Invasive species are recognized as one of the most significant hazards to biodiversity (Elton, 1958; Su et al., 2021) and managing them presents a substantial challenge (Lipták et al., 2024). Globally, hundreds of fish species have successfully established populations outside their native ranges, resulting in diverse negative costly impacts (Diagne et al., 2020; Haubrock et al., 2022). Freshwater systems are particularly vulnerable to invasive species (Kinlock et al., 2020). The changes in hydrology caused by dam construction, land-use modifications both nearshore and in catchments, and increased nutrient runoff from agricultural practices disrupt the trophic interactions in these ecosystems, facilitating invasions (Havel et al., 2015, 2005; Kovalenko, 2019). Further, the connectivity sometimes enhanced by hydrological alterations can facilitate the spread of non-native species, which when combined with late detection, may result in multiple adverse effects on aquatic assemblages (Júlio Jr et al., 2009). Invasive freshwater fish can impact natural ecosystems by interacting with native species through competition, predation, spread of diseases and parasites, hybridization, and others (Bernery et al., 2022; Chakraborty, 2024; ISSG, 2025). Additionally, invasive fish can alter native habitats through behaviors such as substrate disturbance, grazing, and browsing, all of which may disrupt physical and chemical habitat characteristics (Bernery et al., 2022; Chakraborty, 2024; ISSG, 2025).

Bighead carp (Hypophthalmichthys nobilis) and silver carp (H. molitrix), collectively referred to as bigheaded carp, are native to east Asia. They were introduced to the United States in the 1970s to consume and regulate plankton in aquaculture and wastewater treatment facilities (Kelly et al., 2011; Kolar et al., 2007). However, by the 1990s, they had escaped confinement and began expanding into the Mississippi River basin, where they now have established populations (Chick and Pegg, 2001; Kolar et al., 2007). Since then, they have been slowly extending their range and can now be found in most tributaries of the Mississippi River (Miranda, 2023).

Bigheaded carp have specific traits and plasticity that contribute to their successful introduction into non-native habitats (Birdsall, 2023). They migrate long distances to spawn, have high fecundity and growth rates, and have high tolerances to environmental variations (Kolar et al., 2007; Lenaerts et al., 2023). Silver carp primarily feed on phytoplankton, while bighead carp mainly consume zooplankton (Cremer and Smitherman, 1980; Dong and Li, 1994; Ivan et al., 2020). However, bighead carp can also consume detritus and algae when other food sources are scarce (Cremer and Smitherman, 1980; Jennings, 1988; Kolar et al., 2007; Sass et al., 2014). These planktivorous species have the potential to disrupt food webs by competing for plankton with native filter-feeding and planktivorous fish, which rely on plankton for part of their life cycle (Eagles-Smith et al., 2008; Harris et al., 2022; Sass et al., 2014). Several studies have indicated that bigheaded carp exhibit some degree of dietary overlap with native planktivores such as bigmouth buffalo (Ictiobus cyprinellus), gizzard shad (Dorosoma cepedianum), and emerald shiner (Notropis atherinoides), with declines in the abundance and condition of these species (Haupt and Phelps, 2016; Irons et al., 2007; Pendleton et al., 2017; Phelps et al., 2017; Wang et al., 2018). Additionally, juvenile American paddlefish (Polyodon spathula), an iconic recreational and commercial filter-feeder, has been shown to be outcompeted by bighead carp in ponds (Schrank et al., 2003). Similar effects may also be impacting many other species, though these have yet to be studied (Wang et al., 2018).

Estimating competition between bigheaded carps and native species in a multispecies assemblage presents challenges due to the complexity and multidimensional nature of ecological competition, which is seldom direct or easily observable (Ross, 1986; Schoener, 1974). Fish frequently employ resources that are overlapping but not identical. For instance, they may consume distinct species within the same taxonomic group or graze for the same species in distinct habitats (Ríha et al., 2025; Ross, 1986). Therefore, in diverse assemblages, a fish species may compete weakly with many other fishes rather than strongly with one or a few (Ross, 1986). Moreover, many fish exhibit dietary flexibility, ontogenetic shifts, or habitat switching as they develop (Sánchez-Hernández et al., 2019; Winemiller, 1989). These adaptations allow fish to minimize resource overlaps under scarcity, hence altering competition intensity (Polis, 1984; Wisheu, 1998). Competition can vary seasonally, annually, and among habitats within the same water body (Bloomfield et al., 2022). Short-term studies may miss these dynamics; however, long-term or whole-system studies are often logistically challenging. Thus, because direct estimation of competition among species in fish assemblages is impractical, fish ecologists often turn to indices or proxies that may reflect the potential for competition rather than measure competition directly (Wallace Jr., 1981). These indices may integrate dietary and habitat utilization to estimate tentative and multifaceted competitive overlap.

The Tennessee River basin has the highest fish biodiversity in North America and is currently experiencing an increase in bigheaded carp populations within its system of impoundments (Ridgway and Bettoli, 2017; Vallazza et al., 2025). This expansion has the potential to restructure food webs and thereby threaten native biodiversity by disrupting the equilibrium of interactions that support ecological communities (Altenritter et al., 2022; Chick et al., 2020; DeBoer et al., 2018; Harris et al., 2022; Sampson et al., 2009; Yallaly et al., 2014). Some studies have explicitly linked the rise in bigheaded carp densities to a decline in the physical condition of a handful of native planktivores (Haupt and Phelps, 2016; Irons et al., 2007; Pendleton et al., 2017; Phelps et al., 2017), but this may only be the tip of the iceberg. This threat has mobilized natural resource conservation sectors to invest hundreds of millions of dollars to address observed and anticipated detriment (Miranda, 2023). In this context, the aim of this study is to identify the species most vulnerable to competition based on their dietary and habitat overlaps. Our objective is to provide a first-pass view of the functional overlap in dietary and habitat use between bigheaded carp and the diverse fish species assemblage in impounded segments of the Tennessee River basin.

Overall, 255 fish species were reported by the SFC database in the Tennessee River basin. Between 1990 and 2022, 150 of these species were collected in reservoirs through TVA electrofishing and gill-net sampling. Of the 150 species captured in reservoirs, 127 were native and 23 were non-native (i.e., from outside the Tennessee River basin) and were excluded from analysis. An additional 12 native species were excluded because trait data were unavailable or taxonomically uncertain in the FishTraits database, and 15 rare and restricted native species were removed because they were confined to upper-basin reservoirs inaccessible to bigheaded carp. Of the remaining 100 native species collected in reservoirs, 41 were classified as common, 31 as rare and widespread, and 28 as rare and restricted (Table 2).

The Sørensen similarity coefficients between bigheaded carp and native species deviated relative to dietary and habitat traits (Figure 2). Mean similarities were 0.67 (Min-Max, 0.29–1; SD, 0.16) for dietary traits and 0.53 (Min-Max, 0–0.8; SD, 0.13) for habitat traits. Dietary and habitat traits were significantly different (pseudo-F = 49.0, P = 0.001). This distinction suggests that native species and bigheaded carp differed less in dietary traits and more in habitat traits. Moreover, there was no significant correlation between the dietary and habitat similarity coefficients (Spearman r = 0.083, P = 0.413), indicating that when native species and bigheaded carp overlapped in diet often did not overlap in habitat, and vice versa.

The competition index was correlated with the dietary component (Spearman r = 0.94, P > 0.01) and habitat component (Spearman r = 0.37, P > 0.01). As intended by the differential weighting assigned to the competition index components, CI best represents dietary overlap. Moreover, the dietary overlap exhibited slightly greater variability than the habitat overlaps, as evidenced by their standard deviations, hence further amplifying the dominance of diet in the CI. Overall, 19 species demonstrated high competition index (i.e., CI > 0.75; Figure 2). The highest CIs included bigmouth buffalo, threadfin shad (Dorosoma petenense), black redhorse (Moxostoma duquesnii), bluntnose minnow (Pimephales notatus), highfin carpsucker (Carpiodes velifer), and gizzard shad. Most species (N = 69) exhibited low to moderate competition indices (i.e., CI < 0.75). Prominent among these were all black basses (Micropterus), temperate basses (Morone), sunfishes (Lepomis), crappies (Pomoxis), gars (Lepisosteus), silversides (Atherinopsidae), and paddlefish. Darter species, sculpins, and pickerels exhibited the lowest CIs. We note that there was no species with a CI of 0, and that the minimum CI estimated was 0.35, suggesting that perhaps our CI estimations may be conservative.

The sensitivity analysis revealed that CI remained consistent regardless of the weighting scheme. Spearman rank correlations between CI values derived from the baseline (0.7/0.3) and CI values derived from alternative weightings remained high (i.e., Spearman r > 0.9) for weightings of 0.5/0.5 (Figure 3). Thus, although the CI absolute values change depending on the weighting scheme, the rank order of CI values remains largely stable. This indicates that our competition index exhibits limited sensitivity to reasonable weighting variations and can be regarded as robust with respect to weighting choice.

The mean CIs were 0.66 (0.49–0.90), 0.64 (0.38–0.90), and 0.58 (0.38–0.90) for the common, rare and spread, and rare and restricted species, respectively (Figure 4). These CIs differed statistically among rarity status (pseudo-F = 3.4, P = 0.042), with pairwise tests indicating a significant difference between the common and the rare and restricted species (pseudo-t = 2.6, P = 0.013), but no difference between the common and the rare and spread species (pseudo-t = 0.6, P = 0.524). Differences between the rare and spread and rare and restricted species were not statistically significant, though the comparison indicated a modest separation (pseudo-t = 1.7, P = 0.088). We emphasize that these comparisons are exploratory and depending on how they are applied, may require further investigation with additional data. From an alternative point of view, out of the 100 species studied, 41% were common, yet 58% of the 19 high CI species were common. These findings, while ambiguous due to heavy overlaps, may point to a gradient of potentially stronger competition between bigheaded carp and common species, but lesser competition with the rarest of species.

Our assessment of dietary and habitat overlaps between bigheaded carp and native species revealed a key pattern: dietary similarities generally exceeded habitat similarities, and these two axes of overlap were uncorrelated. This decoupling is consistent with classic niche theory, which recognizes food and space as largely independent dimensions along which species may partition resources to facilitate coexistence (Schoener, 1974; Chase and Leibold, 2003). Accordingly, high trophic similarity does not necessarily imply strong interaction if spatial segregation limits encounter rates (Holt, 1987; Coulter et al., 2019). At the same time, trait-based indices quantify potential interactions, and strong competitive effects may still emerge for some species or life stages when resources become limiting or when seasonal hydrology and temperature compress habitat use and increase co-occurrence. Habitat management that preserves spatial heterogeneity and structural complexity may therefore help maintain segregation and dampen competitive interactions (Almany, 2004). However, such separation may erode in heavily invaded systems where bigheaded carp expand across habitats and reorganize food webs (Coulter et al., 2018; DeBoer et al., 2018; Hochstrasser and Collins, 2024).

The lack of correlation between dietary and habitat overlaps also bears on long-standing suggestions that bigheaded carp exploit a relatively underused pelagic–planktonic foraging pathway in North America (Kolar et al., 2007; Irons et al., 2007). Their tendency to occupy the water column and efficiently filter plankton has sometimes been framed as use of an “empty” or weakly exploited niche relative to benthic-oriented native assemblages. Our results are consistent with this hypothesis in the limited sense that trophic similarity need not translate into spatial overlap, potentially reducing encounter rates with many native species. Nevertheless, multiple lines of evidence caution against interpreting this pattern as proof of a truly vacant trophic niche. Many native fishes depend on plankton during early life stages or seasonally, and isotope-based studies have documented substantial trophic overlap and reorganization of native niches across carp density gradients (DeBoer et al., 2018; Harris et al., 2022). Moreover, experimental syntheses indicate that carp can generate both direct and indirect effects across habitats, including pelagic–benthic coupling via recycling and shunting of consumed organic matter to benthic environments (Hochstrasser and Collins, 2024). Together, these studies suggest that rather than occupying a wholly empty niche, bigheaded carp may be functionally specialized along a trophic axis while remaining partially segregated along a habitat axis. Thus, invasion-driven food-web alteration can redistribute native resource use and habitat associations and potentially intensify competition in habitats not typically associated with carp.

The CIs aligned with empirically documented interactions in field investigations and with rarity classifications, as common species are expected to have broader niches. Species such as bigmouth buffalo, gizzard shad, and emerald shiner have consistently shown evidence consistent with competition and food-web impacts in invaded rivers, including declines in body condition and shifts in diets following changes in zooplankton community structure (Harris et al., 2022; Hayer et al., 2014; Irons et al., 2007; Minder and Pyron, 2018; Pendleton et al., 2017; Shields et al., 2021; Wang et al., 2018). Our analysis listed these three species as having high competition risk (Figure 2). Similarly, paddlefish has been reported to have only modest diet overlap with bigheaded carp (Wang et al., 2018). Accordingly, our analysis listed paddlefish as moderate competition risk. Chick et al. (2020), Fletcher et al. (2019), and Tillotson et al. (2023) provided empirical evidence of a negative effect of invasive silver carp on native fish (e.g., bluegill Lepomis macrochirus, yellow perch Perca flavescens, freshwater drum Aplodinotus grunniens, crappies) that they attributed to competition for zooplankton during early life stages. Our index indicated that few gamefish species as high competition risk since the Frimpong and Angermeier (2009) dataset does not distinguish between seasons and life stages, instead presents aggregated traits that incorporate multiple ontogenetic stages, but they do not precisely distinguish among them. This discrepancy underscores a limitation of our estimated CIs but also highlights the need for re-evaluating our findings with juvenile fish traits. Furthermore, our index offers a basis for expanding the species to monitor in field surveys and empirical research. Lastly, common species are predominantly generalists, particularly in reservoirs (Miranda et al., 2025). Consistent with Gaston (2011), common species are likely to overlap with bigheaded carp across a broader range of their dietary and habitat niches, resulting in higher average CIs among common species.

Because competition requires both shared resources and opportunities for encounter, overlap along dietary or habitat axes does not translate directly into ecological interaction when resources are abundant or species remain spatially segregated. Resource availability varies longitudinally in river systems (Vannote et al., 1980), including the impounded Tennessee River (Miranda et al., 2008), but reservoirs modify these gradients through serial discontinuities (Ward and Stanford, 1995), replacing smooth upstream–downstream transitions with a chain of semi-lentic systems separated by riverine reaches whose dimensions depend on local physiography. Each impoundment establishes its own riverine–transitional–lacustrine gradient, resetting conditions downstream of dams (Miranda and Dembkowski, 2016) and altering retention time, light penetration, and primary productivity (Kimmel and Groeger, 1984; Soballe and Kimmel, 1987; USGS, 2018; Naymik et al., 2023; Kortmann, 2025), collectively transforming the Tennessee River into a sequence of hydrologically connected but ecologically semi-isolated reservoirs. Impoundments also create lentic–lotic “edge effects” at reservoir–tributary interfaces, where reservoir-associated generalists interact with rheophilic tributary fauna; seasonal movements of generalists into tributaries can intensify competition and predation and influence fish communities far upstream of impounded reaches (Šmejkal et al., 2023). Upstream reservoirs tend to be smaller, whereas downstream reservoirs are larger and more connected to floodplains and embayments that support substantial plankton production, generating testable predictions within our trait framework: overlap potential should be greatest where residence time is long and water-column habitat and plankton production are highest—typically in transitional and lacustrine zones and embayments—and where low-velocity off-channel habitats increase co-occurrence with pelagic planktivores (USGS, 2018; DeGrandchamp et al., 2008). At the same time, bigheaded carp recruitment depends on turbulent riverine reaches that keep semi-buoyant eggs suspended during downstream drift, linking ontogenetic requirements to the availability of flowing habitat within regulated systems (Chapman and George, 2011; Garcia et al., 2013). In the Tennessee River, this lentic–lotic juxtaposition is most evident in downstream reservoirs where silver carp show tailwater residency and where removal programs are concentrated (Budnick et al., 2025).

Nevertheless, the methods we applied have several limitations that necessitate cautious interpretation of our findings. First, we relied on a database that measures dietary and habitat traits at coarse resolution such as detritus, phytoplankton, pelagic, or benthic but did not capture finer scales such as temporal or ontogenetic shifts (Sanchez-Hernandez et al., 2022). Because traits are aggregated across life stages, our index cannot resolve early-life plankton dependence or indirect food-web effects. Also, the traits database we applied was developed for more natural environments and may not transfer exactly to anthropogenically modified rivers. In addition, trait uncertainty is higher for some taxa (especially rare species), so overlap estimates should be interpreted cautiously and validated. Consequently, using coarse traits could have overestimated or underestimated competition. Second, overlap indices, such as Sørensen similarity index applied in our study, are also coarse approximations. Because they are binary, they give equal weight to all utilizations, disregarding extent of resource use and resource availability, and offer a static depiction of a species, overlooking ontogeny and adaptation responses to changing environmental conditions. Furthermore, since our traits are binary and some dietary categories are broad (e.g., general invertebrate groupings), CI can overestimate overlap among species that partition resources at finer taxonomic or life-stage scales. Thus, our CIs should be interpreted as a preliminary metric of functional proximity rather than a direct measure of competitive strength. Third, 12 native species were excluded from this analysis because traits data were unavailable. These species included mostly uncommon cyprinids, catostomids, and two marine species that migrate up the Mississippi River and into the Tennessee River's lower reservoirs. Fourth, the competition index constructed from the dietary and habitat similarity coefficients required subjective weighting influenced by a static scheme that is likely spatially and temporally dynamic. Nonetheless, our sensitivity analysis mitigated this concern by indicating that the ordinal ranking of competition indices remained relatively stable across different weighting schemes. Overall, because our competition index aggregates trait information across seasons and life stages, it captures broad trophic similarity but may underestimate episodic spatial overlap during critical periods such as larval feeding or low-flow habitat compression. Given these limitations, our competition index is best viewed as a rudimentary indicator appropriate for first-pass estimations, not a precise measure of the intensity of ecological interaction.