relative role of pore water versus ingested sediment in bioavailability of organic contaminants in...

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Environmental Toxicology and Chemistry, Vol. 17, No. 12, pp. 2453–2462, 1998q 1998 SETAC

Printed in the USA0730-7268/98 $6.00 1 .00

RELATIVE ROLE OF PORE WATER VERSUS INGESTED SEDIMENT INBIOAVAILABILITY OF ORGANIC CONTAMINANTS IN MARINE SEDIMENTS

THOMAS L. FORBES,*† VALERY E. FORBES,‡ ANDERS GIESSING,†§ RIKKE HANSEN,† and LIV K. KURE††Department of Marine Ecology and Microbiology, National Environmental Research Institute, PO Box 358,

Frederiksborgvej 399, DK-4000 Roskilde, Denmark‡Department of Oceanography, University of Maine, Darling Center for Marine Research, Walpole, Maine USA

§Department of Life Sciences and Chemistry, Roskilde University Center, PO Box 260, DK-4000 Roskilde, Denmark

(Received 17 August 1997; Accepted 15 April 1998)

Abstract—Experimental data for fluoranthene and feeding selectivity in combination with reaction-diffusion modeling suggest thatingestion of contaminated sediment may often be the dominant uptake pathway for deposit-feeding invertebrates in sediments. Adietary absorption efficiency of 56% and accompanying forage ratio of 2.4 were measured using natural sediment that had beendual-labeled (14C:51Cr) with fluoranthene and fed to the marine deposit-feeding polychaete Capitella species I. Only 3 to 4% ofthe total absorption could be accounted for by desorption during gut passage. These data were then used as input into a reaction-diffusion model to calculate the importance of uptake from ingested sediment relative to pore-water exposure. The calculationspredict a fluoranthene dietary uptake flux that is 20 to 30 times greater than that due to pore water. Factors that act to modify orcontrol the formation of local chemical gradients, boundary layers, or dietary absorption rates including particle selection or burrowconstruction will be important in determining the relative importance of potential exposure pathways. From a chemical perspective,the kinetics of the adsorption and desorption process are especially important as they will strongly influence the boundary layerimmediately surrounding burrowing animals or irrigated tubes. The most important biological factors likely include irrigationbehavior and burrow density and size.

Keywords—Deposit feeding invertebrates Bioavaliability Pore water Ingested sediment Capitella species I

INTRODUCTION

The bioavailable fraction of a material is generally definedas that portion that is available for uptake by biota [1–3]. Thisdeceptively simple definition masks the difficulty of accuratelypredicting the exposure and bioavailability of organic contam-inants in sediments. Three factors are primarily responsiblefor the present degree of difficulty. First, the organic geo-chemistry of sediment is complex. The geochemical detailsconcerning both pore water and particulate organic matter arenot well understood, making predictions regarding chemicalspeciation and kinetics difficult [4,5]. The second difficulty isdue to the myriad ways organisms interact with and influencetheir local geochemical environment while making a living insediment. This strong interaction between an organism and thelocal geochemical environment means that organisms can in-fluence contaminant fate and therefore their own exposure inimportant ways. Third, biological variability is manifest in adiversity of biochemical responses that can come into playafter a contaminant enters an organism—often making sim-plifying assumptions based on body burdens or thermodynam-ic equilibrium difficult to justify.

In aquatic systems, a large amount of research has beendevoted to the development of quality criteria for contaminantsin the water column. Much simplification could be achievedif ecotoxicity results obtained for pelagic (mostly freshwater)systems could be used to assess contamination in sediments.Water quality criteria are available for a large number of chem-icals, and the partitioning data for sediments can be calculatedor produced quickly in the laboratory, obviating the need for

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

expensive chemical studies on pore water. Chemical analysesof solid-phase contaminants are typically simpler and less ex-pensive than analysis of pore-water concentrations. Thus, ifcontaminant concentrations in pore water, biota, and sedimentwere in thermodynamic equilibrium, a great deal of time, ex-pense, and energy could be saved in the development of sed-iment quality criteria. This rationale partly underlies the pres-ent strong emphasis on the development and validation ofthermodynamic equilibrium partitioning (EqP) models [5,6].In contrast to most current regulatory practice, a number ofbiotic factors strongly suggest that thermodynamic equilibriummay be a rare occurrence in natural sediments at scales relevantto bioavailability and it may be unwise to uncritically assumethat pore-water contaminant is the most bioavailable. If ther-modynamic equilibrium at the scale of the individual organismis rare, then determining the pathway of uptake becomes im-portant in the prediction of uptake rates and body burdens.The present contribution is focused on uptake pathway. Theimplications of scale in controlling uptake pathway and equi-librium are explored elsewhere [7].

Several factors related to the ecology of deposit-feedinginvertebrates suggest that feeding-related accumulation of par-ticle-bound contaminants and the details of particle sorptionkinetics may be important factors controlling contaminant up-take. Both in terms of biomass and abundance, deposit-feedinginvertebrates are frequently the dominant form of metazoanlife in marine sediments [8,9]. It has been known for quitesome time that, with the exception of environments experi-encing extremely high levels of organic enrichment, the bio-mass of deposit-feeding animals is highly correlated with theorganic content of sediments [10,11]. Within any local geo-

2454 Environ. Toxicol. Chem. 17, 1998 T.L. Forbes et al.

graphic region experiencing relatively homogeneous contam-inant loading, sedimentary organic content correlates stronglywith organic contaminant concentration. In addition to the rateof contaminant input, this coupling is a function of both localoceanographic conditions and contaminant chemistry. The twomost important physicochemical factors controlling the de-position of both natural and contaminant organic matter areparticle size (and its covariate surface area) and contaminantsurface chemistry. Organic material, (both contaminant andnatural) tends to accumulate in quiet or depositional environ-ments. Relatively small particles preferentially accumulatingin muddy, organic-rich environments have relatively high sur-face area to volume ratios and thus more specific surface areafor contaminant adsorption and/or bonding. In addition, or-ganic contaminants, as organic compounds themselves, tendto associate with natural organic material. These phenomenaunderlie the need to relate contaminant concentrations to or-ganic carbon when employing equilibrium partitioning modelsof bioavailability [5].

Sediment-dwelling animals that ingest sedimentary organicmatter for a living are frequently able to do so in a highlyselective manner. Although the exact mechanism is still unclearand may differ across taxa, most deposit feeders probably se-lectively ingest the organic-rich fraction of the available sed-iment [8]. The degree to which they can concentrate ingestedorganic matter varies from species to species, but concentrationfactors are typically on the order of two or more [9,12–14].If animals do not discriminate between contaminant and nat-ural organic matter, the concentration of the contaminant with-in an animals’ gut will often be two or more times that of thebulk sediment. Deposit feeders also exhibit extemely highweight-specific sedimentary ingestion rates. These are vari-able, but measured values range from several to greater than100 body weights per day [8]. These values make even verylow contaminant absorption efficiencies potentially significantwith regard to total exposure.

Pore water is inherently toxic to most benthic infaunal an-imals. High concentrations of toxic microbial metabolic endproducts such as sulfide (S22, HS2, H2S) and ammonium( ) resulting from organic matter decomposition combined1NH4

with the need for oxygen have led to a tendency for sediment-dwelling animals to actively isolate themselves from pore wa-ter. Animals can achieve control over exposure to pore-watersolutes in two ways. The first is the construction of burrowswith organic linings that are capable of acting as molecularsieves [15]. The second is through active irrigation of theseburrows with overlying water [16,17]. In combination thesetwo activities control the chemical composition of the burrowwater and typically cause it to differ significantly from that ofthe surrounding pore water. Thus, the most successful pore-water geochemical modeling efforts incorporate the effects ofinfaunal burrow construction and assume animal burrow watercomposition is equivalent to that of the overlying water column[18,19].

The physical and chemical evolution of a sedimentary de-posit is strongly influenced by the organisms inhabiting it[20,21]. With the exception of azoic habitats, this interactionbetween organism and environment is what ultimately deter-mines contaminant behavior and effect in marine sediments[21,22]. Four important characteristics of marine sedimentsthat are known to be strongly influenced or regulated by in-faunal animals are (1) sediment bulk composition and physicalproperties, (2) pore-water solute profiles, (3) reaction rate dis-

tributions as a function of depth within the sediment, and (4)material fluxes across the sediment–water interface [18,23,24].

This contribution addresses these problems and attempts toisolate the most important factors influencing organic contam-inant bioavailability in sediments. Our goal is to more clearlyfocus future research efforts. We performed some simple mod-el calculations that shed light on the factors controlling therelative importance of ingested sediment in contaminant ex-posure. Model results are then evaluated in terms of mea-surements of the dietary uptake of three particle-reactive or-ganic contaminants.

The experimental organism: Capitella species I

The marine infaunal polychaete Capitella species I waschosen to illustrate the approach because it is known to occurin organically enriched polycyclic aromatic hydrocarbon(PAH)-contaminated sediments [11], and its feeding biologyand morphology have been described in detail [25,26]. Cap-itella species I is one of a complex of sibling species, allformerly classified as C. capitata, each of which exhibits dis-tinctive life history characteristics. The sibling species are dis-tinquishable through electrophoretic analysis of cytosolic en-zymes [27–30]. Throughout this paper, we adopt the conven-tion of referring to the worm by sibling species where it hasbeen determined and as Capitella capitata when referring tostudies where the species is unknown.

Capitella capitata typically occurs, occasionally in greatnumbers (.500,000 m22), in organically enriched nearshoremarine environments [11,31–36]. The genus Capitella hasbeen termed an ‘‘enrichment opportunist’’ by Pearson and Ro-senberg [11], and laboratory studies have documented thatsmall individuals of Capitella sp. I can attain weight-specificgrowth rates as high as 21% d21 [37].

Worms were originally collected from a shallow subtidalsite in Setauket Harbor on the north shore of Long Island, NewYork, USA. Animals were electrophoretically identified to sib-ling species (by J.P. Grassle) and grown in mass culture indefaunated mud (.250 mm, 20–238C, 25–30 psu) until use inthe experiments.

MODEL DEVELOPMENT AND CALCULATIONS

Model assumptions

Available data on benthic biology and contaminant chem-istry are employed to isolate the most important factors con-trolling sedimentary contaminant bioavailability using calcu-lations based on a series of three simple reaction-diffusionmodels [19]. One steady-state and two transient-state modelsare constructed. The analysis required several simplifying as-sumptions. Because of the present focus on the use of EqPmodels for setting sediment quality criteria, and the prevalentview that uptake of contaminants from pore water across ex-ternal body surfaces is the primary route of exposure to mostinfaunal benthos, the assumptions we have made are generallyconservative in the sense of favoring the role of pore waterrelative to ingested sediment, that is, we have constructed ‘‘endmember’’ models to estimate the maximum potential uptakefrom pore water of a dissolved contaminant. We have chosenthe PAH fluoranthene as a model organic contaminant to il-lustrate the calculation. This choice allows direct comparisonof model results with the dietary absorption efficiency andsedimentary desorption experiments described below using thepolychaete Capitella sp. I and fluoranthene.

Pore water vs ingested sediment Environ. Toxicol. Chem. 17, 1998 2455

Assumptions for calculation of pore water uptake. (1) Dif-fusion within the pore water is assumed to be Fickian, that is,the net flux into the organism is proportional to the concen-tration gradient in the boundary layer at the point of contactbetween the surface of the organism and the surrounding porewater. The constant of proportionality, DFLU,S, is the diffusioncoefficient for fluoranthene within the pore water. (2) The abil-ity of the organism to transport or otherwise take up fluor-anthene across the external body surface is unlimited. This isrepresented mathematically by setting and holding the con-centration of fluoranthene at the surface of the organism tozero. (3) Except for the formation of a boundary layer nearthe surface of the organism due to the local balance betweenthe diffusion and sorption of fluoranthene, the organism cannotdeplete the pore water of contaminant. Thus, the pore-waterconcentration is determined by the sediment-bound concen-tration, worm uptake, and the kinetics of the sorption process.These assumptions mean that the only constraints on the rateof fluoranthene flux into the organism are those determinedby the diffusion coefficient for fluoranthene in sediment andthe thickness of the boundary layer around the organism. Wehave also assumed that (4) animals do not construct tubes,burrows, or other barriers to exposure to pore water and areat all times in direct contact with sediment and pore water.The above assumptions are quite conservative in the sense thatthey will tend to bias the calculation in favor of pore wateras a route of uptake—especially assumptions 3 and 4, whichare known to be violated to varying degrees.

The calculation of uptake from pore water into an individualworm required the additional assumptions that (5) uptake oc-curs across the entire external epithelial surface of the wormand that (6) that surface is smooth. Two important pointsshould be noted. The first is that the role of ingested pore wateris neglected. Ingestion and absorption of dissolved contami-nant will increase the relative role of deposit feeding in de-termining exposure. Second, to our knowledge there is nothingpresently known concerning the detailed mechanisms under-lying uptake of fluoranthene by Capitella sp. I across the ep-ithelial surface. At one extreme is the possibility that the con-taminant can enter worms only through a limited number ofsites on the body surface. Alternatively, uptake could occuracross the entire body surface, which is not smooth. Given thecurrent lack of knowledge and to simplify the calculations, wehave chosen to model the worm as a smooth cylinder with thetotal body surface available for uptake of contaminant.

Uptake via pore water

Three reaction-diffusion models, one steady and two tran-sient state, were examined to investigate the details of the pore-water profiles of dissolved fluoranthene expected to surrounda worm under a range of conditions. All models consisted ofboth a reaction and diffusion term and differed only with regardto time dependency and initial conditions. Degradation offluoranthene was assumed negligible over the relatively shorttime scale of the modeled periods (minutes to hours), and thusthe reaction term included only sorption kinetics. The reactionterm was constructed based on data from the desorption ex-periment described below.

Desorption experiment. A desorption experiment was con-ducted to more accurately predict contaminant kinetics overtime scales relevant to worm feeding and irrigation behavior.The same batch of radiolabeled sediment was used for boththe desorption and dietary absorption efficiency measure-

ments. The desorption kinetics data was then incorporated intothe models describing pore-water profiles surrounding a worm.

Sediment for both experiments was collected from a shal-low subtidal site in the Isefjord (Denmark, station 63 [38]) inFebruary 1995. Sediment was frozen immediately after col-lection and passed through a 63-mm sieve prior to radioisotopelabeling. Six months before the desorption experiment, 100mCi 3-[14C] fluoranthene (Sigma No. F6147; 45 mCi mmol21

was added to 2 ml sediment slurry in a glass centrifuge tube,which was maintained in the dark (58C) and periodically shak-en. Six days after 14C addition, 200 mCi51Cr (Dupont No.NEZ020, chromic chloride in 0.5 M HCl) and an equivalentvolume of 0.5 M NaOH were added the the sediment, whichwas stored at room temperature in the dark with periodic shak-ing. Labeled sediment was then stored for 6 months in darknessat 48C.

Immediately preceding the experiments, labeled sedimentwas rinsed in glass fiber filtered seawater (Whatman GF/C)and centrifuged 53 (3,000 rpm, 5 min, 208C) to remove un-bound isotope. Following the experiments, 51Cr activities weredetermined on a Packard instruments g counter (Cobra II Mod-el 5003). After g counting, sediment and/or worm sampleswere prepared for liquid scintillation counting (LSC) of 14Cby the addition of 1 ml tissue solubilizer (Soluene, Packard)following overnight incubation and vigorous mixing by theaddition of 20 ml Ultima-Gold scintillation cocktail (Packard).14C activities were determined on a Wallac Model 1409 liquidscintillation counter and were corrected for background andquench. A truncated 14C channel was used to eliminate inter-ference from 51Cr in the 14C counts.

Eighty microliters of radiolabeled sediment was carefullypipetted into 5 ml of glass fiber filtered (GF/C) seawater(32‰, 208C) in each of four replicate glass petri dishes. Four80-ml samples of labeled sediment were also taken for spe-cific activity calculation. Specific activity of the fluoranthenestock solution (45 mCi/mmol) and bulk sediment specificactivity (mean 5 32,047 DPM per 80 ml, n 5 4) were usedto calculate the concentration of fluoranthene attributible toadded radiotracer with the following two assumptions. Weassumed an experimental sediment porosity (f) of 0.65 anda solid particle density of 2.1 g/cm. This gave a fluorantheneconcentration of 0.902 mg (per gram of dry weight sediment)for the labeled sediment.

After addition of the labeled sediment to the petri dishes,0.5 ml of overlying water was sampled without replacementat five separate times (0, 13, 31, 48, and 89 min), pressurefiltered (Whatman GF/C, 25 mm), and prepared for LSC asdescribed above.

Incorporation of the experimental data into the reactionterm of the model describing pore-water concentration arounda single worm requires that desorption kinetics be expressedas a rate. A first-order equation was chosen because it provideda good fit to the data:

dCpw5 k(C 2 C ) (1)max pwdt

where Cpw is the pore-water fluoranthene concentration (mi-crograms per liter), Cmax is the experimentally determined as-ymptotic pore-water fluoranthene concentration, and k is theexperimental desorption rate constant (time).

Changing overlying water fluoranthene concentrations ver-sus desorption time were fit to the solution of Equation 1:


Fig. 1. Plot of fluoranthene concentration versus time for the de-sorption experiment. Solid line is a nonlinear curve fit to the datawhere [FLU]porewater 5 0.059 3 (1 2 exp[20.404 3 time]).

Fig. 2. Steady-state fluoranthene profile surrounding an individualworm. See text for details.

Cpw(t) 5 Cmax(1 2 e2kt) (2)

with the initial condition Cpw(0) 5 0. Equation 2 was fit to thedata using nonlinear estimation (Quasi-Newton, least squares[39]). Estimated values for k and Cmax were 0.0585 min21 and0.4044 g/L, respectively. Desorption data and the best fit so-lution to Equation 2 are plotted in Figure 1.

Steady-state pore-water profile and uptake flux. The sim-plest possible case involves a worm in sediment exposed di-rectly to pore water where uptake is assumed to have achievedsteady state. Given the assumptions outlined above and themeasured desorption kinetics, the steady-state profile sur-rounding a stationary individual worm can be calculated as[40]

]CD ] pwF,S r 1 k(C 2 C ) 5 0 (3)max pw[ ]r ]r ]r

where r is radial distance away from the surface of the worm(centimeters), DF,S is the diffusion coefficient for fluoranthenein the experimental sediment corrected for porosity and tor-tuosity (square centimeters per minute), and Cpw is the con-centration of fluoranthene in pore water. All other parametersare as defined previously with boundary conditions:

]Cpwr 5 1; 5 0 (4)]r

r 5 0; C 5 0 (5)pw

DF,S was calculated by estimating the diffusion coefficientof fluoranthene in water (DF,W) and then correcting for f andtortuosity. DF,W (square centimeters per second) was calculatedas [19,41]

2513.26 3 10D . (6)F,W 1.14 0.589m V

where m is the temperature-corrected value for the viscosityof water (1.119 centipoise) and V is the molar volume offluoranthene (cubic centimeters per mole) giving DF,W 5 0.396

cm2/d. Correcting for porosity gives DF,S 5 DF,W (0.65)2 50.167 cm2/d 5 1.160 3 1024 cm2/min [42]. The steady-stateand subsequent models were solved numerically using the Gal-erkin finite element method [43,44]. The model profile forfluoranthene surrounding an individual worm at steady stateis shown in Figure 2. Note that the maximum boundary layerthickness at steady state is on the order of 1 mm. The boundaryis also very nonlinear, with a steep gradient developing within100 mm of the surface of the worm (DCpw/Dx ø 3.1 3 1023

mg/L mm).Worm surface area. The calculation of fluoranthene flux

into an individual worm requires an estimate of worm surfacearea. This was calculated assuming a worm is a smooth cyl-inder:

Surface area 5 2prh (7)

where surface area is in units of square centimeters, r 5 wormradius (centimeters), and h 5 worm length (centimeters). Thus,the calculated surface area for a sexually mature worm witha total body volume of 2.5 mm3, a radius of 0.04 cm, and alength of 0.5 cm [26] is equal to 0.126 cm2.

Flux calculations. Fick’s first equation was used to cal-culate the flux of fluoranthene from pore water across the bodysurface of a worm:

]CJ 5 2D (8)worm F,S ]x

where Jworm is equal to the flux into the worm (nanograms offluoranthene per square centimeter per day) and DF,S is thediffusion coefficient for fluoranthene corrected for the tortu-osity of the sedimentary diffusion path (square centimeters perday). Because of the steep nonlinear gradient near the wormsurface and our use of numerical solution techniques, we haveestimated the gradient empirically using the four points nearestthe surface of the worm given by the numerical solution. Thesefour boundary points were then fit to a polynomial with aknown derivative, which was then substituted into Equation 8to calculate flux into the worm. The same procedure for uptakeflux calculation was used for all three models. The equation


Fig. 3. Example nonlinear curve fit to the four numerically calculatedboundary layer points closest to the surface of an individual worm.This curve was then used to calculate the gradient at the worm–pore-water interface. See text for detailed explanation.

Fig. 4. Time evolution of the fluoranthene profile surrounding an individual worm calculated from the transient-state model described by Equation10 (a). Initial fluoranthene pore-water concentration was set to 0 g/L. Profiles are plotted for times 0, 10, 20, 30, 40, 50, 60, 70, 80, and 90 min.Solid horizontal line at [FLU]porewater 5 0.404 mg/L is the asymptotic value calculated from the desorption experiment. See text for additionaldetails. Time evolution of the fluoranthene profile surrounding an individual worm calculated from the transient state model described by Equation10 (b). Initial fluoranthene pore-water concentration was set to the asymptotic value of 0.404 mg/L. Profiles are plotted for times 0, 10, 20, 30,40, 50, 60, 70, 80, and 90 min. Solid horizontal line at [FLU]porewater 5 0.404 g/L is the asymptotic value calculated from the desorption experiment.See text for additional details.

a3C 5 a 1 a distance 1 a distancepw 0 1 2 (9)

where distance is given in centimeters and is positive awayfrom the body surface provided an extremely good fit to allboundary layer data. The gradient is then given by the deriv-ative of Equation 9 with respect to distance evaluated at x 50 or dCpw/d distance 5 a1. Gradients calculated in this way

are a factor of 3 to 4 lower than those estimated assuming alinear boundary layer of thickness 100 mm. Using the moreaccurate nonlinear approximation of the surface boundary lay-er, we then calculate the total flux to be 1.39 ng (per squarecentimeter worm surface per day or 0.175 ng (worm per day)(Jw,T) for a 0.126 cm2 worm. The boundary layer profile andresulting fit for the steady-state solution to Equation 3 areshown in Figure 3 to illustrate the approach.

Transient-state profile—zero initial pore-water fluoran-thene. Transient-state profiles were modeled by adding a time-dependent term to Equation 3 such that

]C ]CD ] pw pwF,Sr 1 k(C 2 C ) 5 (10)max pw[ ]r ]r ]r ]t

All terms, variables, and parameters are as defined above. Thetwo transient-state models differ only in the nature of the initialcondition. In the first transient-state model, boundary condi-tions were identical to those of Equation 3, i.e., Equations 4and 5, with the addition of the initial condition

t 5 0; Cpw 5 0 (11)

indicating that all subsequent fluoranthene in pore water arisesonly from the desorption of particle-bound contaminant. Thetime course of fluoranthene profiles from 0 to 90 min is shownin Figure 4a, where the time interval between successivecurves is 10 min. Two features are readily apparent from thisplot. First, one sees a rapid initial release of fluoranthene topore water that declines with time during the approach tosteady state. Second, the pore-water gradient nearest the worm,which is important in controlling uptake flux into the worm,increases with time. This also occurs at a decreasing rate assteady state is approached. A very nonlinear boundary layerof approximately 1,000 mm in thickness exists after 90 min.


Fig. 5. Plot of the time dependence of the approach to steady statefor the two transient state models described by Equation 10. M 5initial pore-water fluoranthene concentration 0.404 mg/L, # 5 initialpore-water concentration 0.0 mg/L. See text for details of calculation.

Fig. 6. Plot of the time evolution of fluoranthene flux into an indi-vidual worm. Calculations are based on the transient state model (Eqn10) where the initial pore-water concentration of fluoranthene 5 0.404mg/L. Horizontal baseline is the calculated steady-state flux of 0.175ng worm per day. See text for additional details.

Ninety percent steady state is achieved within approximately40 min.

Transient-state profile—steady-state initial fluorantheneconcentration. To investigate the kinetics of exposure of aworm entering a region of pore water currently at steady statewith respect to the readily desorbed pool of fluoranthene, theinitial condition for Equation 10 was changed to

t 5 0; Cpw 5 0.4044 mg/L (12)

The evolution of the pore-water profile immediately surround-ing a worm is shown in Figure 4b. Steady state was achievedrapidly reaching 85% within approximately 20 min. As in theprevious model, the boundary layer is quite nonlinear andextends outward from the surface of the worm approximately800 to 1,000 mm (Fig. 4b). A rapid initial change in pore-water concentration near the worm is again followed by suc-cessively smaller total changes as time progesses from 0 to 90min. The time evolution of the boundary layer gradient nearthe worm is more complex than the first transient-state caseand is described in detail below.

Time to steady state. Worm behavior is expected to exerta major effect on pore-water profiles near the worm. This willbe true even in the absence of any effect due to the chemicalcompostion of a burrow wall or lining. Thus, it is of interestto estimate the rate at which steady state is approached overtime for later comparison with worm behavior. To quantify theprogress of the transient-state models toward steady state, wehave calculated progress as follows. For the zero pore-waterfluoranthene initial condition, we calculated the ratio of thearea under the pore-water profile for a given time to that ofsteady state (to a distance of 1 cm) and plotted it as a functionof time (Fig. 5, open diamonds). For the steady-state initialcondition we used the formula

area 2 areat50 t5if 5 (13)ss area 2 areat50 ss

where areat50 is the area at time zero, areat5i is the area at timei, and areass is the steady-state area. The relatively rapid ap-proach to steady state for the situation where C0 5 Cmax isshown in Figure 5 (open squares).

Instantaneous flux. We have also calculated instantaneousfluxes for the transient-state model in which the initial con-centration of pore-water fluoranthene is at steady state. Thisillustrates the short-term changes in flux into a worm duringthe transition to equilibrium conditions over the first 60 min.Fluxes were calculated in an identical manner to that of thesteady-state flux described previously. Flux values are initiallyquite high after only 5 min (;0.18 ng worm per day) and reachsteady-state values quickly after the decay of an initial transient‘‘bump’’ occurring at approximately 20 min (Fig. 6). Note thatbecause the diffusion coefficient for fluoranthene is constant,a plot of the worm–surface boundary gradient with time willhave the same shape as the curve for total flux (Eqn. 8).

Uptake via ingested sediment

Absorption efficiency experiment. To compare the role ofdietary uptake of fluoranthene with that estimated to occurexternally via pore water, a twin tracer absorption efficiencyexperiment was performed to measure the fraction of sediment-bound fluoranthene retained by a worm after a single gut pas-sage [45]. The sediment for the absorption experiment was thesame dual-labeled batch as that used for the sediment desorp-tion experiment, and thus labeling followed the previouslydescribed procedure.

After the final sediment centrifugation and rinse to removeunincorporated isotope, 12 worms with body volumes rangingfrom 2 to 3 mm3 were placed in small (5-cm diameter) glasspetri dishes containing 0.2 mm filtered seawater. The wormswere divided equally into absorption and ingestion groups. Athin layer (80 ml) of 14C/51Cr-labeled sediment was then addedto each dish. Both groups were allowed to feed on labeledsediment for approximately 1 h (i.e., less than one gut passagetime under experimental conditions), a time period in which,based on the desorption experiment described previously, onewould expect approximately 4% of the particle-bound fluor-


Fig. 7. Plot of calculated dietary fluoranthene absorption rate versusabsorption efficiency. n 5 nonselective feeding. — 5 calculatedsteady-state flux of 0.175 ng worm per day due only to external porewater exposure. V 5 Selective feeding with a forage ratio of 2.4 forfluoranthene. See text for details.

anthene to desorb (Fig. 1). After the 1-h feeding period, bothingestion and absorption group animals were removed to un-contaminated 0.2-mm filtered seawater and cleaned of adheringradioactive sediment. Ingestion group worms were then placedindividually into 25-mm GF/C filters, placed in g vials, andimmediately frozen in preparation for radioisotope determi-nation. The absorption group worms were transferred to newglass petri dishes containing fresh 0.2 mm filtered seawaterand 250 ml of unlabeled sediment. Absorption group wormswere allowed to feed on unlabeled sediment for 5 h, after whichtime all ingested 51Cr had been egested, indicating that theingested mineral grains had passed through the animals. Ab-sorption group worms and particulate material (containing sed-iments and isotope released in feces and mucus) were collectedseparately on 25 mm GF/C filters and prepared for radioisotopecounting as described previously.

Fluoranthene absorption efficiency (AE) was calculated as

AE 5 100 3 [1 2 (14C/51Crfeces)/(14C/51Cringested)]

where 14C/51Crfeces is the average for the absorption groupworms and 14C/51Cringested is the average measured in the in-gestion group [46]. Because the ratio data was nonnormallydistributed, the error estimates (SD) were calculated using abootstrap resampling method in which the ratio data were re-sampled 10,000 times [47]. Absorption efficiency calculatedin this manner was 55.6% 6 3.36 (SD).

Assumptions and calcululation of dietary uptake rates. Thecentral assumption for the calculation of dietary fluorantheneuptake rate is that animals do not selectively ingest fluoran-thene. Model values were calculated for an adult worm (2.5mm3 body volume) feeding at a typical rate of 3.5 mg dryweight sediment per worm per day (158C) [25] on muddysediment of porosity 0.65. The calculations for both pore waterand ingested uptake rates are described in detail below.

Uptake of fluoranthene from ingested sediment was cal-culated as egestion rate of sediment (258C) multiplied by theabsorption efficiency of the contaminant (AE 3 ER) wherethe allometric relationship for egestion rate is given as [25]

ER 5 77.45WV0.701 (14)

where ER is egestion rate (microgram dry weight sedimentper hour) and WV is worm volume (cubic millimeters). Formost purposes, ingestion rate can be accurately estimated asegestion rate for deposit feeders. This is because by far thegreatest fraction (;95–99.9%) of the material ingested bythese animals consists of undigestible mineral grains [8]. Anyunderestimate of feeding rate due to this approximation willincrease the relative importance of external pore-water uptake.Given an animal with a body volume of 2.5 mm3 and con-verting units we calculate an egestion rate of 0.0035 g dryweight sediment worm per day.

Nonselective feeding. Nonselective feeding at a fixed eges-tion rate means that the most important parameter controllingcontaminant uptake via ingested sediment is absorption effi-ciency. In Figure 7 we plot model calculations for efficienciesranging from 1 to 50% for a single gut passage of sedimentthrough the worm along with the estimated steady-state porewater value of 0.175 ng worm per day for comparison (hor-izontal line). Uptake fluxes into the worm via ingested sedi-ment increase linearly from 0.035 to 1.75 ng worm per dayfor absorption efficiencies of 1 and 50% respectively (opentriangles). Thus, at absorption efficiencies of greater than ap-proximately 5%, we would expect dietary uptake to dominate

even in the absence of selective ingestion or tube and burrowconstruction. For absorption efficiencies on the order of thosemeasured for fluoranthene (;50–60%), calculated dietary up-take flux is an order of magnitude greater than pore-water flux(Fig. 7). The predicted effect of absorption efficiency on therelative importance of the two modes of uptake can be seenclearly in a plot of relative uptake ratio (dietary flux:pore-water flux) versus absorption efficiency (Fig. 8, open trian-gles). This ratio increases linearly from approximately 0.2 to10 as AE increases from 1 to 50%.

Selective feeding. The dual-labeling method allows the cal-culation of a selectivity or forage ratio for the ingested fluor-anthene [13]. The forage ratio for fluoranthene is calculatedto be 2.4 (SD 5 1.83, n 5 11), indicating that the worms areable to more than double their intake of fluoranthene relativeto nonselective ingestion. The consequences of this selectivityare depicted in Figures 7 and 8. Adding the effect of selectivity,we see that total fluoranthene absorption rate increases evenmore rapidly as absorption efficiency increases (Fig. 7, opencircles). Thus, at experimentally measured absorption effi-ciencies, individual worms are capable of absorption rates ofapproximately 4 to 5 ng/d. These rates are 23 to 29 timesgreater than is estimated to occur via pore water alone (Fig.8, open circles). Moreover, one would expect the relative rolesof pore water versus dietary uptake to be equal at absorptionefficiencies lower that are currently measurable (,1%).

DISCUSSION

Relative roles of pore water and ingested sediment in theuptake of fluoranthene

Our analysis suggests that, for deposit feeders, uptake oforganic contaminants may often be dominated by the ingestionof sediment. In addition to the experimental and modeling data


Fig. 8. Ratio of uptake flux of fluoranthene due to ingested sedimentrelative to that taken up from pore water versus absorption efficiency.V 5 Selective feeding with a forage ratio of 2.4 for fluoranthene. n5 nonselective feeding. See text for details.

presented above, this conclusion follows from considerationof both the life history and bioenergetic characteristics of theseorganisms, the most important of which are the following. (1)Deposit feeders dominate communities in organic-rich habitatswhere contaminants accumulate. Their abundances are oftenstrongly correlated with organic contaminant concentrations—by far the greatest fraction of which is associated with particles.(2) Deposit feeders exhibit extremely high mass-specific ratesof ingestion. Ingestion rates can be greater than 100 bodyweights per day [8]. (3) Many of these animals are capable ofa substantial degree of selection for sedimentary organic matterincluding contaminants [8, this study]. Selective ingestion actsto increase the concentration of contaminant within the di-gestive system of the worm (Cworm). This, in combination withthe multiplicative nature of dietary uptake flux (uptake 5 ab-sorption 3 ingestion rate), leads to a divergence in the esti-mates of absorption rate for different selectivities as absorptionefficiency increases (Fig. 8). This means that a given changein selectivity will have greater effect on the relative uptakeratio at higher absorption efficiencies. Finally, points 2 and 3indicate that even low absorption efficiencies (a few percent)can lead to high uptake rates via ingestion of contaminatedparticles, quickly overshadowing those estimated for pore wa-ter (Fig. 8). Recent independent estimates of absorption effi-ciencies for fluoranthene in other deposit-feeding invertebratesare simlar to that of this study—on the order of 20 to 50%for a single passage of sediment through the gut [45].

A complication that our simplified analysis has not ad-dressed concerns the fact that infaunal behavior often modifiesthe local environment in ways that may strongly influence totalexposure. This modification will be driven primarily by feed-ing, burrowing, and irrigation activities [48,49]. The net resultwill be a dynamic relationship strongly influenced by the localinhabitants. Burrow and tube construction in combination withrespiratory ventilation significantly modifies the chemicalcomposition of burrow water in comparison with bulk pore

water [50]. This effect is so pronounced that the current gen-eration of geochemical models can quite accurately predict thesedimentary depth profiles of, for example, nutrients assumingburrow water concentration is equal to that of the overlyingwater [18,51]. If, as is the case for many other dissolved con-stituents, the burrow water concentration of organic contam-inants more closely resembles that of the overlying water, therole of dietary uptake will predominate. Future work shouldinclude more detailed investigations into the chemical com-position of burrow water and role of such behaviors as dis-continuous irrigation in determining total exposure.

Our analysis reinforces appreciation for the importance ofsome factors that are currently well appreciated but sheds lighton several more that deserve greater attention. Factors that actto modify or control the efficiency of dietary contaminantabsorption or the formation of a local gradient or boundarylayer surrounding an organism will be especially important(e.g., selection and burrow construction). Of the more impor-tant chemical factors affecting bioavailability, the role of sorp-tion kinetics should receive much greater attention in the fu-ture. Although equilibrium partitioning considerations may beimportant, the kinetics of adsorption and desorption can strong-ly influence the boundary layer immediately surrounding bur-rowing organisms or irrigated tubes. Important but currentlyneglected biological aspects include irrigation behavior andthe size, composition, and density of tubes and burrows. Thesepopulation- and community-level characteristics will act tomodify the diffusion and transport of contaminants into andout of sediments and may often create a dynamic relationshipstrongly influencing and being influenced by the sorption ki-netics for a particular compound.

Instantaneous flux calculations suggest an initial and rapidincrease in uptake flux followed by a rapid decline to steady-state values if a worm is suddenly exposed to contaminant-equilibrated pore water (Fig. 6). These steady-state fluxes orpore-water profiles are relatively short-term phenomena anddo not indicate the presence of thermodynamic equilibrium.Although it at first appears large, the initial peak value of0.1817 ng worm per day is less than a 4% increase in fluxover the steady-state value. This suggests that changes inboundary layer thickness caused by worm movement alonewould not be expected to significantly increase total uptakeflux. Burrow construction and irrigation will act to displacethe diffusive boundary layer outward from the worm by ex-tending it to the burrow wall or lining. This means that, for agiven contaminated sediment, the temporal and spatial char-acteristics of the boundary are primarily controlled by irri-gation activity and the chemical composition of the burrow.

In contrast to the data for desorption kinetics and the struc-ture of organism-scale boundary layers, data exist for the ab-sorption efficiencies of fluoranthene and a number of otherorganic contaminants by deposit feeders. Forbes and Forbes[45] measured fluoranthene absorption efficiencies of 22 to46% in two clones of the deposit-feeding gastropod Pota-mopyrgus antipodarum using a modern dual-labeling method.Somewhat higher values (36–60%) were obtained for Capi-tella species I using the same dual-labeling methodology (thisinvestigation). For the experiments with Capitella, the highabsorption efficiencies were obtained for radiolabeled sedi-ment that had been aged in darkness for 8 months at 48C,suggesting a substantial digestive capability for this animal inthe 1 to 1 1/2 h required for a single gut passage of ingested


sediment. This is in marked contrast to the experimental de-sorption estimate of 3 to 4% over the same time period.

Recent investigations of the digestive fluids of a numberof deposit-feeding invertebrates demonstrate solubilities ofboth inorganic and organic contaminants that are often manytimes greater than predicted from partitioning models for or-dinary pore-water–sediment systems [52]. High digestive fluidconcentrations of solubilizing agents such as amino acids (met-als) and surfactants (PAH) may underlie this enhanced diges-tive capability.

Direct measures of uptake rates from both pore water andingested sediment are necessary to fully evaluate the potentialeffects of a given chemical or mixture of chemicals on anyparticular organism. Additional metabolism (at some cost)once the chemical is taken up may mask detection of potentialeffects and determination of causal relationships. When in-dividuals of Capitella sp. I were exposed to sediment contam-inated with 360 mg (per gram of dry wt sediment), body bur-dens rose to approximately 650 mg (per gram of dry wt worm)after 1 d but then declined to levels below detection by day7, even though the worms remained in contaminated sediment[45]. Measurements of body burdens after day 7 would suggesta decreased effect, perhaps erroneously if worms must con-tinue to expend energy to maintain low body burdens. Alter-natively, the model results discussed previously suggest thepossibility that the initial high rate of intake recorded for Cap-itella may have been due, at least in part, to initial directexposure to contaminated pore water before tube and burrowconstruction could be performed. If this was the case, then theinitial depuration of the fluoranthene would not have beenfollowed by a continued high metabolic cost.

Finally, we have left totally unexplored the relative im-portance of pore water versus dietary uptake during organismdevelopment. Much is known, particularly for Capitella sp. I,of the body size scaling of physiological and morphologicalcharacteristics that are expected to relate to bioavailability[25,26,53–56]. It is well established that younger or smallerorganisms tend to have higher specific rates of feeding andmetabolism (i.e., per rate weight1) in combination with greaterspecific surface areas or areas for respiratory or other materialexchange with the external environment [57]. It seems verylikely that such activities as irrigation will also occur at higherspecific rates for younger/smaller organisms. Thus, uptake,metabolism, and the influence of organisms on the fate of aparticular contaminant should be enhanced in smaller organ-isms, all other factors being equal, making earlier life stagesat greater potential risk.

Experimental data in combination with model calculationssuggest that ingestion of contaminated sediment may often bethe dominant uptake pathway for deposit-feeding invertebratesin sediments. For organic contaminants, there is no chemicallywell-defined, bioavailable fraction that can be unambiguouslyquantified—even given perfect geochemical knowledge. Bio-availability is a function of organism–sediment interactions.This coupling between biological effect and chemical fate isexpected to generate potentially important feedback relation-ships between the biota and contaminant geochemistry. Theseconclusions bring into question the validity of many currentresearch programs in environmental chemistry geared towardchemically perceiving sediments as organisms do without fo-cusing in more detail on how deposit feeders interact withsediment. A promising current approach involves the use or

simulation of the gut environment of deposit feeders to esti-mate dietary bioavailability [52].

The processes of uptake and metabolism by organisms arerelatively free of the strict constraints that operate in the purelychemical realm. Furthermore, these behavioral and physiolog-ical processes are expected to be coupled with organism sizeor stage of development with smaller/younger organisms in-teracting at greater specific rates with the external environmentand thus at potentially greater risk. Different uptake mecha-nisms and metabolic pathways have evolved across taxa, anddifferent taxa often respond differently to the same chemicalchallenge. Nevertheless, our analysis reveals that certain keyaspects of organism–sediment relations, especially organismbehavior and contaminant sorption kinetics, can be used toimprove our current picture of bioavailability considerably.

Acknowledgement—We gratefully acknowledge the support of theDanish National Science Council with grants No. 11-1047-1 to T.L.Forbes and 11-0954-1 to Ole Anderson. The ideas and thoughts pre-sented here have benefited significantly from the suggestions andcomments of Larry Mayer, Marianne Thomsen, and two anonymousreviewers.

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