SEARCH

SEARCH BY CITATION

Keywords:

  • Sediment;
  • Contamination;
  • Toxicity;
  • Estuaries;
  • Salinity

Abstract

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. UNIQUE CHARACTERISTICS OF ESTUARIES
  5. Unique biological characteristics of estuaries
  6. SALINITY AS A CONTROLLING FACTOR
  7. ASSESSMENT TECHNIQUES: CHEMISTRY
  8. ASSESSMENT TECHNIQUES: BIOLOGICAL SURVEYS
  9. ASSESSMENT TECHNIQUES: TOXICITY TESTS
  10. SUMMARY AND CONCLUSIONS: THE WAY FORWARD
  11. Acknowledgements
  12. REFERENCES

Historic and ongoing sediment contamination adversely affects estuaries, among the most productive marine ecosystems in the world. However, all estuaries are not the same, and estuarine sediments cannot be treated as either fresh or marine sediments or properly assessed without understanding both seasonal and spatial estuarine variability and processes, which are reviewed. Estuaries are physicochemically unique, primarily because of their variable salinity but also because of their strong gradients in other parameters, such as temperature, pH, dissolved oxygen, redox potential, and amount and composition of particles. Salinity (overlying and interstitial) varies spatially (laterally, vertically) and temporally and is the controlling factor for partitioning of contaminants between sediments and overlying or interstitial water. Salinity also controls the distribution and types of estuarine biota. Benthic infauna are affected by interstitial salinities that can be very different than overlying salinities, resulting in large-scale seasonal species shifts in salt wedge estuaries. There are fewer estuarine species than fresh or marine species (the paradox of brackish water). Chemical, toxicological, and community-level assessment techniques for estuarine sediment are reviewed and assessed, including chemistry (grain size effects, background enrichment, bioavailability, sediment quality values, interstitial water chemistry), biological surveys, and whole sediment toxicity testing (single-species tests, potential confounding factors, community level tests, laboratory-to-field comparisons). Based on this review, there is a clear need to tailor such assessment techniques specifically for estuarine environments. For instance, bioavailability models including equilibrium partitioning may have little applicability to estuarine sediments, appropriate reference comparisons are difficult in biological surveys, and there are too few full-gradient estuarine sediment toxicity tests available. Specific recommendations are made to address these and other issues.


INTRODUCTION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. UNIQUE CHARACTERISTICS OF ESTUARIES
  5. Unique biological characteristics of estuaries
  6. SALINITY AS A CONTROLLING FACTOR
  7. ASSESSMENT TECHNIQUES: CHEMISTRY
  8. ASSESSMENT TECHNIQUES: BIOLOGICAL SURVEYS
  9. ASSESSMENT TECHNIQUES: TOXICITY TESTS
  10. SUMMARY AND CONCLUSIONS: THE WAY FORWARD
  11. Acknowledgements
  12. REFERENCES

Sediment contamination is receiving increasing attention from the scientific community. However, most sediment assessment methods developed to date are applicable primarily to fresh or salt waters, not to waters that are truly estuarine. Yet estuaries receive significant anthropogenic inputs from both point and nonpoint sources upstream and from metropolitan areas and industries located on or near estuaries. Over 30 years ago, they were called “the septic tank of the megalopolis” [1]; historic contamination remains a significant concern for many estuarine sediments [2–5]. Sediment contamination in estuaries has been associated not only with effects on benthic species [6] but also with effects on water column species, including migrating salmon [7].

Estuaries are among the most productive marine ecosystems in the world [8] and are critical to the life history and development (e.g., rearing, feeding, migration routes, and nursery grounds) of many aquatic species. Thus, it is critical that sediment contamination in estuaries and its biological and ecological significance be properly and fully assessed.

Toxicity and other assessment methods developed specifically for estuarine sediments are few and relatively new, and there is a disturbing tendency, in the case of both chemical and biological assessment methods, to treat estuaries as either freshwater tainted by salt or the reverse. Many studies and researchers do not appear to properly appreciate the unique and dynamic nature of estuarine ecosystems.

The purpose of this paper is to provide that appreciation as part of a review of the state-of-the-art with regard to estuarine sediment contamination assessment methods. The paper begins with a description of the unique characteristics of estuaries. It then proceeds to review available assessment techniques beginning with sediment chemistry and continuing through biological surveys and sediment toxicity testing. The paper concludes by suggesting the way forward.

UNIQUE CHARACTERISTICS OF ESTUARIES

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. UNIQUE CHARACTERISTICS OF ESTUARIES
  5. Unique biological characteristics of estuaries
  6. SALINITY AS A CONTROLLING FACTOR
  7. ASSESSMENT TECHNIQUES: CHEMISTRY
  8. ASSESSMENT TECHNIQUES: BIOLOGICAL SURVEYS
  9. ASSESSMENT TECHNIQUES: TOXICITY TESTS
  10. SUMMARY AND CONCLUSIONS: THE WAY FORWARD
  11. Acknowledgements
  12. REFERENCES

What is an estuary?

The term “estuary” has historically referred to the lower tidal reaches of a river. Pritchard [9] defines an estuary as follows: “An estuary is a semi-enclosed coastal body of water which has a free connection with the open sea and within which sea water is measurably diluted with fresh water derived from land drainage.” He further divides estuaries into four classes based on physical characteristics (Table 1). Caspers [10] only partly agrees with this definition, arguing that tidal changes are also necessary for estuaries; in this context, the Mediterranean, Black, and Baltic Seas, for instance, would contain no estuaries.

Intuitively, it seems that an estuary could be defined most simply based on salinity, specifically as an area where freshwater enters saline water and where the salinity is, during at least some period of time, neither truly saline (>30 g/L [ppt]) nor truly fresh (0 g/L). However, there are areas that fit this definition that are not estuaries. Examples include groundwater flow into the sea (e.g., Biscayne Bay, FL, USA) and anthropogenic point-source discharges, such as sewage effluent. Further, the word “estuary” is derived from the Latin aestus, the tide. Thus, our suggested definition of an estuary is based on Pritchard's [9] physical definition but also includes mention of tidal influences: “An estuary is a semi-enclosed and tidal coastal body of water which has a free connection with the open sea and within which sea water is measurably diluted with fresh water derived from land drainage.” We define estuarine sediments as sediments whose interstitial salinities are neither truly fresh nor truly saline; that is, they range above 1 and below 30 g/L. This latter distinction is particularly important as sediments with lesser or greater interstitial salinities can and do occur within regions that are designated as estuaries but are, respectively, more appropriately subject to freshwater and marine assessment methods.

Table Table 1.. Estuaries defined by physical characteristics (adapted from Pritchard [9])
Type of estuaryDescriptionExample(s)
Drowned river valley FjordFound along coastlines with wide coastal plains; only a portion of the area affected by tides is estuarine based on salinity diluted by freshwater Generally U-shaped, gouged out by glaciers; mouth often has a shallow sill preceding a deep basin (e.g., >100 m depth, Fig. 1)Chesapeake Bay, Maryland, USA Norwegian and British Columbia (Canada) coastlines
Bar-builtOffshore barrier sand islands and sand spits build above sea level and extend between headlands in a chain broken by one or more inlets; enclosed area generally elongated relative to shoreline; may be more than one river drainage; reduced tidal action; wind-driven circulationPamlico Estuary, North Carolina, USA
TectonicCoastal indentures formed by faulting or local subsistence, with an excess freshwater flowSan Francisco Bay, California, USA

In addition to physical characteristics (Table 1), estuaries differ based on the relationship and mixing between fresh and salt water. Relative to the sediments, there are effectively three different types of estuaries, characterized as follows.

Salt wedge estuaries (Fig. 1). River flow is dominant. Salt water moves up the estuary in the form of a defined wedge whose upstream penetration is greatest during periods of low river flow (e.g., winter or low-rainfall periods) and least during periods of high river flow (e.g., freshet). Salt wedge estuaries include the mouths of some major rivers (e.g., the Fraser River, BC, Canada). In salt wedge estuaries (see Assessment Techniques: Biological Surveys), seasonal shifts occur in the distribution of benthic infaunal species related to seasonal, not diurnal, changes in interstitial salinities [11]. These same salinity shifts can affect the availability of sediment contaminants (see Assessment Techniques: Chemistry).

Shallow, partially mixed estuaries and fjords (Fig. 1). Similar to salt wedge estuaries, there are two layers of flow: salt water at depth and freshwater on the surface. However, river flow is modified by tidal currents. Shallower estuaries, including bar-built estuaries, are subject to vertical mixing; deeper estuaries such as fjords are subject to entrainment (one-way movement of saline waters into surface freshwaters). The estuary may be partially mixed or highly stratified. In these estuaries, the area of the bottom that is transitional between arine and freshwater environments does not demonstrate the large-scale seasonal movements up- and downstream typical of salt wedge estuaries. Examples include the estuaries of the Chesapeake Bay system (e.g., the James River, VA) in the United States and the River Tees and the Thames River in the United Kingdom. Some fjords, because of restricted bottom-water circulation associated with sills at their mouths (Fig. 1) and that are relatively deep, suffer from oxygen deficiency and increased hydrogen sulfide concentrations in the bottom waters.

thumbnail image

Figure Fig. 1.. Major estuary types related to bottom-water salinities. Not shown: vertically homogeneous estuary. Arrows indicate direction of water movement.

Download figure to PowerPoint

Vertically homogeneous estuaries. Tidal currents predominate, and vertical salinity differences are less than 1 g/L because of intense vertical mixing (e.g., the Severn Estuary, UK). There is, as for all estuaries, a horizontal gradient of salinity, increasing from the head to the mouth. However, lateral variation can occur where the ratio of width to depth is sufficiently large, such that the right-hand-side of the estuary (looking to sea) will contain lower-salinity water than on the left-hand side. Shallow, well-mixed estuaries exhibit intense coupling between benthic and pelagic systems [12].

There is, of course, a spectrum of estuarine patterns and gradual transitions between the three previously mentioned major types, and estuaries may differ seasonally. For example, the Vellar estuary in India is a salt wedge estuary during periods of high river flow but can become homogeneous in the hot season with no river flow; the Columbia River estuary, Washington, USA, can vary from a salt wedge to a highly stratified estuary [13]. There are, in addition, exceptional cases where, for instance, nearly all mixing occurs in a very limited area. An example is the reversing falls in the gorge of the St. John River, New Brunswick, Canada. Constrictions and promontories can also greatly affect estuarine circulation patterns, and complex circulation patterns occur where a number of estuaries are tributary to a large estuarine system, such as the Bay of Fundy, New Brunswick, Canada.

Estuaries are ephemeral phenomena in the context of geological time. Advancing river deltas tend to destroy estuaries. Historic estuaries include, for example, the Rio Tapajos, off the Amazon River (Brazil, South America); the Latmian Gulf in Turkey, which is now Lake Bafa; and the former Lake Atchafalaya, off the Mississippi River, Louisiana, USA [14]. Further, because flow, tidal range, and sediment distribution are continually changing, estuaries are not steady-state systems, nor is one estuary the same as another, which makes the problem of detecting subtle anthropogenic changes (both spatial and temporal) particularly challenging.

The bottom line is that estuaries are dynamic, complex, and unique systems. Truly estuarine sediments cannot be treated as either freshwater or marine systems, and they cannot be properly assessed by means of only snapshot-in-time sampling without understanding both seasonal and spatial variability and processes, in particular bottom-water and sediment interstitial salinity variations.

Unique physicochemical characteristics of estuaries

As noted previously, estuaries are associated with rivers or other forms of runoff from land. They are the immediate recipients of sediment carried by those rivers, as manifest by the formation of river deltas. However, this does not imply that sediments in estuaries are all fluvial in origin. Depending on the sediment load of the entering river and on the estuarine circulation patterns, sediments in estuaries can come from inland and/or from the sea [15]. Estuaries with entering rivers with high sediment loads may be filled rapidly with fluvial sediments. In contrast, estuaries with entering streams with low sediment loads may be filled solely or primarily by marine sediments. This can occur because of the penetration of salt water (as in salt wedge estuaries; Fig. 1) and/or diffusion of marine suspended matter. As a result, deposition of marine sediments can occur upriver of saltwater penetration [16].

Although many estuaries are the recipients of deposited sands, the most common sediment deposits for estuaries with limited wave action are fine-grained muds that form shoals and tidal flats. Particle settling can be enhanced by the change from fresh to salt water, the rise and fall of the water level with the tides, and the presence of turbidity maxima during slack tides. Cohesive sediment fluxes can result in both deposition and erosion [17]; tidal and wave action across inter-tidal areas can result in complex suspended sediment dynamics [18]. As noted by McManus [19] relative to estuarine sedimentation, “Change is the norm within the estuarine system.”

Because they provide an interface between fresh and salt waters, estuaries have strong gradients in many physical and chemical variables, including salinity, temperature, pH, dissolved oxygen, redox potential, nutrients, and amount and composition of particles. Gradients exist not only along their length, from river to sea, and laterally but also vertically (from water column to sediment), particularly in stratified and partially stratified estuaries such as salt wedge estuaries and fjords. These gradients may also be subject to seasonal and other temporal variations, which can have a wide variation in influence on many biogeochemical processes occurring in estuaries. Under anoxic conditions, for example, some metals, such as Fe and Mn, are mobilized from reducing sediments and remain dissolved in the water column [20], whereas other metals, such as Cd, Cu, Zn, and Cr, may be removed from the water column by sulfide precipitation or by reduction to insoluble solids [21]. Effects of temporal variability can be far from negligible when assessing sediment quality in estuarine wetland areas [22].

Above all, the most unique characteristic that distinguishes estuaries from fresh and salt waters is their variable salinity. This is discussed in detail later in this paper.

Unique biological characteristics of estuaries

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. UNIQUE CHARACTERISTICS OF ESTUARIES
  5. Unique biological characteristics of estuaries
  6. SALINITY AS A CONTROLLING FACTOR
  7. ASSESSMENT TECHNIQUES: CHEMISTRY
  8. ASSESSMENT TECHNIQUES: BIOLOGICAL SURVEYS
  9. ASSESSMENT TECHNIQUES: TOXICITY TESTS
  10. SUMMARY AND CONCLUSIONS: THE WAY FORWARD
  11. Acknowledgements
  12. REFERENCES

Estuaries are extremely productive and important feeding, migration, and rearing zones. However, they are also transitional areas that are challenging to both residents and immigrants alike. Ecologically, estuaries are zones of reduced interspecific (but not intraspecific) biotic competition due to overriding physical-chemical factors, in particular salinity. Benthic fauna in estuaries either filter their food directly from the water column or depend on the physical deposition of food particles onto the sediment surface or incorporated into the sediment matrix [12]. Faunal distributions in estuaries are controlled primarily by salinity and secondarily by factors such as substrate, temperature, dissolved oxygen, and anthropogenic pollution [23–26]. In many cases, there is a complex correlation between temperature and salinity, each of which is capable of modifying biological tolerances to the other factor [27]. Variations in salinity in both the overlying waters and the sediment interstitial waters occur vertically, horizontally and with time. Organisms can survive in estuaries by one or a combination of the following strategies: avoiding estuarine conditions (e.g., saltwater organisms remaining within the salt wedge in salt wedge estuaries, freshwater organisms remaining above), reducing contact with inimical environments (e.g., interstitial waters of muddy sediments can have very different salinities than overlying waters [11]), adaptation (e.g., ion regulation, volume regulation, or osmoregulation [28]), or acclimation. However, not all organisms living in estuaries live under optimal conditions, which results in natural bioenergetic stress to those organisms [29]. For example, periodic seasonal cycles of anoxic and oxic conditions in bottom waters of partially stratified estuaries (e.g., in Chesapeake Bay) result in a corresponding cycle of mortality and recolonization by benthic macrofauna [30]. Estuarine organisms living in such stressful natural conditions may be more (or less) susceptible to anthropogenic stress.

thumbnail image

Figure Fig. 2.. Portion of the North Arm of the Fraser River, British Columbia, Canada, for which bottom sediment interstitial salinities vary seasonally related to up- and downstream movements of the salt wedge [11,31].

Download figure to PowerPoint

SALINITY AS A CONTROLLING FACTOR

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. UNIQUE CHARACTERISTICS OF ESTUARIES
  5. Unique biological characteristics of estuaries
  6. SALINITY AS A CONTROLLING FACTOR
  7. ASSESSMENT TECHNIQUES: CHEMISTRY
  8. ASSESSMENT TECHNIQUES: BIOLOGICAL SURVEYS
  9. ASSESSMENT TECHNIQUES: TOXICITY TESTS
  10. SUMMARY AND CONCLUSIONS: THE WAY FORWARD
  11. Acknowledgements
  12. REFERENCES

Spatial and temporal fluctuations of salinity in estuarine sediments

As mentioned previously, salinity can fluctuate in both overlying waters and sediment interstitial waters spatially (horizontally and vertically) and temporally. It is not unusual for tens of square kilometers of bottom sediments in salt wedge estuaries to show seasonal interstitial salinity differences related to seasonal differences in the extent and duration of the salt wedge (e.g., the Columbia River, WA, USA [13], and the Fraser River, BC, Canada [11]; Fig. 2). Changes in interstitial salinities related to overlying water salinities are a function of both sediment type and duration of exposure. Figure 3 illustrates three scenarios: sediments that are primarily sands and sediments that are primarily silts and clays, both with highly saline interstitial salinities exposed to overlying freshwater, and the same exposure involving sediments that are primarily silts and clays but with intermediate interstitial salinities. Exchange and equilibration between interstitial and overlying water is fast in sands but slow in sediments containing high proportions of silts and clays. Surface sediments are more rapidly affected than bottom sediments, and differences can persist for days between interstitial salinities at different sediment depths for sediments with high proportions of silts and clays. These patterns are related not only to differences between overlying and interstitial salinities but also to differences in the porosity and permeability of sediments [31,32].

thumbnail image

Figure Fig. 3.. Relative changes to saline sediment interstitial salinities with exposure to less saline overlying waters. (A) sediments >90% sand with 20-g/L interstitial salinities exposed to fresh overlying water. (B) Same as (A) but for sediments on the order of 80% silt and clay. (C) Same sediments and scenario as (A) except initial interstitial salinities 10 g/L. Top of each curve: interstitial salinities in the top centimeter of sediment; bottom of each curve: interstitial salinities in the bottom 6 cm of sediment. Schematic developed based on actual laboratory and field measurements [31,32]. A similar but reverse pattern occurs when sediments interstitial waters are fresh and overlying waters are saline.

Download figure to PowerPoint

Porosity, the fraction of sediment volume occupied by water, usually decreases exponentially with depth because of compaction [33]. Permeability represents a proportionality factor between water pressure gradient and water flow. Voids between larger particles are more available in sands than in muds. In the latter case, these voids tend to become clogged by finer sediments, thus reducing permeability and the mixing of interstitial and overlying waters (Fig. 3).

Salinity as a controlling factor for contaminant partitioning and bioavailability

Unlike freshwaters, where pH is the controlling factor, in estuaries salinity is the controlling factor for the partitioning of contaminants between sediments and overlying or interstitial waters. The partitioning coefficient (Kd) of a contaminant is defined as the ratio of a contaminant concentration in the sediment to that dissolved in the overlying or interstitial water. High ionic strengths in estuarine waters can salt out hydrophobic organic chemicals from the water to the sediment phase [34,35]. In addition, increasing salinity enhances the removal of dissolved organic matter from the water to the sediment phase and the formed particulate organic matter can effectively sorb hydrophobic chemicals [35]. As a result, an increase in salinity generally results in an increase in KdS for hydrophobic chemicals. In contrast, KdS for metals may decrease (e.g., Cd, Zn), increase (e.g., Fe), or be constant (e.g., Ir) when salinity increases, depending on the relative importance of the two counteractive processes [36–38]: (1) desorption due to increasing complexation with seawater anions (Cl- and SO2−4) and/or increasing competition for particle sorption sites with seawater cations (Na+, K+, Ca2+, Mg2+) and (2) coagulation, flocculation, and precipitation. Light rare earth elements (e.g., La, associated with zeolites from oil refineries) tend to be less efficiently trapped by low-salinity sediments [39].

Because it affects the partitioning of contaminants between sediments and overlying or interstitial water, salinity also affects the bioavailability of contaminants in estuarine sediments. For example, Schlekat et al. [40] found that bioavailability of cadmium associated with bacterial exopolymer sediment coatings to the amphipod Leptocheirus plumulosus was dependent on both seawater salinity (greatest at estuarine salinities) and cadmium concentrations. Partitioning to particles favors uptake by sediment feeders [41], whereas desorption to the water (overlying or interstitial) favors uptake via dermal exchange surfaces such as gills. However, while desorption of metals with increasing salinity can increase metal concentrations in water, concurrent increases in Ca2+ and Cl- can decrease water bioavailability [42,43].

It is well established that dissolved organic matter can affect the bioavailability of inorganic [44,45] and organic [46,47] contaminants in fresh and salt waters. Complexation kinetics, which are important in freshwater [44], are no less important related to salinity [48,49]. Exposing freshwater organisms to some contaminants (e.g., metals other than mercury) in waters of increasing salinity (below toxicity thresholds for salinity) demonstrably increases the toxicity threshold for such contaminants [50–52]. Similarly, decreasing the salinity content of exposure waters increases the toxicity of metals such as nickel, zinc, and chromium to saltwater species [53,54].

Since changes in the bioavailability and toxicity of contaminants in estuarine sediments can occur because of changes in the ionic strength (salinities) of the exposure waters [55–57], it is imperative to measure overlying and interstitial salinities when assessing sediment contamination.

Salinity as a controlling factor for estuarine biota

Salinity is also the most important natural factor controlling the distribution of estuarine organisms [23,24,58]. This is true not only for water column biota but also for benthic biota [59], even though some researchers refer to salinity (together with depth and sediment grain size) as nuisance variables [60]. The bottom line is that benthic organisms burrowing in the sediments can be exposed to very different salinity regimes than if they were on the sediment surface. The extent of exposure differences depends on factors such as sediment type, the difference between overlying and interstitial water salinities, duration of exposure to different overlying water salinities, and the organism's depth in the sediment. This has significant implications both to the distributions of organisms in such estuarine areas (see Assessment Techniques: Biological Surveys) and to sediment toxicity tests of these sediments (see Assessment Techniques: Toxicity Tests). Further, not all species of the same estuarine genus have similar salinity tolerances [27].

Salinity, of course, can be toxic. In many ways, estuaries are not dissimilar to freshwaters subject to nonpoint-source runoff. In both cases the biota are exposed in a nonconstant manner to a factor (variable salinity in estuaries) and factors (contaminants in nonpoint-source runoff) that can be toxic. Studies of nonpoint-source runoff have shown that laboratory toxicity tests involving constant exposures, effectively a snapshot in time, cannot adequately predict in situ toxicity. Depending on the situation, the laboratory may under- or over-predict toxicity [61]. There is no reason to expect this situation would be any different in estuaries. In fact, there is every reason to expect a similar situation.

ASSESSMENT TECHNIQUES: CHEMISTRY

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. UNIQUE CHARACTERISTICS OF ESTUARIES
  5. Unique biological characteristics of estuaries
  6. SALINITY AS A CONTROLLING FACTOR
  7. ASSESSMENT TECHNIQUES: CHEMISTRY
  8. ASSESSMENT TECHNIQUES: BIOLOGICAL SURVEYS
  9. ASSESSMENT TECHNIQUES: TOXICITY TESTS
  10. SUMMARY AND CONCLUSIONS: THE WAY FORWARD
  11. Acknowledgements
  12. REFERENCES

From a chemical standpoint, an estuary is a reaction vessel where chemically very different fresh and salt water are drastically mixed. Contaminants in estuaries are mainly transported from rivers and/or from direct effluents located on or near estuaries. Their mixing with salt water is often far beyond a simple dilution (behaving nonconservatively [62]). An increase of ionic strength from approx. 0 to about 0.7 mol/L, together with the change in water composition, has a wide array of influences on transport and transformation processes of contaminants in estuaries, including adsorption or desorption, coagulation, flocculation, and precipitation and biotic assimilation or excretion. Depending on the nature of the contaminant and on the estuarine condition, removal from water to sediments (sediments as a sink) or addition to water from sediments (sediments as a source) can occur during the mixing. Estuarine sediments are efficient and effective traps for hydrophobic chemicals mainly because of the salt-out effect [35,63]. Although desorption of metals from sediments can occur during estuarine mixing, release of metals to water is counteracted by enhanced flocculation [36]. Hence, estuarine sediments serve as a filter for many contaminants between land and sea [64]. Only those metals that form very strong complexes [65] and organic chemicals that are less hydrophobic [66] may be transported out of estuaries to the ocean. This natural mechanism renders estuaries more susceptive to contamination.

Grain size effects

Chemistry-based approaches for assessing sediment contamination are based on reliable measurement and interpretation of contaminant concentrations in the sediments. While the overlying water in estuaries can be heterogeneous because of different mixtures of fresh and saline water (Fig. 1), a much higher degree of heterogeneity and variability exists within estuarine sediments not only because of the salinity differences in the pore waters but also because of the diverse and complicated composition of the sediments. Different sediments can have significantly different capacities for collecting contaminants. For instance, the grain size distribution of a sediment is probably the most important factor controlling sediment metal concentrations; correlations commonly exist between decreasing grain size and increasing metal concentrations [67,68]. Hence, any assessment approach based on sediment metal concentrations needs to consider grain size differences between sediment samples. If such differences are not taken into account, apparent concentration differences between sites may reflect differences in grain size rather than the extent of contamination.

Numerous efforts have been made to normalize grain size effects relative to sediment contaminant concentrations. These efforts can be generally grouped as physical separations (measuring sediment contaminant concentrations in selected, physically separated grain size fractions of the sediment) and mathematical normalizations (measuring contaminant concentrations in whole sediments followed by normalizing to appropriate sediment constituents). In the physical separation method, the <63-μm fraction is favored by many authors [69,70]. However, problems with the efficiency and feasibility of mechanical separation in many cases (particularly when sample sizes are small) render mathematical normalizations the best way to correct for grain size effects [70].

Many sediment constituents have been used for normalization of grain size effects in estuarine sediments. These include grain size [71–73], clay minerals [74], surface area [75], and conservative elements, such as Cs [72], Al [76–78], Fe [79], Li [78], Rb [72], and Sc [72]. The normalization procedure generally involves calculation of ratios or regression analysis [67], with the latter being preferable [80]. Among regression analyses, robust regression, such as least-absolute-values regression, may be more powerful and less subjective than least-squares regression [80].

The most important factor controlling concentrations of hydrophobic organic chemicals in sediments, however, appears to be the fraction of organic carbon [81,82]. Normalization to the organic carbon fraction is hence necessary when comparing concentrations of hydrophobic chemicals in different estuarine sediments.

Background enrichment

Background enrichment assesses the relative contribution of anthropogenic sources of a substance by comparing its concentration at a site to its natural background concentration. This approach applies only to naturally occurring substances such as metals and metalloids. Since this approach does not take into account the biological effects of the substance, enrichment does not necessarily imply biological effects.

The simplest approach for assessing enrichment is a point comparison, which takes the form of the ratio of the sediment contaminant concentration at the site of concern to that at a reference site [83]. Prior to calculating the ratio, both concentrations need to be corrected for grain size effects unless the sediments are of similar composition. For such a point comparison to be meaningful, the geological setting of the reference site must be representative of the site of concern. Such a point comparison may be practicable, at best, only within small areas of fresh and marine sediments [84]; it is often not feasible in estuaries. Since the bulk of the sediment in an estuary can be derived from areas with different geology (e.g., the sediments in the upper reaches may be fluvial in origin, those in the lower estuary largely of marine provenance, and those between may comprise different mixtures of fluvial and marine sediments), it is often impossible to obtain a single representative reference site, even for a small area of the estuary.

Alternatively, a line comparison can be conducted for estuarine sediments. For example, in assessing the spatial scale of trace metal contamination in estuarine and coastal sediments in the United States, Daskalakis and O'Connor [85] developed baseline regression lines between concentrations of trace metals and a conservative element (Fe or Al) from sites relatively free of human influence. Concentrations of trace metals and the conservative element from other sites were then compared with the baseline regression lines to determine whether they were enriched.

While the enrichment approaches have no predictive power for biological effects, they fully consider background concentrations of naturally occurring substances and provide important insights into sources of contamination, both of which are often underestimated or ignored by other approaches. Delineation of the extent of metal enrichment may also be used as a screening tool to promote cost-effective use of sediment toxicity tests (e.g., only those with metal concentrations exceeding expected natural background ranges will be subjected to further toxicity tests [77]).

Bioavailability

The total concentration of a contaminant in sediments can be conveniently measured; however, it alone provides little information as to the possible biological effects of that contaminant [45]. Bioavailability of sediment-bound contaminants is determined by sediment constituents, overlying and interstitial water chemistry, and the behavior of organisms [78]. Main binding phases include organic carbon for hydrophobic chemicals [81] and sulfide, organic matter, and iron and manganese oxyhydroxyides for metals [45].

No consensus exists as to appropriate analytical methods for determining contaminant bioavailability in estuarine sediments [86]. Chemical extractions of sediments have been traditionally used. For example, metal concentrations in estuarine benthic invertebrates have been found correlated to dilute acid extractable metals [83,87], metal/Fe, or metal/organic matter ratios [83] in sediments. Schlekat et al. found that metals associated with bacterial exopolymer sediment coatings were assimilated by the estuarine amphipod Leptocheirus plumulosus with higher efficiency than those associated with recalcitrant organic carbon, iron oxides, or phytoplankton [40,88]. Recent studies, however, indicate that the bioavailability of sediment-bound contaminants is more closely related to the digestive systems of benthic animals. For example, in vitro digestive fluid extractions have been shown to provide a better indicator of the bioavailability of sediment contaminants than chemical extractions [89–93].

Equilibrium partitioning models have been postulated to estimate the bioavailability of contaminants in fresh and salt sediments, for example, the organic carbon sorption model for hydrophobic organic chemicals [81], the iron and manganese oxyhydroxyides and organic carbon model for metals in anoxic sediments [94], and the acid volatile sulfide/simultaneously extracted metals model for metals in anoxic sediments [95]. These models, however, may have little application to estuarine sediments since the very dynamic physical and biogeochemical nature of in situ estuarine sediments overturns the fundamental assumption (a quasi-equilibrium state being achieved between contaminants in sediments and in water) involved in these models. River flow [96], tidal flushing [97], and other sediment resuspension events [98] can have significant influences on the partitioning and bioavailability of contaminants in estuarine sediments. In addition, as noted earlier, salinity also significantly affects bioavailability.

Even if an equilibration of contaminants between sediments and interstitial water exists in estuaries, which is unlikely, bioavailability of sediment-bound contaminants is not necessarily attributable only to interstitial water. For example, exposure of many estuarine and coastal deposit feeders may occur principally through ingestion of particles, and hence metal bioaccumulation cannot be fully accounted for by the acid volatile sulfide/simultaneously extracted metals model [87]. Exposure of benthic organisms may also occur through bioirrigation of the overlying water, as demonstrated for many freshwater tube-dwellers [99,100] as well as estuarine benthic invertebrates [101].

Sediment quality values

Sediment quality values (SQVs) can be used to screen sediment contamination by comparing sediment contaminant concentrations with the corresponding SQVs. The feasibility and reliability of such comparisons are, of course, greatly dependent on the availability and reliability of SQVs. Deriving SQVs requires integrated information on sediment chemistry, toxicology, and biology [102]. While dozens of approaches have been employed for deriving SQVs for fresh and marine sediments [103], no specific efforts have been made to develop SQVs for estuarine sediments. Many jurisdictions classify aquatic ecosystems into either freshwater or marine ecosystems, ignoring the unique characteristics of estuaries. For example, databases for estuarine and saltwater sediments are often lumped to derive SQVs for marine sediments [104]. Similar problems exist with water quality guidelines (WQGs), which have been traditionally derived for fresh and marine waters. The importance of deriving specific WQGs (water quality guidelines) and SQVs for estuaries was recently recognized by Australian and New Zealand Environment and Conservation Council (ANZECC)&Agriculture and Resource Management Council of Australia and New Zealand (ARM-CANZ) [105] (http://www.environment.gov.au/science/water/index.html) in deriving WQGs for Australia and New Zealand. The WQGs for estuarine waters for physical and chemical stressors (e.g., dissolved oxygen, total nitrogen) were derived. However, no WQGs were derived for toxicants for estuarine waters because of insufficient data. This was identified as an area for further research [105].

Interstitial water chemistry

Measurements of interstitial water chemistry assess sediment contamination by measuring contaminant concentrations in the interstitial water rather than in the sediments, followed by a comparison of the measured interstitial water concentrations with WQGs or other toxicity thresholds. While these approaches effectively eliminate grain size effects associated with sediment chemistry, several key issues remain. First, as noted earlier, interstitial water chemistry may not fully account for biological effects on benthic organisms, as fractions of contaminants in sediment particles [87] and/or in the overlying water [99,100] may also be bioavailable to benthic organisms. Second, not all the dissolved contaminants in interstitial water are bioavailable. For example, the most bioavailable forms of contaminants in the interstitial waters are thought to be free metal ions for many metals [45,95] and free dissolved fractions (not complexed by dissolved organic carbon) for hydrophobic organic chemicals [81]. Above all, comparing estuarine interstitial water concentrations with saltwater or freshwater WQCs is, at best, questionable because of the differences in both chemistry and biology between the interstitial and overlying water as well as the differences between estuarine waters compared to salt and fresh waters. Salinity or, preferably, salinity profiles of interstitial water should always be measured and reported.

Furthermore, there is no consensus as to the best methods for sampling and analyzing estuarine interstitial waters. In situ interstitial water sampling techniques (such as in situ dialysis samplers or peepers) may not be feasible because of the very dynamic physical nature of estuaries. Laboratory interstitial water extraction (such as squeezing and centrifugation) substantially destroys natural gradients in redox potential, salinity, and other parameters. Instead, a combination of in situ and laboratory techniques is often employed. For example, field-collected sediment cores are brought to the laboratory, where, under a nitrogen atmosphere, gel probes [106] or multilevel suction samplers (or sippers [107]) are inserted to take interstitial water samples.

ASSESSMENT TECHNIQUES: BIOLOGICAL SURVEYS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. UNIQUE CHARACTERISTICS OF ESTUARIES
  5. Unique biological characteristics of estuaries
  6. SALINITY AS A CONTROLLING FACTOR
  7. ASSESSMENT TECHNIQUES: CHEMISTRY
  8. ASSESSMENT TECHNIQUES: BIOLOGICAL SURVEYS
  9. ASSESSMENT TECHNIQUES: TOXICITY TESTS
  10. SUMMARY AND CONCLUSIONS: THE WAY FORWARD
  11. Acknowledgements
  12. REFERENCES

Types and distribution of estuarine benthos

Estuaries typify the paradox of brackish water [108]. Specifically, although for most ecological factors the largest number of species occurs at intermediate values, this is not the case for salinity. In the case of salinity, the greatest number of species occurs in fresh and in marine waters, with fewer numbers of species at intermediate salinities (Fig. 4). This paradox appears to be due to the unstable and unpredictable behavior of estuarine environmental factors that decrease the probability of speciation and increase the probability of extinction (particularly for estuaries since they are geologically ephemeral phenomena) while excluding most marine and freshwater species [25,109]. Estuarine salinities are generally considered to be above about 1 and below about 30 g/L (freshwater organisms can generally survive, grow, and reproduce in a few g/L salinity; similarly, salinities of 30 g/L do not constitute a major hardship for most marine organisms). Estuarine benthos species richness or diversity are least within a critical salinity range of about 5 to 8 g/L, which reflects the inabilities of many organisms to tolerate salinity stress and to undergo extensive cell volume regulation [59].

thumbnail image

Figure Fig. 4.. Illustration of Remane's [108] “paradox of brackish water.” Species numbers and diversity are lower in estuarine than in fresh or marine waters.

Download figure to PowerPoint

Conflicting concepts exist regarding estuarine biological communities [110]. For instance, Remane [108] considered brackish waters as a distinct biological dominion, situated between marine and freshwater dominions, with a unique mix of species. In contrast, Barnes [111] denied the existence of an identifiable brackish assemblage and suggested that the estuarine environment is simply populated by a small group of marine species. Based on comparative analyses of very fine morphophysiological changes and genetic analyses, Cognetti and Maltagliati [110] recently concluded that the estuarine environment should not be considered as a marginal, transitional environment but rather as a single and well-defined habitat with its own fauna. They also concluded that in brackish waters, a given species of marine origin often consists of many different forms at various levels of differentiation.

Estuarine benthic species tend to be r-selected. They typically exhibit low species diversity, small body size, short life cycle, early reproduction, rapid development, variable population size, density-independent mortality, low competitive ability, and high reproduction [112]. K-strategy benthic species, however, can also be found in brackish environments [110].

Many estuarine benthic communities exhibit symptoms of disturbance without anthropogenic influence [113,114]. Under extreme physical and chemical stresses, such as unstable sediments and periodic changes of anoxic and oxic conditions, only a few species can survive. In quiet, muddy, organically rich environments, the benthos is typically dominated by small deposit feeders characteristic of polluted sites [12].

Estuaries contain a varied fauna, ranging from oligohaline (<5 g/L salinity) at the head to mesohaline (5–18 g/L salinity) to euhaline at the mouth (>18 g/L salinity). Crustacea, Mollusca, and Polychaeta are well represented, as are opportunistic species such as Mulinia lateralis, Capitalla capitata, and Polydora ligni, whose densities fluctuate greatly [59]. Truly estuarine species are generally restricted to between about 3 and 20 g/L salinity, with penetration of estuarine sediments by both freshwater and euryhaline opportunists (Fig. 4). As noted previously, salinity is the most important natural factor controlling the distribution of estuarine organisms. However, organisms burying into muddy sediments are directly affected not by overlying water salinities but rather by interstitial salinities, which change much more slowly in muddy than in sandy sediments ([32]; Fig. 3). In salt wedge estuaries, seasonal variations in interstitial salinities occur, resulting in seasonal shifts in benthic infaunal distributions. Basically, the range of oligohaline, mesohaline, and euhaline species in salt wedge estuaries varies seasonally, with oligohaline populations extending their range seaward during periods of high freshwater flow and low tidal influence and the reverse occurring during periods of low freshwater flow and high tidal influence related to both tidal and riverine transport [115]. These species shifts have been shown to cover distances exceeding 10 km in at least one salt wedge estuary ([11,31]; Fig. 2). Populations involved in such species shifts do not reproduce throughout their entire estuarine range but rather are sustained by periodic recruitment (drift or migration, depending on the species) from centers of abundance to the extremes of the population's estuarine range. Such species shifts are dependent on the availability of larvae, juveniles, or adults in the water column; suitable conditions for settlement (low current velocities and substrate); and suitable conditions for continued survival and residence [12]. The latter requirements include suitable salinity, habitat, and low bottom shear stress. Depending on species-specific salinity tolerances, in some cases immature organisms may not survive in areas where both mature and immature organisms are deposited.

Because of these species shifts in salt wedge estuaries, usage of habitat by different species is actually greater than appears to be the case from snapshot estimates. Thus, in fact, on a habitat usage basis, numbers of species in these estuarine sediments are not as low relative to fresh and marine waters as described by Remane ([108]; Fig. 4).

Hypoxia is also a feature of some estuaries, related to circulation of bottom waters, which are primarily saline [116,117]. Effects of hypoxia on estuarine benthos, independent of other stresses (e.g., variable salinity, anthropogenic inputs), will change community composition and reduce diversity and biomass [117–119].

Biological surveys of estuarine benthos

Benthic organisms are relatively sedentary (avoidance responses are limited), have relatively long life spans (indicate and integrate conditions), comprise different species or functional groups with different tolerances to stress, can be commercially important or are important food sources for economically or commercially important species, and tend to have an important role in cycling nutrients and contaminants between the sediments and water column [120]. Thus, a great deal of effort has been spent on evaluating benthic community structure in estuaries in relation to anthropogenic contamination as well as natural factors. A focus has been attempts to develop benthic community indices particularly related to sediment quality values and sediment toxicity test results [120–126]. However, most authors of studies developing indices acknowledge that these are primarily management tools, not definitive assessment tools.

As previously noted, estuaries tend to be unique physically and chemically. They also tend to have relatively low species numbers and diversity compared to fresh or marine waters (Fig. 4), and differences between interstitial and overlying waters within salt wedge estuaries result in seasonal movements up-and downstream of benthic fauna. These realities render biological surveys in truly estuarine sediments particularly difficult for several reasons. First, reference sites may be so different from exposed sites as to render comparisons meaningless, and gradient approaches must deal with salinity differences as well as the usual confounding factors (e.g., sediment grain size, total organic carbon [TOC]). Second, seasonal variations related to freshwater discharge and salinity intrusions render the benthos more variable than in consistently marine or freshwater sediments.

Thus, it is not surprising that Hyland et al. [126], in attempting to predict stress to benthic communities in southeastern U.S. estuaries, found a correlation between degraded benthos and lower salinity (as well as to higher contaminant levels, muddier sediments, higher TOC levels, and slightly greater depths). Hyland et al. [126] also report a lack of concordance between sediment toxicity tests using marine organisms at marine salinities and estuarine benthos living at much lower salinities. It is entirely possible that concordance would have improved if estuarine sediments had been tested at their in situ salinities, using estuarine organisms. Similarly, estimates made by Long et al. [127,128] of the spatial extent of sediment toxicity in U.S. estuaries might have been very different if the sediments had not been tested at marine salinities using marine organisms (see Assessment Techniques: Toxicity Tests). Detailed, extensive toxicity testing of San Francisco Bay sediments observed wet-dry seasonal fluctuations in both sediment contamination and toxicity, which the authors ascribed to estuarine processes [129].

Ecological stress, from any source, is best measured using multiple variables, methods, or analyses and not necessarily limiting these to the same or a few assumptions. Particularly useful are combinations of sediment toxicity, contamination, and estuarine community structure [130–132], and in estuarine sediments, natural variability in salinity must, in contrast to freshwater or marine sediments, be considered and accounted for. This has been done by some authors [123–125,133], but not all. A further problem is that the tolerances of many estuarine organisms to contaminants are relatively unknown as compared to fresh and marine ecosystems.

Finally, sufficient data are required to adequately determine different habitats and to determine the status, separately, of each of these habitats, which, of course, are specific to individual estuaries. For instance, Weisberg et al. [124] defined seven habitats in Chesapeake Bay (MD, USA) based on salinity and substrate. In some estuaries, waves, currents, and depth are also important factors defining benthos habitats; in all estuaries, biotic interactions (competition, predation) are important [59]. Given the complexity and diversity of estuaries, accurate and precise generic indices are difficult without accounting for habitat differences. This requires very large data and information sets that are not presently available for many estuaries. Thus, on a broad scale, only relatively simplistic indices are possible (e.g., northern Gulf of Mexico estuaries [125]). Such indices cannot distinguish between natural and anthropogenic stresses. For instance, they cannot by themselves distinguish hypoxia that is anthropogenic in origin from natural hypoxic events.

ASSESSMENT TECHNIQUES: TOXICITY TESTS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. UNIQUE CHARACTERISTICS OF ESTUARIES
  5. Unique biological characteristics of estuaries
  6. SALINITY AS A CONTROLLING FACTOR
  7. ASSESSMENT TECHNIQUES: CHEMISTRY
  8. ASSESSMENT TECHNIQUES: BIOLOGICAL SURVEYS
  9. ASSESSMENT TECHNIQUES: TOXICITY TESTS
  10. SUMMARY AND CONCLUSIONS: THE WAY FORWARD
  11. Acknowledgements
  12. REFERENCES

Sediment toxicity tests can be conducted using a variety of exposure techniques: whole sediments, pore (interstitial) waters, elutriates, and extracts. Whole sediment exposures allow for the widest variety of possible exposure routes and result in the least changes to sediment physicochemical conditions. They also comprise the primary tool for sediment toxicity assessments [134,135]. Thus, this section focuses on whole sediment toxicity tests.

Single-species whole sediment toxicity tests

Estuarine sediments vary considerably with regard to salinity (interstitial and overlying), as previously noted. They also vary considerably with regard to particle size, temperature, and TOC content. In some cases, these differences have been ignored, and sediment toxicity has been evaluated using marine species exposed to interstitial or elutriate waters [135–138] or to sediments whose salinity has been artificially increased [126–128,139]. In other cases, freshwater species have been exposed to estuarine sediments overlain with freshwater [56]. In all these cases, potential differences in contaminant bioavailability and toxicity due to the original, unadjusted interstitial salinities (see Assessment Techniques: Chemistry) have not been accounted for. Further, such manipulations tend to dilute the contaminant load. The only reasonable way to determine the toxicity and bioavailability of contaminants in estuarine sediments is to test the sediments as received, using estuarine organisms capable of tolerating the full range of estuarine conditions, in particular salinity.

Whether estuarine species in general are more or less tolerant than freshwater or marine species is a topic of debate among ecologists. It has been suggested [140,141] that because estuarine species live near the limit of their tolerance range, they are more likely to be susceptible to any additional stress. It has also been suggested [142] that because estuarine species must have a high tolerance to abiotic factors such as salinity, they may be preadapted to tolerate pollution stress. To date, spiked sediment, contaminated sediment, and water-only toxicity tests indicate that estuarine species can be as sensitive as freshwater or marine species [143,144] and presumably also as variable in their relative sensitivities to different toxicants.

Single-species estuarine whole sediment toxicity tests can be divided into two basic categories: tests not originally intended for but adapted for estuarine conditions (first generation) and tests designed specifically for estuarine conditions (second generation). The latter tests in particular include both organisms tolerant of the entire range of estuarine salinities (from fresh to oceanic) as well as organisms that can tolerate only a limited range of estuarine salinities. Single-species tests for the toxicity of whole estuarine sediments are listed in Table 2.

Some interesting and puzzling observations can be made based on Table 2. In particular, the importance of salinity in conducting sediment toxicity tests is not fully recognized. There are few whole organism tests presently in use for estuarine whole sediment toxicity tests that can survive the full range of estuarine salinities. There are none that can successfully reproduce across the full estuarine salinity range. In addition, it appears that not all authors appreciate the importance of detailing their test species' tolerances to salinity (or to other modifying factors such as sediment grain size or TOC). This is particularly the case for first-generation tests adapted from fresh and marine sediment tests in which interstitial salinities would normally be ignored. For estuarine sediments, interstitial salinities are sometimes measured and adjusted through addition of higher or lower salinity water either directly (mixing the sediments in different salinity waters) or indirectly (changing the salinity of overlying waters) [133,145]. Prolonged exposure to overlying waters of different salinity will change the salinity of interstitial waters even in low permeability muds (Fig. 3). Salinity variations can also change the bioavailability and route of exposure of contaminants in those sediments (see Unique Physicochemical Characteristics of Estuaries). To our knowledge, there have been no publications in which the effects of adjusting interstitial salinities on bio-availability and toxicity have been directly assessed (though there have been such studies done in water-only situations). Similarly, there have been few publications of studies attempting to maintain similar interstitial and overlying water exposures in the laboratory as found in situ. Instead, there are arguments presented for testing under defined conditions, including salinity [56,146,147].

Table Table 2.. Estuarine Sediment Toxicity Test Organisms. Single-species, whole sediment exposures
Test organismTest end point(s)Test generationaSalinity range (survival, g/L)Salinity range (reproduction, g/L)References
  1. aTest generation-first: techniques and/or organisms originally developed for saline or freshwaters; second: techniques and/or organisms specifically developed for estuarine conditions.

Fish (spot) Leiostomus xanthurusSurvival, fin erosion, lesionsSecondEstuarine; collected at 20, held and tested at 15[221,222]
Amphipod (crustacean) Hyalella aztecaSurvival, reproductionFirst0–17<10[56,223,224]
Amphipod (crustacean) Eohaustorius estuariusSurvival, avoidanceSecond0–34Unknown[143,146,223,225]
Amphipod (crustacean) Gammarus duebeniPleopod beat frequency, swimming enduranceSecond“A truly estuarine species”; held and tested at 15[226,227]
Amphipod (crustacean) Leptocheirus plumulosusSurvival, growth, reproductionSecond0–335–20[146,152,153,223,228]
Amphipod (crustacean) Lepidoctylus dytiscusSurvival, growthSecondNo information; held and tested at 15[155]
Amphipod (crustacean) Melita nitidaSurvival, reproduction, abnormal brood pouch setae, intermolt periodSecondTypically found in salinities at 3 to 20[229]
Amphipod (crustacean) Corophium mutisetosumMortality, growthSecond0–351–15[230]
Amphipod (crustacean) C. sp.SurvivalSecond0.1–241–15 +[144]
Mysid (crustacean) Mysidopsis bahiaMortalityFirstNo information; held and tested at 21[139]
Waterflea (crustacean) Daphnia magnaMortality, reproductionFirst0-<50-<5[231]
Grass shrimp (crustacean) Palaemonetes pugioMortalityFirstNo information; tested at 20[232]
Copepod (crustacean) Amphiascus tenuiremisMortality, reproductionFirstNo information; tested at 28[205,232]
Polychaete Streblospio benedictiSurvival, growthSecondNo information; held and tested at 15[155]
Mollusk Tapes semidecussatusSurvival, behaviorSecondDescribed as “true estuarine species” but held and tested at 24[233]
Mollusk Scrobicularia planaSurvival, behaviorSecond   
Mollusk Mya arenariaBurrowing speedSecondTested at overlying water salinities varying from 2 to 32[234,235]
Mollusk Crassostrea gigasPediveliger larvae survival, settlement and metamorphosisSecondSurvival: >5; metamorphosis: >23[199]
Bacteria MicrotoxEnzyme functionFirst0–35+0–35 +[135,173,236,237]

Given that interstitial salinities can be very different from overlying salinities for truly estuarine sediments, both should always be measured prior to testing (during sample collection). They should also be measured during testing. Such measurements are important not only in terms of test organism sensitivities to salinity but also in terms of the potential effects of salinity on bioavailability (or not) of sediment contaminants. Whereas pore-water collections for chemical contaminants pose difficulties [148,149], pore water for salinity analyses can be adequately collected relatively simply by squeezing or centrifuging sediments [31,32], though this does effectively destroy natural salinity gradients in the sediment.

Ideally, estuarine testing should be conducted with a reasonable knowledge of the estuary in question and in particular of the potential range of seasonal interstitial salinities at all test sites. Again ideally, testing should cover the range from worst to best cases related to salinity effects on contaminant bioavailability. At the very least, the effects of varying salinities on bioavailability need to be considered when interpreting testing done at a specific salinity (a snapshot in time of a dynamic and unique system). McGee et al. [150] measured interstitial salinities and provided matching overlying water salinities in Leptocheirus plumulosus tests of Baltimore Harbor (MD, USA) sediments. This is one of the only studies of estuarine sediment toxicity to have taken the approach of measuring and matching natural salinity conditions rather than ignoring them and/or imposing arbitrary salinity conditions.

Another interesting observation from Table 2 is that sediment tests of estuarine sediments have, to date, been heavily weighted to crustaceans, in particular, amphipods. Of the 19 different estuarine test species in Table 2, there are 12 crustaceans (of which eight are amphipods) and a fish, a poly-chaete, four mollusks, and a bacterium. Other organisms reported in the literature as used in whole sediment toxicity tests with estuarine sediments are marine rather than estuarine and require marine salinities during testing (e.g., the polychaete Arenicola marina [135]). Phylogenetically, the list of available test species is more limited than should be the case. It also comprises few examples of true deposit feeders, such as the mollusks Tapes semidecussatus and Scrobicularia plana. Further, estuarine whole sediment toxicity tests appear to be better developed as follows: North America > Europe > Australia > rest of the world.

DeWitt et al. [143] note that, ideally, estuarine test organisms should live at or below the sediment/water interface and have broad salinity tolerance, high sensitivity to sediment contamination, low mortality under reference conditions, low sensitivity to modifying factors such as sediment particle size and TOC, and a broad geographic range. They should also be amenable to handling in the laboratory, be ecologically important in estuarine systems, and either be readily collected from the field or amenable to laboratory culture. DeWitt et al. [143] favor use of amphipods as test organisms because of their relative sensitivity, as a group, to toxicants. The majority of estuarine sediment toxicity test organisms used to date are in fact amphipods (Table 2).

The amphipod tests are the most widely applied toxicity tests performed with estuarine sediments in North America [127,128]. Two of the most commonly used estuarine amphipods in such tests are Eohaustorius estuarius (free burrowing) and L. plumulosus (open tube dweller) [146,147]. These amphipods are found, respectively, on the west and east coasts of North America. Both species can be collected in large numbers from the field. Ten-day acute toxicity tests with these two species provide reasonable discrimination of contaminated sediments except where sediment contamination is relatively low, and this problem can be minimized by increasing the number of test replicates [151]. However, to date only L. plumulosus is readily amenable to laboratory culture such that, together with its relatively short life history (30–40 d compared to annual for E. estuarius), it can be used in chronic testing focused on growth and reproductive endpoints [152,153].

In addition, E. estuarius is relatively insensitive to copper [154]. Whether it is insensitive to other metals or organic contaminants is unknown. Such is also unknown for many contaminants for all sediment test organisms (freshwater, estuarine, or marine). Investigators need to ensure that test organisms are appropriately sensitive to contaminants of concern in sediments if meaningful results are to be derived.

Not all possible amphipods or other estuarine test species have been used for toxicity tests with estuarine sediments. For instance, DeWitt et al. [143] suggested that four haustoriid amphipods be considered as good candidates for sediment toxicity tests in addition to E. estuarius: Haustorius canadensis, Neohaustorius schmitzi, Lepidactylus dytiscus, and Lepidactylus triarticulatus. Of these, to our knowledge, only L. dytiscus has been used for estuarine sediment toxicity tests, and only in Chesapeake Bay [155], and this amphipod is now no longer being used in Chesapeake Bay because of its preference for sandy rather than muddy sediments (J. Winfield, personal communication). Similarly, although amphipods of the genus Corophium are typically estuarine, they have not been widely used for estuarine sediment toxicity tests. The type species C. volutator, which has a wide geographic distribution and a wide tolerance to salinity (tolerates 2–59 g/L, prefers 10–30 g/L [156,157], has been used primarily in marine sediment toxicity tests [158]. C. spinicorne has similarly been used in marine sediment toxicity tests [159]. C. triaenonyx, which tolerates the full range of estuarine conditions, though reproduction is limited to 7.5 to 37.5 g/L [160], has not been used in estuarine (or marine) sediment toxicity tests. C. arenarium has been used in tests with sediments the authors referred to as estuarine but whose salinities were marine (generally 30–40 g/L [135]). Corophium are tube builders, living in U-shaped burrows that arguably may not result in direct exposure to sediment interstitial waters but rather to overlying waters [99]. However, the significance of this fact remains to be evaluated. The widely used estuarine sediment test species Leptocheirus plumulosus lives in open tubes, as does the less widely used spionid polychaete Streblospio benedicti. This latter species has also been used more commonly in marine sediment bioassays than in estuarine sediment bioassays; in the former but not the latter, end points included reproduction [161]. However, the utility of this polychaete is arguable since it is considered more tolerant of contaminated sediments than many other species [162].

It should be noted that two other amphipods, Rhepoxynius abronius and Ampelisca abdita, while often used in estuarine sediment toxicity tests [126,127,146,147,163], are indeed marine or euryhaline. The former is found where in situ salinity is higher than 20 g/L [163], whereas the latter is found where in situ salinity is higher than 10 g/L [164]. Caution is advised when using results from such tests for assessing estuarine sediment contamination.

Chironomidae are not included among potential estuarine test species. This is a surprising omission since a variety of freshwater chironomids live in natural salinity gradients ranging from 0 to almost 10 g/L [165] and since intertidal species can be naturally exposed to and tolerate the full gamut of estuarine conditions (0 to ≥20 g/L salinity; P.M. Chapman, unpublished data). Further, tests with representatives of this group of organisms are well developed [166].

Similarly, it is also surprising that aquatic oligochaetes are not included among potential estuarine test species. Acute and chronic test procedures with freshwater species such as Tubifex tubifex are well developed [167,168], and this and other freshwater oligochaete species can survive in salinities up to about 10 g/L [169,170]. Estuarine oligochaete species are amenable to acute toxicity testing and can survive across the full gamut of estuarine salinities [170,171]. Clearly there are opportunities to develop estuarine sediment toxicity tests using both chironomids and oligochaetes.

As shown in Table 2, there is only one fish test included. However, this may be an artifact arising from authors not distinguishing whether test sediments are estuarine or marine. For example, work by Nagler and Cyr [172] showing reduced hatching success for American plaice exposed to contaminated sediments involved an estuarine fish and estuarine sediments. However, no information was provided on salinities other than that the sediments were considered marine by the authors. Also, testing of such mobile species must distinguish whether these organisms are likely residents or simply periodic exploiters entering the estuary from outside (e.g., with salt wedge intrusions). Exposure to estuarine contamination will be very different for periodic exploiters than for residents.

Testing using a toxicity test kit, such as Microtox® (Azur Environmental, Carlsbad, CA, USA), clearly offers some advantages in the estuarine environment. For instance, these tests are not directly affected by salinity. However, they are affected by other factors; for instance, the Microtox (Azur) solid-phase test shows decreased toxicity in finer sediments because the test organisms (bacteria) are adsorbed to silt-clay particles [173]. It is to be expected that other toxicity test kits will be applied to estuarine sediments. For instance, the Mutatox™ bioassay (Azur) has not yet been applied to whole sediments but has been applied to fractionated sediments [174]. The relevance of data from such tests to estuarine species and populations is arguable.

Among organisms being developed for sediment toxicity testing are some (e.g., the amphipod Gammarus locusta; [175]) that are abundant in and collected from estuarine sediments, yet the test developed with these organisms is marine. Whether these organisms can also be tested in more estuarine salinities should be determined; the greater the salinity range of test organisms available for estuarine sediments, the better.

A final interesting observation from Table 2 is that a relatively high proportion of estuarine sediment behavioral tests exist (5 of 19 tests). In fact, there may be more estuarine sediment behavioral tests available than noted because some authors are not specific as to the salinity of their test sediments. For instance, burrowing behaviors have been studied for intertidal sediments by some authors but without providing any information as to salinities [176,177].

Potential confounding factors

Ammonia. Ammonia has been implicated as a source of toxicity in some sediment toxicity tests [178–180] and has been shown to interfere in 10-d but not 28-d tests with L. plumulosus [181]. Ammonia in sediments can originate from anthropogenic effluents and wastes but can also originate from natural decomposition processes. Ammonia toxicity is influenced by both pH and hardness/salinity [182–184]. Thus, in estuarine sediments, if interstitial salinities change, it is likely that the risk of ammonia toxicity will also be affected. At this time, no broad generalizations are possible, particularly given differential sensitivities to ammonia by, for instance, different amphipod species [185]. Ammonia can also produce analogous effects with other contaminants such as silver and copper on nitrogen metabolism of freshwater and marine fish [186,187].

Sulfide. While the confounding factor of ammonia in sediment toxicity tests has been recognized by many studies, the role of sulfide in determining sediment toxicity has only recently been fully addressed [188]. Sulfide influences sediment toxicity in three major ways [188]: as a toxicant in its own right, by reducing metal toxicity by forming metal sulfide solids and/or complexes, and by affecting animal behavior, which in turn can alter the toxicity of not only sulfide but also other sediment contaminants. Estuarine sediments, particularly those in stratified estuaries such as salt wedge estuaries and fjords, are particularly rich in sulfide because of the high influxes of both organic matter (from river water) and sulfate (from sea-water). Sulfide levels as high as 100 mg/L are common in estuarine sediments [189,190]. Thus, any estuarine sediment toxicity tests that do not adequately consider sulfide effects risk misestimating toxicity and misidentifying the causative agents [188].

Grain size, organic matter, and other physicochemical characteristics. Sediment physicochemical characteristics such as grain size and organic matter, in addition to modifying the bioavailability of contaminants (see Assessment Techniques: Chemistry), can also affect some organisms in sediment toxicity tests. For example, the amphipod Rhepoxynius abronius, often used in marine sediment toxicity tests, may suffocate when attempting to burrow in fine-grained sediments [191]. Lacey et al. [192] showed that the growth of Chironomus tentans, often used in freshwater sediment toxicity tests, can be significantly affected by both the quantity and the quality of organic matter. To reduce the impacts of these confounding factors, the selection of an appropriate control sediment for sediment toxicity tests is critical. Ideally, the control sediment should be identical to the test sediment in grain size distribution, organic matter content, and other physicochemical characteristics [192]. This is, however, almost impossible. Instead, development and usage of formulated reference sediments may assist in reducing the influence of these factors on sediment toxicity test results [192–194].

Community-level toxicity tests

In addition to single-species whole sediment toxicity tests, community-level toxicity tests have also been developed for estuarine sediments. Some of the earliest such tests were conducted by Tagatz and coworkers [195,196], who developed a laboratory method for exposing previously frozen sediments to pelagic larvae and measuring differences in colonization patterns.

Austen and Somerfield [197] tested previously frozen contaminated field sediments using a simple laboratory microcosm system (570-ml glass bottles) with addition of meiofauna-rich sediment at 20-g/L salinity. After a two-month exposure period, meiofaunal community structure was determined and compared to natural assemblages. These approaches are based on the hypothesis that settlement and postsettlement stages of free-swimming benthic larvae are the most vulnerable life-cycle stages and most likely to be affected by anthropogenic pollution [198,199].

The possible implications on sediment structure and function of freezing sediment to kill resident organisms remains to be fully assessed. Thus, Watzin et al. ([200]; see the following discussion) chose to use artificial sediments in their spiking experiments.

Watzin et al. [200] exposed Zn-spiked artificial sediment for one week in the field in 100-cm2 containers and used the abundance and diversity (species richness) of colonizing species as indicators of sediment quality. In subsequent work, Watzin and Roscigno [201] applied this technique to two sites with differing overlying water salinities (5 and 15 g/L) and recorded resultant changes to the benthos, However, these authors were not able to separate out effects of site from effects of salinity. Hall and Frid [202] undertook a two-year microcosm experiment in the lower Tyne Estuary (UK), dosing the sediments with copper and then determining faunal effects and patterns of recovery. Interestingly, although chemical recovery was rapid, faunal recovery was delayed because of differential recolonization rates. Millward and Grant [203] used pollution-induced community tolerance (PICT) to evaluate the biological impact of chronic anthropogenic copper exposures on another UK estuary. Inherited tolerance to copper had previously been shown in this estuary for the polychaete Nereis diversicolor [204].

Chandler et al. [205] tested intact 1,750-cm3 flow-through sediment microcosms containing their original meiobenthic fauna (primarily copepods) and also conducted spiked sediment tests with chlorpyrifos for 21 d. They argue that such tests are more conservative, informative, and predictive than using single-species sediment toxicity tests in isolation and argue further that both approaches should be used, collecting and culturing taxa and microcosms from actual field sites of concern.

Kurtz et al. [206] also used microcosms of estuarine sediments during 7-d measurements. However, their emphasis was on microbial communities. They point out that, despite the importance of bacteria to aquatic systems and processes, there is a lack of procedures to test bacterial responses to contaminated estuarine sediments.

These experimental approaches retain more environmental realism than single-species toxicity tests while still permitting manipulation in either laboratory or field settings. However, as noted by Chandler et al. [205], they are best interpreted in combination with well-designed single-species sediment toxicity tests.

Laboratory-to-field comparisons

The laboratory does not and cannot mimic the field situation but rather stands alone as an assessment tool [207]. In the case of whole effluent toxicity (WET) tests, laboratory results can be overprotective, underprotective, or of an unknown degree of protection compared to the field situation [208]. The same is probably true of estuarine sediment toxicity tests. For example, McGee et al. [209] found that field-collected L. plu-mulosus were typically more sensitive to cadmium than laboratory animals, although their sensitivities varied seasonally. However, in a comparison between L. plumulosus sediment toxicity test results and populations of these amphipods in Baltimore Harbor, Maryland, USA, McGee et al. [150] found that test results were predictive of population-level effects. Similarly, Swartz et al. [210] found that 10-d acute toxicity tests with E. estuarius provided reliable evidence of biologically adverse sediment contamination in the field. McGee [211] has attempted to integrate laboratory sediment toxicity data for L. plumulosus into a field-based population model to project population-level consequences of sediment contamination. Further development of such a model, including other test species, would greatly progress assessment of estuarine sediment contamination.

Difficulties in predicting what will actually occur in field situations are exacerbated by interactions between biotic and chemical stresses. For instance, Linke-Gamenick et al. [212] investigated the combined effects of toxicant stress (fluoran-thene) and density dependence to the polychaete Capitella sp. M in a 134-d life-table response experiment. They found that food limitations did not affect population dynamics at low levels of toxicant stress but that at higher toxicant exposures toxicant effects were exacerbated. They suggest that given that food limitation in nature is the norm (whereas in the laboratory food is generally provided in abundance), the maximum toxicant concentration at which populations can persist is less than previously believed. This experiment appears to refute the hypothesis proposed by Calow et al. [213] and Grant [214] that conditions of high population density buffer negative toxicant effects on population dynamics. However, whether the responses of Capitella sp. M populations are typical remains to be determined.

SUMMARY AND CONCLUSIONS: THE WAY FORWARD

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. UNIQUE CHARACTERISTICS OF ESTUARIES
  5. Unique biological characteristics of estuaries
  6. SALINITY AS A CONTROLLING FACTOR
  7. ASSESSMENT TECHNIQUES: CHEMISTRY
  8. ASSESSMENT TECHNIQUES: BIOLOGICAL SURVEYS
  9. ASSESSMENT TECHNIQUES: TOXICITY TESTS
  10. SUMMARY AND CONCLUSIONS: THE WAY FORWARD
  11. Acknowledgements
  12. REFERENCES

Specific concluding comments are provided followed by recommendations.

General comments

Estuaries are complex, dynamic, unique, and very individual tidally influenced environments (change is the norm). They are defined primarily by their variable salinities, both overlying and interstitial, and include fresh, marine, and truly estuarine fauna. In the context of geological time, they are also ephemeral phenomena. This may partially account for the fact that estuaries, in contrast to other intermediate ecological environments, have fewer numbers of species than their bordering environments (fresh and salt water). Estuarine sediments originate from fresh or marine waters but are not and cannot be treated as fresh or marine sediments. Estuarine physicochemical gradients (longitudinal, lateral, and vertical; salinity, temperature, pH, dissolved oxygen, redox potential, amount and composition of particles) tend to be strong and to influence estuarine biogeochemical processes, including contaminant binding and release from sediments. Salinity is the controlling factor for partitioning of contaminants between sediments and overlying or interstitial water and affects contaminant bio-availability.

Variable estuarine salinities provide stressful conditions for resident fauna whose survival strategies include avoidance, reduced contact, adaptation, and acclimation. Interstitial salinities tend to differ from overlying salinities, particularly in salt wedge estuaries, where seasonal, large-scale species shifts occur. Specifically, freshwater species extend their range downstream during periods when freshwater flow dominates over tidal intrusions; saltwater species extend their range upstream during periods when tidal intrusions dominate over freshwater flow. Such species shifts can occur over longitudinal distances on the order of 10 km.

Recommendations. All assessment techniques need to consider the unique and complex dynamic processes that characterize that estuary, in particular salinity effects. At a minimum, interstitial and overlying salinities need to be measured during sediment collection, and data interpretation should occur in the context of knowledge of seasonal and spatial variations in both interstitial and overlying salinities as well as other key estuarine physicochemical processes.

Chemical assessments

Chemical assessment techniques need to consider grain size effects, normalizing these relative to sediment contaminant concentrations. Mathematical normalizations are preferred over physical separations. Other normalizations for grain size effects are possible (e.g., surface area, organic carbon, conservative elements) and may be useful in specific situations.

Background enrichment chemical assessments apply only to naturally occurring substances such as metals and metalloids. Although not predictive for biological effects, they are useful for assessing contaminant sources. In this regard, point comparisons are rarely possibly in estuaries because of the difficulty in obtaining a representative reference site. Regression line comparisons are the most appropriate approach.

There is no consensus as to the most appropriate methods to determine bioavailability in estuarine sediments. In vivo digestive fluid extractions offer advantages over traditional chemical extractions. Bioavailability models may not be applicable to estuarine sediments because the fundamental assumption of a quasi-equilibrium state is not valid. Such models include equilibrium partitioning, the acid volatile sulfide/simultaneously extracted metals, organic carbon, and iron and manganese oxyhydroxyides.

Sediment quality guideline values have been developed for fresh and marine waters. They have not been specifically developed for estuarine waters. It is inappropriate to apply fresh or marine values to estuaries. Interstitial water collection and analysis for chemical contaminants is of questionable use in estuaries for reasons including inappropriate comparisons to fresh and marine water quality values and a lack of consensus on methods for collection.

Recommendations. Sediment chemistry should be normalized at least for grain size effects using mathematical techniques. For naturally occurring substances, regression lines should be used to correct for background enrichment. Models and approaches specific to estuaries need to be developed to assess bioavailability. Similarly, sediment quality values specific to estuaries need to be developed (and used only for screening). The applicability and relevance of interstitial water contaminant analyses in estuaries need to be determined.

Biological surveys

Much greater numbers of species occur in fresh and marine waters than in estuarine waters. Estuarine species tend to be r-selected with two critical salinity ranges: about 5 to 8 g/L and about 15 to 20 g/L. Sediment infauna are not affected by overlying salinities but rather by interstitial salinities, which are only slowly affected by overlying salinities in the case of muddy sediments. As a result, salt wedge estuaries in particular typically exhibit large-scale seasonal movements of fresh and marine species up and down the estuary. Biological surveys within about 3 to 20 g/L are difficult because of such factors; the variable nature of the benthos makes reference comparisons difficult if not impossible.

Recommendations. Investigators conducting biological surveys need to understand their estuaries' basic physicochemical characteristics and the possible longitudinal, lateral, and seasonal variations that can be typical of estuarine benthic infaunal populations. Habitat differences need to be understood and documented; in particular, overlying and interstitial salinities should be measured when samples are collected and then reported. In many cases, gradient approaches will be more realistic than reference comparisons. Simple estuarine benthic indices can be useful for initial assessments; however, definitive assessments are best done using the sediment quality triad, which consists of chemical analyses, toxicity tests, and metrics of benthic structure [131,215,216]. When constructing sediment quality triad triaxial plots for estuarine sediments, Del Valls et al. [217,218] noticed that the benthic community alteration axis can be variable and indeterminate, possibly because of salinity stress, emphasizing the importance of interstitial and overlying salinity measurements in estuaries.

Toxicity tests

Single-species whole sediment estuarine toxicity tests can be either first generation (adapted for estuaries) or second generation (designed for estuaries). Although there appear to be more of the latter generation tests available, there are few tests that span a large range of salinities. In too many cases, investigators have ignored salinity differences, conducting testing of estuarine sediments in fresh or marine waters with corresponding nonestuarine organisms. Changing salinity can change the bioavailability and route of exposure of sediment contaminants.

Laboratory toxicity tests involve a snapshot in time. As such, they cannot adequately predict in situ estuarine toxicity where major salinity variations occur in the sediments, for instance, in the mesohaline portions of salt wedge estuaries.

Estuarine single-species tests to date have been heavily weighted to crustaceans, particularly amphipods. There is not good taxonomic representation of estuarine fauna among estuarine test species; some obvious candidate test species have not been used. Further, the relative sensitivity of test species to particular contaminants is relatively unknown.

In addition to salinity, potential confounding factors need to be fully considered. Such factors include ammonia, sulfide, grain size effects, and TOC.

Community-level toxicity tests have been developed. These range from laboratory to in situ exposures and are arguably more realistic than single-species tests. Laboratory single-species tests are least predictive of field responses.

Recommendations. Estuarine sediments (especially over the range of about 3–20 g/L salinity) need to be tested as received, measuring overlying and interstitial salinities during sample collection and testing, and using estuarine organisms capable of tolerating those conditions. Studies should be conducted to determine the effects of changing interstitial salinities on contaminant bioavailability and toxicity relative to realistic toxicity testing for estuaries that experience large-scale seasonal salinity changes. A wider range of second-generation test species that are truly estuarine is required, and community-level testing should be further developed. The relative tolerances of new and existing test organisms to contaminants need to be determined. Most important, toxicity test approaches need to be integrated into predictive population models.

The way forward

Within estuaries in particular, subsurface sediment contamination may not be isolated from biological organisms because of the dynamic nature of estuarine processes. There is a clear need to tailor chemical, biological, and toxicological assessment techniques specifically for estuarine environments; fresh and marine water techniques are not generally applicable to estuaries. However, this will not be possible immediately. Less specific and realistic methods will continue to be used, if only for pragmatic reasons, such as cost. However, measuring interstitial and overlying salinities during collection of any and all samples (chemical, biological, toxicological) is neither difficult nor expensive and should be immediately implemented.

Where detailed, specific assessments (e.g., use of the sediment quality triad, testing with estuarine fauna without changing interstitial salinities) are not possible, a tiered approach should be followed based on the ecological risk assessment framework for sediments [219]. Simple techniques, such as benthic indices, toxicity tests with marine or freshwater species, and altered salinities, should be considered solely an initial assessment. Such approaches would be equivalent to the problem formulation stage of an ecological risk assessment. Depending on the level of detail [220], more detailed studies could be equivalent to either a screening level or a detailed ecological risk assessment. Results should be reported and interpreted within this context and framework. Given the dynamic nature of estuaries, studies that are only snapshots in time and that did not take into account temporal/seasonal variability would at best be considered equivalent to the screening level of an ecological risk assessment.

As scientific understanding and assessment tools improve, so too will our ability to properly assess and eventually predict the outcome of sediment contamination in estuaries. However, only estuarine studies that mimic conditions typical of estuaries will yield results that can be widely extrapolated to nature.

Acknowledgements

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. UNIQUE CHARACTERISTICS OF ESTUARIES
  5. Unique biological characteristics of estuaries
  6. SALINITY AS A CONTROLLING FACTOR
  7. ASSESSMENT TECHNIQUES: CHEMISTRY
  8. ASSESSMENT TECHNIQUES: BIOLOGICAL SURVEYS
  9. ASSESSMENT TECHNIQUES: TOXICITY TESTS
  10. SUMMARY AND CONCLUSIONS: THE WAY FORWARD
  11. Acknowledgements
  12. REFERENCES

We thank Steve Klaine for asking us to write this review paper, two anonymous reviewers, and EVS Consultants for their support. Some useful ideas originated via discussions with Joe Winfield. However, the opinions expressed herein are solely the responsibility of the authors.

REFERENCES

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. UNIQUE CHARACTERISTICS OF ESTUARIES
  5. Unique biological characteristics of estuaries
  6. SALINITY AS A CONTROLLING FACTOR
  7. ASSESSMENT TECHNIQUES: CHEMISTRY
  8. ASSESSMENT TECHNIQUES: BIOLOGICAL SURVEYS
  9. ASSESSMENT TECHNIQUES: TOXICITY TESTS
  10. SUMMARY AND CONCLUSIONS: THE WAY FORWARD
  11. Acknowledgements
  12. REFERENCES
  • 1
    Hedgpeth JW., 1967. The sense of the meeting. In LauffGH, ed, Estuaries. American Association for the Advancement of Science, Washington, DC, pp 707710.
  • 2
    Nichols FH, Cloern JE, Luoma SN, Peterson DH., 1986. The modification of an estuary. Science 231: 567573.
  • 3
    Vallette-Silver NJ, Bricker SB, eds., 1993. Historical trends in contamination of estuarine and coastal sediments. Estuaries 16: 75696.
  • 4
    French PW., 1993. Post-industrial pollutant levels in contemporary Severn Estuary intertidal sediments, compared to pre-industrial levels. Mar Pollut Bull 26: 3035.
  • 5
    Virkanen J., 1998. Effect of urbanization on metal deposition in the Bay of Töölönlahti, Southern Finland. Mar Pollut Bull 36: 729738.
  • 6
    Varanasi U, Reichert WL, Stein JE, Brown DW, Sanborn HR., 1985. Bioavailability and biotransformation of aromatic hydrocarbons in benthic organisms exposed to sediments from an urban estuary. Mar Environ Res 19: 36841.
  • 7
    Stein JE, Hom T, Collier TK, Brown DW, Varanasi U., 1995. Contaminant exposure and biochemical effects in outmigrant juvenile Chinook salmon from urban and nonurban estuaries of Puget Sound, Washington. Environ Toxicol Chem 14: 10191029.
  • 8
    Underwood GJC, Kromkamp J., 1999. Primary production by phytoplankton and microphytobenthos in estuaries. Adv Ecol Res 29: 3153.
  • 9
    Pritchard DW., 1967. What is an estuary: Physical viewpoint. In LauffGH, ed, Estuaries. American Association for the Advancement of Science, Washington, DC, pp 35.
  • 10
    Caspers H., 1967. Estuaries: Analysis of definitions and biological considerations. In LauffGH, ed, Estuaries. American Association for the Advancement of Science, Washington, DC, pp 68.
  • 11
    Chapman PM, Brinkhurst RO., 1981. Seasonal changes in interstitial salinities and seasonal movements of subtidal benthic invertebrates in the Fraser River estuary, BC. Estuar Coast Mar Sci 12: 4966.
  • 12
    Herman PMJ, Middleburg JJ, van de Koppel J, Heip CHR., 1999. Ecology of estuarine benthos. Adv Ecol Res 29: 195240.
  • 13
    Dyer KR., 1973. Estuaries: A Physical Introduction. John Wiley & Sons, Toronto, ON, Canada.
  • 14
    Russell RJ., 1967. Origins of estuaries. In LauffGH, ed, Estuaries. American Association for the Advancement of Science, Washington, DC, pp 9399.
  • 15
    Guilcher A., 1967. Origin of sediments in estuaries. In LauffGH, ed, Estuaries. American Association for the Advancement of Science, Washington, DC, pp 149157.
  • 16
    Postma H., 1967. Sediment transport and sedimentation in the estuarine environment. In LauffGH, ed, Estuaries. American Association for the Advancement of Science, Washington, DC, pp 158179.
  • 17
    Wu Y, Falconer RA, Uncles RJ., 1998. Modelling of water flows and cohesive sediment fluxes in the Humber Estuary, UK. Mar Pollut Bull 37: 182189.
  • 18
    Black KS., 1998. Suspended sediment dynamics and bed erosion in the high shore mudflat region of the Humber Estuary, UK. Mar Pollut Bull 37: 122133.
  • 19
    McManus J., 1998. Temporal and spatial variations in estuarine sedimentation. Estuaries 21: 622634.
  • 20
    Lewis BL, Landing WM., 1991. The biogeochemistry of manganese and iron in the Black Sea. Deep-Sea Res 38: S773S803.
  • 21
    Brügmann L, Bernard PV, van Grieken R., 1992. Geochemistry of suspended matter from the Baltic sea. Mar Chem 38: 303323.
  • 22
    Lau SSS., 2000. The significance of temporal variability in sediment quality for contamination assessment in a coastal wetland. Water Res 34: 387394.
  • 23
    Kinne O., 1966. Physiological aspects of animal life in estuaries with special reference to salinity. Neth J Sea Res 3: 227244.
  • 24
    Boesch DF., 1977. A new look at the distribution of benthos along the estuarine gradient. In CoullBC, ed, Ecology of Marine Benthos. University of South Carolina Press, Raleigh, SC, USA, pp 245266.
  • 25
    Wolff WJ., 1983. Estuarine benthos. In KetchumBH, ed, Ecosystems of the World 26: Estuaries and Enclosed Seas. Elsevier, New York, NY, USA, pp 151181.
  • 26
    Berger VY, Naumov AD, Babkov AI., 1995. The relationship of abundance and diversity of marine benthos to environmental salinity. Russ J Mar Biol 21: 4146.
  • 27
    Browne RA, Wanigasckerra G., 2000. Combined effects of salinity and temperature on survival and reproduction of five species of Artemia. J Exp Mar Biol Ecol 244: 2944.
  • 28
    Icely JD, Nott JA., 1984. On the morphology and fine structure of the alimentary canal of Corophium volutator. Phil Trans R Soc Lond B 306: 4978.
  • 29
    Vernberg FJ, Piyatiratitivorakul S., 1998. Effects of salinity and temperature on the bioenergetics of adult stages of the grass shrimp (Palaemonetes pugio Holthuis) from the North Inlet estuary, South Carolina. Estuaries 21: 76193.
  • 30
    Riedel GF, Sanders JG, Osman RW., 1997. Biogeochemical control on the flux of trace elements from estuarine sediments: Water column oxygen concentrations and benthic infauna. Estuar Coast Shelf Sci 44: 2338.
  • 31
    Chapman PM., 1979. Seasonal movements of subtidal benthic communities in a salt wedge estuary as related to interstitial salinities. PhD thesis. University of Victoria, Victoria, BC, Canada.
  • 32
    Chapman PM., 1981. Measurements of the short-term stability of interstitial salinities in subtidal estuarine sediments. Estuar Coast Shelf Sci 12: 6781.
  • 33
    Berner RA., 1980. Early Diagenesis: A Theoretical Approach. Princeton University Press, Princeton, NJ, USA.
  • 34
    Means JC., 1995. Influence of salinity upon sediment-water partitioning of aromatic hydrocarbons. Mar Chem 51: 316.
  • 35
    Brunk BK, Jirka GH, Lion LW., 1997. Effects of salinity changes and the formation of dissolved organic matter coatings on the sorption of phenanthrene: Implications for pollutant trapping in estuaries. Environ Sci Technol 31: 119125.
  • 36
    Li YH, Burkhardt L, Teraoka H., 1984. Desorption and coagulation of trace elements during estuarine mixing. Geochim Cosmochim Acta 48: 16591664.
  • 37
    Comans RNJ, van Dijk CPJ., 1988. Role of complexation processes in cadmium mobilization during estuarine mixing. Nature 336: 151154.
  • 38
    Turner A, Millward GE., 1994. Partitioning of trace metals in a macrotidal estuary: Implications for contaminant transport models. Estuar Coast Shelf Sci 39: 558.
  • 39
    Ravichandran M., 1996. Distribution of rare earth elements in sediment cores of Sabine-Neches estuary. Mar Pollut Bull 32: 719726.
  • 40
    Schlekat CE, Decho AW, Chandler GT., 1999. Dietary assimilation of cadmium associated with bacterial exopolymer sediment coatings by the estuarine amphipod Leptocheirus plumulosus: Effects of Cd concentration and salinity. Mar Ecol Prog Ser 183: 205216.
  • 41
    Forbes TL, Forbes VE, Giessing A, Hansen R, Kure LK., 1998. Relative role of pore water versus ingested sediment in bio-availability of organic contaminants in marine sediments. Environ Toxicol Chem 17: 24532462.
  • 42
    Paalmann MHH, Van der Weijden CH, Loch JPG., 1995. Sorption of cadmium on suspended matter under estuarine conditions: Competition and complexation with major sea-water ions. Water Air Soil Pollut 73: 4960.
  • 43
    Fisher NS, Reinfelder JR., 1995. The trophic transfer of metals in marine systems. In TessierA, TurnerDR, eds, Metal Speciation and Bioavailability in Aquatic Systems. John Wiley & Sons, New York, NY, USA, pp 363406.
  • 44
    Kim SD, Ma H, Allen HE, Cha DK., 1999. Influence of dissolved organic matter on the toxicity of copper to Ceriodaphnia dubia: Effect of complexation kinetics. Environ Toxicol Chem 18: 24332437.
  • 45
    Chapman PM, Wang F, Janssen C, Persoone G, Allen HE., 1998. Ecotoxicology of metals in aquatic sediments: Binding and release, bioavailability, risk assessment, and remediation. Can J Fish Aquat Sci 55: 22212243.
  • 46
    DeWitt TH, et al., 1990. The influence of organic matter quality on the toxicity and partitioning of sediment-associated fluor-anthene. Environ Toxicol Chem 11: 197208.
  • 47
    Gundersen JL, MacIntyre WG, Hale RC., 1997. pH-Dependent sorption of chlorinated guiacols on estuarine sediments: The effects of humic acids and TOC. Environ Sci Technol 31: 188193.
  • 48
    Sunda W, Guillard RRL., 1976. The relationship between cupric ion activity and the toxicity of copper to phytoplankton. J Mar Res 34: 511529.
  • 49
    Sunda WG, Engel DW, Thuotte RM., 1978. Effects of chemical speciation on toxicity of cadmium to grass shrimp, Palaemonetes pugio: Importance of free cadmium ion. Environ Sci Technol 12: 409413.
  • 50
    Jones MB., 1975. Synergistic effects of salinity, temperature and heavy metals on mortality and osmoregulation in marine and estuarine isopods (Crustacea). Mar Biol 30: 1320.
  • 51
    Frank PM, Robertson PB., 1979. The influence of salinity on toxicity of cadmium and chromium to the blue crab, Callinectes sapidus Bull Environ Contam Toxicol 21: 7478.
  • 52
    Chapman PM, Farrell MA, Brinkhurst RO., 1982. Relative tolerances of selected aquatic oligochaetes to combinations of pollutants and environmental factors. Aquat Toxicol 2: 6978.
  • 53
    Bryant V, McLusky DS, Roddie K, Newberry DM., 1984. Effect of temperature and salinity on the toxicity of chromium to three estuarine invertebrates (Corophium volutator, Macoma balthica, Nereis diversicolor). Mar Ecol Prog Ser 20: 137149.
  • 54
    Bryant V, Newberry DM, McLusky DS, Campbell R., 1985. Effect of temperature and salinity on the toxicity of nickel and zinc to two estuarine invertebrates (Corophium volutator, Macoma balthica). Mar Ecol Prog Ser 24: 139153.
  • 55
    DeLisle PF, Roberts MH Jr., 1988. The effects of salinity on cadmium toxicity to the estuarine amphipod, Mysidopsis bahia: Role of chemical speciation. Aquat Toxicol 12: 357370.
  • 56
    Nebeker AV, Miller CE., 1988. Use of the amphipod crustacean Hyalella azteca in freshwater and estuarine sediment toxicity tests. Environ Toxicol Chem 7: 10271033.
  • 57
    Dwyer FJ, Burch SA, Ingersoll CG, Hunn JB., 1992. Toxicity of trace element and salinity mixtures to striped bass (Morone saxatilis) and Daphnia magna. Environ Toxicol Chem 11: 513520.
  • 58
    Remane A, Schlieper C., 1971. Biology of Brackish Water, 2nd ed. Wiley Interscience, New York, NY, USA.
  • 59
    Kennish MJ., 1990. Ecology of Estuaries, Vol 2. CRC, Boca Raton, FL, USA.
  • 60
    Drake P, Baldo F, Saenz V, Arias AM., 1999. Mocrobenthic community structure in estuarine pollution assessment on the Gulf of Cadiz (SW Spain): Is the phylum-level meta-analysis approach applicable? Mar Pollut Bull 38: 10381047.
  • 61
    Tucker KA, Burton GA Jr., 1999. Assessment of nonpoint-source runoff in a stream using in situ and laboratory approaches. Environ Toxicol Chem 18: 7972803.
  • 62
    Liss PS., 1976. Conservative and non-conservative behaviour of dissolved constituents during estuarine mixing. In BurtonJD, LissPS, eds, Estuarine Chemistry. Academic, London, UK, pp 93130.
  • 63
    Bates TS, Murphy PP, Curl HC Jr, Feely RA., 1987. Hydrocarbon distributions and transport in an urban estuary. Environ Sci Technol 21: 193198.
  • 64
    Schubel JR, Kennedy VS., 1984. The estuary as a filter: An introduction. In KennedyVS, ed, The Estuary as a Filter. Academic, Orlando, FL, USA, pp 111.
  • 65
    Turekian KK., 1977. The fate of metals in the oceans. Geochim Cosmochim Acta 41: 11391144.
  • 66
    Zhou JL, Fileman TW, House WA, Long JLA, Mantoura RFC, Meharg AA, Osborn D, Wright J., 1998. Fluxes of organic contaminants from the river catchment into, through and out of the Humber Estuary, UK. Mar Pollut Bull 37: 330342.
  • 67
    Forstner U., 1989. Contaminated Sediments. Springer-Verlag, Berlin, Germany.
  • 68
    Horowitz AJ., 1991. A Primer on Sediment-Trace Element Chemistry, 2nd ed. Lewis, Chelsea, MI, USA.
  • 69
    Forstner U, Salomons W., 1980. Trace metal analysis on polluted sediments. I. Assessment of sources and intensities. Environ Technol Lett 1: 494505.
  • 70
    Horowitz AJ, Elrick KA., 1988. Interpretation of bed sediment trace metal data: Methods for dealing with the grain size effect. In LichtenbergJJ, WinterJA, WeberCI, FradkinL, eds, Chemical and Biological Characterization of Sludges, Sediments, Dredge Spoils, and Drilling Muds. STP 976. American Society for Testing and Materials, Philadelphia, PA, pp 114128.
  • 71
    Smith JD, Nicholson RA, Moore PJ., 1973. Mercury in sediments from the Thames Estuary. Environ Pollut 4: 153157.
  • 72
    Ackermann F., 1980. A procedure for correcting the grain size effect in heavy metal analyses of estuarine and coastal sediments. Environ Technol Lett 1: 518527.
  • 73
    Ackermann F, Bergmann H, Schleichert V., 1983. Monitoring of heavy metals in coastal and estuarine sediments-A uestionof grain size: < 20μm versus < 607μm. Environ Technol Lett 4: 317328.
  • 74
    Queralt I, Barreiros MA, Carvalho ML, Costa MM., 1999. Application of different techniques to assess sediment quality and point source pollution in low-level contaminated estuarine recent sediments (Lisboa coast, Portugal). Sci Total Environ 241: 3951.
  • 75
    Mayer LM, Fink LK Jr., 1980. Granulometric dependence of chromium accumulation in estuarine sediments in Maine. Estuar Coast Mar Sci 11: 491503.
  • 76
    Goldberg ED, Griffin JJ, Hodge V, Koide M, Windom H., 1979. Pollution history of the Savannah River estuary. Environ Sci Technol 13: 588594.
  • 77
    SchroppSJ, WindomHL, eds., 1988. A guide to the interpretation of metal concentrations in estuarine sediments. Coastal Zone Management Section, Florida Department of Environmental Regulation, Tallahassee, FL, USA.
  • 78
    Loring DH., 1991. Normalization of heavy-metal data from estuarine and coastal sediments. ICES J Mar Sci 48: 101115.
  • 79
    Morse JW, Presley BJ, Taylor RJ, Benoit G, Santshi P., 1993. Trace metal chemistry of Galveston Bay: Water, sediments and biota. Mar Environ Res 36: 137.
  • 80
    Grant A, Middleton R., 1998. Contaminants in sediments: Using robust regression for grain-size normalization. Estuaries 21: 197203.
  • 81
    Di Toro DM, et al., 1991. Technical basis for establishing sediment quality criteria for nonionic organic chemicals by using equilibrium partitioning. Environ Toxicol Chem 10: 15411583.
  • 82
    Di Toro DM, De Rosa LD., 1998. Equilibrium partitioning and organic carbon normalization. In National Sediment Bioaccumulation, 1996. EPA 823-R-98–002. U.S. Environmental Protection Agency, Washington, DC, pp 36.
  • 83
    Bryan GW, Langston WJ., 1992. Bioavailability, accumulation and effects of heavy metals in sediments with special reference to United Kingdom estuaries: A review. Environ Pollut 76: 89131.
  • 84
    Müller G, 1979. Schwermetalle in den Sedimenten des Rheins- Veranderungen seit 1971 Umsch Wiss Tech 79: 778783.
  • 85
    Daskalakis KD, O'Connor TP., 1995. Normalization and elemental sediment contamination in the coastal United States. Environ Sci Technol 29: 470477.
  • 86
    Luoma SN., 1989. Can we determine the biological availability of sediment-bound trace metals? Hydrobiologia 176/177: 379396.
  • 87
    Lee B-G, Griscom SB, Lee J-S, Choi HJ, Koh C-H, Luoma SN, Fisher NS., 2000. Influences of dietary uptake and relative sulfides on metal bioavailability from aquatic sediments. Science 287: 282284.
  • 88
    Schlekat CE, Decho AW, Chandler GT, 2000. Bioavailability of particle-associated silver, cadmium, and zinc to the estuarine amphipod Leptocheirus plumulosis through dietary ingestion. Limnol Oceanogr 45: 1121.
  • 89
    Mayer LM, et al., 1996. Bioavailability of sedimentary contaminants subject to deposit-feeder digestion. Environ Sci Technol 30: 26412645.
  • 90
    Weston DP, Mayer LM., 1998a. In vitro digestive fluid extraction as a measure of the bioavailability of sediment-associated poly-cyclic aromatic hydrocarbons: Sources of variation and implications for partitioning models. Environ Chem Toxicol 17: 820829.
  • 91
    Weston DP, Mayer LM., 1998b. Comparison of in vitro digestive fluid extraction and traditional in vivo approaches as measures of polycyclic aromatic hydrocarbon bioavailability from sediments. Environ Toxicol Chem 17: 830840.
  • 92
    Chen Z, Mayer LM., 1998. Mechanisms of Cu solubilization during deposit feeding. Environ Sci Technol 32: 770775.
  • 93
    Chen Z, Mayer LM., 1999. Assessment of sedimentary Cu availability: A comparison of biomimetric and AVS approaches. Environ Sci Technol 33: 650652.
  • 94
    Tessier A, Couillard Y, Campbell PGC, Auclair JC., 1993. Modeling Cd partitioning in oxic lake sediments and Cd concentrations in the freshwater bivalve Anodonta grandis. Limnol Oceanogr 38: 1017.
  • 95
    Ankley GT, Di Toro DM, Hansen DJ, Berry WJ., 1996. Technical basis and proposal for deriving sediment quality criteria for metals. Environ Toxicol Chem 15: 20562066.
  • 96
    Geesey GG, Borstad L, Chapman PM., 1984. Influence of flow-related events on concentration and phase distribution of metals in the lower Fraser River and a small tributary stream in British Columbia, Canada. Water Res 18: 233238.
  • 97
    Caetano M, Falcao M, Vale C, Bebianno MJ., 1997. Tidal flushing of ammonium, iron and manganese from intertidal sediment pore waters. Mar Chem 58: 203211.
  • 98
    Simpson SL, Apte SC, Batley GE., 1998. Effect of short-term resuspension events on trace metal speciation in polluted anoxic sediments. Environ Sci Technol 32: 620625.
  • 99
    Warren LA, Tessier A, Hare L., 1998. Modeling cadmium accumulation by benthic invertebrates in situ: The relative contributions of sediment and overlying water reservoirs to organism cadmium concentrations. Limnol Oceanogr 43: 4421454.
  • 100
    Wang F, Tessier A, Hare L., 2000. Oxygen measurements in freshwater insect burrows. Freshw Biol (in press).
  • 101
    Rasmussen AD, Banta GT, Andersen O., 2000. Cadmium dynamics in estuarine sediments: Effects on salinity and lugworm bioturbation. Environ Toxicol Chem 19: 380386.
  • 102
    Chapman PM., 1989. Current approaches to developing sediment quality criteria. Environ Toxicol Chem 8: 589599.
  • 103
    Chapman PM, Wang F, Adams W, Green A., 1999. Appropriate applications of sediment quality values for metals and metalloids. Environ Sci Technol 33: 39373941.
  • 104
    Long ER, MacDonald DD, Smith SL, Calder FD., 1995. Incidence of adverse biological effects within ranges of chemical concentrations in marine and estuarine sediments. Environ Manage 19: 8197.
  • 105
    Australia and New Zealand Environment and Conservation Council and Agriculture and Resource Management Council of Australia and New Zealand., 1999. Australian and New Zealand guidelines for fresh and marine water quality. Draft, July 1999 Environment Australia, Community Information Unit, Canberra, ACT.
  • 106
    Mortimer RJG, Krom MD, Hall POJ, Hulth S, Ståhl H., 1998. Use of gel probes for the determination of high resolution solute distributions in marine and estuarine pore waters. Mar Chem 63: 119129.
  • 107
    Watson PG, Frickers TE., 1990. A multilevel, in situ pore-water sampler for use in intertidal sediments and laboratory microcosms. Limnol Oceanogr 35: 13811389.
  • 108
    Remane A., 1934. Die Brackwasserfauna. Verh Dtsch Zool Ges 36: 3474.
  • 109
    Sanders HL., 1968. Marine benthic diversity: A comparative study. Am Nat 102: 243282.
  • 110
    Cognetti G, Maltagliati F., 2000. Biodiversity and adaptive mechanisms in brackish water fauna. Mar Pollut Bull 40: 714.
  • 111
    Barnes RSK., 1989. What, if anything, is a brackish fauna? Trans R Soc Edinb: Earth Sci 80: 235240.
  • 112
    Jones NV., 1981. Feeding and survival strategies of estuarine organisms. In JonesNV, WolffWJ, eds, Feeding and Survival Strategies of Estuarine Organisms. Plenum, New York, NY, USA, pp 291293.
  • 113
    Beukema JJ., 1988. An evaluation of the ABC-method (abundance/biomass comparison) as applied to macrozoobenthic communities living in tidal flats in the Dutch Wadden Sea. Mar Biol 99: 425433.
  • 114
    Craeymeersch JA., 1991. Application of the abundance/biomass comparison method to detect pollution effects on intertidal ma-crobenthic communities. Hydrobiol Bull 24: 133140.
  • 115
    Power JH., 1997. Time and tide wait for no animal: Seasonal and regional opportunities for tidal stream transport or retention. Estuaries 20: 312318.
  • 116
    Kuo AY, Neilson BJ., 1987. Hypoxia and salinity in Virginia estuaries. Estuaries 10: 277283.
  • 117
    Dauer DM, Rodi AJ Jr, Ranasinghe JA., 1992. Effects of low dissolved oxygen events on the macrobenthos of the Lower Chesapeake Bay. Estuaries 15: 384391.
  • 118
    Harper DE Jr, McKinney LD, Salzer RR, Case RJ., 1981. The occurrence of hypoxic bottom water off the upper Texas coast and its effects on the benthic bioata. Contrib Mar Sci 24: 5379.
  • 119
    Holland AF, Shaughnessy AT, Hiegel H., 1987. Long-term variation in mesohaline Chesapeake Bay macrobenthos: Spatial and temporal patterns. Estuaries 10: 227245.
  • 120
    Dauer DM., 1993. Biological criteria, environmental health and estuarine macrobenthic community structure. Mar Pollut Bull 26: 249257.
  • 121
    Dauer DM, Ewing RM, Rodi AJ., 1987. Macrobenthic distribution within the sediment along an estuarine salinity gradient. Int Rev Ges Hydrobiol 72: 529538.
  • 122
    Engle VD, Summers JK, Gaston GR., 1994. A benthic index of environmental condition of Gulf of Mexico estuaries. Estuaries 17: 372384.
  • 123
    Deegan LA, Finn JT, Awazian SG, Ryder-Kieffer CA, Buon-accorsi J., 1997. Development and validation of an estuarine biotic integrity index. Estuaries 20: 601617.
  • 124
    Weisberg SB, Ranasinghe JA, Dauer DM, Schaffner LC, Diaz RJ, Frithsen JB., 1997. An estuarine benthic index of biotic integrity (B-IBI) for Chesapeake Bay. Estuaries 20: 49158.
  • 125
    Engle VD, Summers JK., 1999. Refinement, validation and application of a benthic condition index for Northern Gulf of Mexico estuaries. Estuaries 22: 624635.
  • 126
    Hyland JL, Van Dolah RF, Snoots TR., 1999. Predicting stress in benthic communities of southeastern U.S. estuaries in relation to chemical contamination of sediments. Environ Toxicol Chem 18: 25572564.
  • 127
    Long ER, Robertson A, Wolfe DA, Hameedi J, Sloane GM., 1996. Estimates of the spatial extent of sediment toxicity in major U.S. estuaries. Environ Sci Technol 30: 35853592.
  • 128
    Long ER., 2000. Degraded sediment quality in U.S. estuaries: A review of magnitude and ecological implications. Ecol Appl 10: 338349.
  • 129
    Thompson B, Anderson B, Hunt J, Taberski K, Phillips B., 1999. Relationship between sediment contamination and toxicity in San Francisco Bay. Mar Environ Res 48: 285309.
  • 130
    Schlekat CE, McGee BL, Howard DM, Reinharz E, Velinsky DJ, Wade TL., 1994. Tidal river sediments in the Washington, D.C. area. III. Biological effects associated with sediment contamination. Estuaries 17: 334344.
  • 131
    Chapman PM., 1996. Presentation and interpretation of sediment quality triad data. Ecotoxicology 5: 327339.
  • 132
    Krantzberg G, Hartig JH, Zarull MA., 2000. Sediment management: Deciding when to intervene. Environ Sci Technol 34: 22A27A.
  • 133
    Hall JA, Frid CLL, Gill ME., 1997. The response of estuarine fish and benthos to an increasing discharge of sewage effluent. Mar Pollut Bull 34: 527535.
  • 134
    HillIR, MatthiessenP, HeimbachF, eds., 1994. Guidance Document on Sediment Toxicity Tests and Bioassays for Freshwater and Marine Environments. Society of Environmental Toxicology and Chemistry Europe, Brussels, Belgium.
  • 135
    Matthiessen P, et al., 1998. An assessment of sediment toxicity in the River Tyne estuary, UK, by means of bioassays. Mar Environ Res 45: 115.
  • 136
    Hoss DE, Coston LC, Schaaf WE., 1974. Effects of sea water extracts of sediments from Charleston Harbor, SC, on larval estuarine fishes. Estuar Coast Mar Sci 2: 323328.
  • 137
    Tietjen JH, Lee JJ., 1984. The use of free living nematodes as a bioassay for estuarine sediments. Mar Environ Res 11: 33251.
  • 138
    Matthiessen P, Thain JI, Law RJ, Fileman TW., 1993. Attempts to assess the environmental hazard posed by complex mixtures of organic chemicals in UK estuaries. Mar Pollut Bull 26: 9095.
  • 139
    Norton BL, Lewis MA, Mayer FL., 1999. Storage duration and temperature and the acute toxicities of estuarine sediments to Mysidopsis bahia and Leptocheirus plumulosus. Bull Environ Contam Toxicol 63: 157166.
  • 140
    McLusky D, Bryant V, Campbell R., 1986. The effects of temperature and salinity on the toxicity of heavy metals to marine and estuarine invertebrates. Oceanogr Mar Biol Annu Rev 24: 481520.
  • 141
    McLusky D., 1989. The Estuarine Ecosystem, 2nd ed. Blackie and Sons, London, UK.
  • 142
    Jernelov A, Rosenberg R., 1976. Stress tolerance of ecosystems. Environ Conserv 3: 4346.
  • 143
    DeWitt TH, Swartz RC, Lamberson JO., 1989. Measuring the acute toxicity of estuarine sediment. Environ Toxicol Chem 8: 10351048.
  • 144
    Hyne RV, Everett DA., 1998. Application of a benthic euryhaline amphipod, Corophium sp., as a sediment toxicity testing organism for both freshwater and estuarine systems. Arch Environ Contam Toxicol 34: 2633.
  • 145
    Wolfe DA, Long ER, Thursby GB., 1996. Sediment toxicity in the Hudson-Raritan estuary: Distribution and correlations with chemical contamination. Estuaries 19: 901912.
  • 146
    U.S. Environmental Protection Agency. 1994. Methods for assessing the toxicity of sediment-associated contaminants with estuarine and marine amphipods. EPA-600-R-94–025. Washington, DC.
  • 147
    American Society for Testing and Materials. 1999. Standard guide for conducting 10-d static sediment toxicity tests with estuarine and marine amphipods. In 1999 Annual Book of ASTM Standards. E 1367–92. American Society for Testing and Materials, Philadelphia, PA, pp 733758.
  • 148
    Bufflap SE, Allen HE., 1995. Sediment pore water collection methods for trace metal analysis: A review. Water Res 29: 165177.
  • 149
    Angelidis TN., 1997. Comparison of sediment pore water sampling for specific parameters using two techniques. Water Air Soil Pollut 99: 179185.
  • 150
    McGee BL, Fisher DJ, Yonkos LT, Ziegler GP, Turley S., 1999. Assessment of sediment contamination, acute toxicity, and population viability of the estuarine amphipod Leptocheirus plu-mulosus in Baltimore Harbor, Maryland, USA. Environ Toxicol Chem 18: 21512160.
  • 151
    Schlekat CE, et al., 1995. Interlaboratory comparison of a 10-day sediment toxicity test method using Ampelisca abdita, Eoh-austorius estuarius and Leptocheirus plumulosus. Environ Toxicol Chem 12: 21632174.
  • 152
    McGee BL, Schlekat CE, Reinharz E., 1993. Assessing sublethal levels of sediment contamination with the estuarine amphipod Leptocheirus plumulosus. Environ Toxicol Chem 12: 577588.
  • 153
    Emery VL Jr, Moore DW, Gray BR, Duke BM, Gibson AB, Wright RB, Farrar JD., 1997. Development of a chronic sublethal sediment bioassay using the estuarine amphipod Leptocheirus plumulosus (Shoemaker). Environ Toxicol Chem 16: 19121920.
  • 154
    McPherson CA, Chapman PM., 2000. Copper effects on potential sediment test organisms: The importance of appropriate sensitivity. Mar Pollut Bull 40: 656665.
  • 155
    Hall LW Jr, Alden RW III., 1997. A review of concurrent ambient water column and sediment toxicity testing in the Chesapeake Bay watershed: 1990–1994. Environ Toxicol Chem 16: 16061617.
  • 156
    McLusky DS., 1967. Some effects of salinity on the survival, moulting and growth of Corophium volutator (Amphipoda). J Mar Biol Assoc UK 48: 607617.
  • 157
    McLusky D., 1970. Salinity preference in Corophium volutator. J Mar Biol Assoc UK 50: 747752.
  • 158
    van den Hurk P, Chapman PM, Roddie B, Swartz R., 1992. A comparison of North American and Western European infaunal amphipod species in a toxicity test on North Sea sediments. Mar Ecol Prog Ser 91: 37243.
  • 159
    Swartz RC, Schults DW, DeWitt TH, Ditsworth GR, Lamberson JO., 1990. Toxicity of fluoranthene in sediment to marine amphipods: A test of the equilibrium partitioning approach to sediment quality criteria. Environ Toxicol Chem 9: 10711080.
  • 160
    Shyamasundari K., 1973. Studies on the tube building amphipods (Corophium triaenonyx) (Stebbing) from Visakanpatnam Harbor: Effects of salinity and temperature. Biol Bull 144: 503510.
  • 161
    Ferguson PL, Chandler GT., 1998. A laboratory and field comparison of sediment polycyclic aromatic hydrocarbon bioaccumulation by the cosmopolitan estuarine polychaete Streblospio benedicti (Webster). Mar Environ Res 45: 387401.
  • 162
    Bridges TS, Levin LA, Cabrera D, Plaia G., 1994. Effects of sediment amended with sewage, algae, or hydrocarbons on growth and reproduction in two opportunistic polychaetes. J Exp Mar Biol Ecol 177: 99119.
  • 163
    Environment Canada., 1992. Biological test method: Acute test for sediment toxicity using marine or estuarine amphipods. EPS 1/RM/26. Ottawa, ON.
  • 164
    Hyland JL., 1981. Comparative structure and response to (petroleum) disturbance in two nearshore infaunal communities. PhD thesis. University of Rhode Island, Kingston, RI, USA.
  • 165
    Bendell-Young LI., 1999. Application of a kinetic model of bioaccumulation across a pH and salinity gradient for the prediction of cadmium uptake by the sediment dwelling Chironomidae. Environ Sci Technol 33: 15011508.
  • 166
    Environment Canada., 1997. Biological test method: Test for survival and growth in sediment using the larvae of freshwater midges (Chironomus tentans or Chironomus riparius). Final Report. EPS 1/RM/32. Ottawa, ON.
  • 167
    Reynoldson TB, Thompson SP, Bamsey JL., 1991. A sediment bioassay using the tubificid oligochaete worm Tubifex tubifex. Environ Toxicol Chem 10: 10611072.
  • 168
    Reynoldson TB., 1994. A field test of a sediment bioassay with the oligochaete worm Tubifex tubifex (Muller, 1774). Hydro-biologia 278: 223230.
  • 169
    Hunter J., 1981. Survival strategies of tubificids in the Thames and other estuaries. In JonesNV, WolffWJ, eds, Feeding and Survival Strategies of Estuarine Organisms. Plenum, New York, NY, USA, pp 5363.
  • 170
    Chapman PM, Brinkhurst RO., 1984. Lethal and sublethal tolerances of aquatic oligochaetes with reference to their use as a biotic index of pollution. Hydrobiologia 115: 139144.
  • 171
    Chapman PM, Farrell MA, Brinkhurst RO., 1982. Relative tolerances of selected aquatic oligochaetes to individual pollutants and environmental factors. Aquat Toxicol 2: 4767.
  • 172
    Nagler JJ, Cyr DG, 1997. Exposure of male American plaice (Hippoglossoides platessoides) to contaminated marine sediments decreases the hatching success of their progeny. Environ Toxicol Chem 18: 17331738.
  • 173
    Ringwood AH, DeLorenzo ME, Ross PE, Holland AF., 1997. Interpretation of Microtox® solid-phase toxicity tests: The effects of sediment composition. Environ Toxicol Chem 16: 11351140.
  • 174
    Ho KTY, Quinn JG, 1993. Bioassay-directed fractionation of organic contaminants in an estuarine sediment using the new mutagenic bioassay, Mutatox™. Environ Toxicol Chem 12: 823830.
  • 175
    Costa FO, Correia AD, Costa M II., 1998. Acute marine sediment toxicity: A potential new test with the amphipod Gammarus locusta. Toxicol Environ Saf 40: 8187.
  • 176
    Phelps HL, Pearson WH, Hardy JT, 1985. Clam burrowing behaviour and mortality related to sediment copper. Mar Pollut Bull 16: 309313.
  • 177
    Williamson RB, Wilcock RJ, Wise BE, Pickmere SE., 1999. Effect of burrowing by the crab Helice crassa on chemistry of intertidal muddy sediments. Environ Toxicol Chem 18: 20782086.
  • 178
    Ankley GT, Katko A, Authur JW, 1990. Identification of ammonia as an important sediment-associated toxicant in the lower Fox River and Green Bay, Wisconsin. Environ Toxicol Chem 9: 313322.
  • 179
    Burgess RM, Schweitzer KA, McKinney RA, Phelps DK., 1993. Contaminated marine sediments: Water column and interstitial toxic effects. Environ Toxicol Chem 12: 127138.
  • 180
    Whiteman FW, Ankley GW, Kahl MD, Rau DM, Balcer MD., 1996. Evaluation of interstitial water as a route of exposure for ammonia in sediment tests with benthic macroinvertebrates. Environ Toxicol Chem 15: 794801.
  • 181
    Moore DW, Bridges TS, Gray BR, Duke BM., 1997. Risk of ammonia toxicity during sediment bioassays with the estuarine amphipod Leptocheirus plumulosus. Environ Toxicol Chem 16: 10201027.
  • 182
    Hazel CR, Thomsen W, Meith SJ., 1971. Sensitivity of striped bass and stickleback to ammonia in relation to temperature and salinity. Calif Fish Game 57: 154161.
  • 183
    Ankley GT, Shubauer-Berigan MK, Monson PD., 1995. Influence of pH and hardness on toxicity of ammonia to the amphipod Hyalella azteca. Can J Fish Aquat Sci 52: 20782083.
  • 184
    Schubauer-Berigan MK, Monson PD, West CW, Ankley GT., 1995. Influence of pH on the toxicity of ammonia to Chironomus tentans and Lumbriculus variegatus. Environ Toxicol Chem 14: 713717.
  • 185
    Kohn NP, Word JQ, Niyogi DK, Ross LT, Dillon T, Moore DW., 1994. Acute toxicity of ammonia to four species of marine amphipod. Mar Environ Res 38: 115.
  • 186
    Beaumont MW, Butler PJ, Taylor EW., 1995. Exposure of brown trout, Salmo trutta, to sublethal copper concentrations and its effects upon sustained swimming performance. Aquat Toxicol 33: 4563.
  • 187
    Shaw JR, Wood CM, Birge WJ, Hogstrand C., 1998. Toxicity of silver to the marine teleost (Oligocottus maculosus): Effects of salinity and ammonia. Environ Toxicol Chem 17: 594600.
  • 188
    Wang F, Chapman PM., 1999. Biological implications of sulfide in sediment-A review focusing on sediment toxicity. Environ Toxicol Chem 18: 25262532.
  • 189
    MacCrehan W, Shea D., 1995. Temporal relationship of thiols to inorganic sulfur compounds in anoxic Chesapeake Bay sediment porewater. In VairavamurthyMA, SchoonenMAA, eds, Geo-chemical Transformation of Sedimentary Sulfur. ACS Symposium Series 612. American Chemical Society, Washington, DC, pp 294310.
  • 190
    Wharfe J., 1977. The intertidal sediment habitats of the lower Medway estuary in Kent. Environ Pollut 13: 7991.
  • 191
    Dewitt TH, Ditsworth GR, Swartz RC., 1988. Effects of natural sediment features on survival of the phoxocephalid amphipod, Rhepoxynius abronius. Mar Environ Res 25: 99124.
  • 192
    Lacey R, Watzin MC, McIntosh AW., 1999. Sediment organic matter content as a confounding factor in toxicity tests with Chironomus tentans. Environ Toxicol Chem 18: 231236.
  • 193
    Suedel BC, Rodgers JH Jr., 1994. Development of formulated reference sediments for freshwater and estuarine sediment testing. Environ Toxicol Chem 13: 11631175.
  • 194
    Kemble NE, Dwyer FJ, Ingersoll CG, Dawson TD, Norberg-King TJ., 1999. Tolerance of freshwater test organisms to formulated sediments for use as control materials in whole-sediment toxicity tests. Environ Toxicol Chem 18: 222230.
  • 195
    Tagatz ME, Ivey JM, Gregory NR, Oglesby JL., 1981. Effects of pentachlorophenol on field- and laboratory-developed estuarine benthic communities. Bull Environ Contam Toxicol 26: 37143.
  • 196
    Tagatz ME, Deans CH., 1983. Comparison of field- and laboratory-developed estuarine benthic communities for toxicant-exposure studies. Water Air Soil Pollut 20: 199209.
  • 197
    Austen MC, Somerfiled PJ., 1997. A community level sediment bioassay applied to an estuarine heavy metal gradient. Mar Environ Res 43: 315328.
  • 198
    Menzie CA., 1984. Diminishment of recruitment: A hypothesis concerning impacts on benthic communities. Mar Pollut Bull 15: 127128.
  • 199
    Phelps HL, Warner KA., 1990. Estuarine sediment bioassay with oyster pediveliger larvae (Crassostrea gigas). Bull Environ Contam Toxicol 44: 197204.
  • 200
    Watzin, MC, Poscigno PF, Burke WD., 1994. Community-level field method for testing the toxicity of contaminated sediments in estuaries. Environ Toxicol Chem 13: 1871193.
  • 201
    Watzin MC, Roscigno PR., 1997. The effects of zinc contamination on the recruitment and early survival of benthic invertebrates in an estuary. Mar Pollut Bull 34: 443455.
  • 202
    Hall JA, Frid CLJ., 1995. Responses of estuarine benthic macrofauna in copper-contaminated sediments to remediation of sediment quality. Mar Pollut Bull 30: 694700.
  • 203
    Millward RN, Grant A., 1995. Assessing the impact of copper on nematode communities from a chronically metal-enriched estuary using pollution-induced community tolerance. Mar Pollut Bull 30: 701706.
  • 204
    Grant A, Hateley JG, Jones NV., 1989. Mapping the ecological impact of heavy metals on the estuarine polychaete Nereis diversicolor using inherited metal tolerance. Mar Pollut Bull 20: 235238.
  • 205
    Chandler GT, Coull BC, Schizas NV, Donelan TL., 1997. A culture-based assessment of the effects of chlorpyrifos on multiple meiobenthic copepods using microcosms of intact estuarine sediments. Environ Toxicol Chem 16: 23392346.
  • 206
    Kurtz JC, Devereux R, Barkay T, Jonas RB., 1998. Evaluation of sediment slurry microcosms for modelling microbial communities in estuarine sediments. Environ Toxicol Chem 17: 12741281.
  • 207
    Chapman PM., 1995. Do sediment toxicity tests require field validation? Environ Toxicol Chem 14: 14511453.
  • 208
    Chapman PM., 2000. Whole effluent toxicity testing-Usefulness, level of protection, and risk assessment. Environ Toxicol Chem 19: 313.
  • 209
    McGee BL, Wright DA, Fisher DJ., 1998. Biotic factors modifying acute toxicity of aqueous cadmium to the estuarine amphipod Leptocheirus plumulosus. Arch Environ Contam Toxicol 34: 3440.
  • 210
    Swartz RC, Cole FA, Lamberson JO, Ferraro SP, Schults DW, DeBen WA, Lee H II, Ozretich RJ., 1994. Sediment toxicity, contamination and amphipod abundance at a DDT- and dieldrin-contaminated site in San Francisco Bay. Environ Toxicol Chem 13: 949962.
  • 211
    McGee BL., 1998. Population dynamics of the estuarine amphipod Leptocheirus plumulosus: Implications for sediment toxicity tests. PhD thesis. University of Maryland, College Park, MD, USA.
  • 212
    Linke-Gamenick I, Forbes VE, Sibly RM., 1999. Density-dependent effects of a toxicant on life-history traits and populations dynamics of a capitellid polychaete. Mar Ecol Prog Ser 184: 139148.
  • 213
    Calow P, Sibly RM, Forbes VE., 1997. Risk assessment on the basis of simplified life-history scenarios. Environ Toxicol Chem 16: 19831989.
  • 214
    Grant A., 1998. Population consequences of chronic toxicity: Incorporating density dependence into the analysis of life table response experiments. Ecol Model 105: 325335.
  • 215
    Long ER, Chapman PM., 1985. A sediment quality triad: Measures of sediment contamination, toxicity and infaunal community composition in Puget Sound. Mar Pollut Bull 16: 405415.
  • 216
    Chapman PM, Dexter RN, Long ER., 1987. Synoptic measures of sediment contamination, toxicity and infaunal community composition (the Sediment Quality Triad) in San Francisco Bay. Mar Ecol Prog Ser 37: 7596.
  • 217
    Del Valls TA, Conradi M, Garcia-Adiego E, Forja JM, Gómez-Parra A., 1998. Analysis of macrobenthic community structure in relation to different environmental sources of contamination in two littoral ecosystems from the Gulf of Cádiz (SW Spain). Hydrobiolgia 385: 5970.
  • 218
    Del Valls TA, Forja JM, Gómez-Parra A., 1998. Integrative assessment of sediment quality in two littoral ecosystems from the Gulf of Cádiz, Spain. Environ Toxicol Chem 17: 10731084.
  • 219
    IngersollCG, DillonT, BiddingerGR, eds., 1997. Ecological Risk Assessment of Contaminated Sediments. Society of Environmental Toxicology and Chemistry, Pensacola, FL, USA.
  • 220
    Hill RA, Chapman PM, Mann GL, Lawrence GS., 2000. Level of detail in ecological risk assessments. Mar Pollut Bull 40: 471477.
  • 221
    Hargis WJ Jr, Roberts MH Jr, Zwerner DE., 1984. Effects of contaminated sediments and sediment-exposed effluent water on an estuarine fish: Acute toxicity. Mar Environ Res 14: 337354.
  • 222
    Roberts MH Jr, Hargis WJ Jr, Strobel CJ, De Lisle PF., 1989. Acute toxicity of PAH contaminated sediments to the estuarine fish, Leiostomus xanthurus. Bull Environ Contam Toxicol 42: 142149.
  • 223
    Davies TT, Davis DG, Elmore JP., 1993. Technical panel recommendations concerning the use of acute amphipod tests in evaluation of dredged material. Technical Memorandum. U.S. Environmental Protection Agency, Washington, DC.
  • 224
    Winger PV, Lasier PJ, Geitner H., 1993. Toxicity of sediments and pore water from Brunswick estuary, Georgia. Arch Environ Contam Toxicol 25: 371376.
  • 225
    Kravitz MJ, Lamberson JO, Ferraro SP, Swartz RC, Boese BL, Specht DT., 1999. Avoidance response of the estuarine amphipod Eohaustorius estuarius to polycyclic aromatic hydrocarbon-contaminated field-collected sediments. Environ Toxicol Chem 18: 12321235.
  • 226
    Lawrence AJ, Poulter C., 1996. The potential role of the estuarine amphipod Gammarus duebeni in sub-lethal ecotoxicology testing. Water Sci Technol 34: 93100.
  • 227
    Lawrence AJ, Poulter C., 1998. Development of a sub-lethal pollution bioassay using the estuarine amphipod Gammarus due-beni. Water Res 32: 569578.
  • 228
    Schlekat CE, McGee BL, Reinhard E., 1992. Testing sediment toxicity in Chesapeake Bay with the amphipod Leptocheirus plumulosus: An evaluation. Environ Toxicol Chem 11: 225236.
  • 229
    Borowsky B, Aitken-Ander P, Tanacredi JT., 1997. Changes in reproductive morphology and physiology observed in the amphipod crustacean, Melita nitida Smith, maintained in the laboratory on polluted estuarine sediments. J Exp Mar Biol Ecol 214: 8595.
  • 230
    Quintino V, Re A., 1995. A potential new estuarine amphipod test species from Europe. Abstracts, 2nd Society of Environmental Toxicology and Chemistry World Congress, Vancouver, Canada, November 5–9, p 314.
  • 231
    Schuytema GS, Nebeker AV, Stutzman TW., 1997. Salinity tolerance of Daphnia magna and potential use for estuarine sediment toxicity tests. Arch Environ Contam Toxicol 33: 194198.
  • 232
    Wirth EF, Fulton MH, Chandler GT, Key PB, Scott GI., 1998. Toxicity of sediment associated PAHs to the estuarine crustaceans, Palaemonetes pugio and Amphiascus tenuiremis. Bull Environ Contam Toxicol 61: 637644.
  • 233
    Byrne PA, O'Halloran JO., 1999. Aspects of assaying sediment toxicity in Irish estuarine ecosystems. Mar Pollut Bull 39: 97105.
  • 234
    Phelps HL., 1989. Clam burrowing bioassay for estuarine sediment. Bull Environ Contam Toxicol 43: 838845.
  • 235
    Phelps HL., 1990. Development of an estuarine sediment burrowing bioassay for shipboard use. Bull Environ Contam Toxicol 45: 722728.
  • 236
    Schiewe MH, Hawk EG, Actor DI, Krahn MM., 1985. Use of a bacterial bioluminescence assay to assess toxicity of contaminated marine sediments. Can J Fish Aquat Sci 42: 12441248.
  • 237
    Chapman PM, Swartz RC, Roddie B, Phelps H, van den Hurk P, Butler R., 1992. An international comparison of sediment toxicity tests in the North Sea. Mar Ecol Prog Ser 91: 253264.