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A Comparison of Resident Fish Assemblages in Managed and Unmanaged Coastal Wetlands in North Carolina and South Carolina
Kelly F. Robinson and Cecil A. Jennings

Southeastern Naturalist, Volume 13, Issue 2 (2014): 237–260

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Southeastern Naturalist 237 K.F. Robinson and C.A. Jennings 22001144 SOUTHEASTERN NATURALIST 1V3o(2l.) :1233,7 N–2o6. 02 A Comparison of Resident Fish Assemblages in Managed and Unmanaged Coastal Wetlands in North Carolina and South Carolina Kelly F. Robinson1,2 and Cecil A. Jennings3,* Abstract - The dominant fish species within impounded coastal wetlands in the southeastern US may be different from the species that dominate natural marshes. We tested the hypothesis that resident fish assemblages inhabiting impounded coastal wetlands in South Carolina would differ from resident assemblages in natural marshes of the southeastern United States. We used rarefied species richness, Shannon’s H' diversity, J' evenness, Morisita’s index of similarity, and the percent similarity index to compare resident fish assemblages from two impoundments to 12 open-marsh resident fish assemblages from previously published studies in North and South Carolina. We used rotenone to sample fish assemblages in impoundments. The assemblages in natural marsh habitat had been sampled with rotenone and seines. We classified comparisons yielding a similarity index ≥0.50 as moderately similar and those with an index ≥0.75 as very similar. Fifty-three percent of the among-impoundment comparisons (Morisita’s index) were at least moderately similar, whereas 7% of impoundment–natural marsh comparisons were moderately similar. A difference in tidal influence was the only parameter in the best-fitting model describing the observed Morisita’s indices. The index of similarity decreased by 63% when tidal influence differed between compared assemblages. Species richness and diversity were greater in impoundments than natural marshes, but evenness was similar between habitat types. Our results support the hypothesis that resident fish assemblages in impounded wetlands and natural marshes are different, and suggest that a degree of tidal influence is the most important factor behind the difference. Introduction In the late 18th century, thousands of hectares of estuarine marsh along the southeastern coast of the United States were ditched, drained, and impounded for rice cultivation and mosquito control. Although a large number of these impoundments have been abandoned and now are subject to tidal inundation, about 28,000 ha of coastal impoundments are still intact in South Carolina (DeVoe et al. 1987, Kelley 1999, Tiner 1977). Water-level manipulation strategies in South Carolina impoundments, which are designed to benefit migratory waterfowl, greatly reduce 1Georgia Cooperative Fish and Wildlife Research Unit, Warnell School of Forestry and Natural Resources, 180 East Green Street, The University of Georgia, Athens, GA 30602- 2152. 2Current address - New York Cooperative Fish and Wildlife Research Unit, Department of Natural Resources, B02 Bruckner Hall, Cornell University, Ithaca, NY 14853. 3US Geological Survey, Georgia Cooperative Fish and Wildlife Research Unit, Warnell School of Forestry and Natural Resources, 180 East Green Street, The University of Georgia, Athens, GA 30602-2152. *Corresponding author - jennings@uga.edu. Manuscript Editor: Andrew Rypel Southeastern Naturalist K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 238 tidal exchange within these structures and can affect the fish and invertebrate assemblages within the managed area. For example, in a comparison of former rice fields on the Cooper River, SC, the amount of water export from an intact impoundment was between 0.56% and 1.01% of the amount of water export from breached former rice fields (Joyner 2007). Reduced tidal exchange decreases fish and invertebrate access to managed wetlands, creates unfavorable abiotic conditions such as salinity extremes and low dissolved oxygen, and reduces population exchange with adjacent unaltered environments (Wenner et al. 1986). Additionally, the manipulation of water levels causes the plant community to shift from emergent halophytic macrophytes to submerged aquatic vegetation (SAV) found more commonly in freshwater and brackish environments (Portnoy 1999, Rogers et al. 1994, Rozas and Minello 1999). Water-level manipulations in impoundments encourage vegetative communities that are beneficial for waterfowl (Tiner 1977) and that probably favor certain species of crustaceans and fish (Rogers et al. 1994). Although many studies (McGovern and Wenner 1990, Robinson and Jennings 2012, Rogers et al. 1992, Rozas and Minello 1999, Wenner et al. 1986) have documented the negative effects of impoundments on marine-transient fishes (i.e., those species that use the estuary for a portion of their life cycle), fewer studies have focused on the resident fishes within these structures. Impoundments may provide some potential benefits to resident fishes. These benefits include greater abundances of prey associated with SAV, reduced competition for food resources, and decreased predation risk because of the lower overall abundance of fish in the impoundments (Hoese and Konikoff 1995, Rozas and Minello 1999). For example, phytoplankton and zooplankton are often more readily available in impoundments and may therefore represent more stable environments for certain types of fishes (Herke et al. 1992, Milgarese and Sandifer 1982, Talbot et al. 1986). Therefore,environmental conditions in impounded marshes, whether beneficial or detrimental to resident fish species, have the potential to alter assemblage structure. Reduced immigration into impounded wetlands (McGovern and Wenner 1990, Rozas and Minello 1999, Wenner et al. 1986), as well as impoundment-specific environmental characteristics, may alter the resident fish assemblages within these structures (Rogers et al. 1994, Rozas and Minello 1999). For example, previous researchers have suggested that resident fish assemblages within estuarine impoundments that lack diel tidal cycles may become dominated by poeciliid fishes at the expense of killifishes (Kneib 1997). Poeciliids are often dominant or codominant in lower-latitude, low-energy marshes, such as those on the coasts of the Gulf of Mexico and Florida, but are not as abundant in the marshes of the Carolinas (see Nordlie 2003 for review). Considering these documented effects of impoundments on marsh-fish assemblages, our objectives for this study were to: 1) compare the resident fish assemblages in two mesohaline waterfowl impoundments along the Combahee River, SC, with resident assemblages in natural marsh habitats along the southeastern coast of the Southeastern Naturalist 239 K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 US, and 2) determine the factors that influence any observed differences in assemblage structure. We predicted that the resident fish assemblages within our study impoundments would differ from assemblages collected in natural marsh habitats. For example, we expected that stress-tolerant species would be more abundant in the impoundments than in the natural marsh habitats, where they do not have a competitive advantage. We also expected that the differences in tidal influence and salinity would drive any observed dissimilarities in fish assemb lages. Field-site Description The two mesohaline study impoundments, Nieuport and Big Rice Field (BRF), were located on the banks of the Combahee River in Beaufort County, SC, ≈48–53 river-km from the coast (Fig. 1). Nieuport was much larger (≈118 ha) than BRF (≈48 ha). The Combahee River is part of the ACE Basin watershed, so named because it drains the Ashepoo, Combahee, and Edisto rivers. Much of the land around the ACE Basin is protected through efforts of private and public entities, and as such, this study system has experienced fewer recent anthropogenic influences than other estuarine areas in coastal South Carolina. Both study impoundments were managed to enhance the growth of forage, including Ruppia maritima L. (Widgeongrass), for migratory waterfowl and other shore birds and water birds. These impoundments were of the same basic design: an Figure 1. Map of the study impoundments (Nieuport and Big Rice Field) along the Combahee River, SC, sampled 2008–2010. Southeastern Naturalist K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 240 elevated central marsh composed of open-water areas and marsh edge, surrounded by canals that were maintained at desired water levels via a system of dikes and trunks. In each impoundment, two rice-field trunks (McGovern and Wenner 1990) spanned the dikes and were used to lower water levels in the spring to expose the central marsh, which allowed seed germination and oxidation to occur. The impoundments were reflooded throughout the summer to reach a depth of ≈0.5 m above the interior marsh in the fall. Additionally, the trunks remained slightly open throughout the year, and riser boards placed on the impoundment side of each trunk maintained the desired water level. This procedure allowed for some water exchange to limit hypoxia. Within each impoundment, we chose three sections of the perimeter canal as our study sites. Site selection was constrained primarily by our ability to sample by boat, and the sites were as close to equidistant from one another as possible. During initial site delineation, when water was contained completely in the canals, Nieuport sites were 0.09, 0.10, and 0.27 ha, respectively, with average depths of 0.80 (0.11 SD), 1.48 (0.05), and 1.64 m (0.09), respectively. Each site at BRF was 0.11 ha, and the three sites had average depths of 0.82 (0.12), 1.00 (0.60), and 2.08 m (0.11). Methods Data collection We sampled each impoundment with rotenone in May from 2008 to 20 10 when water was confined to the canals. For each sampling effort, we blocked off each of the 6 canal sections with two 6.4-mm mesh block-nets (4.3 m deep x 30.5 or 45.7 m wide). We used a battery-operated pesticide sprayer to apply liquid rotenone (Prenfish Toxicant®; Prentiss Incorporated, Floral Springs, NY; 1 ppb) under the surface within each canal section. We used dip nets (6.4-mm mesh) to collect all fish as they rose to the surface. Fish collection continued until less than 3 specimens rose to the surface within a 20-min time period. We applied potassium permanganate (Argent Chemical Laboratories, Redmond, WA; 3 ppb) within the sampled areas to neutralize the rotenone. The total area sampled with rotenone represented less than 1.2% of the total impoundment area in each of the impoundments. We grouped all collections by species and counted the number of fish per species. We placed any fish that we could not identify in the field in 95% ethanol and transported them back to the laboratory for identification. Generally during rotenone sampling, researchers retrieve fish from the water both the day the rotenone is applied and again the next day. However, we observed Alligator mississipiensis (Daudin) (American Alligator) and many species of water birds consuming the limited fish kill that surfaced outside the block nets and we became concerned that many of the fish that otherwise would have been collected on the second day would be consumed. Thus, for this study, we confined our sampling efforts to a first day pick-up to avoid the sample bias resulting from fish consumption by wildlife. During each sampling effort, we saw American Alligators consuming fish at all sample sites, which may have introduced some bias into our sampling. Because Southeastern Naturalist 241 K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 this disturbance occurred in all impoundments and we did not see more than one American Alligator in any canal section, this effect was probably equal across sites. We measured the water quality of each canal section within each impoundment during each sampling event. We measured temperature, salinity, and dissolved oxygen (DO) with a YSI-85® (Yellow Springs Instruments, Inc., Yellow Springs, OH) handheld meter. We took all measurements just below the water surface and reported the averages of the water-quality measurements taken in each of the 3 canal sections within each impoundment. We performed a nested analysis of variance (ANOVA) on each water-quality variable to analyze differences among impoundments while taking into account differences among sampling years. When ANOVAs produced significant differences (α = 0.05), we further analyzed the results with Tukey’s HSD. We used R (R Development Core Team, 2010) to perform these analyses. We used data from a total of 12 fish assemblages sampled from North Carolina and South Carolina for comparison to our impoundment resident assemblages. Weinstein et al. (1980) used block nets and either 1.0-mm mesh seines or rotenone to collect fish-assemblage data from 11 tidal-creek stations in the Cape Fear River estuary in North Carolina monthly for 20 months. Additionally, Wenner et al. (1986) performed a study of tidal creek and impoundment fish communities in the North Inlet estuary of South Carolina. We used the fish assemblage data collected from their tidal-creek rotenone survey. Wenner et al. (1986) used 0.8-mm mesh block-nets and rotenone to conduct monthly surveys (n = 17) in three small subtidal creeks that drained Chainey Creek. We chose these assemblages for comparison to our impoundment samples for many reasons. These studies represented two of the few studies available that have quantified resident fish abundances in North and South Carolina estuaries. We focused solely on tidal creeks in North and South Carolina because the tidal ranges in these regions were similar. Florida’s topography, which also contained many impounded wetlands, created a very different tidal hydrology. The studies we chose provided the most recent estimations of fish-assemblage abundances in the Carolinas (Nordlie 2003). Rotenone was used to collect a portion of the fish samples in these studies, which allowed us to reduce gear bias in our comparisons. Rotenone sampling has largely been replaced by other methods (e.g., gill nets, electrofishing, and seines) in recent studies. Additionally, all three of the locations used in our analyses were located within the same biogeographic region, and salinities and temperatures recorded in the three river systems were similar (Table 1). Salinities in the estuarine portion of the Combahee River ranged from 2.0 to 20.7 psu (K.F. Robinson and C.A. Jennings, unpubl. data; Upchurch and Wenner 2008), whereas salinities in the North Carolina study ranged from 1.7 to 22.4 psu (Weinstein et al. 1980) and salinities in the South Carolina study ranged from 0.0 to 31.9 psu (Wenner et al. 1986). The vegetative communities of the tidal marshes in the three study locations were dominated by Spartina alterniflora Loisel (Smooth Cordgrass) and Juncus roemerianus Scheele (Needlegrass Rush) (McGovern and Wenner 1990, Upchurch and Wenner 2008, Weinstein 1979), and the sample stations were Southeastern Naturalist K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 242 located in first-order tidal creeks. The study in South Carolina provided data on fish assemblages sampled adjacent to waterfowl impoundments in a wildlife preserve, which provided a similarly low level of anthropogenic influence as was observed in our study sites in the ACE Basin. The estuarine portion of the Cape Fear River was ≈45 river km in length (Weinstein 1979), which was similar to the estuarine portion of the Combahee River (≈61 river km; Upchurch and Wenner 2008), and the Cape Fear River estuary contained the largest tidal salt marsh in North Carolina. Although these two studies provided pooled abundances of resident fishes from 17 (Wenner et al. 1986) to 20 months (Weinstein et al. 1980) of sampling, the resident assemblages sampled in these published studies did not vary greatly in species proportions over the course of sampling. Assemblage comparisons and model selection We retained only those species described as permanent resident or freshwater transient in Nordlie (2003) for all analyses. We calculated a suite of community indices to describe the 18 fish assemblages from impounded and tidal creek Table 1. Sampling method, mean salinity, temperature, and DO (1 SD) for each of the natural and impounded marsh locations sampled for fish-assemblage comparisons. Impoundment samples were from spring sampling in Nieuport and Big Rice Field (BRF), on the Combahee River, SC, 2008–2010. For each impoundment, values are the average of point estimates from three sites. The Chainey Creek values are the range from 24 months of sampling in a tidal creek in North Inlet Estuary, SC. Data are from the study of Wenner et al. (1986). Averages for Baldhead Creek, Barnard’s Creek, Cape Creek, Dutchman Creek, Governor’s Creek, Hechtic Creek, Town Creek, and Walden Creek are from monthly measurements over 20 months from samples in tidal creeks in the Cape Fear River Estuary, NC. Data are from the study of Weinstein et al. (1980). NA = data not given in original paper. Sampling Location Habitat type method Salinity (psu) Temp.(°C) DO (mg l-1) Nieuport 2008 Impounded Rotenone 5.0 (0.7) 26.7 (1.4) 3.2 (1.5) Nieuport 2009 Impounded Rotenone 7.5 (0.5) 25.8 (0.5) 9.5 (2.2) Nieuport 2010 Impounded Rotenone 3.0 (1.3) 31.0 (1.3) 5.8 (4.3) BRF 2008 Impounded Rotenone 13.0 (1.8) 26.4 (1.7) 8.5 (1.4) BRF 2009 Impounded Rotenone 5.4 (1.1) 25.1 (1.7) 9.0 (0.9) BRF 2010 Impounded Rotenone 10.8 (0.2) 27.5 (2.2) 7.4 (1.2) Chainey Creek Natural Rotenone 0.0–31.9A 4.0–32.4A NA Baldhead Creek Natural Rotenone 22.0 (6.4) 19.9 (8.4) NA Barnard’s Creek Natural Rotenone 1.7 (2.3) 17.4 (7.3) NA Cape Creek Natural Rotenone 22.4 (5.5) 21.1 (8.6) NA Dutchman Creek Natural Rotenone 14.2 (6.8) 18.8 (7.8) NA Dutchman Creek Natural Seine 19.3 (5.5) 18.4 (8.1) NA Governor’s Creek Natural Seine 2.2 (2.3) 19.7 (8.4) NA Hechtic Creek Natural Seine 2.5 (3.2) 19.0 (7.9) NA Town Creek Natural Rotenone 2.2 (3.3) 18.1 (9.3) NA Walden Creek Natural Rotenone 4.8 (3.0) 20.1 (8.4) NA Walden Creek 1 Natural Seine 5.0 (3.2) 19.9 (8.0) NA WaldenCreek 2 Natural Seine 4.4 (2.7) 19.3 (8.2) NA AMean salinity and temperature were not given, only the range of measurements for the entire study were reported. Southeastern Naturalist 243 K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 habitats. We calculated rarefied species richness curves for each assemblage and standardized the species richness for each assemblage to 100 individuals because the smallest number of fish collected at an individual site was 112. Additionally, we used Shannon’s index (H') to assess assemblage diversity, s H' = -Σ (pi)(loge pi), i = 1 where S = total number of species and pi = the proportion of the total sample represented by the ith species (Kwak and Peterson 2007). We calculated H' and rarefied species richness curves with the vegan package (Oksanen et al. 2010) for R (R Development Core Team 2010). We calculated the evenness of each assemblage, based on the Shannon index: H' J ' = logeS We compared rarefied species richness, H', and J' between habitat types (impoundments versus tidal creeks) with Welch’s two-sample t-tests (α = 0.05). We used Morisita’s index of similarity and percent similarity to compare the resident assemblages of our study impoundments with those of the previous studies of natural marshes. Similarity indices have been shown to diverge from one another (Bloom 1981); we calculated two separate indices of similarity to account for this potential divergence. Morisita’s index ranges from 0 (assemblages are completely dissimilar) to 1 (assemblages are identical). Percent similarity ranges from 0% (zero species in common) to 100% (assemblages are identical). We chose these two indices because they have been described as the most robust and best measures of similarity (Krebs 1998, Kwak and Peterson 2007). Our assemblages contained size displacements and many species collected in small numbers; unlike the more widely used Bray-Curtis coefficient, neither of the indices is sensitive to these characteristics of the data (Kwak and Peterson 2007, Wolda 1981). Additionally, Morisita’s index provides a measure of similarity in terms of a probability: the ratio of the probability that an individual drawn from assemblage 1 and an individual drawn from assemblage 2 will be the same species to the probability that two individuals drawn from either assemblage 1 or 2 will be the same species (Krebs 1998, Kwak and Peterson 2007). Based on this ratio, we chose similarity cutoffs for both indices of 0.25 (somewhat similar), 0.50 (moderately similar), and 0.75 (very similar) to describe the similarities of the compared assemblages. Morisita’s index uses abundance data (total number of individuals) while percent similarity uses relative species abundances (Kwak and Peterson 2007, Wolda 1981). We included rare species in the analysis because indices of similarity did not differ between calculations in which rare species were included or excluded. We used the vegan package (Oksanen et al. 2010) for R (R Development Core Team 2010) to calculate Morisita’s index of similarity. We conducted pairwise comparisons of the resident assemblages from each of three years (2008–2010) of May samples from Nieuport and BRF, and the 12 assemblages from North Carolina (Weinstein Southeastern Naturalist K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 244 et al. 1980) and South Carolina (Wenner et al. 1986). The output from R for the calculation of Morisita’s index provided the dissimilarity index, so we subtracted all calculated indices from one to obtain the similarity index. For brevity, we refer to each impoundment collection by the first letter of the impoundment (B or N) and the last two digits of the collection year (08, 09, or 10). We created a set of hypotheses and a corresponding candidate set of linear models that might explain the observed similarity indices, and chose to use Morisita’s index for this analysis because it provides similarities in terms of a probability. The models used linear regression analysis (Neter et al. 1990) to describe the relationships among the observed Morisita’s indices and the following variables: the absolute difference in average salinity between compared sites (maximum recorded salinity was used for the data from Wenner et al. 1986), a categorical variable to describe differences in fish-collection methods (0 = same methods between compared sites, 1 = different methods between compared sites), and a categorical variable to describe differences in tidal influence between sites (0 = both sites impounded or natural, 1 = impounded to natural site comparison). We used an information theoretic approach to assess the relative fit of each candidate model. The global model, which contained all of the variables thought to influence assemblage similarity, was used to create the candidate set of models, which included all combinations of the above variables. We assessed this global model for goodness-of-fit by examination of the residuals and normal probability plots. The pairwise comparisons of sites might not be independent because all pairs that include a particular site may have similar Morisita values, which would require the use of hierarchical linear models. An analysis of variance (ANOVA) of the residuals of the global linear model by site was not significant (P = 0.61), which indicated that the use of a linear model was appropriate in our case. We evaluated the fit of each model in the candidate set by calculating Akaike’s information criterion (AIC; Akaike 1973) with the small-sample bias adjustment (AICc; Hurvich and Tsai 1989). We assessed Pearson correlations for all pairs of predictor variables to avoid multicollinearity, and used only uncorrelated predictor variables (r < 0.40) in the candidate models (all predictor variables were uncorrelated in this case; Moore and McCabe 1993). We calculated Akaike weights (wi, range = 0–1; Burnham and Anderson 2002) to assess the relative fit of each of the candidate models; the best-fitting model had the highest Akaike weight. We incorporated model-selection uncertainty by calculating the modelaveraged estimates of the linear regression model coefficients and standard errors (Burnham and Anderson 2002). We created a confidence set of models that contained all models whose weights were within 10% of the highest weight (Thompson and Lee 2000). The coefficients of the estimates and their standard errors were weighted by their Akaike weights (wi) and summed across the confidence set of models to obtain the composite model. Our interpretations were based on this composite model. We calculated 95% confidence intervals based on the t-statistic with n - 1 degrees of freedom to assess the precision of our model-averaged coefficient estimates. Any confidence intervals that spanned zero Southeastern Naturalist 245 K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 indicated that the associated parameter’s relationship with similarity indices (positive or negative) could not be inferred. All statistical analyses were performed in R (R Development Core Team 2010). Results Data collection The average spring salinity in our study impoundments ranged from 3.0 psu (1.3 SD) to 13.0 psu (1.8) throughout the three years of fish collections (Table 1). Salinity ranged from 1.7 psu (2.3) to 19.3 psu (5.5) in the North Carolina study areas (Weinstein et al. 1980) and from 0.0 psu to 31.9 psu in the South Carolina study areas (Wenner et al. 1986). Temperature ranged from 25.1 °C (1.7) to 31.0 °C (1.3) in our study impoundments, from 17.4 °C (7.3) to 21.1 °C (8.6) in the North Carolina tidal creeks (Weinstein et al. 1980), and from 4.0 °C to 32.4 °C in the South Carolina tidal creek (Wenner et al. 1986). DO ranged from 3.2 mg/L (1.5) to 9.5 mg/L (1.5) in our study impoundments, but DO was not reported for the other study areas. Nieuport and Big Rice Field were statistically similar with respect to all water-quality variables, as indicated by the results of the nested ANOVAs for each variable (salinity [F = 3.02, df = 1, P = 0.157], temperature [F = 0.71, df = 1, P = 0.447], DO [F = 1.30, df = 1, P = 0.318]). Salinity varied among sampling years (F = 26.43, df = 4, P < 0.001), as did temperature (F = 5.88, df = 4, P = 0.007). DO did not vary among years (F = 3.20, df = 4, P = 0.053). We collected 7946 resident fishes representing 12 species in Nieuport and 1258 resident fishes representing 15 species in BRF (Table 2). The most abundant resident species varied among years in each impoundment, though four species— Table 2. Numbers of resident species captured during spring rotenone sampling in Nieuport and Big Rice Field impoundments in the Combahee River, SC, 2008–2010. Big Rice Field Nieuport Species Common name 2008 2009 2010 2008 2009 2010 Lepisosteus osseus (L.) Longnose Gar 2 4 7 2 18 10 Myrophis punctatus Lütken Speckled Worm-eel 0 0 2 0 0 0 Dorosoma cepedianum (Lesueur) Gizzard Shad 1 3 0 4 7 14 Carassius auratus (L.) Goldfish 0 154 0 0 0 0 Cyprinus carpio L. Common Carp 0 2 0 0 1 0 Notemigonus crysoleucas (Mitchill) Golden Shiner 0 0 0 0 0 1 Ameiurus catus (L.) White Catfish 0 0 1 1 13 24 Ameiurus nebulosus (Lesueur) Brown Bullhead 0 0 3 0 0 0 Menidia beryllina (Cope) Inland Silverside 54 56 42 461 683 224 Fundulus heteroclitus L. Mummichog 0 0 1 0 0 1 Lucania parva (Baird and Girard) Rainwater Killifish 4 0 15 3 0 1115 Gambusia holbrooki Girard Eastern Mosquitofish 16 203 35 272 117 346 Poecilia latipinna (Lesueur) Sailfin Molly 132 501 5 89 3365 1062 Cyprinodon variegatus (Lacepède) Sheepshead Minnow 1 1 0 0 40 72 Syngnathus scovelli (Evermann and Gulf Pipefish 10 1 0 0 1 0 Kendall) Gobiosoma bosc (Lacepède) Naked Goby 0 1 1 0 0 0 Southeastern Naturalist K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 246 Gambusia holbrooki (Eastern Mosquitofish), Lucania parva (Rainwater Killifish), Poecilia latipinna (Sailfin Molly), and Menidia beryllina (Inland Silverside)— tended to dominate in our samples. In Nieuport, Inland Silverside had the greatest density (0.08/m2) in 2008, though overall, all species had low densities that year. Sailfin Molly had the greatest density (0.99/m2) in Nieuport in 2009, and Rainwater Killifish (0.38/m2) and Sailfin Molly (0.32/m2) dominated samples of resident species in 2010 (Fig. 2a). In BRF, Sailfin Molly had the greatest density in 2008 (0.04/m2) and 2009 (0.16/m2), whereas Eastern Mosquitofish (0.01/m2) and Inland Silverside (0.01/m2) had the greatest densities in 2010 (Fig. 2b). Figure 2. Densities (±1 SE) of the four most abundant resident fishes—Gambusia holbrooki (Eastern Mosquitofish), Lucania parva (Rainwater Killifish), Poecilia latipinna (Sailfin Molly), and Menidia beryllina (Inland Silverside)— collected in spring rotenone samples in a) Nieuport and b) Big Rice Field impoundments, Combahee River, SC, 2008 – 2010. Note the difference in y-axis ranges in the two plots. Southeastern Naturalist 247 K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 Assemblage comparisons Rarefied species richness (n = 100) ranged from 1.0 to 9.7 species (Fig. 3a). Mean rarefied species richness was significantly greater in impoundments (6.0 [2.0 SD]), than in tidal creeks (3.4 [1.9]; t = 2.6, df = 9.58, P = 0.026). H' diversity ranged from 0.00 to 1.60 (Fig. 3b), and J' evenness ranged from 0.16 to 0.93 (Fig. 3c). Mean diversity was significantly greater in impoundments (1.17 [0.31]) Figure 3. a) Rarefied species richness (n = 100), b) Shannon (H') diversity, c) J' evenness calculated for fish assemblages captured during spring rotenone sampling in Big Rice Field (B) and Nieuport (N) impoundments, Combahee River, SC, 2008–2010, and fish assemblages sampled with seine (S) and rotenone (R) for previous studies of tidal creeks in Cape Fear River estuary, NC, (Hechtic, Governor’s, Walden, Dutchman, Barnard’s, Town, Cape, and Baldhead creeks) and North Inlet estuary, SC, (Chainey Creek). Evenness was not calculated for Dutchman (R) or Walden (R) because only one species was collected in these sites. Data from North Carolina are from the study of Weinstein et al. (1980) and data from South Carolina are from the study of Wenner et al. (1986). Southeastern Naturalist K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 248 than in tidal creeks (0.68 [0.41]; t = 2.8, df = 12.82, P = 0.016), but evenness did not differ between habitat types (impoundments = 0.53 [0.13], tidal creeks = 0.55 [0.32]; t = 0.1, df = 12.84, P = 0.899). Similarity indices for comparisons of resident fish assemblages among impoundments ranged from 0.23 to 0.96 for Morisita’s index and from 24 to 80% for percent similarity (Appendices A, B). Overall, 40.0% (percent similarity) to 53.3% (Morisita’s index) of the among-impoundment comparisons were at least moderately similar (Table 3). Comparisons among all tidal creek assemblages, regardless of sampling method, ranged from 0.39 to 1.00 for Morisita’s index and from 32 to 100% for percent similarity (Appencies C, D). Overall, 63.6% (percent similarity) to 85.7% (Morisita’s index) of the comparisons among all tidal creek assemblages were at least moderately similar (Table 3). Resident fish assemblages sampled with rotenone in impoundments and natural marshes were dissimilar; only 2.4% (percent similarity) to 4.8% (Morisita’s index) of these comparisons were at least moderately similar (Table 3). The exceptions were the comparisons between N08 and Town Creek (Morisita index = 0.75, percent similarity = 55%), and B10 and Town Creek (Morisita index = 0.62) (Appendices E, F). In contrast, 52.4% (percent similarity) to 76.2% (Morisita index) of the comparisons among rotenone-sampled natural marsh assemblages were at least moderately similar (Table 3). When comparing the rotenone-sampled resident assemblages in impoundments to natural marsh assemblages sampled by seine, 6.7% of the comparisons (Morisita’s index and percent similarity) were moderately similar. N08 (Morisita index = 0.83, percent similarity = 68%) and B10 (Morisita index = 0.79, percent similarity = 61%) were moderately to very similar to the Governor’s Creek resident assemblage (Apendices E, F). Conversely, 70.0% (percent similarity) to 90.0% (Morisita’s index) of the comparisons among tidal creek assemblages Table 3. Percent of fish-assemblage comparisons using Mosita’s index and percent similarity index: Imp = among our study impoundments (Big Rice Field and Nieuport, Combahee River, SC); Imp– NM = between our study impoundments and natural marsh habitat from previously published studies (Cape Fear River estuary, NC, and North Inlet estuary, SC); and NM = among natural marshes from previously published studies, whose similarity indices fall in the categories of ≥0.75, 0.50–0.75, 0.25–0.50, and 0.00–0.25 for samples collected by rotenone and seine. All NM = the comparisons of marshes sampled by both rotenone and seine. Rotenone Seine Index Imp Imp–NM NM Imp–NM NM All NM Morisita index ≥0.75 26.7 2.4 47.6 6.7 60.0 56.1 0.50–0.75 26.7 2.4 28.6 0.0 30.0 27.3 0.25–0.50 40.0 2.4 23.8 3.3 10.0 16.7 0.00–0.25 6.7 92.9 0.0 90.0 0.0 0.0 Percent similarity index ≥75% 6.7 0.0 4.8 0.0 30.0 18.2 50–75% 33.3 2.4 47.6 6.7 40.0 45.5 25–50% 53.3 2.4 47.6 6.7 30.0 36.4 0–25% 6.7 95.2 0.0 86.7 0.0 0.0 Southeastern Naturalist 249 K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 that were sampled with seines were moderately similar (Table 3). Among all comparisons, the observed percent similarities were slightly lower than the Morisita’s indices, but the patterns of similarity were the same between the two methods (Appendices A–F). Model selection The model containing only a difference in tidal influence was the most plausible linear regression model explaining the observed Morisita’s indices (wi = 0.765; Table 4). The model of a difference in tidal influence was 6.1 times more likely to provide the best explanation for similarity indices than was the global model, which was the next most plausible model. These two models and the model that considered a difference in tidal influence and salinity made up the confidence set of models; the other models did not have sufficient evidence to be considered as plausible explanations for differences among resident fish assemblages. Because the global model was a part of the confidence set of models, all parameters were included in the composite model (intercept = 0.72, 95% CI = 0.67–0.77; coefficients: tidal influence = -0.63, 95% CI = -0.69 to -0.57; method = 0.07, 95% CI = -0.08–0.21; salinity = 0.00, 95% CI = 0.00– 0.01). The observed Morisita’s index of similarity values were most affected by tidal influence. The estimated coefficient for tidal influence indicated that Morisita’s index of similarity decreased by 0.63 when comparisons were made between impounded and natural marshes. Sampling method had a very weak positive influence on the index of similarity (0.07), but the confidence intervals spanned zero, which indicates that comparisons between our samples collected by rotenone and natural marsh assemblages collected by seine were valid. Salinity appeared to have very little to no influence on assemblage similarity. Discussion The results of our study indicated that, as expected, the resident fish assemblages in our study impoundments were dissimilar to resident assemblages in natural tidal creek systems along the southeastern coast of the United States. In addition to the habitat-specific differences indicated by the estimates of two indices of similarity, rarefied species richness and H' diversity were significantly greater in the impoundments than the tidal creeks. Poeciliids were the most abundant species in the study impoundments in all samples except N08 and B10, whereas Fundulus Table 4. Candidate models of parameter influences on Morisita’s index of similarity values obtained for comparisons of resident fish assemblages in impounded and natural wetlands, number of parameters in each model (K), deviations from best-fitting model (Δ i), and Akaike weights (wi). Model K Δi wi Difference in tidal influence 3 0.000 0.765 Differences in tidal influence, salinity and method 5 3.629 0.125 Differences in tidal influence and salinity 4 3.864 0.111 Difference in method 3 194.885 0.000 Difference in salinity 3 194.567 0.000 Southeastern Naturalist K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 250 heteroclitus (L.) (Mummichogs) were most abundant in all but two (Governor’s and Town creeks) of the natural marsh assemblages (Weinstein et al. 1980, Wenner et al. 1986). Inland Silverside was most abundant in N08, B10, and Governor’s and Town creeks, and these four assemblages ranged in similarity (Morisita’s index) from 0.62 to 0.93. Results of our model-selection analysis showed that the index of similarity between two resident fish assemblages decreased by more than half of the range of the index when there was a difference in tidal influence. We hypothesized that salinity would also play a role in the observed differences among fish assemblages, but the results of our model-selection analysis did not support this expectation. Our results corroborated the assertion of Weinstein et al. (1980) that gear bias was negligible for seine- and rotenone-sampled fish assemblages collected in Cape Fear River tidal creeks because 85.7% of the comparisons among tidal creeks had Morisita’s index values ≥0.50. The results of our model-selection analysis indicated that tidal influence was the factor driving differences in fish-assemblage structure between our study impoundments and natural marsh habitats in North and South Carolina. Tidal influence was drastically reduced in the impoundments because managers held the interior marsh of these structures in the early stages of the drought-recovery cycle that is characteristic of temporary wetlands (Batzer et al. 2006). This reduction in tidal influence can alter estuarine-community structure by changing colonization/ extinction dynamics and by shifting the dominant vegetation in the impoundments from emergent-marsh vegetation to submerged aquatic vegetation (SAV; Rozas and Minello 1999, Weaver and Holloway 1974). The difference in tidal influence manifests partially as an issue of connectivity. The lack of connectivity of our impoundments to the salt marsh for most of the year means that colonization might be an important factor influencing impoundment resident-fish assemblage structure. Species had a brief window of time to enter impoundments during fall flooding, and species that were most abundant during this time immigrated into the impoundments in greater numbers than those that were less abundant (Rogers et al. 1992, Wenner et al. 1986). Connectivity, through reduced tidal influence, is a large-scale process that helps determine the local species available for establishment in impoundments (Baber et al. 2002, Snodgrass et al. 1996, Taylor 1997). This process provides a potential explanation for why salinity was not driving the difference in assemblage structure between residents in impounded and natural marsh habitats: it likely was overshadowed by tidal influence. For example, the bulk of the tidal creek assemblages were dominated by Mummichogs (Weinstein et al. 1980, Wenner et al. 1986), which were almost completely absent from the impoundment resident assemblages. Mummichogs are a dominant member in the tidal creek communities of the ACE Basin. In a study in the Ashepoo River (part of the ACE Basin) concurrent with our impoundment sampling, Goldman et al. (2010) used minnow traps set in tidal creeks to capture >140,000 Mummichogs during a 9-mo period. Their catch per unit effort was lowest in September, which is typically when impoundments are flooded. In addition, we captured very few Southeastern Naturalist 251 K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 Mummichogs in baited minnow traps set in tidal creeks across the river from Nieuport and Big Rice Field in September, whereas this species dominated springtime trap-sets in these low salinity tidal creeks (2–6 psu; K.F. Robinson and C.A. Jennings, unpubl. data). If this seasonal pattern of Mummichog abundance is typical of the ACE Basin, the lack of Mummichogs in our study impoundments could be a direct result of the lack of diel tidal exchange. The current water-management regime could be selecting against the most dominant member of the natural marsh resident-fish community. In conjunction with colonization, extinction-like processes also might have influenced the structure of the resident fish assemblages in our study impoundments. In the summer, the majority of the fish in these structures perish (Robinson 2011). This large-scale mortality can be attributed to poor-quality habitat (e.g., declining DO levels) and predation by birds and other aquatic animals (Robinson 2011). Conditions in natural marsh habitats can be stressful for most fish species, but unlike fish in natural marshes, fish in impoundments are unable to emigrate out of harsh conditions (Rogers et al. 1992, Wenner et al. 1986). This yearly disturbance created an assemblage structure with characteristics of a community living in a frequently disturbed habitat (Menge and Sutherland 1976, Poff and Ward 1989, Snodgrass et al. 1996, Wiens 1984). In frequently disturbed communities, biotic interactions are overshadowed by the disturbance (Snodgrass et al. 1996) and colonization–extinction dynamics become important. In addition to encouraging colonization–extinction dynamics, the water-level management scheme practiced in the study impoundments used reduced tidal influence to replace the emergent-marsh vegetation found in natural salt marshes with SAV, mostly Widgeongrass. SAV is one of the most important variables in predicting relative abundances of small resident fish species such as Sailfin Molly and Eastern Mosquitofish in natural marsh and impounded habitats (Castellanos and Rozas 2001, Harrington and Harrington 1982, Kanouse et al. 2006, Rozas and Minello 1999). In the study impoundments, four species were most abundant each year: Eastern Mosquitofish, Sailfin Molly, Rainwater Killifish, and Inland Silverside. These species have been described as rapid colonizers of habitat because of their omnivorous feeding habits and ability to withstand stressful conditions such as hypoxia and extremes in salinity and temperature (Felley and Daniels 1992, Martin et al. 2009, Meffe and Snelson 1989, Ruetz et al. 2005). In numerous studies in impounded wetlands as well as in low-energy Gulf of Mexico marshes, Eastern Mosquitofish (Kanouse et al. 2006, Rozas and Minello 1999, Weaver and Holloway 1974), Sailfin Molly (Kanouse et al. 2006, Rogers et al. 1994, Rozas and Minello 1999, Weaver and Holloway 1974), and Rainwater Killifish (Castellanos and Rozas 2001, Rogers et al. 1994, Rozas and Minello 1999) were much more abundant in SAV than in emergent marsh or open water habitats. Although we did not quantify SAV in the study impoundments, some interesting trends in the fish assemblages among years follow our qualitative observations of SAV (Widgeongrass) cover. The resident fish assemblages in Nieuport in 2008 and Big Rice Field in 2010 were more similar to each other (0.93) than they were Southeastern Naturalist K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 252 to any other impoundment sample (0.23–0.53). These two assemblages had lower resident-fish densities and were uniquely dominated by Inland Silverside. We observed much lower cover of SAV in these two impoundments in 2008 and 2010. Inland Silverside prefers emergent-marsh vegetation to SAV (Castellanos and Rozas 2001) and has been classified as an edge-marsh user because it rarely moves into interior marshes (Peterson and Turner 1994). The proportion of interior and edge marsh, as well as the amount of SAV in an impoundment, probably influenced which of the suite of stress-tolerant species was most dominant in a given year and helped structure the resident assemblages in impounded wetlands . Each year, we observed gravid females of each of the four species most common in our impoundments— Eastern Mosquitofish, Sailfin Molly, Rainwater Killifish, and Inland Silverside—which suggests that relative abundances of these species may change because of differences in reproduction and not solely through stresstolerance and immigration. To reproduce, Mummichogs require access to marsh pools that are accessible only on very high tides (Taylor et al. 1979), and, therefore, any Mummichogs that entered the impoundments likely could not reproduce, further reducing the relative abundance of the species there. In the future, examining a larger suite of environmental characteristics in conjunction with fish-assemblage sampling may lead to a better understanding of what factors drive the change in resident-assemblage structure that we observed in impounded wetlands as compared to natural marsh habitat. Although the natural marsh studies that we used for our comparisons were older than we would have preferred, the temporal difference likely did not have a major influence on our results. The high degree of similarity observed for comparisons between North Carolina resident assemblages and the assemblage sampled from Chainey Creek, SC, (Morisita index = 0.65–0.94) indicated that resident assemblages from North and South Carolina were comparable, despite the studies being separated by six years and over 100 miles of coastline. Additionally, these two studies represented the most recent assessments of estuarine fish assemblage structure in the Carolinas. Unlike most studies of fish in these habitats, the authors of our comparison studies provided complete descriptions of the fish assemblages; these same papers were recently used to characterize estuarine fish assemblages in the eastern United States (Nordlie 2003). These studies also provided descriptions of southeastern estuarine fish assemblages from relatively pristine habitats. Although the assemblages in specific locations may have been subject to anthropogenic influences in recent years, when these studies were performed, the habitat structure was quite similar to the current state of estuarine habitat in the relatively pristine ACE Basin. The effects of changes in hydrologic connections on estuarine fishes in the southeastern US are understudied. Our results indicate that reduced connectivity and tidal influence can greatly change the composition of resident fish assemblages. Reduced connectivity probably changes the assemblage structure through a combination of reduced migration and alterations in the environment that favor certain species over others. We found that four small resident species comprised the bulk of Southeastern Naturalist 253 K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 the fish within these impoundments and that different species dominated in different years. As habitat fragmentation in estuarine systems along the Atlantic and Gulf coasts of the United States intensifies, the habitat characteristics that create these assemblage-level changes within managed wetlands may become more prevalent in other fragmented marshes. Fragmentation through reduced hydrologic connectivity can affect fish assemblages beyond changes in resident-species composition. For example, these systems act as sinks for many species that use estuaries as nursery grounds (Robinson and Jennings 2012, Wenner et al. 1986). These differences between natural and altered estuarine habitats can result in reduced productivity and reduced export of nutrients from estuaries to coastal waters (Kneib 1997, Robinson 2011). For these reasons, future research seeking to understand the dynamics of species assemblages within impounded wetlands may provide us with valuable information about fish-community dynamics in other fragmented sy stems. Acknowledgments We thank M. Alber, J. Peterson, and S. Schweitzer for input into experimental design and comments on this manuscript. We also thank B. Carswell, G. Crouch, P. Dimmick, J. Dycus, P. Ely, M. Homer, J. Kirsch, E. Mills, M. Mundy, R. Peterson, J. Robinson, J. Ruiz, E. Wiggers, and S. Zimpfer for assistance in field collections and logistical support. Further, we thank A. Overton, J. Robinson, and A. Rypel, and two anonymous reviewers for helpful comments on earlier drafts of this manuscript. 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Pp. 439–457, In D.R. Strong, D. Simberloff, L.G. Abele, and A.B. Thistle (Eds.). Ecological Communities: Conceptual Issues and Evidence. Princeton University Press, Princeton, NJ. 613 pp. Wolda, H. 1981. Similarity indices, sample size, and diversity . Oecologia 50:296–302. Southeastern Naturalist 257 K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 Appendix A. Morisita’s index of similarity calculated for comparisons of resident fish assemblages captured during spring rotenone sampling in Big Rice Field (BRF) and Nieuport impoundments, Combahee River, SC, 2008–2010. Values followed by a = similarity ≥0.75; b = similarity is 0.50–0.75. Nieuport Nieuport BRF 2008 BRF 2009 BRF 2010 2008 2009 BRF 2009 0.89a BRF 2010 0.42 0.37 Nieuport 2008 0.53b 0.41 0.93a Nieuport 2009 0.96a 0.87a 0.23 0.34 Nieuport 2010 0.70b 0.68b 0.48 0.34 0.64b Appendix B. Percent similarity index calculated for comparisons of resident fish assemblages captured during spring rotenone sampling in Big Rice Field (BRF) and Nieuport impoundments, Combahee River, SC, 2008–2010. Values followed by a = similarity ≥0.75; b = similarity is 0.50–0.75. Nieuport Nieuport BRF 2008 BRF 2009 BRF 2010 2008 2009 BRF 2009 68.4 BRF 2010 39.0 33.0 Nieuport 2008 43.6 39.2 73.9b Nieuport 2009 79.9a 63.7b 24.0 30.1 Nieuport 2010 55.2b 55.9b 38.9 31.8 49.3 Southeastern Naturalist K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 258 Appendix C. Morisita’s index of similarity calculated for comparisons of fish assemblages sampled in previous studies of tidal creeks in the Cape Fear River estuary, NC, and the North Inlet estuary, SC. Five North Carolina sites (Dutchman, Walden 1 and 2, Governor’s, and Hechtic creeks) were sampled with seines (S) and six North Carolina sites (Baldhead, Cape, Walden, Dutchman, Town, and Barnard’s creeks) were sampled with rotenone (R). The South Carolina site (Chainey Creek) was sampled with rotenone. Values followed by a = similarity ≥0.75; b = similarity is 0.50–0.75. Data from the Cape Fear River estuary are from Weinstein et al. (1980) and data from the North Inlet estuary are from Wenner et al. (1986). Sampled with seines Sampled with rotenone Dutchman Governor’s Hechtic Walden 1 Walden 2 Chainey Baldhead Cape Walden Dutchman Town Governor’s (S) 0.49 Hechtic (S) 0.93a 0.67b Walden 1 (S) 0.93a 0.52b 0.97a Walden 2 (S) 0.93a 0.57b 0.98a 1.00a Chainey (R) 0.90a 0.68b 0.97a 0.92a 0.94a Baldhead (R) 0.89a 0.45 0.80a 0.78a 0.79a 0.87a Cape (R) 0.87a 0.50b 0.90a 0.88a 0.90a 0.94a 0.89a Walden (R) 0.92a 0.47 0.95a 1.00a 0.99a 0.89a 0.76a 0.86a Dutchman (R) 0.92a 0.47 0.95a 1.00a 0.99a 0.89a 0.76a 0.86a 1.00a Town (R) 0.45 0.90a 0.65b 0.48 0.52b 0.65b 0.39 0.42 0.45 0.45 Barnard’s (R) 0.66b 0.51b 0.71b 0.66b 0.68b 0.75b 0.59b 0.66b 0.64b 0.64b 0.48 Southeastern Naturalist 259 K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 Appendix D. Percent similarity index calculated for comparisons of fish assemblages sampled in previous studies of tidal creeks in the Cape Fear River estuary, NC, and the North Inlet estuary, SC. Five North Carolina sites (Dutchman, Walden 1 and 2, Governor’s, and Hechtic creeks) were sampled with seines (S) and six North Carolina sites (Baldhead, Cape, Walden, Dutchman, Town, and Barnard’s creeks) were sampled with rotenone (R). The South Carolina site (Chainey Creek) was sampled with rotenone. Values followed by a = similarity ≥0.75; b = similarity is 0.50–0.75. Data from the Cape Fear River estuary are from Weinstein et al. (1980) and data from the North Inlet estuary are from Wenner et al. (1986). Sampled with seines Sampled with rotenone Dutchman Governor’s Hechtic Walden 1 Walden 2 Chainey Baldhead Cape Walden Dutchman Town Governor’s (S) 32.4 Hechtic (S) 75.4b 50.6b Walden 1 (S) 75.3b 37.8 82.1a Walden 2 (S) 73.9b 44.1 84.8a 93.6a Chainey (R) 74.2b 32.0 54.2b 53.9b 52.5b Baldhead (R) 61.7b 32.0 62.0b 61.6b 61.6b 63.5b Cape (R) 73.5b 32.0 77.4a 92.6a 87.8a 72.6b 68.1b Walden (R) 73.5b 32.0 77.4a 92.6a 87.8a 65.1b 52.5b 61.6b Dutchman (R) 34.9 77.6a 52.4b 37.1 41.2 65.1b 52.5b 61.6b 100.0a Town (R) 45.4 43.1 56.6b 47.5 51.6b 49.8 34.7 34.7 34.7 34.7 Barnard’s (R) 65.8b 47.7 81.9a 68.3b 72.3b 59.1b 45.1 45.1 45.1 45.1 45.7 Southeastern Naturalist K.F. Robinson and C.A. Jennings 2014 Vol. 13, No. 2 260 Appendix E. Morisita’s index of similarity calculated for comparisons of fish assemblages captured during spring rotenone sampling in Big Rice Field (BFR) and Nieuport impoundments, Combahee River, SC, 2008–2010, and fish assemblages sampled with seine (S) and rotenone (R) for previous studies of tidal creeks in Cape Fear River estuary, NC, (Hechtic, Governor’s, Walden, Dutchman, Barnard’s, Town, Cape, and Baldhead creeks) and North Inlet estuary, SC, (Chainey Creek). Values followed by a = similarity ≥0.75; b = similarity is 0.50–0.75. Data from North Carolina are from the study of Weinstein et al. (1980) and data from South Carolina are from the study of Wenner et al. (1986). Nieuport Nieuport Nieuport BRF 2008 BRF 2009 BRF 2010 2008 2009 2010 Hechtic Creek (S) 0.08 0.02 0.17 0.19 0.05 0.03 Governor’s Creek (S) 0.33 0.21 0.79a 0.83a 0.16 0.19 Walden Creek 1 (S) 0.01 0.01 0.05 0.04 0.01 0.01 Walden Creek 2 (S) 0.03 0.03 0.10 0.10 0.02 0.02 Dutchman Creek (S) 0.00 0.00 0.02 0.00 0.00 0.00 Barnard’s Creek (R) 0.07 0.02 0.14 0.15 0.03 0.02 Town Creek (R) 0.33 0.09 0.62b 0.75a 0.18 0.12 Dutchman Creek (R) 0.00 0.00 0.01 0.00 0.00 0.00 Walden Creek (R) 0.00 0.00 0.01 0.00 0.00 0.00 Cape Creek (R) 0.00 0.00 0.02 0.00 0.00 0.00 Baldhead Creek (R) 0.00 0.00 0.02 0.00 0.00 0.00 Chainey Creek (R) 0.09 0.03 0.18 0.19 0.05 0.04 Appendix F. Percent similarity index calculated for comparisons of fish assemblages captured during spring rotenone sampling in Big Rice Field (BFR) and Nieuport impoundments, Combahee River, SC, 2008–2010, and fish assemblages sampled with seine (S) and rotenone (R) for previous studies of tidal creeks in Cape Fear River estuary, NC, (Hechtic, Governor’s, Walden, Dutchman, Barnard’s, Town, Cape, and Baldhead creeks) and North Inlet estuary, SC, (Chainey Creek). Values followed by a = similarity ≥0.75; b = similarity is 0.50–0.75. Data from North Carolina are from the study of Weinstein et al. (1980) and data from South Carolina are from the study of Wenner et al. (1986). Nieuport Nieuport Nieuport BRF 2008 BRF 2009 BRF 2010 2008 2009 2010 Hechtic Creek (S) 18.8 7.2 19.7 18.8 16.9 9.0 Governor’s Creek (S) 31.9 28.1 60.7b 67.9b 18.9 20.0 Walden Creek 1 (S) 5.8 5.8 6.6 5.8 5.4 5.8 Walden Creek 2 (S) 12.1 11.7 13.0 12.1 9.3 12.1 Dutchman Creek (S) 0.4 0.5 1.3 0.3 0.4 0.4 Barnard’s Creek (R) 11.0 6.0 11.9 11.0 11.0 7.8 Town Creek (R) 24.5 6.0 38.4 55.4b 16.1 7.8 Dutchman Creek (R) 0.0 0.0 0.9 0.0 0.0 0.0 Walden Creek (R) 0.0 0.0 0.9 0.0 0.0 0.0 Cape Creek (R) 0.0 0.1 1.8 0.0 0.0 0.0 Baldhead Creek (R) 0.0 0.1 1.8 0.0 0.0 0.0 Chainey Creek (R) 16.8 7.1 18.9 16.2 16.5 9.6