventer and Platts 1985, Zippin 1958). These values were divided by the total area sampled at each site to estimate the number or the biomass (g) of fish in the local community or species population per unit area. Two components of ecosystem diversity, breadth or size (total species richness) and heterogeneity (Simpson’s diversity index), were also generated for each survey (Simpson 1949, Whittaker 1975). Richness (S) is the number of different fish species collected in the sampled area at each site. Simpson’s diversity index is commonly used to characterize community biodiversity, and it employs the number of species present and the relative abundance of each species to calculate a metric that ranges from 0 to 1. When calculated as 1 - D, zero indicates no diversity (e.g., 0–1 species), and values close to 1 indicate a large number of species and proportionally similar numbers of each. We described the status of resident-fish populations and communities, and spatial and temporal variations in key metrics graphically. We evaluated the potential effects of the portal on fish assemblages with 4 related methods. Longitudinal and annual trends in metrics (total community density, biomass, richness, and diversity, as well as density and biomass for trout populations) were evaluated through graphical analyses. The significance of changes or differences in metrics between or among individual sites and groups of sites within (a) tributaries, (b) upstream of the portal, and (c) downstream of the portal were evaluated using the upper and lower 95% CIs (which equal 1.96 times the standard errors), and parametric analysis of variance (ANOVA) if variance and normality assumptions were met, or non-parametric Kruskal-Wallis tests if data were not normally distributed. We considered differences in means and medians to be significant when the CIs did not overlap (Cumming et al. 2007) or when P-values for appropriate statistical tests were ≤0.05. We completed additional analyses of spatial patterns in fish-community composition and classifications (groupings of sites with similar assemblages) through nonmetric multidimensional scaling (MDS) ordination of taxon-density data that were square-root transformed (Kruskal 1964, Shepard 1962). MDS ordination generates an arrangement of samples in “species space” according to the non-parametric ranks of their Bray-Curtis similarities. We employed a similarity profile test (SIMPROF) to identify statistically significant (a priori unstructured) clusters within the larger dendrogram (Clarke and Warwick 2001). The significant groups were labeled as such on a non-metric MDS ordination to visually express the relationships among dissimilar groups. Subsequently, we used a similarity-percentage (SIMPER) analysis to determine the contribution of individual species to the overall dissimilarity between groups of samples (Clarke and Warwick 2001). The SIMPER breaks down the original Bray-Curtis similarities between samples into percent contributions of each species to those sample similarities, and identifies the species that are most responsible for defining site and group dissimilarities. Finally, we used an analysis of similarities (ANOSIM) to test a priori hypotheses that species assemblages in various reaches (e.g., sites from tributary, upstream, or downstream reaches) did not differ significantly (Clarke and Warwick 2001). Northeastern Naturalist Vol. 22, No. 2 B.P. Baldigo, S.D. George, and W.T. Keller 2015 351 Results Site names and the codes we used for our analyses and to report our results are provided in Table 1. Study sites on the mainstem of the Esopus upstream from the portal are collectively referred to as upstream sites and those downstream from the portal are referred to as downstream sites in the remainder of this manuscript. Fish communities Species richness ranged from as low as 4 species at 3 of the small drainage-area (DA) tributaries (fox, peck, and bush) during 1 or 2 years, to a maximum of 15 species at the largest DA site (esop6) during 2009 and 2010 (Tables 1, 2; Fig. 2A). Total community richness was strongly related to the drainage area of study sites (Fig. 3). Mean richness at all tributary sites (DA < 100 km2) over all years was 7.0 species, whereas mean richness at mainstem sites (DA > 100 km2) averaged 9.2 species upstream from the portal and 11.1 species downstream from the portal (Table 3). Median richness differed significantly among tributary, upstream, and downstream sites (Table 3). With one exception (birch during 2010), richness at the 6 tributary sites with DAs less than 40 km2 ranged from 4 to 7 species. Richness reached an asymptote at about 10 species as DA increased above 40 km2. Influx of nonresident species from Schoharie Reservoir through the portal may have increased richness at 2 sites closest to the portal. Richness at esop6 was larger than anticipated and inconsistent with the asymptotic pattern—a finding probably related to the influx of several species from the nearby Ashokan Reservoir. Mean richness across all sites decreased slightly during each year of the study, yet median richness was relatively consistent at 8.0, 7.5, and 8.0 species in 2009, 2010, and 2011, respectively, and did not differ between years. Table 2. Fish-community metrics from all sites surveyed in the Upper Esopus Creek during 2009–2011. Density Biomass Index of diversity Site code Richness (S) (fish/0.1ha) (g/0.1ha) (1 - D) Evenness (J') 2009 surveys fox 4 1707 13,517 0.35 0.48 peck 5 1396 9180 0.56 0.67 broad 7 881 4837 0.71 0.71 bush 4 2485 14,451 0.52 0.66 birch 7 1525 12,799 0.59 0.58 lbeav 9 657 4830 0.67 0.63 wood 7 2964 15,336 0.79 0.87 beav 11 1985 9891 0.72 0.65 stoc0 10 1790 10,749 0.74 0.66 stoc1 8 1440 7192 0.77 0.78 esop0 6 3474 9372 0.40 0.47 esop2 9 1517 40,843 0.79 0.80 esop3 8 2655 14,984 0.76 0.79 esop3a 10 2428 8307 0.80 0.79 esop3b 13 1260 3308 0.79 0.72 esop4 8 879 8743 0.72 0.72 esop4a 10 821 4696 0.81 0.79 esop6 15 1168 4993 0.69 0.62 Northeastern Naturalist 352 B.P. Baldigo, S.D. George, and W.T. Keller 2015 Vol. 22, No. 2 Community diversity, estimated using 1 - D, generally reflected the same patterns observed for species richness (Fig. 2B). Diversity scores ranged from 0.15 at bush (DA = 29.5 km2) in 2011 to 0.83 at esop6 (DA = 497.3 km2) in 2011 (Table 2). Mean diversity scores for all 3 years were lower at all tributary sites (0.56) than they were at all downstream (0.62) and upstream (0.74) mainstem sites (Table 3). Like richness, mean diversity increased between tributary and mainstem sites, but unlike richness, the median values at upstream sites (0.76) did not differ significantly from median values at downstream sites (0.79). Furthermore, the average diversity at 3 sites (esop2, esop3, and esop3a) immediately upstream from the portal Table 2, continued. Density Biomass Index of diversity Site code Richness (S) (fish/0.1ha) (g/0.1ha) (1 - D) Evenness (J') 2010 surveys fox 6 1567 147,06 0.66 0.68 peck 4 792 5638 0.57 0.77 broad 7 922 4248 0.77 0.81 bush 7 2375 9028 0.45 0.47 birch 9 1417 16,445 0.62 0.57 lbeav 10 1118 4943 0.77 0.73 wood 7 1215 7544 0.75 0.82 beav 9 2902 4533 0.69 0.63 stoc0 6 834 12,526 0.66 0.71 stoc1 7 817 3197 0.72 0.77 esop0 6 2219 11,032 0.22 0.30 esop2 9 1496 14,339 0.71 0.71 esop3 8 2436 11,205 0.70 0.69 esop3a 10 1301 5703 0.67 0.64 esop3b 11 533 1106 0.80 0.73 esop4 7 731 14,125 0.70 0.77 esop4a 9 2432 6854 0.79 0.77 esop6 15 850 2447 0.82 0.75 2011 surveys fox 4 516 3758 0.40 0.52 peck 5 703 4732 0.32 0.41 broad 6 622 4304 0.53 0.59 bush 7 3483 10,247 0.15 0.19 birch 6 1115 9554 0.53 0.60 lbeav na na na na na wood 7 497 2216 0.66 0.64 beav 10 2440 4643 0.50 0.45 stoc0 8 1029 4376 0.18 0.23 stoc1 8 438 2498 0.58 0.62 esop0 6 785 2801 0.48 0.48 esop2 8 1103 13,007 0.75 0.77 esop3 10 2570 9690 0.76 0.72 esop3a 11 1020 4222 0.76 0.73 esop3b 13 1072 2439 0.80 0.68 esop4 8 693 16,536 0.72 0.76 esop4a na na na na na esop6 13 680 3175 0.83 0.75 Northeastern Naturalist Vol. 22, No. 2 B.P. Baldigo, S.D. George, and W.T. Keller 2015 353 was comparable to the 4 downstream sites (Fig. 2B). Average and median speciesdiversity values at all Esopus sites were 0.59 and 0.72 in 2009, 0.67 and 0.70 in 2010, and 0.56 and 0.56 in 2011, respectively; medians did not differ significantly among years (Table 3). Figure 2. Mean richness (A), diversity (B), density (C), and biomass (D) and corresponding 95% confidence intervals (CIs) for resident-fish communities at all sites surveyed in the Upper Esopus Creek, 2009–2011. The locations of the sites (i.e., in a tributary and in the mainstem, either upstream or downstream from the portal) are denoted below the site identification codes; site esop0 functions as a tributary and a mainstem site for selected analyses. Northeastern Naturalist 354 B.P. Baldigo, S.D. George, and W.T. Keller 2015 Vol. 22, No. 2 Figure 3. Relationship between community richness and drainage area, and 95% confidence intervals (CIs) for all sites surveyed in the Upper Esopus Cree k, 2009–2011. Table 3. Mean and median fish-community and trout-population metrics grouped by site type and survey year and the significance (P-values) for the differences in medians as determined by Kruskall- Wallis non-parametric tests. Significantly different group-medians are denoted by non-similar letters below each median. YOY = young-of-year, Trib = tributary sites, US = sites upstream from portal, DS = sites downstream from portal. Site type Survey year Metric P-value Trib US DS P-value 2009 2010 2011 Community richness mean 7.0 9.2 11.1 8.4 8.2 8.1 median less than 0.001 7.0 9.0 11.0 0.9243 8.0 7.5 8.0 similar a b c a a a Community diversity (D) mean 0.56 0.74 0.62 0.59 0.67 0.56 median less than 0.001 0.59 0.76 0.79 0.3487 0.72 0.70 0.56 similar a b b a a a Community density (fish/0.1 ha) mean 1503 1836 1011 1724 1442 1173 median 0.0313 1306 1517 850 0.0253 1521 1258 903 similar ab a b a ab b Community biomass (g/0.1 ha) mean 7973 13,589 6220 11,002 8312 6137 median 0.0508 7368 11,205 4696 0.0310 9276 7199 4340 similar a b a a ab b All trout density (fish/0.1 ha) mean 165.0 319.0 158.0 280.6 205.2 72.1 median 0.0697 138.6 357.7 148.9 0.0001 264.0 183.3 45.9 similar a b a a a b Northeastern Naturalist Vol. 22, No. 2 B.P. Baldigo, S.D. George, and W.T. Keller 2015 355 Table 3, continued. Site type Survey year Metric P-value Trib US DS P-value 2009 2010 2011 All trout biomass (g/0.1 ha) mean 2621 3953 2326 3629 2926 1689 median 0.1791 2031 2804 3006 0.0037 3012 2031 1972 similar a a a a ab b Brown Trout density (fish/0.1 ha) mean 91.1 101.7 24.9 137.7 65.9 27.6 median 0.0069 64.0 68.8 25.1 0.0001 138.0 54.6 21.3 similar a a b a b b Brown Trout biomass (g/0.1 ha) mean 2124 2941 1863 3175 2314 1007 median 0.2395 1399 2747 857 0.0014 2742 1399 581 similar a a a a ab b Mature Brown Trout density (fish/0.1 ha) mean 41.5 21.7 13.2 42.2 41.8 9.9 median 0.0112 36.4 22.2 11.8 0.0001 39.0 36.9 10.7 similar a ab b a a b YOY Brown Trout density (fish/0.1 ha) mean 49.6 80.0 11.7 95.5 24.1 17.7 median 0.0157 20.3 33.9 9.3 0.0003 81.6 9.5 8.2 similar ab a b a b b Percent YOY Brown Trout (%) mean 38.0 75.2 36.9 62.6 28.9 40.8 median 0.007 38.0 83.0 30.0 0.0053 65.9 31.2 34.7 similar a b a a b b Rainbow Trout density (fish/0.1 ha) mean 71.3 213.3 132.5 140.4 135.8 43.1 median 0.0254 42.8 210.8 127.3 0.0075 114.9 127.8 27.5 similar a b ab a a b Rainbow Trout biomass (g/0.1 ha) mean 475.7 975.0 461.9 432.5 595.2 661.1 median 0.5264 414.8 815.3 257.0 0.5769 372.3 507.8 376.7 similar a a a a a a Mature Rainbow Trout density (fish/0.1 ha) mean 13.8 18.2 5.1 6.4 18.5 13.4 median 0.0309 12.2 16.6 2.0 0.0525 2.7 21.7 10.4 similar ab a b a b ab YOY Rainbow Trout density (fish/0.1 ha) mean 57.5 195.2 127.4 134.0 117.3 29.8 median 0.0137 26.5 193.4 125.7 0.0030 105.3 111.8 13.9 similar a b b a a b Percent YOY Rainbow Trout (%) mean 62.1 76.7 92.6 77.4 80.2 53.6 median less than 0.001 71.4 92.1 15.1 0.0083 94.4 86.4 50.0 similar a ab b a a b Northeastern Naturalist 356 B.P. Baldigo, S.D. George, and W.T. Keller 2015 Vol. 22, No. 2 The spatial trend in density of fish communities was not as strongly related to DA as were richness and diversity (Fig. 2C). Total density ranged from 438 fish/0.1 ha at stoc1 in 2011 to 3483 fish/0.1 ha at bush in 2011 (Table 2). Mean density at all tributary sites (1503 fish/0.1 ha) differed little from that at upstream (1836 fish/0.1 ha) and downstream (1011 fish/0.1 ha) sites; however, the median density at downstream sites (850 fish/0.1 ha) was significantly lower than the median density at upstream sites (1517 fish/0.1 ha) (Table 3). Total community density declined at many sites (individually) and overall throughout the 3-year study. Mean and median density for all surveys at all sites ranged from 1724 and 1521 fish/0.1 ha in 2009, 1142 fish/0.1 ha and 1258 fish/0.1 ha in 2010, and 1173 fish/0.1 ha and 903 fish/0.1 ha in 2011, respectively; median density at all sites was significantly lower during 2011 than during 2009. Community density decreased from 2009 to 2011 in about 80% of the tributaries and half of the mainstem sites (see Fig. B in Supplemental File 1, available online at https://www.eaglehill.us/NENAonline/suppl-files/n22- 2-N1280-Baldigo-s1, and for BioOne subscribers, at http://dx.doi.org/10.1656/ N1280.s1). Based on 95% CIs, the 2011 estimates were significantly lower than in 2009 at most sites, and densities for 2010 were generally intermediate and not different from the 2009 estimates. Community biomass (Fig. 2D), like density, was not strongly related to DA. Biomass ranged from 1106 g/0.1 ha at esop3b in 2010 to 40,843 g/0.1 ha at esop2 in 2009 (Table 2). Biomass averages and medians (g/0.1 ha) were, respectively: 7973 and 7368 in tributaries, 13,589 and 11,205 at upstream sites, and 6220 and 4696 at downstream sites. Estimates for median biomass at downstream sites were significantly lower than that for upstream sites (Table 3). Total biomass averaged about 9019 g/0.1 ha at sites esop3 and esop3a, immediately above the portal, and 7710 g/0.1 ha at sites esop3b and esop4, immediately below the portal. The lowest mean biomass for any site over most years was found at site esop3b, just downstream of the portal (Table 2; see also Fig. B in Supplemental File 1, available online at https://www.eaglehill.us/NENAonline/suppl-files/n22-2-N1280-Baldigos1, and for BioOne subscribers, at http://dx.doi.org/10.1656/N1280.s1). The low mean biomass (across all years) and the narrow CIs (±1255 g/0.1 ha) at this site (Fig. 2D) suggest that community biomass was significantly lower than at the next 2 upstream and next 2 downstream sites. Community biomass declined at many sites throughout the study period: it averaged 11,002 g/0.1 ha, 8312 g/0.1 ha, and 6137 g/0.1 ha for all surveys in 2009, 2010, and 2011, respectively. However, median biomass at upstream and downstream sites did not differ significantly (Table 3; see also Fig. B in Supplemental File 1, available online at http:// www.eaglehill.us/NENAonline/suppl-files/n22-2-N1280-Baldigo-s1, and for BioOne subscribers, at http://dx.doi.org/10.1656/N1280.s1). Biomass estimates were lowest at 11 sites during 2011. Total biomass decreased from 2009 to 2011 in all of the tributaries, and in about 80% of the mainstem sites; these declines were significant between 2009 and 2011, and data for 2010 were generally intermediate or not different from 2009 or 2011. Northeastern Naturalist Vol. 22, No. 2 B.P. Baldigo, S.D. George, and W.T. Keller 2015 357 Species populations Changes in the distribution of several fish species accounted for most of the differences in community metrics across sites. Species distributions (richness) appeared to be influenced largely by stream size, which can be classified by DA. Small DA sites (tributaries with DAs of less than 40 km2) tended to be dominated by Cottus cognatus Richardson (Slimy Sculpin), trout, and Rhinichthys cataractae Valenciennes (Longnose Dace) (see Fig. C in Supplemental File 1, available online at https://www.eaglehill.us/NENAonline/suppl-files/n22-2-N1280-Baldigo-s1, and for BioOne subscribers, at http://dx.doi.org/10.1656/N1280.s1), which are characteristic species in cold-water communities. Few other species contributed strongly to community density (see Table D in Supplemental File 1, available online at https://www.eaglehill.us/NENAonline/suppl-files/n22-2-N1280-Baldigo-s1, and for BioOne subscribers, at http://dx.doi.org/10.1656/N1280.s1) or biomass at small DA sites (see Table E in Supplemental File 1, available online at http://www. eaglehill.us/NENAonline/suppl-files/n22-2-N1280-Baldigo-s1, and for BioOne subscribers, at http://dx.doi.org/10.1656/N1280.s1). Trout became more sporadically distributed and Slimy Sculpin was partially replaced by minnow and sucker species at tributary sites with moderately sized (40 km2–100 km2) DAs (see Fig. C in Supplemental File 1, available online at https://www.eaglehill.us/NENAonline/ suppl-files/n22-2-N1280-Baldigo-s1, and for BioOne subscribers, at http://dx.doi. org/10.1656/N1280.s1). With minor exceptions, species density and biomass at most mainstem sites with DAs >100 km2 generally decreased from upstream to downstream sites. Biomass of trout and Slimy Sculpin populations were very high at the 2 upper-most mainstem sites, esop2 and esop3, and declined at the sites immediately above and below the portal (at esop3a and esop3b). Seven species—Luxilus cornutus (Mitchill) (Common Shiner), Etheostoma olmstedi Storer (Tessellated Darter), Noturus insignis (Richardson) (Margined Madtom), Notemigonus crysoleucas Mitchill (Golden Shiner), Lepomis cyanellus Rafinesque (Green Sunfish), Micropterus salmoides Lacépède (Largemouth Bass), and Alosa pseudoharengus (Wilson) (Alewife)—were first encountered at these 2 sites, i.e., they were absent from all mainstem sites upstream of esop3a (see Table D in Supplemental File 1, available online at https://www.eaglehill.us/NENAonline/suppl-files/n22-2-N1280-Baldigos1, and for BioOne subscribers, at http://dx.doi.org/10.1656/N1280.s1). Species populations at esop4 also did not reflect the expected community composition given its drainage area. Trout and Slimy Sculpin density and biomass increased markedly and density and biomass of minnow species declined at esop4 relative to the 2 prior upstream sites (see Fig. C in Supplemental File 1, available online at http://www. eaglehill.us/NENAonline/suppl-files/n22-2-N1280-Baldigo-s1, and for BioOne subscribers, at http://dx.doi.org/10.1656/N1280.s1). Increases in water velocity and decreases in temperature (related to inputs from the portal) may be partially responsible for observed community changes. Species diversity and richness increased and additional centrarchids, as well as other minnow species, were present at the 2 lower mainstem sites, esop4a and esop6. Four species—Micropterus dolomieu Lacépède Northeastern Naturalist 358 B.P. Baldigo, S.D. George, and W.T. Keller 2015 Vol. 22, No. 2 (Smallmouth Bass), Ambloplites rupestris (Rafinesque) (Rock Bass), Semotilus corporalis Mitchill (Fallfish), and Pimephales notatus (Raphinesque) (Bluntnose Minnow)—were only collected at esop6. Trout populations Understanding spatial and temporal patterns in trout populations was a primary focus of this study. Spatial trends in total density and biomass of trout populations at all tributary and mainstem sites were not generally predictable based on DA (Fig. 4). In fact, DA only explained 2–3% of the variability in the density or biomass of all trout (pooled) populations at all sites in the basin. Trout density averaged 200 fish/0.1 ha at the 6 small tributary sites (DA < 40 km2), 120 fish/0.1 ha at the 5 large tributary (DA = 40 km2–100 km2), and 230 fish/0.1 ha at the 7 mainstem (DA > 100 km2) sites (excluding esop0) (Fig. 4A). Trout biomass averaged 3015, 2113, and 3058 g/0.1 ha at the small tributary, large tributary, and mainstem sites, respectively (Fig. 4B). Median trout density was significantly lower at downstream sites (149 fish/0.1 ha) than at upstream sites (358 fish/0.1 ha) (Table 3). Median trout biomass, however, did not differ significantly between mainstem sites located upstream and downstream from the portal. We collected Brown Trout and Rainbow Trout at most sites. Estimates of total biomass for both species collected from all sites each year during 2009–2011 are summarized in Figure 5. The average density of Rainbow Trout populations at all sites (169 fish/0.1 ha) was about 3-fold higher than that of Brown Trout (59 fish/0.1 ha), however average biomass of Brown Trout populations at all sites Figure 4. Estimates of total density (A) and biomass (B) and 95% confidence intervals (CIs) from all trout at sites surveyed in the Upper Esopus Creek during 2009, 2010, and 2011. Northeastern Naturalist Vol. 22, No. 2 B.P. Baldigo, S.D. George, and W.T. Keller 2015 359 (2347 g/0.1 ha) was about 3-fold greater than that of Rainbow Trout populations (693 g/0.1 ha) (Table 3). The portal may affect Brown Trout and Rainbow Trout populations indirectly by altering survival of their early life stages (e.g., eggs, swim-up fry, and youngof- year), thus altering the proportions of juvenile and mature fish. We used nonparametric Kruskal-Wallis tests to assess differences in median density and biomass for both species and for differences in median densities and percentage of youngof- year (YOY) Brown Trout and Rainbow Trout at 3 upstream, 4 downstream, and 11 tributary sites (Table 3). The median density of Brown Trout populations was significantly lower at downstream sites (25 fish/0.1 ha) than at upstream sites (69 fish/0.1 ha), but median biomass did not differ significantly among the 3 site types. Median densities of mature Brown Trout did not differ significantly between downstream and upstream sites. Median density of YOY Brown Trout was significantly lower at downstream sites (9 fish/0.1 ha) than upstream sites (34 fish/0.1 ha). The YOY constituted 38%, 75%, and 37% of Brown Trout populations at tributary, upstream, and downstream sites, respectively, and the median percentages were significantly lower at downstream sites (30%) than at upstream sites (83%). Median density and biomass estimates for Rainbow Trout populations did not differ significantly between downstream and upstream sites (Table 3). Median density of mature Rainbow Trout was significantly lower at downstream sites (2 fish/0.1 ha) than at upstream sites (17 fish/0.1 ha). Median densities of YOY Rainbow Trout did not differ significantly between downstream and upstream sites. The YOY constituted 62%, 77%, and 93% of Rainbow Trout populations at tributary, upstream, and downstream sites, respectively, and median percentages did not differ significantly between downstream and upstream sites. Figure 5. Estimates of Brown Trout biomass (A) and Rainbow Trout biomass (B) and 95% confidence intervals (CIs) at sites surveyed in the Upper Esopus Creek during 2009–2011. Northeastern Naturalist 360 B.P. Baldigo, S.D. George, and W.T. Keller 2015 Vol. 22, No. 2 Temporal and spatial variability in fish assemblages The temporal trends in fish communities and trout populations were relatively similar. Community density and biomass decreased significantly at many sites (e.g., stoc1, fox, and esop3a) (see Fig. B in Supplemental File 1, available online at https://www.eaglehill.us/NENAonline/suppl-files/n22-2-N1280-Baldigo-s1, and for BioOne subscribers, at http://dx.doi.org/10.1656/N1280.s1), and overall at all sites (Table 3) between 2009 and 2011. Temporal trends in density and, to some degree, biomass of trout populations were also significant and relatively consistent at many sites (Table 3, Fig. 4). The density of trout populations at most sites was generally highest in 2009, lower in 2010, and lowest in 2011; many of these differences were significant between years (Fig. 4A). The changes in trout biomass over the 3 periods were comparable to density changes at several sites (e.g., esop3, wood, and stoc0), yet the observed decreases were less regular, often not significant, and sometimes interrupted by increases during 2010 (Fig. 4B). The few exceptions to these temporal trends were generally related to collection of a few large trout, which strongly influenced biomass estimates at individual sites such as esop2 during 2011 and esop4 during 2010. Median density of all trout was significantly lower in 2011 (46 fish/0.1 ha) than in 2009 (264 fish/0.1 ha) and 2010 (183 fish/0.1 ha), and median biomass was significantly lower in 2011 (1972 g/0.1 ha) and 2010 (2031g/0.1 ha) than in 2009 (3012 g/0.1 ha) (Table 3). In general, all mean and median population metrics for Brown Trout decreased significantly from 2009 to 2011; most decreases were also significant between 2009 and 2010 (Fig. 5A). Three median Rainbow Trout population metrics decreased significantly between 2009 and 2011, but only the decrease in density of mature Rainbow Trout was significant between 2009 and 2010 (Fig. 5B). Multidimensional scaling ordinations and cluster analysis using species-density data identified 5 unique site groupings distributed roughly along a stream-size or drainage-area gradient (Fig. 6). The composition of fish assemblages at all sites (and survey years) within each group was 58–72% similar to each other and differed significantly (P < 0.05) among groups. The groups consisted of (1) only small headwater tributaries, which were 67% similar; (2) large tributaries, Broadstreet Hollow Brook (broad), and mainstem sites mostly upstream of esop4a, which were 53% similar; (3) large tributaries and one upstream site (esop3a), which were 58% similar; (4) esop6 during 2010 and 2011, which were 64% similar; and (5) esop6 and esop3a during 2009, which were 72% similar. The last 2 groups were also 55% similar and are treated as 1 group in the discussion. The overall groupings confirm major differences among fish assemblages observed at small tributaries, the largest downstream site, and all other sites, yet they identify no consistent differences between fish assemblages at sites upstream and downstream of the portal. Density bubble-plots for Slimy Sculpin and Exoglossum maxillingua Lesueur (Cutlips Minnow) illustrated pronounced trends along the horizontal axis (Fig. 7). Slimy Sculpin populations were generally most dense at small headwater tributaries (1800–3100 fish/0.1 ha at bush) and essentially absent at large downstream mainstem sites during all years (Fig. 7A), whereas Cutlips Minnow populations were largest (600–800 fish/0.1 ha) at upstream (esop3a) and downstream (esop6) mainstem sites during Northeastern Naturalist Vol. 22, No. 2 B.P. Baldigo, S.D. George, and W.T. Keller 2015 361 2009 (Fig. 7B). A SIMPER analysis confirmed that Slimy Sculpin was responsible for as much as 32.4% of the dissimilarity among small tributary, large tributary, and mainstem groups. Cutlips Minnow was almost as influential as Slimy Sculpin and contributed as much as 19.6% to the dissimilarities among groups. These findings indicate that Slimy Sculpin and Cutlips Minnow may be suitable indicator species for classifying community types in this basin. Discussion In general, the results from this investigation indicate that fish assemblages in the Esopus respond to waters from the portal in a variety of ways. The most important finding was that no adverse changes in fish assemblages could be attributed directly to portal waters, and some of the noted effects might even be considered beneficial. Though fish communities were altered at sites near and downstream from the portal, many changes were positive and could be related to changes in temperature or habitat quality and quantity. One companion study of Figure 6. Non-metricMDS ordination plot of fish assemblages based on square-root-transformed density of all species collected at 16–18 sites in the Upper Esopus Creek annually from 2009–2011. The 5 symbols denote site membership in groups with significantly similar communities (53–72% Bray-Curtis similarity, P ≤ 0.05) based on group-averaged cluster analysis. Site locations are shown in Figure 1. Figure 7 (following page). Density-bubble plots showing the distribution and density of (A) Slimy Sculpin and (B) Cutlips Minnow populations across sites and years (labels on bubbles) in the Upper Esopus Creek, 2009–2011. Site locations are shown in Figure 1. Northeastern Naturalist 362 B.P. Baldigo, S.D. George, and W.T. Keller 2015 Vol. 22, No. 2 Northeastern Naturalist Vol. 22, No. 2 B.P. Baldigo, S.D. George, and W.T. Keller 2015 363 fatty acids and periphyton communities in the upper basin (S.D. George, US Geological Survey, Troy, NY, unpubl. data) detected the only ecological impairment that may be linked directly to the portal: a reduction in the standing crop of primary producers at sites immediately downstream from the portal. Potentially adverse effects were also noted in the density and biomass of juvenile life stages of Brown Trout. Though such effects could be caused by impaired water quality, the portal was not a major contributor to turbidity and suspended sediment loads in the upper basin (McHale and Siemion 2014). Loads typically reflect short-duration (high-flow) events and, thus, may not be as important to trout growth and survival as long-term exposure to moderately elevated levels of suspended sediment and turbidity. During water-years 2010 and 2011, median turbidity from 34 routine grab samples collected on the same dates (every 2–4 weeks) upstream from the portal at esop03a (7.5 NTRUs) and downstream from the portal at esop03b (9.4 NTRUs) (USGS 2014) did not differ significantly (Mann-Whitney U-Test: P = 0.1990), nor did the distributions of turbidity data from both sites (Kolmogorov- Smirnov Test: P = 0.1601). The median turbidity for 24 mean-daily values (estimated monthly from continuous data) at the portal during the same 2-year period was 16.4 NTU (EPA 2014). Although these results indicate that waters from the portal slightly reduced clarity at downstream sites, turbidity levels generally remained below thresholds (approximately 40 NTU) found to impair feeding, growth, and survival of several trout species (Newcombe 2003, Rowe et al. 2003, Stuart-Smith et al. 2004, Sweka and Hartman 2001b), which suggests that these small increases may not be biologically meaningful. When compared to richness at other mainstem sites, the significantly larger number of species at sites esop3a and esop3b, immediately upstream and downstream of the portal, suggests that the portal’s connection to the Schoharie Basin was the source of several fish species. Similarly, richness was high at the furthest downstream site (espo6), near the Ashokan Reservoir where we encountered as many as 15 species. The richness of fish species at sites in the nearby Beaverkill Basin, with the same drainage area as esop6, ranged from 12 to 14, whereas comparable sites in the adjacent Neversink Basin (which are close to its confluence with the Delaware River) reach as high as 18 species (Baldigo, et al., in press). Richness at sites in the Beaverkill with the same drainage area as esop3a and esop3b ranged from 10 to12 species, whereas 7 species were collected at the 1 comparable site in the Neversink, which was upstream of a reservoir. Although the small increases in richness at sites near the portal (and at esop6) did not cause significant differences in diversity, the addition of species creates more complex food webs, which could benefit local communities by making ecosystems less susceptible to the effects of short- and long-term stresses or perturbation (Schaefer et al. 2012). The increased richness at esop6 was likely related to its close proximity to the reservoir and to its habitat, which is more suitable for these lentic species than sites further upstream. Although increased species richness may be beneficial, additional nonnative species could also be detrimental to resident fish communities (Cucherousset and Olden 2011). Northeastern Naturalist 364 B.P. Baldigo, S.D. George, and W.T. Keller 2015 Vol. 22, No. 2 Waters from the portal had a number of measureable effects on fish communities at downstream mainstem sites that suggest the normal upstream–downstream continuum was disrupted. Estimates of total density and biomass for fish communities at individual mainstem sites revealed no significant differences that could be directly attributed to the portal, yet total density and biomass at 4 mainstem sites downstream were significantly lower than at the 3 mainstem sites upstream. Conversely, the ANOSIM detected no significant impacts of the portal on fish communities at mainstem sites downstream. More important, however, may be the increased number of fish species at several sites (e.g., esop3a and esop3b) in the Esopus. Fish assemblages (i.e., the distribution of individual species) in large undisturbed basins normally follow a predictable succession (longitudinal zonation) between small, cold, low-order headwater reaches, and large, warm, high-order reaches reflecting a continuum of abiotic and biotic factors (Vannote et al. 1980). Fish-species richness and diversity normally increase progressively with increasing stream size and order (at low to mid-order reaches), often with additions to, rather than replacement of, existing species (Hutchinson 1993, Whiteside and McNatt 1972, Zalewski et al. 1990). In general, physical and chemical factors including dissolved oxygen, pH, alkalinity, suspended turbidity, and conductivity often change with stream size; however, stream depth, size, order, and habitat heterogeneity typically explain the largest amount of variability in richness and diversity of fish communities (Hutchinson 1993; Schlosser 1987, 1991). In temperate rivers, abiotic factors can limit the diversity of low-order fish communities to relatively few species that are tolerant of lower temperatures, wide fluctuations in flow, small/ shallow channels with homogeneous habitat, and high water-velocities. High rates of primary production in mid-order stream reaches can increase species richness, biodiversity, density, and/or levels of biomass for secondary and tertiary consumers (Ward and Stanford 1983). Other biotic factors, such as competition and predation, typically control richness and diversity of fish communities at larger downstream reaches where relatively stable flows, more heterogeneous stream channels and habitat, and slower and warmer waters permit additional species to coexist (Vannote et al. 1980). Mid-basin changes in richness and diversity indicate a small and localized effect on the Esopus ecosystem, yet decreases in fish-community biomass and density indicate that primary production, along with secondary and top consumers, may have been unusually low at several sites downstream from the portal. Altered composition of periphyton communities could be responsible for decreases in primary production at mid-order sites in the Esopus. Periphyton is responsible for most primary production and form the base of the food web in low- to mid-order streams in temperate climates (Vannote et al. 1980). Chlorophyll a (chl a) concentration and ash-free dry mass (AFDM) quantify the standing crop of primary producers (assumed herein to be a surrogate for primary production) in stream food webs (Rosemond et al. 1993). Although neither chl a nor AFDM differed significantly between groups of upstream and downstream sites, these values were significantly depressed at the first 3 downstream sites (esop3b, esop4, and esop4a) compared to the 3 upstream sites (Smith et al. 2013). Large decreases in Northeastern Naturalist Vol. 22, No. 2 B.P. Baldigo, S.D. George, and W.T. Keller 2015 365 primary production can cause cascading effects at higher trophic levels (e.g., primary consumers, predators; Kurle and Cardinale 2011). Changes in the abundance and/or biomass of certain preferred (or all) macroinvertebrate-prey species could theoretically limit available food resources, and therefore, growth and survival of some fish species, at one or more downstream sites. The associations between standing crop, periphyton communities, and macroinvertebrate communities are relatively strong in the Esopus (Smith et al. 2013) and provide additional evidence that the food web is altered at several downstream sites. Because changes in primary production can cause cascading effects throughout the food web (Henley et al. 2000, Vannote et al. 1980), direct and indirect impacts to predator (fish and macroinvertebrate) populations are possible. The indirect effects of reduced primary production and the direct effects of unusual thermal, hydrologic, and water-quality regimes on the 2 dominant trout species are potentially important, but difficult to isolate in this system because of their dissimilar life-histories. Although both types of effects may be attributed to the portal, the disparate timing for spawning, incubation, and emergence (from gravel beds) make the early life stages for each species differentially vulnerable to storm flows occurring at different times of the year. Nevertheless, the median density of trout at downstream sites was significantly lower than that at sites upstream from the portal, and biomass differences were only nominal. Differences in median Rainbow Trout metrics were generally minor; thus, changes in Brown Trout populations drove differences in overall trout metrics. Median densities of all Brown Trout and YOY Brown Trout were significantly lower at downstream sites than at upstream sites, yet densities of mature Brown Trout did not differ between the 2 reaches. Ross (2012) determined that upstream reaches contained more habitat preferred by juvenile Brown Trout than did downstream reaches, which agrees with the significant differences we detected in densities of YOY Brown Trout between sites from both reaches. Although the 37% decrease in biomass of Brown Trout populations between upstream and downstream sites was not significant, the significantly lower density of YOY Brown Trout suggests that their populations might have been larger if not for their location downstream of the portal. Clearly, the abundance of juvenile Brown Trout was reduced at many downstream Esopus sites. Though the potential effects of the portal on trout provided the impetus for this investigation, a number of related mechanisms could also adversely affect productivity and fish assemblages downstream. This consideration is important because observational studies can only describe the strength of the relationships between or among factors; the findings cannot unequivocally attribute a specific effect to a specific cause. Assuming that turbid waters directly reduce primary productivity, then the portal would only play a nominal role because it has been a minor source of turbidity and suspended sediments since the mid-1990s (CCE 2007). For example, estimates from all samples collected during water-year 2010 for median turbidity (NTRU) varied from 119 at Stony Clove, 26 at Broadstreet Hollow Brook, 41 at Woodland Valley Creek, and 47 at Beaverkill, to 77 at the last mainstem site Northeastern Naturalist 366 B.P. Baldigo, S.D. George, and W.T. Keller 2015 Vol. 22, No. 2 (esop3a) upstream of the portal, and 20 at the first site (esop3b) downstream of the portal (McHale and Siemion 2014, USGS 2014). The median turbidity for 12 mean-daily values (estimated monthly from continuous data) reported for waters discharged from the portal during water-year 2010 was 15.3 NTU (EPA 2014), which suggests that portal waters do not substantially alter natural turbidity levels within the Esopus. Thus, turbidity from portal waters probably does not limit fishspecies distributions and their assemblages at downstream Esopus sites. These data indicate that several tributaries, not the portal, are the primary sources of turbidity in the upper basin. In fact, Stony Clove Creek accounted for more of the total suspended sediment load (30–57%) at the furthest downstream site, esop6, during water-years 2010 and 2012 than did all other tributaries combined (McHale and Siemion 2014). The results from other investigations show that elevated turbidity levels can strongly affect the behavior, growth, and condition of Brown Trout and other salmonids, yet few have detected strong linkages in the wild (Henley et al. 2000). During laboratory experiments, significant changes in behavior and activity levels of Brook Trout only occurred when turbidity levels surpassed 7.1 NTU (Gradall and Swenson 1982). Increasing turbidity levels from 0 to 45 NTU had no significant effect on the rates at which Brook Trout consumed prey because they switched from passive to more active searching and feeding behavior (Sweka and Hartman 2001a). Although growth rates of Brook Trout decreased linearly with increasing turbidity, they did not differ significantly from controls until turbidity levels reached 45 NTU, when they declined by 62% (Sweka and Hartman 2001a). Similarly, the volume of stomach contents and prey diversity for Brown Trout sampled from Lake Sorell, Tasmania, were greater in 1996 when turbidity averaged 26 NTU than they were during 2001 when turbidity averaged 141 NTU (Stuart-Smith et al. 2004). Repeated exposure to 0–6-d pulses of turbid (suspended sediment concentrations of 700 mg/L) water also significantly affected growth of juvenile Rainbow Trout in streamside (caged) experiments (Shaw and Richardson 2001). Model simulations predicted that if Rainbow Trout were exposed to high turbidity (53 NTU), they would occupy shallower- and slower-water habitats, switch from passive to active prey-capture behavior, gain no net energy, and remain in poor condition (Harvey and Railsback 2009). Despite these findings, Rainbow Trout had normal feeding rates (White and Harvey 2007) and wild trout populations generally persisted (Harvey and Railsback 2009) in streams with highly turbid regimes, purportedly because trout can sense and capture prey using non-visual organs (e.g., the lateral line) under highly turbid and low-light conditions (Rowe et al. 2003). The supplemental flows from the portal could potentially benefit the health of resident trout and their populations at downstream Esopus sites in several ways. The intakes for the Shandaken tunnel are deep in the Schoharie Reservoir and withdrawals usually originate in the hypolimnion. Thus, water temperatures for these enhanced flows are typically lower at downstream Esopus sites than they are at upstream sites during the warm months. The positive effect of the supplemental flows on habitat quantity (area and volume) in the upper basin Northeastern Naturalist Vol. 22, No. 2 B.P. Baldigo, S.D. George, and W.T. Keller 2015 367 is illustrated by adding flows at esop3a, upstream of the portal (USGS 2012a), which averaged 97 ft3s-1, to those from the portal (USGS 2012b), which averaged 226 ft3s-1 during summer base-flow periods (July–September) between 1996 and 2011. Simple addition shows that Esopus flows immediately downstream from the portal averaged at least 323 ft3s-1 during these same months, and would have been about 77% lower without supplemental flows from the Schoharie Reservoir. During the warmest months of June, July, and August 2011, water temperatures averaged 11.4, 14.0, and 17.2 °C, respectively at the portal; 14.9, 17.9, and 17.8 °C, respectively at the upstream site esop3a; and 12.8, 15.6, and 17.3 °C, respectively at the downstream site esop3b. Except for August 2010 (when the Schoharie Reservoir reached unusually low levels and epilimnion withdrawals replaced hypolimnion withdrawals), water temperatures during June–August in 2009–2011 averaged 2.7 °C lower at the portal than at esop3a, and temperatures averaged 1.8 °C lower at esop3b than at esop3a. The large volumes of cold water from the portal have an important effect on summer stream temperatures at Esopus reaches downstream from the portal. Such thermal differences are crucial to the viability of Brown Trout populations in reaches where water temperatures approach and sometimes surpass thermal limits for growth and survival (Elliott and Elliott 2010, Wehrly et al. 2007). The findings by Ross (2012) help explain our conflicting results that identified substantial effects on the densities of juveniles (YOY), yet only minor effects on densities of mature Brown Trout at downstream sites. Ross (2012) reported that biomarkers (serum chemistry, gill histology, and water content) and growth rates indicated that Brown Trout were generally stressed at all reaches (upstream, downstream, and further downstream), but that the portal had no significant impact on the species at downstream reaches in the Esopus. In general, Ross (2012) determined that (a) the health and growth rates of adult Brown Trout were poor across the Upper Esopus, (b) the quality of habitat for adult trout was optimal at downstream sites, and (c) the quality of habitat for juvenile trout was poor at downstream sites. These findings partly explain why mature Brown Trout densities were generally unaffected and YOY Brown Trout were less abundant at downstream sites. Possibly more notable was the fact that growth of Brown Trout, during warm months, was less negative in reaches immediately downstream from the portal than growth of trout from upstream and further downstream reaches. The inference is that conditions downstream from the portal were less stressful than conditions at upstream and further downstream reaches during the warm months. These data suggest that a stress-refuge may exist periodically at sites immediately downstream from the portal, yet our results indicate that no broad benefit was conveyed to resident Brown Trout or Rainbow Trout populations at downstream sites. The results from annual surveys generated baseline information for an important natural resource and also detected decreasing temporal trends in many fish-community and population metrics. These data are important because they define the current status of local fisheries and quantify the normal measures of error (natural variability) that may occur in key metrics given the typical year-to-year variations in Northeastern Naturalist 368 B.P. Baldigo, S.D. George, and W.T. Keller 2015 Vol. 22, No. 2 temperature. precipitation, and stream flows. As such, these results provide a baseline dataset that defines the fishery status during the period 2009–2011. More important, these data are now available to quantify the impacts of short-term (e.g., severe flooding or contaminant spills) or long-term (e.g., altered thermal or hydrologic regimes associated with climate change) disturbances on resident-fish assemblages. The decline in many metrics between 2009 and 2011 may reflect natural responses to a drought during summer 2010 and moderate floods during the 2011 water-year (before we conducted our fish surveys). Although speculative, low flows and warmer than normal temperatures during summer 2010 could have caused widespread stress and mortality among resident-fish species, and flood flows during fall 2010 and spring 2011 could have negatively affected the survival of eggs and YOY Brown Trout that spawned in the fall or Rainbow Trout that spawned in the spring. Thus, depressed fish populations during 2010 could have further adversely affected the density and biomass of species populations during 2011. Regardless of the reasons, site-to-site and year-to-year variations in most fishery metrics are much larger and more notable than any effects that the portal may have had on fish populations. In summary, from 2009–2011, the fish assemblages at most sites in the Esopus were comparable to those previously observed in other rivers of the region. Community richness increased near the portal, and median densities and biomass of fish communities generally declined significantly at the downstream sites. Such effects, however, do not suggest a specific cause and could easily be related to a general shift to larger and fewer individuals, some sampling bias, differences in sampling efficiency, changes in habitat, or poor water quality at sites downstream from the portal. A decrease in water quality would implicate tributaries because they were the primary sources of increased turbidity and suspended sediments loads at all mainstem reaches downstream of the portal. The only obvious deleterious effect of the portal seemed to be a large decrease in the standing crop of primary producers as indicated by changes in chl a and AFDM at sites immediately downstream from the portal (Smith et al. 2013). Less conspicuous, the densities of YOY Brown Trout were significantly reduced at several downstream sites, but associated impacts are not transferred to their whole populations. The increased water volume and decreased temperature at downstream reaches during summer months also increase physical habitat and help moderate temperatures that would otherwise be harmful to resident trout. Although such conditions may not truly be beneficial, the conditions at mainstem sites immediately downstream from the portal are clearly less harmful to trout (and their populations) than are conditions at upstream and further downstream sites during warm months. Despite the notable decreases in primary production and density of YOY Brown Trout at several sites, this study shows that fish communities and trout populations across most reaches in the Esopus are generally in good condition and unaffected by the portal. Acknowledgments The authors extend appreciation to David Munsey and Martyn Smith of the USGS; Danyelle Davis and Mark Vian of the NYCDEP; Karen Stainbrook of the NYSDEC; and Northeastern Naturalist Vol. 22, No. 2 B.P. Baldigo, S.D. George, and W.T. 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