Samuel Johnston
Erin Shepta
Caroline C. Arantes
Brent Murry- Davis College of Agriculture and Natural Resources, West Virginia University, Morgantown, WV, United States
Introduction: Zooplankton are critical components of aquatic food webs that transfer pelagic primary production to higher trophic levels. Nearly all fishes rely on zooplankton during their early life history, and many species remain planktivorous throughout their lives, yet zooplankton assemblages are vulnerable to anthropogenic stressors, especially invasive species, and are frequently understudied. Here we examine zooplankton assemblage changes occurring between sampling periods in 1991–1992 and 2021–2022 (a 30-year period) and a more robust analysis of contemporary assemblage change relative to the introduction of aquatic invasive species in the Ohio River, U.S.A.
Methods: Zooplankton were sampled in 8 navigation pools (reaches between locks and dams) in the Ohio River in 2021 and 2022 and (1) qualitatively compare these contemporary samples to historic published data from 1991–1992 and (2) analyze the contemporary data using multivariate statistics to evaluate the impact of invasive carps (silver, Hypophthalmichthys molitrix and bighead, H. nobilis) across a gradient of carp density.
Results: There were notable changes to the zooplankton community from samples taken in 1991–92 to 2021–22, including a shift in large-bodied zooplankton from cladocerans to copepods and among small-bodied zooplankton invasive dreissenid mussel veligers largely replaced rotifers. Peak zooplankton density also advanced a month earlier in contemporary samples relative to the historic peak. Invasive carp have not yet had a measurable impact on the Ohio River zooplankton assemblage.
Discussion and Conclusion: Changes in the zooplankton assemblage were substantial and included taxa-specific phenological shifts that culminated in total assemblage density peak advancing a month earlier relative to historic timing. There were also changes in dominance in both large-bodied and small-bodied taxa. These changes are more consistent with changes observed in other systems relative to climate change and dreissenid mussel invasions than with invasive carps. Contrary to many other rivers in the Mississippi River basin invaded by silver and bighead carp, we observed little to no effect of invasive carps, though it will be important to continue monitoring zooplankton as the carp invasion advances.
Introduction
Zooplankton play an important role in aquatic ecosystems by connecting primary producers (phytoplankton) to higher levels of the food web (Carpenter et al., 2001; Arndt, 1993). Despite ontogenetic shifts later in life, zooplankton serve as the primary food for fish species across every trophic level (Abo-Taleb, 2019). Post-larval fish are especially vulnerable to changes in zooplankton abundance, with low caloric stores and a high cost of movement so that any interruption or decrease in prey availability is often fatal (Tillotson et al., 2024; Rellstab et al., 2004; Cushing, 1990; Hjort, 1914). For many fish species survival past age-0 also depends on high somatic growth during their early life stages (Cushing, 1990). Fluctuations in zooplankton abundance ultimately determine early life survival and recruitment of fishes, with cascading effects on population dynamics and food web functioning.
In recent centuries, natural fluctuations in zooplankton abundance have been increasingly altered by anthropogenic influences, with invasive species emerging as a major driver of zooplankton community shifts and broader food web disruptions (de Arruda Ramos et al., 2025; Camatti et al., 2023; Serranito et al., 2016). The introduction of invasive species has become a major concern within aquatic ecosystems due to their increasing frequency and dramatic impacts (Mayfield et al., 2021; Mainka and Howard, 2010). Invasive species have been shown to have impacts that can span from hybridization with native species to disruption of biochemical processes and alteration of food webs (Cucherousset and Olden, 2011). However, impacts can vary greatly depending on both characteristics of the invader and the recipient community in which it becomes established (Ricciardi et al., 2013). Once established, invasive species are usually difficult and costly to manage and nearly impossible to eradicate (Meronek et al., 1996). Yet rates of introduction of new invasive species and their potential impacts are only expected to increase in over the next several decades (Cuthbert et al., 2021; Pluess et al., 2012; Cucherousset and Olden, 2011; Lodge et al., 2006). Conservation of aquatic systems therefore depends greatly on understanding the magnitude of potential impacts of invasive species after introduction (Cucherousset and Olden, 2011).
Within the Mississippi River basin, two species of invasive carp (silver, Hypophthalmichthys molitrix and bighead, H. nobilis) have spread quickly and become a dominant portion of the biomass in many rivers (Sass et al., 2010). After escaping from aquaculture ponds in the 1970s, filter-feeding invasive carp quickly became established, spreading throughout the Mississippi River and its major tributaries including the Ohio River (Chick and Pegg, 2001; Kolar et al., 2005). Specific to the Ohio River, there were records of invasive carps throughout the lower and mid-river reaches by the 1990s but densities remained low until the 2010s when multiple detections became more regular in the mid- and upper-pools (Mississippi Interstate Cooperative Resource Association, 2017). Their effective establishment has been attributed to their high fecundity, rapid growth rates, ability to swim long distances and to filter a large range of particles sizes (Sass et al., 2014; DeGrandchamp et al., 2008; Williamson and Garvey, 2005). Both species have been shown to have rapid early somatic growth, reaching larger juvenile sizes quicker than many native planktivores, thus lowering their risk of predation relative to native juvenile fish (Michaletz, 1998; Williamson and Garvey, 2005). Bighead carp and silver carp can filter particles down to 50 μm and 3 µm respectively (Havel et al., 2015), which allows them to exploit the full-size range of phytoplankton and zooplankton, essentially sequestering energy and preventing it from transferring to higher trophic levels (Hecky et al., 2004; Sanful et al., 2013).
One of the greatest concerns with the introduction and spread of invasive carp is their potential to alter food webs (DeBoer et al., 2018). Invasive carp have been shown to drastically decrease the abundance of zooplankton in several tributary rivers of the Mississippi River (Sass et al., 2014). For example, the high abundance of both species in the Illinois River resulted in dramatic shifts in the native zooplankton community, including the suppression of larger zooplankton (copepods and cladocerans) density while smaller rotifers increased in density (Sass et al., 2014). However, laboratory microcosm studies conducted in lentic environments had varied effects. When introduced into clay-lined ponds bighead carp were shown to reduce abundance across the entire zooplankton size spectrum (Collins and Wahl, 2018). This reduction in zooplankton is thought to have dramatic impacts on native planktivorous fish species (Samson et al., 2009; Sass et al., 2014). Both species have been found to overlap in diets with native gizzard shad (Dorosoma cepedianum), bigmouth buffalo (Ictiobus cyprinellus), and to a lesser extent paddlefish (Polyodon spathula) (Samson et al., 2009; Coulter et al., 2013). Potential competition has resulted in declines of body condition and juvenile growth rates of native planktivores in some cases as well (Irons et al., 2007; Schrank and Guy, 2002). While their potential to impact native planktivores is well documented, invasive carp reductions of zooplankton are likely to have much broader food-web wide effects as well.
The Ohio River has undergone drastic anthropogenic changes over the past several decades and the introduction and spread of invasive carp is likely to exacerbate these impacts on native food webs. Our objective was to assess invasive carp impacts on zooplankton communities. Specifically, we assessed potential changes in zooplankton community before and after carp establishment by comparing historical (1991–1992) and contemporary (2021–2022) datasets. We also examined how zooplankton communities responded to a gradient of invasive carp abundance based on contemporary data collections.
We compared total zooplankton density, and community structure (i.e., density of cladocerans, copepods, rotifers) across time periods and months to evaluate shifts associated with carp invasion. Additionally, we compared zooplankton density and community structure across the gradient of carp invasion — in areas where carp are present and largely absent. We hypothesized that the establishment of silver and bighead carp would result in (a) reduced total zooplankton density; (b) reduced density of cladocerans and copepods; and (c) an increase of rotifer densities consistent with the changes observed in the Illinois River (Sass et al., 2014). This study directly evaluates impacts of invasive carp on the zooplankton community, with results providing insights on changes throughout the Ohio River food web.
Methods
Study site
The Ohio River is 981 miles long, originating in Pittsburg, PA and ending when it joins with the Mississippi river at Cairo, IL (Figure 1). Alternating between a relatively free flowing river and pools created by 20 traversable dams, the Ohio River is the second largest river by discharge in the United States, averaging 3,273 m3/s at Louisville, KY. Invasive carp densities in the Ohio River existed along a gradient of higher densities at the confluence with the Mississippi River and decreasing densities upriver. Pools of the Ohio River have been categorized into invasion fronts: established, invasion, and presence, which describe this gradient of carp density, reproduction, and recruitment success. The lowest downstream pools (Cannelton and McAlpine) are in the established front, which is defined by higher abundance of adult carp with frequent reproduction and evidence of recruitment. The middle pools (Markland and Meldahl) are in the invasion front and are characterized by lower adult densities, with limited evidence of reproduction and recruitment. The upper pools (Greenup and R.C. Byrd) are the presence front, which have low carp densities and no evidence of reproduction or recruitment (Novak et al., 2024).
Figure 1. Map of the study region, the Ohio River, between Evansville, IN and Point Pleasant, WV). Blue circles represent approximate locations of historic (1991–92) samples, black circles show sampling areas of both 2021 and 2022, and yellow circles show sample locations that were added in 2022. Each section of the river is highlighted by invasive carp front, red is establishment front, yellow is invasion front and green is presence front.
Zooplankton data were collected from the middle of the river at 8 study sites spanning 697 river kilometers and 6 navigable pools (Cannelton to R.C. Byrd) along the Ohio River Sampling locations included: Point Pleasant and Huntington, WV (presence front); Portsmouth, Aberdeen and Cincinnati, OH; Aurora, IN (invasion front); and Louisville and Concordia KY (established front) (Figure 1). All six pools are considered constricted with narrow floodplains generally alternating between right and left banks (Thorp J. H. et al., 1994). The average river width for the 8 sites is 466 m with the narrowest being 380 m at Portsmouth, OH and the widest being 600 m at Louisville, KY.
Historic data collection
Data from the pre-carp establishment was obtained from existing published data (Thorp J. H. et al., 1994) to serve as a baseline for comparison with contemporary patterns. Historical data on zooplankton community composition and density were collected from 1991–1992 by Thorp J. H. et al. (1994) from the Markland, McAlpine and Cannelton pools. Samples were collected once monthly in the McAlpine pool and quarterly in the Cannelton and Markland pools. Samples were collected by dipping a 5 L bucket for a water volume of 10–20 L depending on relative zooplankton abundance. Water was then poured through a 63 µm sieve, transferred to 60 mL bottles, killed with hot water and preserved using a modified Lugols’s solution (Thorp J. H. et al., 1994). We reached out to the authors and the original data was not available. Thus, we are limited to qualitative comparisons of published mean values. We, did however, obtain historic water flow, discharge, data from USGS station 03294500 for the Ohio River at Louisville, KY (Figure 2) (data source: https://waterdata.usgs.gov/monitoring-location/USGS-03294500/) and water temperature data for 1999–2021 from ORSANCO (Figure 3) (https://www.orsanco.org/data/temperature/) to compare environmental conditions across pools and time periods.
Figure 2. Monthly average flow (discharge, cubic feet per second) for (A) historic (1991–1992, Thorp et al., 1994a) and (B) contemporary time periods (2021–2022) for the Ohio River at Louisville, KY (Data derived from USGS: https://waterdata.usgs.gov/monitoring-location/USGS-03294500/).
Figure 3. Monthly mean water temperature recorded by ORSANCO (https://www.orsanco.org/data/temperature/) between 1999 and 2021 for Cannelton, Markland, Greenup, and RC Byrd pools. There was a 0.073 °C/yr increase in March water temperatures across all pools, but no other temporal patterns in water temperature in May, July, or August. Mean monthly water temperatures did not vary significantly across pools.
Contemporary data collection
Data on zooplankton community composition and density were collected from all eight sites and were sampled 1–2 times a month from May to October in 2021 and 2022. We collected samples from the main channel pelagic zone between 7a.m. and 5p.m. Samples were collected right, center, and left sides of the river spanning the entire river width including nearshore and pelagic areas (minimum depth sampled = 1 m, maximum depth = 17 m, mean depth = 10.09 m, median depth sampled was 10 m). Zooplankton samples were collected using an integrated water column tube sampler (Sass et al., 2014) consisting of a pump and hose (2.4 cm inside diameter) with a weight attached to the end that was raised and lowered through the entire water column while pumping. Water was pumped into a premeasured 136 L container using a hand powered diaphragm pump to avoid dismemberment of collected zooplankton. The 136 L sample of water was then pumped through a 64 µm filter before being rinsed into a sample bottle (Chick et al., 2010). All zooplankton collected were then anesthetized with alka-seltzer before being preserved in 70% ethanol. All zooplankton were identified using a stereomicroscope (Amscope SZMT2) at 28-180x using a plankton wheel and measured using an attached calibrated digital camera (Amscope MU1803). Biomass was calculated based on published length weight regressions from Downing and Rigler (1984). Cladocerans, copepods and rotifers were identified to genus while copepod nauplii were identified to class. Depending on relative abundance, samples were either fully counted or subsampled from a known dilution until at least 100 individuals were counted or 10% of the subsample’s volume had been processed. Total counts of subsamples were then extrapolated to full volume. Historical sampling protocols combined immature and adult copepods for data presentation; therefore, we followed the same approach for the historical comparison to ensure consistency and comparability among datasets. However, for analyses assessing contemporary patterns along the carp invasion gradient, we separated immature and adult copepods to better capture size-related differences between these groups.
Data analysis
Discharge was qualitatively compared across years to evaluate the role of potentially large variation in flow regime influencing zooplankton dynamics (Figure 2). Bimonthly water temperature data from select pools were available from ORSANCO (https://www.orsanco.org/data/temperature/) between the years 2000 and 2021 for distinct reaches of the river: Pittsburgh to Greenup, we used to represent RC Byrd pool, ORSANCO reach Greenup to McAlpine we used to represent water temperatures for Markland and Greenup pools, and McAlpine to Cairo were used to represent water temperatures in Cannelton pool. Many years were missing, particularly the range between 2004 and 2010. We examined the months of March, May, July and September individually. March was prior to our sampling period, however spring warming plays a major role in zooplankton phenology (Winder et al., 2012; Schlüter et al., 2010). We also included 1999 for each of the pools and months derived from a different ORSANCO data set (same web source) that provided daily means pooled across 1995–2003 as a single annual average, 1999 is the median year. The historical sampling occurred in 1991–1992, but we could not locate data beyond these described above. We tested for differences among the interaction of years and pools of each month separately using a linear model. We modeled months separately as they represent unique seasons that may show differing long-term dynamics.
We conducted two complementary sets of analyses to evaluate invasive carp effects on zooplankton communities: (1) a temporal comparison between historical and contemporary datasets to assess changes between pre- and post-carp establishment, and (2) an analysis of contemporary spatial patterns along a gradient of carp abundance to examine current ecological responses to carp invasion. Due to the limitations associated with using published data, historical comparisons with Thorp J. H. et al. (1994) were restricted to qualitative and descriptive analyses. Specifically, we compared mean total zooplankton density overall and across months, and examined species structure, including both dominance and community composition.
For our analyses of contemporary spatial patterns, we used a factorial analysis of variance (ANOVA) to test for differences in zooplankton mean density (# zooplankton/L) and mean total biomass (µg dry weight/L) across pools, years, and months. These are inclusive of all taxa combined. When only the main effects were significant, we used a Tukey post hoc test to identify significance among levels of the main effect.
We then used a permutational multivariate analysis of variance (PERMANOVA) to assess differences in zooplankton taxa density and biomass across years, pools, and months. Matrices of taxa density and biomass were the response variables and matrices were Hellinger transformed to address problems associated with taxa absence (i.e., zeros in the data) and for large variances among individual samples and over time. Taxa were grouped by copepods (adults), immature copepods, cladocerans, veligers, and rotifers. The model included the interaction of factors: year, month, and pools. For all statistical analyses, p-values were considered significant at α = 0.05 and analyses were performed in R 4.3.3 (R Core Team, 2023) with packages emmeans (Lenth et al., 2018), ggplot2, vegan (Oksanen et al., 2022), and dplyr (Yarberry, 2021).
Results
Environmental conditions
Qualitatively, the river’s flow regime was relatively consistent 1991–1992, 2021–2022 relative to a long-term average (1928–2025) (Figure 2). The main notable difference between the historic and contemporary sampling was higher spring flows (a spike in May) not observed during the historic period. For the most part seasonal water temperatures remained consistent among sampling periods, with the exception of early spring (March) where we observed an approximate 1.6 °C over the 22-year time series (on average 0.073 °C/yr, Table 1; Figure 3).
Table 1. Statistical outcomes for linear models testing mean monthly water temperature variation over time (1999–2021) across pools (Cannelton, Markland, Greenup, and RC Byrd) representing the range of pools sampled in the historic and contemporary data sets.
Historic comparisons
Thorp et al. (1994b) completed 277 samples between May and November 1991 and 1992, while our study includes 89 samples May through October 2021 and 2022. Mean zooplankton density during our contemporary sampling (2021–2022) appears to have increased substantially compared with historic values (1991–1992) (Figure 4). The maximum historical mean total zooplankton density was 24.4 individuals/L in August while the lowest was 4.7 individuals/L in May. The highest mean total zooplankton density for contemporary sampling was 67.2 individuals/L in July while the lowest was 14.9 individuals/L in May.
Figure 4. Graph of average Ohio River zooplankton taxa density (individual/L) from 1991–1992 (Left) (Thorp et al., 1994b) and 2021–2922 (Right).
In addition to changes in total density, we also observed potential changes in phenology. Mean zooplankton density historically had two density peaks, the highest being 24.4 individuals/L in August and the second highest of 15.6 individuals/L occurring in October. Current zooplankton density showed a similar bimodal peak but occurred a month earlier (July and September) in each case. There was a large first peak of 67.2 individuals/L in July with a secondary smaller peak of 26.6 in September.
Veligers, planktonic juveniles from invasive zebra mussels (Dreissena polymorpha) which were not present during the historic period, were the dominant plankton in contemporary samples. They ranged from a low of 7.9 individuals/L in October to a peak density of 34.6 individuals/L in July. Across all month’s veligers represented on average 59.5% of total zooplankton density (range = 32%–92%). Without the invasive veligers our sampling still had a larger and earlier bimodal peak (36.3 individuals/L in July and 18.0 in October).
Along with the appearance and dominance of dreissenid veligers we also observed other apparent changes in taxonomic composition. Copepod density historically had a single peak in September of 3.2 individuals/L while contemporary sampling showed increased overall abundance as well as altered phenology with two peaks, the largest being 27.2 individuals/L in July with a second smaller peak in September of 11.9 individuals/L. Contemporary copepod density increased 560% across all sampled months compared to historical collections. Rotifer density historically had three seasonal peaks, one small peak of 4.1 individuals/L in May, a seasonal peak of 16.7 individuals/L in August and final medium peak of 11.6 individuals/L in October. Contemporary rotifers showed lower overall density with a 60% decrease over the year and again altered phenology with two small peaks, one of 6.1 individuals/L in June and another in September of 4.8 individuals/L. Finally, cladocerans historically had two peaks, a small early season peak of 3.6 individuals/L in June and a larger mid-season peak of 5.1 individuals/L in August. Contemporary cladocerans were 44% lower than historical and had a single peak of 3.9 individuals/L in July. The largest change in the zooplankton density was the addition and dominance of dreissenid veligers.
Contemporary patterns in zooplankton assemblage dynamics
The mean density of zooplankton (# zooplankton/L across all taxa) showed monthly changes (F5,42 = 2.45, P = 0.0492, Table 2a) with no differences among pools or years (all interactions and other main effects from the full model were non-significant, P > 0.05). Tukey HSD post hoc comparisons for months showed that July had significantly higher total zooplankton density than May (P = 0.0484), June (P = 0.0412), and October (P = 0.0449), all other pairwise comparisons were insignificant (P > 0.05) (Figure 5a).
Table 2. Statistical outcomes for analysis of variance (ANOVA) testing (a) mean zooplankton density (#/L) and (b) mean zooplankton biomass (µg/L).
Figure 5. Mean monthly zooplankton (a) density (#/L) and (b) biomass (µg/L) in the middle Ohio River (R.C. Byrd to Cannelton Pools) in 2021 and 2022.
The density of the zooplankton assemblage (i.e., taxa) differed among years and months, but there were no differences among pools. The 3-way interaction of year*month*pool was non-significant (F6,88 = 0.97, p = 0.46, Table 3a), month*pool and year*pool were similarly non-significant (F23,88 = 0.87, p = 0.72 and F3,88 = 1.02, p = 0.44, respectively) suggesting no effect of invasive carp in the lower pools. In contrast, the interaction of year*month was significant (F3,88 = 7.86, p = 0.01). Seasonal density dynamics were generally driven by invasive dreissenid mussel veligers which were numerous in all months and both years (Figure 6).
Table 3. Statistical outcomes for permutational multivariate analysis of variance (PERMANOVA) testing (a) mean zooplankton taxa density (#/L) and (b) mean zooplankton taxa biomass (µg/L).
Figure 6. Monthly taxa (a,c) density (#/L) and (b,d) biomass (µg/L) in the middle Ohio River in 2021 and 2022. “Immature” refers to immature copepods.
Zooplankton biomass followed the same patterns as density, where the only differences were seasonal among months (F5,42 = 2.59, P = 0.0397, Table 2b) with no differences among pools or years (all interactions and other main effects from the full model were non-significant, P > 0.05). Tukey HSD post hoc comparisons for months showed that July had significantly higher zooplankton density than May (P = 0.0278) and June (P = 0.0049), all other pairwise comparisons were insignificant (P > 0.05) (Figure 5b). The distribution of zooplankton biomass among taxa differed among years and months (year*month; F3,88 = 11.12, P = 0.01), but no other interaction terms were significant (P > 0.05 in all cases, Table 3b). In contrast to density fluctuations, biomass seasonal dynamics were driven by large bodied, adult copepods which made up a small fraction of the total density but carried a high biomass per individual.
Discussion
Our study provides a comparison of Ohio River zooplankton pre- (1991–1992) and post-invasive carp introductions (2021–2022) spanning a 30-year time period along with a spatial comparison of contemporary zooplankton assemblage taxonomic structure across a gradient of invasive carp abundance. Major differences in contemporary samples compared to historical samples included phenological changes and shifts in densities of dominant taxonomic groups. However none of these changes appear to be directly attributed to invasive carp introductions. We observed phenological shifts in zooplankton densities, where peak densities emerged over a month earlier in contemporary samples compared to historic samples. This change in phenological zooplankton densities may be due to changes in climate (Richardson, 2008) rather than the presence of invasive carp. Average zooplankton densities were also higher in contemporary surveys; however, this was primarily due to the large increase in zebra mussels (D. polymorpha) veligers, the planktonic larval stage of these invasive mussels. We also observed a significant increase in copepod density and decrease in cladocerans and rotifers relative to the earlier samples. This shift was consistent across the gradient of carp abundance, and more so resembles zooplankton community restructuring observed in other systems invaded by zebra mussels (DeBoer et al., 2018; Sass et al., 2014; Tillotson et al., 2022) rather than the impacts of invasive carp (Sass et al., 2014; Collins and Wahl, 2018). While invasive carp may not be the dominant driver reshaping zooplankton communities in the Ohio River, we still found dramatic shifts in zooplankton communities that could have serious impacts on native planktivores and the Ohio River food web (Tillotson et al., 2024; DeBoer et al., 2018).
We acknowledge potential sampling bias between the historic surface samples and our contemporary integrated tube samples. However, Jack et al. (2006) found in the Ohio River that copepods tended to migrate to the surface during the day and drop to the bottom overnight (contrary to typical lake migrations), such that the historic samples should have over-sampled copepods and contemporary may have under-sampled by similar logic, but we found a far higher proportion of copepods than historical. Similarly, the observed replacement of abundant rotifers by veligers is certainly not a result of gear differences. The river is also turbulent which increases vertical mixing. Overall, while sampling gear may have had some level of influence, we suggest that the observed changes far exceed what we might expect based on differences in gears.
Contrary to our original prediction and to trends observed in the Illinois and Upper Mississippi Rivers (Tillotson et al., 2022; DeBoer et al., 2018; Sass et al., 2014), we did not find evidence to support the hypothesis that invasive carp are driving changes in zooplankton composition. Previous studies found decreases in copepods and cladocerans because of carp introduction in the Illinois and Upper Mississippi Rivers (Tillotson et al., 2022; DeBoer et al., 2018; Sass et al., 2014). This shift was linked to high carp density as invasive carp either directly consume zooplankton or compete with them for the consumption of phytoplankton (Tumolo and Flinn, 2017; Samson et al., 2009; Radke and Kahl, 2002). While we observed a decrease in cladoceran density post-carp invasion, there was a simultaneous increase in copepod density in contemporary samples. Tillotson et al. (2022) reported that zooplankton communities in the Upper Mississippi River were dominated by rotifers and ostracods in areas with dense carp populations. While rotifers dominated the zooplankton community in our historic data set (Thorp J. H. et al., 1994; pre-silver and bighead carp invasion) rotifer densities were overall very low in our contemporary samples, though replacement by zebra mussel veligers in this same size class was evident.
The lack of consistency between our results and previous studies demonstrating strong impacts of invasive carp on zooplankton communities (DeBoer et al., 2018; Sass et al., 2014) may be explained by differences in carp population density among systems. The Ohio River does not appear to support carp populations as dense as those reported in other invaded rivers. It is likely that invasive carp in this system have not yet reached density thresholds sufficient to alter zooplankton community structure (Novak et al., 2024; Erickson et al., 2021). Novak et al. (2024) examined changes in fish community size structure and reached the same conclusion, suggesting that carp populations in the Ohio River may not have attained densities high enough to elicit expected ecological impacts on fish communities. They suggest that carp relative abundance may need be at least 24% of the total fish biomass to significantly impact the food web, however these abundances have not yet been consistently reached in the Ohio River (Novak et al., 2024). While invasive carp may not be driving these shifts, we still observed considerable changes between the two surveys that are likely due to other anthropogenic factors negatively impacting this system.
Invasive zebra mussels appear to be the most likely explanation for the observed changes in the zooplankton community of the Ohio River. Zebra mussels had been introduced but were not well established in the Ohio River during historical sampling (Thorp J. H. et al., 1994). Zebra mussel veligers were observed for the first time in 1992 and reached a seasonal high late in the sampling season (August-October) with small maximum densities of 0.15 individuals/L (Thorp J. H. et al., 1994). In the contemporary sampling, veligers were the dominant taxon of the overall zooplankton density throughout most of the year. The increased abundance of zebra mussel veligers contributed to the higher overall zooplankton densities observed in contemporary surveys, replacing rotifer density in this system (Bowen et al., 2018). We generally observed declines in rotifers and cladocerans with increases in copepods and zebra mussel veligers, but will briefly note that during both sampling periods the use of 63–64 µm mesh underestimates total rotifer abundance by inadequately sampling the smallest rotifers, but the results are comparable among time periods. In the Hudson River, zebra mussel introduction resulted in decline of microzooplankton including rotifers, likely due to direct consumption from zebra mussels, while the macrozooplankton did not significantly decrease (Pace et al., 1998). However, with veligers to prey upon instead of rotifers, copepods may be able maintain a higher population than cladocerans and rotifers, preventing similar declines (Williamson and Butler, 1986). Similar results have been observed in the Great Lakes, where the introduction of zebra mussel and later quagga mussel (Dreissena rostriformis) was followed by a shift in zooplankton communities from cladoceran to copepod dominance, potentially driven in part by dreissenid-induced oligotrophication (Barberio et al., 2019; Sterner, 2011). While invasive zebra mussels may be disrupting native zooplankton communities, the massive number of veligers in the water column may provide a benefit to the native zooplankton community by mitigating the predation effects of invasive carp (Stoeckel et al., 1997). A large abundance of veligers, such as those observed in our samples, might provide a buffer for native zooplankton against predation from invasive carp. Anderson et al. (2016) reported a personal communication from Dr. István Tátrai, who documented that the majority of animal portion of the diet of bigheaded carp were veligers in Lake Balaton, Hungary. This potential buffering effect could further explain the apparent lack of impacts of invasive carp observed in our contemporary surveys. However, further research is needed to confirm the impacts of zebra mussel introduction and their potential interactions with invasive carp on the Ohio River zooplankton community.
Another major result from our study was the observed shifts in zooplankton phenology in contemporary samples compared to our historic reference. While this is not likely due to the presence of invasive species, climate change and in particular earlier spring warming, may be a contributing factor to the changes we found in copepod and overall zooplankton dynamics. Gerten and Adrian (2002) observed similar trends in community dynamics where certain copepods emerged earlier in the year and had higher annual peak abundance in response to increased annual water temperature. However, studies have shown warmer temperatures associated with climate change may promote zooplankton community shifts towards smaller individuals and taxa which in general means a shift towards cladocerans (Rasconi et al., 2015; Beaver et al., 2014; Gillooly, 2000). While in contemporary surveys copepods emerged earlier with an earlier peak abundance, cladoceran densities decreased with one peak in the middle of the season instead of an early and late peak observed historically and commonly (Jansson et al., 2020; Carter et al., 2017; Adrian et al., 2006). However, potential effects of climate change may be confounded by effects of zebra mussels on the community especially considering the replacement of rotifers by veligers that cladocerans are less able to exploit (O’Malley and Bunnell, 2014; Vadadi-Fülöp et al., 2012; Liebig and Vanderploeg, 1995). Future research is needed to address the potential impacts of climate change on the Ohio River zooplankton community.
Conclusion
The zooplankton community of the Ohio River has changed considerably in composition, density and phenology over the course of 30 years. Zooplankton play an important role in aquatic ecosystems by connecting primary producers (phytoplankton) to higher levels of the food web (Carpenter et al., 2001; Arndt, 1993). Within that context, zooplankton are known to serve as an important food resource for many native planktivores, as well as for early life stages (i.e., larvae and fry) of many other fish species. The seasonal use of zooplankton by fish early life stages can become disconnected through changes in zooplankton phenology (Tillotson et al., 2024; Abo-Taleb, 2019; Balachandran and Peter, 1989). Due to limited larval fish data in the Ohio River, we are uncertain as to whether fish recruitment is tracking changes in zooplankton abundance dynamics. Preventing further introductions and effectively managing current populations of aquatic invasive species is essential for maintaining the stability of native zooplankton communities. Preserving native zooplankton communities is essential for ensuring the health and diversity of native aquatic ecosystems in the future.
Data availability statement
The raw data supporting the conclusions of this article will be made available by the authors, without undue reservation.
Ethics statement
The manuscript presents research on animals that do not require ethical approval for their study.
Author contributions
SJ: Conceptualization, Resources, Formal Analysis, Validation, Writing – review and editing, Methodology, Supervision, Writing – original draft, Investigation. ES: Writing – original draft, Formal Analysis, Writing – review and editing, Data curation, Investigation, Validation, Conceptualization, Methodology. CA: Writing – original draft, Formal Analysis, Validation, Writing – review and editing, Methodology. BM: Methodology, Conceptualization, Validation, Visualization, Formal Analysis, Supervision, Writing – original draft, Funding acquisition, Data curation, Investigation, Writing – review and editing, Resources.
Funding
The author(s) declared that financial support was received for this work and/or its publication. This material is based upon work that was partially supported by the National Institute of Food and Agriculture, U.S. Department of Agriculture, McIntire Stennis project under 1026001 (BM) and 1026124 (CA). This project was largely funded by the United States Department of the Interior, Fish and Wildlife Service Grant No. F21AP03188-00, F23AP00140-00, and F24AP00729-00 (BM).
Acknowledgements
The findings and conclusions in this article are those of the authors and do not necessarily represent the views of the U.S. Fish and Wildlife Service.
Conflict of interest
The author(s) declared that this work was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.
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Keywords: climate change, invasive carp, phenology, zebra mussels, zooplankton
Citation: Johnston S, Shepta E, Arantes CC and Murry B (2026) Temporal changes in zooplankton assemblage and the complex role of aquatic invasive species in the Ohio River, USA. Front. Environ. Sci. 13:1736157. doi: 10.3389/fenvs.2025.1736157
Received: 30 October 2025; Accepted: 22 December 2025;
Published: 27 January 2026.
Edited by:
Hao Yu, Michigan State University, United StatesReviewed by:
Tan Lu, Chinese Academy of Sciences (CAS), ChinaMolly Sobotka, Missouri Dept of Conservation, United States
Copyright © 2026 Johnston, Shepta, Arantes and Murry. This is an open-access article distributed under the terms of the Creative Commons Attribution License (CC BY). The use, distribution or reproduction in other forums is permitted, provided the original author(s) and the copyright owner(s) are credited and that the original publication in this journal is cited, in accordance with accepted academic practice. No use, distribution or reproduction is permitted which does not comply with these terms.
*Correspondence: Samuel Johnston, c2owMDAyMUBtaXgud3Z1LmVkdQ==; Brent Murry, YnJlbnQubXVycnlAbWFpbC53dnUuZWR1