Freshwater Bivalve Ecotoxoicology - Chapter 8 ppsx

BIOCONCENTRATION The accumulation of contaminants from the water column by bivalves is referred to as tration.” Bioconcentration is defined as the partitioning of a contaminant from an aq

8 Toxicokinetics of Environmental

Contaminants in Freshwater Bivalves

Waverly A Thorsen, W Gregory Cope, and Damian Shea

INTRODUCTION

Bivalves have been used for decades as sentinel organisms to monitor pollution in the aquaticenvironment (Foster and Bates 1978; Farrington et al 1983; Colombo et al 1995; Peven, Uhler, andQuerzoli 1996; Blackmore and Wang 2003) Many different classes of chemicals have been studied

in this way including hydrophobic organic contaminants (HOCs), such as polycyclic aromatichydrocarbons (PAHs), polychlorinated biphenyls (PCBs), and organochlorine (OC) pesticides,

as well as inorganic contaminants such as the heavy metals cadmium (Cd), lead (Pb) andmercury (Hg) and the radionuclides plutonium (239,240Pu) and cesium (137Cs) The use of bivalvesfor biomonitoring of environmental pollution addresses difficulties associated with determiningaqueous contaminant concentrations (Farrington et al 1983) Many HOCs exhibit very low watersolubilities (e.g., coronene: 1.4!10K4mg/L, at 258C), which require large sample sizes foradequate instrumental analysis Moreover, trace metals require “ultraclean” techniques and arealso frequently found in very low concentrations in the aqueous phase, sometimes at levels close

to instrument detection limits (i.e., pg/L) Additionally, random water sampling may not capturereal trends in pollutant concentrations over an integrated time scale

In an attempt to overcome these obstacles, native bivalves are frequently collected worldwide,extracted, and analyzed for pollutant tissue burdens to provide preliminary information at sitessuspected of contamination or to monitor chemical and waste discharge effluents However, toeffectively understand and correlate the relationship between concentrations of pollutants in theaquatic environment to concentrations in bivalve tissue and potential toxic effects, it is best to have

an understanding of the kinetics involved in the uptake, distribution, and elimination of pollutantsby/from mussel tissues Additionally, this information is required to understand and predict concen-trations in other environmental compartments, such as predicting aqueous or sediment exposureconcentrations from bivalve tissue burdens (Neff and Burns 1996)

Traditionally, marine bivalves such as the blue mussel, Mytilus edulis, have been used forenvironmental monitoring due to concern for pollution in coastal and estuarine areas (Farrington

et al 1983; Salanki and Balogh 1989; Beliaeff et al 2002) However, more recently (1980s) water bivalves have been increasingly utilized to assess the quality of lakes, rivers, and streams ofconcern, not only for the protection of human health, but also to better explain recent major declines

fresh-of many North American freshwater mussel populations (e.g., Keller and Zam 1991; Naimo 1995;Jacobson et al 1997) Generally, information gleaned from freshwater bivalves has demonstratedsimilarities to marine bivalves; however, physiologies can vary greatly between species, age, bodysize, ingestion rate, reproductive state, stress, and location, among other factors (Landrum et al 1994;

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Naimo 1995; Morrison et al 1996) Therefore, in an attempt to better evaluate pollutant fate and toeffectively protect and remediate the natural environment, it would be beneficial to understand thetoxicokinetics of both marine and freshwater mussels The intent of this chapter is to present back-ground information and to assess the toxicokinetic information available for freshwater bivalves(mussels and clams) Where data are limited, information on marine bivalves will be presented and, insome cases, will be presented in tandem with freshwater bivalve information in a comparativecontext This chapter is not meant to be an exhaustive review of the literature pertaining to theseissues, but rather it is meant to aid researchers, managers, and others, in understanding the bioaccu-mulation of organic and inorganic contaminants in freshwater bivalves.

UPTAKE ANDELIMINATION

Bivalves are exposed to and take up pollutants in tandem with their primary respiratory and feedingmechanisms; chemicals enter mussels actively and passively as they filter water through their gillsfor respiration and feeding (dietary exposure), or in the case of inorganic contaminants such asmetals, through facilitated diffusion, active transport, or endocytosis (Marigomez et al 2002).Additionally, some bivalve species are exposed to pollutants through pedal feeding or gut ingestion

of sediment (McMahon and Bogan 2001) Therefore, chemical uptake can occur in a direct fashionwhen mussels draw large quantities of water (up to 11 L/mussel/day for Unionidae, Naimo 1995)into their gills or, in an indirect fashion, when ingestion of sediment occurs and chemicals desorb(passively or through facilitated desorption) from the sediment particles into the bivalve gut andbecome assimilated Once chemicals enter the organism, they partition into or associate withtissues For example, heavy metals will accumulate primarily in muscles and organ (soft) tissues(Plette et al 1999; Markich, Brown, and Jeffree 2001; Marigomez et al 2002) and organic pollu-tants will accumulate in the lipid (Farrington et al 1983; Di Toro et al 1991) Generally, uptake isvery rapid when the bivalve is first exposed and then levels off, sometimes requiring extensive timeperiods for an equilibrium state to be reached (Figure 8.1a) A similar trend (Figure 8.1b) isobserved for the elimination process, which may be rapid at first and then level off, somecompounds never being fully eliminated (i.e., some compounds with half-lives of 20 years).Uptake and elimination rates for both HOCs and metals can be determined through field and/orlaboratory studies One potential concern in these types of studies is the possibility that the bivalvesstop siphoning Although this is more likely to influence studies of shorter duration, it should betaken into consideration when analyzing the data A typical uptake/elimination experiment consists

of “clean” bivalves (referenced or depurated prior to commencement of the study) exposed to aconstant chemical concentration in water, and sampled at increasing time intervals, to determine thechemical concentrations in tissue over time For example, bivalves can be collected from a rela-tively uncontaminated field reference site, and deployed at a contaminated field site, or broughtback to the laboratory for contaminant exposure After sufficient exposure time, the organisms areremoved and placed in clean water for measurement of the elimination (depuration) rate of thecompounds In the natural environment, elimination of certain chemicals might require extensivetime periods In locations where exposure levels are constant or increasing, bivalves may noteliminate the chemicals In many instances, bivalves will accumulate contaminants to levels sig-nificantly higher than those in the water column This can pose toxicity risks to the mussel andpredatory animals or can result in biomagnification and subsequent increases in contaminantconcentrations progressively up the food web

BIOCONCENTRATION

The accumulation of contaminants from the water column by bivalves is referred to as tration.” Bioconcentration is defined as the partitioning of a contaminant from an aqueous phaseinto an organism and will occur when the contaminant uptake rate is greater than that for

“bioconcen-elimination Typically, this leads to high concentrations of chemicals in bivalve tissues For HOCs,partitioning generally occurs between the dissolved phase of the water and the organism lipid Themost basic example of partitioning is defined as the octanol–water partition coefficient, or Kow:

KowZ ½contaminantoctanol=½contaminantwaterThe Kowis a measurement of a chemical’s affinity for octanol versus water In many cases,octanol is used as a surrogate for the organism lipid A chemical with a lesser Kowvalue (less than100) will partition less into the lipid than a chemical with a greater Kow(greater than 1,000) This type

of partitioning will occur between the aqueous phase and bivalve lipid until a steady-state conditionhas been reached (i.e., the concentration in the organism relative to the exposure system is unchan-ging with time) Once steady-state or equilibrium has been reached, it is generally referred to as

“equilibrium partitioning.” In a simple system, equilibrium partitioning can be modeled bycomparing the affinities (i.e., solubilities and fugacities) of a chemical for bivalve lipid versuswater(Figure 8.2).To determine the extent of bioconcentration of a chemical in tissues, a “biocon-centration factor” or BCF can be calculated The BCF is defined as the pollutant concentration in thebivalvel tissue (Ctissue) divided by the dissolved aqueous pollutant concentration (Cwater) at steady-state:

BCF Z Ctissue=Cwater

0 2 4 6 8 10 12 14

100 hours of exposure The rate of elimination is also rapid and is essentially the reverse of the uptake curve.When placed in clean water, the mussels initially depurate the contaminant rapidly from their tissues and thenreach a plateau, where no further elimination occurs on this time scale

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The BCF can also be determined by dividing the empirically derived contaminant uptake rateconstant (k1) by the empirically derived elimination rate constant (k2):

BCF Z k1=k2

In general, the BCF is related to the hydrophobic character of the contaminant In this way, BCFvalues typically correlate in a linear fashion to Kowvalues (Geyer et al 1982; Mackay 1982; Hawkerand Connell 1986; Pruell et al 1986; Schuurmann and Klein 1988; Thorsen 2003) (Figure 8.3)

In many cases, a steady state bioconcentration regression equation can be developed by linearlyregressing a log BCF versus a log Kowplot The resulting equation for the line takes the form of

log BCF Z m log KowCbwhere m and b are the slope and y-intercept of the line, respectively This equation can model thebioconcentration of hydrophobic organic pollutants by bivalves and can be used to predict aqueousexposure concentrations

y = 1.024 x − 1.8183

R2 = 0.8741 0

1 2 3 4 5 6

The “partitioning” of metals, however, generally refers to the adsorption of metals onto activesites in/on target tissues, such as anionic sites on bivalve gills (Kramer et al 1997; Marigomez et al.2002), rather than absorption into a bivalve lipid A bioconcentration factor, though slightly lessutilitarian than for HOCs due to very slow uptake rate constants, can similarly be computed by

BCFmetalZ Ctissue=Cwater

where Ctissueis the moles of metal per gram of soft weight tissue and Cwateris the moles of metaldissolved per mL (or L) of water This BCF value must also be calculated when the system hasreached steady-state More complex equations exist for predicting bioconcentration (and uptake,elimination rates) when a system is not at steady state and are discussed elsewhere (Russell andGobas 1989; Butte 1991) The bioconcentration of metals is affected by many factors, includingwater pH, hardness, alkalinity, conductivity, and dissolved organic and inorganic matter, which will

be discussed in following sections

BIOACCUMULATION

While bioconcentration refers only to the uptake of chemicals directly from the water, the termbioaccumulation does not differentiate between uptake media and includes chemical accumulationinto organisms from both abiotic (i.e., water and sediment) and biotic (i.e., food) sources Forexample, bivalves can bioaccumulate chemicals and metals from the water column and thesediment phase in the natural environment Typically, scientists may model this relationship bycalculating either a bioaccumulation factor (BAF) or a biota-sediment accumulation factor (BSAF).The BAF includes exposure due to water and food sources, whereas the BSAF (only used forHOCs) models the partitioning/association of a chemical between the lipid phases in the organismand the sediment, where the sediment “lipid” phase is considered to be organic carbon The BAF isrepresented by

BAF Z Ctissue=CfoodCCwaterCCother exposureswhereas the BSAF is mathematically defined as

BSAF Z ðCtissue=lipid fractionÞ=ðCsediment=organic carbon fractionÞwhere the chemical concentration in the bivalve (Ctissue) and sediment (Csediment) are normalized

to the mass fraction of organism lipid and sediment organic carbon, respectively Similar to theBCF calculation, a BSAF value is calculated when the chemical has reached a steady-statewithin the study system Theoretically, BSAF values will equal unity or one However, BSAFvalues may be less than one if the bivalve metabolizes the chemical or the system has notreached steady-state (chemicals may not be fully available to the organism due to very slowdesorption or very strong binding) BSAF values can also be greater than one because organiccarbon is generally less “lipid-like” than the organism lipid due to hydrophilic components

of natural organic matter (Di Toro et al 1991) The calculation of BSAF values can lendinformation about a particular chemical’s bioavailability (see Bioavailability and BioticLigand Models)

Metals do not interact with organisms in the environment in the same way that HOCs do Aspreviously mentioned, while HOCs generally partition (absorb) into the lipid phase of a bivalve,metals adsorb to the gill and other anionic sites on tissue surfaces or are actively transported viamembrane pumps For example, metals such as cadmium can enter a bivalve by binding toToxicokinetics of Environmental Contaminants in Freshwater Bivalves 173

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membrane transport ligands Bioaccumulation of metals, including filtration of water and ingestion

of food particles, in bivalves can be similarly measured through the use of a BAF:

BAF Z Ctissue=Cwater;dissolvedBioaccumulation factors for metals are more difficult to interpret than for organics because theinteractions between a target site (biological organism) and the metal are complicated by compe-tition for binding sites and many more environmental variables than simply dissolved or particulateorganic carbon For all chemicals and metals, bioaccumulation is the balance between all means ofchemical uptake and all means of elimination

METABOLISM ANDBIOTRANSFORMATION

For those contaminants that bivalves are capable of metabolizing, BCF, BAF, and BSAF valueswill be decreased In general, the lesser metabolic capacities in bivalves makes them adequatesentinels of aquatic environmental pollution (James 1989); however, bivalves have been shown tometabolize certain classes of compounds better than others For example, mussels possess onlyminimal abilities to biotransform PAHs, and therefore, are good sentinels of the accumulation ofPAHs Some marine mussels (M edulis), however, have been shown to metabolize the PCB,hexachlorobiphenyl (HCBP) (Bauer, Weigelt, and Ernst 1989), and therefore, will exhibit lowerBCF values Additionally, bivalves have been shown to possess detoxification systems includinglow molecular weight proteins like metallothionein (MT) and lysosomal granules that makemetals complex and chelate, thereby altering the metal uptake/distribution/eliminationkinetics (Naimo 1995; Tessier and Blais 1996; Vesk and Byrne 1999; Byrne and Vesk 2000;Baudrimont et al 2002)

BIOAVAILABILITY ANDBIOTICLIGANDMODELS

Underlying all of the previous concepts is the notion of bioavailability Bioavailability can bedefined as the percentage of a chemical fully available for uptake by an organism Differentchemicals and inorganic contaminants have unique bioavailabilites, which will depend on manyfactors including water conditions such as hardness, pH, temperature, and turbidity, as well as thephysical–chemical characteristics of the compound such as water solubility, vapor pressure, andspeciation (ionic state) For example, chemicals that exhibit very low water solubilities readily sorb

to organic carbon phases in the water column, such as particulate or dissolved organic carbon (POC,DOC) The rate of desorption and co-occurrence of the mussel with the particle(s) partiallydetermines the chemical’s bioavailability If the rate of desorption is rapid relative to the co-occur-rence of the particle and the organism, the chemical may be fully bioavailable However, if the rate

of desorption is very slow, the chemical may not be readily available HOCs may frequentlybecome associated with natural organic matter in the aqueous and sediment phases, whereasmetals may become complexed to various organic (DOC) and inorganic compounds present inthe water such as calcium and potassium carbonates (CaCO3, KCO3)

The bioavailability of a chemical is important to understand both to ensure the protection ofaquatic organisms and to implement effective and cost-efficient remediation techniques This

is particularly important because underpredictions of toxicity can result in unacceptable risks toorganisms, whereas overpredictions of toxicity can require costly practices for clean-up Forinstance, bivalve tissue burdens are traditionally compared directly to total aqueous or sediment-contaminant concentrations, without regard for the bioavailable fraction This method can over-predict the actual exposure concentrations bivalves (and other aquatic organisms) receive and mayresult in costly, yet ineffective, remediation of a site Moreover, sediment concentrations of total

metal do not always correlate well with bivalve tissue burdens Rather, it may be the speciation ofthe metal (e.g., Hg2Cversus CH3Hg), or ratio of metal concentration to the amount of acid-volatilesulfate in the sediment (Di Toro et al 1992), that best determines the metal concentration in andsubsequent toxicity to the bivalve One can see the problems that may arise when regulatory andremediation techniques are based on incorrect assessments of chemical bioavailability.

HYDROPHOBIC ORGANIC CONTAMINANTS

UPTAKE

As previously stated, HOCs primarily partition into a bivalve lipid, which is considered essentially an

“infinite sink” whereby saturation of the pool does not occur The uptake of a hydrophobic organicchemical into bivalve tissues can be defined mathematically as

dCtissue=dt Z k1CwaterKk2Ctissuewhere dCtissue/dt is the change in bivalve contaminant concentration over change in time (t), k1is theuptake rate constant of the chemical, Cwateris the aqueous chemical concentration, k2is the elimin-ation rate (seeElimination),and Ctissueis the concentration of chemical in the bivalve (see Landrum,Lee, and Lydy 1992 for a review of toxicokinetic models) If the concentration of the pollutant in thewater column changes, this change will be mirrored in the bivalve over several days to weeks Thisprocess is considered first-order on a natural log (ln) basis By integration, the above equationbecomes

CtissueZ ðk1=k2ÞCwaterð1KeKk 2 tÞ:

Bivalves primarily take up HOCs directly from the water column (Thomann and Komlos 1999;Birdsall, Kukor, and Cheney 2001) through their gills, although some studies have suggestedadditional chemical inputs from dietary exposure (Brieger and Hunter 1993; Gossiaux, Landrum,and Fisher 1996; Bjork and Gilek 1997; Raikow and Hamilton 2001), and direct sediment ingestionvia pedal feeding mechanisms (McMahon and Bogan 2001; Raikow and Hamilton 2001) There isdebate in the literature over the relative contribution of each of these uptake routes; however, it should

be noted that once the system has attained steady-state (dC/dtZ0), the route of contaminant exposure

is irrelevant (Di Toro et al 1991) Because of their minimal metabolic capabilities for metabolizingthe majority of HOCs (Farrington et al 1983; James 1989), bivalves accumulate these contaminants

to high levels in their lipid tissues, which can often reach many orders of magnitude greater than thecorresponding concentrations in water or sediment phases Despite the common use of freshwaterbivalves for monitoring aquatic environments, relatively little information is known regarding HOCuptake rate constants, compared with that for marine bivalves Moreover, much of the freshwater andmarine data represent only a few species For instance, the majority of the freshwater uptake studiesfocus on Dreissena polymorpha, whereas the majority of marine uptake studies use M edulis.There are various ranges in reported k1values for freshwater bivalves depending on species, andstudy variables such as temperature, exposure environment, mussel size, and lipid content

(Table 8.1a,b;Table 8.2a,bfor study summaries, Fisher et al 1993; Bruner, Fisher, and Landrum1994; Gossiaux, Landrum, and Fisher 1996; Fisher et al 1999) However, based on the available data,most k1values compare well, with only a few exceptions (Table 8.1a) Many studies demonstrateinitial rapid uptake during initial exposure for both freshwater and marine species (Lee, Sauerheber,and Benson 1972; Obana et al 1983; Bjork and Gilek 1997; Birdsall, Kukor, and Cheney 2001) Forexample, Birdsall, Kukor, and Cheney (2001) reported rapid uptake of the PAHs naphthalene (N0),anthracene (AN), and chrysene (C0) by Elliptio complanata gills Their data demonstrated that theToxicokinetics of Environmental Contaminants in Freshwater Bivalves 175

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Published Uptake and Elimination Rate Constants for Various Freshwater and Marine Bivalve Species as a Function of Chemical Class

and Water Solubility

C leana Pesticides (3) 3.22(OX)–4.22(TBC) 24.2(TBC)–338.0(CNF) 0.0450(CNF)–0.0600(TBC) Uno et al (1997)

D polymorpha PAH (2) 5.18(PY)–6.04(BaP) 672.0(BaP)–32,737.0(BaP) 0.0240–0.3840(BaP) Gossiaux, Landrum, and

Fisher (1996) PCB (2) 5.90(PCP)–6.90(HCBP) 2,280.0(PCP)–

26,448.0(HCBP) 0.0240(HCBP)–0.1920(PCP)

D polymorpha TCBT (8) 6.73(28)–7.54(25) 683.3(52)–848.7(80) 0.0052(27)–0.0226(21) Van Haelst et al (1996a)

D polymorpha PAH (2) 5.18(PY)–6.04(BaP) 7,680.0(PY)–31,200.0(BaP) 0.1920(BaP)–0.5760(PY) Bruner, Fisher, and Landrum

(1994) PCB (2) 5.9(TCBP)–6.9(HCBP) 9,120.0–40,320.0(HCBP) 0.1200(HCBP)–

0.5040(TCBP)

20,112.0(BaP) 0.0090(PY, BaP) Fisher et al (1993)PCB (2) 5.90(TCBP)–6.90(HCBP) 4,008.0–25,752.0(HCBP) 0.0040(HCBP)–

0.0170(TCBP)

D polymorpha PCB (2) 6.36(77)–7.42(169) 551.0(77)–1,480.0(169) 0.0340(169)–0.0350(77) Brieger and Hunter (1993)

E complanata HCB, OCS 5.45(HCB)–6.29(OCS) 650.0(HCB)–1,010.0(OCS) 0.4100(HCB)–0.1600(OCS) Russell and Gobas (1989)

M edulis PCB (3) 5.67(31)–6.92(153) 2,160.0–168,000.0(153) 0.0288(153)–0.1368(31) Bjork and Gilek (1997)

(1996)

(1991)

C virginica PAH (14) 4.57(P0)–6.50(BghiF) 330.0(P0)–2,365.0(MPY) 0.0090(BF)–0.1180(FL) Bender et al (1988)

M mercenaria PAH (14) 4.57(P0)–6.50(BghiF) 187.0(MP0)–2,842.0(BaA) 0.0870(BaP)–0.2130(FL)

DA=diben-a Number in parentheses refers to total number of chemicals studied within the chemical class.

Published Solubility Values, Bioconcentration Factors, and Half-Lives for Various Freshwater and Marine Bivalves as a Function of

Chemical Class

Freshwater

PAH (38) 3.37(N0)–7.64(CO) 1.90(N0)–5.20(CO) PAH (45) 3.37(N0)–7.64(CO) 1.60(AN)–5.51(C4) 2.60(26DMN0)–16.50(PE)

C leana Pesticides (3) 3.22(OX)–4.22(TBC) 2.34(OX)–4.14(CNF) 11.60(TBC)–15.40(CNF) Uno et al (1997)

D polymorpha PAH (2) 5.18(PY)–6.04(BaP) 4.34(PY)–5.43(BaP) 1.75(BaP)–28.80(BaP) Gossiaux, Landrum, and

Fisher (1996) PCB (2) 5.90(PCP)–6.90(HCBP) 4.00(PCP)–5.74(HCBP) 3.60(PCP)–28.80(HCBP)

D polymorpha TCBT (8) 6.73(28)–7.54(25) 4.43(80)–5.19(27) 18.60(80)–71.80(22) Van Haelst et al (1996a,

1996b)

D polymorpha PAH (2) 5.18(PY)–6.04(BaP) 4.11(PY)–4.92(BaP) 1.20(PY)–3.60(BaP) Bruner, Fisher, and Landrum

(1994) PCB (2) 5.90(TCBP)–6.90(HCBP) 4.32(TCBP)–5.38(HCBP) 1.40(TCBP)–5.80(HCBP)

D polymorpha PAH (2) 5.18(PY)–6.04(BaP) 4.65(PY)–4.88(BaP) 2.60(BaP)–3.00(PY) Fisher et al (1993)

PCB (2) 5.90(TCBP)–6.90(HCBP) 4.62(HCBP)–5.43(HCBP) 1.70(TCBP)–7.20(HCBP)

D polymorpha PCB (2) 6.36(77)–7.42(169) 4.02(77)–4.45(169) 19.80(77)–20.40(169) Brieger and Hunter (1993)

E complanata HCB, OCS 5.45(HCB)–6.29(OCS) 3.56(HCB)–4.16(OCS) 1.70(HCB)–4.30(OCS) Russell and Gobas (1989) Marine

M edulis PCB (3) 5.67(31)–6.92(153) 4.70(49)–6.80(153)BAFs 5.00(31)–24.20(153) Bjork and Gilek (1997)

PCB (9) 22.00(26)–130.00(149)

(1996)

C virginica PAH (14) 4.57(P0)–6.50(BghiF) 3.20(P0)–4.90(BF) 5.90(FL)–77.00(BF) Bender et al (1988)

M mercenaria PAH (14) 4.57(P0)–6.50(BghiF) 3.20(MP0)–4.40(BghiF) 3.30(FL)–8.00(BaP)

D polymorpha PAH, PCB Fisher et al (1993)

D polymorpha Water, sediment,

food PCB, 3 21- to 100-dayexposure, rapid

complanata Water only PCP Steady-state reachedin 16 h Makela and Oikari(1995)

D polymorpha Water and field PCB, 36 2-day exposure,

16-day elimination Morrison et al (1995)

D polymorpha Water only PAH, PCB 6-hour uptake, 15-day

elimination Gossiaux, Landrum,and Fisher (1996)

D polymorpha Water only PCBs, OCs Chevreuil et al (1996)

D polymorpha Water only TCBTs, 8 21-day uptake, no

steady-state reached

Van Halest et al (1996a, 1996b)

C flumina Water only PCP 96-hour uptake,

D polymorpha Water only PAH, PCB Fisher et al (1999)

E complanata Water only PAH, pesticides Used excised gills Birdsall, Kukor, and

Oysters No 2 Fuel oil 60-day uptake,

(1977)

(continued)

average uptake of AN (log Kow4.54) and C0 (log Kow5.86) was similar, and both were greater thanthat for N0 (log Kow3.37), which was explained by its lower lipid affinity.

Differences in k1can be observed when comparing the same analyte among studies, as well aswhen comparing different analytes with similar physico-chemical parameters However, with a fewexceptions, the differences appear to be relatively small, considering the many variables that canexist between studies For example, k1values measured for benzo(a)pyrene (BaP) and HCBP inboth the field and laboratory over the course of three years and at different temperatures (5–248C)

in D polymorpha compare well(Table 8.1a).Specifically, for BaP the range of uptake rates is from9,960 to 32,736 mL/g day, a factor of 3 difference The differences between highest and lowestupdate rate constants for HCBP, pentachlorophenol (PCP), and pyrene (PY) are even less, at factors

of 2.0, 2.6, and 2.0, respectively Data from two collection timepoints have been omitted for thiscomparison due to very low uptake rate constants, which the authors believed was from over-wintered mussels experiencing stress (both occurred for mussels collected at 48C in the field;however, when mussels were fed while being acclimated to 48C in the laboratory, these effectswere not observed) (Gossiaux, Landrum, and Fisher 1996) Therefore, it is important to considerthat larger differences can occur based on the physiological state of the organism Laboratory-derived k1s for PCP increased from 3,960 mL/g day at 48C to 5,928 mL/g day at 158C, whereasfield-derived k1s showed even less difference with a more dramatic temperature increase from 4 to248C (3,240 versus 2,640 mL/g day, respectively) (Gossiaux, Landrum, and Fisher 1996) These

TABLE 8.2a (Continued)

Species Exposure Chemical Class Duration References Ostrea edulis Flow-through system PAH, N0 Riley et al (1981) Tapes japonica Water and field PAH, 9 7- to 14-day exposure Obana et al (1983) Clam, oyster, mussel Water only PAH, D0-D3 Ogata et al (1984) Oysters PAH 15-day uptake Pittinger et al (1985)

M edulis Sediment dosed PAH, PCB 40-day uptake, 40-day

elimination Pruell et al (1986)

(1986) Mutiple aquatic

organisms Multiple HOCs Hawker and Connell(1986)

P viridis Field PCB, 54 17-day uptake, 32-day

elimination

Tanabe, Tatsukawa, and Phillips (1987)

elimination

Tanacredi and Cardenas (1991) Oysters Water only PCB, 77 Sericano et al (1992)

M edulis Field PCBs 28-day exposure Bergen, Nelson, and

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D polymorpha No change in k 1 within a

season, but change between seasons

Reeders, Bij de Vaate, and Slim (1989)

E complanata BCF, k 2 Russell and Gobas (1989)

D polymorpha k 1 Fisher et al (1993)

D polymorpha k 1 , k 2 , BCF/BAF Brieger and Hunter (1993)

D polymorpha k 1 , k 2 , BCF, T 1/2 k 2 depends on the lipophilicity

of a chemical

Bruner, Fisher, and Landrum (1994)

A anatina,

P complanata BCF Makela and Oikari (1995)

D polymorpha k 2 , T 95 Morrison et al (1995)

D polymorpha k 1 ,k 2 , BCF, T 1/2 Temperature effects,

Chevreuil et al (1996)

D polymorpha Bivalves: have MFO but

capabilities are/fish Van Haelst et al (1996b)

D polymorpha k 1 , k 2 , BCF, T 1/2 Log K ow versus k 2 :

independent; mussel lipid decrease over time

Van Haelst et al (1996a)

C flumina k 2 No extensive phase I

metabolism Basack et al (1997)

C leana k 1 , k 2 , BCF Measured pesticide

concentrations in mussels in rice patties

Uno et al (1997)

D polymorpha k 1 ,k 2 , BCF, T 1/2 Temperature and pH effects Fisher et al (1999)

PAH uptake due to partitioning from water to animal across gill surface

Thomann and Komlos (1999)

E complanata Average uptake of ANZ

COON0 Birdsall, Kukor, andCheney (2001)

E complanata k 2 Linear relationship between

Lee, Sauerheber, and Benson (1972) Oysters Elimination nearly complete

after 28 days

Stegman and Teal (1973)

M edulis k 2 dependent on chemical

lipophilicity

Clark and Findley (1975)

(continued)

authors noted that others (e.g., Reeders, Bij de Vaate, and Slim 1989) have documented a lack ofsubstantial change in D polymorpha filtration activity over a temperature range of 5–208C, whichhelps to explain their data (Gossiaux, Landrum, and Fisher 1996) While k1s for some ofthe compounds in this study increased proportionally with increasing temperature in the field

Table 8.2b (Continued)

Species Variables Measured Primary Findings References

Mussels k 2 dependent on chemical

lipophilicity Dunn and Stich (1976)Clams Slight elimination after

120 days Boehm and Quinn (1977)

M edulis High lipid tissuesZrapid

elimination versus low lipid tissuesZslower

elimination: biphasic

Hansen et al (1978)

O edulis Gill: primary site: uptakeC

accumulation Riley et al (1981)

T japonica Rapid PAH accumulation Obana et al (1983)

Clam, oyster, mussel k 1 , k 2 , BCF Ogata et al (1984)

Oysters Analytes below detection limit

within 4 days of elimination

Pittinger et al (1985)

M edulis k 2 , BCF, T 1/2 Slow elimination observed and

k 2 depends on liophilicity of chemical

Pruell et al (1986)

M edulis High lipid tissuesZrapid

elimination versus low lipid tissuesZslower

Hawker and Connell (1986)

P viridis k 2 , T 1/2 , T 90 Rapid uptake, release of lower

K ow PCBs Tanabe, Tatsukawa, andPhillips (1987)

Oysters Equilibrium attained in 30 days Serciano et al (1992)

M edulis BCF Coplanar PCBs reach

steady-state faster (7 days) than nonplanar PCBs (14–

Gilek, Bjork, and Naef (1996)

M edulis k 2 k 2 unaffected by [POC], and

initial uptake rapid Bjork and Gilek (1996)

C virginica k 2 , T 1/2 Sericano, Wade, and

Brooks (1996)

M edulis k 1 , k 2 , BAF, T 1/2 Physioligically-based model of

bioaccumulation, food ration affected k 1 , but not k 2

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(e.g., BaP and HCBP), the trend was not consistently exhibited over the three-year time frame andled the authors to suggest that uptake kinetics do not change in a proportional manner withtemperature (Gossiaux, Landrum, and Fisher 1996), at least across the range tested AlthoughReeders, Bij de Vaate, and Slim (1989) reported no significant change in uptake rates in

D polymorpha within a season, a significant change between seasons was documented

Variations in uptake rates with D polymorpha body size and lipid content were reported byBruner, Fisher, and Landrum (1994) for HCBP, tetrachlorobiphenyl (TCBP), BaP, and PY.The average uptake rate constant for HCBP over varying mussel lipid and size was23,680 mL/g day (Bruner, Fisher, and Landrum 1994), which compared well with k1s reported

by Gossiaux, Landrum, and Fisher (1996) for D polymorpha over varying temperatures, aging 18,624 mL/g day in the laboratory and 21,000 mL/g day in the field When varying pH isconsidered in combination with changing temperatures, differences in k1s increase but are stillwithin a factor of less than five on average, which translates into about an order of magnitudedifference in BCF values The reported field and laboratory k1s in D polymorpha for PCP(log Kow 5.12) are 2,760 and 4,120 mL/g day (Gossiaux, Landrum, and Fisher 1996), whereasthose reported for varying pH (and averaged over temperature) are lower: 1,657 (pH 6.5), 1,218(pH 7.5), and 868 (pH 8.5) (Fisher et al 1999) The lesser k1s may be due to the dissociable nature

aver-of PCP in the range aver-of ambient pH (pKaZ4.74) or to a combination of effects caused by changing

pH and temperature on mussel filtration rates and subsequent uptake rates When individualvalues are compared, rather than averages, the variation in k1 is increased For instance, thesmaller the mussel size (measured in shell length), the faster the uptake rate (Bruner, Fisher,and Landrum 1994) Large (21 mm) zebra mussels with high lipid content (greater than 9%) hadTCBP uptake rate constants of 10,080 mL/g day, whereas smaller (15 mm) but also higher lipidcontent mussels (greater than 9%), had TCBP uptake rate constants twice that of the largermussels at 23,760 mL/g day (Bruner, Fisher, and Landrum 1994)

In general, uptake rates were directly proportional to compound Kow; as Kow increased, k1increased as well For example, as log Kowvalues increased from 5.18 for PY to 6.90 for HCBP, theaverage uptake rate constant increased from 10,480 to 23,680 mL/g day, respectively Anadditional study reported k1s ranging from 2,976 to 25,752 mL/g day in D polymorpha forPAHs, PCBs, and OCs (DDT) spanning a similar log Kowrange of 5.2–6.7 (Fisher et al 1993).This range is comparable to the other k1s previously listed, when values for DDT are omitted(lowest values) Moreover, k1values reported for HCB (hexachlorobenzene) and OCS (octachlor-ostyrene) in E complanata also increased with increasing log Kow; from 650/day for HCB (log Kow5.45) to 1,010/day for OCS (log Kow6.29) (Russell and Gobas 1989) However, these values aresubstantially less than those reported for D polymorpha

In contrast to the linear relationship between k1and Kowreported by some (Russell and Gobas1989; Bruner, Fisher, and Landrum 1994; Gossiaux, Landrum, and Fisher 1996), uptake rates foreight different TCBT congeners in D polymorpha were independent of Kow (Van Haelst et al.1996a) As log Kowincreased from 6.73 (TCBT # 28) to 7.54 (TCBT # 25), k1s varied little, from

772 to 803 mL/g day (Van Haelst et al 1996a), respectively However, when all TCBT congenerswere included in the log Kowrange, the k1values demonstrated larger variation and ranged from683.3 to 848.8 mL/g day This may be partially explained by the high Kowvalues or the decreasedability of highly hydrophobic compounds to permeate membranes (Van Haelst et al 1996a) More-over, the uptake rates reported for D polymorpha for TCBT congeners are lower than those forPAHs or PCBs with similar hydrophobicity (see previous values) Uptake rate constants for PCBcongener 153 (Bruner, Fisher, and Landrum 1994) and TCBT (tetrachorobenzyltoluene) congener

28 (Van Haelst et al 1996a), which have similar log Kowvalues (6.92 and 6.73, respectively), differ

by as much as a factor of 50, from as low as 771 mL/g day for TCBT congener 28 (Van Haelst et al.1996a) to between 9,120 and 38,592 mL/g day for congener 153 (Bruner, Fisher, and Landrum1994), both for D polymorpha

Bjork and Gilek (1997) reported k1s for three PCB congeners (PCBs 31, 49, 153) in the marinemussel, M edulis, that ranged from 2,160 (PCB153) to 168,000 mL/g day (PCB153) While theupper range is quite large, and is about four times greater than the upper range reported for

D polymorpha, the freshwater mussel k1s are still within these limits The larger k1 values in

M edulis are probably due to the addition of contaminated food in the study conducted by Bjorkand Gilek (1997) In contrast, Ogata et al (1984) reported k1s for parent and various alkylateddibenzothiophenes in a marine short-necked clam, which were significantly less ranging from 33/day for dibenzothiophene to 66/day for dialkylated dibenzothiophene It should be noted that someauthors (e.g., Ogata et al 1984; Russell and Gobas 1989) have reported k1values in reciprocal days,which is assumed to be equivalent to mL/g day (where 1 mLZ1 g) However, this assumption maynot always be valid, which may explain some of the differences observed in k1values

Uptake rates for various pesticides in the Asian clam Corbicula leana (Uno et al 1997)are much lower than those reported in D polymorpha for compounds with similar Kows Whilethe log Kowfor the pesticides thiobencarb, oxadiazon, and chlornitrofen are less than the HOCs, theuptake rate constants are more than proportionally less, ranging from 24.2 for thiobencarb to626.0 mL/g day for chlornitrofen in the field and 140 for thiobencarb to 338 mL/g day for chlorni-trofen in the laboratory (Uno et al 1997) The authors attributed the low uptake rate(s) forthiobencarb to a temperature decrease of 28C over the course of a year causing slower ventilationrates in the mussels In contrast, reports with D polymorpha show that a temperature range of 208Cdoes not cause substantial changes in uptake rates (Reeders, Bij de Vaate, and Slim 1989; Gossiaux,Landrum, and Fisher 1996) The large differences in uptake rates for C leana versus D polymorphaand M edulis are probably due to a combination of species and chemical differences

In summary, uptake rate constants were remarkably similar across temperature, season, pH,chemical, and study variables, although some differences were observed, particularly whencomparing chemicals of similar log Kow (TCBTs versus PCBs), low versus high lipidcontent, and bivalves of differing size and species Large variation in k1 was demonstratedfor stressed mussels (Gossiaux, Landrum, and Fisher 1996), suggesting that bivalve physiologymust be considered when measuring empirical uptake rates or BCFs under adverse conditionssuch as very low temperatures Moreover, k1s were greater for combined food and waterexposures (Bjork and Gilek 1997) The uptake rate constants reported in this chapter representonly those for a few freshwater mussel and clam species, which demonstrates the need forfurther research in this area For instance, while D polymorpha uptake rate constants may notvary substantially with increases or decreases in temperature (over a 208C range) (Gossiaux,Landrum, and Fisher 1996), this may not be the case for other freshwater bivalve species(e.g., Corbicula) (Uno et al 1997)

BIOCONCENTRATION

Gossiaux, Landrum, and Fisher (1996) reported bioconcentration factors in D polymorpha forBaP (log Kow 6.04) that ranged from 4.38 to 5.28 log bioconcentration in field exposures attemperatures from 4 to 248C The BaP log BCF values had a similar range in the laboratory fortemperatures from 4 to 208C (4.60 (48C) to 5.43 (158C),Table 8.1b).The log BCF values for PY(log Kow5.18) in both the field and laboratory ranged from 4.34 to 4.89, over a similar temperaturerange However, the authors were not convinced that steady-state had been reached due to a factor

of 100 difference between BCF values calculated from Ctissue/Cwaterand those calculated from k1/

k2 This implies BCF values in reality would be larger than those reported or that the organismspossess some capacity for metabolism of PY In comparison, Bruner, Fisher, and Landrum (1994)reported similar log BCF values also in D polymorpha for both BaP, ranging from 4.61 to 4.92 and

PY, ranging from 4.11 to 4.54, depending on mussel lipid and size These values compare well,especially when considering the variation in temperature, lipid content, and size

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In contrast, log BCF values reported by Thorsen (2003) for E complanata are lower for bothPAHs, ranging from 3.50 to 4.66 for BaP, and 2.29 to 3.79 for PY, depending on exposure source(water-only versus sediment) The differences between these data may be partially explained by lipidcontent of D polymorpha and E complanata; D polymorpha were generally 7–15% lipid on a dryweight basis (Gossiaux, Landrum, and Fisher 1996) whereas E complanata were much lower,typically 3–4% lipid (Thorsen 2003) The contribution of a lipid can be partly confirmed byresults from Bruner, Fisher, and Landrum (1994) who reported an increase in BCF values withsubsequent increase in mussel lipid content However, this effect was only observed for the higher

Kowcompounds (HCBP and BaP) and not for the lower Kowcompounds (TCBP and PY) Whilebivalve BCF values are not traditionally normalized to lipid content, it would be helpful to report lipidvalues for conversion and comparison The addition of lipid-normalized BCF values may help toexplain variations in bivalve bioconcentration Furthermore, log BCF values determined for HCBand OCS in E complanata (log Kows 5.49 and 6.29, respectively) compare well with those for PAHs

of similar hydrophobicity, ranging from 3.56 to 4.16 (e.g., 3.58 and 3.64 for C2-dibenzothiopheneswith log Kow 5.50, and 4.23 and 4.54 for benzo(e)pyrene with log Kow6.20) (Thorsen 2003).Additional variations in BCFs may be further explained by physiological differences between

E complanata and D polymorpha, differences in study design, or a combination of environmentaland physiological factors

Makela and Oikari (1995) reported BCF values for PCP in two freshwater mussels, Anodontaanatina and Pseudanodonta complanata, which range from 1.9 to 2.1 and 1.8 to 1.9, respectively.These BCF values are much lower than those reported by Gossiaux, Landrum, and Fisher (1996) forPCP in D polymorpha, which ranged from 4.00 to 5.27, depending on the study temperature Incontrast, log BCF values reported for PCP in a different study for D polymorpha with varyingtemperature and pH are mid range between those reported for A anatina, P complanata (with arange of 2.60–3.13) (Fisher et al 1999), and D polymorpha (4.00–5.27) (Gossiaux, Landrum, andFisher 1996,Table 8.1b)

The log BCF values for HCBP determined in two separate studies on D polymorpha comparewell, ranging from 4.79 to 5.38 in one study (Bruner, Fisher, and Landrum 1994) and from 5.24 to5.74 in the second (Gossiaux, Landrum, and Fisher 1996) Brieger and Hunter (1993) reportedlog BAF values for D polymorpha of 4.02 and 4.45 for two PCB congeners, 77 (log Kow6.36)and 169 (log Kow 7.42), which were lower relative to their Kowvalues than those reported forsimilar log Kowcompounds such as TCBT congener 28 (log Kow6.73, log BCF 4.83) (Van Haelst

et al 1996a, 1996b), HCBP (log Kow6.9, log BCF range 4.8–5.7) (Bruner, Fisher, and Landrum1994; Gossiaux, Landrum, and Fisher 1996), and TCBT congener 22 (log Kow7.43, log BCF 4.71)(Van Haelst et al 1996a, 1996b) These differences may simply suggest a lack of steady state, as BAFvalues would be expected to be larger than BCF values from increased exposure tocontaminated food

The values of log BCF for various pesticides including chloronitrofen, thiobencarb, andoxadiazon have been reported for C leana ranging from 2.34 for oxadiazon (log Kow 3.89) to4.14 for chlornitrofen in the field, and from 3.79 for chlornitrofen to 3.45 for thiobencarb(log Kow4.22) in the laboratory (Uno et al 1997) It should be noted that the log BCF values foroxadiazon and thiobencarb increase with corresponding increases in hydrophobicity

Bioconcentration factors determined for PAHs in M edulis (Pruell et al 1986) and a marineshort-necked clam, oyster, and mussel (Ogata et al 1984) compare well to those for E complanata(Thorsen 2003) but are less than those reported for D polymorpha (see previous comparison between

E complanata and D polymorpha) For example, across a log Kowrange of 3.9–6.1, log BCF valuesfor M edulis ranged from 2.0 to 4.4 (Pruell et al 1986), whereas across a similar log Kowrange of3.37–7.60 for E complanata, log BCF values ranged from 1.5 to 5.2 (Thorsen 2003) Moreover, thelog BCFs reported for dibenzothiophene (D0) in marine clam, oyster, and mussel were 2.17, 3.12,and 3.13, respectively (Ogata et al 1984), which are near the range reported for E complanata of

2.69–2.93 (Thorsen 2003) and similar to those reported for thiobencarb (of similar log Kowto D0:4.22 versus 4.49) in C leana of 3.25–3.48 (Uno et al 1997,Table 8.3).

Similar to uptake rate constant data, empirically derived BCF values generally increase withincreasing Kowof the compound (Pruell et al 1986; Brieger and Hunter 1993; Bruner, Fisher, andLandrum 1994; Gossiaux, Landrum, and Fisher 1996; Thorsen 2003) For example, as the log Kow

is increased from 5.18 (PY) to 6.90 (HCBP), the average log BCF values for D polymorphaincrease from 4.28 to 5.14 (Bruner, Fisher, and Landrum 1994) However, exceptions to thistrend have been observed The BCFs for compounds with log Kowvalues greater than 6–7 tend

to level off due to factors such as steric hinderance (reduction of membrane permeation), lack ofsteady state (very long times required to reach equilibrium), and growth dilution Van Haelst et al.(1996a, 1996b) found no correlation between log BCF values for eight TCBT congeners andlog Kow They suggested that this was due to the small range of log Kow(6.73–7.54) compoundsused, as well as the fact that the TCBT congeners all have log KowsO6 (i.e., may be in the linearpart of the curve)

The values of log BCF for PCBs of similar hydrophobicity reported for M edulis were higherthan those for PAHs: ranging from about 5.0 to 5.7 for a corresponding PCB log Kow range

of approximately 6.0–7.0 (Pruell et al 1986) This log BCF range fits within that reported for

D polymorpha (Bruner, Fisher, and Landrum 1994; Gossiaux, Landrum, and Fisher 1996) forvarious PCBs over the same log Kow range of 4.0–6.9 However, the differences between PAHand PCBs for freshwater mussels appear to be less pronounced (Bruner, Fisher, and Landrum 1994;Gossiaux, Landrum, and Fisher 1996) Moreover, a linear relationship between log Kow andlog BCF was observed for both PAHs and PCBs in M edulis (Pruell et al 1986), E complanata(Thorsen 2003), and D polymorpha (Bruner, Fisher, and Landrum 1994) Comparisons of steady-state bioconcentration regression equations(Table 8.4)demonstrate reasonable agreement in PAHaccumulation, with few exceptions For example, Pruell et al (1986) reported a slope of 0.965 and ay-intercept of K1.41 for M edulis, whereas Thorsen (2003) reported a slope of 0.895 and ay-intercept of K1.21 (r2Z0.8325) for E complanata However, Ogata et al (1984) reportedregression equations with slopes much less than one (0.16 for short-necked clams, 0.49 foroysters, and 0.31 for mussels) and positive y-intercepts (1.54, 1.03, 1.63, respectively) Thedifferences may be due to the fact that the regression equations of Ogata et al (1984) werebased on the parent and alkyated homologues of dibenzothiophene only, whereas those of Pruell

et al (1986) and Thorsen (2003) were based on data sets containing greater numbers of PAHs.These data suggest a good correlation between marine and freshwater BCF values, for M edulis,

E complanata, and Mya arenaria

ELIMINATION

The elimination rate constant (k2) can be calculated from an elimination plot of the lipidnormalized, natural log (ln) of the contaminant concentration in bivalve versus time In a first-order, one-compartment kinetic model, k2is the absolute value of the slope of the line, based on theequation

ln CtissueZKk2t Cln Ctissue;0where Ctissue,0is the tissue chemical concentration at elimination time zero(Figure 8.4)

Bivalve elimination rate constants are also fairly consistent, depending on compound, study,and species(Table 8.1a).Elimination rates for HOCs are generally much lower than their counter-part uptake rates but similarly are dependent upon the hydrophobic character of the compounds(Dunn and Stich 1976; Bruner, Fisher, and Landrum 1994; Morrison et al 1995; Gewurtz et al.2002; Thorsen et al 2004a) Gewurtz et al (2002), who calculated k2s for nine PAHs in

E complanata, showed variation from 0.037/day for benzo(k)fluoranthene (BkF) to 0.217/dayToxicokinetics of Environmental Contaminants in Freshwater Bivalves 187

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Comparison of Bioconcentration/Bioaccumulation Factors and Uptake and Elimination Rate Constants for Similar Solubility HOC

Analytes in the Peer-Reviewed Reference

laboratory

Thorsen (2003)

Gilek (1997)

laboratory Thorsen (2003)

laboratory Thorsen (2003)

Pentachlorophenol 5.20 P complanata 1.80–1.90 Water Makela and

a Units not specified in reference.

for fluoranthene (FL) An inverse linear relationship was observed between analyte elimination rateconstant and corresponding log Kow, which the authors attributed to potential passive elimination ofPAHs (Gewurtz et al 2002) This type of response is characteristic of monophasic, first-orderelimination, which has also been reported in D polymorpha for lower Kowcompounds (Gossiaux,Landrum, and Fisher 1996) Additional k2values reported for 45 PAHs in E complanata exposed tosediment during the uptake phase ranged from 0.04 to 0.22/day (Thorsen et al 2004a) The authorsfurther noted that these elimination rate constants were less than those from a water-only exposurestudy and suggested that it may have been due to increased stress on the mussels from an unknownfungal or bacterial growth, and subsequently, increased handling (Thorsen et al 2004a) The k2values for OCS and HCB in E complanata ranged from 0.16 to 0.41/day and were slightly higherwhen compared to similar log KowPAHs (Russell and Gobas 1989).

Moreover, Gossiaux, Landrum, and Fisher (1996) demonstrated slow elimination rate constantsfor D polymorpha in field studies, ranging from 0.024 to 0.096/day for HCBP to 0.024 to 0.384/dayfor BaP For the lower hydrophobic compounds in this study (PCP and PY), elimination was rapidduring the first 24 hours and then leveled off, while elimination of HCBP and BaP was minimal overthe first 24 hours, increased during the following 48–168 hours, and then slowed, suggestive of abiphasic, two-compartment model These authors, however, classified the elimination

as monophasic

Furthermore, k2s from the studies of Gossiaux, Landrum, and Fisher (1996) and Gewurtz et al.(2002), in D polymorpha and E complanata, compared well among HOCs of similar log Kow Forexample, Gewurtz et al (2002) reported a k2for PY of 0.144/day, and this is within the range also

TABLE 8.4

Comparison of Steady-State Bioconcentration Regression Equations for Organic

Contaminants in Freshwater and Marine Mussels

Species ChemicalClass Exposure Slope y-Intercept r 2 References Freshwater

E complanata PAH (34) Water,

laboratory

0.895 K1.21 0.83 Thorsen (2003)

E complanata PAH (35) Sediment, field 0.786 K0.98 0.78 Thorsen (2003)

E complanata PAH (45) Sediment, lab 0.807 K1.12 0.73 Thorsen (2003) Marine

M edulis PAH (6) Sediment,

laboratory 0.965 K1.40 Pruell et al.(1986)

M edulis Multiple HOCs Water,

laboratory 0.858 K0.81 0.96 Geyer et al.(1982)

M arenaria PAH Water, field 1.097 K1.54 0.85 Thorsen (2003)

M arenaria PAH Sediment, field 1.042 K1.28 0.85 Thorsen (2003) Multiple

marine

Multiple HOCs 0.844 K1.23 0.83 Hawker and

Connell (1986) Marine clam PAH (4, all D0) Water,

laboratory 0.163 1.52 0.71 Ogata et al.(1984) Marine oyster PAH (4, all D0) Water,

laboratory

0.494 1.03 0.62 Ogata et al.

(1984) Marine mussel PAH (4, all D0) Water,

laboratory

0.311 1.63 0.64 Ogata et al.

(1984)