annual variations. This outcome strongly suggests that habitat associations have been stable between 1986-87 and 1994-95.


The benthic fauna present in the Lower Bay Complex are typical of estuarine and coastal regions of the Northeast and Mid-Atlantic. They live in a highly fluctuating physical environment and as a result, tend to show dynamic temporal patterns. There was, however, stability or coherence in the faunal associations apparent over the four decades spanned by the regional data sets. The best example is the comparisonbetween 1973 and 1987. A strong faunal association was found in winter between 1973 and 1987, even though the faunain 1973 declined to low abundance and there was marked evidence for habitat differences.

Benthic community structure in the Lower Bay Complex tends to follow in a broad way the sediment and hydrographic regime that divides the area into distinct southern (Raritan and Sandy Hook Bays) and northern regions (Lower Bay). Within this broadscale pattern, habitats have changed over time. Mantel tests indicated that the habitat structure was uncorre-lated when comparing the surveys of 1959-60 and 1986 and when comparing 1973 and 1987. From 1983 onward, spatial associations remained stable. It was apparent, from an examination of Tables 18.3 and 18.4 and the plots in Figure 18.3, that many members of the benthic community were more abundant during the 1980s and 1990s than in 1957-60 and 1973. This trend could not be explained by differences in survey methods.

Further detailed analysis of benthic community structure was hampered by a number of problems: 1) high annual variability, 2) differences in sampling methods among regional studies, 3) goodness-of-fit problems in the multivariate analyses, and 4) weak faunal-environmental relationships. High annual variability is a characteristic of temperate, coastal benthic communities, but little quantitative information exists on the degree of annual change, mainly because few regional studies are carried out at annual or greater time scales (Constable, 1999). In the present study, correlations declined abruptly between seasonal and annual time scales. For example, in examining spatial associations and using species as descriptors, seasonal correlations ranged from 0.65-0.75 while annual correlations were only 0.19-0.41 (Table 18.8). In contrast, McArdle and Blackwell (1989) found a gradual decline in correlation from 0.6 to 0 for Chione stutch-buryi as sampling intervals increased from three to eighteen months. McArdle and Blackwell were, however, examining a single, long-lived species over a small spatial scale (100 m), so it would be expected to behave more predictably than a multispecies assemblage. The cause of high annual variability in the present study was probably episodic recruitment and mortality of several of the dominant species. During 1957-60 (Table 18.3), there were large annual changes in abundance of Polydora ligni, Mya arenaria, Mytilis edulus, Gemma gemma, Ampelisca sp., Unciola serrata, and other dominants. High annual variability obscures relationships with stable environmental factors such as sediment grain-size and hampers attempts to identify structure or pattern across data sets. A practical concern that must be considered in the future is the relevance of single surveys in characterizing community structure or establishing baseline levels.

The different sampling methods used in the re-gionalstudies (Table 18.1) limited the types of comparisons that could be made. Simple comparisons of abundance, relative abundance of each species, species richness, or evenspecies presence/absence among the regional studies would have been very valuable but were avoided for the most part because no reliable corrections for methodological differences exist. To assess long term changes, future regional studies will need to seriously consider matching sampling locations, season of collection, sampling device, and sieve size to prior studies. Establishing type collections and a specimen repository would also be worthwhile, since they would reduce variability in taxonomy among studies.

The ordination and cluster analysis techniques used in this study did not perform well. Goodness-of-fit problems arose in all attempts, indicating that a simple two- or three-dimensional display for ordination or a tree-diagram for cluster analysis would not accurately represent relationships. It was not possible, therefore, to confidently use these methods to identify spatial and faunal associations for the regional studies, let alone use them to assess change in associations between studies. The goodness-of-fit criteria used in this study are an essential part of the process of validation of a multivariate method, that is, determining whether the structure or relationships being displayed are real or an artifact created by the analytical method (Legendre and Legendre, 1998). Although validation is important, validation methods are generally not available in statistical packages (Legendre and Legendre, 1998) and are not commonly examined as a regular step in an analysis.

It is possible that an index other than the Bray-Curtis measure, or a different choice of ordination and clustering algorithm, would have performed better than those selected for the present study. More likely, a different approach is required to handle these large, complex data sets. A rapidly expanding class of multivariate methods, generally called "direct comparison" by Legendre and Legendre (1998) and "direct analysis" by ter Braak (1996), may provide a partial solution. A direct analysis examines faunal data and environmental data together within a single multivariate technique. Canonical correspondence analysis (CCA) is one example (ter Braak 1996). CCA would produce an ordination of stations and/or species arranged along gradients in grain-size, temperature, salinity, and other environmental data. This approach has not been applied to data from the Lower Bay Complex.

Simply applying direct analyses to existing data, however, would not totally resolve the analytical problem, since community structure was only weakly related to sediment grain-size, depth, salinity, and temperature data (Table 18.5), only about 10 to 15 percent of the spatial variation was explained by these variables. This highlights a need for additional environmental data useful in characterizing habitats. Interestingly, collection and use of new environmental data for habitat analyses has already begun in the Lower Bay Complex. Side-scan sonar imagery is available for some areas (Schwab et al., 1997), and NOAA-USACE (2001) has generated very high resolution habitat maps for the Lower Bay Complex based on extensive use of sediment-profile and sediment-surface imagery. By combining these data with water quality, sediment contaminant, and sediment toxicity data, it should be possible to substantially increase the amount of explained variation and make direct analyses meaningful.

What caused the observed habitat change and especially the decline in 1973-74? Diaz and Boesch (1979) and Berg and Levinton (1984) in comparing 1957-60 to 1973-74 data felt that the environmental data available to them, particularly data on pollutants in sediments, were not adequate to assign specific causes. Diaz and Boesch (1979) thought that the cause was anthropogenic, particularly because the numerical dominants (Ampelisca abdita, Cyathura polita, Mya arenaria, and Ilyanassa obsoleta) all declined. They indicated that low abundance was not a typical organic enrichment response; low dissolved oxygen and high sulfides, conditions that are associated with enrichment, generally allow dense populations of a few stress tolerant species to occur. Instead, they suggested that the response was more consistent with toxicants. Ampelisca abdita in particular is sensitive to contaminants and is often used in sediment toxicity tests (e.g., Wolfe, Long, and Thursby, 1996). Other possibilities suggested by investigators included sewage and industrial pollution (Franz, 1982), dredging (Franz, 1982; MacKenzie, 1983), siltation (MacKenzie, 1983), increased salinity (MacKenzie, 1983), heavy metals (Diaz and Boesch, 1979; Steimle and Caracciolo-Ward, 1989), chlorinated pesticides and PCBs (Diaz and Boesch, 1979), and extractable hydrocarbons (Steimle and Caracciolo-Ward, 1989).

Stainken (1984), sampling along two transects in Raritan Bay in June and September-October 1977, reported little relationship betweenbiotic indices of diversity and specific sediment contaminants. He did, however, find a trend of increased abundance and diversity with distance from inner Raritan Bay that correlated with both decreasing silt-clay content and decreasing levels of PAHs, PCBs, and extractable hydrocarbons. Faunal abundances were depressed in general, and perhaps even more significantly, Ampelisca abdita was not reported in any of his samples. Even with the 1.5 mm mesh used in his study, Stainken should have retained some individuals of this species had itbeenpresent. Combining McGrath's 1973-74data (Table 18.3) with Stainken's 1977 study suggests that a dominant mud species, Ampelisca abdita, may have been absent for several years. By 1983, this species had clearly returned as a dominant (Table 18.3).

A specific cause for the observed habitat change will probably never be identified. In fact, Franz (1982) suggested that major habitat change actually began to occur in the Lower Bay Complex by the late 1800s, that is, well before any recent studies. By comparing the number of molluscan species from studies conducted in the late 1800s and 1920s, he concluded that many species were eliminated when eelgrass and oyster beds declined. Extensive oyster beds were present in western Raritan Bay during the late 1800s and once supported an active oyster industry (MacKenzie, 1983).

Adams et al. (1998) reported that the benthos in 1993-94 in the Lower Bay Complex was the least impacted in New York-New Jersey Harbor. The most extensive sediment contaminants in the Harbor were mercury, chlordane, and PCBs, and high concentrations were restricted to muddy areas. Similar results were also obtained for toxicity tests both in 1991 (Wolfe et al., 1996) and 1993-94 (Adams et al., 1998). Overall, contaminant concentrations and sediment toxicity were lower compared to Newark Bay and Upper Bay, but New York-New Jersey Harbor was generally more contaminated than other sites in the mid-Atlantic region from Cape Cod to Chesapeake Bay.

Even without identifying a specific cause for the poor health of the benthic community from the

1950s to the 1970s, improvement in the benthic fauna by the 1980s is almost certainly due to a significant improvement in water quality in the 1970s. This occurred primarily as a result of the Clean Water Act, when existing plants in the region were upgraded to secondary treatment and additional plants were constructed (Brosnan and O'Shea, 1996). Secondary treatment resulted in decreased loadings of organic carbon, phosphorus, metals such as cadmium, copper, and lead, and PCBs. Untreated water discharges into the Lower Hudson River, for example, decreased from 19.7 m3/s in 1970 to 0.2 m3/s in 1988 (Brosnan and O'Shea, 1996). Indicators of water quality, such as fecal coliform and dissolved oxygen, in the Lower Bay Complex showed noticeable improvement from the 1970s to the 1990s (NYCDEP, 1995), and with the exception of some results in western Raritan Bay, the Lower Bay Complex met water quality standards for the fifteen-year period from 1985-2000 (NYCDEP, 2000).

Despite impressive gains in controlling inputs, it is likely that contaminant concentrations and sediment toxicity such as that observed by Adams et al. (1998) will continue to persist in muddy areas for at least several more decades due to the high binding capacity of fine-grained sediments. Improvements in the benthic fauna should continue in the Lower Bay Complex but at a slower rate than seen during the past 20 years. A reasonable goal would be the restoration of eelgrass and oyster beds to the Bay as was typical of a century ago. That goal is certainly obtainable but probably decades off.


Special thanks to Frank Steimle, Pace Wilber, and Robert Will for sharing regional survey data with me, to Mark Wiggins for his meticulous work on the 1986-87 study, to Shawn MacCafferty for early help with data analysis, and to the Hudson River Foundation for providing partial support.


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19 The Benthic Animal Communities of the Tidal-Freshwater Hudson River Estuary

David L. Strayer abstract Benthic animals (those that live in or on sediments or vegetation) are of key importance in the Hudson River ecosystem. They are the major source of food to the Hudson's fish and regulate the abundance and composition of phytoplankton in the river. Benthic animals probably are important in mixing sediments, an activity that may affect the movement and ultimate fate of toxins in the river, although this process is not well studied in the Hudson. The benthic animal community of the Hudson is diverse, containing several hundred species of worms, mollusks, crustaceans, insects, and other invertebrates. These animals represent a wide array of life histories, feeding types, distributions, and adaptations. Community structure and population density vary greatly from place to place in the Hudson, and are determined chiefly by salinity, the presence of rooted plants, and the nature ofthe sediment (hard vs. soft). Nevertheless, agreat deal of site-to-site variation in benthic community structure in the Hudson and other large rivers is unexplained. Human activities (especially water pollution and alteration of the channel for navigation) probably had large effects on the benthic communities ofthe Hudson, but these effects have not been well documented. The recent invasion of the Hudson by the zebra mussel (Dreissena polymorpha) profoundly changed the benthic communities ofthe river, altering their composition and function in the ecosystem.


Benthic animals (collectively called the zooben-thos) are a diverse community of animals living in or on sediments, aquatic plants, or other solid surfaces under the waters. The zoobenthos of the

Hudson is one of the most diverse communities in the river, containing several hundred species of varied habits. Benthic animals play key roles in the river's ecosystem. They are the predominant food for many of the river's fish, regulate populations of phytoplankton and zooplankton, and probably are important in determining the movement and fate of nutrients and toxins in the river. Despite this importance, much remains unknown about the Hudson's benthic animals and their roles in the river's ecosystem. My goal in this chapter is to describe the animals that make up the Hudson's zoobenthos, discuss how different habitats within the river support different kinds of benthic animals, review how benthic animal communities in the river have changed over time, especially in response to the zebra mussel invasion, and evaluate the importance of benthic animals in the Hudson's ecosystem.

Sources of Information

Studies of the Hudson's zoobenthos have been spotty, limiting our insight into this part of the ecosystem. The earliest naturalists collected specimens from the Hudson and made incidental observations on benthic animals (e.g., Say, 1821; Lea, 1829; Dekay, 1844; Gordon, 1986), but the first systematic survey of the Hudson's zoobenthos was done by Townes (1937), who made a few collections from the middle estuary as part of the Conservation Department's survey of New York's fisheries resources. In the 1970s, the Boyce Thompson Institute (Ristich, Crandall, andFortier, 1977;Weinstein, 1977) surveyed the benthos of the lower estuary (Manhattan to Poughkeepsie), and in 1983-84 researchers from the New York State Department of Health (Simpson et al., 1984, 1985, 1986; Bode et al., 1986) made a detailed study of the zoobenthos of the main channel of the freshwater Hudson from Troy to New Hamburg. Two vegetated areas (Bowline Pond - Menzie, 1980, and Tivoli South Bay - Findlay, Schoeberl, and Wagner, 1989) were studied during the same time period. Finally, my colleagues and I have studied the zoobenthos of the freshwater tidal section of the river (Troy to Newburgh) since 1990, a period that included the zebra mussel invasion (Strayer et al., 1994, 1996, 1998; Strayer and Smith, 1996, 2000, 2001).

Figure 19.1. Some benthic animals that are common in the Hudson River. A. the flatworm Hydrolimaxgrisea, B. the nematode Dorylaimus stagnalis, C. a tubificid oligochaete, D. the poly-chaete Eteoneheteropoda,E. the isopod Cyathura polita,F. the amphipod Leptocheirusplumulosus, G. the blue crab Callinectes sapidus, H. the caddisfly Oecetis inconspicua, and I. its case, J. the chi-ronomid Ablabesmyia, K. the phantom midge Chaoborus, L. the snail Amnicola limosa, M. the mussel Anodonta implicata. From Thorne and Swanger (1936), Hyman and Jones (1959), Burch (1975), Weinstein (1977), Oliver and Roussel (1983), Fryer (1991), Jokinen (1992), and Wiggins (1996).

Figure 19.1. Some benthic animals that are common in the Hudson River. A. the flatworm Hydrolimaxgrisea, B. the nematode Dorylaimus stagnalis, C. a tubificid oligochaete, D. the poly-chaete Eteoneheteropoda,E. the isopod Cyathura polita,F. the amphipod Leptocheirusplumulosus, G. the blue crab Callinectes sapidus, H. the caddisfly Oecetis inconspicua, and I. its case, J. the chi-ronomid Ablabesmyia, K. the phantom midge Chaoborus, L. the snail Amnicola limosa, M. the mussel Anodonta implicata. From Thorne and Swanger (1936), Hyman and Jones (1959), Burch (1975), Weinstein (1977), Oliver and Roussel (1983), Fryer (1991), Jokinen (1992), and Wiggins (1996).

In addition to these large studies, a number of studies more limited in scope (e.g., Hirschfield, Rachlin, and Leff, 1966; Howells, Musnick, and Hirschfield, 1969; Williams, Hogan, and Zo, 1975; Crandall, 1977;Yozzo and Steineck, 1994) have contributed information on the Hudson's zoobenthos. Together, these studies offer a moderately clear picture of benthic animal communities of the freshwater tidal river in 1983-2000, a glimpse into communities of the lower river in the mid-1970s, and only hints of the benthic communities that lived anywhere in the river before 1970.

Further, most of the studies in the Hudson have been focused on the macrofauna (animals large enough to be caught on a 0.5-1 mm mesh screen), and have excluded the numerous smaller animals as well as larger mobile forms such as crabs and shrimp. Typically, these excluded forms constitute

5-75 percent of benthic biomass, production, and diversity (e.g., Strayer, 1985; Hakenkamp, Morin, and Strayer, 2002). Consequently, benthic animals in the Hudson are more numerous and more diverse than existing studies on the Hudson suggest.

Biology of the Zoobenthos

Approximately three hundred species of mac-robenthic animals have been recorded from the Hudson River (Ristich et al., 1977; Simpson et al., 1986; Strayer and Smith, 2001). This fauna includes animals with a wide array of body sizes and shapes (Fig. 19.1), life histories, and ecological habits. In terms of numbers, biomass, and species richness, the most important groups in the Hudson's zoobenthos are annelids, mollusks, crustaceans, and insects.

Three major groups of annelids are common in the Hudson: leeches, oligochaetes, and poly-chaetes. Although leeches are well known (and reviled!) as bloodsuckers, only a few species of leeches are parasites of humans and other vertebrates. Most leech species are scavengers or predators of invertebrates. About ten species of leeches have been reported from the freshwater parts of the Hudson. Leech densities usually are low in the Hudson, but these animals may be locally important as predators in plant beds, where their densities are highest. Most oligochaetes and poly-chaetes burrow in soft sediments or crawl on vegetation or rocks and are deposit-feeders, feeding on sediment bacteria and organic matter. Many species are macroscopic, and reach lengths of 3-30 mm as adults. Polychaetes are predominately marine, and are dominant in the polyhaline and mesohaline parts of the Hudson (river kilometer (RKM) 0-75). Only one species (the microscopic Manayunkia speciosa) lives in the freshwater part of the estuary. Oligochaetes live throughout the river, but are especially common in the freshwater estuary (RKM 100-248), where they often constitute >75 percent of macrobenthic animals. Scientists have thus far found twenty to thirty species each of oligochaetes and polychaetes in the Hudson.

Mollusks (clams, mussels, and snails) are among the most familiar of the benthic animals in the Hudson. About fifty species have been reported from the river. Bivalve mollusks (clams and mussels) feed either on phytoplankton and other suspended material (suspension-feed) or on organic matter deposited on the sediments (deposit-feed). While some bivalves are among the largest invertebrates in the river, reaching >10 cm long, others never reach 5 mm long, even as adults. The life cycles of our bivalves are highly varied. Most of the brackish-water species have free-living larvae, but most freshwater species either have larvae that are parasitic on fish (pearly mussels) or no larvae at all (pea clams). The pearly mussels may live for decades. Some of the bivalves in the Hudson are edible (for example, oysters, mussels), and have been fished in prehistoric (e.g., Schaper, 1989) and recent times (because of widespread contamination, it is probably not a good idea to eat mol-lusks from the river today). Most of the Hudson's snails graze on attached algae or deposit feed on organic sediments; a few are able to suspension feed. Several alien mollusk species have been introduced to the Hudson (e.g., the zebra mussel Dreis-sena polymorpha, the dark false mussel Mytilop-sis leucophaeta, the Atlantic rangia Rangia cuneata, the faucet snail, Bithynia tentaculata) and are now common in the river.

Although only about thirty species of benthic crustaceans (isopods, amphipods, barnacles, and decapods) have been reported from the Hudson, the crustaceans are among the most important benthic animals in the river. They often are abundant, and many are especially choice food for fish (Table 19.1). Isopods (relatives of the familiar terrestrial pill bug) are common on unvegetated sediments throughout the river. Amphipods (scuds, sideswimmers) are small shrimplike crustaceans common throughout the river that are one of the most important fish foods in the river (Table 19.1). Barnacles live on rocky shorelines as far north as Beacon (RKM 99). The decapods (crabs, crayfish, and shrimp) are another important fish food, but have received little study in the Hudson. Crayfish live in freshwater habitats, grass shrimp (Paleomonetes) live in brackish habitats, and blue crabs (Callinectes sapidus) migrate from the lower estuary as far north as Troy in some summers. Blue crabs (color plate 7) are widely fished for food in the Hudson and elsewhere; in recent years, the commercial catch in the river was ~40,000 kg/yr (NYSDEC, 1993). Many marine crustaceans have free-swimming larvae, and larval crabs and barnacles are common in the plankton on the lower Hudson. In contrast, most freshwater crustaceans have no larval stage, and develop directly from egg to juvenile to adult.

Benthic insects are common in the Hudson, especially in freshwater habitats. The chirono-mid midges (larvae of non-biting flies) are by far the most abundant and species-rich of the insects (color plate 8). Chironomid densities in the freshwater Hudson typically are ~1,000/m2. More than 70 species of chironomids have been identified from the Hudson, and true diversity probably exceeds 100 species. The chironomids are a diverse group that includes predators, suspensionfeeders, and grazers. Other insects that may be locally abundant in the freshwater Hudson include Ephemeroptera (mayflies), Odonata (damselflies),

Table 19.1. Importance of benthic invertebrates in the diets of some Hudson River fishes.

Fish species

% of diet

Dominant items in diet


Shortnose sturgeon (YOY)

100 (V)


Carlson and Simpson, 1987

Shortnose sturgeon

100 (V)

Chironomids, mollusks,

Curran and Ries, 1937

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