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
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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
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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.
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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
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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).
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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
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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
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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
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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
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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).
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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
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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
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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).
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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,
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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
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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
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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. We also thank Damian Griffin
(Bronx River Alliance) for field assistance. Two anonymous reviewers offered valuable comments
on earlier drafts of this paper.
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