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Macroinvertebrate Assemblages, Stormwater Pollution, and Habitat Stressors in the Bronx River

Alejandro Baladrón1* and David J. Yozzo1

1Great Ecology - 379 W. Broadway, 5th Floor New York, NY 10012 USA. *Corresponding author

Urban Naturalist, No. 31 (2020)

Abstract
The Bronx River, located within the Hudson-Raritan watershed in southeastern New York State, is an urban watercourse affected by stormwater pollution and urban development. Four sites along the river were sampled to examine relationships between 15 environmental variables and 26 benthic metrics. Macroinvertebrate assemblages presented low diversity and were dominated by aquatic worms and midge larvae. Variation in environmental factors among sites coincided with significant differences among benthic metrics. Moderate recovery of macroinvertebrate diversity was observed downstream from an intact riparian corridor (The New York Botanical Gardens) which may be functioning as a disturbance buffer, increasing stream habitat quality within an otherwise urbanized and degraded watercourse.

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Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 1 2020 URBAN NATURALIST No. 31:1–22 Macroinvertebrate Assemblages, Stormwater Pollution, and Habitat Stressors in the Bronx River Alejandro Baladrón1* and David J. Yozzo1 Abstract - The Bronx River, located within the Hudson-Raritan watershed in southeastern New York State, is an urban watercourse affected by stormwater pollution and urban development. Four sites along the river were sampled to examine relationships between 15 environmental variables and 26 benthic metrics. Macroinvertebrate assemblages presented low diversity and were dominated by aquatic worms and midge larvae. Variation in environmental factors among sites coincided with significant differences among benthic metrics. Moderate recovery of macroinvertebrate diversity was observed downstream from an intact riparian corridor (The New York Botanical Gardens) which may be functioning as a disturbance buffer, increasing stream habitat quality within an otherwise urbanized and degraded watercourse. Introduction The most widespread forms of disturbance affecting aquatic systems now come from human activities (Nicacio and Juen 2015), and their effect on the structure of biological communities is highly variable (Battisti et al. 2016, Isaac and Cowlishaw 2004). Benthic metrics are used to assess the stress caused by anthropogenic disturbances by measuring aspects of biological communities that respond in different manners to stressors (Barbour et al.1999, Canobbio et al. 2013, Huang et al. 2014). The present project is a macroinvertebrate-based biomonitoring study conducted at the Bronx River, an urban watercourse mainly affected by two disturbance sources: 1) discharges of polluted water, and 2) degradation of the riparian corridor caused by urban development. Physical and chemical parameters in the Bronx River are affected by periodic discharges of polluted water from combined sewer overflows (CSOs), stormwater runoff from impervious surfaces (NYSDEC 2015a, PlaNYC 2008), sanitary sewer breaks and illegal wastewater connections that occur throughout the River corridor and watershed (Crimmens and Larson 2006). Discharges of polluted water result in water-quality impairment, often affecting the River biota (Corsi et al. 2010, Kayhanian et al. 2008). Another watershed-specific factor affecting the Bronx River is the degradation of surrounding riparian habitat caused by urban development (i.e., road building, shoreline “hardening”, and stream channelization). Notable degradation factors associated with urban development include the removal of riparian vegetation and the deposition of fine sediments in the riverbed (Larsen et al. 2010, Walsh et al. 2005). Riparian communities provide multiple ecosystem services, including streambank stabilization (Easson and Yarbrough 2002, Simon and Collison 2002), water purification (Hatt et al. 2006, Palmer and Richardson 2009), water temperature regulation (Wilkerson et al. 2006), leaf litter supply to the aquatic food chain (Cummins and Klug 1979, Pusey and Arthington 2003), and sediment trapping (Clinton et al. 2010, DeWalle 2010, NYSDEC 2015, Wilkerson et al. 2010). Lack of riparian vegetation may reduce the availability of such services. Additionally, when riparian cover is removed, upland sediment loads can increase by 1Great Ecology - 379 W. Broadway, 5th Floor New York, NY 10012 USA. *Corresponding author - abaladron@greatecology.com. Manuscript Editor: Sylvio Codella Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 2 many orders of magnitude (Wolman and Schick 1967), and the predominant grain size distribution can shift to much finer fractions (Jobson and Carey 1989). The removal of riparian vegetation may also contribute to destabilization of streambanks (Hubble et al. 2010, Shilla and Shilla 2012), favoring accelerated stream bank erosion (Fitzpatrick et al. 1998), and lateral channel migration (Knox 2006, Owen et al. 2011), causing sedimentation problems downstream (Piégay et al. 2005). Sediment deposition has been associated with reduced richness and densities of sensitive taxa (Jones et al. 2011), as well as with altered representation of behavioral traits (Von Bertrab et al. 2013), habitat preference, and life history (Larsen et al. 2010, Wiitala 2013). In this study, four Bronx River sites with different levels of disturbance were assessed for changes in environmental conditions and invertebrate assemblages. We assumed that the quality of physico-chemical, morphohydraulic, and riparian features affecting macroinvertebrate composition would decrease along a downstream urbanization gradient with increasing water pollution and river habitat degradation. We also hypothesized that improvement of water quality and habitat conditions would occur directly downstream from an intact riparian corridor. Materials and Methods Study site The Bronx River basin is part of the New York-New Jersey Harbor Estuary watershed and drains approximately 124 km2 of urbanized land in Westchester and Bronx Counties. The River originates in Valhalla, flows approximately 39 km south past White Plains, and then turns south-southwest through the northern suburbs of New York City, including Edgemont, Tuckahoe, Eastchester, and Bronxville. The River continues through the Bronx, including a portion of the New York Botanical Gardens (NYBG), before reaching its confluence with the East River at Hunt’s Point. Precipitation is generally well distributed throughout the year, with the wettest conditions in April and May and driest in February (WCDP 2007). Although the River’s water quality falls well within acceptable biological thresholds, it is also considered moderately to severely impacted (Hudsonia 1994, NYSDEC 1998). Four reaches of the River were selected for sampling (Fig. 1): Site 1 is adjacent to the North White Plains Metro-North rail station, in central Westchester County. Previous studies near Site 1 indicate that water quality approaches the “slightly impacted” range (Hudsonia 1994; NYSDEC 1998, 2003). For this reason, Site 1 was used to establish reference conditions. Site 2 is at Muskrat Cove, between East 233rd Street and the Bronx River Parkway, in the northern portion of the Bronx. At this site, the River is channelized with extensively hardened shorelines and riparian vegetation is scarce. Site 3 is at Kazimiroff Boulevard, immediately upstream from NYBG. Although the River has a relatively well-structured red maple-hardwood riparian corridor at this location, there is evidence of streambank erosion and increased sediment load on the riverbed (Runfola and Weiss 2007). Finally, Site 4 is located at River Park next to 180th Street, downstream from a waterfall located at the southern boundary of the Bronx Zoo. Sites were selected with the intention of assessing degradation of the stream habitat from Sites 1 to 3, and habitat recovery from Sites 3 to 4 (Fig. 1). There are >100 stormwater and other discharges that flow to the Bronx River from Westchester County to Tremont Avenue (NYCDEC 2015a, Wang 2011). Sites 2, 3, and 4 are located along this section of the river (Fig. 1) and, therefore, are exposed to increasing levels of water pollution, with Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 3 downstream sites potentially receiving higher pollutant loads. Therefore, Sites 2 and 3 were expected to support macroinvertebrate communities that were indicative of poorer habitat quality than the Site 1 community. Site 4, although potentially more exposed than any other site to water pollution resulting from downstream pollutant accumulation, was also located downstream from the New York Botanical Garden (NYBG), where a dense riparian corridor may be functioning as a vegetative buffer. A comparison between macroinvertebrate assemblages at Site 4 and Sites 2 and 3 was intended to determine whether or Figure 1. Sampling site locations. Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 4 not the riparian corridor of the Bronx River acts as a vegetative buffer as it passes through the NYBG. To investigate these assumptions, data corresponding to fifteen environmental factors and thirty benthic metrics were selected to assess changes in the structure of the invertebrate community. Benthic metrics were selected from the Environmental Protection Agency Rapid Bioassessment (EPA RBA) protocol, including richness and diversity measures, dominance, presence/absence of tolerant or intolerant species, functional feeding groups measures, and combination indices: ASPT, BMWP, FBI, CLI (Table 1), and Biological Assessment Profiles (BAPs). Some of these metrics are already used by the New York State Department of Environmental Conservation (NYSDEC) in their regular biomonitoring efforts at the Bronx River. Metrics from the EPA RBA protocol not used by NYSDEC were chosen to compare alternative measures and to investigate if extra value could be added to NYSDEC’s standard Table 1. Combined indices and response to increasing perturbation (modified from Barbour et al. 1999). Benthic metric Definition Predicted response to increasing disturbance Reference BMWP (Biological Monitoring Working Party) This index provides a score that will vary depending on how tolerant are taxa to pollutants. The score is obtained adding all the individual scores corresponding to each invertebrate family detected in a sample; families are assigned a score between 1 and 10 accordingly. Decrease Hawkes 1997, Mason 2002 ASPT (Average Score Per Taxon) Represents the average tolerance score of all taxa within the community, and is calculated by dividing the BMWP by the number of families represented in the sample. A high ASPT score is considered indicative of a clean site containing large numbers of highscoring taxa. Decrease Armitage et al. 1983, Friedrich et al. 1996 FBI (Family Biotic Index) Similar to BMWP, but this index takes also into account how many individuals account for each taxa in the sample, as well as the total number of individuals in the sample. Tolerance scores for each invertebrate family are weighted by the abundance of the invertebrate family and the total abundance of the sample. Increase Plafkin et al. 1989 CLI (Community Loss Index) This index estimates the loss of taxa between comparison samples and reference samples. Increase Plafkin et al. 1989 Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 5 approach. Definitions of the selected benthic metrics and their responses to increasing perturbation can be found in Barbour et al. (1996), Bode et al. (2002), DeShon (1995), Fore et al. (1996), Friedrich et al. (1996), Hawkes (1997), Kerans and Karr (1994), Morin (2011), NYSDEC (2014), Plafkin et al. (1989), and Smith and Voshell (1997). Benthic macroinvertebrate sampling and environmental measurements Field sampling. Five replicate samples of benthic material were collected using a 30 × 30 cm Surber sampler (500 μm mesh, 0.09 m2 sample area) at each sampling site during the first two weeks of April 2012 (N = 20 samples total). Replicates were separated by 5 meters inside a 20-meter sampling reach (Fig. 2). Prior to sample collection, physico-chemical and hydraulic parameters were measured in the water column corresponding to each replicate quadrat. Temperature, dissolved oxygen, pH, conductivity, turbidity, and water velocity were measured using a Hydrolab Quanta Multiparameter Sonde, and water depth was measured using a 1-meter wooden rule. A visual estimation of fine sediments in each quadrat was performed to assign a percentage of embeddedness to each replicate sample. The substrate surface in each quadrat was subsequently disturbed for one minute, allowing invertebrates and benthic material to drift into the net placed downstream of the quadrat. After disturbing the substrate, 30 stones were randomly selected inside the quadrat, and each was measured by its longest axis to obtain a size distribution of stream substrate material. Benthic material trapped in the net was preserved in 70% ethanol. Finally, channel width and canopy cover were measured by establishing five Figure 2. Schematic of the design used to collect environmental data (physico-chemical, hydraulic, substrate composition, tree canopy cover and organic leaf matter data) and macroinvertebrate data (modified from Fitzpatrick et al. 1998). Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 6 transects perpendicular to the channel (Fitzpatrick et al. 1998). Measurements taken within each transect were associated to their corresponding replicate sample (Fig. 2). Channel width was measured with a tape measure, and percent canopy cover was visually estimated from the center of the channel using a clinometer (Fitzpatrick et al. 1998; Fig. 2). Laboratory analyses. For each replicate, macroinvertebrates were identified from 12–13 cm3 sub-samples of benthic material in a time-constrained sorting procedure. At 10-minute intervals, a sub-sample was sorted for invertebrates and discarded in a 500 mm sieve for subsequent ash free dry mass (AFDM) analysis. Invertebrates were retrieved and preserved in 70% ethanol for further identification. The procedure was repeated for two hours. An additional 15 min/sample were spent searching for cryptic specimens and taxa not collected during the two-hour invertebrate extraction (see Rinella et al. 2005). For this task, sub-samples were picked and discarded every 5 min. In all, 15 sub-samples were picked for each replicate sample. Macroinvertebrates sorted from the benthic material were identified to the lowest taxonomic level possible (family and genus for most taxa) using keys by Daly (1996), Peckarsky et al. (1990), and Voshell (2002). Naididae (aquatic worms), Decapoda (crayfishes), Nematoda (roundworms), and Turbellaria (flatworms) were not identified to lower taxonomic levels. Following invertebrate identification, the benthic material was used to quantify the ash free dry mass (AFDM) of leaf litter present. Total benthic organic matter corresponding to each sample was dried at 60 ºC for 24 h, weighed, ashed at 550 ºC for two hours, and then reweighed (Cummins and Klug 1979). The difference between dry mass weight and ash dry mass weight was calculated and expressed as AFDM. Data analysis. Densities of invertebrate taxa were calculated (Table 2), and environmental measures obtained during field sampling were tabulated. Size distributions of stream Table 2. Mean macroinvertebrate densities (No. m-2) with standard error (SE). Taxa Functional group Site 1 Site 2 Site 3 Site 4 Hydropsyche sp. (net-spinning caddisfly) Filtering collectors 436 ± 13.9 93 ± 8.4 0 ± 0.0 271 ± 3.4 Simulium sp. (black fly) Filtering collectors 251 ± 10.3 13 ± 1.2 4 ± 0.4 13± 0.5 Antocha sp. (crane fly) Gatherers/collectors 0 ± 0.0 0 ± 0.0 0 ± 0.0 2 ± 0.2 Chironomidae (non-biting midge) Gatherers/collectors 6196 ± 53.6 5318 ± 45.8 2953 ± 25.5 4129 ± 25.1 Naididae (aquatic worm) Gatherers/collectors 4413 ± 45.1 8304 ± 71.7 11867 ± 102.3 2738 ± 31.1 Podura sp. (Podurid springtail) Gatherers/collectors 9 ± 0.8 24 ± 1.36 0 ± 0.0 4 ± 0.2 Decapoda (crayfish) Scavengers 2 ± 0.2 0 ± 0.0 0 ± 0.0 0 ± 0.0 Gammarus sp. (sideswimmer) Scavengers 171 ± 5.1 293 ± 26.4 93 ± 4.2 11 ± 0.6 Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 7 substrate material corresponding to each sampling replicate were classified in four categories: boulder (>20 cm), cobble (20 cm≤ longest axis >6 cm), gravel-pebble (6 cm≤ longest axis >2 cm), and fine sediments (2 cm≤ longest axis >1 cm), and a percentage of substrate material was assigned to each category. Both taxonomic and functional feeding groups were expressed as relative abundances, the latter following the guidelines of Cummins and Merritt (1996), and benthic metrics were calculated. Biological Assessment Profiles (BAPs) were also calculated to obtain an overall assessment of water quality (non-impacted, slight, moderate, severe; Bode et al. 2002, NYSDEC 2014). Statistical procedures were performed with SPSS v. 17.0. One-way ANOVA was used to compare variables across sites; Kruskal-Wallis tests were used in cases where arcsine transformation failed to resolve heteroscedasticity and/or normality issues. To control for possible Type I errors, a False Discovery Rate (FDR) was applied (see Benjamini and Hochberg 1995, Koen et al. 2005). Benthic metrics and environmental measures showing statistically significant results after applying FDR were analyz ed with Tukey’s tests. Results Environmental differences between sites Nine environmental variables (conductivity, turbidity, water depth, pH, boulders, cobbles, embeddedness, channel width, and canopy) differed significantly among sites (Table 3). Both pH and conductivity were significantly higher at Sites 2 and 3 compared to Sites 1 and 4. Turbidity was significantly lower at Sites 2 and 3 compared to Sites 1 and 4. The relative percentage of boulders/cobbles was significantly higher at Sites 2 and 3 compared to Sites 1 and Table 2. Continued. Taxa Functional Group Site 1 Site 2 Site 3 Site 4 Nematoda (roundworm) Parasites 5082 ± 48.5 8069 ± 69.7 22 ± 1.3 162 ± 0.8 Chelifera sp. (true fly) Predators 0 ± 0.0 0 ± 0.0 0 ± 0.0 2 ± 0.2 Coenagrionidae (pond damsel) Predators 0 ± 0.0 0 ± 0.0 0 ± 0.0 2 ± 0.2 Hemerodromia sp. (dance fly) Predators 0 ± 0.0 16 ± 1.4 0 ± 0.0 13 ± 0.6 Turbellaria (flatworm) Predators 4 ± 0.2 40 ± 1.3 71 ± 1.3 13 ± 0.6 Ancyronyx sp. (Elmid riffle beetle) Scrapers 9 ± 0.8 13 ± 1.2 0 ± 0.0 9 ± 0.4 Baetidae (baetid mayfly) Scrapers 9 ± 0.8 0 ± 0.0 0 ± 0.0 0 ± 0.0 Stenelmis sp. (Elmid riffle beetle) Scrapers 229 ± 8.1 0 ± 0.0 0 ± 0.0 0 ± 0.0 Pyralidae (Pyralid moth) Shredders 0 ± 0.0 0 ± 0.0 9 ± 0.8 0 ± 0.0 Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 8 Table 3. ANOVA and Kruskall-Wallis test results for environmental factors and segregation of sites according to Post-hoc Tukey’s test. Results include associated levels of significance (p values), type of test performed for each environmental factor, and threshold values for post-hoc control of the false discovery rate (FDR). The “Verification value” column corresponds to the subtraction “p values - FDR threshold for p significance (maximum value for any variable to be considered statistically significant)”. Negative values on the “Verification value” column indicate which variables showed statistically significant values between sites. Superscript letters indicate which Sites were significantly different from each other (e.g., Turbidity at sites 4 and 1 are not significantly different between them, but they are significantly different compared to sites 2 and 3). Sites with two superscript letters are “in-between” groups (e.g., % Canopy value at site 2 did not differ significantly to values at sites 3 and 4, but values at sites 3 and 4 were significantly different between each other). Variables without superscript letters did not show dif ferences between Sites. Environmental variable Category p values Test Ranking FDR threshold for p significance Verification value Segregation of sites after Tukey´s test Site 1 Site 2 Site 3 Site 4 pH Water quality variables 0.001 ANOVA 1 0.0036 “-0.0061” 7.75 ± 0.07b 8.11 ± 0.02a 8.24 ± 0.06a 7.83 ± 0.07b Conductivity (ms cm-1) Water quality variables 0.001 ANOVA 2 0.0071 “-0.0097” 0.56 ± 0.00d 0.85 ± 0.00a 0.73 ± 0.00b 0.66 ± 0.00c Turbidity (NTU) Water quality variables 0.001 ANOVA 3 0.0107 0.0330 45.74 ± 3.03a 22.16 ± 3.84b 22.20 ± 4.08b 40.30 ± 3.77a Water depth (cm) Morpho-hydraulic variables 0.001 ANOVA 4 0.0143 “-0.0169” 13.10 ± 2.12b 12.60 ± 2.32b 30.00 ± 0.00a 13.40 ± 1.72b Boulders (%) Morpho-hydraulic variables 0.001 ANOVA 5 0.0178 0.2730 0.00 ± 0.00b 0.00 ± 0.00b 0.00 ± 0.00b 2.00 ± 1.33a Cobbles (%) Morpho-hydraulic variables 0.001 ANOVA 6 0.0214 “-0.0240” 6.00 ± 0.67b 6.67 ± 3.50a 0.00 ± 0.00a 24.00 ± 6.53b Channel width (m) Morpho-hydraulic variables 0.001 ANOVA 7 0.0250 “-0.0276” 6.46 ± 0.56c 7.50 ± 0.00c 12.48 ± 0.69b 20.00 ± 0.00a Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 9 Table 3. Continued. Environmental variable Category p values Test Ranking FDR threshold for p significance Verification value Segregation of sites after Tukey´s test Site 1 Site 2 Site 3 Site 4 Canopy (%) Riparian variables 0.001 ANOVA 8 0.0286 “-0.0311” 84.00 ± 4.00a 1.00 ± 1.00bc 40.00 ± 20.98b 0.00 ± 0.00c Embeddedness (%) Morpho-hydraulic variables 0.027 ANOVA 9 0.0321 “-0.0087” 36.00 ± 6.78 57.00 ± 7.35 70.00 ± 11.29 34.00 ± 5.10 Pebbles (%) Morpho-hydraulic variables 0.061 ANOVA 10 0.0357 0.0217 45.33 ± 4.55 58.67 ± 6.11 10.00 ± 3.33 52.67 ± 5.42 Velocity (m s-1) Morpho-hydraulic variables 0.069 ANOVA 11 0.0393 0.0261 0.36 ± 0.07 0.53 ± 0.09 0.22 ± 0.02 0.40 ± 0.10 AFDM (g) Riparian variables 0.118 Kruskal- Wallis 12 0.0429 0.0716 19.66 ± 7.84 7.16 ± 5.04 5.91 ± 1.78 3.74 ± 0.44 Oxygen (mg l-1) Water quality variables 0.736 ANOVA 13 0.0464 0.6860 8.98 ± 0.27 10.20 ± 0.12 10.94 ± 0.07 10.40 ± 0.20 Gravels (%) Morpho-hydraulic variables 0.822 ANOVA 14 0.0500 0.8220 48.67 ± 4.78 34.67 ± 8.54 90.00 ± 3.33 21.33 ± 5.01 Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 10 4. Water depth was significantly greater at Site 3 compared to Sites 1, 2, and 4. Channel width was significantly greater at Site 4 compared to Site 3, which in turn was also significantly greater than Sites 1 and 2. Tree canopy was significantly denser at Site 1 compared to Sites 2, 3, and 4. The lowest tree canopy density was observed at Site 4 (Table 3). No groups could be differentiated by means of Tukey’s tests for embeddedness. Taxon and functional feeding group composition of the invertebrate community Seventeen macroinvertebrate taxa were collected from the entire study area (Table 2). Organisms tolerant of water pollution comprised the majority of invertebrates collected, with relative abundances ranging between 94.5% and 99.9%. The assemblages were dominated by aquatic worms, non-biting midges, and roundworms, which collectively comprised >93% of the community. Among the remaining taxa, Hydropsyche sp. (netspinning caddisflies) and Gammarus sp. (sideswimmers) were usually the most abundant. Stenelmis sp. and Ancyronyx sp. (elmid riffle beetles) and Simulium sp. (black flies) constituted a small fraction of the community (1–2%) at Site 1 and were absent or rare at the other sites. The most abundant functional feeding group at all sites was gatherers/collectors, with 62–63% at Sites 1 and 2, and 93–99% at Sites 3 and 4. Frequently, aquatic worms were found parasitizing non-biting midges. For analysis purposes, we considered all aquatic worms to be parasites, although some component of the assemblage could represent other feeding groups. This group represented >30% of invertebrates at Sites 1 and 2, but were rare at Sites 3 and 4. Filtering collectors (net-spinning caddisflies, black flies) constituted 4% of the community at Sites 1 and 4. Additional feeding groups such as grazers, predators, scrapers, and shredders comprised a small fraction of the communities at all sites. Differences in benthic metrics between sites Nineteen benthic metrics analyzed showed significant differences among sites (Table 4). Non-biting midges; side-swimmers; black flies; net-spinning caddisflies; flatworms; elmid riffle beetles; Empididae: Chelifera sp. (dance flies) and Hemerodromia (true flies); and Podura (podurid springtails) were taxonomically homogeneous groups that constituted a significant portion of the community, and thus were included as additional benthic metrics to explore differences between sites. Richness, taxa diversity, Biological Monitoring Working Party (BMWP), and Average Score Per Taxon (ASPT) indices decreased from Site 1 to 3 and increased at Site 4 (Table 4). Sites 1 and 4 exhibited significantly higher taxa richness compared to Site 3. Site 2 exhibited lower richness than Sites 1 and 4, but higher than Site 3. Taxa diversity was significantly different among sites with the highest value at Site 1 and the lowest at Site 3. Significant differences were observed among sites for the BMWP Index, the ASPT Index, and the EPT/Ch Index (ratio resulting from the sum of Ephemeroptera (mayflies), Plecoptera (stoneflies) and Trichoptera (caddisflies) taxa divided by Chironomidae (non-biting midges) taxa). The three indices scored highest at Sites 1 and 4, and lowest at Sites 2 and 3. Average BAP values were 2.34 at Site 1, 1.62 at Site 2, 1.07 at Site 3, and 2.05 at Site 4 (Fig. 3). Aquatic worms were significantly more abundant at Site 3 compared to the rest of the sites and significantly lowest in abundance at Site 1. Net-spinning caddisflies and non-biting midges were more abundant at Sites 1 and 4 compared to Sites 2 and 3. Gatherers/collectors were significantly more abundant at Sites 3 and 4 compared to Sites 1 and 2, while Parasites presented the opposite trend. In addition, Filterers were more abundant at Sites 1 and 4 compared to Sites 2 and 3 (Table 4). Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 11 Table 4. ANOVA and Kruskall-Wallis test results for the benthic metrics and segregation of Sites according to Post-hoc Tukey’s test. Results include associated levels of significance (p values), type of test performed for each benthic metric, and threshold values for post-hoc control of the false discovery rate (FDR). The “Verification value” column corresponds to the subtraction “p values - FDR threshold for p significance (maximum value for any variable to be considered statistically significant)”. Negative values on the “Verification value” column indicate which variables showed statistically significant values between sites. Superscript letters indicate which Sites were significantly different from each other (e.g., mean BMWP values at sites 1 and 4 are not significantly different between them, but they are significantly different compared to Sites 2 and 3). Sites with two superscript letters are “in-between” groups (e.g., richness of taxa at site 2 did not differ significantly from Richness at sites 1, 3, and 4, but Richness at site 3 was significantly different from Richness at sites 1 and 4). Variables without superscript letters did not show differences between Sites. Benthic metric p values Test Ranking FDR threshold for p significance Verification value Segregation of sites after Tukey´s test Site 1 Site 2 Site 3 Site 4 % Oligochaeta 0.001 ANOVA 1 0.0017 “-0.0009 26.27 ± 11.75c 37.44 ± 16.74b 78.73 ± 35.21a 36.72 ± 16.42b % Chironomidae 0.001 ANOVA 2 0.0038 “-0.0028 36.91 ± 1.45b 24.09 ± 1.82c 19.82 ± 1.76c 56.34 ± 3.18a % Parasites (Nematoda) 0.001 ANOVA 3 0.0057 “-0.0048 30.27 ± 13.54b 36.38 ± 16.27a 0.15 ± 0.07d 2.27 ± 1.01c % Hydropsyche sp. 0.001 ANOVA 4 0.0077 “-0.0067 2.56 ± 0.99a 0.39 ± 0.39b 0.00 ± 0.00b 3.72 ± 0.58a Taxon diversity 0.001 ANOVA 5 0.0096 “-0.0086 1.33 ± 0.59a 1.15 ± 0.51b 0.58 ± 0.26d 0.94 ± 0.42c % Trichoptera 0.001 ANOVA 6 0.0115 0.0105 2.56 ± 0.99a 0.39 ± 0.39b 0.00 ± 0.00b 3.72 ± 0.58a % Diptera 0.001 ANOVA 7 0.0135 “-0.0125 38.44 ± 17.19b 24.22 ± 10.83c 19.85 ± 8.88c 56.77 ± 25.39a % Filterers 0.001 ANOVA 8 0,0154 “-0.0144” 4.09 ± 1.83a 0.45 ± 0.20b 0.03 ± 0.01b 3.91 ± 1.75a Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 12 Table 4. Continued. Benthic metric p values Test Ranking FDR threshold for p significance Verification value Segregation of sites after Tukey´s test Site 1 Site 2 Site 3 Site 4 % Gatherers/ collectors 0.001 ANOVA 9 0.0173 “-0.0163” 63.23 ± 28.28c 61.65 ± 27.57c 98.54 ± 44.07a 93.14 ± 41.66b BMWP 0.001 ANOVA 11 0.0212 “-0.0202” 25.60 ± 11.45a 12.80 ± 5.72b 11.00 ± 4.92b 24.00 ± 10.73a % Tolerant organisms 0.002 ANOVA 12 0.0231 “-0.0211” 94.46 ± 42.24c 99.38 ± 44.44ab 99.91 ± 44.68a 95.88 ± 42.88bc ASPT 0.002 ANOVA 13 0.0250 “-0.0230” 3.34 ± 1.49a 2.27 ± 1.02b 2.30 ± 1.03b 3.02 ± 1.35ab % Turbellaria 0.006 ANOVA 14 0.0269 “-0.0209” 0.03 ± 0.02b 0.17 ± 0.06ab 0.50 ± 0.11a 0.18 ± 0.09ab % Simulium sp. 0.007 ANOVA 15 0.0288 “-0.0218” 1.53 ± 0.69a 0.06 ± 0.06b 0.03 ± 0.03b 0.19 ± 0.09ab Species Richness 0.007 ANOVA 16 0.0308 “-0.0238” 7.6 ± 3.40a 5.40 ± 2.41ab 4.60 ± 2.06b 8.00 ± 3.58a Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 13 Discussion The low taxa richness, as well as their uneven abundance, suggests that the Bronx River is affected by disturbances of high intensity that may frequently “reset” the composition of the invertebrate community (Hussain and Pandit 2012, Lake 2000). If this occurs on a regular basis, the probability of finding short life cycle organisms (e.g., non-biting midges, aquatic worms; Danks 2006, Nicacio and Juen 2015) will be higher than other invertebrates with longer life cycles, such as mayflies and stoneflies (Hussain and Pandit 2012). Past monitoring efforts in the Bronx River support this conclusion (NYSDEC 1998, 2003). Low invertebrate diversity and dominance of the assemblage by tolerant taxa indicate stress not only downstream from the stormwater outfalls in the mid-basin area, but also in North White Plains and in the lower-basin area downstream from the NYBG. Monitoring efforts conducted by NYC Parks´ Natural Resources Group (NRG) between 2002–2004 found invertebrate communities dominated by non-biting midges, sideswimmers, molluscs, and aquatic worms in upper areas of the river, with aquatic worms increasing drastically (52.9% of individuals) in the NYBG section of the River (NRG 2004). Results from a NYSDEC (2015) biological assessment conducted at five River locations in September 2014 were similar to those reported here: macroinvertebrate communities were dominated by moderately to very tolerant taxa (NRG 2004, NYSDEC 2015) and by other fauna usually present in sewage-impacted communities (Mason 2002). Caddisflies, less tolerant to pollution, also made up a significant portion of the community downstream from the NYBG in both 2012 and 2014. Sensitive taxa, such as mayflies, were found neither in 2012 nor in 2014. Figure 3. Biological Assessment Profile (BAP) of index values, sampling sites at the Bronx River (2012). Values are plotted on a normalized scale of water quality. The BAP is the mean of the four values for each site, representing species richness (SR), EPT richness, Family Biotic Index (FBI), and Percent Model Affinity (PMA). Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 14 BAP profiles provided evidence of “severely impacted” water quality at all four study locations, with Sites 2 and 3 displaying the poorest conditions (Fig. 3). Past NRG (2004) monitoring efforts diagnosed water quality as “severely impacted” in the NYBG section of the Bronx River, and “moderately impacted” in the rest of upper, middle, and downstream sites. NYSDEC (1998, 2003, 2015a) biological stream assessments also reported “moderately impacted” water quality at all upper, middle, and downstream sites, with sensitive taxa reduced, and the distribution of major taxonomic groups significantly different from what is naturally expected. The mean abundance of aquatic worms exceeded 10,000/m2 at both Sites 2 and 3, which is also indicative of severe pollution (Rodriguez and Reynoldson 2011). These results suggest chronic poor water quality conditions in the Bronx River causing overall macroinvertebrate diversity to substantially decrease in response. Results of benthic metrics evidenced differences in the structure of the invertebrate community between sites (Table 4). Taxa richness; taxa diversity; combined indices BMWP, ASPT, and EPT/Chironomidae; relative abundances of aquatic Oligochaeta, Chironomidae, parasites, Diptera, Trichoptera and Turbellaria; and relative abundances of most functional feeding groups (Gatherers/collectors, Filterers, Predators, and Grazers/scrapers) were the most informative benthic metrics in determining differences among sites. The other metrics did not provide additional information because they were either redundant with other metrics or failed to detect differences in invertebrate assemblage composition between sites (Table 4). Decreasing invertebrate diversity and increasing relative abundances of aquatic worms occurred from Sites 1 to 3. These findings, along with lower BMWP and ASPT scores, and lower abundance of net-spinning caddisflies, suggest a higher level of disturbance (Walsh et al. 2005, Walters et al. 2009) in the Bronx compared to North White Plains. Our observations corroborated that stormwater enters at a greater frequency and magnitude progressing downstream from Site 1 to 3, which partially explains the extremely low diversity between North White Plains and the NYBG (NYSDEC 2015a). The decrease of non-biting midges from Site 1 to 3 was unexpected. Although many taxa in this group can thrive in highly polluted environments (Simião-Ferreira et al. 2009), others are sensitive to specific pollutants (Al-Shami et al. 2010, Rosa et al. 2014). We observed color changes and swelling in some of the non-biting midges collected in this study, which suggests physiological stress from chemical pollutant exposure (Bailey and Liu 1980). Past studies (Dieter et al. 1996, Sloof 1983) have found evidence of higher chemical tolerance in aquatic worms compared to most invertebrate taxa, including non-biting midges. Aquatic worms also have mechanisms that may allow them to thrive in the most polluted stretches of the Bronx River, including the capacity to regenerate their posterior region when it is lost through toxicant exposure, as well as the ability to reduce bioaccumulation of toxins (Rodriguez and Reynoldson 2011). Additionally, bioturbation by aquatic worms (Matisoff et al. 1999, Wang and Matisoff 1997) may explain decreasing numbers of non-biting midges from Sites 1 to 3. The activity of aquatic oligochaetes through burrowing, feeding, deposition of fecal pellets on the sediment surface, and their respiratory movement, all contribute to the bioturbation effect, which may release contaminants into the water column (Rodriguez and Reynoldson 2011). The presence of greater amounts of fine sediments at Sites 2 and 3 may promote bioturbation by aquatic worms, which may be harmful to the most sensitive nonbiting midges, as well as to other taxa. An increase in diversity at Site 4 compared to Site 3 suggests that improvement of water quality and habitat conditions occurs downstream from the NYBG. One possible explanation is that NYBG is acting as a disturbance buffer (Clinton et al. 2010, DeWalle 2010, Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 15 Hatt et al. 2006, McMillan et al. 2014, NYSDEC 2015a, Wilkerson et al. 2010) and thus increasing habitat quality. Vegetation can enhance nutrient transformations via maintenance of streambank stability, large wood contributions that increase channel complexity, modulation of organic matter sources, and temperature regulation (Dosskey et al. 2010, McMillan et al. 2014). Hydraulic conditions, such as strong water turbulence, may also reduce contaminant load. Additionally, a dam located at 182nd Street, 30 meters upstream from Site 4, may act as a trap for sediments and associated contaminants, thus reducing their downstream flow (Ypsilanti 2016). Reduction of sediment loads reaching Site 4 may ameliorate the decline of taxa sensitive to water pollution (Ku zmanović et al. 2016). BAP results indicate that no differences can be established in terms of water quality designation between sites. Potential buffer effects provided by the NYBG riparian corridor may not be enough to counterbalance the presumably intense stream disturbance that occurs in the mid and lower-basin areas of the River. No groups could be differentiated by means of Tukey’s tests for embeddedness despite that field observations evidenced higher embeddedness at Site 3. One of the five replicates at Site 3 registered an extremely low value of embeddedness compared to the other four replicates, which explains why Tukey was not able to segregate Site 3 from the rest of sites. Highest embeddedness at Site 3 corresponded to the lowest taxa richness and diversity. Fine sediment deposited within river channels can result in altered benthic community structure (Walters et al. 2009), as a direct result of smothering of the substratum and the clogging of interstices (Russell et al. 2017, Wood and Armitage 1999). A combination of body form and borrowing allows tolerant organisms such as aquatic worms, roundworms, and non-biting midges to survive when exposed to fine sediments (Extence et al. 2011). Although pH, conductivity, and turbidity were implicated as potential stressors, none of them were elevated enough to cause a large impact on invertebrates. Little research exists on the impact of elevated pH on invertebrates (WDFW 2009), and existing studies (Bowman and Bailey 1998, Cheng and Chen 2000) only found negative impacts at higher or lower pH levels than those found in this study. Conductivity exceeded 500 μhos/cm at Sites 1, 2, and 3, indicating that the water is not suitable for certain species of freshwater fish or invertebrates (Timpano et al. 2011; USEPA 1997, 2011). High conductivity at Site 4 was likely related to tidal inputs (NYCDEC 2015a). Turbidity values were low, especially at Sites 2 and 3. Turbulent flow downstream from the dam located at 182nd Street may limit settling of suspended sediments (Kondolf et al. 2014), which may explain the higher turbidity observed at Site 4 compared to Sites 2 and 3. The Bronx River is not an especially turbid system compared to other water bodies in the region, like the Wallkill River to the northwest of the study area, or the Esopus Creek in the Catskills, where water may become turbid due to significant stream bank erosion (NYSDEC 2015b). In addition, the weather was moderately dry during the first quarter of 2012 (Utah Climate Center 2018). Reduced stormwater inputs prior to data collection may explain, in part, the low values recorded for some water quality variables. The dominant macroinvertebrate functional feeding group found in the Bronx River was the Gatherers/Collectors. Large amounts of Fine Particulate Organic Matter (FPOM) present in sewer discharges may increase the abundance of Collectors in the River (Moreyra and Padovesi-Fonseca 2015). Other feeding behaviors, such as filtering, were much less common. The presence of Filterers in the study area may be regulated by a trade-off between availability of feeding resources and adequate substrate (Baer et al. 2001). For example, high embeddedness at Sites 2 and 3 may limit the capacity of caddisflies and black flies for settling spin nets and attaching appendages on the substrate, respectively. The functional feeding dominance of Gatherers/Collectors and the scarcity of other functional feeding Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 16 groups in the Bronx River indicate the existence of an unbalanced representation of feeding modes that may affect in complex ways ecosystem services, such as stream nutrient cycling (Baer et al. 2001, Moreyra and Padovesi-Fonseca 2015). Feeding modes such as scraping or shredding were absent at all the study sites. Different environmental constraints may explain the absence of Shredders, including very poor water quality and a small proportion of boulders and cobbles, which may limit the capacity of substrate to retain leaf litter, the main food resource for this group. Absence of Scrapers was unexpected at Sites 2 and 4, since low levels of canopy cover usually favor the presence of these taxa (Cummins and Merritt 1996). Sampling dates may explain this absence, since Scrapers couple their life cycle with the presence of algae, which should be more abundant with increasing temperature and hours of light (Townsend et al. 2008). Reduction of phosphorous resulting from accumulation above the dam (Ypsilanti 2016) located at 182nd Street might affect primary production in Site 4, which can have an impact on certain groups of invertebrates, including Scrapers. The absence of Scrapers at Site 3 may be due to high embeddedness. Fine sediment deposited within the channel of rivers result in mobile substrata, which may constrain algal growth (Davies-Colley et al. 1992, Parkhill and Gulliver 2002). The main cause for adverse disturbance effects on stream invertebrate assemblages was not identified in this study (Herringshaw et al. 2011). Many investigators have recognized the difficulty of identifying causal linkages among environmental stressors and biotic assemblage structure because it is impossible to account for all factors which operate independently and in combination to structure assemblages over variable spatial and temporal scales (Herringshaw et al. 2011, Paul and Meyer 2001). Results indicate that the Bronx River is likely affected by impacts related to both stormwater and riparian urbanization. A combined strategy, including watershed restoration efforts aimed at reducing stormwater runoff volume rates and riparian restoration actions to enhance the biofiltering capacity of the Bronx River, may constitute a sound approach to improve water quality and habitat conditions for the aquatic biota. Future studies could benefit from using the benthic metrics included in the New York State’s standard operating procedure (SOP) for stream biomonitoring (NYSDEC 2014), along with some of the metrics from EPA RBA protocol not used by NYSDEC that have been calculated in this study. Where contamination is severe, as it is the case of the Bronx River, the simplest indices, such as density, richness, the family proportion, and species dominance, are adequate to characterize system response (Rodriguez and Reynoldson 2011). Aquatic worms monitoring should also be considered because of its value in making distinctions at “lower levels” of pollution (Martins et al. 2008, Rodriguez and Reynoldson 2011). Combination indices included in the EPA RBA protocol that may provide complementary information include BMWP and ASPT. BMWP is an index widely used to detect mainly organic pollution (Alba-Tercedor 1996) and, in comparison with other metrics, it may provide better results (Guimarães et al. 2009, Silva et al. 2011). A weakness of the BMWP approach in common with many other score systems is the effect of sampling effort. When samples are taken quickly, ASPT (BMWP Score/ number of taxa) may provide better results than BMWP (Dorji 2016). Therefore, ASPT might be a better fit than BMWP for rapid biomonitoring assessments. Indices not calculated in this study and included in the New York State’s SOP, like the NBI-P (Nutrient Biotic Index – Phosphorus; Smith et al. 2007), may help in assessing relationships between stream nutrient enrichment and macroinvertebrate composition (NYSDEC 2014). Although often considered biologically impaired due to low macroinvertebrate diversity (Walsh et al. 2005, 2012), urban streams such as the Bronx River can exhibit rates Urban Naturalist A. Baladrón and D.J. Yozzo 2020 No. 31 17 of ecosystem functions equivalent to, or higher than, agricultural or reference streams (Bernot et al. 2010, Reisinger et al. 2017). Considering the long history of human impact, modification and degradation, the Bronx River might be a more resilient system than originally thought. Acknowledgments We thank J. Alan Clark, John D. Wehr, and Sarah Whorley for technical guidance, field assistance, and access to laboratory facilities at Fordham University’s Calder Center. 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