Factors controlling the bioaccumulation of mercury and methylmercury by the estuarine amphipod Leptocheirus plumulosus

Environmental Pollution 111 (2001) 217±231

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Factors controlling the bioaccumulation of mercury and methylmercury by the estuarine amphipod Leptocheirus plumulosus A.L. Lawrence, R.P. Mason * Chesapeake Biological Laboratory, Center for Environmental Science, The University of Maryland, PO Box 38, Solomons, MD 20688, USA Received 24 March 1999; accepted 17 January 2000

``Capsule'': Accumulation of HgI and MMHg by benthic organisms cannot be predicted using total concentrations in sediments as the sole factor. Abstract The bioaccumulation of inorganic mercury (HgI) and monomethylmercury (MMHg) by benthic organisms and subsequent trophic transfer couples the benthic and pelagic realms of aquatic systems and provides a mechanism for transfer of sedimentary contaminants to aquatic food chains. Experiments were performed to investigate the bioavailability and bioaccumulation of particle-associated HgI and MMHg by the estuarine amphipod Leptocheirus plumulosus to further understand the controls on bioaccumulation by benthic organisms. HgI and MMHg are particle reactive and have a strong anity for organic matter, a potential food source for amphipods. Microcosm laboratory experiments were performed to determine the e€ects of organic matter on Hg bioaccumulation and to determine the major route of Hg uptake (i.e. sediment ingestion, uptake from water/porewater, or uptake from `food'). Amphipods living in organic-rich sediment spiked with Hg accumulated less Hg than those living in sediments with a lower organic matter content. Feeding had a signi®cant impact on the amount of HgI and MMHg accumulated. Similarly, amphipods living in water with little organic matter accumulated more Hg than those living in water with a greater percentage of organic matter. MMHg was more readily available for uptake than HgI. Experimental results, coupled with results from a bioaccumulation model, suggest that accumulation of HgI and MMHg from sediment cannot be accurately predicted based solely on the total Hg, or even the MMHg, concentration of the sediment, and sediment-based bioaccumulation factors. All routes of exposure need to be considered in determining the accumulation of HgI and MMHg from sediment to benthic invertebrates. # 2000 Elsevier Science Ltd. All rights reserved. Keywords: Mercury; Bioaccumulation; Bioavailability; Amphipods; Organic matter; Sediment

1. Introduction Anthropogenic inputs of mercury (Hg) to the aquatic environment have substantially increased in the last century (Mason et al., 1994; Engstrom and Swain, 1997; Fitzgerald et al., 1998; US EPA, 1997a). In coastal environments, urban runo€, industrial discharge and atmospheric deposition of Hg have resulted in elevated Hg levels in sediment, the main repository for Hg in these environments (e.g. US EPA, 1997b; Mason and Lawrence, 1999). Sediments are the dominant location for Hg methylation in estuaries (Benoit et al., 1998; Gilmour and Henry, 1991) and, thus, as a result of the anthropogenic inputs, benthic and epibenthic organisms * Corresponding author. Tel.: +1-410-326-3787; fax: +1-410-3267341. E-mail address: [email protected] (R.P. Mason).

in coastal waters often contain elevated levels of Hg, especially methylmercury (MMHg) (e.g. US EPA, 1997b; Mason and Lawrence, 1999). Accumulation of Hg into higher trophic level aquatic organisms is a cause for concern and a current regulatory focus (Clarkson, 1990; US EPA, 1996) because Hg, principally as MMHg, bioaccumulates through all levels of the aquatic food chain (Lindqvist et al., 1991; Watras and Bloom, 1992; Kidd et al., 1995) and ®sh consumption is the dominant source of Hg to humans (Clarkson, 1990). To evaluate the biological signi®cance of Hg contamination in coastal sediments, this study focused on determining the factors a€ecting the bioavailability and bioaccumulation of inorganic Hg (HgI) and MMHg to the estuarine amphipod Leptocheirus plumulosus. Amphipods were chosen because they: (1) dwell at the sediment±water interface; (2) feed by ®lter feeding on suspended particulate matter and by deposit feeding on detritus and sediment

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(DeWitt et al., 1992); (3) are a major component of the estuarine aquatic food web; (4) are sensitive to sediment-bound contaminants (Schlekat et al., 1992); and (5) have been chosen as an EPA sediment toxicity test species (DeWitt et al., 1992). Lower trophic levels play a major role in Hg bioaccumulation into ®sh as, for example, in pelagic systems, the greatest bioconcentration occurs between the water and phytoplankton (Lindqvist et al., 1991; Watras and Bloom, 1992; Mason et al., 1996). For benthic organisms, however, there is little information on the factors controlling HgI and MMHg accumulation from their surrounding environment, and the importance of sediment and/or water chemistry in controlling bioaccumulation (Gagnon and Fisher, 1997). As organic matter (OM) complexation plays an important role in controlling Hg speciation in aquatic systems (Hudson et al., 1994; Hintelman et al., 1995), we hypothesized that OM would in¯uence Hg accumulation at the base of estuarine benthic food chains. When assessing the potential exposure of benthic organisms, especially amphipods which both suspension and deposit feed, it is necessary to examine four exposure routes: water, porewater, inplace sediment, and suspended particulate matter/food. We present here the results of these experiments plus a simple model developed to assess the importance of each exposure route to MMHg bioaccumulation into amphipods under ®eld conditions. 2. Experimental design and methods Three types of microcosm laboratory experiments were performed to investigate accumulation: a sediment exposure, a feeding (sediment/algae) exposure, and a water exposure. Sediment was collected from Fishing Bay, MA, USA, an unpolluted site; and sand from the Patuxent River mouth. Experimental treatments of varying OM (nine sediment mixtures) were obtained by diluting Fishing Bay sediment with mu‚ed (550 C for 24 h) sand, both sieved to less than 250 mm. These mixtures were then spiked with solutions made from certi®ed reference standards of Hg (Fisher Scienti®c, Inc.) and MMHg (Brooks Rand Ltd.) and stored in the dark at 4 C for 4 days. The target concentrations were 500 ng gÿ1 HgI and 50 ng gÿ1 MMHg. A 3-cm layer (300 g) of sediment mixture was added to each experimental microcosm (2-l beaker) over which 1.5 l of ®ltered (0.2 mm) ambient seawater (10 ppt) was added. The experimental microcosms were kept in a temperature-controlled room (25 C) and exposed to a photoperiod of 16 h of light, continually aerated and allowed to equilibrate for 24 h prior to the addition of amphipods. These protocols were used for all experiments discussed in this paper. Two sediment exposure experiments were performed. The ®rst covered the full range of OM encountered in the estuarine

environment, while the second experiment, based on results from the ®rst, focused on low OM sediments. Amphipods were obtained from cultures maintained by the Maryland Department of Natural Resources at the Chesapeake Biological Laboratory (CBL) following recommended methods from DeWitt et al. (1992). Amphipods (>500 mm, at least 7 days old; Schlekat, personal communication), sieved from their stock tank, were allowed to depurate for 4 h before addition to the microcosms (40 amphipods/microcosm). This depuration time is two to three times the mean gut passage time of L. plumulosus (95 min; measured by using 10 mm plastic beads impregnated with 57Co; Schlekat, personal communication). The amphipods were not fed and the water not changed over the 6-day duration of the sediment-only experiment to avoid any confounding factors associated with changes in sediment geochemistry or unwanted algal growth. Samples of water, sediment, and amphipods were collected initially: on day 3 for Experiment 1; and at the termination of the experiment (Experiments 1 and 2). The amphipods were depurated for 4 h before being frozen for subsequent analysis. Porewater was collected at the termination of Experiment 2 only from one of each of the triplicate exposures. The overlying water was siphoned from the beaker, the sediment transferred to an acid-cleaned suction ®ltration device with a 0.8mm membrane and the porewater collected by suction ®ltration. After re®ltering through a 0.2-mm membrane, the samples were stored frozen in Te¯on1 vials for subsequent analysis. During the experiments water quality parameters (temperature, salinity, pH, dissolved oxygen, and ammonia) of overlying water were monitored (ASTM, 1990) and remained relatively constant. The average pH and dissolved oxygen were 7.9 and 7.1 mg lÿ1, respectively. The concentration of ammonia was below 2 mg lÿ1. Each experimental treatment was performed in triplicate. Two sets of Hg-free controls were also run in triplicate, typically under the end-member conditions (i.e. pure sand/mud in Experiment 1; low and high OM in Experiment 2). No dissolved organic carbon (DOC) measurements were made in Experiment 1. In Experiment 2, DOC measurements were compromised in the spiked treatments by the fact that the MMHg spike solution was a mixture of propanol and water. Concentrations were, however, measured in the controls (5.4 mg/lÿ1 in the water and 41 mg/lÿ1 in the porewater for the 3.85% OM control; 4.4 and 36.3 mg/l in the 1.24% OM control). In the feeding experiments, water column DOC values were similar, but porewater values were somewhat lower (Table 2). For the feeding exposures, sediment mixtures were spiked in a manner similar to the sediment exposure study. Cultures of Isochrysis (TISO), an autotrophic ¯agellate, were grown in glass carboys according to

A.L. Lawrence, R.P. Mason / Environmental Pollution 111 (2001) 217±231

standard algal culture methods (Guillard, 1975) and spiked with Hg and MMHg in late exponential growth phase. The cultures were stored at 20 C under constant light for 24 h to allow assimilation of Hg to occur. Given typical growth rates of TISO (2 doublings dayÿ1), algae would have assimilated 80% of their steady-state concentration after 24 h. Aliquots of algae (50 ml) were placed in acid-cleaned polycarbonate centrifuge tubes and centrifuged at 2700 g for 30 min at room temperature to separate the algal cells from the water. The overlying supernatant was removed and the algal pellet frozen. The pellet was thawed and resuspended in 3.5 ml of ®ltered, deionized water before addition to the experimental microcosms. For the sediment/feeding experiments, microcosms were assembled and maintained in a fashion similar to that in the sediment-only exposures. The OM at the end of the experiments was similar in all beakers (4.1‹2.0% OM). Samples of water, sediment, and amphipods were collected initially and at the termination of the experiment (Day 8). During this experiment, the average pH and dissolved oxygen were 7.9 and 7.3 mg lÿ1, respectively. The concentration of ammonia was below 2 mg lÿ1. The feeding exposure design included four experimental treatments in triplicate; two spiked and two unspiked sediment treatments to which spiked or unspiked algae was added. Algae was added every other day for a total of four feedings (2.1 mg unspiked algae/ amphipod/feeding and 2.5 mg spiked algae/amphipod/ feeding), every other day; a rate similar to that recommended by DeWitt et al. (1992) for a 10-day standard toxicity test. Note that the amount of OM added during the feeding Ð around 400 mg Ð was trivial compared to the total OM the sediment (300 g4% OM=12 g of OM). Thus, addition of food did not dramatically change the sediment OM characteristics. A further experiment was performed to determine the amount of HgI and MMHg that could be accumulated by amphipods from water. The design included four experimental treatments of varying dissolved OM content at di€erent Hg concentrations and two Hg-free controls. Each microcosm (2-l beaker), containing 2 l of ®ltered (0.2 mm) ambient seawater (10 ppt) and 50 amphipods, was spiked with Hg solutions. The water was completely renewed and respiked daily to control water column concentrations. The experimental microcosms were continually aerated and kept at 25 C. The dissolved OM added was Aldrich humic acid solution. Although Aldrich humic acid is terrestrially derived, it was used in these experiments because it was readily available, homogenous in source, and was used successfully in previous Hg binding studies (Lawson and Mason, 1998). Samples of amphipods were taken daily during the 5-day experiment. Water samples were taken on Days 1 and 4 and food was not added over the duration of the experiment. Water quality parameters

219

were again monitored; the average pH and dissolved oxygen were 8.0 and 7.9 mg lÿ1, respectively. The water and sediment-only exposures were of similar duration (5 and 6 days, respectively) and the results of these experiments are, therefore, directly comparable. The food experiment was extended to 8 days to ensure that the amphipods accumulated signi®cant Hg and MMHg from the food, as feeding was only every other day. As discussed below, the di€erences in experiment duration do not have a signi®cant e€ect on the ability to compare across treatments, given the length of time of these experiments relative to the time expected to reach steady state accumulation. All samples were stored frozen in acid-cleaned Te¯on1 vials. A composite of ®ve amphipods for each analysis were pre-weighed using a precision analytical balance. Samples for Hg were digested overnight in a VWR Scienti®c forced air oven at 60 C in a 70% sulfuric acid/ 30% nitric acid solution to ensure complete digestion of OM (Bloom and Crecelius, 1987). Under cleanroom conditions, the samples were further oxidized with bromine monochloride and the excess oxidant neutralized with 10% hydroxylamine hydrochloride. The samples were then reduced with tin chloride, sparged, and the elemental Hg trapped on a gold column for determination via cold vapor atomic ¯uorescence detection (CVAFS) (Bloom and Fitzgerald, 1988) in accordance with protocols outlined in EPA Method 1631 (US EPA, 1995). Samples were generally analyzed in duplicate and the error between these duplicates was 5% on average. Samples for MMHg were distilled with a 50% sulfuric acid/20% potassium chloride solution (Horvat et al., 1993). The distillate was reacted with a sodium tetraethylborate solution to convert the nonvolatile MMHg to gaseous methylethylmercury. The volatile adduct was then purged from solution and recollected on a graphitic carbon column at room temperature. The methylethylmercury was thermally desorbed from the column and analyzed by isothermal gas chromatography with CVAFS detection (Bloom, 1989). Reported values are the averages of concentrations of the triplicate exposures, except for pore water and the water exposure, in which only one sample was available. A standard calibration curve with an r2 of at least 0.999 was achieved daily. Hg standards, MMHg standards, laboratory duplicates, and standard reference material (SRM) of known Hg and MMHg concentrations were analyzed to ensure accuracy of our results. Duplicate analysis of 10% of the MMHg samples yielded no signi®cant di€erence, and >90% of all SRM replicates analyzed for Hg and MMHg fell within the certi®ed ranges. Spike recoveries for water, sediment, and tissue measurements ranged from 75 to 120% for Hg and 80 to 110% for MMHg. Detection limits for Hg and MMHg were: total Hg 0.27 ng lÿ1, MMHg 0.02 ng lÿ1 for aqueous samples; 1.3 ng gÿ1 Hg and 0.02 ng gÿ1

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MMHg for sediments; and 0.06 ng gÿ1 Hg and 0.015 ng gÿ1 MMHg for amphipods and algae (wet wt.). HgI was determined by di€erence (total HgÿMMHg). To calculate the dry/wet ratio of the sediment, weighed sediment samples were dried at 60 C and reweighed. After determining the water content of the sediment, the samples were placed in a furnace and heated to 550 C overnight. The samples were then reweighed and the percent OM in the sediment was determined by di€erence as loss on ignition. Organic carbon content was estimated assuming OM to be CH2O. Dry/wet ratio and percent OM determinations were generally analyzed in duplicate and the error between these duplicates was <10% on average. DOC measurements were performed on an OI Model 700 Carbon Analyzer. Following persulfate digestion (Menzel and Vaccaro, 1964), DOC was determined by the measurement of CO2 released from the chemical oxidation of organic carbon in the sample. CO2 was purged from solution, concentrated by trapping and detected via non dispersive infra red analysis (NDIR). Acid-volatile sulfur (AVS) analysis was performed using a modi®ed version of the EPA method (Cornwell and Morse, 1987). Wet sediments (2.0‹0.2 g) were digested, in a system ¯ushed with nitrogen, using degassed cold 6 N HCl. The evolved H2S was collected in a deaerated solution of zinc acetate and sodium acetate bu€er. The precipitated sul®de was then measured using a sul®de probe with a Pb titration. Detection limits for AVS were 0.01 mmol gÿ1 (dry wt.). 3. Results 3.1. Amphipod survival Amphipods were active and burrowing in the sediment during all the sediment exposure experiments.

There was a signi®cant overall increase (above baseline levels) in HgI and MMHg concentration in the amphipods exposed to spiked sediment, algae or water. However, there was no signi®cant di€erence between average amphipod survival in spiked (91%) and Hg-free treatments (90%) in the sediment-only exposures; nor in amphipod survival across sediment treatments of various OM content (range: 83±98%). Although the amphipods were not fed during the sediment-only experiments, our previous laboratory experiments, and the 10-day amphipod toxicity test standard procedures (ASTM, 1990), do not suggest that starvation signi®cantly a€ects amphipod survival over this time period (590% average survival). In the sediment plus algae experiment, the percent recovery of amphipods in all treatments exceeded 90%, similar to that for the experiments where amphipods were not fed. Survival was similar in the water exposure experiments (>88% survival in all treatments). 3.2. Sediment only exposure The concentration of HgI and MMHg in ambient seawater was <1 ng lÿ1 and <0.1 ng lÿ1 at the beginning of the exposure experiments and the amphipods concentrations averaged 24 ng gÿ1 HgI and 1.6 ng gÿ1 MMHg (wet wt.). At the termination of Experiment 1, the concentration of total Hg in overlying water ranged from 62 to 83 ng lÿ1 for the high OM sediments and was 960 ng lÿ1 for the sand-only experiment (Table 1); MMHg water measurements were compromised by contamination during analysis for this experiment. Additionally, there were no porewater collections in Experiment 1 and no DOC measurements were made. The concentrations of total Hg in water for Experiment 2 were of similar magnitude to those of Experiment 1: <104 ng lÿ1 (Table 1). MMHg concentrations ranged from 3.7 to 7.3 ng lÿ1 in spiked microcosms, and were <2.5 ng lÿ1 in the controls (Table 1). Porewater measurements showed

Table 1 Concentrations of inorganic mercury (HgI), and methylmercury MMHg in overlying water (OW) and porewater (PW) for the various exposure conditions in Experiments 1 and 2a Exp. no.

%OM in sediment

Total Hg in OW* (ng lÿ1)

MMHg in OW (ng lÿ1)

Total Hg in PW* (ng lÿ1)

MMHg in PW (ng lÿ1)

1 2 2 2 2 2 1 1 1

0 0.53 1.24 1.46 2.14 3.85 6.7 9.8 12.9

960 49 57 109 83 38 83 82 62

± 4.5 4.6 5.2 3.7 7.3 ± ± ±

± 27 49 64 170 85 ± ± ±

± 10.0 10.0 13.6 9.9 6.5 ± ± ±

Control Control a

1.24 3.85

35 6.7

2.4 2.0

22.6 16.4

2.0 2.3

The total (Hg) concentrations are given as MMHg values are not available for Experiment 1. Inorganic Hg (HgI)=Total HgÿMMHg. * OW=overlying water; PW=porewater.

A.L. Lawrence, R.P. Mason / Environmental Pollution 111 (2001) 217±231

values up to three times higher than the overlying water (Table 1). As discussed above, DOC concentrations were compromised in Experiment 2 except for the control exposures and ranged from 4 to 6 mg/lÿ1 in the overlying water; from 36 to 41 mg lÿ1 in porewater (Table 1). The sediment AVS concentrations were low (<0.1 mmol g ÿ1, dry wt.). In the feeding experiments (Table 2), DOC values in overlying water were similar but DOC concentrations in porewater were lower. Accumulation into amphipods did not signi®cantly deplete the sediment of HgI or MMHg Ð mass balance calculations indicate <20% accumulation of the sedimentary HgI and MMHg into amphipods in low OM treatments and <0.1% at high OM. Based on our measurements of sediment concentration at Days 3 and 6, the concentration of HgI and MMHg in the sediments did not change signi®cantly during the experiments. However, the exposure concentration in each treatment was not the same (Fig. 1), even given the experimental protocol that all the sediment mixtures were initially spiked to the same concentration. For both HgI and MMHg, the concentrations varied across treatments by more than an order of magnitude, increasing with sediment OM (Fig. 1a,b). The loss of HgI and MMHg must have occurred during sediment preparation and was likely due to absorption to the mixing container walls, as well as loss into overlying water which was discarded after this initial equilibration. A positive relationship between concentration in amphipods and sediment was not apparent in these

221

Table 2 Concentrations of inorganic mercury (HgI) and methylmercury (MMHg) in the various phases during the experiment with sediment plus algaea Parameter

HgI

MMHg

Unspiked sediment Spiked sediment Unspiked algae (dry wt.) Spiked algae (dry wt.) Water at end of experiment (un®ltered) Ð unspiked algae Water at end of experiment (un®ltered) Ð spiked algae Porewater at end Ð unspiked sediment Porewater at end Ð spiked sediment

2.38‹0.04 127‹32 100 2800 135‹23

0.11‹0.07 (4.6%) 33.7‹7.4 (21%) 12.5 (13%) 460 (16%) 17.8‹2.4 (12%)

250‹128

165‹50 (40%)

174

3.4 (1.9%)

645‹335

26.5‹23.9 (4.0%)

DOC of overlying water (mg lÿ1) DOC of porewater (mg lÿ1)

5.3‹0.3 6.6‹1.4

a Concentrations in ng lÿ1 for water or ng gÿ1 wet wt. for sediment and algae, unless otherwise stated. Values in parentheses are the percent MMHg for that phase. The values given are the average values for each type of sediment and algae Ð spiked and unspiked Ð as these did not vary signi®cantly as a result of the combination of treatments. See text for details. DOC, dissolved organic carbon.

laboratory exposures for either HgI or MMHg (Fig. 1c, d). A signi®cant correlation was, however, seen between sedimentary OM and the resultant concentration of both HgI and MMHg in the sediment, which was repeatable across experiments (Fig. 1a,b). This suggests

Fig. 1. Comparison of (a) sediment mercury and (b) methylmercury concentrations with sediment organic matter content and (c) amphipod mercury and methylmercury (d) concentrations in the sediment exposure.

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that OM is controlling the sediment `binding capacity' for HgI and MMHg, as has been suggested by others and by ®eld data (Mason and Lawrence, 1999; Muhaya et al., 1997; Langstrom, 1986; Nuutinen and Kukkonen, 1998). To remove the in¯uence of sediment concentration, bioaccumulation factors (SBAFs) Ð the concentration of Hg in the organism (ng gÿ1 wet wt.) relative to the concentration of Hg in the sediment (ng gÿ1 wet wt.) Ð were calculated; SBAFHgI is the bioaccumulation factor for HgI, SBAFMMHg for MMHg. Among treatments which only varied in Hg concentration (not OM), similar SBAFs were found, as expected. Both SBAFHgI and SBAFMMHg decreased as sediment OM increased (Fig. 2), with the largest decrease in sediments containing less than 2% OM.

Fig. 2. Comparison of amphipod sediment bioaccumulation factors and sediment organic content in the sediment exposure.

3.3. Feeding exposure At the beginning of this experiment, the concentration of HgI and MMHg in the amphipods was 27 and 2.4 ng gÿ1 wet wt., respectively. The concentrations of Hg and MMHg in algae, sediment and water are given in Table 2. The percent MMHg in the spiked algae was 16 and 13% in the unspiked algae, compared to 21% for spiked sediment and 4.6% for unspiked sediment (Table 2). All AVS concentrations were once again low (<0.1 mmol gÿ1, dry wt.). The exposure conditions were unspiked sediment/unspiked algae (US/UF), unspiked sediment/ spiked algae (US/SF), spiked sediment/unspiked algae (SS/UF), and spiked sediment/spiked food (SS/SF). As noted above, the addition of food did not signi®cantly change the sediment OM concentration, which were, respectively, 4.1, 5.1, 2.7 and 2.5%. The lowest HgI and MMHg amphipod concentration occurred in the control treatment (US/UF), but there was a signi®cant overall increase in HgI and MMHg concentration in the amphipods exposed to spiked sediment and/or algae (Fig. 3). The highest amphipod concentration occurred in the SS/SF treatment. The concentration of MMHg in the amphipods was greater than the concentration of HgI in all treatments except the SS/UF exposure (Fig. 3), even though the concentration of MMHg was <20% of the concentration of total Hg in both sediment and algae in all instances (Table 2). To further demonstrate this, the SBAF values for the two sets of experiments, sediment-only and sediment plus food are compared for sediments of similar OM (Fig. 4). The comparison reinforces the notion that addition of spiked food led to a much higher accumulation of MMHg compared to the experiment where amphipods were not fed.

Fig. 3. Concentration of inorganic mercury and methylmercury in amphipods exposed to spiked and unspiked sediment and algae for 8 days.

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223

Fig. 4. Amphipod sediment bioaccumulation factors with varying algal inputs in spiked sediment treatments of similar organic content.

3.4. Water exposure The amphipod concentrations, 28.6 ng gÿ1 HgI and 1.4 ng gÿ1 MMHg at the beginning of the experiment, did not change in the control microcosms during the experiment. The water exposure also illustrates the e€ect of OM on dissolved HgI and MMHg concentration, although the e€ect is less dramatic. As seen in Fig. 5a, measured HgI and MMHg concentrations were low in the ambient DOC treatments (4.7 mg lÿ1 DOC; 284 ng lÿ1 HgI; 10 ng lÿ1 MMHg) and highest in the high DOC microcosms (7.1 mg lÿ1 DOC; 320 ng lÿ1 HgI; 18 ng lÿ1 MMHg) even though each treatment was spiked at the same concentration. These results suggest that there was di€erential adsorption to the container walls. A positive relationship between concentration of HgI and MMHg in amphipods and water was not seen for the water exposures (Fig. 6), which agrees with results from the sediment exposure. During the water exposure, accumulation appeared to be initially rapid in most exposures and leveled o€, suggesting that a quasi steady state concentration had been reached after 4 days for both HgI and MMHg (Fig. 6). Mass balance calculations indicate that 10% or less of the total Hg was taken up by the amphipods. For MMHg, the percent accumulated was much higher; up to 60% of the added MMHg was accumulated by the amphipods. The calculation of water-based BAFs (WBAFHgI and WBAFMMHg) for the water exposure experiment was based on the average water and amphipod concentrations for Days 3±5 of the experiment. The

Fig. 5. Comparison of (a) inorganic mercury and methylmercury concentrations in the water column for beakers spiked at the same concentrations and (b) amphipod water-based bioaccumulation factors with varying dissolved organic carbon.

amphipod WBAFHgI and WBAFMMHg decreased with increasing DOC (Fig. 5b), illustrating the in¯uence of dissolved organic matter on Hg bioaccumulation. Once again, WBAFMMHg was greater than WBAFHgI (Fig. 5b).

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Fig. 6. Bioaccumulation of inorganic mercury and methylmercury by amphipods from water with varying dissolved organic carbon concentrations with time.

3.5. Methylmercury bioaccumulation model To further assess the importance of various uptake routes for the amphipod, a simple bioaccumulation model was developed for MMHg. The aim of the model was to estimate the relative importance of the four exposure routes (water, porewater, in-place sediment, food); i.e the results should be interpreted in a semiquantitative sense only as not all parameters are well known. The basic model was constructed using the results from the sediment exposure assays with amphipods in which there was no food addition, in conjunction with the results from water-only exposure trials. For the sediment exposures, the total amphipod concentration (AmpT) is given by: AmpT ˆ Ampsed ‡ Ampwater ‡ Amppore

…1†

where Ampsed is the ng gÿ1 accumulated from solid sediment ingestion; Ampwater that accumulated via uptake from water; and Amppore that accumulated from porewater; or: ÿ1 AmpT ˆ SS BAF  …ng gÿ1 sed † ‡ WBAFwater  …ng lwater † ‡ WBAFpore  …ng lÿ 1pore †

…2†

where SSBAF is the bioaccumulation factor from ingestion of solid sediment only. The exposure due to overlying water and porewater in the sediment exposures was estimated using the available data for HgI, MMHg and DOC in water and porewater (Table 1) and the laboratory-derived relationship between WBAF and water column DOC (Fig. 5b): log WBAFMMHg ˆ 1:74 ÿ 0:173‰DOC…mg lÿ1 †Š; r2 ˆ 0:999

…3†

The water-based accumulation Ð from both overlying water and from porewater Ð was then subtracted from the total accumulation during the experiment for each exposure to estimate bioaccumulation from sediment ingestion only. It is necessary to assign the relative time the amphipods spent exposed to water (x) and to porewater (1ÿx), as the animals are not exposed to both simultaneously. Now, rearranging Eq. (2) and substituting: SS BAFˆ ‰AmpT ÿxWBAFwater  …ng lÿ1 water † ÿ1 ÿ …1ÿx†  WBAFpore  …ng lÿ1 pore †Š=ng gsed

…4†

Based on observation during the experiments, where it was noted that the amphipods spent time on the sediment surface, and given that they actively irrigate their burrows, for the model it was assumed that the amphipods were exposed to overlying water 80% of the time

A.L. Lawrence, R.P. Mason / Environmental Pollution 111 (2001) 217±231

(x=0.8). The impact of this choice on the overall conclusion of the modeling is limited, as discussed below. Given the increase in both MMHg and DOC in porewater in these experiments, the accumulation of MMHg from porewater is always less than that from overlying water. Thus, by using a value of 0.8 for x, the importance of uptake from water is likely to be conservative (i.e. overestimated). Furthermore, as the model results for the feeding exposures show (see below), <30% of the total accumulation into amphipods is from the combination of the water exposure routes. The estimated SSBAFMMHg for each exposure was plotted against sediment particulate organic carbon (POC; mg C gÿ1 dry sediment) and the resultant curve was modeled using two exponential equations: For POC<10 mg C gÿ1 sediment; log SS BAFMMHg …solid† ˆ 3:04 ÿ 0:24‰POCŠ; r2 ˆ 0:70 …5† For POC > 10 mg C gÿ1 sediment; log SS BAFMMHg …solid† ˆ 1:06 ÿ 0:039‰POCŠ; r2 ˆ 0:50 …6†

These equations [Eq. (3), (5) and (6)] were also used to derive a bioaccumulation factor for food (FBAF), derived from the data in Fig. 4 and in Table 3, assuming algae as 90% water. From the data in Fig. 4, ng gÿ1 assimilated from food (Ampfood) was estimated as: Ampfood ˆ SBAF  …ng gÿ1 sed …spiked algae†† ÿ SBAF  …ng gÿ1 sed …no algae††

…7†

Table 3 Model parameter equations for the modeling of the ®eld data from Lavaca Baya For organic carbon, a distribution coecient (KD) of 125 l kgÿ1 is assumed (Thomann et al., 1992; Bloom et al., 1999); i.e. DOC (mg C lÿ1) in porewater=[mg C gÿ1 sediment]8 A base level of 4 mg lÿ1 DOC was assumed in the water with the following relationship between porewater and overlying water concentrations, again based on ®eld data: DOC (mg C lÿ1) in overlying water=4+(0.2[mg C gÿ1 sediment]). For MMHg, the KD was made a function of sediment OM, based on Bloom et al. (1999): Log KD for MMHg=2+0.03[mg C gÿ1 dry sediment] and water, MMHg concentration was related to sediment in the following fashion: Dissolved ng lÿ1 MMHg in porewater=([ng MMHg gÿ1 wet sediment]1000R)/KD where R is the sediment dry/wet ratio; and Dissolved ng lÿ1 water=0.02+0.01[ng MMHg lÿ1 porewater]. a

DOC, dissolved organic carbon; MMHg, monomethyl mercury; OM, organic matter.

225

Then, FBAF ˆ Ampfood =ng gÿ1 algae

…8†

A value for FBAF (wet wt.) of 70 [(FBAF(dry wt.)=7.8] was derived. 4. Discussion Preliminary sediment exposure experiments, where amphipod concentrations were measured after 3 and 6 days, suggested that there was no signi®cant di€erence in amphipod concentration after 3 days. Theoretical considerations, however, suggest that the amphipod concentrations should take longer to reach steady state. The steady state concentration, CSS, is given by: CSS ˆ Cex …k1 =k2 †…1 ÿ exp…ÿk2 t††; where k1 is the combined accumulation rate from all phases, Cex is the e€ective exposure concentration and k2 is the overall rate of loss. Thus, the time to steady state is essentially controlled by the rate of loss (growth dilution plus elimination/depuration); i.e. t / 1/k2). The growth rate of L. plumulosus has been estimated as 0.04±0.1 dayÿ1 (Emery et al., 1997; Merten, 1999) and thus, steady state due to growth dilution during uptake would take 50 days or more. Depuration rates of MMHg from bivalves and ®sh are small compared to growth while depuration of HgI is more signi®cant (e.g. Cunningham and Tripp, 1975; Boudou and Ribeyre, 1985). A similar case is likely for amphipods. Studies of the uptake/depuration of Cd by Hyalella showed that steady state was attained in about 2 weeks (Stephenson and Turner, 1993) with a loss rate of 0.36 dayÿ1. Based on these data, we assume that the rate of depuration of MMHg by amphipods will be small compared to growth. For HgI, depuration is likely to be important compared to growth dilution. Moulting and reproduction would enhance the overall measured loss rates from amphipods but, as our previous experiments with copepods (Lawson and Mason, 1998) have shown, essentially all the MMHg was associated with soft tissue, so moulting is an unimportant loss mechanism for MMHg; it could be for HgI (15% of the total Hg was found in carapace of copepods exposed to Hg). Given the animal sizes used in the experiments, production of o€spring is likely insigni®cant (Merten, 1999). Overall, therefore, the rate of loss of MMHg via depuration or other loss mechanisms is likely small or similar to growth in these experiments. For a growth rate of 0.1 dayÿ1, the amphipods would have accumulated about 45% of their steady state concentration in 6 days; i.e. the measured concentrations are likely within a factor of 2±3 of the steady state values, given the rate of the other loss mechanisms besides growth dilution. The results from the water experiment indicate that a quasi steady state had been reached during the

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experimental period (Fig. 5c). As uptake of Hg/MMHg from water likely occurs across the body wall and gills, the accumulated burden is partitioned into the organism di€erently than it is from uptake via food, and higher depuration rates have been measured for ®sh after water exposure compared to exposure through food (Ribeyre and Boudou, 1989). Overall, given the above considerations, the di€erent exposure times of the experiments should not negate comparison of uptake between experiments. The duration of each experiment was chosen to minimize artifacts. Longer experiments without feeding or water changes are problematic because of the potential for starvation (e.g. Emery et al., 1997; Merten, 1999) and water quality deterioration, and because growth and reproduction are confounding factors in assessing bioaccumulation. Also, if HgI and MMHg release from sediments into water occurs, the need for water changes results in less control over the actual amount of exposure due to Hg in the overlying water during the sediment exposures. With the additional impact of feeding, discussed below, the overall potential artifacts that exist with exposures of sucient duration to allow for steady state accumulation did not warrant this approach. Thus, as steady state was not likely attained in the sediment exposure experiments, the results indicate the relative accumulation from sediment. Our results suggest that sediment OM is an important factor in determining the sorption properties of sediment for HgI and MMHg and appears to be, for the low AVS sediments studied here, the most important factor controlling the amount of HgI and MMHg bound to the sediment, as found in ®eld studies (Langston, 1986; Muhaya et al., 1997; Bloom et al., 1999; Mason and Lawrence, 1999). Other factors besides OM content may also be a€ecting bioaccumulation, such as the behavior and feeding of the amphipod. Emery et al. (1997) showed that grain size is more important than sediment composition (i.e percent clay or sand) in determining amphipod growth rate and percent survival, and thus growth should not have been a€ected by the treatments used in this study. Although growth rates and lipid content of the amphipods were not measured in these experiments, previous experiments with two di€erent species of amphipods of similar size class did not show signi®cant changes in whole body lipid values during 8- and 10-day exposures (Meador et al., 1997). It was, therefore, concluded that variations in sediment characteristics did not a€ect amphipod growth or survival. Results from the sediment exposure concur with our ®eld results from Baltimore Harbor and around HartMiller Island, MD (Mason and Lawrence, 1999), and those of others (Langston, 1986; Muhaya et al., 1997; Nuutinen and Kukkonen, 1998; Bloom et al., 1999), which show a signi®cant positive correlation between sediment HgI and MMHg concentration and sediment OM. For ®eld sites, the correlation is not as strong as seen in the

laboratory exposures because of di€erences in sediment characteristics (e.g. grain size, sand/clay content) and due to the in¯uence of point source inputs. The correlation between HgI and MMHg and OM for Baltimore Harbor and at Hart-Miller Island is, however, stronger than that of other factors (e.g. Fe, AVS, %sand; Mason and Lawrence, 1999), again suggesting that organic content to a large extent controls Hg and MMHg concentration and subsequent bioaccumulation from sediment. Further evidence to support the role of organic carbon in controlling Hg, and especially MMHg, bioaccumulation comes from digestive ¯uid solubilization studies with the intestinal ¯uid of benthic invertebrates (Chen and Mayer, 1998; Lawrence et al., 1999; McAloon et al., 1999). In these experiments, the sediment is incubated in vitro with the intestinal ¯uid of an invertebrate and the fraction of the Hg or MMHg release to solution measured. It has been determined for a number of contaminants that this technique provides a representative measure of bioaccumulation (Weston and Mayer, 1998) and our preliminary results suggest this is also true for Hg and MMHg (McAloon et al., 1999). Our solubilization studies show a strong inverse correlation between the amount of MMHg, and to a lesser extent Hg, released from sediment and the organic content of the sediment (Lawrence et al., 1999). Indeed, the shape of the resultant curve bears a striking similarity to that of the amphipod bioaccumulation curve. These results suggest that OM is binding Hg and MMHg within the sediment and rendering it unavailable for solubilization within the intestinal tract, and thus for bioaccumulation, given the accepted paradigm of digestion that only soluble compounds are absorbed across the gut lining. Overall, therefore, there are multiple lines of evidence to support the contention that OM in sediment plays an important role in controlling Hg and MMHg bioavailability to benthic organisms. The relative magnitude of the SBAFHgI to that of SBAFMMHg (Fig. 2) is in agreement with our understanding of the relative strength of the complexes formed between HgI and MMHg and natural OM (Hudson et al., 1994; Hintelman et al., 1995; Mason et al., 1996). Comparison of the relative magnitude of the equilibrium constants for HgI and MMHg binding to OM (respectively, 1019 and 1014.6 l kgÿ1 C for Aldrich humic acid; Lawson and Mason, 1998) with those for binding to hydroxide (a surrogate measure of anity for oxides phases in sediments; Dzomback and Morel, 1990), of 1010.1 and 109.4 Mÿ1, respectively, suggests that HgI will be more strongly bound to particulate OM than MMHg, even taking into account any uncertainties associated with the values for the thermodynamic equilibrium constants. At low OM, we suggest that as both HgI and MMHg are largely bound to inorganic complexes and phases, SBAFs are high, indicating high bioavailability of both

A.L. Lawrence, R.P. Mason / Environmental Pollution 111 (2001) 217±231

forms of Hg (Fig. 2). These results agree, for example, with studies of HgI and MMHg bioaccumulation by algae (Mason et al., 1996) which showed that HgCl2 and CH3HgCl had similar accumulation rates and octanol± water partition coecients. In contrast, at high organic content, both SBAFHgI and SBAFMMHg are small, indicating that under these conditions both forms of Hg are tightly bound and relatively unavailable for assimilation. Our solubilization studies (Lawrence et al., 1999; McAloon et al., 1999) support this contention as they suggest that it is competitive ligand interactions (the strength of binding to the sediment) that control the degree of solubilization within the intestinal ¯uid, as Chen and Mayer (1998) have found for copper. These results provide a reasonable explanation for both the laboratory results and ®eld studies and suggest that, over the range of OM found in the environment, bioaccumulation of HgI should be hindered to a larger degree than MMHg. Field-determined SBAFs for amphipods, along with other benthic organisms, from Baltimore Harbor (Mason and Lawrence, 1999) and from Lavaca Bay, TX (Mason et al., 1998) also displayed a trend of decreasing SBAF with increasing sediment OM, and were of the same magnitude as those determined in this experiment (Fig. 7; Lavaca Bay data).

227

The magnitude of the di€erence in the SBAFMMHg compared to SBAFHgI at intermediate OM in our experiments was similar to that found in the ®eld. Given that MMHg is typically <1% of total Hg in estuarine sediments (e.g. Benoit et al., 1998; Mason and Lawrence, 1999) and that in the experimental system SBAFMMHg is about 10 times that of SBAFHgI over intermediate sediment OM (1.5±7%), it should be expected that MMHg would account for approximately 10% of the total Hg in amphipods in the ®eld. Field data from Lavaca Bay and Baltimore Harbor, however, show much higher %MMHg in amphipods (Mason et al., 1998; Mason and Lawrence, 1999). The di€erences in%MMHg between laboratory exposures and the ®eld data may be attributed to: (1) an alternative food source besides sediment present in the ®eld (i.e. algal inputs); (2) di€ering depuration rates for HgI and MMHg; or (3) the lack of attainment of steady state accumulation in the laboratory exposures. The importance of the various depuration rates and the lack of steady state in the laboratory exposures have been discussed above. Both factors would tend to enhance the relative concentration of MMHg and could account for the di€erence in %MMHg in laboratory versus ®eld amphipods. However, the feeding exposure

Fig. 7. Comparison of various benthic organism's sediment bioaccumulation factors with sediment organic content in Lavaca Bay, TX.

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results suggest that `food' is potentially an important contamination route in both the laboratory exposures and in the ®eld, i.e. newly deposited OM in the ®eld. Comparing SBAFs from the sediment/algae experiment (Fig. 3) to those of the sediment-only exposure at similar sedimentary OM (Fig. 2) illustrates that the addition of spiked algae increased both SBAFHgI and SBAFMMHg (Fig. 4). Amphipods living in an environment with spiked sediment and spiked algae accumulated relatively more HgI and MMHg than those living in spiked sediment with no algae present. When unspiked algae was added to the spiked sediment microcosms, both SBAFHgI and SBAFMMHg decreased below levels found in the treatments which had the addition of spiked algae. The modeling estimations give a FBAF of 70 which is greater than that for accumulation from sediment alone (i.e. the SBAF), except for sediment of very low OM (Fig. 2). Thus, the changes in the presence of algae could be accounted for by selective feeding on algae rather than sediment particles, combined with the probable greater bioavailability of both HgI and MMHg from algae, although potentially confounding factors cannot be resolved. For example, it is not known if changes in accumulation due to algae addition are a result of DOC changes in the water column, although the calculations given above strongly suggest that this is not the case. However, in corroboration with our experimental studies, a comparison of toxicity tests with amphipods exposed to the same sediment, but under di€erent feeding regimes (Fisher, personal communication) noted that amphipod survival was enhanced when L. plumulosus was fed algae (a preferred food) compared to Tetramin1 (ground ®sh food). These results, and our data, point towards the importance of food in mediating contaminant bioaccumulation from sediments by amphipods. The water column experiment illustrates the e€ect of organic matter on HgI and MMHg bioaccumulation from water (Fig. 5b). However, the e€ect is small, suggesting that the added DOC is not changing the water column speciation to a dramatic degree at the OM ranges examined (3.2±7.1 mg lÿ1), but that the HgI and MMHg is becoming less available, as expected from an thermodynamic equilibrium standpoint, and from our bioaccumulation studies with phytoplankton (Lawson and Mason, 1998). The experimental treatments in the water experiment were an ambient natural DOC exposure (4.7 mg lÿ1), plus two higher DOC exposures attained by the addition of Aldrich humic acid to the natural seawater (Fig. 5a). DOC levels were those as measured at the end of the experiment. The results obtained suggest that the type of DOC does not signi®cantly in¯uence the Hg/MMHg binding capacity or the relative bioaccumulation (Fig. 5). The magnitude of the WBAFMMHg is about 50 times that of WBAFHgI across the DOC gradient used in this exposure. In the water column, the relative binding is

driven by the ratio of the organic complexation constant to that of chloride, the dominant inorganic ligand in 10 ppt seawater. Chloride complexes of HgI are relatively much stronger than for MMHg due to the formation of higher chloride complexes (Thomann et al., 1992; Mason et al., 1996; Stumm and Morgan, 1996). The overall binding capacity (Ki[Cl]i; Ki=individual equilibrium constants) is about 1012.8 for inorganic HgCli compared to 104.4 for MMHgCl (K[Cl]). However, OM should still complex most (>90%) of the MMHg and HgI (>80%) in the water column experiments at the organic levels used. The MMHg model was extended to include uptake from algae to further investigate the importance of algal/fresh OM inputs in amphipod bioaccumulation. The data in Table 2 were used as input to the model. The bioaccumulation factors were those derived previously, using the appropriate values for POC and DOC and MMHg concentration (Table 2).The comparison of the model predictions for each uptake route for the various laboratory exposure conditions discussed above (the combination of spiked and unspiked sediment and algae; Fig. 3) is shown in Fig. 8a in terms of amphipod concentration. In the model calculations, the percentage of solid material ingested that was sediment (y), or algae (1ÿy), was not known so a variety of values were used

Fig. 8. (a) Comparison of the model methylmercury results with results from the feeding exposure in terms of amphipod concentration. (b) Model estimate of the importance of the various sources in uptake of methylmercury.

A.L. Lawrence, R.P. Mason / Environmental Pollution 111 (2001) 217±231

in a sensitivity analysis. The value of y was manipulated in an attempt to generate results that best ®tted the laboratory data, and the value with the best ®t of the model data (y=0.4) was used in generating Fig. 8a. It should be noted that increasing the value of y above 0.5 does not provide a reasonable representation of the laboratory data. Thus, this modeling exercise suggests that, to achieve the concentrations measured in the laboratory experiments, amphipods must have been consuming algae in preference to sediment solids. There is good agreement overall between the model results and the experimental data over the various treatments (Fig. 8). The largest discrepancy in terms of concentration appears for the SS/SF treatment. The reason for the synergistic increase in accumulation in the laboratory SS/SF exposure is not known but for the spiked food exposures, higher accumulation would be predicted by the model under the scenario of a higher proportion of the ingested solid as algae (i.e. y<0.4). The relative importance of each pathway was also estimated (Fig. 8b). The model results con®rm the laboratory data and suggest that ingestion of fresh OM (algae) is the dominant source of MMHg to the amphipods in the experiments where spiked food was added. In the presence of unspiked food, uptake from water and porewater increased. Thus, direct uptake from water cannot be ignored as it can be important under speci®c conditions. In the ®eld, it is uncertain whether accumulation from ingestion of `fresh' OM is more important than accumulation from sediment. The data from Lavaca Bay (Fig. 6) were used to investigate this possibility. Recent studies in Lavaca Bay have developed relationships between sediment and porewater concentration, and sediment±overlying water relationships for this shallow, wellmixed estuary (Mason et al., 1998; Bloom et al., 1999). The water concentrations were modeled using these relationships (Table 3) as water/porewater data were not collected concurrently with the biota sampling. The KD values chosen for carbon partitioning (Table 3) are similar to values reported in the literature (Thomann et al., 1992) and the KD values for MMHg sediment partitioning and water column partitioning are similar to those derived for other estuarine systems (e.g. Benoit et al., 1998; Mason et al., 1999). The empirical relationship of water column concentration against sediment concentration for MMHg (ng lÿ1 water=a+b [ng MMHg gÿ1 sediment]) was derived from the results of ®eld studies in Lavaca Bay (Bloom et al., 1999). The model developed here, which strongly links water column concentrations to sediment concentrations, is valid for Lavaca Bay, and especially for the grass/marsh sites where the amphipods were collected (Bloom et al., 1999). As algal MMHg concentrations were not measured, they were estimated using a relationship between dissolved water column concentration and algal concentration, essentially a KD of 5104 (i.e. ng gÿ1 algae

229

(dry wt.)=50[ng lÿ1 water]), derived from the data of Mason et al. (1996). This formulation yielded algal concentrations that were up to three times higher than the sediment concentration, on a per gram dry weight basis. Again, this relative magnitude of concentration mimics the ®eld situation (Mason et al., 1998; Bloom et al., 1999). For this simulation, the FBAF derived from the laboratory exposures was used. This value corresponds relatively closely to the range in values measured in the ®eld for trophic transfer factors between trophic levels (Lindqvist et al., 1991; Watras and Bloom, 1992) and for laboratory-derived values from feeding experiments (Mason et al., 1996; Lawson and Mason, 1998). As the model application to Lavaca Bay was used primarily to assess the relative importance of fresh OM as a food source, the model was manipulated, in terms of the relative food source, until a reasonable agreement between model amphipod concentrations and measured ®eld concentrations was obtained (Fig. 9a). To obtain values similar to the ®eld data it was necessary to include a substantial amount of fresh algal input in the diet. Clearly, the relative importance of each source is

Fig. 9. Model results for Lavaca Bay, TX. (a) Comparison of model and ®eld data for methylmercury. (b) Model estimation of the importance of sources of methylmercury to amphipods.

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indicative of what is occurring in the ®eld but the absolute number for the model run plotted in Fig. 9 (50% of solid intake from sediment) should not be over interpreted. The model results suggest that the amphipods are obtaining their MMHg both from in-place sediments and from fresh algal matter in the ®eld (Fig. 9b). At the lower sediment POC concentrations, the role of sediment and porewater as sources of MMHg increases and thus, for sediment of less than 2±3 g C/kg sediment, sediment ingestion appears to be the major vector for MMHg accumulation. These results are in agreement with the laboratory data. The notion that amphipods may preferentially feed on new organic inputs rather than in-place sediments has implications for the outcome of standard amphipod toxicity tests, when combined with the fact that uptake from porewater is small for MMHg and Hg; and likely so for other strongly-bound contaminants. Thus, there is the potential that toxicity may be alleviated by addition of low MMHg concentration food, as suggested by the decreased SBAF in the SS/UF exposures, and by the impact of food type on the outcome of amphipod toxicity tests (Fisher, personal communication). While the di€erences in toxicity could be explained by other factors, the results are suggestive of selective feeding of the amphipods on the algae, which leads to an alleviation of the overall toxicity of their environment. Overall, the results of the laboratory exposures described here, coupled with the model results, suggest that bioaccumulation by amphipods and other organisms inhabiting the sediment±water interface is a complex interaction where the importance of fresh OM input cannot be ignored as a source of contaminants to benthic organisms. Clearly, further determinations of Hg bioaccumulation under other environmental conditions, especially those which would greatly a€ect Hg speciation, are necessary. For example, investigation at high levels of sedimentary AVS would enhance our understanding of Hg bioaccumulation from sediment to benthic organisms. In shallow, productive estuarine systems, the incorporation of seasonal cycles, including algal blooms, is also necessary to accurately assess the impact of most particle-associated contaminants. For example, a bloom of relatively uncontaminated algae serves as a pulse of fresh food from the surface, which may be a dominant food source for benthic organisms living in contaminated sediment. Therefore, these organisms would have lower body concentrations of the contaminant than predicted from sediment concentration. This benthic±pelagic interaction might be an important exposure route for benthic organisms, especially in more oligotrophic, deeper environments. For example, in Lake Michigan, deposited algae from a spring diatom bloom can account for 30% of the energy requirements of Diporeia species, the dominant amphipod (Fitzgerald et al., 1993). Thus, in this instance,

pelagic sources would have a large impact on contaminant concentrations found in benthic organisms. Results from this study suggest that spiked algae was the most important route of HgI and MMHg uptake by amphipods in these studies. Overall, MMHg was more available for bioaccumulation than HgI. With regard to management, these ®ndings illustrate that decisions based solely on the concentration of Hg in sediment, excluding sediment characteristics and the concentration of Hg in organisms at the base of the food chain, will be inaccurate. Further studies should continue to examine the importance of water column and sediment speciation of Hg and MMHg in controlling bioaccumulation at the sediment±water interface. Acknowledgements This research was partially funded by grants from the National Oceanic and Atmospheric Administration and the Environmental Protection Agency through the Multiscale Ecosystem Environmental Research Center. We thank the Maryland Department of Natural Resources, for amphipods, Analytical Services for DOC analysis and Tyler Bell, Cindy Gilmour, and the laboratory of J. Cornwell for their aid in AVS analysis. We also wish to thank Jani Benoit, Nicole Lawson, Kelly McAloon, Beth McGee, Amy Merten, and Jean-Michael Laporte for their help and support of this work. This is manuscript No. 3273 of the Center for Environmental Science of the University of Maryland. References ASTM, 1990. Standard guide for conducting 10-day static toxicity tests with marine and estuarine amphipods (E1367-90). In: Annual Book of ASTM Standards. American Society for Testing Materials, Philadelphia, PA, pp. 1±24. Benoit, J.M., Gilmour, C.C., Mason, R.P., Riedel, G.S., Riedel, G.F., 1998. The sources and cycling of mercury in the Patuxent estuary. Biogeochemistry 40, 249±265. Bloom, N.S., 1989. Determination of picogram levels of methylmercury by aqueous phase ethylation followed by cryogenic gas chromatography with cold vapor atomic ¯uorescence detection. Can. J. Fish. Aquat. Sci. 46, 1131±1140. Bloom, N.S., Crecelius, E.A., 1987. Distribution of silver, mercury, lead, copper and cadmium in central Puget Sound sediments. Mar. Chem. 21, 347±377. Bloom, N.S., Gill, G.A., Cappellino, S., Dobbs, S., McShea, L., Driscoll, C. et al., 1999. Speciation and cycling of mercury in Lavaca Bay, Texas, sediments. Environ. Sci. Technol. 33, 7±13. Bloom, N.S., Fitzgerald, W.F., 1988. Determination of volatile mercury species at the picogram level by low temperature gas chromatography with cold vapor atomic ¯uorescence detection. Anal. Chim. Acta 208, 151±161. Boudou, A., Ribeyre, F., 1985. Experimental study of trophic contamination of Salmo gairdneri by two mercury compounds Ð HgCl2 and CH3HgCl. Analysis at the organism and organ levels. Water Air Soil Poll. 26, 137±148.

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