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LIVE-DEAD MISMATCH OF MOLLUSCAN ASSEMBLAGES INDICATES DISTURBANCE FROM ANTHROPOGENIC EUTROPHICATION IN THE BARNEGAT BAY-LITTLE EGG HARBOR ESTUARY.

INTRODUCTION

Barnegat Bay-Little Egg Harbor is a 280 k[m.sup.2] estuary located on the southern coast of New Jersey, separated from the Atlantic Ocean by a barrier island system (Fig. 1). This estuary, hereafter simply Barnegat Bay, is particularly prone to eutrophication as it is shallow (typically <2 m) and poorly flushed (~ 13 days) (Kennish 2001, Defne & Ganju 2014). Recent studies have indicated signs of a stressed ecosystem, including extensive brown tide blooms in the early 2000s and a near 60% decline of submerged aquatic vegetation from 2003 to 2009 (Kennish 2001, Gastrich et al. 2004, Pecchioli et al. 2006, Lathrop & Haag 2011).

Because of these ecological changes and pressure from the public, New Jersey Governor Chris Christie implemented a comprehensive action plan in 2010 that allocated more than 100 million U.S. dollars to address the ecological decline of Barnegat Bay. Under this plan, Taghon et al. (2015) assessed the health of the bay over a 3-y period (2012 to 2014) (1) using the multivariate version of the AZTI Marine Biotic Index, M-AMBI (Muxika et al. 2007). The M-AMBI approach, which integrates the AZTI Marine Biotic Index (AMBI)--a biotic index based on the relative abundance of living taxa in five ecological groups (EG) that indicate the sensitivity/tolerance of a species to organic enrichment--with diversity and richness, is based on community-level sampling of benthic invertebrates and results in marine and coastal systems being ranked in one of five ecological status categories, which include, from best to worst: "High," "Good," "Moderate," "Poor," and "Bad" (Sigovini et al. 2013). Taghon et al. (2015) indicated that Barnegat Bay was largely undisturbed in 2014, with 88% (n = 91) of sites having "High" or "Good" ecological status classifications. Eleven of 91 sites sampled, however, scored high enough to indicate disturbance, being classified as "Moderate" or "Poor," with eight of them located in the northernmost section of the bay, which Taghon et al. (2015) designated as low salinity.

In Barnegat Bay, the largest rivers--the Toms River and the Metedeconk River--enter the estuary in the northern half of the bay (Fig. 1). These two rivers, fed primarily by groundwater (Guo et al. 2004), supply most of the freshwater and keep salinities low in the northernmost section of the bay. As a result, there is a general north-south gradient of increasing salinity in Barnegat Bay (Taghon et al. 2015) (Fig. 1). In addition, because of the presence of Barnegat Inlet and Little Egg Harbor Inlet on the eastern side of the bay, which connect the bay to high-salinity ocean water, there is also a weaker west-east gradient of increasing salinity in Barnegat Bay (Chizmadia et al. 1984) (Fig. 1). Salinity levels have been relatively consistent since the formation of the bay in the Holocene epoch approximately 11,000 y ago (Widmer 1964), with a few exceptions. Oyster Creek Nuclear Generating Station was constructed in the 1960s and the southern branch of the smaller Forked River was redirected as an intake for the station. As a result, the salinity of this river has increased with this redirection of flow (Chizmadia et al. 1984). In addition, the creation of the Point Pleasant Canal in the northern section of the bay in the 1920s, connecting the Manasquan River and Barnegat Bay, similarly raised the salinity of the northern part of the bay. In both of these locations, however, salinity is still low and considered either "low" (14.9-22.6 ppt bottom salinity) or "transitional" (22.6-30.3 ppt bottom salinity) (Taghon et al. 2015).

Although salinity variation in the bay has been relatively consistent, nutrient loads have not. Eutrophication is a recent problem (Kennish & Townsend 2007). Diatom assemblages preserved in sediment cores show evidence of a shift toward high-nutrient conditions starting in the 1940s and 1950s for the northern part of the bay and the 1990s for the southern part of the bay (Velinsky et al. 2011). Roughly half of the nutrients come from surface water inflow, but atmospheric deposition (~39%) and direct ground-water discharge (~ 11 %) also significantly contribute to the nutrient load (Kennish et al. 2007). The northern section of the bay is the most affected by eutrophication because the watershed feeding the estuary in this area is the most densely populated (Kennish 2001, Kennish & Fertig 2012). In this section of the bay, the Toms River and the Metedeconk River deliver the highest concentration of nutrients to the estuary, more than 60% of its nitrogen load, derived primarily from fertilizers (Wieben & Baker 2009).

Taghon et al. (2015) noted that the relative abundance of species that are "very sensitive to organic enrichment" is both inversely correlated with nitrogen concentration and positively correlated with salinity. As a result, the presence of a north-to-south gradient of increasing salinity and decreasing nutrients in Barnegat Bay complicates the interpretation of the M-AMBI scores that Taghon et al. (2015) found in their study. M-AMBI scores indicating disturbance could have resulted from a high level of organic enrichment or simply because the salinity was low. Indeed, this type of bias is widely acknowledged to be a problem when reference conditions have not been defined in naturally stressful low-salinity habitats (Elliott & Quintino 2007, Dauvin & Ruellet 2009, Gillett et al. 2015).

Taghon et al. (2015) recognized salinity as a confounding factor and used a stepwise linear regression with backward selection to assess its effect on the composition of benthic communities. Using this approach, they identified dissolved oxygen saturation and total nitrogen, which explained more than 90% of the variation in their dataset, as drivers in the proportion of "very sensitive" taxa and eliminated salinity, temperature, dissolved oxygen concentration, total phosphorous, and chlorophyll a as important driving factors. Although this approach is commonly used in ecology, its inherent biases and drawbacks are well documented and include an inappropriate dependence on a single best model and biases associated with parameter entry and order (Whittingham et al. 2006). These biases are especially present when predictors are correlated (Whittingham et al. 2006), which is the case for salinity and total nitrogen in Barnegat Bay. Because of these potential biases it may be premature to discount salinity as an important driving factor of benthic community composition. Here, geohistorical information derived from molluscan death assemblages is used to disentangle the effects of salinity and eutrophication on benthic community composition in Barnegat Bay to test the robustness of the conclusions drawn in the Taghon et al. (2015) study.

Death Assemblages and Predictions

Death assemblages are the "dead remains sieved from the upper mixed-zone of the sedimentary column" (Kidwell 2007; p. 17, 701). Molluscan death assemblages can be used as an ecological baseline because they are time averaged; that is, death assemblages are continuously adding and losing shells, while retaining shells from the past several decades and centuries. Time-averaging has the advantage of diminishing short-term ecological variation, producing average conditions of the benthic community (and by extension environmental conditions) (Kowalewski et al. 1998). In lagoonal systems such as Barnegat Bay, maximum shell ages can be up to thousands of years old but most shells are from the most recent decades and centuries (Kidwell 2013). In the absence of significant environmental change and taphonomic biases (e.g., out-of-habitat transport and differential preservation), the molluscan death assemblage (DA) and living assemblage (LA) will generally match in terms of species composition and abundance (Kidwell 2007). When the two assemblages diverge (or are "mismatched"), anthropogenic impact, particularly from eutrophication, has likely occurred (Kidwell 2007). This live-dead mismatch approach has been used in bays and estuaries around the world to assess the degree of anthropogenic impact (e.g., Albano & Sabelli 2011, Casey et al. 2014, Negri et al. 2015, Leshno et al. 2016, Dietl & Smith 2017, Tomasovych & Kidwell 2017). Although death assemblages are composed primarily of the hard parts of molluscs, they have been found to be good surrogates for the whole living benthic community. In a meta-analysis of 12 benthic community datasets from European waters, nearly 80% of stations had matching mollusc-only and whole-community AMBI ecological status classifications, illustrating the usefulness and viability of mollusc-only studies (Dietlet al. 2016).

If excess nutrient load is the main factor driving the molluscan community composition and disturbed M-AMBI scores in the Taghon et al. (2015) study, then the most pronounced differences between the LA and DA are likely to occur in the northernmost section of the bay where nutrient loads are the greatest (Prediction 1). If, however, salinity is the main factor in shaping the molluscan community composition, then all sites should have similar living and death assemblages as the salinity gradient typical of the bay today predates influences of modern industrial society (Prediction 2).

MATERIALS AND METHODS

Sampling and Data Collection

Nineteen samples of both molluscan living and death assemblages were collected May-October 2015 and January 2017 at four locations along a north-south gradient of decreasing eutrophication and increasing salinity: four from Mantoloking, three from Forked River, eight from Island Beach, and four from Little Egg Harbor (Fig. 1). These four sampling sites were adjacent to sites sampled by Taghon et al. (2015) and fall into three salinity zones designated by Taghon et al. (2015) as "low" (14.9-22.6 ppt bottom salinity), "transitional" (22.6-30.3 ppt bottom salinity), or "high" (26.5-30.3 ppt bottom salinity). The northernmost site, Mantoloking, was located in the low-salinity zone and was adjacent to site BB14-013 in the Taghon et al. (2015) study, which was classified as "Moderate" on the M-AMBI scale (Table 1). The second site along the salinity gradient was Forked River, located just north of the mouth of the Forked River. This site was located in the low-salinity zone of the bay, adjacent to a Taghon et al. (2015) site (BB 14-023) which was classified as "Moderate" on the M-AMBI scale (Table 1). The third sampling site, Island Beach, was located just west of the Island Beach State Park. A nearby Taghon et al. (2015) site (BB14-050) was classified as "Good" on the M-AMBI scale (Table 1). Although the Forked River and Island Beach sites were close to each other in terms of position along the north-south gradient, the proximity of the Island Beach site to the Barnegat Inlet, which is connected to the open ocean, results in a much higher average salinity. The southernmost site, Little Egg Harbor--located at the very southern tip of Barnegat Bay--was in a high-salinity zone adjacent to one of the Taghon et al. sites (BB14-083) that was classified as "Good" on the M-AMBI scale (Table 1).

Samples were taken using an Ekman grab (15.25 X 15.25 X 15.25 cm in size at the Forked River site only) or from 1 [m.sup.2] by 30-cm deep plots that were excavated by hand with a shovel from soft sandy bottom locations, some of which were adjacent to sparse to dense eelgrass beds (e.g., Island Beach) (Fig. 1; see also Fig. 18 in Bricelj et al. 2012). Samples taken by Ekman grab were all from locations where water depth prevented hand excavation. Samples were washed through a 5-mm sieve on site. This sieve size is larger than typical benthic community studies but was chosen to minimize potential taphonomic biases in the DA that could influence results. For instance, the potential bias of out-of-habitat transport of shells is mitigated by the use of a larger sieve size because individuals less than 1 mm in size are more prone to being transported between habitats (Kidwell 2001). By looking at larger and typically adult individuals, this sampling approach also minimized the possibility of differential preservation bias against smaller, possibly more fragile shells and differential turnover between taxa because adult communities are less sensitive to seasonal variation (Kidwell 2007).

Live and dead gastropod and bivalve specimens were sorted out from assemblages, cleaned, and identified to species level. Only right valves with intact hinges were used to count individual bivalves in the death assemblages. Only gastropod shells in the death assemblages with the spire present were counted.

Data Analysis

To assess the role of salinity in influencing the composition of molluscan communities, a two-pronged live-dead approach was used. For the first approach, Jaccard-Chao index of taxonomic similarity was plotted against rank-order abundance (indexed by Spearman's rho). The Jaccard-Chao index assesses taxonomic similarity between the LA and DA, with values ranging from 0 to 1, with highly similar assemblages scoring closer to 1 (Chao et al. 2005). Spearman's rho estimates the rank-order agreement in species abundance between the living and death assemblages and ranges from -1 to 1, with inversely ranked assemblages scoring -1 and identically ranked assemblages scoring 1. In a plot of Jaccard-Chao and Spearman's rho, sites that plot in the upper right-hand quadrant have high similarity between the living and death assemblages. Sites with high levels of anthropogenic eutrophication will generally plot away from this upper right-hand quadrant. According to Kidwell (2007), Jaccard-Chao values under 0.7 and Spearman's rho values under 0.2 tend to indicate ecological change related to anthropogenic eutrophication. All mollusc species appearing live or dead were included in this analysis.

The second approach consisted of applying the M-AMBI index (Muxika et al. 2007) to the LA and DA data. M-AMBI values were calculated (2) using the AMBI 5.0 software (and the November 2014 species list), with the "bad" reference condition set as the theoretical minimum (or maximum in the case of AMBI) for each metric that is integrated into the M-AMBI index [i.e., AMBI = 6, richness (S) = 0, and Shannon diversity (H') = 0], and the "high" reference condition corresponding to the highest (or lowest in case of AMBI) value within the dataset for each of the metrics (AMBI = 0.9, H' = 2.58, S = 17, which corresponded to the DA from the Island Beach site) (Sigovini et al. 2013). M-AMBI values range from 0 to 1 and correspond to five ecological status classifications: "Poor" (0-0.2), "Bad" (0.2-0.39), "Moderate" (0.39-0.53), "Good" (0.53-0.77), and "High" (0.77-1). Ecosystems that score above the "Good" to "Moderate" boundary are considered undisturbed, and those that score below it are considered disturbed.

RESULTS

A total of 146 live individuals and 625 dead individuals were collected (Table 2). Eight species (six bivalves and two gastropods) were identified in the living assemblages and 19 species (15 bivalves and four gastropods) were identified in the death assemblages. The living assemblages were composed mostly of mud snails [Ilyanassa obsoleta (Say, 1822)], hard clams [Mercenaria mercenaria (Linnaeus, 1758)], and dwarf surf clams [Mulinia lateralis (Say, 1822)]. The death assemblages were primarily composed of razor clams [Ensis directus (Conrad, 1844)], mud snails (I. obsoleta), and soft-shelled clams [Mya arenaria (Linnaeus, 1758)] (Table 2). The soft-shell clam M. arenaria was only found in the death assemblages.

Although none of the four sites had both Jaccard-Chao taxonomic similarity and Spearman's rho values (under 0.70 and 0.20, respectively) that indicated a significant compositional change in the benthic community (Table 1; Fig. 2), there was a spatial pattern in the results that is consistent with human impact (e.g., eutrophication) in the north but not the south. The northernmost sites, Mantoloking and Forked River, both had Jaccard-Chao values of 0.59. Spearman's rho values for these sites were 0.29 and 0.42, respectively. The two southernmost sites, with higher salinities, however, had Jaccard-Chao values (0.79 and 0.85 for Island Beach and Little Egg Harbor, respectively) that were on average 1.4 times higher, and Spearman's rho values (0.70 and 0.63 for Island Beach and Little Egg Harbor, respectively) that were on average 1.9 times higher than the two northern sites.

The M-AMBI analysis also showed a pattern that was consistent with greater ecological impact of high nutrient loads in the northern part of the bay on the benthic community. The ecological status classification of the LA for the two northernmost sites was "Moderate" (M-AMBI = 0.4 and 0.41 for Mantoloking and Forked River, respectively; Table 1; Fig. 3), but the DA M-AMBI values for these two sites (0.69 and 0.62 for Mantoloking and Forked River, respectively) corresponded with a "Good" ecological status classification, indicating an undisturbed past and recent shift to a disturbed benthic community. There was no evidence of a disturbed benthic community in either of the two southern sites. The Island Beach site had a LA M-AMBI value (0.68) that corresponded with a "Good" ecological status classification and a DA M-AMBI value (1.0) that corresponded with a "High" ecological status classification. Similarly, the southernmost site along the north-south gradient, Little Egg Harbor, had both a LA M-AMBI value (0.6) and a DA M-AMBI value (0.7) that corresponded with a "Good" ecological status classification (Table 1; Fig. 3).

DISCUSSION

Although the two northernmost sites did not have Jaccard-Chao and Spearman's rho values lower than the designated cutoffs for confidently declaring anthropogenic eutrophication, the Jaccard-Chao and Spearman's rho values for the two northernmost sites sampled in Barnegat Bay were notably lower than the southernmost sites (Fig. 2). In addition, the M-AMBI analysis showed that there were also shifts in ecological status classifications between the living and death assemblages in the north (from undisturbed to disturbed) but not the south. Both the lower agreement in the north in terms of taxonomic similarity and rank-order abundance between the living and death assemblages as well as shifting ecological statuses (from undisturbed to disturbed in the death to living assemblages, respectively) in the north support the Taghon et al. (2015) hypothesis that anthropogenic eutrophication (Prediction 1), and not salinity, has influenced the molluscan assemblage in the northernmost section of the bay causing disturbed M-AMBI scores. Positive correlations between the Jaccard-Chao, Spearman's rho, and M-AMBI values derived from the present live-dead study and the corresponding M-AMBI values from Taghon et al. (2015) (r = 0.69, 0.53 and 0.82, respectively; Fig. 4) further strengthens confidence in the conclusions drawn from the independent M-AMBI evidence presented by Taghon et al. (2015).

Specifically, the northernmost site, Mantoloking, had low agreement in taxonomic similarity and rank-order abundance between the LA and DA, with the lowest Jaccard-Chao and the second lowest Spearman's rho values (0.59 and 0.42, respectively; Table 1). The M-AMBI analysis also classified the ecological status of the LA at the Mantoloking site as being more degraded than the DA ("Moderate" versus "Good" classification, respectively; Table 1; Fig. 3). These results correspond well with the disturbed M-AMBI classification of "Moderate" found at site BB14-013 in the Taghon et al. (2015) study (Table 1). At the Forked River site, the second northern site, low agreement in taxonomic similarity and rank-order abundance was found between the LA and DA, with the second lowest Jaccard-Chao and lowest Spearman's rho values (0.59 and 0.29, respectively). M-AMBI analysis again also showed evidence of recent disturbance, with the ecological status of the LA being classified as "Moderate" and the DA as "Good" (Table 1; Fig. 3). These results correspond well with the disturbed "Moderate" M-AMBI classification of the adjacent Taghon et al. (2015) site BB14-023 (Table 1). Island Beach, the third site along the salinity gradient, had very high agreement in taxonomic similarity and rank-order abundance between the LA and DA, with the second highest Jaccard-Chao value and the highest Spearman's rho value (0.79 and 0.68, respectively). This site was also classified as undisturbed by the M-AMBI analysis (ecological status = "Good" and "High" for the LA and DA, respectively), which matches the "Good" ecological status classification that the Taghon et al. (2015) M-AMBI analysis found at site BB14-050 (Table 1; Figs. 2 and 3). Finally, there was also good agreement between this study and the results that Taghon et al. (2015) found at the southernmost site, Little Egg Harbor. This site had the highest Jaccard-Chao and second highest Spearman's rho values (0.85 and 0.63, respectively), and displayed no shift between the LA and DA M-AMBI ecological status classification, both being classified as "Good," which agrees well with the "Good" ecological status that Taghon et al. (2015) found at site BB14-083 (Table 1; Figs. 2 and 3).

Species-Specific Agreement and Disagreement Between JC-rho and M-AMBI Analyses

The two live-dead approaches used in this study are fundamentally different. M-AMBI integrates an ataxic index (AMBI) based on the relative abundance of different ecological groups, indicating the degree of sensitivity/tolerance to organic enrichment, in a benthic sample, with Shannon diversity and species richness. A live-dead comparison using the AMBI index could result in very similar ecological status classifications even if the species composition of the LA and DA was different. The Jaccard-Chao taxonomic similarity-Spearman's rank-order abundance (JC-rho, hereafter) approach is dependent on the identity of the species in the live-dead assemblages being compared. Despite this fundamental difference, there was good agreement between the two approaches.

A closer examination of the species present in the live-dead samples indicates that many of the low- and high-ecological group (EG) species (i.e., those most and least sensitive to disturbance, respectively) responded predictably under eutrophic conditions. Of 19 species found in this study, 10 were represented by greater than 10 individuals in either the living or the death assemblages. Of these 10 species, six had EG assignments that corresponded well with their rank-order abundance in the living and death assemblages. For example, two EG I bivalve species, Ensis directus and Spisula solidissima (Dillwyn, 1822) (Table 2), which are predicted to be the most sensitive to disturbance (Grall & Glemarec 1997), each dropped by three or more ranks in the LA compared with the DA (Table 2). In addition, the rank abundance of the bivalve Tellina agilis (Stimpson, 1857), an EG II species (Table 2), remained relatively consistent between the LA and DA (Table 2). EG II species are defined as species indifferent to nutrient enrichment (Grall & Glemarec 1997), so the observed consistency in rank abundance corresponds well with its EG assignment.

Although only represented by a few individuals in the live-dead samples, a species of particular importance in Barnegat Bay--the bay scallop Argopecten irradians (Lamarck, 1819), another EG I species (Table 2) that is very sensitive to organic enrichment--also displays a shift that corresponds well with its EG assignment. Only three dead individuals of A. irradians were found at one site, Island Beach (Table 2). This site is located within the Sedge Island Marine Conservation Zone (MCZ), a 4.5+-k[m.sup.2] area, where commercial harvesting of shellfish and personal watercraft use are prohibited. The Sedge Island MCZ is composed of more than 1 k[m.sup.2] of submerged aquatic vegetation (Bricelj et al. 2012), which both protects scallops from being washed ashore during storms and acts as critical habitat for juvenile bay scallops (MacKenzie 2008). The lack of A. irradians in the LA samples, particularly at the Island Beach site, echoes the recent decline of bay scallops in Barnegat Bay. This decline is tied to the loss of eelgrass, which experienced a nearly 60% decline from 2003 to 2009 (Bologna et al. 2000).

Even though the response of six of the 10 species with greater than 10 individuals aligned well with their EG assignments, four species did not (Table 2). One of these species--the bivalve Mya arenaria--is an EG II species (Table 2), which is predicted to be indifferent to disturbance resulting from nutrient enrichment (Grall & Glemarec 1997). This species was the third most commonly found dead at three of the four sampling sites (Table 2), with nearly 80% of individual specimens (>25% of DA) found at the northernmost site, Mantoloking; however, it was not found alive at any sampling site. Absence of this EG II species in the LA may reflect the recent increase in nutrients and macroalgae in Barnegat Bay, particularly in the north; studies have shown that macroalgal mats can prevent up to 100% of M. arenaria larval settlement (Olafsson 1988). Indeed, sampling of macroalgae at 120 sites throughout Barnegat Bay from 2004 to 2010 found 55 occurrences of early bloom (70%-80% cover) and full bloom (>80% macroalgal cover) incidents, with Ulva lactuca Linnaeus, 1753 and Gracilaria tikvahiae (McLachlan, 1979) being the two most abundant algal species found (Kennish et al. 2011). In addition to preventing larval settlement, sheet-forming algae, such as U. lactuca, are known for their capacity to cause suffocation from low dissolved oxygen concentrations, reduce burial depth, and inhibit growth of M. arenaria (Auffrey et al. 2004). Although no living M. arenaria individuals were found in this study, this does not imply that the species no longer lives in Barnegat Bay. For instance, Taghon et al. (2017; table 8, p. 20) documented the presence of M. arenaria in their 2012 to 2014 survey of the macroinvertebrate benthic fauna of the bay, and the species is well documented within Barnegat Inlet (Fig. 1), most likely because of larval transport and settlement from nearshore ocean currents entering through the inlet (M. Kennish, personal communication, Rutgers University, December 12, 2017). Therefore, the pattern of high abundance in the death assemblages and absence in the living assemblages (Table 2) observed in this study may indicate that the negative indirect effects of macroalgal mat growth on M. arenaria may far outweigh this species' capacity to tolerate high-nutrient conditions as predicted by its EG II assignment (raising some concern about EG reliability; see also Leonardsson et al. 2015).

Another notable species that does not shift in abundance as expected is the commercially valuable hard clam, Mercenaria mercenaria (Table 2). The hard clam M. mercenaria is an EG II species (Table 2), so, as with Mya arenaria, this species' response is predicted to be indifferent to nutrient enrichment. In the southern section of Barnegat Bay, where M. mercenaria is most often found owing to its intolerance of low-salinity conditions (<20 ppt) in the north (Malouf & Bricelj 1989; Bricelj et al. 2012), a well-documented 68% decline in the hard clam population has occurred between 1980 and 2001 (Bricelj et al. 2012). The main stressor implicated in this recent decline is summer blooms of harmful algae, which are more prevalent and intense in the southern half of the bay (see review in Bricelj et al. 2012). Blooms of harmful microalgal species, such as Aureococcus anophagefferens (Hargraves & Sieburth, 1988), can inhibit the growth of larvae and juveniles (Greenfield & Lonsdale 2002), given their relatively small (about 2 urn in diameter) cell size (Grizzle et al. 2001), which hard clams cannot filter (Tracey et al. 1988), as well as inhibit reproduction of adult hard clams (Bricelj et al. 2012).

By contrast to data linking harmful algal blooms associated with eutrophication to a declining Mercenaria mercenaria population in Barnegat Bay, the rankings of this species (second in the DA versus fourth in the LA; Table 2) in this live-dead study suggests a recovering population, and perhaps an improving ecological status of the bay. Interpreting the live-dead data in this way, however, may be inappropriate. Ongoing hard clam restoration efforts within Barnegat Bay likely explain this counterintuitive finding. In the live-dead samples, over 80% of M. mercenaria individuals were found at one site: Island Beach, which is within the Sedge Island MCZ (Table 1). Between 2010 and 2011, over 60,000 [m.sup.2] of the MCZ were seeded with nearly four million hard clams in efforts to bolster the M. mercenaria population (Bricelj et al. 2012). It is quite possible that any live-dead sampling in this zone would be biased in terms of living hard clams.

Although some of the species present in the live-dead samples do not match perfectly with their assigned EG assignments, the overall agreement between the JC-rho and M-AMBI analyses is reassuring given the fundamental differences between the two approaches and the documented bias in the performance of the latter in low-salinity environments; that is, as mentioned earlier, its potential to score sites with no observed eutrophication problems as having degraded benthic communities (Elliott & Quintino 2007, Dauvin & Ruellet 2009, Gillett et al. 2015). For example, the presence of the bivalve Mulinia lateralis--a small-sized, short-lived opportunistic EG IV species (Table 2) known to experience rapid population growth following environmental disturbances (Cerrato et al. 2013)--in the northern half of Barnegat Bay is consistent with the influence of recent anthropogenic eutrophication. This species was more abundant in the LA than in the DA at the northernmost site, Mantoloking, and the most common species in the LA and in the DA at the second northernmost site, Forked River (Table 2). The dwarf surf clam, M. lateralis, however, is also known for its capacity to reproduce successfully under low-salinity conditions. For instance, in a series of experiments, Calabrese (1969) established that embryos of M. lateralis developed normally (>70% success) between 22.5 and 30 ppt and some embryos even developed in salinities as low as 15 ppt. Because M. lateralis was only found at the low-salinity sites (save one live individual at Island Beach), its presence cannot be conclusively interpreted (despite its EG IV assignment) as a signal of a highly eutrophic system. Indeed, dense populations of M. lateralis have long been documented in Barnegat Bay since as early as the 1960s, particularly in the low-salinity Oyster Creek area (Loveland & Vouglitois 1984). This potential bias in a M-AMBI analysis reflects the difficulties associated with defining a baseline in naturally stressful habitats, such as low-salinity estuaries, which the live-dead geohistorical approach used in this study overcomes (see also Dietl et al. 2016), further strengthening confidence in the conclusion that anthropogenic eutrophication (Prediction 1), and not salinity, is a primary factor in driving disturbed M-AMBI scores in the northern half of Barnegat Bay.

Next Steps/Applications

One of the 10 points of Governor Chris Christie's comprehensive action plan to address and improve the ecological quality of Barnegat Bay was to fill in gaps in research. (3) Benthic community monitoring in Barnegat Bay is a prime example. For instance, before the Taghon et al. (2015) study, only a handful of surveys had been performed in limited areas of the bay over the last 60 y. Phillips (1972) and Loveland et al. (1972, 1974) conducted benthic invertebrate surveys in the western section of the bay surrounding Oyster Creek between 1965 and 1973 in an effort to monitor if and how the community might change with installation of the Oyster Creek Nuclear Generating Station (see Loveland & Vouglitois, 1984, for more details). Although such sampling was extensive and occurred over multiple years, sampling was only performed in the western part of Barnegat Bay. Moser (1997) sampled benthic invertebrates at one location in Barnegat Bay and one location in Little Egg Harbor. More recently, MacKenzie (2003) surveyed densities of benthic invertebrates at three sites in Barnegat Bay in a comparative study to evaluate the role of overharvesting in explaining declines in hard clam populations. Ninety-six sites throughout Barnegat Bay were sampled in 2001 as part of the U.S. Environmental Protection Agency Regional Environmental Monitoring and Assessment Program and National Coastal Assessment efforts to develop benthic indicators to assess the current condition of, and forecast future risks for, ecological resources. (4) The Taghon et al. (2015) study, which was designed to capture a comparable sampling density as the Environmental Protection Agency Regional Environmental Monitoring and Assessment Program and National Coastal Assessment efforts in 2001, is the most recent and most comprehensive, with 100 sampling locations spread throughout Barnegat Bay and sampling occurring over a 3-y period. Although this study elaborates on the current status of Barnegat Bay, a lack of an ecological baseline limits understanding of how Barnegat Bay has changed. One-time geohistorical sampling can provide this baseline. Combining this ecological baseline with ongoing monitoring can guide ecological assessment and restoration efforts and provide key information on the performance of restoration efforts (Dietl & Smith 2017). In October 2017, an additional $20 million in funding was allocated "to improve water quality, protect natural resources and restore ecological balance" in Barnegat Bay, which includes continued benthic invertebrate monitoring by the New Jersey Department of Environmental Protection (NJDEP 2017). Without much additional effort, geohistorical data can be integrated into this effort, simply by analyzing the dead component of each benthic sample that is collected, to assess how human activities have impacted Barnegat Bay and measure the success of restoration efforts.

CONCLUSIONS AND RECOMMENDATIONS

This study corroborates the hypothesis that anthropogenic eutrophication, and not salinity, is driving the disturbed M-AMBI scores observed by Taghon et al. (2015) in the northernmost section of the Barnegat Bay estuary. In addition, there are key species-specific differences between the living and death assemblages that may also suggest disturbance in the southern part of the bay. Additional live-dead sampling in this portion of the bay is recommended to permit a better understanding of the effects of recent eutrophication in the south, and comprehensive sampling across the entire bay to establish ecological baselines of what benthic communities in Barnegat Bay looked like before anthropogenic impact. Adding this geohistorical approach to ongoing benthic community monitoring efforts in Barnegat Bay provides a unique opportunity to fill research gaps and plan, guide and evaluate the success of restoration efforts.

ACKNOWLEDGMENTS

We thank John Wnek for his generous help both in the field and with acquiring equipment, Gary Taghon for providing data, and Michael Kennish for discussing the ecological status of mollusc populations in Barnegat Bay. Additional thanks to Morgan Barney and Lorraine Tweitmann for their assistance in the field, Carrie DePasquale for her help with data analysis and an anonymous reviewer for their helpful comments on the manuscript. This work was partially supported by funding to AT from Save Barnegat Bay.

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ANNALEE TWEITMANN (1*) AND GREGORY P. DIETL (2,3)

(1) Department of Natural Resources, Cornell University, 226 Mann Drive, Ithaca, NY 14853; (2) Paleontological Research Institution, 1259 Trumansburg Road, Ithaca, NY 14850; (3) Department of Earth and Atmospheric Sciences, Cornell University, 112 Hollister Drive, Ithaca, NY 14853

(*) Corresponding author. E-mail: amt223@cornell.edu

(1) Additional field sampling conducted in 2016 showed that benthic community structure was similar to that in 2012 to 2014; see Taghon et al. (2017), for more details.

(2) No correction (sensu Dietl et al. 2016) was applied to the mollusc-only AMBI values used in the M-AMBI calculation; regression equations to estimate whole-community AMBI values from mollusc-only values are currently available for European molluscan taxa only.

(3) http://www.state.nj.us/dep/barnegatbay/bbfh.htm.

(4) https://archive.epa.gov/emap/archive-emap/web/html/index-171.html.

DOI: 10.2983/035.037.0314
TABLE 1. Comparison of live-dead sampling sites in this study with
adjacent sampling sites from Taghon et al. (2015).


                            Our sampling sites
                            Mantoloking

Coordinates                 40.05[degrees] N 74.05[degrees] W
Jaccard-Chao                 0.59
Spearman's rho               0.42
LA ([dagger]) M-AMBI Score   0.4
DA ([double dagger])
M-AMBI Score                 0.69
LA Ecological Status        Moderate
DA Ecological Status        Good
                            Adjacent Taghon et al. (2015)
                            sampling sites
                            BB14-013
Coordinates                 40.06[degrees] N 74.08[degrees] W
M-AMBI Score                 0.5
M-AMBI Ecological Status    Moderate
Salinity                    Low

                            Our sampling sites
                            Forked River

Coordinates                 39.83[degrees] N 74.15[degrees]W
Jaccard-Chao                 0.59
Spearman's rho               0.29
LA ([dagger]) M-AMBI Score   0.41
DA ([double dagger])
M-AMBI Score                 0.62
LA Ecological Status        Moderate
DA Ecological Status        Good
                            Adjacent Taghon et al. (2015)
                            sampling sites
                            BB14-023
Coordinates                 39.84[degrees] N 74.13[degrees] W
M-AMBI Score                 0.53
M-AMBI Ecological Status    Moderate
Salinity                    Transitional

                            Our sampling sites
                            Island Beach

Coordinates                 39.78[degrees] N 74.10[degrees] W
Jaccard-Chao                 0.79
Spearman's rho               0.68
LA ([dagger]) M-AMBI Score   0.68
DA ([double dagger])
M-AMBI Score                 1
LA Ecological Status        Good
DA Ecological Status        High
                            Adjacent Taghon et al. (2015)
                            sampling sites
                            BB 14-050
Coordinates                 39.79[degrees] N 74.13[degrees] W
M-AMBI Score                 0.56
M-AMBI Ecological Status    Good
Salinity                    High

                            Our sampling sites
                            Little Egg Harbor

Coordinates                 39.56[degrees] N 74.34[degrees] W
Jaccard-Chao                 0.85
Spearman's rho               0.63
LA ([dagger]) M-AMBI Score   0.6
DA ([double dagger])
M-AMBI Score                 0.7
LA Ecological Status        Good
DA Ecological Status        Good
                            Adjacent Taghon et al. (2015)
                            sampling sites
                            BB 14-083
Coordinates                 39.57[degrees] N 74.32[degrees] W
M-AMBI Score                 0.55
M-AMBI Ecological Status    Good
Salinity                    High

([dagger]) LA, living assemblage; ([double dagger]) DA, death
assemblage.

TABLE 2. Mollusc species and number of individuals found at each
live--dead sampling site. Numbers in parentheses indicate abundance
ranking. EG indicates ecological group assignment. (*)


                                                    Mantoloking
                             EG                     Live  Dead

Bivalvia
Anadara transversa           IV                      0      0
Anomia simplex               I                       0      0
Argopecten irradians         I ([dagger])            0      0
Ensis direclus               I                       1     10
Gemma gemma                  II                      0      0
Geukensia demissa            III                     0     16
Laevicardium mortoni         I                       0      0
Mercenaria mereenaria        II                      0      0
Mulinia lateralis            IV                      5      6
Mya arenaria                 II                      0     56
Mytilus edulis               III                     0      0
Solemya velum                I                       0      0
Spisula solidissima          I                       0      0
Tagelus divisus              III                     4     12
Tellina agilis               II                      0      0
Gastropoda
Crepidula convexa            II                      0      2
Crepidula plana              III                     0      2
Ilyanassa obsoleta           III                    42     82
Urasalpinx cinerea           N/A ([double dagger])   0      1
Total Number of Individuals                         56    196
Total Number of Species                              4      9

                                                         Little Egg
                             Forked River  Island Beach  Harbor
                             Live  Dead    Live  Dead    Live  Dead

Bivalvia
Anadara transversa            0     0       0      3      0     0
Anomia simplex                0     0       0      5      0     0
Argopecten irradians          0     0       0      3      0     0
Ensis direclus                0     0       0    152      1     2
Gemma gemma                   0     0       0      1      0     0
Geukensia demissa             0     2       0      2      0     1
Laevicardium mortoni          0     0       0      1      0     0
Mercenaria mereenaria         0     1      17     35      5     5
Mulinia lateralis             9     9       0      1      0     0
Mya arenaria                  0     0       0     14      0     1
Mytilus edulis                0     0       0      2      0     0
Solemya velum                 0     0       7      6      4     0
Spisula solidissima           0     0       0     30      0     0
Tagelus divisus               0     4       0      1      1    10
Tellina agilis                7     1       2     16      0     0
Gastropoda
Crepidula convexa             0     2       0      0      0     3
Crepidula plana               0     0       0      0      0     0
Ilyanassa obsoleta            1     3       1     63     18    18
Urasalpinx cinerea            0     0       0      3      1     4
Total Number of Individuals  20    29      40    355     30    45
Total Number of Species       3     7       5     17      6     9


                             All Sites
                             Live    Dead

Bivalvia
Anadara transversa             0       3(14)
Anomia simplex                 0       5(13)
Argopecten irradians           0       3(15)
Ensis direclus                10(5)  164(2)
Gemma gemma                    0       1(18)
Geukensia demissa              0      21(7)
Laevicardium mortoni           0       1(19)
Mercenaria mereenaria         22(2)   41(4)
Mulinia lateralis             14(3)   16(9)
Mya arenaria                   0      71(3)
Mytilus edulis                 0       2(17)
Solemya velum                 11(4)    6(12)
Spisula solidissima            0      30(5)
Tagelus divisus                5(7)   27(6)
Tellina agilis                 9(6)   17(8)
Gastropoda
Crepidula convexa              0       7(11)
Crepidula plana                0       2(16)
Ilyanassa obsoleta            62(1)  166(1)
Urasalpinx cinerea             1(8)    8(10)
Total Number of Individuals  146     625
Total Number of Species        8      19

(*) EG assignments were obtained using the AMBI 5.0 software (and the
November 2014 species list). Following Grail and Glemarec (1997), the
most sensitive of the ecological groups is EG I; these species are
described as very sensitive to organic enrichment and present in normal
conditions. EG II species are indifferent to organic enrichment and are
always present in low densities. EG III describes species that are
tolerant of organic enrichment. EG IV is composed of second-order
opportunistic species, which are small species with a short life cycle
that are adapted to a life in reduced sediments. EG V species are
first-order opportunists (e.g., deposit feeders that proliferate in
reduced sediments).
([dagger]) This species was not listed in the AMBI November 2014
species list; the EG for Argopecten ventricosis (G. B. Sowerby II,
1842) was used as a substitute.
([double dagger]) There was no EG listed for this species and no
genus-level substitutions in the AMBI species list version of November
2014.
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