The ecological consequences of nonindigenous Corbicula fluminea establishment on a benthic macroinvertebrate community

The frequently high abundances achieved by the nonindigenous Asian clam, Corbicula fluminea, has caused concern regarding the potential negative ecological impact the clams may have on native benthic macroinvertebrate communities. This study examined the ecological consequences of C. fluminea establishment on the benthic macroinvertebrate community in a New Hampshire, USA, river by comparing the benthic community before and after establishment and comparing sites with and those without C. fluminea. The nonmolluscan benthic communities were sampled using a ponar grab and compared using community metrics of macroinvertebrate densities, taxa richness, Shannon diversity, Hilsenhoff Biotic Index, and Bray-Curtis similarities cluster analysis and MSD ordination. Native bivalves were sampled by divers excavating 0.25 m quadrats and compared using abundance and size-frequency distributions. No consistent significant difference was seen when comparing macroinvertebrate metrics before vs. after C. fluminea establishment or when comparing sites with and without C. fluminea for any of the metrics used. Metrics were often similar or improved at sites with C. fluminea relative to years before establishment or when compared to reference sites lacking C. fluminea. Bray-Curtis cluster analysis and MDS ordination failed to separate sites with vs. those without C. fluminea. Similarly, native bivalve abundances, namely Elliptio complanata and Sphaeriidae, were similar at sites with vs. sites without C. fluminea. Elliptio complanata size-frequency distributions did not differ significantly when compared across sites with and without C. fluminea. Rather than having negative consequences on the benthic macroinvertebrate community as many have proposed, it appears that C. fluminea may either have no effect or positive effects on the macrobenthos.


Introduction
The potential ecological consequences of nonindigenous Asian clam, Corbicula fluminea (O. F. Müller, 1774), populations have been discussed for years (Cooper et al. 2005;Ilarri and Sousa 2012;Sickel 1973;Sousa et al. 2005Sousa et al. , 2008aStrayer 1999, Vaughn and. The frequently high population abundances achieved when C. fluminea invade a new area have led many to conclude that C. fluminea negatively impact abundance and diversity of benthic macroinvertebrate communities, including native bivalves, in North America (e.g., Hakenkamp et al. 2001;Strayer 1999;Sousa et al. 2008a;Vaughn and Hakenkamp 2001;Williams et al. 1993). The potential impact of C. fluminea on indigenous benthic communities continues to be of concern as these clams expand their range northward in North America and Europe (e.g., Bódis et al. 2012;Caffrey et al. 2011;Domagala et al. 2004;Mackiewicz 2013;Morgan et al. 2003Morgan et al. , 2004Simard et al. 2012;Smagula 2016). Unfortunately, little empirical evidence is available documenting the impact, or lack thereof, of the nonindigenous C. fluminea on benthic invertebrate communities (Ilarri and Sousa 2012;Strayer 1999;Vaughn and Hakenkamp 2001). As C. fluminea expand their range northward, more evidence is needed to elucidate the impact of C. fluminea on indigenous benthic communities.
Originally native to southeast Asia, C. fluminea has been introduced to North and South America, Europe, Africa, and the Pacific islands (e.g., Ilarri and Sousa 2012;Clavero et al. 2012;Morgan et al. 2003;Müller and Baur 2011;Strayer 1999;McMahon 1999McMahon , 2002McMahon and Bogan 2001). Corbicula fluminea has experienced considerable geographic dispersion in just the past few decades (Ilarri and Sousa 2012;McMahon 1999;Morgan et al. 2003;Sousa et al. 2008a). In North America, live C. fluminea were first documented in British Columbia in 1924 (Counts 1981). By 1953, the clams had spread through much of the U.S., especially the Southeast (McMahon 1983;Simard et al. 2012), and can now be found in most of the lower 48 states and Hawaii (Strayer 1999). Corbicula fluminea have continued to spread in North America expanding their range northward into cooler waters once thought too cold for their survival (Smagula 2016;A. Benson, USGS pers. comm. 21 August 2017). For example, C. fluminea have recently spread into numerous areas of Colorado, Connecticut, Massachusetts, Minnesota, Michigan, New Hampshire, New York, and Utah where low water temperatures and ice cover are common (Colwell et al. 2017;Cordeiro et al. 2007;Morgan et al. 2003;Richards 2018;Smith et al. 2018;Wick 2017;USGS 2019;T. Richardson pers. observ. July and August 2017). Asian clams were first reported in Connecticut in 1990 (Morgan et al. 2003) and appeared in Massachusetts by 2001 (Colwell et al. 2017). By 2007 Asian clams had appeared in New Hampshire in the Merrimack River (Smagula 2018) and in Vermont by 2016 (Colwell et al. 2017). Asian clams have since spread into the greater southeastern area of New Hampshire including the numerous sites in the Merrimack River and in several lakes (T. Richardson pers. observ. July and August 2017).
The reasons for the northern range extension of C. fluminea into areas with low water temperatures and winter ice formation is a matter of considerable scientific uncertainty. Often, such expansion has been attributed to thermal plumes from cooling water discharge (Morgan et al. 2003(Morgan et al. , 2004Simard et al. 2012). Compared to other bivalve species, C. fluminea has a limited temperature tolerance and has been widely perceived to be limited in range due to intolerance of cold water < 2 °C and warm water > 36 °C (Cairns and Cherry 1983;Mattice and Dye 1976;McMahon 1979McMahon , 1983McMahon , 2002Rosa et al. 2012;Smith et al. 2018;Verbrugge et al. 2012;Werner and Rothhaupt 2008a). For example, French and Schloesser (1991) saw extirpation of complete C. fluminea populations due to winter mortality in the St. Clair River, Michigan and Smith et al. (2018) suspected back-to-back harsh winters contributed to overwinter mortality in the Fox River in Wisconsin. However, Müller and Baur (2011) found ≥ 75% survival when C. fluminea was exposed to 0 °C water for up to 4 weeks and that 17.5% of clams survived 0 °C exposure for 9 weeks. Similarly, populations of Asian clams were able to rebound after harsh winter conditions caused a die-off in Lake Constance, Germany (Werner and Rothhaupt 2008a). The rebound was attributed to reproduction from the few surviving individuals. Additionally, Morgan et al. (2004) concluded that the effects of thermal discharge on winter survival in the Connecticut River were minimal. Cordiero et al. (2007) saw successful overwintering of Asian clams throughout Colorado in the absence of warm water discharges. In a study examining C. fluminea distribution and thermal discharge, Castañeda (2012) cited human population density as being a more important factor than thermal discharge in C. fluminea establishment and density. This evidence combined with known distribution expansion into areas with low water temperature and ice formation suggests that C. fluminea have the genetic variability and capacity for adaptation to withstand cold winter temperatures and are able to establish in a much wider range of rivers and lakes than previously assumed.
Corbicula fluminea populations are known to frequently reach abundances > 1,000 clams/m 2 and may exceed 5,000 clams/m 2 in some locations (e.g., Caffrey et al. 2011;Morgan et al. 2004;Simard et al. 2012;Sousa et al. 2008b;Strayer 1999;Vaughn and Spooner 2006). Such high Asian clam abundances have led some researchers to suggest that these clams could have a negative impact on the abundance and diversity of the indigenous benthic community including native unionid bivalves (Cordeiro et al. 2007;Strayer 1999;Sousa et al. 2008c;Vaughn and Hakenkamp 2001;Williams et al. 1993). With few exceptions, the potential negative ecological consequences of C. fluminea establishment on the ecosystems they invade have not been confirmed or validated. For example, Vaughn and Hakenkamp (2001) point out establishment of C. fluminea has only been speculated to have negatively impacted native bivalve abundance and distribution. Strayer (1999) recognizes that evidence for impacts of C. fluminea on native bivalves is derived largely from examining non-overlapping spatial distributions of bivalves or, less frequently, from changes in populations of native bivalves over time. Most of this evidence is anecdotal making it difficult to ascertain the impacts of C. fluminea on native bivalves. Studies that investigated the interaction between native unionids and C. fluminea have suggested no significant effects of the clams on native bivalves or other invertebrates at Asian clam densities < 100 clams/m 2 up to 3,000/m 2 (Belanger et al. 1990;Beran 2006;Karatayev et al. 2003;Leff et al. 1990). For example, Karatayev et al. (2003) found no significant correlations between low C. fluminea densities (36-43 clams/m 2 ) or the density of their shells and any invertebrate taxon studied. An experiment by Hakenkamp et al. (2001) indicated that C. fluminea at densities near 1,900 clams/m 2 had no effect on benthic protists and invertebrates. Unfortunately, most studies of C. fluminea impacts on native assemblages of benthic organisms are indirectly correlative or largely examined non-overlapping spatial distributions of bivalves or investigated changes in populations over time following C. fluminea establishment (sensu Strayer 1999).
Rather than negatively impacting indigenous benthic communities, some authors have suggested that C. fluminea may actually have a positive effect through ecosystem engineering (Gutiérrez et al. 2003;Jones and Gutiérrez 2007;Sousa et al. 2009). Corbicula fluminea movement within the top layer of sediments leads to bioturbation. Such bioturbation contributes to substantial changes in abiotic conditions like dissolved oxygen, redox potential, amount of organic matter, and particle size in a manner typically enhancing habitat conditions for other organisms (Ilarri and Sousa 2012;Werner and Rothhaupt 2007). Additionally, invasive bivalve species tend to have positive impacts on invertebrate density, biomass, and species richness through enhanced habitat heterogeneity, provide refugia from predators and abiotic stress, fluid transport and organic matter accumulation, and sediment reworking (Gutiérrez et al. 2003;Sousa et al. 2009 and references therein). In general, studies on the ecosystem engineering of bivalves, including C. fluminea, suggest they may either have no effect on native benthic invertebrates or mainly positive effects on native benthic invertebrates. As C. fluminea continue to spread northward into previously unoccupied areas, it becomes increasingly important to determine what impacts, positive or negative, the clams may have on indigenous benthic communities in these areas.
The purpose of this study was to determine if any negative effect of C. fluminea presence on native benthic invertebrates and native bivalves could be detected. To achieve this, multiple community metrics (e.g., density, richness, diversity, and resilience) were enumerated. The metrics were used to compare sites with and without C. fluminea across years as well as before and after invasion by C. fluminea. It was hypothesized that if C. fluminea negatively affect native species there should be a detectable decrease in native species abundance, richness, diversity, and thus a reduction in resilience to disturbance. A lack of a decrease in these metrics would suggest that C. fluminea invasion is not negatively impacting native invertebrate species.

Study Site
The current study was conducted in Hooksett Pool, Amoskeag Pool downstream of Hooksett Dam, and at reference sites in Garvins Pool located immediately upstream of Hooksett Pool all in the Merrimack River, New Hampshire, USA. The Merrimack River is a relatively large river with average discharge at Goffs Falls below Manchester, NH near 215 m 3 /sec during the time of the study (USGS 2020). Channel width ranged from approximately 100 m in Garvins Pool to near 230 m in lower Hooksett Pool with depth at the study sites ranging from 2-4 m (Normandeau 2012).

Quantitative Macroinvertebrate Sampling
Quantitative macroinvertebrate sampling was used to determine the current extent of the distribution and abundance of C. fluminea, and composition and abundance of benthic macroinvertebrates within Hooksett Pool. In 1972, Normandeau (2012 1972(Normandeau 2012. In 2011, 2014, and 2016, C. fluminea were not found in Garvins Pool or upstream of Merrimack Station in Hooksett Pool. As a result of the current study using the same sampling locations, the benthic communities before and after C. fluminea establishment in Hooksett Pool could be directly compared. Furthermore, using the 2014 and 2016 data from these stations, the benthic community from upstream reference areas and areas lacking C. fluminea could be directly compared to downstream areas with C. fluminea. Such before and after C. fluminea establishment, and presence vs. absence studies provided strong comparisons for assessing the ecological impact of C. fluminea establishment on the indigenous benthic invertebrate community. At each station a transect line perpendicular to Merrimack River flow was established. Along each river bank-to-bank transect line three, one quarter distance locations were sampled: West, Middle, and East. A 23 × 23 cm standard ponar grab sampler was used to collect five replicate samples in 2014 and three replicate samples in 2016 at each of the three locations along each transect (Normandeau 2012). For the 2014 and 2016 study, the following ten stations were sampled (listed from upstream to downstream): Garvins Pool Stations USR and DSR (reference stations), Hooksett Pool Stations N-10 and N-5 upstream of Merrimack Station, and S-0, S-4, S-8, S-17, S-24 and Amoskeag Pool, AMOS downstream of Merrimack Station ( Figure 1). The substrata in Garvins, Hooksett, and Amoskeag Pools at the sampling transects were similar and uniformly sand and silt with some cobble/gravel mixture (unpublished data).
A total of 150 replicate samples were taken in 2014 and 90 were taken in 2016. Each replicate sample was initially sieved in the field through a 0.6 mm mesh sieve, preserved in an individual sample container, labeled with a unique sample number, replicate number, collection date, station and location. Samples were taken to the laboratory for sorting, identification, and enumeration of macroinvertebrates. The GPS coordinates of each station and location sampled were recorded on the field data sheet along with the sample label information.
In 2014, three of the five replicate samples from each station and location were randomly selected for further processing in the laboratory. In 2016, two of the three replicate samples were randomly selected for further processing. The remaining two samples from 2014 and the single remaining sample for 2016 were archived for potential future use if necessary. Each of the selected replicate samples was sorted, invertebrates identified to the lowest distinguishable taxon, and each taxon enumerated. All laboratory sorting, identification, and counting were subjected to a quality control (QC) inspection to insure an average outgoing quality limit of 10% or better. This QC inspection indicates that the data originating from laboratory processing were certified by independent statistical re-inspection at a sampling frequency to document that less than one record (one line of data) out of every ten records was outside of the established precision and accuracy for all contents of that record.

Native Freshwater Bivalve Quantitative Sampling
Compared to other benthic invertebrates, freshwater bivalves in the family Unionidae are typically found in low abundance and are not sampled adequately by the ponar method (T. Richardson pers. observ.). Quantitative sampling using divers to excavate 15 cm of substratum from a quadrat of known area (usually 0.25 m 2 ) is preferred for assessing unionid abundance (Miller and Payne 1993). Diver sampling was used coincident with the 2014 and 2016 quantitative macroinvertebrate sampling to adequately quantify the abundance of native bivalves found in Hooksett Pool during November/December 2014 and July 2016. The quantitative unionid samples were taken in the same stations and locations after quantitative benthic macroinvertebrate ponar grab sampling was completed. Care was taken that the diver quadrats were established adjacent to, but not within the exact footprint of the substratum disturbed by ponar sampling. Only stations N-10, S-0, S-4, and S-24 were sampled by divers for native bivalves in 2014 and 2016. At each station and location, two quantitative samples were taken for a total of 24 samples in 2014 with only one sample per station (n = 12) in 2016. Each sample was taken by randomly placing a 0.25 m 2 quadrat on the river bottom, excavating the substratum to a depth of 15 cm, and placing the quadrat contents in a 19-L plastic pail. In the laboratory, one of the two replicate samples in 2014 from each station and location was randomly selected for further processing. The remaining sample was archived for potential future use if necessary. Each of the 12 replicate samples to be processed was immediately sieved first over a 25 mm mesh sieve followed by 6 mm mesh then finally a 3 mm mesh sieve. All bivalves (unionids and sphaeriids) were removed, identified, and counted.

Data Analyses
Data from the quantitative macroinvertebrate samples were used to calculate densities of C. fluminea for each sample. These samples were also used to compute density, taxa richness, Shannon Community Diversity Index, and Hilsenhoff Biotic Index (HBI) of the benthic macroinvertebrate community for each sample using standard methods. At sites with Asian clams, C. fluminea was excluded from macroinvertebrate estimates for direct comparisons to the benthic macroinvertebrate community at sites without C. fluminea. The Shannon Community Diversity Index focuses on quantifying the uncertainty in predicting the species identity of an individual that is taken at random from the data set; similar communities will have similar Shannon Diversity. The HBI estimates the overall pollution tolerance of the community in a sampled area, weighted by the relative abundance of each taxonomic group, i.e., the HBI takes into account resiliency of the community. Lower HBI's indicate a less pollution tolerant benthic community and, therefore, a "healthier" benthic community i.e., reflects the community response to C. fluminea presence.
Data analyses of abundance and community metrics comparing stations or years were conducted using QI-Macro (2015) 2-way ANOVA with replication. Each null hypothesis tested was that there were no differences among stations or among years in the metric of interest. When comparing among year differences, only the two stations with highest C. fluminea abundances were used because the greatest potential for C. fluminea influence should have been at those stations. Similarly, when comparing among stations, only data from 2014 and 2016 were used because these should have reflected the greatest potential for C. fluminea impact on the benthic community and provided the most complete data set across all stations. When significant terms existed, single factor ANOVA with post hoc comparison was used to determine significant differences among station or year means. Post-hoc ANOVA comparisons of means were conducted with LSD comparisons. LSD post hoc comparisons were chosen as the least conservative estimate to show differences among sites or years with and without C. fluminea. Size-frequency distributions of native bivalves were analyzed using Kolmogorov-Smirnoff. Non-normal data sets were also analyzed using conservative nonparametric statistics (Kruskal-Wallis or Friedman as appropriate) and produced similar results with respect to the null hypotheses tested using parametric statistics. All statistical analyses used α = 0.05 to test for significant differences.
Additionally, multivariate analyses were performed using data from the 90 quantitative macroinvertebrate samples from 2014. Prior to multivariate analyses, the taxon-mean counts per ponar grab sample were computed across the three replicate grab samples from each station and location. This reduced the input dataset for multivariate analyses to the mean count per taxon and grab sample for 30 station-locations. Data handling and preparation for multivariate analyses were completed using SAS software (version 9.3). Multivariate analyses were then performed using PRIMER v6 (Plymouth Routines in Multivariate Ecological Research) software, following standard techniques for the evaluation of spatial patterns in the distribution of faunal assemblages (Clarke 1993;Warwick 1993;Clarke and Green 1988;Clarke and Warwick 2001). These analyses included classification (cluster analysis) by hierarchical agglomerative clustering with group average linking and ordination by non-metric multidimensional scaling (MDS). Bray-Curtis Similarity was used as the basis for both classification and ordination. Prior to analyses, faunal abundance data (i.e., mean count per taxon and grab sample) were square-root transformed to ensure that all taxa, not just the numerical dominants, would contribute to similarity measures. Bray-Curtis Community Similarity cluster analysis separates sites with dissimilar benthic invertebrate community composition. Likewise, MDS Ordination based on Bray-Curtis Similarity clusters separates sites with differences in community similarity. The "similarity profile test" (SIMPROF) was used to provide statistical support for the identification of faunal assemblages (i.e., selection of cluster groups). SIMPROF is a permutation test of the null hypothesis that the groups identified by cluster analysis (samples included under each node in the dendrogram) do not differ from each other in multivariate structure. The "similarity percentages" (SIMPER) analysis was used to identify contributions from individual taxa to the overall dissimilarity between cluster groups. This analysis was used to identify the contribution of macroinvertebrate taxa (including C. fluminea) to the overall dissimilarity between cluster groups.

Corbicula fluminea
Corbicula fluminea densities were compared among years only for the three sites with clams for which there was complete data for 2011, 2013, 2014, and 2016, i.e

Discussion
The potential ecological consequences of nonindigenous C. fluminea populations and their adaptations for rapid establishment have been discussed for years (Colwell et al. 2017;Cooper et al. 2005;Ilarri and Sousa 2012;McMahon 1983McMahon , 1999McMahon , 2002Sickel 1973;Sousa et al. 2005Sousa et al. , 2008aStrayer 1999;Vaughn and Hakenkamp 2001). This study investigated the ecological impact of C. fluminea on the indigenous benthic community of Hooksett Pool, Merrimack River, New Hampshire, USA, in order to determine the ecological consequences of C. fluminea establishment on the benthic community. At the three stations in Hooksett Pool with multiple year data, C. fluminea densities fluctuated widely between 2011 and 2016. When C. fluminea were first discover in Hooksett Pool in 2011 (Normandeau 2012), densities averaged nearly 1,800 clams•m -2 , declined to 123•m -2 in 2013, rebounded by 2014 to a high of nearly 3,400•m -2 only to fall again to about 300•m -2 in 2016. Such rapid population growth of C. fluminea following establishment or some density reduction event is not uncommon. Corbicula fluminea populations may rapidly reach high abundances, but a low juvenile survivorship and a high mortality rate throughout adult life leads to considerable annual, seasonal, and site-to-site variability and fluctuations in abundances and frequent population mortality events, especially in sites experiencing cold water temperatures (e.g., French and Schloesser 1991;Ilarri et al. 2011;Morgan et al. 2003Morgan et al. , 2004Smith et al. 2018;Vohmann et al. 2010;Werner and Rothhaupt 2008a). Contributing to this variability is the relatively low physiological tolerances of the C. fluminea and its dependence on elevated fecundity for invasive success and rapid population recovery (McMahon 2002). Such elevated fecundity by C. fluminea is due, in part, to its high allocation of energy to growth and reproduction which is typical of such opportunistic and invasive species (McMahon 2002). The population of C. fluminea in Hooksett Pool appears to adhere to the typical population variability centered on the capacity of the clam to rapidly re-establish populations after density reduction (McMahon 2002

Asian Clams Present
Population variability notwithstanding, the frequently high population abundances achieved when C. fluminea invade a new area has led many to conclude that C. fluminea negatively impact abundance and diversity of benthic macroinvertebrates communities, including native bivalves (e.g., Araujo et al. 1993;Hakenkamp et al. 2001;Strayer 1999;Sousa et al. 2005Sousa et al. , 2008dVaughn and Hakenkamp 2001;Williams et al. 1993). However, experimental evidence for the impact of C. fluminea on benthic macroinvertebrates is rare (Vaughn and Hakenkamp 2001;Karatayev et al. 2005;Rothhaupt 2007, 2008b). Most studies rely on examination of spatial distributions, comparing systems with C. fluminea to systems lacking the clams. No studies use multiple community metrics or compare benthic communities before and after establishment of C. fluminea (sensu Sousa et al. 2008b). This study used density and commonly accepted macroinvertebrate community metrics to compare invertebrate communities with and without C. fluminea, and before vs. after C. fluminea establishment. Benthic macroinvertebrate density and community metrics in Hooksett Pool showed no clear pattern between areas with and those without C. fluminea. During 2014 and 2016, macroinvertebrate densities were the same or higher in areas with C. fluminea. Likewise, benthic invertebrate density at the two stations with highest C. fluminea densities (S-4 and S-17) was similar before and after C. fluminea establishment (1972 and 1973 vs. 2011-2016). Lacking a consistent significant difference in invertebrate density among stations with and without C. fluminea for either 2014 or 2016 and before and after Asian clam establishment suggests C. fluminea had no negative effect on the benthic macroinvertebrate community density. Like this study, Werner and Rothhaupt (2007) found no significant difference in macroinvertebrate density between experimental boxes with and without C. fluminea at densities similar to those found in Hooksett Pool. Werner and Rothhaupt (2007) also found that some macroinvertebrate species actually increased in abundance and attributed this positive affect to the addition of shell to the sandy substratum after mass mortality of C. fluminea. Similarly, using a combination of field and laboratory experiments, Hakenkamp et al. (2001) found that abundance of C. fluminea was negatively associated with the abundance of benthic bacteria and flagellates, but saw no effect on other protists or meiofauna. Taken together, this information strongly supports that benthic macroinvertebrate abundance in large rivers like the Merrimack River is not adversely affected by the establishment and persistence of C. fluminea and that Asian clams may actually have positive impacts on invertebrate abundance. Such evidence supports the assertions of Gutiérrez et al. (2003) and Sousa et al. (2009) that nonindigenous invasive species like C. fluminea may actually enhance abundance through ecosystem engineering, i.e., organisms that can physically modify the environment.
Various metrics of benthic macroinvertebrate communities are often used to examine community responses to stressors (e.g., pollution) by comparing them to reference sites and other areas lacking stressors or before and after some stressor is applied. In 2014 and 2016, benthic invertebrate taxa richness was similar between areas with and without C. fluminea and higher at some sites compared to reference sites. Shannon diversity was similar at many sites with C. fluminea compared to sites without clams including reference sites during 2014 and 2016. HBI during 2014 and 2016 was similar between sites with and those without C. fluminea and was even lower at some sites with clams compared to reference sites. Similarly, benthic invertebrate richness and diversity at the two stations with highest C. fluminea densities (S-4 and S-17) were similar before and after C. fluminea establishment (1972 and 1973 vs. 2011-2016). HBI was either similar or lower following C. fluminea establishment than prior to establishment. Bray-Curtis similarities cluster analysis and MDS based on Bray-Curtis similarities was unable to distinguish between macroinvertebrate communities at reference sites and other sites with vs. those without C. fluminea as was the case with Sousa et al. (2008c). Others have seen similar beneficial effects of introduced C. fluminea (Sousa 2008c, d;Werner and Rothhaupt 2007), but lacked the comparison of indices like Shannon diversity and HBI. The use of such community metrics is a powerful tool for assessing the effects that C. fluminea establishment and abundance may have on benthic invertebrate communities. The use of such metrics and the findings here conclusively support the idea that nonindigenous C. fluminea are not having a negative effect on macrobenthic communities and may possibly be enhancing the community through ecosystem engineering (Gutiérrez et al. 2003;Sousa et al. 2009).
In addition to the benthic macroinvertebrate community in general, specific concern has been raised about the effects C. fluminea may have on native bivalves, specifically members of the Unionidae (e.g., Strayer 1999;Sousa et al. 2008c;Vaughn and Hakenkamp 2001;Williams et al. 1993). Using diver excavated samples, this study compared E. complanata, the dominant unionid in Hooksett Pool, and a sphaeriid clam at sites with and sites without C. fluminea in 2014 and 2016. Although differences among years were seen in native bivalve density, lack of a significant difference between sites with and without Asian clams supported that in 2014 and again in 2016 the presence vs. absence of C. fluminea had no impact on native bivalve densities. Size-frequency of E. complanata was also compared between sites with and without C. fluminea and no difference was observed. Native bivalve density was unaffected by presence of C. fluminea and E. complanata size-frequency distribution was similar between sites with and those without C. fluminea. However, E. complanata is the most widespread, ubiquitous, and common unionid species in New England and the effects of C. fluminea on rare species of Unionidae may differ. Nonetheless, the frequently high abundances reached by C. fluminea has led to the supposition that C. fluminea may have greater impacts on native bivalves than any other nonindigenous species except for the zebra mussel, Dreissena polymorpha (Pallas, 1771) (Strayer 1999). This purported impact is generally thought to happen through competition (e.g., depletion of phytoplankton and pedal feeding food resources, and space utilization/reduction); ingestion of sperm, glochidia and juveniles; and ammonia production and oxygen consumption following mass mortality events (see Strayer 1999 and references therein). However, Leff et al. (1990) found that while C. fluminea appeared to cause localized reduction of seston and rapidly cleared the sediment boundary layer of food, there was no evidence of a negative impact on the distribution of the native E. complanata. Similarly, in this study C. fluminea and native bivalves were abundant at the same sites and no effect of C. fluminea presence on native bivalve abundance was seen. Such overlapping distributions at sites with relatively high C. fluminea abundances suggest the clams are not crowding out native bivalves, and ammonia and oxygen consumption following die-offs appears to have not been an issue. Similarly, if C. fluminea were affecting native unionid bivalves through competitive interactions, a difference in population size structure between stations with and without C. fluminea might be expected. Competitive interactions through feeding would tend to lead to reduced growth and an overall smaller shell size-frequency distribution. Conversely, ingestion of sperm, glochidia or juveniles would tend towards a larger shell size-frequency distribution. Because E. complanata size-frequency distributions were not different between sites with and without C. fluminea, neither competition through feeding activities nor consumption of sperm and propagules seem to be affecting this native unionid.
As C. fluminea spread into previously unoccupied areas, the impacts, if any, that C. fluminea may have on indigenous benthic communities continues to be of concern. Before and after C. fluminea establishment, and presence vs. absence studies combined with community metrics provide strong evidence for assessing the ecological impact of C. fluminea establishment on indigenous macroinvertebrate communities. Clearly, in Hooksett Pool the presence of nonindigenous C. fluminea is not having negative consequences on the indigenous benthic macroinvertebrates or native bivalves as demonstrated through density and various community metrics. Indeed, C. fluminea may be beneficial to the native invertebrates through ecosystem engineering leading to more habitat heterogeneity, refugia from predators and abiotic stress, fluid transport and organic matter accumulation, and sediment reworking (see Gutiérrez et al. 2003;Sousa et al. 2009 and references therein). Nonetheless, Sousa et al. (2009) rightly point out that beneficial effects on some species may mask negative consequences experienced by declines in other species. However, similar or improved HBI metrics and the inability of community similarity indices to separate communities with C. fluminea in this study indicate this is not the case in Hooksett Pool. Nonindigenous C. fluminea may not have the negative impacts on benthic macroinvertebrate communities as was once thought and, in some cases, may be having beneficial effects on the communities. Invasivity inherently implies that the nonnative species is causing harm to the indigenous community. Lacking any evidence of harm to the indigenous benthic invertebrate community, the Asian clam in the Merrimack River may not therefore be an invasive species per se, but may simply be an addition to the indigenous community.