Assessing sediment contamination in estuaries

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Environmental Toxicology and Chemistry, Vol. 20, No. 1, pp. 3–22, 2001q 2001 SETAC

Printed in the USA0730-7268/01 $9.00 1 .00

Annual Review

ASSESSING SEDIMENT CONTAMINATION IN ESTUARIES

PETER M. CHAPMAN* and FEIYUE WANGEVS Environment Consultants, 195 Pemberton Avenue, North Vancouver, British Columbia, V7P 2R4, Canada

(Received 15 February 1999; Accepted 10 June 2000)

Abstract—Historic and ongoing sediment contamination adversely affects estuaries, among the most productive marine ecosystemsin the world. However, all estuaries are not the same, and estuarine sediments cannot be treated as either fresh or marine sedimentsor 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 inother 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 ofcontaminants between sediments and overlying or interstitial water. Salinity also controls the distribution and types of estuarinebiota. 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 paradoxof brackish water). Chemical, toxicological, and community-level assessment techniques for estuarine sediment are reviewed andassessed, including chemistry (grain size effects, background enrichment, bioavailability, sediment quality values, interstitial waterchemistry), biological surveys, and whole sediment toxicity testing (single-species tests, potential confounding factors, communitylevel tests, laboratory-to-field comparisons). Based on this review, there is a clear need to tailor such assessment techniquesspecifically for estuarine environments. For instance, bioavailability models including equilibrium partitioning may have littleapplicability to estuarine sediments, appropriate reference comparisons are difficult in biological surveys, and there are too fewfull-gradient estuarine sediment toxicity tests available. Specific recommendations are made to address these and other issues.

Keywords—Sediment Contamination Toxicity Estuaries Salinity

INTRODUCTION

Sediment contamination is receiving increasing attentionfrom the scientific community. However, most sediment as-sessment methods developed to date are applicable primarilyto fresh or salt waters, not to waters that are truly estuarine.Yet estuaries receive significant anthropogenic inputs fromboth point and nonpoint sources upstream and from metro-politan areas and industries located on or near estuaries. Over30 years ago, they were called ‘‘the septic tank of the meg-alopolis’’ [1]; historic contamination remains a significant con-cern for many estuarine sediments [2–5]. Sediment contami-nation in estuaries has been associated not only with effectson benthic species [6] but also with effects on water columnspecies, including migrating salmon [7].

Estuaries are among the most productive marine ecosystemsin the world [8] and are critical to the life history and devel-opment (e.g., rearing, feeding, migration routes, and nurserygrounds) of many aquatic species. Thus, it is critical that sed-iment contamination in estuaries and its biological and eco-logical significance be properly and fully assessed.

Toxicity and other assessment methods developed specifi-cally for estuarine sediments are few and relatively new, andthere is a disturbing tendency, in the case of both chemicaland biological assessment methods, to treat estuaries as eitherfreshwater tainted by salt or the reverse. Many studies andresearchers do not appear to properly appreciate the uniqueand dynamic nature of estuarine ecosystems.

The purpose of this paper is to provide that appreciationas part of a review of the state-of-the-art with regard to es-

* To whom correspondence may be addressed([email protected]).

tuarine sediment contamination assessment methods. The pa-per begins with a description of the unique characteristics ofestuaries. It then proceeds to review available assessment tech-niques beginning with sediment chemistry and continuingthrough biological surveys and sediment toxicity testing. Thepaper concludes by suggesting the way forward.

UNIQUE CHARACTERISTICS OF ESTUARIES

What is an estuary?

The term ‘‘estuary’’ has historically referred to the lowertidal reaches of a river. Pritchard [9] defines an estuary as fol-lows: ‘‘An estuary is a semi-enclosed coastal body of waterwhich has a free connection with the open sea and within whichsea water is measurably diluted with fresh water derived fromland drainage.’’ He further divides estuaries into four classesbased on physical characteristics (Table 1). Caspers [10] onlypartly agrees with this definition, arguing that tidal changes arealso 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 mostsimply based on salinity, specifically as an area where fresh-water enters saline water and where the salinity is, during atleast some period of time, neither truly saline (.30 g/L [ppt])nor truly fresh (0 g/L). However, there are areas that fit thisdefinition that are not estuaries. Examples include groundwaterflow into the sea (e.g., Biscayne Bay, FL, USA) and anthro-pogenic point-source discharges, such as sewage effluent. Fur-ther, the word ‘‘estuary’’ is derived from the Latin aestus, thetide. Thus, our suggested definition of an estuary is based onPritchard’s [9] physical definition but also includes mentionof tidal influences: ‘‘An estuary is a semi-enclosed and tidalcoastal body of water which has a free connection with the

4 Environ. Toxicol. Chem. 20, 2001 P.M. Chapman and F. Wang

Table 1. Estuaries defined by physical characteristics (adapted from Pritchard [9])

Type of estuary Description Example(s)

Drowned river valley Found along coastlines with wide coastal plains; only a portion of the area affectedby tides is estuarine based on salinity diluted by freshwater

Chesapeake Bay, Mary-land, USA

Fjord Generally U-shaped, gouged out by glaciers; mouth often has a shallow sill preced-ing a deep basin (e.g., .100 m depth, Fig. 1)

Norwegian and BritishColumbia (Canada)coastlines

Bar-built Offshore barrier sand islands and sand spits build above sea level and extend be-tween headlands in a chain broken by one or more inlets; enclosed area generallyelongated relative to shoreline; may be more than one river drainage; reduced tid-al action; wind-driven circulation

Pamlico Estuary, NorthCarolina, USA

Tectonic Coastal indentures formed by faulting or local subsistence, with an excess freshwaterflow

San Francisco Bay,California, USA

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

open sea and within which sea water is measurably dilutedwith fresh water derived from land drainage.’’ We define es-tuarine sediments as sediments whose interstitial salinities areneither truly fresh nor truly saline; that is, they range above1 and below 30 g/L. This latter distinction is particularly im-portant as sediments with lesser or greater interstitial salinitiescan and do occur within regions that are designated as estuariesbut are, respectively, more appropriately subject to freshwaterand marine assessment methods.

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

Salt wedge estuaries (Fig. 1). River flow is dominant. Saltwater moves up the estuary in the form of a defined wedgewhose upstream penetration is greatest during periods of lowriver flow (e.g., winter or low-rainfall periods) and least during

periods of high river flow (e.g., freshet). Salt wedge estuariesinclude the mouths of some major rivers (e.g., the Fraser River,BC, Canada). In salt wedge estuaries (see Assessment Tech-niques: Biological Surveys), seasonal shifts occur in the dis-tribution of benthic infaunal species related to seasonal, notdiurnal, changes in interstitial salinities [11]. These same sa-linity shifts can affect the availability of sediment contami-nants (see Assessment Techniques: Chemistry).

Shallow, partially mixed estuaries and fjords (Fig. 1). Sim-ilar to salt wedge estuaries, there are two layers of flow: saltwater at depth and freshwater on the surface. However, riverflow is modified by tidal currents. Shallower estuaries, in-cluding bar-built estuaries, are subject to vertical mixing; deep-er estuaries such as fjords are subject to entrainment (one-waymovement of saline waters into surface freshwaters). The es-tuary may be partially mixed or highly stratified. In theseestuaries, the area of the bottom that is transitional between

Contamination in estuaries Environ. Toxicol. Chem. 20, 2001 5

marine and freshwater environments does not demonstrate thelarge-scale seasonal movements up- and downstream typicalof salt wedge estuaries. Examples include the estuaries of theChesapeake Bay system (e.g., the James River, VA) in theUnited States and the River Tees and the Thames River in theUnited 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 andincreased hydrogen sulfide concentrations in the bottom wa-ters.

Vertically homogeneous estuaries. Tidal currents predom-inate, and vertical salinity differences are less than 1 g/L be-cause 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 var-iation can occur where the ratio of width to depth is sufficientlylarge, such that the right-hand-side of the estuary (looking tosea) will contain lower-salinity water than on the left-handside. Shallow, well-mixed estuaries exhibit intense couplingbetween benthic and pelagic systems [12].

There is, of course, a spectrum of estuarine patterns andgradual transitions between the three previously mentionedmajor types, and estuaries may differ seasonally. For example,the Vellar estuary in India is a salt wedge estuary during pe-riods of high river flow but can become homogeneous in thehot season with no river flow; the Columbia River estuary,Washington, USA, can vary from a salt wedge to a highlystratified estuary [13]. There are, in addition, exceptional caseswhere, for instance, nearly all mixing occurs in a very limitedarea. An example is the reversing falls in the gorge of the St.John River, New Brunswick, Canada. Constrictions and prom-ontories can also greatly affect estuarine circulation patterns,and complex circulation patterns occur where a number ofestuaries are tributary to a large estuarine system, such as theBay of Fundy, New Brunswick, Canada.

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

The bottom line is that estuaries are dynamic, complex, andunique systems. Truly estuarine sediments cannot be treatedas either freshwater or marine systems, and they cannot beproperly assessed by means of only snapshot-in-time samplingwithout understanding both seasonal and spatial variability andprocesses, in particular bottom-water and sediment interstitialsalinity variations.

Unique physicochemical characteristics of estuaries

As noted previously, estuaries are associated with rivers orother forms of runoff from land. They are the immediate re-cipients of sediment carried by those rivers, as manifest bythe formation of river deltas. However, this does not implythat sediments in estuaries are all fluvial in origin. Dependingon the sediment load of the entering river and on the estuarinecirculation patterns, sediments in estuaries can come from in-land and/or from the sea [15]. Estuaries with entering rivers

with high sediment loads may be filled rapidly with fluvialsediments. In contrast, estuaries with entering streams withlow sediment loads may be filled solely or primarily by marinesediments. This can occur because of the penetration of saltwater (as in salt wedge estuaries; Fig. 1) and/or diffusion ofmarine suspended matter. As a result, deposition of marinesediments can occur upriver of saltwater penetration [16].

Although many estuaries are the recipients of depositedsands, the most common sediment deposits for estuaries withlimited wave action are fine-grained muds that form shoalsand tidal flats. Particle settling can be enhanced by the changefrom fresh to salt water, the rise and fall of the water levelwith the tides, and the presence of turbidity maxima duringslack tides. Cohesive sediment fluxes can result in both de-position 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 sedi-mentation, ‘‘Change is the norm within the estuarine system.’’

Because they provide an interface between fresh and saltwaters, estuaries have strong gradients in many physical andchemical variables, including salinity, temperature, pH, dis-solved oxygen, redox potential, nutrients, and amount andcomposition of particles. Gradients exist not only along theirlength, from river to sea, and laterally but also vertically (fromwater column to sediment), particularly in stratified and par-tially stratified estuaries such as salt wedge estuaries andfjords. These gradients may also be subject to seasonal andother temporal variations, which can have a wide variation ininfluence on many biogeochemical processes occurring in es-tuaries. Under anoxic conditions, for example, some metals,such as Fe and Mn, are mobilized from reducing sedimentsand remain dissolved in the water column [20], whereas othermetals, such as Cd, Cu, Zn, and Cr, may be removed from thewater column by sulfide precipitation or by reduction to in-soluble solids [21]. Effects of temporal variability can be farfrom negligible when assessing sediment quality in estuarinewetland areas [22].

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

Unique biological characteristics of estuaries

Estuaries are extremely productive and important feeding,migration, and rearing zones. However, they are also transi-tional areas that are challenging to both residents and immi-grants alike. Ecologically, estuaries are zones of reduced in-terspecific (but not intraspecific) biotic competition due tooverriding physical-chemical factors, in particular salinity.Benthic fauna in estuaries either filter their food directly fromthe water column or depend on the physical deposition of foodparticles onto the sediment surface or incorporated into thesediment matrix [12]. Faunal distributions in estuaries are con-trolled primarily by salinity and secondarily by factors suchas substrate, temperature, dissolved oxygen, and anthropogenicpollution [23–26]. In many cases, there is a complex corre-lation between temperature and salinity, each of which is ca-pable of modifying biological tolerances to the other factor[27]. Variations in salinity in both the overlying waters andthe sediment interstitial waters occur vertically, horizontallyand with time. Organisms can survive in estuaries by one ora combination of the following strategies: avoiding estuarineconditions (e.g., saltwater organisms remaining within the saltwedge in salt wedge estuaries, freshwater organisms remaining


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

Fig. 3. Relative changes to saline sediment interstitial salinities withexposure to less saline overlying waters. (A) sediments .90% sandwith 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 salinities10 g/L. Top of each curve: interstitial salinities in the top centimeterof sediment; bottom of each curve: interstitial salinities in the bottom6 cm of sediment. Schematic developed based on actual laboratoryand field measurements [31,32]. A similar but reverse pattern occurswhen sediments interstitial waters are fresh and overlying waters aresaline.

above), reducing contact with inimical environments (e.g., in-terstitial waters of muddy sediments can have very differentsalinities than overlying waters [11]), adaptation (e.g., ion reg-ulation, volume regulation, or osmoregulation [28]), or accli-mation. However, not all organisms living in estuaries liveunder optimal conditions, which results in natural bioenergeticstress to those organisms [29]. For example, periodic seasonalcycles of anoxic and oxic conditions in bottom waters of par-tially stratified estuaries (e.g., in Chesapeake Bay) result in acorresponding cycle of mortality and recolonization by benthicmacrofauna [30]. Estuarine organisms living in such stressfulnatural conditions may be more (or less) susceptible to an-thropogenic stress.

SALINITY AS A CONTROLLING FACTOR

Spatial and temporal fluctuations of salinity in estuarinesediments

As mentioned previously, salinity can fluctuate in bothoverlying waters and sediment interstitial waters spatially (hor-izontally and vertically) and temporally. It is not unusual fortens of square kilometers of bottom sediments in salt wedgeestuaries to show seasonal interstitial salinity differences re-lated to seasonal differences in the extent and duration of thesalt wedge (e.g., the Columbia River, WA, USA [13], and theFraser River, BC, Canada [11]; Fig. 2). Changes in interstitialsalinities related to overlying water salinities are a function ofboth sediment type and duration of exposure. Figure 3 illus-trates three scenarios: sediments that are primarily sands andsediments that are primarily silts and clays, both with highlysaline interstitial salinities exposed to overlying freshwater,and the same exposure involving sediments that are primarilysilts and clays but with intermediate interstitial salinities. Ex-change and equilibration between interstitial and overlyingwater is fast in sands but slow in sediments containing high


proportions of silts and clays. Surface sediments are morerapidly affected than bottom sediments, and differences canpersist for days between interstitial salinities at different sed-iment depths for sediments with high proportions of silts andclays. These patterns are related not only to differences be-tween overlying and interstitial salinities but also to differencesin the porosity and permeability of sediments [31,32].

Porosity, the fraction of sediment volume occupied by wa-ter, usually decreases exponentially with depth because of com-paction [33]. Permeability represents a proportionality factorbetween water pressure gradient and water flow. Voids betweenlarger particles are more available in sands than in muds. Inthe latter case, these voids tend to become clogged by finersediments, thus reducing permeability and the mixing of in-terstitial and overlying waters (Fig. 3).

Salinity as a controlling factor for contaminantpartitioning and bioavailability

Unlike freshwaters, where pH is the controlling factor, inestuaries salinity is the controlling factor for the partitioningof contaminants between sediments and overlying or intersti-tial waters. The partitioning coefficient (Kd) of a contaminantis defined as the ratio of a contaminant concentration in thesediment to that dissolved in the overlying or interstitial water.High ionic strengths in estuarine waters can salt out hydro-phobic organic chemicals from the water to the sediment phase[34,35]. In addition, increasing salinity enhances the removalof dissolved organic matter from the water to the sedimentphase and the formed particulate organic matter can effectivelysorb hydrophobic chemicals [35]. As a result, an increase insalinity generally results in an increase in Kds for hydrophobicchemicals. In contrast, Kds for metals may decrease (e.g., Cd,Zn), increase (e.g., Fe), or be constant (e.g., Ir) when salinityincreases, depending on the relative importance of the twocounteractive processes [36–38]: (1) desorption due to increas-ing complexation with seawater anions (Cl2 and ) and/or22SO4

increasing competition for particle sorption sites with seawatercations (Na1, K1, Ca21, Mg21) and (2) coagulation, floccula-tion, and precipitation. Light rare earth elements (e.g., La,associated with zeolites from oil refineries) tend to be lessefficiently trapped by low-salinity sediments [39].

Because it affects the partitioning of contaminants betweensediments and overlying or interstitial water, salinity also af-fects the bioavailability of contaminants in estuarine sedi-ments. For example, Schlekat et al. [40] found that bioavail-ability of cadmium associated with bacterial exopolymer sed-iment coatings to the amphipod Leptocheirus plumulosus wasdependent on both seawater salinity (greatest at estuarine sa-linities) and cadmium concentrations. Partitioning to particlesfavors uptake by sediment feeders [41], whereas desorption tothe water (overlying or interstitial) favors uptake via dermalexchange surfaces such as gills. However, while desorption ofmetals with increasing salinity can increase metal concentra-tions in water, concurrent increases in Ca21 and Cl2 can de-crease water bioavailability [42,43].

It is well established that dissolved organic matter can affectthe 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 importantrelated to salinity [48,49]. Exposing freshwater organisms tosome contaminants (e.g., metals other than mercury) in watersof increasing salinity (below toxicity thresholds for salinity)demonstrably increases the toxicity threshold for such con-

taminants [50–52]. Similarly, decreasing the salinity contentof exposure waters increases the toxicity of metals such asnickel, zinc, and chromium to saltwater species [53,54].

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

Salinity as a controlling factor for estuarine biota

Salinity is also the most important natural factor controllingthe distribution of estuarine organisms [23,24,58]. This is truenot only for water column biota but also for benthic biota [59],even though some researchers refer to salinity (together withdepth and sediment grain size) as nuisance variables [60]. Thebottom line is that benthic organisms burrowing in the sedi-ments can be exposed to very different salinity regimes thanif they were on the sediment surface. The extent of exposuredifferences depends on factors such as sediment type, the dif-ference between overlying and interstitial water salinities, du-ration of exposure to different overlying water salinities, andthe organism’s depth in the sediment. This has significant im-plications both to the distributions of organisms in such es-tuarine areas (see Assessment Techniques: Biological Surveys)and to sediment toxicity tests of these sediments (see Assess-ment Techniques: Toxicity Tests). Further, not all species ofthe same estuarine genus have similar salinity tolerances [27].

Salinity, of course, can be toxic. In many ways, estuariesare not dissimilar to freshwaters subject to nonpoint-sourcerunoff. In both cases the biota are exposed in a nonconstantmanner 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 laboratorytoxicity tests involving constant exposures, effectively a snap-shot in time, cannot adequately predict in situ toxicity. De-pending on the situation, the laboratory may under- or over-predict toxicity [61]. There is no reason to expect this situationwould be any different in estuaries. In fact, there is everyreason to expect a similar situation.

ASSESSMENT TECHNIQUES: CHEMISTRY

From a chemical standpoint, an estuary is a reaction vesselwhere chemically very different fresh and salt water are dras-tically mixed. Contaminants in estuaries are mainly transportedfrom rivers and/or from direct effluents located on or nearestuaries. Their mixing with salt water is often far beyond asimple dilution (behaving nonconservatively [62]). An in-crease of ionic strength from approx. 0 to about 0.7 mol/L,together with the change in water composition, has a widearray of influences on transport and transformation processesof contaminants in estuaries, including adsorption or desorp-tion, coagulation, flocculation, and precipitation and biotic as-similation or excretion. Depending on the nature of the con-taminant and on the estuarine condition, removal from waterto sediments (sediments as a sink) or addition to water fromsediments (sediments as a source) can occur during the mixing.Estuarine sediments are efficient and effective traps for hy-drophobic chemicals mainly because of the salt-out effect[35,63]. Although desorption of metals from sediments canoccur during estuarine mixing, release of metals to water iscounteracted by enhanced flocculation [36]. Hence, estuarinesediments serve as a filter for many contaminants betweenland and sea [64]. Only those metals that form very strong


complexes [65] and organic chemicals that are less hydropho-bic [66] may be transported out of estuaries to the ocean. Thisnatural mechanism renders estuaries more susceptive to con-tamination.

Grain size effects

Chemistry-based approaches for assessing sediment con-tamination are based on reliable measurement and interpre-tation of contaminant concentrations in the sediments. Whilethe overlying water in estuaries can be heterogeneous becauseof different mixtures of fresh and saline water (Fig. 1), a muchhigher degree of heterogeneity and variability exists withinestuarine sediments not only because of the salinity differencesin the pore waters but also because of the diverse and com-plicated composition of the sediments. Different sediments canhave significantly different capacities for collecting contami-nants. For instance, the grain size distribution of a sedimentis probably the most important factor controlling sedimentmetal concentrations; correlations commonly exist between de-creasing grain size and increasing metal concentrations[67,68]. Hence, any assessment approach based on sedimentmetal concentrations needs to consider grain size differencesbetween sediment samples. If such differences are not takeninto account, apparent concentration differences between sitesmay reflect differences in grain size rather than the extent ofcontamination.

Numerous efforts have been made to normalize grain sizeeffects relative to sediment contaminant concentrations. Theseefforts can be generally grouped as physical separations (mea-suring sediment contaminant concentrations in selected, phys-ically separated grain size fractions of the sediment) and math-ematical normalizations (measuring contaminant concentra-tions in whole sediments followed by normalizing to appro-priate sediment constituents). In the physical separationmethod, the ,63-mm fraction is favored by many authors[69,70]. However, problems with the efficiency and feasibilityof mechanical separation in many cases (particularly whensample sizes are small) render mathematical normalizationsthe best way to correct for grain size effects [70].

Many sediment constituents have been used for normali-zation of grain size effects in estuarine sediments. These in-clude 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 pro-cedure generally involves calculation of ratios or regressionanalysis [67], with the latter being preferable [80]. Amongregression analyses, robust regression, such as least-absolute-values regression, may be more powerful and less subjectivethan least-squares regression [80].

The most important factor controlling concentrations of hy-drophobic organic chemicals in sediments, however, appearsto be the fraction of organic carbon [81,82]. Normalization tothe organic carbon fraction is hence necessary when comparingconcentrations of hydrophobic chemicals in different estuarinesediments.

Background enrichment

Background enrichment assesses the relative contributionof anthropogenic sources of a substance by comparing its con-centration at a site to its natural background concentration.This approach applies only to naturally occurring substancessuch as metals and metalloids. Since this approach does not

take into account the biological effects of the substance, en-richment does not necessarily imply biological effects.

The simplest approach for assessing enrichment is a pointcomparison, which takes the form of the ratio of the sedimentcontaminant concentration at the site of concern to that at areference site [83]. Prior to calculating the ratio, both con-centrations need to be corrected for grain size effects unlessthe sediments are of similar composition. For such a pointcomparison to be meaningful, the geological setting of thereference site must be representative of the site of concern.Such a point comparison may be practicable, at best, onlywithin small areas of fresh and marine sediments [84]; it isoften not feasible in estuaries. Since the bulk of the sedimentin an estuary can be derived from areas with different geology(e.g., the sediments in the upper reaches may be fluvial inorigin, those in the lower estuary largely of marine provenance,and those between may comprise different mixtures of fluvialand marine sediments), it is often impossible to obtain a singlerepresentative reference site, even for a small area of the es-tuary.

Alternatively, a line comparison can be conducted for es-tuarine sediments. For example, in assessing the spatial scaleof trace metal contamination in estuarine and coastal sedimentsin the United States, Daskalakis and O’Connor [85] developedbaseline regression lines between concentrations of trace met-als and a conservative element (Fe or Al) from sites relativelyfree of human influence. Concentrations of trace metals andthe conservative element from other sites were then comparedwith the baseline regression lines to determine whether theywere enriched.

While the enrichment approaches have no predictive powerfor biological effects, they fully consider background concen-trations of naturally occurring substances and provide impor-tant insights into sources of contamination, both of which areoften underestimated or ignored by other approaches. Delin-eation of the extent of metal enrichment may also be used asa screening tool to promote cost-effective use of sedimenttoxicity tests (e.g., only those with metal concentrations ex-ceeding expected natural background ranges will be subjectedto further toxicity tests [77]).

Bioavailability

The total concentration of a contaminant in sediments canbe conveniently measured; however, it alone provides littleinformation as to the possible biological effects of that con-taminant [45]. Bioavailability of sediment-bound contaminantsis determined by sediment constituents, overlying and inter-stitial water chemistry, and the behavior of organisms [78].Main binding phases include organic carbon for hydrophobicchemicals [81] and sulfide, organic matter, and iron and man-ganese oxyhydroxyides for metals [45].

No consensus exists as to appropriate analytical methodsfor determining contaminant bioavailability in estuarine sed-iments [86]. Chemical extractions of sediments have been tra-ditionally used. For example, metal concentrations in estuarinebenthic invertebrates have been found correlated to dilute acidextractable metals [83,87], metal/Fe, or metal/organic matterratios [83] in sediments. Schlekat et al. found that metals as-sociated with bacterial exopolymer sediment coatings wereassimilated by the estuarine amphipod Leptocheirus plumu-losus with higher efficiency than those associated with recal-citrant organic carbon, iron oxides, or phytoplankton [40,88].Recent studies, however, indicate that the bioavailability of


sediment-bound contaminants is more closely related to thedigestive systems of benthic animals. For example, in vitrodigestive fluid extractions have been shown to provide a betterindicator of the bioavailability of sediment contaminants thanchemical extractions [89–93].

Equilibrium partitioning models have been postulated toestimate the bioavailability of contaminants in fresh and saltsediments, for example, the organic carbon sorption model forhydrophobic organic chemicals [81], the iron and manganeseoxyhydroxyides and organic carbon model for metals in anoxicsediments [94], and the acid volatile sulfide/simultaneouslyextracted metals model for metals in anoxic sediments [95].These models, however, may have little application to estuarinesediments since the very dynamic physical and biogeochemicalnature of in situ estuarine sediments overturns the fundamentalassumption (a quasi-equilibrium state being achieved betweencontaminants in sediments and in water) involved in thesemodels. River flow [96], tidal flushing [97], and other sedimentresuspension events [98] can have significant influences on thepartitioning and bioavailability of contaminants in estuarinesediments. In addition, as noted earlier, salinity also signifi-cantly affects bioavailability.

Even if an equilibration of contaminants between sedimentsand interstitial water exists in estuaries, which is unlikely,bioavailability of sediment-bound contaminants is not neces-sarily attributable only to interstitial water. For example, ex-posure of many estuarine and coastal deposit feeders may occurprincipally through ingestion of particles, and hence metalbioaccumulation cannot be fully accounted for by the acidvolatile sulfide/simultaneously extracted metals model [87].Exposure of benthic organisms may also occur through bioir-rigation of the overlying water, as demonstrated for manyfreshwater tube-dwellers [99,100] as well as estuarine benthicinvertebrates [101].

Sediment quality values

Sediment quality values (SQVs) can be used to screen sed-iment contamination by comparing sediment contaminant con-centrations with the corresponding SQVs. The feasibility andreliability of such comparisons are, of course, greatly depen-dent on the availability and reliability of SQVs. Deriving SQVsrequires integrated information on sediment chemistry, toxi-cology, and biology [102]. While dozens of approaches havebeen employed for deriving SQVs for fresh and marine sed-iments [103], no specific efforts have been made to developSQVs for estuarine sediments. Many jurisdictions classifyaquatic ecosystems into either freshwater or marine ecosys-tems, ignoring the unique characteristics of estuaries. For ex-ample, databases for estuarine and saltwater sediments areoften lumped to derive SQVs for marine sediments [104]. Sim-ilar problems exist with water quality guidelines (WQGs),which have been traditionally derived for fresh and marinewaters. The importance of deriving specific WQGs (waterquality guidelines) and SQVs for estuaries was recently rec-ognized by Australian and New Zealand Environment andConservation Council (ANZECC) & Agriculture and ResourceManagement 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 chemicalstressors (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 anarea for further research [105].

Interstitial water chemistry

Measurements of interstitial water chemistry assess sedi-ment contamination by measuring contaminant concentrationsin the interstitial water rather than in the sediments, followedby a comparison of the measured interstitial water concentra-tions with WQGs or other toxicity thresholds. While theseapproaches effectively eliminate grain size effects associatedwith sediment chemistry, several key issues remain. First, asnoted earlier, interstitial water chemistry may not fully accountfor biological effects on benthic organisms, as fractions ofcontaminants in sediment particles [87] and/or in the overlyingwater [99,100] may also be bioavailable to benthic organisms.Second, not all the dissolved contaminants in interstitial waterare bioavailable. For example, the most bioavailable forms ofcontaminants in the interstitial waters are thought to be freemetal ions for many metals [45,95] and free dissolved fractions(not complexed by dissolved organic carbon) for hydrophobicorganic chemicals [81]. Above all, comparing estuarine inter-stitial water concentrations with saltwater or freshwater WQCsis, at best, questionable because of the differences in bothchemistry and biology between the interstitial and overlyingwater as well as the differences between estuarine waters com-pared to salt and fresh waters. Salinity or, preferably, salinityprofiles of interstitial water should always be measured andreported.

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

ASSESSMENT TECHNIQUES: BIOLOGICAL SURVEYS

Types and distribution of estuarine benthos

Estuaries typify the paradox of brackish water [108]. Spe-cifically, although for most ecological factors the largest num-ber of species occurs at intermediate values, this is not thecase for salinity. In the case of salinity, the greatest numberof species occurs in fresh and in marine waters, with fewernumbers of species at intermediate salinities (Fig. 4). Thisparadox appears to be due to the unstable and unpredictablebehavior of estuarine environmental factors that decrease theprobability of speciation and increase the probability of ex-tinction (particularly for estuaries since they are geologicallyephemeral phenomena) while excluding most marine andfreshwater species [25,109]. Estuarine salinities are generallyconsidered to be above about 1 and below about 30 g/L (fresh-water organisms can generally survive, grow, and reproducein a few g/L salinity; similarly, salinities of 30 g/L do notconstitute a major hardship for most marine organisms). Es-tuarine benthos species richness or diversity are least withina critical salinity range of about 5 to 8 g/L, which reflects the


Fig. 4. Illustration of Remane’s [108] ‘‘paradox of brackish water.’’Species numbers and diversity are lower in estuarine than in fresh ormarine waters.

inabilities of many organisms to tolerate salinity stress and toundergo extensive cell volume regulation [59].

Conflicting concepts exist regarding estuarine biologicalcommunities [110]. For instance, Remane [108] consideredbrackish waters as a distinct biological dominion, situated be-tween marine and freshwater dominions, with a unique mixof species. In contrast, Barnes [111] denied the existence ofan identifiable brackish assemblage and suggested that the es-tuarine environment is simply populated by a small group ofmarine species. Based on comparative analyses of very finemorphophysiological changes and genetic analyses, Cognettiand Maltagliati [110] recently concluded that the estuarineenvironment should not be considered as a marginal, transi-tional environment but rather as a single and well-defined hab-itat with its own fauna. They also concluded that in brackishwaters, a given species of marine origin often consists of manydifferent forms at various levels of differentiation.

Estuarine benthic species tend to be r-selected. They typ-ically exhibit low species diversity, small body size, short lifecycle, early reproduction, rapid development, variable popu-lation size, density-independent mortality, low competitiveability, and high reproduction [112]. K-strategy benthic spe-cies, however, can also be found in brackish environments[110].

Many estuarine benthic communities exhibit symptoms ofdisturbance without anthropogenic influence [113,114]. Underextreme physical and chemical stresses, such as unstable sed-iments and periodic changes of anoxic and oxic conditions,only a few species can survive. In quiet, muddy, organicallyrich environments, the benthos is typically dominated by smalldeposit 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, Mol-lusca, and Polychaeta are well represented, as are opportunisticspecies such as Mulinia lateralis, Capitalla capitata, and Po-lydora ligni, whose densities fluctuate greatly [59]. Truly es-tuarine species are generally restricted to between about 3 and20 g/L salinity, with penetration of estuarine sediments by bothfreshwater and euryhaline opportunists (Fig. 4). As noted pre-

viously, salinity is the most important natural factor controllingthe distribution of estuarine organisms. However, organismsburying into muddy sediments are directly affected not byoverlying water salinities but rather by interstitial salinities,which change much more slowly in muddy than in sandy sed-iments ([32]; Fig. 3). In salt wedge estuaries, seasonal varia-tions in interstitial salinities occur, resulting in seasonal shiftsin benthic infaunal distributions. Basically, the range of oli-gohaline, mesohaline, and euhaline species in salt wedge es-tuaries varies seasonally, with oligohaline populations extend-ing their range seaward during periods of high freshwater flowand low tidal influence and the reverse occurring during pe-riods of low freshwater flow and high tidal influence relatedto both tidal and riverine transport [115]. These species shiftshave been shown to cover distances exceeding 10 km in atleast one salt wedge estuary ([11,31]; Fig. 2). Populationsinvolved in such species shifts do not reproduce throughouttheir entire estuarine range but rather are sustained by periodicrecruitment (drift or migration, depending on the species) fromcenters of abundance to the extremes of the population’s es-tuarine range. Such species shifts are dependent on the avail-ability of larvae, juveniles, or adults in the water column;suitable conditions for settlement (low current velocities andsubstrate); and suitable conditions for continued survival andresidence [12]. The latter requirements include suitable salin-ity, habitat, and low bottom shear stress. Depending on species-specific salinity tolerances, in some cases immature organismsmay not survive in areas where both mature and immatureorganisms are deposited.

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

Hypoxia is also a feature of some estuaries, related to cir-culation of bottom waters, which are primarily saline[116,117]. Effects of hypoxia on estuarine benthos, indepen-dent of other stresses (e.g., variable salinity, anthropogenicinputs), will change community composition and reduce di-versity and biomass [117–119].

Biological surveys of estuarine benthos

Benthic organisms are relatively sedentary (avoidance re-sponses are limited), have relatively long life spans (indicateand integrate conditions), comprise different species or func-tional groups with different tolerances to stress, can be com-mercially important or are important food sources for eco-nomically or commercially important species, and tend to havean important role in cycling nutrients and contaminants be-tween the sediments and water column [120]. Thus, a greatdeal of effort has been spent on evaluating benthic communitystructure in estuaries in relation to anthropogenic contami-nation as well as natural factors. A focus has been attemptsto develop benthic community indices particularly related tosediment quality values and sediment toxicity test results[120–126]. However, most authors of studies developing in-dices acknowledge that these are primarily management tools,not definitive assessment tools.

As previously noted, estuaries tend to be unique physicallyand chemically. They also tend to have relatively low speciesnumbers 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 bi-ological surveys in truly estuarine sediments particularly dif-ficult for several reasons. First, reference sites may be so dif-ferent from exposed sites as to render comparisons meaning-less, and gradient approaches must deal with salinitydifferences as well as the usual confounding factors (e.g., sed-iment grain size, total organic carbon [TOC]). Second, sea-sonal variations related to freshwater discharge and salinityintrusions render the benthos more variable than in consistentlymarine or freshwater sediments.

Thus, it is not surprising that Hyland et al. [126], in at-tempting to predict stress to benthic communities in south-eastern U.S. estuaries, found a correlation between degradedbenthos and lower salinity (as well as to higher contaminantlevels, muddier sediments, higher TOC levels, and slightlygreater depths). Hyland et al. [126] also report a lack of con-cordance between sediment toxicity tests using marine organ-isms at marine salinities and estuarine benthos living at muchlower salinities. It is entirely possible that concordance wouldhave improved if estuarine sediments had been tested at theirin situ salinities, using estuarine organisms. Similarly, esti-mates made by Long et al. [127,128] of the spatial extent ofsediment toxicity in U.S. estuaries might have been very dif-ferent if the sediments had not been tested at marine salinitiesusing marine organisms (see Assessment Techniques: ToxicityTests). Detailed, extensive toxicity testing of San FranciscoBay sediments observed wet-dry seasonal fluctuations in bothsediment contamination and toxicity, which the authors as-cribed to estuarine processes [129].

Ecological stress, from any source, is best measured usingmultiple variables, methods, or analyses and not necessarilylimiting these to the same or a few assumptions. Particularlyuseful are combinations of sediment toxicity, contamination,and estuarine community structure [130–132], and in estuarinesediments, natural variability in salinity must, in contrast tofreshwater or marine sediments, be considered and accountedfor. This has been done by some authors [123–125,133], butnot all. A further problem is that the tolerances of many es-tuarine organisms to contaminants are relatively unknown ascompared to fresh and marine ecosystems.

Finally, sufficient data are required to adequately determinedifferent habitats and to determine the status, separately, ofeach of these habitats, which, of course, are specific to indi-vidual estuaries. For instance, Weisberg et al. [124] definedseven habitats in Chesapeake Bay (MD, USA) based on sa-linity and substrate. In some estuaries, waves, currents, anddepth are also important factors defining benthos habitats; inall estuaries, biotic interactions (competition, predation) areimportant [59]. Given the complexity and diversity of estu-aries, accurate and precise generic indices are difficult withoutaccounting for habitat differences. This requires very largedata and information sets that are not presently available formany estuaries. Thus, on a broad scale, only relatively sim-plistic indices are possible (e.g., northern Gulf of Mexico es-tuaries [125]). Such indices cannot distinguish between naturaland anthropogenic stresses. For instance, they cannot by them-selves distinguish hypoxia that is anthropogenic in origin fromnatural hypoxic events.

ASSESSMENT TECHNIQUES: TOXICITY TESTS

Sediment toxicity tests can be conducted using a variety ofexposure techniques: whole sediments, pore (interstitial) wa-

ters, elutriates, and extracts. Whole sediment exposures allowfor the widest variety of possible exposure routes and resultin the least changes to sediment physicochemical conditions.They also comprise the primary tool for sediment toxicityassessments [134,135]. Thus, this section focuses on wholesediment toxicity tests.

Single-species whole sediment toxicity tests

Estuarine sediments vary considerably with regard to sa-linity (interstitial and overlying), as previously noted. Theyalso vary considerably with regard to particle size, tempera-ture, and TOC content. In some cases, these differences havebeen ignored, and sediment toxicity has been evaluated usingmarine species exposed to interstitial or elutriate waters [135–138] or to sediments whose salinity has been artificially in-creased [126–128,139]. In other cases, freshwater species havebeen exposed to estuarine sediments overlain with freshwater[56]. In all these cases, potential differences in contaminantbioavailability and toxicity due to the original, unadjusted in-terstitial salinities (see Assessment Techniques: Chemistry)have not been accounted for. Further, such manipulations tendto dilute the contaminant load. The only reasonable way todetermine the toxicity and bioavailability of contaminants inestuarine sediments is to test the sediments as received, usingestuarine organisms capable of tolerating the full range ofestuarine conditions, in particular salinity.

Whether estuarine species in general are more or less tol-erant than freshwater or marine species is a topic of debateamong ecologists. It has been suggested [140,141] that becauseestuarine 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 speciesmust 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 tox-icity tests indicate that estuarine species can be as sensitiveas freshwater or marine species [143,144] and presumably alsoas variable in their relative sensitivities to different toxicants.

Single-species estuarine whole sediment toxicity tests canbe divided into two basic categories: tests not originally in-tended for but adapted for estuarine conditions (first genera-tion) and tests designed specifically for estuarine conditions(second generation). The latter tests in particular include bothorganisms tolerant of the entire range of estuarine salinities (fromfresh to oceanic) as well as organisms that can tolerate only alimited range of estuarine salinities. Single-species tests for thetoxicity of whole estuarine sediments are listed in Table 2.

Some interesting and puzzling observations can be madebased on Table 2. In particular, the importance of salinity inconducting sediment toxicity tests is not fully recognized.There are few whole organism tests presently in use for es-tuarine whole sediment toxicity tests that can survive the fullrange of estuarine salinities. There are none that can success-fully reproduce across the full estuarine salinity range. In ad-dition, it appears that not all authors appreciate the importanceof detailing their test species’ tolerances to salinity (or to othermodifying factors such as sediment grain size or TOC). Thisis particularly the case for first-generation tests adapted fromfresh and marine sediment tests in which interstitial salinitieswould normally be ignored. For estuarine sediments, intersti-tial salinities are sometimes measured and adjusted throughaddition of higher or lower salinity water either directly (mix-ing the sediments in different salinity waters) or indirectly


Table 2. Estuarine Sediment Toxicity Test Organisms. Single-species, whole sediment exposures

Test organism Test end point(s)

Testgenera-

tionaSalinity range(survival, g/L)

Salinity range(reproduction,

g/L) References

Fish (spot)Leiostomus xanthurus

Survival, fin erosion, lesions Second Estuarine; collected at 20,held and tested at 15

[221,222]

Amphipod (crustacean)Hyalella azteca

Survival, reproduction First 0–17 ,10 [56,223,224]

Amphipod (crustacean)Eohaustorius estuarius

Survival, avoidance Second 0–34 Unknown [143,146,223,225]

Amphipod (crustacean)Gammarus duebeni

Pleopod beat frequency, swimmingendurance

Second ‘‘A truly estuarine species’’;held and tested at 15

[226,227]

Amphipod (crustacean)Leptocheirus plumulosus

Survival, growth, reproduction Second 0–33 5–20 [146,152,153,223,228]

Amphipod (crustacean)Lepidoctylus dytiscus

Survival, growth Second No information; held andtested at 15

[155]

Amphipod (crustacean)Melita nitida

Survival, reproduction, abnormalbrood pouch setae, intermolt peri-od

Second Typically found in salinitiesat 3 to 20

[229]

Amphipod (crustacean)Corophium mutisetosum

Mortality, growth Second 0–35 1–15 [230]

Amphipod (crustacean)C. sp.

Survival Second 0.1–24 1–151 [144]

Mysid (crustacean)Mysidopsis bahia

Mortality First No information; held andtested at 21

[139]

Waterflea (crustacean)Daphnia magna

Mortality, reproduction First 0–,5 0–,5 [231]

Grass shrimp (crustacean)Palaemonetes pugio

Mortality First No information; tested at 20 [232]

Copepod (crustacean)Amphiascus tenuiremis

Mortality, reproduction First No information; tested at 28 [205,232]

PolychaeteStreblospio benedicti

Survival, growth Second No information; held andtested at 15

[155]

MolluskTapes semidecussatus

Survival, behavior Second Described as ‘‘true estuarinespecies’’ but held and test-ed at 24

[233]

MolluskScrobicularia plana

Survival, behavior Second

MolluskMya arenaria

Burrowing speed Second Tested at overlying water sa-linities varying from 2 to32

[234,235]

MolluskCrassostrea gigas

Pediveliger larvae survival, settle-ment and metamorphosis

Second Survival: .5; metamorpho-sis: .23

[199]

BacteriaMicrotox

Enzyme function First 0–351 0–351 [135,173,236,237]

a Test generation—first: techniques and/or organisms originally developed for saline or freshwaters; second: techniques and/or organisms spe-cifically developed for estuarine conditions.

(changing the salinity of overlying waters) [133,145]. Pro-longed exposure to overlying waters of different salinity willchange the salinity of interstitial waters even in low perme-ability muds (Fig. 3). Salinity variations can also change thebioavailability and route of exposure of contaminants in thosesediments (see Unique Physicochemical Characteristics of Es-tuaries). To our knowledge, there have been no publicationsin which the effects of adjusting interstitial salinities on bio-availability and toxicity have been directly assessed (thoughthere have been such studies done in water-only situations).Similarly, there have been few publications of studies attempt-ing to maintain similar interstitial and overlying water expo-sures in the laboratory as found in situ. Instead, there arearguments presented for testing under defined conditions, in-cluding salinity [56,146,147].

Given that interstitial salinities can be very different fromoverlying salinities for truly estuarine sediments, both shouldalways be measured prior to testing (during sample collection).They should also be measured during testing. Such measure-ments are important not only in terms of test organism sen-sitivities 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 contaminantspose difficulties [148,149], pore water for salinity analyses canbe adequately collected relatively simply by squeezing or cen-trifuging sediments [31,32], though this does effectively de-stroy natural salinity gradients in the sediment.

Ideally, estuarine testing should be conducted with a rea-sonable knowledge of the estuary in question and in particularof the potential range of seasonal interstitial salinities at alltest sites. Again ideally, testing should cover the range fromworst to best cases related to salinity effects on contaminantbioavailability. At the very least, the effects of varying salin-ities on bioavailability need to be considered when interpretingtesting done at a specific salinity (a snapshot in time of adynamic and unique system). McGee et al. [150] measuredinterstitial salinities and provided matching overlying watersalinities in Leptocheirus plumulosus tests of Baltimore Har-bor (MD, USA) sediments. This is one of the only studies ofestuarine sediment toxicity to have taken the approach of mea-suring and matching natural salinity conditions rather thanignoring them and/or imposing arbitrary salinity conditions.


Another interesting observation from Table 2 is that sedi-ment tests of estuarine sediments have, to date, been heavilyweighted to crustaceans, in particular, amphipods. Of the 19different estuarine test species in Table 2, there are 12 crus-taceans (of which eight are amphipods) and a fish, a poly-chaete, four mollusks, and a bacterium. Other organisms re-ported in the literature as used in whole sediment toxicity testswith estuarine sediments are marine rather than estuarine andrequire marine salinities during testing (e.g., the polychaeteArenicola marina [135]). Phylogenetically, the list of availabletest species is more limited than should be the case. It alsocomprises few examples of true deposit feeders, such as themollusks Tapes semidecussatus and Scrobicularia plana. Fur-ther, estuarine whole sediment toxicity tests appear to be betterdeveloped as follows: North America . Europe . Australia. rest of the world.

DeWitt et al. [143] note that, ideally, estuarine test organ-isms should live at or below the sediment/water interface andhave broad salinity tolerance, high sensitivity to sediment con-tamination, low mortality under reference conditions, low sen-sitivity to modifying factors such as sediment particle size andTOC, and a broad geographic range. They should also be ame-nable to handling in the laboratory, be ecologically importantin estuarine systems, and either be readily collected from thefield or amenable to laboratory culture. DeWitt et al. [143]favor use of amphipods as test organisms because of theirrelative sensitivity, as a group, to toxicants. The majority ofestuarine sediment toxicity test organisms used to date are infact amphipods (Table 2).

The amphipod tests are the most widely applied toxicitytests performed with estuarine sediments in North America[127,128]. Two of the most commonly used estuarine amphi-pods in such tests are Eohaustorius estuarius (free burrowing)and L. plumulosus (open tube dweller) [146,147]. These am-phipods are found, respectively, on the west and east coastsof North America. Both species can be collected in large num-bers from the field. Ten-day acute toxicity tests with these twospecies provide reasonable discrimination of contaminatedsediments except where sediment contamination is relativelylow, and this problem can be minimized by increasing thenumber of test replicates [151]. However, to date only L. plu-mulosus is readily amenable to laboratory culture such that,together with its relatively short life history (30–40 d com-pared to annual for E. estuarius), it can be used in chronictesting 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 organiccontaminants is unknown. Such is also unknown for manycontaminants for all sediment test organisms (freshwater, es-tuarine, or marine). Investigators need to ensure that test or-ganisms are appropriately sensitive to contaminants of concernin sediments if meaningful results are to be derived.

Not all possible amphipods or other estuarine test specieshave been used for toxicity tests with estuarine sediments. Forinstance, DeWitt et al. [143] suggested that four haustoriidamphipods be considered as good candidates for sediment tox-icity tests in addition to E. estuarius: Haustorius canadensis,Neohaustorius schmitzi, Lepidactylus dytiscus, and Lepidac-tylus triarticulatus. Of these, to our knowledge, only L. dy-tiscus has been used for estuarine sediment toxicity tests, andonly in Chesapeake Bay [155], and this amphipod is now nolonger being used in Chesapeake Bay because of its preference

for sandy rather than muddy sediments (J. Winfield, personalcommunication). Similarly, although amphipods of the genusCorophium are typically estuarine, they have not been widelyused for estuarine sediment toxicity tests. The type species C.volutator, which has a wide geographic distribution and a widetolerance to salinity (tolerates 2–59 g/L, prefers 10–30 g/L[156,157], has been used primarily in marine sediment toxicitytests [158]. C. spinicorne has similarly been used in marinesediment toxicity tests [159]. C. triaenonyx, which toleratesthe full range of estuarine conditions, though reproduction islimited to 7.5 to 37.5 g/L [160], has not been used in estuarine(or marine) sediment toxicity tests. C. arenarium has beenused in tests with sediments the authors referred to as estuarinebut whose salinities were marine (generally 30–40 g/L [135]).Corophium are tube builders, living in U-shaped burrows thatarguably may not result in direct exposure to sediment inter-stitial waters but rather to overlying waters [99]. However, thesignificance of this fact remains to be evaluated. The widelyused estuarine sediment test species Leptocheirus plumulosuslives in open tubes, as does the less widely used spionid poly-chaete Streblospio benedicti. This latter species has also beenused more commonly in marine sediment bioassays than inestuarine sediment bioassays; in the former but not the latter,end points included reproduction [161]. However, the utilityof this polychaete is arguable since it is considered more tol-erant of contaminated sediments than many other species[162].

It should be noted that two other amphipods, Rhepoxyniusabronius and Ampelisca abdita, while often used in estuarinesediment toxicity tests [126,127,146,147,163], are indeed ma-rine or euryhaline. The former is found where in situ salinityis higher than 20 g/L [163], whereas the latter is found wherein situ salinity is higher than 10 g/L [164]. Caution is advisedwhen using results from such tests for assessing estuarine sed-iment contamination.

Chironomidae are not included among potential estuarinetest species. This is a surprising omission since a variety offreshwater chironomids live in natural salinity gradients rang-ing from 0 to almost 10 g/L [165] and since intertidal speciescan be naturally exposed to and tolerate the full gamut ofestuarine conditions (0 to $20 g/L salinity; P.M. Chapman,unpublished data). Further, tests with representatives of thisgroup of organisms are well developed [166].

Similarly, it is also surprising that aquatic oligochaetes arenot included among potential estuarine test species. Acute andchronic test procedures with freshwater species such as Tubifextubifex are well developed [167,168], and this and other fresh-water oligochaete species can survive in salinities up to about10 g/L [169,170]. Estuarine oligochaete species are amenableto acute toxicity testing and can survive across the full gamutof estuarine salinities [170,171]. Clearly there are opportunitiesto develop estuarine sediment toxicity tests using both chi-ronomids and oligochaetes.

As shown in Table 2, there is only one fish test included.However, this may be an artifact arising from authors notdistinguishing whether test sediments are estuarine or marine.For example, work by Nagler and Cyr [172] showing reducedhatching success for American plaice exposed to contaminatedsediments involved an estuarine fish and estuarine sediments.However, no information was provided on salinities other thanthat the sediments were considered marine by the authors.Also, testing of such mobile species must distinguish whetherthese organisms are likely residents or simply periodic ex-


ploiters entering the estuary from outside (e.g., with salt wedgeintrusions). Exposure to estuarine contamination will be verydifferent for periodic exploiters than for residents.

Testing using a toxicity test kit, such as Microtoxt (AzurEnvironmental, Carlsbad, CA, USA), clearly offers some ad-vantages in the estuarine environment. For instance, these testsare not directly affected by salinity. However, they are affectedby other factors; for instance, the Microtox (Azur) solid-phasetest shows decreased toxicity in finer sediments because thetest organisms (bacteria) are adsorbed to silt–clay particles[173]. It is to be expected that other toxicity test kits will beapplied to estuarine sediments. For instance, the Mutatoxybioassay (Azur) has not yet been applied to whole sedimentsbut has been applied to fractionated sediments [174]. The rel-evance of data from such tests to estuarine species and pop-ulations is arguable.

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

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

Potential confounding factors

Ammonia. Ammonia has been implicated as a source oftoxicity in some sediment toxicity tests [178–180] and hasbeen shown to interfere in 10-d but not 28-d tests with L.plumulosus [181]. Ammonia in sediments can originate fromanthropogenic effluents and wastes but can also originate fromnatural decomposition processes. Ammonia toxicity is influ-enced by both pH and hardness/salinity [182–184]. Thus, inestuarine sediments, if interstitial salinities change, it is likelythat the risk of ammonia toxicity will also be affected. At thistime, no broad generalizations are possible, particularly givendifferential sensitivities to ammonia by, for instance, differentamphipod species [185]. Ammonia can also produce analogouseffects with other contaminants such as silver and copper onnitrogen metabolism of freshwater and marine fish [186,187].

Sulfide. While the confounding factor of ammonia in sed-iment toxicity tests has been recognized by many studies, therole of sulfide in determining sediment toxicity has only re-cently been fully addressed [188]. Sulfide influences sedimenttoxicity in three major ways [188]: as a toxicant in its ownright, by reducing metal toxicity by forming metal sulfide sol-ids and/or complexes, and by affecting animal behavior, whichin turn can alter the toxicity of not only sulfide but also othersediment contaminants. Estuarine sediments, particularly thosein stratified estuaries such as salt wedge estuaries and fjords,are particularly rich in sulfide because of the high influxes ofboth organic matter (from river water) and sulfate (from sea-water). Sulfide levels as high as 100 mg/L are common inestuarine sediments [189,190]. Thus, any estuarine sedimenttoxicity tests that do not adequately consider sulfide effects

risk misestimating toxicity and misidentifying the causativeagents [188].

Grain size, organic matter, and other physicochemicalcharacteristics. Sediment physicochemical characteristicssuch as grain size and organic matter, in addition to modifyingthe bioavailability of contaminants (see Assessment Tech-niques: Chemistry), can also affect some organisms in sedi-ment toxicity tests. For example, the amphipod Rhepoxyniusabronius, often used in marine sediment toxicity tests, maysuffocate when attempting to burrow in fine-grained sediments[191]. Lacey et al. [192] showed that the growth of Chiron-omus tentans, often used in freshwater sediment toxicity tests,can be significantly affected by both the quantity and the qual-ity of organic matter. To reduce the impacts of these con-founding factors, the selection of an appropriate control sed-iment for sediment toxicity tests is critical. Ideally, the controlsediment should be identical to the test sediment in grain sizedistribution, organic matter content, and other physicochem-ical characteristics [192]. This is, however, almost impossible.Instead, development and usage of formulated reference sed-iments may assist in reducing the influence of these factors onsediment 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 forestuarine sediments. Some of the earliest such tests were con-ducted by Tagatz and coworkers [195,196], who developed alaboratory method for exposing previously frozen sedimentsto pelagic larvae and measuring differences in colonizationpatterns.

Austen and Somerfield [197] tested previously frozen con-taminated field sediments using a simple laboratory microcosmsystem (570-ml glass bottles) with addition of meiofauna-richsediment at 20-g/L salinity. After a two-month exposure pe-riod, meiofaunal community structure was determined andcompared to natural assemblages. These approaches are basedon the hypothesis that settlement and postsettlement stages offree-swimming benthic larvae are the most vulnerable life-cycle stages and most likely to be affected by anthropogenicpollution [198,199].

The possible implications on sediment structure and func-tion of freezing sediment to kill resident organisms remainsto be fully assessed. Thus, Watzin et al. ([200]; see the fol-lowing discussion) chose to use artificial sediments in theirspiking experiments.

Watzin et al. [200] exposed Zn-spiked artificial sedimentfor one week in the field in 100-cm2 containers and used theabundance and diversity (species richness) of colonizing spe-cies as indicators of sediment quality. In subsequent work,Watzin and Roscigno [201] applied this technique to two siteswith differing overlying water salinities (5 and 15 g/L) andrecorded resultant changes to the benthos, However, these au-thors were not able to separate out effects of site from effectsof salinity. Hall and Frid [202] undertook a two-year micro-cosm experiment in the lower Tyne Estuary (UK), dosing thesediments with copper and then determining faunal effects andpatterns of recovery. Interestingly, although chemical recoverywas rapid, faunal recovery was delayed because of differentialrecolonization rates. Millward and Grant [203] used pollution-induced community tolerance (PICT) to evaluate the biologicalimpact of chronic anthropogenic copper exposures on anotherUK 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-throughsediment microcosms containing their original meiobenthicfauna (primarily copepods) and also conducted spiked sedi-ment tests with chlorpyrifos for 21 d. They argue that suchtests are more conservative, informative, and predictive thanusing single-species sediment toxicity tests in isolation andargue further that both approaches should be used, collectingand culturing taxa and microcosms from actual field sites ofconcern.

Kurtz et al. [206] also used microcosms of estuarine sed-iments during 7-d measurements. However, their emphasis wason microbial communities. They point out that, despite theimportance of bacteria to aquatic systems and processes, thereis a lack of procedures to test bacterial responses to contam-inated estuarine sediments.

These experimental approaches retain more environmentalrealism than single-species toxicity tests while still permittingmanipulation in either laboratory or field settings. However,as noted by Chandler et al. [205], they are best interpreted incombination with well-designed single-species sediment tox-icity tests.

Laboratory-to-field comparisons

The laboratory does not and cannot mimic the field situationbut rather stands alone as an assessment tool [207]. In the caseof whole effluent toxicity (WET) tests, laboratory results canbe overprotective, underprotective, or of an unknown degreeof protection compared to the field situation [208]. The sameis probably true of estuarine sediment toxicity tests. For ex-ample, McGee et al. [209] found that field-collected L. plu-mulosus were typically more sensitive to cadmium than lab-oratory animals, although their sensitivities varied seasonally.However, in a comparison between L. plumulosus sedimenttoxicity test results and populations of these amphipods inBaltimore Harbor, Maryland, USA, McGee et al. [150] foundthat test results were predictive of population-level effects.Similarly, Swartz et al. [210] found that 10-d acute toxicitytests with E. estuarius provided reliable evidence of biolog-ically adverse sediment contamination in the field. McGee[211] has attempted to integrate laboratory sediment toxicitydata for L. plumulosus into a field-based population model toproject population-level consequences of sediment contami-nation. Further development of such a model, including othertest species, would greatly progress assessment of estuarinesediment contamination.

Difficulties in predicting what will actually occur in fieldsituations are exacerbated by interactions between biotic andchemical 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 thatfood limitations did not affect population dynamics at lowlevels of toxicant stress but that at higher toxicant exposurestoxicant effects were exacerbated. They suggest that given thatfood limitation in nature is the norm (whereas in the laboratoryfood is generally provided in abundance), the maximum tox-icant concentration at which populations can persist is lessthan previously believed. This experiment appears to refutethe hypothesis proposed by Calow et al. [213] and Grant [214]that conditions of high population density buffer negative tox-icant effects on population dynamics. However, whether the

responses of Capitella sp. M populations are typical remainsto be determined.

SUMMARY AND CONCLUSIONS: THE WAY FORWARD

Specific concluding comments are provided followed byrecommendations.

General comments

Estuaries are complex, dynamic, unique, and very individ-ual tidally influenced environments (change is the norm). Theyare defined primarily by their variable salinities, both overlyingand interstitial, and include fresh, marine, and truly estuarinefauna. In the context of geological time, they are also ephem-eral phenomena. This may partially account for the fact thatestuaries, in contrast to other intermediate ecological environ-ments, have fewer numbers of species than their borderingenvironments (fresh and salt water). Estuarine sediments orig-inate from fresh or marine waters but are not and cannot betreated as fresh or marine sediments. Estuarine physicochem-ical gradients (longitudinal, lateral, and vertical; salinity, tem-perature, pH, dissolved oxygen, redox potential, amount andcomposition of particles) tend to be strong and to influenceestuarine biogeochemical processes, including contaminantbinding and release from sediments. Salinity is the controllingfactor for partitioning of contaminants between sediments andoverlying or interstitial water and affects contaminant bio-availability.

Variable estuarine salinities provide stressful conditions forresident fauna whose survival strategies include avoidance,reduced contact, adaptation, and acclimation. Interstitial salin-ities tend to differ from overlying salinities, particularly in saltwedge estuaries, where seasonal, large-scale species shifts oc-cur. Specifically, freshwater species extend their range down-stream during periods when freshwater flow dominates overtidal intrusions; saltwater species extend their range upstreamduring periods when tidal intrusions dominate over freshwaterflow. Such species shifts can occur over longitudinal distanceson the order of 10 km.

Recommendations. All assessment techniques need to con-sider the unique and complex dynamic processes that char-acterize that estuary, in particular salinity effects. At a mini-mum, interstitial and overlying salinities need to be measuredduring sediment collection, and data interpretation should oc-cur in the context of knowledge of seasonal and spatial var-iations in both interstitial and overlying salinities as well asother key estuarine physicochemical processes.

Chemical assessments

Chemical assessment techniques need to consider grain sizeeffects, normalizing these relative to sediment contaminantconcentrations. Mathematical normalizations are preferredover physical separations. Other normalizations for grain sizeeffects are possible (e.g., surface area, organic carbon, con-servative elements) and may be useful in specific situations.

Background enrichment chemical assessments apply onlyto naturally occurring substances such as metals and metal-loids. Although not predictive for biological effects, they areuseful for assessing contaminant sources. In this regard, pointcomparisons are rarely possibly in estuaries because of thedifficulty in obtaining a representative reference site. Regres-sion line comparisons are the most appropriate approach.

There is no consensus as to the most appropriate methodsto determine bioavailability in estuarine sediments. In vivo


digestive fluid extractions offer advantages over traditionalchemical extractions. Bioavailability models may not be ap-plicable to estuarine sediments because the fundamental as-sumption of a quasi-equilibrium state is not valid. Such modelsinclude equilibrium partitioning, the acid volatile sulfide/si-multaneously extracted metals, organic carbon, and iron andmanganese oxyhydroxyides.

Sediment quality guideline values have been developed forfresh and marine waters. They have not been specifically de-veloped for estuarine waters. It is inappropriate to apply freshor marine values to estuaries. Interstitial water collection andanalysis for chemical contaminants is of questionable use inestuaries for reasons including inappropriate comparisons tofresh and marine water quality values and a lack of consensuson methods for collection.

Recommendations. Sediment chemistry should be normal-ized at least for grain size effects using mathematical tech-niques. For naturally occurring substances, regression linesshould be used to correct for background enrichment. Modelsand approaches specific to estuaries need to be developed toassess bioavailability. Similarly, sediment quality values spe-cific to estuaries need to be developed (and used only forscreening). The applicability and relevance of interstitial watercontaminant analyses in estuaries need to be determined.

Biological surveys

Much greater numbers of species occur in fresh and marinewaters than in estuarine waters. Estuarine species tend to ber-selected with two critical salinity ranges: about 5 to 8 g/Land about 15 to 20 g/L. Sediment infauna are not affected byoverlying salinities but rather by interstitial salinities, whichare only slowly affected by overlying salinities in the case ofmuddy sediments. As a result, salt wedge estuaries in particulartypically exhibit large-scale seasonal movements of fresh andmarine species up and down the estuary. Biological surveyswithin about 3 to 20 g/L are difficult because of such factors;the variable nature of the benthos makes reference comparisonsdifficult if not impossible.

Recommendations. Investigators conducting biological sur-veys need to understand their estuaries’ basic physicochemicalcharacteristics and the possible longitudinal, lateral, and sea-sonal variations that can be typical of estuarine benthic infau-nal populations. Habitat differences need to be understood anddocumented; in particular, overlying and interstitial salinitiesshould be measured when samples are collected and then re-ported. In many cases, gradient approaches will be more re-alistic than reference comparisons. Simple estuarine benthicindices can be useful for initial assessments; however, defin-itive assessments are best done using the sediment qualitytriad, which consists of chemical analyses, toxicity tests, andmetrics of benthic structure [131,215,216]. When constructingsediment quality triad triaxial plots for estuarine sediments,Del Valls et al. [217,218] noticed that the benthic communityalteration axis can be variable and indeterminate, possibly be-cause of salinity stress, emphasizing the importance of inter-stitial and overlying salinity measurements in estuaries.

Toxicity tests

Single-species whole sediment estuarine toxicity tests canbe either first generation (adapted for estuaries) or second gen-eration (designed for estuaries). Although there appear to bemore of the latter generation tests available, there are few teststhat span a large range of salinities. In too many cases, in-

vestigators have ignored salinity differences, conducting test-ing of estuarine sediments in fresh or marine waters with cor-responding nonestuarine organisms. Changing salinity canchange the bioavailability and route of exposure of sedimentcontaminants.

Laboratory toxicity tests involve a snapshot in time. Assuch, they cannot adequately predict in situ estuarine toxicitywhere major salinity variations occur in the sediments, forinstance, in the mesohaline portions of salt wedge estuaries.

Estuarine single-species tests to date have been heavilyweighted to crustaceans, particularly amphipods. There is notgood taxonomic representation of estuarine fauna among es-tuarine test species; some obvious candidate test species havenot been used. Further, the relative sensitivity of test speciesto particular contaminants is relatively unknown.

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

Community-level toxicity tests have been developed. Theserange from laboratory to in situ exposures and are arguablymore realistic than single-species tests. Laboratory single-spe-cies tests are least predictive of field responses.

Recommendations. Estuarine sediments (especially over therange of about 3–20 g/L salinity) need to be tested as received,measuring overlying and interstitial salinities during samplecollection and testing, and using estuarine organisms capableof tolerating those conditions. Studies should be conducted todetermine the effects of changing interstitial salinities on con-taminant bioavailability and toxicity relative to realistic tox-icity testing for estuaries that experience large-scale seasonalsalinity changes. A wider range of second-generation test spe-cies that are truly estuarine is required, and community-leveltesting should be further developed. The relative tolerances ofnew and existing test organisms to contaminants need to bedetermined. Most important, toxicity test approaches need tobe integrated into predictive population models.

The way forward

Within estuaries in particular, subsurface sediment contam-ination may not be isolated from biological organisms becauseof the dynamic nature of estuarine processes. There is a clearneed to tailor chemical, biological, and toxicological assess-ment techniques specifically for estuarine environments; freshand marine water techniques are not generally applicable toestuaries. However, this will not be possible immediately. Lessspecific and realistic methods will continue to be used, if onlyfor pragmatic reasons, such as cost. However, measuring in-terstitial and overlying salinities during collection of any andall samples (chemical, biological, toxicological) is neither dif-ficult nor expensive and should be immediately implemented.

Where detailed, specific assessments (e.g., use of the sed-iment quality triad, testing with estuarine fauna without chang-ing interstitial salinities) are not possible, a tiered approachshould be followed based on the ecological risk assessmentframework for sediments [219]. Simple techniques, such asbenthic indices, toxicity tests with marine or freshwater spe-cies, and altered salinities, should be considered solely aninitial assessment. Such approaches would be equivalent tothe problem formulation stage of an ecological risk assessment.Depending on the level of detail [220], more detailed studiescould be equivalent to either a screening level or a detailedecological risk assessment. Results should be reported andinterpreted within this context and framework. Given the dy-


namic nature of estuaries, studies that are only snapshots intime and that did not take into account temporal/seasonal var-iability would at best be considered equivalent to the screeninglevel of an ecological risk assessment.

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

Acknowledgement—We thank Steve Klaine for asking us to write thisreview paper, two anonymous reviewers, and EVS Consultants fortheir support. Some useful ideas originated via discussions with JoeWinfield. However, the opinions expressed herein are solely the re-sponsibility of the authors.

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Assessing sediment contamination in estuaries