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Comparison of fish community within the Blackwater River watershed before and after establishment of Northern Snakehead Channa argus

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Northern Snakehead is an invasive species initially discovered in the Potomac River in 2004, but has since spread to most major river systems of the Chesapeake Bay. In 2012, Northern Snakehead was first reported from the Blackwater River drainage on the eastern shore of Maryland. Fish community surveys were conducted in Blackwater River and Little Blackwater River in 2006 and 2007, before the establishment of Northern Snakehead there. Because of minimal habitat changes owed to protection by Blackwater National Wildlife Refuge, this dataset enabled us to document changes in the fish community that could be attributed to the establishment of Northern Snakehead. We replicated the 2006 and 2007 surveys (pre-Snakehead) over a year from 2018-2019 (post-Snakehead). Over all sampling periods we caught 35 species (32 fish species and 3 invertebrate species) totaling over 50,000 individuals. Of 21 species that were captured both pre- and post-Snakehead, 17 declined in relative abundance with percent reductions ranging from 30%-97%. We found that five of six sites had significantly different fish communities when comparing pre-Snakehead and post-Snakehead surveys. The main difference in fish communities was a reduction in overall biomass of most fish. Species dominance during the post-Snakehead period was significantly higher for both Blackwater and Little Blackwater River. Pre-Snakehead surveys were more evenly distributed and dominated by White Perch, Black Crappie, and Brown Bullhead, while post-Snakehead surveys were less even and dominated by Common Carp and Gizzard Shad. This study is the first to document major shifts in a fish community following establishment of Northern Snakehead. Further investigation into ongoing fish community changes and continued vigilance in minimizing spread and population growth of Northern Snakehead is warranted.
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Comparison of fish community within the Blackwater River
watershed before and after establishment of Northern Snakehead
Channa argus
By: Joshua J. Newhard1 and Joseph W. Love2
1U.S. Fish and Wildlife Service, Maryland Fish and Wildlife Conservation Office, 177 Admiral
Cochrane Dr., Annapolis, MD 21401
2Maryland Department of Natural Resources, Fishing and Boating Services, Freshwater
Fisheries, 580 Taylor Ave. B-2, Annapolis, MD 21401
Summary
Northern Snakehead is an invasive species initially discovered in the Potomac River in
2004, but has since spread to most major river systems of the Chesapeake Bay. In 2012,
Northern Snakehead was first reported from the Blackwater River drainage on the eastern shore
of Maryland. Fish community surveys were conducted in Blackwater River and Little
Blackwater River in 2006 and 2007, before the establishment of Northern Snakehead there.
Because of minimal habitat changes owed to protection by Blackwater National Wildlife Refuge,
this dataset enabled us to document changes in the fish community that could be attributed to the
establishment of Northern Snakehead. We replicated the 2006 and 2007 surveys (pre-
Snakehead) over a year from 2018-2019 (post-Snakehead). Over all sampling periods we caught
35 species (32 fish species and 3 invertebrate species) totaling over 50,000 individuals. Of 21
species that were captured both pre- and post-Snakehead, 17 declined in relative abundance with
percent reductions ranging from 30%-97%. We found that five of six sites had significantly
different fish communities when comparing pre-Snakehead and post-Snakehead surveys. The
main difference in fish communities was a reduction in overall biomass of most fish. Species
dominance during the post-Snakehead period was significantly higher for both Blackwater and
Little Blackwater River. Pre-Snakehead surveys were more evenly distributed and dominated by
White Perch, Black Crappie, and Brown Bullhead, while post-Snakehead surveys were less even
and dominated by Common Carp and Gizzard Shad. This study is the first to document major
shifts in a fish community following establishment of Northern Snakehead. Further investigation
into ongoing fish community changes and continued vigilance in minimizing spread and
population growth of Northern Snakehead is warranted.
Background
Throughout the world, introduced fishes have been a significant source of concern where
they become established (Cucherousset and Olden 2011). The focus of research for introduced
species often includes investigation into control and management methods, determining current
negative impacts to invaded ecosystems, and modeling potential future community and/or
species changes due to the invader (Simberloff et al. 2013). Effects of introduced fishes can vary
from minimal to extreme, such as causing extirpation or extinction of native species. Not all
introductions lead to negative consequences (Cucherousset and Olden 2011; Martin and
Valentine 2019), but generalized impacts can include: reductions in abundance of native fauna,
loss of biodiversity (Gurevitch and Padilla 2004), or alteration of trophic dynamics (Britton et al.
2010, Gallardo et al. 2016, Kramer et al. 2019). As invasive species become abundant in an
area, they could potentially use resources otherwise available to native fishes, which can reduce
biomass and growth of native fish (Carey and Wahl 2010, Hughes and Herlihy 2012, Kramer et
al. 2019). Experimentally, Carey and Wahl (2010) showed that increasing density of invasive
carp reduced native fish growth. In addition to declines in biomass owed to reduced growth,
declines in biomass can occur when the number of individuals declines via predation. Invasive
lionfishes (Pterois volitans and P. miles) in the Western Atlantic have been shown to cause
declines in prey fish biomass and/or their competitors for those prey (Green et al. 2012, Albins
2013). Such changes in biodiversity and biomass contribute to biotic homogenization, which
could reduce resiliency of the ecosystem to natural or human disturbances and significantly alter
the evolutionary trajectory of native species within a community (Olden et al. 2004; Petsch
2016). As introduced fishes expand their range and become more abundant, research
investigating community level impacts have become of greater and broader need.
Northern Snakehead (Channa argus) is a freshwater fish native to parts of Asia and
Russia (Courtenay and Williams 2004). It was first reported from the Potomac River in 2004
(Odenkirk and Owens 2005) and has since spread to nearly every major tributary of the
Chesapeake Bay (Fuller et al. 2019). Its introduction and subsequent spread has caused concern
among state and federal agencies about its impact to native ecosystems. To date, most studies
have focused on potential competition with sportfish in Potomac River (Saylor et al. 2012, Love
et al. 2015) and modeling potential ecological consequences (Love and Newhard 2012). No
empirical research has addressed actual population or community level changes in response to
Northern Snakehead establishment. In addition, the research that has been done has primarily
focused on the population within the Potomac River watershed due to Northern Snakehead’s
initial establishment and population growth there (Odenkirk and Owens 2005, 2007, Odenkirk
and Chapman 2019). Because climatic factors and natural ability have catalyzed the spread of
Northern Snakehead beyond Potomac River and into the Chesapeake Bay watershed (Love and
Newhard 2018), there is a need to continue research beyond Potomac River and into other
ecosystems that differ in habitat type.
Northern Snakehead was first reported from the Blackwater River drainage (eastern
Chesapeake Bay watershed) in 2012 (Fuller et al. 2019). Reports of Northern Snakehead in the
area coincide with their expansion in the eastern Chesapeake Bay watershed from illegal
introductions (King and Johnson 2011, Wegleitner et al. 2016) in Delaware, where the species
was first observed in 2010 (unpublished observation, C. Martin, Delaware Department of Natural
Resources and Environmental Control). Abundance of Northern Snakehead in Blackwater
National Wildlife Refuge (BNWR) and Dorchester County has apparently increased dramatically
since Northern Snakehead was first established and has led to a popular recreational fishery
(Ciekot 2019) that has been supported by state agencies (Love and Genovese 2019). The rapid
increase in abundance of Northern Snakehead in eastern shore rivers somewhat mimics its
increase in abundance following establishment in the Potomac River (Odenkirk and Owens
2007), where a recreational and commercial fishery also now exists (Newhard et al. 2019).
Prior to the introduction, establishment and rapid population growth of Northern
Snakehead on the eastern shore of Maryland, fish community surveys in the Blackwater River
drainage were conducted to inventory fish communities from various habitats that differed in
salinity and position in the stream gradient (Love et al. 2008; Newhard et al. 2012). When long-
term monitoring data is unavailable, comparison of current community data with historical data
is a recommended strategy (Gallien and Carboni 2017). Because of minimal habitat changes
owed to protection by BNWR, this dataset enabled us to document changes in the fish
community that could be attributed primarily to the establishment of Northern Snakehead. Our
objective for this study was to compare the fish community before and after Northern Snakehead
introduction for sites in Blackwater River and Little Blackwater River.
Methods
Fish collection
We sampled three sites each within Blackwater River and Little Blackwater River (Fig.
1), areas that were inventoried before Northern Snakehead was introduced into the Blackwater
River drainage (Love et al. 2008, Newhard et al. 2012). Similar to our reference study (Love et
al. 2008), sites were sampled monthly from June 2018 to May 2019 (hereafter, post-Snakehead).
Prior data were available for most sites for two years, in 2006 and 2007 (hereafter, pre-
Snakehead), with the exception being Buttons Creek where data were only collected in 2007.
Analyzing data for these two years afforded the opportunity to examine annual variability in
community structure. Not all sites could be sampled during winter due to icing of the river, and
no sites were sampled during January 2019 due to a federal government shutdown. At each site,
fyke nets (1.25 cm mesh, 15.2 m lead line) were set for approximately 24 hours. Net set and pull
times were recorded to standardize catch by hours fished (i.e., catch per unit effort or CPUE).
Upon retrieving the nets, all fish were identified to species and individually counted. Any
Northern Snakeheads captured were measured for total length and placed on ice for later
dissection. Water quality was measured at each sampling event using a YSI Pro 2030. Water
quality variables included: temperature (℃), dissolved oxygen (DO; mg/L), salinity (psu) and
ambient conductivity (µs).
Analysis
We calculated CPUE for each species by dividing monthly abundances by number of net
hours fished. We used monthly CPUE as a measure for species’ relative abundances. Because
biologists had wanted to identify specific times for habitat use by anadromous fish, surveys were
conducted weekly during spring in 2006 and 2007. To help facilitate comparisons, we averaged
CPUE across weeks for each month in spring 2006 and 2007 when multiple weeks within a
month were sampled. In addition to relative abundance, we also analyzed the frequency of
occurrence among months for captured species. The number of months when a species was
present at a site was divided by the total number of months sampled (i.e., 12) to calculate the
percentage of months the species was present both pre-Snakehead and post-Snakehead for all
sites within the Blackwater River and Little Blackwater River.
To summarize and analyze data for relative abundances of all species caught within a site
over time, we used non-metric multidimensional scaling (NMS). Each sample of speciesrelative
abundances for a site was assigned a site score using NMS. Site scores were calculated using
Sӧrenson (Bray-Curtis) distances that measured dissimilarity in ranked, relative abundances of
species among monthly surveys. The NMS analysis is a multidimensional ordination analysis
that ranks and places site scores on k-axes in order to reduce stress of the final k-dimensional
configuration (McCune and Grace 2002). Final stress values ranging between 10-20 are
common with community data and are generally considered acceptable (McCune and Grace
2002). We used 250 runs of our data and 250 Monte Carlo simulations to assess the significance
of obtaining an equal or lesser stress value than our observed final stress value. The final k-
dimensional configuration was used to determine whether site scores produced pre-Snakehead
differed from those produced post-Snakehead.
We learned which species were responsible for changes in aquatic communities between
periods by conducting a Spearman’s rank correlation analysis on axis site scores and total
monthly relative abundance of species at a site for: all fish; ubiquitous sunfish species [Black
Crappie (Pomoxis nigromaculatus), Bluegill (Lepomis macrochirus), Pumpkinseed (L.
gibbosus)]; White Perch (Morone americana); Northern Snakehead; Gizzard Shad (Dorosoma
cepedianum); Common Carp (Cyprinus carpio); and Brown Bullhead (Ameiurus nebulosus). We
also calculated Shannon-Weiner diversity indices for each site and monthly sample to determine
if changes in NMS scores were correlated with changes in species diversity. Because water
quality differences between periods may also influence community structure, environmental data
(water temperature, dissolved oxygen, and salinity) were correlated with site scores for all fish
using Spearman’s rank correlations. Owed to relatively weak correlations (<0.50) of the
variables and site scores within a year, environmental variables were not considered important
predictors of changes in community structure. Using results from the reduced species list, we
interpreted axes produced from each site-specific NMS.
We tested the hypothesis that aquatic communities significantly differed between pre-
Snakehead and post-Snakehead periods using a multi-response permutation procedure (MRPP).
The MRPP is a non-parametric analysis method designed to test the null hypothesis of no
differences between communities of species (McCune and Grace 2002). The level of difference
in between communities was computed using Sӧrenson distance. The analysis provided a test
statistic (T), a measure of “effect size” (A), and a p-value. The T is the difference between
observed and expected within-group distances among scores, divided by the square-root in
variance of within-group distances. The A was measured by the chance-corrected within-group
agreement, which describes within group homogeneity compared to a random expectation. The
random expectation was determined by permuting the matrix of relative abundance and assigning
each to either a pre-Snakehead or post-Snakehead group, then determining the mean distance in
scores for each group. Species-specific monthly CPUE data were used to compare fish
assemblages pre-Snakehead to post-Snakehead. All NMS and MRPP analyses were conducted
using PC-ORD (Version 7.07, McCune and Mefford 2018). Spearman’s Rank correlations were
conducted using SYSTAT for Windows (Version 13.0, SYSTAT Software Inc.).
We used rarefaction to compare two measures of species diversity, species richness (i.e.,
the number of different species) and dominance (i.e., the fraction represented by the most
common species), between pre-Snakehead and post-Snakehead periods for Blackwater River and
Little Blackwater River. Because measurements of species diversity can increase with increasing
abundance in the sample, we used rarefaction to compare species diversity at the lowest,
common abundance (Sanders 1968; Hurlbert 1971). Individuals were drawn at random from
each community of organisms until reaching a user-specified number of individuals (i.e., lowest,
common abundance). Diversity measures were then computed for that level of abundance. This
process was iterated 1000 times in order to generate a mean and variance for each diversity
metric. When iterations were finished we compared metrics for the least abundant community to
the 95% confidence interval generated for more abundant communities to determine if diversity
significantly differed between the communities. Rarefaction was performed using EcoSim
(Version 7.0; Gotelli and Entsminger 2001).
Results
Across all survey years we captured 35 species (32 fish species and 3 invertebrate
species; Table 1) totaling 51,781 individuals. Prior to the introduction of Northern Snakehead,
White Perch, Brown Bullhead, and Black Crappie were the most abundant species (in rank
order). In fact, the species that were 90% of annual catch, were the same between 2006 and
2007 (White Perch, Brown Bullhead, Black Crappie, Pumpkinseed, and Bluegill). The rank order
of dominant species also did not change between 2006 and 2007, indicating a fairly stable
assemblage of aquatic organisms between years. Furthermore, site scores generated for those
two years overlapped in distribution for most sites surveyed here (Figure 2), with a notable
exception at LB1. Change in White Perch abundance is likely the driver of differences at LB1
between 2006 and 2007, where abundances dropped from 3,296 individuals in 2006 to 486 in
2007. After the establishment of Northern Snakehead, the three most abundant species (in rank
order) were Common Carp, Gizzard Shad, and White Perch. Additionally, there were nine
species captured in the pre-Snakehead period that were not captured in the post-Snakehead
period, while there were 5 species captured post-Snakehead that were not captured pre-
Snakehead (Table 1). Water quality conditions were generally similar between the two, though
salinity was generally lower for 2018 and 2019 then it was for 2006 or 2007, especially for
Blackwater River and Buttons Creek (Table 2).
During post-Snakehead surveys we captured 125 Northern Snakehead in fyke nets (Table
1). They were encountered at the same rate (45% of months sampled; Table 4) in both
Blackwater River and Little Blackwater River, though they were more abundant in Little
Blackwater River (N=110) than Blackwater River (N=15). Total length range of all Northern
Snakehead captured was 121-680 mm (Mean=351, SD ±78). Of 31individuals that were
dissected for gut contents, 77% had empty stomachs. Prey items from remaining stomachs
included (in rank order): Gizzard Shad, Bluegill, American Eel (Anguilla rostrata), dragonfly
larvae, and an unidentifiable fish species.
In general, relative abundances of most species were lower than they were pre-Snakehead
(Table 3). Of all 35 species observed pre-Snakehead, 26 of them either had lower relative
abundances or were not collected. The largest declines in relative abundance were observed for
White Perch, Brown Bullhead, Atlantic Silverside (Menidia menidia), and Black Crappie. Of the
21 species that were captured in both pre-Snakehead and post-Snakehead periods, 17 species
exhibited declines in average relative abundance. Percent declines in relative abundance of those
species ranged from 30% to 97% (Average=58%, SD±20). There were 8 species that increased
in relative abundance post-Snakehead, with only 3 of those being captured in both survey periods
(Atlantic Menhaden (Brevoortia tyrannus), Gizzard Shad, and Common Carp). Common Carp
had the largest increase in relative abundance (286%), followed by Atlantic Menhaden (78%),
Bay Anchovy (Anchoa mitchilli) (72%), and Gizzard Shad (53%). These changes in relative
abundance resulted in a modest change in the species that were 90% of the annual catch during
2018-2019. These species were Common Carp, Gizzard Shad, White Perch, Pumpkinseed,
Black Crappie, and Bluegill.
Killifishes were less frequently collected post-Snakehead than pre-Snakehead (Table 4).
During the pre-Snakehead period, three species of killifish had been collected: Banded Killifish
(Fundulus diaphanus), Mummichog (F. heteroclitus), and Striped Killifish (F. majalis). Two of
these species were not collected post-Snakehead and Mummichog was only collected in Little
Blackwater River with a frequency that declined from 16.7% of surveys to 9.1% of surveys. The
frequencies in collecting most other species were similar between pre-Snakehead and post-
Snakehead periods. In some cases, the prevalence of species increased. In Blackwater River, for
example, the prevalence of diadromous fish such as river herring (Alosa spp.), Yellow Perch
(Perca flavescens), and American Eel increased by at least 11%. However, this was not true for
Little Blackwater River where prevalence for these species decreased by at least 22%.
Examination of assemblages at each site using NMS were generally summarized by two-
dimensional solutions and had final stress values within acceptable range (12.0-16.6, Table 5).
For Buttons Creek (site BC), a three dimensional solution was preferred, but we opted to only
interpret axes 1 and 2 to ease interpretation. Final stress for the two-dimensional solution was
still within acceptable range (Table 5). While there was some separation of monthly site scores
along axes 1 and 2 for Buttons Creek, there was only a marginally significant difference in fish
communities before and after Northern Snakehead introduction (p = 0.06; A = 0.052; Table 5).
Buttons Creek scores for NMS axis 1 increased from pre-Snakehead to post-Snakehead and were
positively correlated to sunfish and White Perch abundance (Table 6, Fig. 2). The relative
abundance of sunfish and White Perch increased over time in Buttons Creek. While NMS axis 2
for Buttons Creek was most correlated to total fish abundance, no strong patterns in score values
were evident for this axis.
Fish communities at all other sites significantly differed between pre-Snakehead and
post-Snakehead periods (Table 5). Differences in fish communities were largely driven by
differences in overall abundance and changes in abundance of dominant species. The NMS Axis
1 scores were strongly correlated to overall abundance (Table 6, Fig. 2), with the exception of
BC. Sites sampled during the post-Snakehead period had species with lower overall abundances
compared to the pre-Snakehead period. The NMS Axis 2 scores were differently correlated with
species relative abundances among sites, but were generally associated with changes in
abundance of dominant species. The NMS Axis 2 scores for BW1 and BW2 were most strongly
correlated to monthly relative abundance of White Perch. At LB1, NMS Axis 2 correlated to
relative abundances of White Perch and Gizzard Shad, while axis 2 scores for LB2 were
correlated to relative abundance of White Perch and Brown Bullhead. The strongest correlation
for axis 2 scores at LB3 was for Shannon-Weiner diversity index (Table 6). At BW1, BW2, LB1,
and LB2, relative abundances of dominant species were significantly lower after establishment
of Northern Snakehead. And at LB3, diversity increased after establishment of Northern
Snakehead.
After standardizing sample size to a common abundance via rarefaction, we learned that
species richness for Blackwater River was significantly higher before the introduction of
Northern Snakehead (Fig. 3). Species richness pre-Snakehead averaged 26.7 (95% CI: 25 – 27),
which decreased to 21 post-Snakehead (lowest common abundance = 4,602). However, there
was no significant difference in species richness for Little Blackwater River following Northern
Snakehead establishment (Fig. 3). The number of species post-Snakehead (23) at the lowest
common abundance of 3,313, was not statistically different than the number of species estimated
for the pre-Snakehead period (average = 22.0, 95% CI: 19 – 25). Species dominance was
significantly higher for the post-snakehead period for both rivers, partially owed to dominance
by Common Carp and Gizzard Shad. Blackwater River had the largest change in species
dominance, which increased from 0.38 (95% CI: 0.38 – 0.39) during the pre-Snakehead period to
0.60 (lowest common abundance = 4,602) during the post-Snakehead period. A similar, but less
dramatic increase in dominance was observed for the fish community in Little Blackwater River.
Dominance increased from 0.28 (95% CI: 0.20 – 0.29) during the pre-Snakehead period to 0.32
(lowest common abundance = 3,313) during the post-Snakehead period.
Discussion
We found significant changes in aquatic community structure for fish and invertebrate
fauna in the Blackwater River drainage since the introduction and establishment of Northern
Snakehead. These changes were evidenced by both significant differences in ranked abundance
and relative abundance for multiple species, with differences leading to measurable differences
in fundamental attributes of species diversity. These differences can be explained by the
introduction of a top predator (Northern Snakehead) and the installation of a water control
structure and fresher water. Major reductions occurred for the overall abundance of several fish
species after Northern Snakehead was introduced and became abundant. There was also a
reduction in abundance of dominant species, and a shift in the dominant species that make up the
fish community. Surveys completed in 2006 and 2007, before Northern Snakehead were known
to inhabit waters on the eastern shore of Maryland, were dominated by an abundance of White
Perch, Brown Bullhead, and a few sunfish species (F. Centrarchidae). Additional species, such as
Banded Killifish, had been frequently caught before snakeheads were introduced, but were not
observed in 2018 and 2019, suggesting a much reduced relative abundance and/or distribution.
Banded Killifish, in addition to sunfish and perch, is a principal prey item of Northern
Snakeheads in Potomac River (Saylor et al. 2012, Isel and Odenkirk 2019, Lapointe et al. 2019).
While not found in guts during our study, killifish species (Fundulus) were the numerically
dominant prey item in Northern Snakehead guts sampled from Little Blackwater River and
nearby rivers on the eastern shore of Maryland (J. Thompson, Maryland Department of Natural
Resources, unpublished data). The loss of these prey species could be in part due to presence of
an additional top predator like Northern Snakehead.
In general, results of our surveys are similar to other assessments of community level
impacts of aquatic invasive species. One of the most pronounced differences we observed was in
the reduction in abundance of many fish and invertebrate species. Invasive fishes can cause
reductions in biomass where they become established (Albins 2013, Carey and Wahl 2010,
Gallardo et al. 2016). The observed reductions in relative abundance of most species is likely
why we observed a significant difference in species dominance for both Blackwater and Little
Blackwater Rivers. Reductions in abundances and loss of native species can lead to biotic
homogenization of fish communities (Rahel 2007). Introduced top-predators (such as Northern
Snakehead) have been shown to be drivers of such homogenization and may facilitate successful
future invasions of other species (Daga et al. 2015, Gallien and Carboni 2016). Bezerra et al.
(2019) demonstrate that introduced Largemouth Bass (a top-predator) and African Cichlids
(omnivore-detritivore) in neotropical lakes and reservoirs caused trophic downgrading and biotic
homogenization of fish communities. While it is too early to determine if such effects are
occurring in Blackwater River and Little Blackwater River, similarities exist with the
establishment of Northern Snakehead alongside an apparently abundant population of Common
Carp, an introduced detritivore. We observed fewer species post-Snakehead, along with
significantly reduced biomass of most fish, at minimum suggesting homogenization may be
occurring.
There was a significant difference in species richness for Blackwater River, but this may
not indicate extirpation of the species from the drainage. One potential reason species richness
was lower during the post-Snakehead period is the lower salinities observed in Blackwater River,
likely a result of a water control structure. There was a loss of some euryhaline species (e.g.,
Spot (Leiostomus xanthurus); Atlantic Silverside (Menidia menidia)) in Blackwater River in
2018-2019. The highest recorded salinity in Blackwater River in 2006 or 2007 was 14.7 psu,
while the maximum salinity observed was 6.2 psu during 2018-2019. The U.S. Fish and Wildlife
Service completed a water control structure in upper portions of the Blackwater River in 2007
for the purpose of limiting saltwater intrusion into the watershed (Love et al. 2008). There was
also significant rainfall during much of 2017 and 2018, which likely led to reduced salinities in
Blackwater River. The fresher water may explain the increased prevalence of diadromous fishes
in Blackwater River. Because similarly greater prevalence was not observed in Little Blackwater
River, it is possible that the observed increase in Blackwater River is not due to increased
precipitation but is instead the result of the water control structure.
In addition to a loss of euryhaline species, there was also a decline in freshwater-
dependent species in Blackwater River, despite the greater availability of freshwater. Unlike
during the pre-Snakehead period, we did not collect Largemouth Bass, Redfin and Chain
Pickerel (Esox americanus and E. niger), or Channel Catfish (Ictalurus punctatus). These species
were not numerically abundant in surveys completed prior to Northern Snakehead introduction,
suggesting some gear avoidance or low catchability relative to other species. However, based on
percent occurrence data, we determined that these species were encountered occasionally during
the pre-Snakehead period and it is unknown why these species were not collected during the
post-Snakehead period. Anglers reportedly caught Largemouth Bass and Channel Catfish in
Blackwater River drainage in 2018-2019. It is possible that these species currently have a lower
relative abundance or a different distribution than a decade ago.
Species dominance significantly differed between survey periods. Comparison of the fish
community before Northern Snakehead were present showed high abundances (>1,000
individuals) of several species captured within a year, indicating a more even assemblage of
fishes. However, surveys conducted after Northern Snakehead introduction showed that fewer
species had high abundances (Common Carp, Gizzard Shad, and White Perch) along with much
lower catches of some previously abundant species, such as Brown Bullhead, Black Crappie, and
Bluegill. The reductions in abundance of these species likely explains lower evenness and greater
dominance. Invasive fishes have been shown to reduce abundances of smaller bodied native
fishes whether through competition, direct predation, or habitat alteration (Albins 2013, Pelicice
et al. 2015, Kramer et al. 2019). The potential alteration of the fish community within the
Blackwater River drainage warrants further investigation. If native fish diversity changes or is
already low, an invasive fish may have greater success within an ecosystem (Carey and Wahl
2010).
This study compared two time periods that bracketed an intense community disturbance,
the establishment of Northern Snakehead, in Blackwater River drainage. Currently Blackwater
River drainage’s fish community is dominated by Common Carp and Gizzard Shad. While White
Perch and centrarchids remained relatively abundant in the post-Snakehead period, their biomass
was significantly lower than during the pre-Snakehead period. Because there was no control or
reference system for this analysis, we acknowledge that other factors may have caused changes
in abundance and species diversity. Two common ecological factors that could affect aquatic
communities include habitat changes and natural, interannual fluctuations in abundance of
individual species. Our study periods were separated by over a decade, when habitat conditions
and land use changes in the watershed could have altered the fish fauna. However, with the
exception of the aforementioned water control structure, habitat conditions have remained
largely unchanged during the period between surveys. Land use was relatively unchanged for
Dorchester County between 2002 and 2010 (Dorchester County 2019). A significant portion of
the Blackwater River watershed is owned and managed by the U.S. Government and no new
land was developed. Additionally, the watershed of the Little Blackwater appears to be largely
unchanged with no new significant changes over the last decade (M. Whitbeck, Supervisory
Wildlife Biologist, BNWR, personal comment). Also, two sites on the upper Little Blackwater
River (LB1, LB2) had significantly different fish communities over time, and water quality was
not different in these mainly freshwater habitats. With the exception of salinity, habitat
measurements (dissolved oxygen and temperature) recorded during the post-Snakehead period
were generally not different from the pre-Snakehead period.
Changes in habitat conditions other than water quality and land use can affect abundance
of species. de Mutsert et al. (2017) monitored the fish community in a Potomac tributary for
over 30 years, which included several years with an established Northern Snakehead population.
Banded Killifish increased in abundance over time and the increase was attributed to an increase
of submerged aquatic vegetation (SAV) over time in their study area. Because SAV is not
abundant in Blackwater and Little Blackwater River (J. Newhard pers. obs.; Orth et al. 2018), the
growth of SAV is unlikely to influence the abundance of aquatic species in Blackwater River.
Though their study was not designed to study the effects of Northern Snakehead, de Mutsert et
al. (2017) also observed decreases over time for White Perch and Bluegill, as well as an increase
in Gizzard Shad. Moreover, their results show an increase in species dominance over time,
especially during the period with Northern Snakehead presence (Figure 6 in de Mutsert et al.
(2017)). Differences in their study cannot necessarily be attributed Northern Snakehead alone, as
habitat changed, and other introduced species such as Blue Catfish (Ictalurus furcatus) increased
in abundance during that time. In another study, Isel and Odenkirk (2019) found that relative
abundance of Bluegill (another common prey item of Northern Snakehead) was unchanged in
one recently invaded Virginia lake with Northern Snakehead. Of critical importance when
comparing results of studies is referencing density of Northern Snakehead and length of time
since establishment of Northern Snakehead, as well as the resiliency of the aquatic community.
Our study and the studies referenced here demonstrate that impacts to fish communities where
Northern Snakehead are established may vary, especially as aquatic communities respond to
biotic and anthropogenic factors alongside establishment of an invasive species.
Fish community changes observed in our study could be due to other processes not
measured. For example, nutrient availability within an ecosystem can alter community dynamics
separately from, or in conjunction with invasive species (Preston et al. 2018). There are also
other introduced species present in Blackwater River drainage that can alter their environment.
Common Carp are considered “ecosystem engineers” that alter native habitats through
consumption of aquatic vegetation which can influence primary production, water clarity and
nutrient cycling, among others (Carey and Wahl 2010, Cucherousset and Olden 2011). Common
Carp were present in 2006 and 2007 surveys at consistent levels (7th most abundant in both
years) and were the most abundant species in 2018-2019 surveys. While the reason for the
increase in relative abundance of Common Carp is not known, it could be explained by a change
in commercial fishing pressure in Blackwater River drainage whereby 5-times the number of
carp were harvested between 2006 and 2008 than 2016 and 2018 (unpublished data, Maryland
Department of Natural Resources). The addition of Northern Snakehead to the fish fauna as well
as Common Carp could have additive effects that cause reductions in abundance and/or species
diversity (Didham et al. 2007).
Continued monitoring of the fish community in Blackwater River and Little Blackwater
River is warranted, especially if Northern Snakehead abundance continues to increase. Our
results support well-established research that new and abundant aquatic predators can cause
changes in aquatic communities, particularly via predation. They also further support risk
assessments (Courtenay and Williams 2004) that Northern Snakehead is an invasive species that
should be controlled. Management of Northern Snakehead should be a top priority to slow
population growth in Blackwater and Little Blackwater Rivers, as they are important nursery
habitats for resident and migratory fish species (Love et al. 2008, Newhard et al. 2012).
Freshwater nursery habitats within the drainage have already been reduced in size due to sea
level rise (Rogers and McCarty 2000) and may continue with climate change, which can also
increase the spread of aquatic invasive species (Rahel and Olden 2008). Control options for
Northern Snakehead are somewhat limited, especially in Blackwater and Little Blackwater River
where electrofishing (one of the main control methods in many locations) is difficult, thus
limiting managers to traditional fish collection methods such as nets and traps which are not as
efficient at capturing Northern Snakehead. Additional control mechanisms could include
recreational and commercial fishing operations, though these fishing sectors likely need to
remove a significant portion of the population to cause noticeable declines (Newhard et al.
2019).
There are benefits to controlling the biomass and spread of invasive species. Reducing
biomass of an invasive species can cause increases in biomass of native fish fauna (Abekura et
al. 2004). Removal rates should be sufficiently high in order to see desirable outcomes (Kramer
et al. 2019). While not measured here, the current fishing mortality in areas of the Potomac River
did not cause significant declines in populations of Northern Snakehead (Newhard et al. 2019).
However, modeling work has shown that reduction of Northern Snakehead adults can lead to
reduced reproduction and potentially smaller populations (Hoff and Odenkirk 2019). Efforts to
increase mortality of Northern Snakehead will likely need to be taken within Blackwater and
Little Blackwater River, where the fishery is relatively new and fishing mortality is therefore
assumed to be low. Such effort can be limited by public access and public apathy, thereby
requiring effort from public agencies to achieve any established biomass reduction targets.
Successful efforts may not only increase biomass of native fish fauna, but could also buffer
against ecosystem changes and limit impacts of other invasive species (Carey and Wahl 2010).
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Table 1. List of all species captured at sites within Blackwater and Little Blackwater Rivers from
2006, 2007, and 2018-2019. Species with superscript “A” were captured during surveys before
snakehead introduction (2006 & 2007) but not after (2018-2019). Species with superscript “B”
were captured post-snakehead introduction but not before.
Common Name
Scientific Name
2006
2007
2018-2019
American Eel
Anguilla rostrata
88
41
33
Atlantic Menhaden
Brevoortia tyrannus
8
- -
31
Atlantic Silverside
Menidia menidia
115
66
2
Banded KillifishA
Fundulus diaphanus
169
3
- -
Bay Anchovy
Anchoa mitchilli
- -
1
2
Black Crappie
Pomoxis nigromaculatus
3,329
2,665
750
Blue Crab
Callinectes sapidus
370
404
105
Bluegill
Lepomis macrochirus
2,127
1,955
450
Bluespotted SunfishB
Enneacanthus gloriosus
- -
- -
1
Brown Bullhead
Ameiurus nebulosus
4,365
3,776
282
Chain PickerelA
Esox niger
- -
1
- -
Channel CatfishA
Ictalurus punctatus
55
1
- -
Common Carp
Cyprinus carpio
812
681
2,900
Crayfish spp.A
5
4
- -
Eastern Mudminnow
Umbra pygmaea
2
6
1
Gizzard Shad
Dorosoma cepedianum
891
978
1,171
Golden Shiner
Notemigonus crysoleucas
21
17
1
Hogchoker
Trinectes maculatus
5
53
3
Inland Silverside
Menidia beryllina
9
- -
1
Largemouth BassA
Micropterus salmoides
14
19
- -
Longear SunfishA
Lepomis megalotis
21
- -
- -
Mud Crab spp.B
- -
- -
1
Mummichog
Fundulus heteroclitus
2
38
1
Northern SnakeheadB
Channa argus
- -
- -
125
Pumpkinseed
Lepomis gibbosus
3,227
1,310
774
Redbreast SunfishB
Lepomis auritus
- -
- -
4
Redfin PickerelA
Esox americanus
12
4
- -
River Herring
Alosa aestivalis/pseudoharengus
20
10
7
Silver PerchB
Bairdiella chrysoura
- -
- -
2
SpotA
Leiostomus xanthurus
2
6
- -
Striped Bass
Morone saxatilis
4
- -
1
Striped KillifishA
Fundulus majalis
10
65
- -
White Catfish
Ameiurus catus
26
139
6
White Perch
Morone americana
11,538
4,548
1,039
Yellow Perch
Perca flavescens
27
7
16
Table 2. Summary of yearly water quality measurements from all sites sampled within
Blackwater River and Little Blackwater River. Measurements below indicate average (min-max)
and include: water temperature (), dissolved oxygen (DO; mg/L), and salinity (psu).
Site
Water Temp.
DO
Salinity
BC
19.7 (4.5-30.1)
8.8 (4.5-13.8)
5.1 (0.1-11.3)
17.2 (3.9-29.1)
8.0 (4.7-11.2)
0.84 (0.1-3.1)
BW1
18.3 (5.7-25.0)
7.9 (4.0-13.7)
11.8 (7.8-14.3)
19.2 (7.1-28.5)
7.5 (4.1-12.5)
4.8 (2.6-12.9)
18.2 (4.5-28.9)
7.4 (4.2-11.6)
2.9 (0.8-6.2)
BW2
18.0 (3.7-31.2)
7.8 (2.6-13.5)
9.0 (3.4-14.7)
17.6 (4.0-29.3)
6.4 (3.4-10.5)
3.0 (0.8-9.2)
17.0 (3.1-29.1)
6.6 (2.2-11.2)
1.5 (0.6-3.9)
LB1
16.1 (5.2-29.3)
6.5 (0.4-11.8)
0.2 (0.04-0.7)
16.4 (1.9-28.7)
6.1 (1.6-9.5)
0.5 (0.0-3.5)
16.8 (5.6-26.4)
5.4 (0.5-10.4)
0.1 (0.0-0.1)
LB2
19.3 (6.6-29.2)
6.7 (1.8-11.0)
0.7 (0.1-1.5)
17.3 (3.4-29.3)
7.3 (3.4-13.8)
0.9 (0.0-5.2)
18.5 (6.6-28.6)
5.7 (0.7-10.0)
0.1 (0.0-0.1)
LB3
16.9 (7.2-27.4)
9.6 (5.7-13.4)
5.0 (0.4-9.5)
17.4 (2.3-30.2)
9.0 (6.1-13.0)
4.0 (0.1-13.2)
18.3 (3.8-31.6)
9.1 (6.5-11.2)
1.0 (0.2-3.3)
Table 3. Monthly catch per-unit-effort (fish per hour) for all species captured at all sites sampled
before Northern Snakehead introduction (2006 and 2007) compared to after Northern Snakehead
introduction (2018-2019).
Species
Pre
Post
Difference
American Eel
0.15
0.08
-0.07
Atlantic Menhaden
0.18
0.32
0.14
Atlantic Silverside
1.41
0.04
-1.37
Banded Killifish
0.81
-0.81
Bay Anchovy
0.05
0.08
0.03
Black Crappie
1.68
0.63
-1.05
Blue Crab
0.43
0.24
-0.19
Bluegill
1.18
0.4
-0.78
Bluespotted Sunfish
0.05
0.05
Brown Bullhead
2.27
0.35
-1.92
Chain Pickerel
0.05
-0.05
Channel Catfish
0.15
-0.15
Common Carp
0.62
2.39
1.77
Crayfish
0.08
-0.08
Eastern Mudminnow
0.12
0.04
-0.08
Gizzard Shad
0.81
1.24
0.43
Golden Shiner
0.11
0.04
-0.07
Hogchoker
0.16
0.07
-0.09
Inland Silverside
0.1
0.04
-0.06
Largemouth Bass
0.05
-0.05
Longear Sunfish
0.15
-0.15
Mud Crab
0.04
0.04
Mummichog
0.32
0.04
-0.28
Northern Snakehead
0.24
0.24
Pumpkinseed
1.51
0.99
-0.52
Redbreast Sunfish
0.06
0.06
Redfin Pickerel
0.1
-0.1
River Herring
0.08
0.05
-0.03
Silver Perch
0.08
0.08
Spot
0.05
-0.05
Striped Bass
0.06
0.04
-0.02
Striped Killifish
1.02
-1.02
White Catfish
0.43
0.04
-0.39
White Perch
4
0.86
-3.14
Yellow Perch
0.09
0.06
-0.03
Table 4. Monthly percent frequency of occurrence of species captured in fyke nets from
Blackwater and Little Blackwater Rivers before Northern Snakehead introduction (2006 and
2007) and after Northern Snakehead introduction (2018-2019).
Species
Blackwater
Little Blackwater
Before
After
Before
After
American Eel
0.500
0.636
0.833
0.455
Atlantic Menhaden
0.167
0.182
0.000
0.091
Atlantic Silverside
0.250
0.000
0.167
0.182
Banded Killifish
0.250
0.000
0.083
0.000
Bay Anchovy
--
--
0.083
0.091
Black Crappie
0.917
1.000
1.000
1.000
Blue Crab
0.750
0.818
0.750
0.545
Bluegill
0.917
0.909
1.000
1.000
Bluespotted Sunfish
0.000
0.091
--
--
Brown Bullhead
0.917
0.727
1.000
0.909
Chain Pickerel
0.083
0.000
--
--
Channel Catfish
0.083
0.000
0.667
0.000
Common Carp
0.667
1.000
1.000
1.000
Crayfish
--
--
0.333
0.000
Eastern Mudminnow
--
--
0.250
0.091
Gizzard Shad
0.583
0.818
1.000
1.000
Golden Shiner
0.250
0.091
0.583
0.000
Hogchoker
0.417
0.091
0.167
0.091
Inland Silverside
0.167
0.091
0.083
0.000
Largemouth Bass
0.500
0.000
0.750
0.000
Longear Sunfish
0.250
0.000
0.167
0.000
Mud Crab
--
--
0.000
0.091
Mummichog
0.250
0.000
0.167
0.091
Northern Snakehead
0.000
0.455
0.000
0.455
Pumpkinseed
0.917
1.000
1.000
0.727
Redbreast Sunfish
0.000
0.182
--
--
Redfin Pickerel
0.167
0.000
0.333
0.000
River Herring
0.167
0.273
0.500
0.273
Silver Perch
--
--
0.000
0.091
Spot
0.250
0.000
0.250
0.000
Striped Bass
0.083
0.000
0.167
0.091
Striped Killifish
0.167
0.000
0.083
0.000
White Catfish
0.167
0.182
0.583
0.364
White Perch
1.000
1.000
1.000
1.000
Yellow Perch
0.167
0.545
0.667
0.273
Table 5. Stress values from non-metric multidimensional scaling analysis (NMS Stress) and
multi-response permutation procedure results (MRPP A) from comparison of fish communities
before (2006 and 2007) and after (2018-2019) Northern Snakehead introduction at all sites in
Blackwater River and Little Blackwater River.
Site
NMS Stress
MRPP A
p-value
BC
15.8
0.052
0.065
BW1
10.6
0.152
0.000
BW2
13.3
0.042
0.043
LB1
12.0
0.058
0.013
LB2
13.3
0.061
0.020
LB3
16.6
0.050
0.013
Table 6. Spearman’s Rank correlations for non-metric multidimensional scaling (NMS) axis
scores with total monthly fish abundance (ALL), Shannon-Weiner diversity index (H’), and
monthly abundances for White Perch (WP), Northern Snakehead (NSH), Gizzard Shad (GS),
Common Carp (CC), all Centrarchidae spp. (SUN), and Brown Bullhead (BB).
NMS
Axis
Site
ALL
H'
WP
NSH
GS
CC
SUN
BB
1
BC
0.748
0.039
0.54
-0.022
-0.238
-0.454
0.925
0.34
BW1
0.665
0.265
0.059
0.525
0.539
0.253
0.915
0.25
BW2
-0.978
0.164
-0.822
0.151
-0.184
-0.041
-0.784
-0.432
LB1
0.957
-0.25
0.416
-0.215
0.046
-0.43
0.918
0.704
LB2
-0.938
0.41
-0.309
0.092
-0.005
-0.031
-0.927
-0.165
LB3
-0.885
0.222
-0.93
0.15
0
0.177
-0.602
-0.469
2
BC
0.553
0.185
0.424
0.32
0.336
0.253
0.257
0.302
BW1
-0.465
0.162
-0.888
0.391
0.055
0.205
0.248
-0.299
BW2
0.018
0.037
0.479
-0.312
0.05
-0.34
-0.254
0.368
LB1
-0.185
-0.259
-0.639
0.337
-0.736
-0.348
0.252
-0.415
LB2
-0.27
0.154
-0.654
0.248
-0.456
-0.493
0.234
-0.824
LB3
-0.038
0.538
-0.191
-0.132
-0.251
0.187
0.193
0.452
Figure 1. Map of sites for fish community surveys completed in 2006, 2007 and 2018-2019 in
the Blackwater (sites BC, BW1, BW2) and Little Blackwater Rivers (LB1-3).
Figure 2. Non-metric multidimensional scaling analysis for fish community surveys conducted at
sites in the Blackwater (BC, BW1, BW2) and Little Blackwater (LB1, LB2, LB3) Rivers before
(2006, 2007) and after (2018-2019) establishment of Northern Snakehead (NSH). Axes were
correlated to overall abundance of fish, as well as abundances of certain species, including:
White Perch (WP), Gizzard Shad (GS), Brown Bullhead (BB), and sunfish spp. (Centrarchidae).
Figure 3. Comparison of diversity measures (species richness (upper panel) and species
dominance (lower panel)) for Blackwater River (right) and Little Blackwater River (left) before
and after establishment of Northern Snakehead.
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Northern Snakehead Channa argus, a species native to parts of Asia, became established in the Potomac River drainage prior to 2004. Removals by agencies appeared to do little to control abundance or limit spread into new waterways. As such, Northern Snakehead has become widespread throughout much of the Potomac River drainage in addition to many other river systems throughout the Chesapeake Bay watershed. As abundance increased within the Potomac watershed, new recreational and commercial fisheries were developed by encouragement of state and federal agencies to increase harvest. A mark–recapture program to examine growth and movement of Northern Snakehead began in 2009, as population density appeared to be increasing. In 2013, tagging methods changed to allow population size of Northern Snakehead to be estimated within selected tributaries (Little Hunting Creek (LHC) and Upper Anacostia River (UA)). From 2009–2017 we used mark–recapture angler returns and agency sampling data to view population size in context with changes in fishing mortality. The UA population linearly declined with increasing fishing mortality, while the LHC population changed very little in response to fishing mortality except for 2016 which had the lowest population estimate and highest fishing mortality. We are cautiously optimistic that exploitation may help control population growth, but recreational fishing alone is unlikely to cause significant declines in Northern Snakehead populations. Furthermore, well-established populations are likely to require high (>25%) exploitation rates to be effective.
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The invasion of non-native fishes has caused a great detriment to many of our native fishes. Since the introduction of invasive carps, such as Silver, Bighead, Common and Grass Carp, managers and researcher have been struggling to remove these species while also hypothesizing the detriment of further invasion. This study developed a food web model of four locations on the Mississippi River and used those models to assess the impacts of two scenarios: carp removal and carp invasion. In the Middle Mississippi River where these invasive carps are already present, the models found that it would take a sustained exploitation of up to 30% of initial biomass over an extended period to remove Grass Carp and up to 90% removal of initial biomass to remove Silver and Bighead Carp. In the locations where Silver, Bighead, and Grass Carp are not yet established (i.e., Pools 4,8, and 13) the invasion of these species could cause declines from 10 to 30% in initial biomass of native fishes as well as already established nonnative invasive species.
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Northern Snakehead Channa argus, a nonnative species to North America, was discovered in 2004 in tidal freshwater of the Potomac River, the second largest drainage of the Chesapeake Bay watershed. Since then Northern Snakehead has expanded its range throughout much of Maryland's portion of the Chesapeake Bay watershed. We estimated that the species has spread beyond its introduced range at a rate of about 2.7 subwatersheds per year. If that rate is maintained, the species will have spread its range throughout the entire watershed in 52 years. This rate has been consistent over time except for short periods of heightened expansion that followed an introduction of snakeheads into Delaware waters and in years with greater levels of spring precipitation. The expansion by Northern Snakehead is more widespread than that of other invasive fishes of Maryland's portion of the Chesapeake Bay, which include Blue Catfish Ictalurus furcatus and Flathead Catfish Pylodictus olivaris. Attempts to control the spread of the species have included public education, incentives for harvest, agency surveys, and law enforcement. The rapid natural and human-aided expansion of Northern Snakehead throughout the country's largest estuary highlights the importance of laws that prevent live possession and importation of this species.
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Conservation of Largemouth Bass (LMB) Micropterus salmoides populations requires an understanding of population dynamics that are influenced by environmental challenges, such as the spread of invasive species. We used an age-structured population model to compare population growth rates (λ) between a simulated population that included invasive Northern Snakehead (NSH) Channa argus as a competitor and predator and a simulated population that did not. We then assessed the sensitivity of our results to natural variation in LMB recruitment. When recruitment of LMB was already poor, there was a high risk of population decline that did not depend on whether NSH was included in the model scenario. When recruitment of LMB was high, the risk of population decline was only 40% when NSH was not included in the model scenario; however, predation by and competition with NSH caused a higher risk of LMB population decline. Regardless of the level of LMB recruitment, the size of the LMB population at equilibrium was 20% lower (on average) when including NSH in the model. We conclude that when habitat conditions do not already significantly limit recruitment, populations of LMB may be adversely affected by cohabitation with NSH. Preventing the spread of NSH will lessen ecosystem pressures that negatively affect the LMB population. We encourage continued vigilance in conserving LMB populations by encouraging actions that promote recruitment and prevent spread of invasive species.