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Uptake of 109Cd from sediments by the bivalves Macoma balthica and Protothaca staminea
Aquatic Toxicology 47 (2000) 147 – 159 www.elsevier.com/locate/aquatox
Uptake of
109
Cd from sediments by the bivalves Macoma balthica and Protothaca staminea J.R. Pierre Stecko, L.I. Bendell-Young *
Dept. of Biological Sciences, Simon Fraser Uni6ersity, 8888 Uni6ersity A6e., Vancou6er, V5A 1S6, Canada Received 23 April 1998; received in revised form 8 March 1999; accepted 30 March 1999
Abstract This study (1) compared the relative importance of suspended particulate matter (SPM) and deposited sediment (DS) as a source of cadmium exposure to two sediment ingesting bivalves, and (2) determined the importance of feeding behaviour on cadmium uptake from SPM and DS by comparing metal uptake from ingested sediments by a facultative and an obligate filter-feeder. Two types of sediment, deposited (DS) and suspended (SPM) sediments were sampled from the Fraser River Estuary and geochemically characterized with respect to amounts of easily reducible manganese and reducible iron (manganese and iron oxides respectively), organic matter, and amounts of cadmium associated with each of the three components. Sediments were radiolabeled with 109Cd, with the labelled sediments then fed to Macoma balthica and Protothaca staminea, a facultative filter-feeder and an obligate filter-feeder respectively. Amounts of radiotracer accumulated by the bivalves from the two types of sediment over an 8-day period were then assessed. The geochemistry of DS was distinct from that of SPM with DS containing almost half the concentrations of oxides of manganese and iron as compared to SPM. Natural cadmium was recovered primarily from the easily reducible (associated with oxides of manganese) followed by the reducible (associated with oxides of iron) sediment component, for both SPM and DS. In contrast, 109Cd was recovered from the reducible fraction, followed by the easily reducible fraction from the two types of sediment. No natural cadmium or 109Cd was recovered from the organic component of either sediment. For both bivalves, uptake of 109Cd from sediment as compared to water accounted for ca. 80% of accumulated radiotracer. Uptake of 109Cd from DS was significantly greater than from SPM, most notably for P. staminea. For M. balthica, the amount of accumulated 109Cd was best explained by the degree of isotope desorption from DS. In contrast, 109Cd accumulation by P. staminea was best described by its feeding behaviour. Metal accumulation by sediment-ingesting organisms is not simply related to sediment metal concentrations but will also depend on the strength of association of the metal within the sediment (for facultative feeders), and the feeding behaviour of the bivalve. 109Cd partitioning in DS and SPM did not parallel the field partitioning of cadmium which precluded our objective of assessing the relative importance of SPM and DS as a source of cadmium to facultative filter-feeders. To elucidate the role of complex sediment geochemistry in influencing
* Corresponding author. Tel.: + 1-604-291-5621; fax: + 1-604-291-3496. E-mail address: [email protected] (L.I. Bendell-Young) 0166-445X/00/$ - see front matter © 2000 Elsevier Science B.V. All rights reserved. PII: S 0 1 6 6 - 4 4 5 X ( 9 9 ) 0 0 0 2 3 - 5
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metal availability to sediment ingesting organisms, further study is needed on ways to either label or spike natural complex sediments such that the speciation achieved under laboratory conditions is that which is observed in the field. © 2000 Elsevier Science B.V. All rights reserved. Keywords: Sediment geochemistry; Bivalve; Filter-feeding behaviour; Metal uptake
1. Introduction The behaviour of metals in aquatic systems has been of increased interest due to the global anthropogenic alteration of trace metal cycles (Forstner and Wittman, 1981; Salomons and Forstner, 1984). Anthropogenic metals introduced into the aquatic environment are either in particulate form or are rapidly sorbed to particles (Jenne and Zachara, 1987; Regnier and Wollast, 1993). Once sorbed, they contribute to the pool of metals associated with suspended particulate matter (SPM) within the water column or are ultimately deposited to the sediments as deposited sediments (DS). For organisms that ingest either DS and/or SPM as a food source, these anthropogenic metals have the potential for causing severe adverse effects on the organism plus potentially being bioaccumulated by the organism possibly leading to the food-chain transfer of the trace metal to higher trophic levels. The availability of anthropogenic metals to estuarine organisms from the sediment is dependent in part on the presence of the trace metal (exposure), plus the geochemical phase with which the trace metal is associated (Tessier et al., 1984; Bendell-Young and Harvey, 1991; Rule and Allen, 1996). Several studies have suggested that the easily reducible metal is most available to organisms. For example, Rule and Allen (1996) and Thomas and Bendell-Young (1998) under laboratory and field conditions respectively, found that metal associated with the easily reducible fraction of sediment (i.e. associated with oxides of manganese) was the best predictor of tissue cadmium concentrations in the sediment ingesting Macoma balthica. Recently, Stecko and Bendell-Young (1999) contrasted the geochemistry of SPM versus DS in the Fraser River Estuary and found that the geochemistry of SPM was distinct from that of
DS; SPM contained greater amounts of organic matter, reducible iron content (iron oxide) and total metal concentrations (zinc, copper, cadmium and lead) with a greater proportion of the metal present in an easily reducible form compared to DS. These differences in the geochemistry and potential metal bioavailability between SPM and DS are important in regards to sediment ingesting organisms that feed on either SPM or DS, or organisms that are capable of exploiting both types of sediments as potential food sources. For example, the suspension-feeding bivalves Mya arenaria and M. balthica are facultative filter-feeders, capable of feeding on deposited as well as suspended sediment with the ‘choice’ of sediment being dependent on the quantity of particles (Olafsson, 1986). In contrast, Protothaca staminea and Cerastoderma edule are obligate filter-feeders filtering SPM from the water column relying on inhalant water currents to take up food (Crecelius et al., 1982). If metal bioavailability is dependent on sediment geochemistry as found for example, by Rule and Allen (1996) and Thomas and Bendell-Young (1998), and if the organism has the option of switching feeding modes depending on particle flux, then there is the potential for metal exposure and subsequent availability to sediment ingesting organisms, specifically facultative filter-feeders, to differ between DS and SPM with differences being dependent on feeding behaviour. In contrast, for obligate filter-feeders, exposure should occur only through one sediment type. Hence, the objectives of this study were twofold: 1. to compare the relative importance of SPM and DS as a source of trace metal exposure to two sediment ingesting bivalves and 2. to determine the importance of feeding behaviour on metal uptake from SPM and DS by comparing metal uptake by a facultative and an obligate filter-feeder.
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To meet these objectives, two types of natural sediment, collected from the field, both DS and SPM, were radiolabeled with 109Cd. Labelled sediments were then fed to M. balthica and P. staminea, a facultative filter-feeder and an obligate filter-feeder respectively. Amounts of radiotracer accumulated over an 8-day period were then assessed. Ultimately, we hope that the results of our study will lead to a greater predictive capacity for determining which sediment type (in this case, SPM versus DS) and the feeding behaviour (facultative or obligate filter-feeder) that results in the greatest metal exposure to sediment ingesting organisms via their diet. 2. Materials and methods
2.1. Field collection of sediments and bi6al6es Deposited sediment (DS) and suspended particulate matter (SPM) were collected from a mid-estuarine site of the Fraser River Estuary, B.C., Canada. Because the B63-mm fraction of suspended sediment is not depth-dependent, suspended sediment recovery for geochemical purposes can be accomplished by sampling close to the water surface (Ongley et al., 1982). To collect the evenly distributed B63 mm fraction of SPM, near surface water (4000 l total) was collected on mudflats at 1.5 m of water depth. Visible resuspensions were carefully avoided. For DS, several sediment cores (using a Wildco® hand corer) were taken at the sample site; the oxic (surficial) portion of these cores (above the visible redox colour change) were put into acid-washed glass jars for transport to the laboratory. The oxic portion generally varied from 2 to 8 cm depth. Organisms were collected from the intertidal region of the Fraser River Estuary. P. staminea, the common Pacific littleneck clam, was collected by raking the mudflats and obtaining organisms ranging in size from 3.5 to 5 cm (approximately 4 – 6 years old). M. balthica, the Baltic tellin, was collected by gently hand raking the sediment, feeling for the small (8 – 14 mm) clams. Organisms were stored in buckets containing sea-water and were transported to the lab where they were acclimatized and depurated in a filtered sea-water and
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dechlorinated water mixture for a minimum of 72 h before the initiation of the uptake experiments. 2.2. Sediment geochemistry Sediment samples were immediately transported to the laboratory where they were kept refrigerated at 5°C until processing (always within 72 h). SPM was separated from the water by flow through centrifugation at 43 807.5× g using the Sorvall® TZ-28 system in GK continuous flow mode in a Sorvall® RC5C Centrifuge with a flow rate of between 0.6 and 0.75 l/min. Sediment samples were simultaneously extracted after methods of Bendell-Young et al. (1992) which separates sediment into operationally defined components including: (1) easily reducible manganese and metal associated with manganese oxides, (2) reducible iron and metal associated with iron oxides, and, (3) organically bound metal (Fig. 1). Organic content was determined by loss on ignition (%LOI) at 600°C for 1 h in a blast furnace. All sediment samples were treated with a 15 to 1 ratio of analytical-grade extractant (Baker Instra Analyzed or BDH Analar) to sediment. Supernatants were removed from the extraction tubes through Millipore Millex-HV sterilizing filter units (0.45 mm) using acid washed 10-ml syringes. Metal analyses for manganese and iron were carried out by flame atomic absorption spectrophotometry (FAAS) on a Perkin-Elmer® 1100B FAAS. Analyses for cadmium was done by graphite furnace atomic absorption (GFAA), at the University of British Columbia, Department of Oceanography. Quality assurance/quality control (QA/QC) was assured through inclusion of NRC reference standards (MESS-2) and reagent blanks. Reference standards fell within 10% of certified values.
2.3. Isotope partitioning To determine how the isotope distributed within the sediments, in triplicate, 1.000 g of wet DS and 1.605 g of wet SPM to achieve an equivalent dry weight of 0.6 g (where the dry/wet ratios for each sediment are 0.61 and 0.38 for DS and SPM respectively) were added to centrifuge tubes,
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with 25 ml of water (a mixture of sea-water and dechlorinated distilled deionised water to achieve a salinity of 14 ppt), and spiked with 27.8 kBq of 109 Cd (\99% purity, t1 453 days) giving a final 2 concentration of 1112 Bq/ml (12.15 ppb) 109Cd. Slurries were mixed several times daily and sampled at 1, 3 and 7 weeks. At the end of the interval, the slurries were centrifuged, extracted by methods of Bendell-Young et al. (1992), and the extracts counted for radioactivity on instrumentation as outlined below.
2.4. Exposures The experimental design used three replicate 38 l aquaria with 36 P. staminea and 20 M. balthica per aquaria. Controls were run in the exposure aquaria using 0.45-mm Millipore filters as dividers (Harvey and Luoma, 1985a,b). To ensure that the control side was in equilibrium throughout the exposure experiments, filtered (0.45-mm Millipore filter) water samples from the control side of the aquaria were analysed for dissolved 109Cd to com-
pare with dissolved 109Cd (0.45-mm Millipore filter) in the exposure side of the aquaria. Equilibrium of dissolved 109Cd between the control and exposure sides of the tank was achieved within 48–96 h of addition of the radiolabelled sediments to the exposure side of the tank (Fig. 2). This experimental design therefore allowed half the organisms to be exposed to the particulate plus the dissolved radioisotope, whereas the other half was exposed only to the free (dissolved) radioisotope. Salinity was held at levels approximating levels found at the collection site ( 14 ppt) using an appropriate mixture of filtered seawater (1 mm) and dechlorinated distilled deionized freshwater. Temperature was set at 12°C. To ensure equal exposures of the bivalves to the two sediment types, radiolabeling of the sediments for the exposures with 109Cd was done on an equivalent dry weight of 5.00 g of sediment; 8.19 g of wet DS and 13.16 g of wet SPM, at dry/wet ratios of 0.61 and 0.38 respectively. To each of three high density polyethylene (HDPE) bottles were added the sediment and 205 ml of sea-water
Fig. 1. Extraction scheme of Bendell-Young et al. (1992) to determine the geochemical characteristics of DS and SPM.
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Fig. 2. Desorption of 109Cd from SPM and DS as a function of time. Values are means of three measurements9 1 S.E. (a) Dissolved 109Cd in the exposure versus the control tank for the DS exposures; (b) dissolved 109Cd in the exposure versus the control tank for the SPM exposures. Note that equilibrium between the exposure and exposure tanks is reached at approximately 96 h for DS and SPM. E, exposure side of tank; C, control side of tank.
(14 ppt) taken from the sampling site. To these bottles, 2268 kBq of 109Cd was added yielding an
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interim slurry of 10.84 kBq/ml 109Cd. Sediments were shaken every day for 12 days to allow for desorption of most of the loosely bound 109Cd prior to being added to the exposure aquaria. In the exposure aquaria, the total isotope concentration was diluted to 222 Bq/ml (equivalent concentrations of 2.43 ppb) 109Cd. Eighteen P. staminea and 10 M. balthica were used on each side of the aquarium. The majority of added DS and SPM was kept suspended by the turbulence generated by oxygen stones, although the heavier fraction of DS tended to settle out onto the bottom of the aquaria. This settling of DS but not SPM had important implications in determining final tissue radiotracer levels in the bivalves. Bivalves were added to the aquaria 24 h after the addition of the radiolabeled sediments, as the two sides of the aquaria were coming into equilibrium. Exposure of the bivalves to the radiolabeled sediments occurred for a period of 8 days, followed by 6 days of depuration where bivalves were exposed to non-labelled ‘cold’ sediments. Three organisms were removed from each side of the aquaria, control and exposure, on days 1, 2, 4, and 8. The mussels were depurated for 24 h in uncontaminated sediment before determining isotope activity. After 8 days of exposure, the remaining mussels were transferred to an uncontaminated environment to examine elimination of 109Cd P. staminea were sacrificed, the tissue removed from the shell and rinsed in EDTA (to remove any externally absorbed metal), blotted and weighed before analysis. M. balthica were rinsed with EDTA, blotted and weighed before analysis whole; following analysis they were returned to the appropriate aquaria. Simple uptake calculations involved the determination of the amount of radioisotope in the organisms in Bq/g wet wt., at the various time intervals.
2.5. Mass balance of radiotracer Following completion of the exposure experiments, swipe tests of the aquaria (aquaria walls, Millipore membrane), and sediments and water
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were sampled to allow for a mass balance of the radiotracer added to the aquaria (exclusive of bivalves).
2.6. Instrumentation Cd was detected at 88 keV g emission using an automatic g counting system. The system uses Nucleus Analyser Software, and measures g radiation with a sodium iodide crystal. Canberra preamplifiers and amplifiers supply the signal to the computer. The system was optimized using standards of the isotope with optimal setting at 820 V for the power supply and an amplifier setting of 3 K. Counting efficiency for 109Cd was 4.33%. 109
2.7. Statistical analyses Statistical analyses were carried out using SAS, version 6.08 (SAS, 1990). Comparison of amounts of dissolved radiotracer in the exposure versus the control side of the tank at t \96 h was done using simple Student t-tests; comparison of sediment characteristics between SPM and DS was done using simple Students t-tests; comparison of amount of radiotracer sorbed to the three sediment geochemical components was assessed through a one-way ANOVA. Radiotracer uptake from the two types of sediment for each sampling day by the two bivalves was analysed through two-way ANOVA with bivalve and sediment type as the two co-factors (SAS, 1990). Significance for all tests was accepted at P B0.05.
ever, in both cases, reducible iron and easily reducible manganese concentrations were significantly greater in SPM compared to DS (P B 0.001). Concentrations of natural Cd were sevenfold greater in SPM versus that of DS. Although there were differences in total concentrations, partitioning of Cd among the three sediment components was similar for both sediment types, with the majority of natural speciated Cd being recovered from the easily reducible component of both sediments (Fig. 3). No Cd was recovered from the organic phase of sediment.
3.2. Isotope partitioning and desorption in SPM and DS The partitioning of the radioisotope in SPM and DS following 1, 3 and 7 weeks of equilibration indicated that despite the differences in geochemistry of sediment used in the exposures, 109 Cd partitioning in the two sediments did not differ (Fig. 4; ANOVA; P\0.05). For both sediments, the same amount of radiolabel was recovered from the reducible fraction, followed by the easily reducible fraction of sediment. Only trace amounts of the radiolabel were recovered from Table 1 Geochemical characteristics of deposited sediments (DS) and suspended particulate matter (SPM) sampled from (a) Fraser River Estuary (August 1995) and (b) those subsequently used for the radiolabelling and exposure experimentsa SPM
3. Results
3.1. Sediment geochemistry Geochemical characteristics of SPM and DS sampled from a mid-estuarine site in the Fraser River Estuary and the geochemical characteristics of the sediments used in the isotope partitioning and desorption experiment and the exposures are outlined in Table 1. For sediments used for geochemical characterization, isotope experiments and the exposures, there was no significant difference between the organic content (LOI) of the two sediments (Students t-test; P =0.1432), how-
(a) Fraser Ri6er Estuary N LOI (%) RFe (mg/kg) ERMn (mg/kg) (b) Radiolabelling and exposure experiments N LOI (%) RFe (mg/kg) ERMn (mg/kg)
DS
9 6.62 91.58 9280 9783 293 934
9 4.41 90.35 50609301 184920
5 6.5 91.29 12 320 9478 418 98
3 4.19 90.21 6545 9293 197 935
Values are mean 9S.E.; LOI (%), loss on ignition (organic matter); RFe, reducible iron (oxides of iron); ERMn, easily reducible manganese (oxides of manganese). a
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Although the partitioning of the radiolabel did not differ between the two types of sediment, greater amounts of 109Cd desorbed from DS than from SPM (Students t-test; PB 0.05) (Fig. 2). After allowing for a 96-h equilibration period, average (9 1 S.E.) dissolved activity of 109Cd for the DS exposures was 62.59 0.89 (Bq/ml), whereas for the SPM exposures, average dissolved 109 Cd activity was 39.691.17 (Bq/ml) (Fig. 2). The mass balance of the isotope indicated that only 0.7% of the added radiotracer was associated with the walls of the aquaria, the Millipore membrane and its frame. Approximately 73 and 26% were associated with the sediment and water respectively.
3.3. Exposures Fig. 3. Distribution of natural cadmium among the three sediment components in SPM and DS sampled from the Fraser River Estuary. Values are means of nine measurements9 1 S.E. ER, easily reducible; R, reducible; O, organic component of sediment.
the organic component of sediment. The partitioning of the radiolabel also remained consistent over the 7-week period (Figs. 4 and 5).
Uptake of 109Cd from SPM and DS was measured in M. balthica (Fig. 6a) and P. staminea (Fig. 6b). The concentration of isotope in both the sediment (350 kBq/g or 3.9 mg/g) and water (30 Bq/ml or 0.33 ng/ml) and the exposure conditions of 14 ppt salinity and 12°C were environmentally realistic. Activities reported for P. staminea are for the tissue of depurated organisms, and the activities reported for M. balthica
Fig. 4. Distribution of 109Cd among the three sediment components in SPM and DS at 1, 3 and 7 weeks. Values are means of three measurements91 S.E. ER, easily reducible; R, reducible; O, organic component of sediment.
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Fig. 5. Relative proportion of the radiolabel recovered from the three sediment fractions from DS and SPM at 1, 3 and 7 weeks.
are for whole depurated organisms. Two-way ANOVA indicated that the effect of sediment type was significant (N =48, F = 6.44, P = 0.015) whereas organism type and the interaction were not. For both organisms, uptake from DS was significantly greater than that from SPM, with this difference most notable for P. staminea. For M. balthica, most 109Cd uptake occurred over the first 2 days of both experiments with subsequent days showing either lower net uptake or net loss. Following the removal from the exposure vessel on day 8, 109Cd levels in M. balthica decreased. This depuration is evident in both the particulate and water, and the water only exposures. In the DS exposure, most of the uptake by P. staminea occurred over the first 2 days, with consistently low uptake on subsequent days. Decreased 109Cd levels were observed following depuration. During the SPM exposure, P. staminea showed a consistently low uptake with 109 Cd concentrations increasing slightly during depuration. In the DS exposures, M. balthica took up 86.894.0% (mean 9 S.E.) of its 109Cd from the particulate phase, while P. staminea took up 90.6 92.1% from the particulate phase. In the SPM exposures, M. balthica took up 78.9 93.8% 109 Cd from the particulate phase, while P. staminea took up 80.897.3% from the particulate
phase. Because the quantity of food consumed by these organisms was not measured, no comparisons based upon total uptake can be made, although the amount of radiotracer accumulated by the two bivalves over the 8-day period was similar.
4. Discussion
4.1. Sediment geochemistry of SPM
109
Cd in DS and
Environmental concentrations of cadmium were recovered primarily from the easily reducible followed by the reducible sediment components in both sediment types. No natural cadmium was recovered from the organic phase of sediments. Riise et al. (1994) also noted that 109Cd partitioned onto the easily reducible phase of soil (rather than sediment) supporting our field results. In contrast, 109Cd was found to associate predominantly with the reducible and to a lesser extent, the easily reducible phase with the same amount of radioactivity recovered from each sediment component for both SPM and DS. This partitioning of 109Cd was surprising given the differences in sediment geochemistry between the SPM and DS (i.e. SPM contained greater
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amounts of reducible iron and easily reducible manganese as compared to DS). Based on the behaviour of cadmium in the natural environment, more of the radiolabel should have partitioned onto the easily reducible components of sediment. This lack of a partitioning difference between the two sediments plus the equivalent amount of radioactivity recovered from each sediment component suggests that in-
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troduction of the radiotracer to sediments under laboratory conditions (i.e. addition of a relatively large amount at one time) and the subsequent partitioning of the radiolabel among the sediment components does not allow for the same processes and hence the same partitioning as that observed in nature. Unlike experimental conditions, in nature, low concentrations of the trace metal are added over a long term to an environment where
Fig. 6. 109Cd activity in M. balthica (a) and P. staminea (b) for both SPM and DS exposures at days 1, 2, 4, 8, 10 and 14. Values are means of three measurements 9 1 S.E.
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the geochemical composition of the suspended and deposited sediment are undergoing constant change. In the natural environment, both iron and manganese undergo a continuous redox cycle which leads to freshly precipitated oxides of manganese and iron, either within the oxic water column or at the sediment/water interface. The oxidation of Mn occurs primarily on the surfaces of particles via oxidizing bacteria (Santschi et al., 1990), whereas the oxidation of Fe is primarily accomplished without the mediation of oxidizing bacteria (Egeberg et al., 1988; Santschi et al., 1990). These freshly precipitated oxides present in SPM and DS will co-precipitate trace metals such as cadmium. What we recover from the natural sediment sampled from the field are these freshly precipitated oxides of Mn and Fe (as easily reducible Mn and Fe, i.e. oxides of Mn and the amorphous oxides of Fe), and cadmium which has been scavenged by these oxides. In contrast, under laboratory conditions, sediments have been taken from their dynamic environment; rather than having the trace metal co-precipitate with the freshly precipitating Mn and Fe oxides, the added radiotracer will sorb to all existing binding sites, which include both oxides of Mn (easily reducible manganese) and Fe (reducible iron). This difference in partitioning under laboratory and field situations has important implications for those studies which are attempting to assess the availability of metals from natural sediments (e.g. Gagnon and Fisher, 1997). Further studies need to focus on ways to ensure that laboratory speciation is indicative of what is actually occurring in the environment. However, although the partitioning of the radiotracer did not differ between the two types of sediment, over a period of 4 days, a greater amount of 109Cd desorbed from DS than from SPM. This difference in the desorption of 109Cd from DS versus that of SPM suggests some dissimilarities in the association of the radiolabel between the two types of sediment; specifically, a greater amount of weakly sorbed tracer present on the DS as compared to SPM. The amount of radiotracer that desorbed from the SPM and DS and the differences in the amount of radiotracer
desorbed over time was an order of magnitude less than that which was recovered via the various extracts from the three sediment components of the DS and SPM. Hence, it is possible that the simultaneous extraction procedures as applied here are not sensitive enough to detect such small differences in trace metal partitioning in the two types of sediment which appeared to have occurred as indicated by the differences in amounts of radiotracer desorbed from the sediments over time. Given the almost twofold greater concentrations of easily reducible manganese and reducible iron in the SPM versus the DS, there are most likely a greater number of available binding sites for 109Cd on the SPM versus that of the DS. This could result in greater desorption of less strongly associated 109Cd from the DS as compared to the SPM.
4.2. Uptake of 109Cd by the two bi6al6es from the two types of radiolabeled sediments The majority of isotope uptake in both organisms occurred over the first 2 days and subsequently decreased. Similar observations were made by Harvey and Luoma (1985b) for 109Cd uptake by M. balthica from bacteria Pseudomonas atlantic and by Absil et al. (1996), who note that most 64Cu uptake occurred in the first 5–10 h of exposure in M. balthica. Both test organisms took up more isotope from the particulate phase than from the water; M. balthica took up 87% of its 109 Cd from the particulate phase in the DS exposures and 79% in the SPM exposures. P. staminea took up 91% of its 109Cd from the particulate phase in the DS exposures and 81% in the SPM exposures. The proportional values of the amount of 109Cd taken up via water or sediment determined here are similar to the 94% particulate uptake estimates by Thomann et al. (1995). They suggested that this high proportion could be due to the induction of metallothionein (i.e. Roesijadi, 1992), resulting in high assimilation and low depuration rates. Indeed, this is supported by the observed high biological half time (3–7 months) of Cd (Borchardt, 1983; Viarengo et al., 1985). However, Harvey and Luoma (1985b) noted that up-
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take from particulates contributed 67% of the total metal uptake after 2 days, 50% of the total after 8 days, but only 39% of the total after 14 days. They suggested that this was due to isotope depletion in the bacteria they were using as particulate food for M. balthica. However, in their experiment, depletion of 109Cd was much greater than in the present study with only 6.7% of the original 109Cd activity being associated with bacteria after 4 days. In comparison, in the current study, 77% of the 109Cd added to SPM and 65% of the 109Cd added to DS was still associated with the sediments after 4 days. In comparing uptake from the water phase with uptake from particulate phase, it is important to note that control organisms were not exposed to sediments. The absence of the sediment could alter the filtering activity of the organisms. A difference in filtering activity will affect how much dissolved metal an organism is exposed to. If sediment were added to the exposure vessel, dissolved metal would sorb to the sediment and uptake via the particulate route would occur precluding a perfect control for this experiment. The absence or shortage of food has been observed to strongly reduce ventilation rates (Riisgard and Randlov, 1981; Hummel, 1985) and therefore lower uptake of dissolved metals. Hence, it is likely that the controls represent an underestimation of 109Cd uptake via the water route in the treatment organisms.
4.3. Comparing uptake of 109Cd by two bi6al6es; the influence of sediment geochemistry and feeding beha6iour Luoma and Jenne (1977) found that uptake of isotopes of Ag, Co and Zn from water and particulate in M. balthica was a strong function of sediment–solute distribution. They concluded that the distribution ratios and therefore metal uptake were controlled by binding intensity. Similarly, 20 years later, Gagnon and Fisher (1997) noted that the assimilation of the radiotracers, Cd, Ag and Co by M. edulis was dependent on a combination of the extent of metal desorption from the labelled sediment particles under acidic
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conditions and the retention time of the ingested food in the digestive tract. For M. balthica uptake of the radiotracer from SPM was approximately two-thirds of that taken up from DS. For P. staminea, uptake of the radiolabel from SPM was negligible compared to DS. For M. balthica, assuming that filter-feeding by the bivalve was the same under both sediment exposures, i.e. M. balthica ingesting both SPM and DS in equal amounts, differences in amounts of radiotracer desorbed from SPM versus DS could have resulted in greater uptake of the tracer from DS versus that of SPM. However, this explanation cannot be as readily applied to differences in uptake of the 109Cd from SPM versus DS for P. staminea. The feeding characteristics of the two clams differ substantially. M. balthica is a facultative filter feeder, that can ‘vacuum’ bottom sediments with its long siphon with some selectivity (Absil et al., 1996). M. balthica is also able to filter food from the overlying water column while the inhalant siphon is just at the sediment surface (Brafield and Newell, 1961; Hummel, 1985). P. staminea is an obligate filter-feeder that relies exclusively upon inhalant water currents to take up food via its short siphons (George and George, 1979; Crecelius et al., 1982). It is highly sensitive to overlying water column particulate concentrations and will stop filter-feeding if particle concentrations are too high. During the SPM exposure experiment, P. staminea accumulated only a fraction of the radiotracer as compared to M. balthica or to the DS exposures. A probable reason for this is that in the SPM exposure, and as evidenced by the almost negligible uptake of the radiotracer during the SPM exposure, under less than optimal filter-feeding conditions (i.e. a high, fine, silt load) P. staminea simply ceased to filter-feed. In the DS experiments a greater portion of the sediment settled to the bottom of the aquaria, providing more optimal filter-feeding conditions in the overlying water, which in turn, allowed P. staminea to filter in an uninterrupted manner. In contrast, throughout both exposures, M. balthica continued to ‘vacuum’ the bottom of the aquaria (personal observation) and did not rely upon inhalant cur-
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rents, hence, it was able to maintain consistent uptake in both experiments. 5. Summary and conclusions Uptake of the radiotracer 109Cd by the bivalves M. balthica and P. staminea was dependent on the degree of isotope desorption from the sediment (for M. balthica) and feeding behaviour (for P. staminea). M. balthica accumulated radiotracer from both SPM and DS with a greater amount of uptake occurring from the DS. This uptake could be explained by the higher rates of radiotracer desorption from DS versus that of SPM. P. staminea accumulated a minimal amount of radiotracer from the SPM, most likely a consequence of high sediment loads in the test aquaria which precluded optimum filter-feeding conditions. However, significant amounts of radiotracer were taken up from DS compared to water only, indicating that sediments are an important source of contaminants to this obligate filter-feeder as well. Natural cadmium was primarily associated with the easily reducible components of suspended and deposited sediment as compared to the reducible and organic component of sediment. In contrast, the radioisotope partitioned primarily on the reducible component of both types of sediment. Hence, in these experiments the geochemical behaviour of the radiotracer did not parallel that which was observed in the natural environment. This lack of agreement between natural and experimental partitioning of cadmium among the three main sediment components of SPM and DS precluded our first objective of assessing the relative importance of SPM and DS as a source of contaminants to facultative filter-feeders. Further studies need to focus on ways to spike or label natural complex sediments, such that the partitioning achieved under laboratory conditions reflects that which is observed in the field. Acknowledgements Funding support to L. Bendell-Young in the form of an NSERC operating grant is gratefully acknowledged.
References Absil, M.C.P., Berntssen, M., Gerringa, L.J.A., 1996. The influence of sediment, food and organic ligands on the uptake of copper by sediment-dwelling bivalves. Aquat. Toxicol. 34, 13 – 29. Bendell-Young, L.I., Harvey, H.H., 1991. Metal concentrations in chironomids in relation to geochemical characteristics of surficial sediments. Arch. Environ. Contam. Toxicol. 21, 202 – 211. Bendell-Young, L.I., Dutton, M., Pick, F.R., 1992. Contrasting two methods for determining trace metal partitioning in oxidized lake sediments. Biogeochemistry 17, 205 – 219. Borchardt, T., 1983. Influence of food quantity on the kinetics of cadmium uptake and loss via food and seawater in Mytilus edulis. Mar. Biol. 76, 67 – 76. Brafield, A.E., Newell, G.E., 1961. The behaviour of Macoma balthica (L.). J. Mar. Biol. Assoc. UK 41, 81 – 87. Crecelius, E.A., Hardy, J.T., Bobson, C.I., Schmidt, R.L., Apts, C.W., Gurtisen, J.M., Joyce, S.P., 1982. Copper bioavailability to marine bivalves and shrimp. Relationship to cupric ion activity. Mar. Environ. Res. 6, 13 – 26. Egeberg, P.K., Schaanning, M., Naes, K., 1988. Modelling the manganese cycling in two stratified fjords. Mar. Chem. 23, 383 – 391. Forstner, U., Wittman, G., 1981. Metal Pollution in the Aquatic Environment, 2nd edn. Springer, New York. Gagnon, C., Fisher, N., 1997. The bioavailability of sedimentbound Cd, Co, and Ag to the mussel Mytilus edulis. Can. J. Fish. Aquat. Sci. 54, 147 – 156. George, J.D., George, J.J., 1979. Marine Life: An Illustrated Encyclopaedia of Invertebrates in the Sea. Wiley, New York. Harvey, R.W., Luoma, S.N., 1985a. Separation of solute and particulate vectors of heavy metal uptake in controlled suspension feeding experiments with Macoma balthica. Hydrobiologia 121, 97 – 102. Harvey, R.W., Luoma, S.N., 1985b. Effect of adherent bacteria and bacterial extracellular polymers upon assimilation by Macoma balthica of sediment bound Cd, Zn, and Ag. Mar. Ecol. Prog. Ser. 22, 281 – 289. Hummel, H., 1985. Food intake of Macoma balthica (Mollusca) in relation to seasonal changes in its potential food on a tidal mudflat in the Dutch Wadden Sea. Neth. J. Sea Res. 19, 52 – 76. Jenne, E.A., Zachara, J.M., 1987. Factors influencing the sorption of metals. In: Dickson, K.L., Maki, A.W., Brungs, W.A. (Eds.), Fate and Effects of Sediment-bound Chemicals in Aquatic Systems. Pergamon, New York, pp. 83 – 98. Luoma, S.N., Jenne, E.A., 1977. The availability of sedimentbound cobalt, silver, and zinc to a deposit feeding clam. In: Wildung, R.E., Drucker, H. (Eds.), Biological Implications of Metals in the Environment. NTIS CONF-750920, Springfield, pp. 213 – 230. Olafsson, E.B., 1986. Density dependence in suspension feeding and deposit feeding populations of the bivalve Macoma balthica: a field experiment. J. Anim. Ecol. 55, 517 – 526.
J.R.P. Stecko, L.I. Bendell-Young / Aquatic Toxicology 47 (2000) 147–159 Ongley, E.D., Byne, M.C., Percival, J.B., 1982. Physical and geochemical characteristics of suspended solids, Wilton Creek, Ontario. Hydrobiologica 91, 41–57. Regnier, P., Wollast, R., 1993. Distribution of trace metals in suspended matter of the Scheldt estuary. Mar. Chem. 43, 3–19. Riise, G., Salbu, B.R., Singh, Steinnes, E., 1994. Distribution of 109Cd among different soil fractions studied by a sequential extraction technique. Water Air Soil Pollut. 73, 285 – 295. Riisgard, H.U., Randlov, A., 1981. Energy budgets, growth and filtration rates in Mytilus edulis at different algal concentrations. Mar. Biol. 61, 227–234. Roesijadi, G., 1992. Metallothioneins in metal regulation and toxicity in aquatic animals. Aquat. Toxicol. 22, 81–114. Rule, J.H., Allen, R.W., 1996. Interactions of Cd and Cu in aerobic estuarine sediments. II Bioavailability, body burdens and respiration effects as related to geochemical partitioning. Environ. Toxicol. Chem. 15, 466–471. Salomons, W., Forstner, U., 1984. Metals in the Hydrocycle. Springer, New York, p. 349. Santschi, P., Hohener, P., Benoit, G., Buchholtz-ten Brink, M., 1990. Chemical processes at the sediment–water interface. Mar. Chem. 30, 269–315.
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SAS, 1990. SAS Users Guide, Version 6.08. SAS Institute Inc., Cary, NC. Stecko, J.R.P., Bendell-Young, L.I., 1999. Contrasting the geochemistry of suspended particulate matter and deposited sediment within an estuary. Appl. Geochem., in press. Tessier, A., Campbell, P.G.C., Auclair, A.J., Bisson, M., 1984. Relationships between the partitioning of trace metals in sediments and their accumulation in the tissues of the freshwater mollusc Elliptio complanata in a mining area. Can. J. Fish. Aquat. Sci. 41, 1463 – 1472. Thomann, R.V., Mahoney, J.D., Mueller, R., 1995. Steadystate model of biota sediment accumulation factor for metals in two marine bivalves. Environ. Toxicol. Chem. 14 (11), 1989 – 1998. Thomas, C., Bendell-Young, L.I., 1998. Linking the sediment geochemistry of an intertidal region to metal availability in the bivalve Macoma balthica. Mar. Ecol. Prog. Ser. 173, 197 – 213. Viarengo, A., Palermo, S., Zanicchi, G., Capelli, R., Vaissiere, R., Orunesu, M., 1985. Role of metallothioneins in Cu and Cd accumulation and elimination in the gill and digestive gland cells of Mytilus gallopro6incialis L. Mar. Environ. Res. 16, 23 – 36.
Uptake of
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Cd from sediments by the bivalves Macoma balthica and Protothaca staminea J.R. Pierre Stecko, L.I. Bendell-Young *
Dept. of Biological Sciences, Simon Fraser Uni6ersity, 8888 Uni6ersity A6e., Vancou6er, V5A 1S6, Canada Received 23 April 1998; received in revised form 8 March 1999; accepted 30 March 1999
Abstract This study (1) compared the relative importance of suspended particulate matter (SPM) and deposited sediment (DS) as a source of cadmium exposure to two sediment ingesting bivalves, and (2) determined the importance of feeding behaviour on cadmium uptake from SPM and DS by comparing metal uptake from ingested sediments by a facultative and an obligate filter-feeder. Two types of sediment, deposited (DS) and suspended (SPM) sediments were sampled from the Fraser River Estuary and geochemically characterized with respect to amounts of easily reducible manganese and reducible iron (manganese and iron oxides respectively), organic matter, and amounts of cadmium associated with each of the three components. Sediments were radiolabeled with 109Cd, with the labelled sediments then fed to Macoma balthica and Protothaca staminea, a facultative filter-feeder and an obligate filter-feeder respectively. Amounts of radiotracer accumulated by the bivalves from the two types of sediment over an 8-day period were then assessed. The geochemistry of DS was distinct from that of SPM with DS containing almost half the concentrations of oxides of manganese and iron as compared to SPM. Natural cadmium was recovered primarily from the easily reducible (associated with oxides of manganese) followed by the reducible (associated with oxides of iron) sediment component, for both SPM and DS. In contrast, 109Cd was recovered from the reducible fraction, followed by the easily reducible fraction from the two types of sediment. No natural cadmium or 109Cd was recovered from the organic component of either sediment. For both bivalves, uptake of 109Cd from sediment as compared to water accounted for ca. 80% of accumulated radiotracer. Uptake of 109Cd from DS was significantly greater than from SPM, most notably for P. staminea. For M. balthica, the amount of accumulated 109Cd was best explained by the degree of isotope desorption from DS. In contrast, 109Cd accumulation by P. staminea was best described by its feeding behaviour. Metal accumulation by sediment-ingesting organisms is not simply related to sediment metal concentrations but will also depend on the strength of association of the metal within the sediment (for facultative feeders), and the feeding behaviour of the bivalve. 109Cd partitioning in DS and SPM did not parallel the field partitioning of cadmium which precluded our objective of assessing the relative importance of SPM and DS as a source of cadmium to facultative filter-feeders. To elucidate the role of complex sediment geochemistry in influencing
* Corresponding author. Tel.: + 1-604-291-5621; fax: + 1-604-291-3496. E-mail address: [email protected] (L.I. Bendell-Young) 0166-445X/00/$ - see front matter © 2000 Elsevier Science B.V. All rights reserved. PII: S 0 1 6 6 - 4 4 5 X ( 9 9 ) 0 0 0 2 3 - 5
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metal availability to sediment ingesting organisms, further study is needed on ways to either label or spike natural complex sediments such that the speciation achieved under laboratory conditions is that which is observed in the field. © 2000 Elsevier Science B.V. All rights reserved. Keywords: Sediment geochemistry; Bivalve; Filter-feeding behaviour; Metal uptake
1. Introduction The behaviour of metals in aquatic systems has been of increased interest due to the global anthropogenic alteration of trace metal cycles (Forstner and Wittman, 1981; Salomons and Forstner, 1984). Anthropogenic metals introduced into the aquatic environment are either in particulate form or are rapidly sorbed to particles (Jenne and Zachara, 1987; Regnier and Wollast, 1993). Once sorbed, they contribute to the pool of metals associated with suspended particulate matter (SPM) within the water column or are ultimately deposited to the sediments as deposited sediments (DS). For organisms that ingest either DS and/or SPM as a food source, these anthropogenic metals have the potential for causing severe adverse effects on the organism plus potentially being bioaccumulated by the organism possibly leading to the food-chain transfer of the trace metal to higher trophic levels. The availability of anthropogenic metals to estuarine organisms from the sediment is dependent in part on the presence of the trace metal (exposure), plus the geochemical phase with which the trace metal is associated (Tessier et al., 1984; Bendell-Young and Harvey, 1991; Rule and Allen, 1996). Several studies have suggested that the easily reducible metal is most available to organisms. For example, Rule and Allen (1996) and Thomas and Bendell-Young (1998) under laboratory and field conditions respectively, found that metal associated with the easily reducible fraction of sediment (i.e. associated with oxides of manganese) was the best predictor of tissue cadmium concentrations in the sediment ingesting Macoma balthica. Recently, Stecko and Bendell-Young (1999) contrasted the geochemistry of SPM versus DS in the Fraser River Estuary and found that the geochemistry of SPM was distinct from that of
DS; SPM contained greater amounts of organic matter, reducible iron content (iron oxide) and total metal concentrations (zinc, copper, cadmium and lead) with a greater proportion of the metal present in an easily reducible form compared to DS. These differences in the geochemistry and potential metal bioavailability between SPM and DS are important in regards to sediment ingesting organisms that feed on either SPM or DS, or organisms that are capable of exploiting both types of sediments as potential food sources. For example, the suspension-feeding bivalves Mya arenaria and M. balthica are facultative filter-feeders, capable of feeding on deposited as well as suspended sediment with the ‘choice’ of sediment being dependent on the quantity of particles (Olafsson, 1986). In contrast, Protothaca staminea and Cerastoderma edule are obligate filter-feeders filtering SPM from the water column relying on inhalant water currents to take up food (Crecelius et al., 1982). If metal bioavailability is dependent on sediment geochemistry as found for example, by Rule and Allen (1996) and Thomas and Bendell-Young (1998), and if the organism has the option of switching feeding modes depending on particle flux, then there is the potential for metal exposure and subsequent availability to sediment ingesting organisms, specifically facultative filter-feeders, to differ between DS and SPM with differences being dependent on feeding behaviour. In contrast, for obligate filter-feeders, exposure should occur only through one sediment type. Hence, the objectives of this study were twofold: 1. to compare the relative importance of SPM and DS as a source of trace metal exposure to two sediment ingesting bivalves and 2. to determine the importance of feeding behaviour on metal uptake from SPM and DS by comparing metal uptake by a facultative and an obligate filter-feeder.
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To meet these objectives, two types of natural sediment, collected from the field, both DS and SPM, were radiolabeled with 109Cd. Labelled sediments were then fed to M. balthica and P. staminea, a facultative filter-feeder and an obligate filter-feeder respectively. Amounts of radiotracer accumulated over an 8-day period were then assessed. Ultimately, we hope that the results of our study will lead to a greater predictive capacity for determining which sediment type (in this case, SPM versus DS) and the feeding behaviour (facultative or obligate filter-feeder) that results in the greatest metal exposure to sediment ingesting organisms via their diet. 2. Materials and methods
2.1. Field collection of sediments and bi6al6es Deposited sediment (DS) and suspended particulate matter (SPM) were collected from a mid-estuarine site of the Fraser River Estuary, B.C., Canada. Because the B63-mm fraction of suspended sediment is not depth-dependent, suspended sediment recovery for geochemical purposes can be accomplished by sampling close to the water surface (Ongley et al., 1982). To collect the evenly distributed B63 mm fraction of SPM, near surface water (4000 l total) was collected on mudflats at 1.5 m of water depth. Visible resuspensions were carefully avoided. For DS, several sediment cores (using a Wildco® hand corer) were taken at the sample site; the oxic (surficial) portion of these cores (above the visible redox colour change) were put into acid-washed glass jars for transport to the laboratory. The oxic portion generally varied from 2 to 8 cm depth. Organisms were collected from the intertidal region of the Fraser River Estuary. P. staminea, the common Pacific littleneck clam, was collected by raking the mudflats and obtaining organisms ranging in size from 3.5 to 5 cm (approximately 4 – 6 years old). M. balthica, the Baltic tellin, was collected by gently hand raking the sediment, feeling for the small (8 – 14 mm) clams. Organisms were stored in buckets containing sea-water and were transported to the lab where they were acclimatized and depurated in a filtered sea-water and
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dechlorinated water mixture for a minimum of 72 h before the initiation of the uptake experiments. 2.2. Sediment geochemistry Sediment samples were immediately transported to the laboratory where they were kept refrigerated at 5°C until processing (always within 72 h). SPM was separated from the water by flow through centrifugation at 43 807.5× g using the Sorvall® TZ-28 system in GK continuous flow mode in a Sorvall® RC5C Centrifuge with a flow rate of between 0.6 and 0.75 l/min. Sediment samples were simultaneously extracted after methods of Bendell-Young et al. (1992) which separates sediment into operationally defined components including: (1) easily reducible manganese and metal associated with manganese oxides, (2) reducible iron and metal associated with iron oxides, and, (3) organically bound metal (Fig. 1). Organic content was determined by loss on ignition (%LOI) at 600°C for 1 h in a blast furnace. All sediment samples were treated with a 15 to 1 ratio of analytical-grade extractant (Baker Instra Analyzed or BDH Analar) to sediment. Supernatants were removed from the extraction tubes through Millipore Millex-HV sterilizing filter units (0.45 mm) using acid washed 10-ml syringes. Metal analyses for manganese and iron were carried out by flame atomic absorption spectrophotometry (FAAS) on a Perkin-Elmer® 1100B FAAS. Analyses for cadmium was done by graphite furnace atomic absorption (GFAA), at the University of British Columbia, Department of Oceanography. Quality assurance/quality control (QA/QC) was assured through inclusion of NRC reference standards (MESS-2) and reagent blanks. Reference standards fell within 10% of certified values.
2.3. Isotope partitioning To determine how the isotope distributed within the sediments, in triplicate, 1.000 g of wet DS and 1.605 g of wet SPM to achieve an equivalent dry weight of 0.6 g (where the dry/wet ratios for each sediment are 0.61 and 0.38 for DS and SPM respectively) were added to centrifuge tubes,
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with 25 ml of water (a mixture of sea-water and dechlorinated distilled deionised water to achieve a salinity of 14 ppt), and spiked with 27.8 kBq of 109 Cd (\99% purity, t1 453 days) giving a final 2 concentration of 1112 Bq/ml (12.15 ppb) 109Cd. Slurries were mixed several times daily and sampled at 1, 3 and 7 weeks. At the end of the interval, the slurries were centrifuged, extracted by methods of Bendell-Young et al. (1992), and the extracts counted for radioactivity on instrumentation as outlined below.
2.4. Exposures The experimental design used three replicate 38 l aquaria with 36 P. staminea and 20 M. balthica per aquaria. Controls were run in the exposure aquaria using 0.45-mm Millipore filters as dividers (Harvey and Luoma, 1985a,b). To ensure that the control side was in equilibrium throughout the exposure experiments, filtered (0.45-mm Millipore filter) water samples from the control side of the aquaria were analysed for dissolved 109Cd to com-
pare with dissolved 109Cd (0.45-mm Millipore filter) in the exposure side of the aquaria. Equilibrium of dissolved 109Cd between the control and exposure sides of the tank was achieved within 48–96 h of addition of the radiolabelled sediments to the exposure side of the tank (Fig. 2). This experimental design therefore allowed half the organisms to be exposed to the particulate plus the dissolved radioisotope, whereas the other half was exposed only to the free (dissolved) radioisotope. Salinity was held at levels approximating levels found at the collection site ( 14 ppt) using an appropriate mixture of filtered seawater (1 mm) and dechlorinated distilled deionized freshwater. Temperature was set at 12°C. To ensure equal exposures of the bivalves to the two sediment types, radiolabeling of the sediments for the exposures with 109Cd was done on an equivalent dry weight of 5.00 g of sediment; 8.19 g of wet DS and 13.16 g of wet SPM, at dry/wet ratios of 0.61 and 0.38 respectively. To each of three high density polyethylene (HDPE) bottles were added the sediment and 205 ml of sea-water
Fig. 1. Extraction scheme of Bendell-Young et al. (1992) to determine the geochemical characteristics of DS and SPM.
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Fig. 2. Desorption of 109Cd from SPM and DS as a function of time. Values are means of three measurements9 1 S.E. (a) Dissolved 109Cd in the exposure versus the control tank for the DS exposures; (b) dissolved 109Cd in the exposure versus the control tank for the SPM exposures. Note that equilibrium between the exposure and exposure tanks is reached at approximately 96 h for DS and SPM. E, exposure side of tank; C, control side of tank.
(14 ppt) taken from the sampling site. To these bottles, 2268 kBq of 109Cd was added yielding an
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interim slurry of 10.84 kBq/ml 109Cd. Sediments were shaken every day for 12 days to allow for desorption of most of the loosely bound 109Cd prior to being added to the exposure aquaria. In the exposure aquaria, the total isotope concentration was diluted to 222 Bq/ml (equivalent concentrations of 2.43 ppb) 109Cd. Eighteen P. staminea and 10 M. balthica were used on each side of the aquarium. The majority of added DS and SPM was kept suspended by the turbulence generated by oxygen stones, although the heavier fraction of DS tended to settle out onto the bottom of the aquaria. This settling of DS but not SPM had important implications in determining final tissue radiotracer levels in the bivalves. Bivalves were added to the aquaria 24 h after the addition of the radiolabeled sediments, as the two sides of the aquaria were coming into equilibrium. Exposure of the bivalves to the radiolabeled sediments occurred for a period of 8 days, followed by 6 days of depuration where bivalves were exposed to non-labelled ‘cold’ sediments. Three organisms were removed from each side of the aquaria, control and exposure, on days 1, 2, 4, and 8. The mussels were depurated for 24 h in uncontaminated sediment before determining isotope activity. After 8 days of exposure, the remaining mussels were transferred to an uncontaminated environment to examine elimination of 109Cd P. staminea were sacrificed, the tissue removed from the shell and rinsed in EDTA (to remove any externally absorbed metal), blotted and weighed before analysis. M. balthica were rinsed with EDTA, blotted and weighed before analysis whole; following analysis they were returned to the appropriate aquaria. Simple uptake calculations involved the determination of the amount of radioisotope in the organisms in Bq/g wet wt., at the various time intervals.
2.5. Mass balance of radiotracer Following completion of the exposure experiments, swipe tests of the aquaria (aquaria walls, Millipore membrane), and sediments and water
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were sampled to allow for a mass balance of the radiotracer added to the aquaria (exclusive of bivalves).
2.6. Instrumentation Cd was detected at 88 keV g emission using an automatic g counting system. The system uses Nucleus Analyser Software, and measures g radiation with a sodium iodide crystal. Canberra preamplifiers and amplifiers supply the signal to the computer. The system was optimized using standards of the isotope with optimal setting at 820 V for the power supply and an amplifier setting of 3 K. Counting efficiency for 109Cd was 4.33%. 109
2.7. Statistical analyses Statistical analyses were carried out using SAS, version 6.08 (SAS, 1990). Comparison of amounts of dissolved radiotracer in the exposure versus the control side of the tank at t \96 h was done using simple Student t-tests; comparison of sediment characteristics between SPM and DS was done using simple Students t-tests; comparison of amount of radiotracer sorbed to the three sediment geochemical components was assessed through a one-way ANOVA. Radiotracer uptake from the two types of sediment for each sampling day by the two bivalves was analysed through two-way ANOVA with bivalve and sediment type as the two co-factors (SAS, 1990). Significance for all tests was accepted at P B0.05.
ever, in both cases, reducible iron and easily reducible manganese concentrations were significantly greater in SPM compared to DS (P B 0.001). Concentrations of natural Cd were sevenfold greater in SPM versus that of DS. Although there were differences in total concentrations, partitioning of Cd among the three sediment components was similar for both sediment types, with the majority of natural speciated Cd being recovered from the easily reducible component of both sediments (Fig. 3). No Cd was recovered from the organic phase of sediment.
3.2. Isotope partitioning and desorption in SPM and DS The partitioning of the radioisotope in SPM and DS following 1, 3 and 7 weeks of equilibration indicated that despite the differences in geochemistry of sediment used in the exposures, 109 Cd partitioning in the two sediments did not differ (Fig. 4; ANOVA; P\0.05). For both sediments, the same amount of radiolabel was recovered from the reducible fraction, followed by the easily reducible fraction of sediment. Only trace amounts of the radiolabel were recovered from Table 1 Geochemical characteristics of deposited sediments (DS) and suspended particulate matter (SPM) sampled from (a) Fraser River Estuary (August 1995) and (b) those subsequently used for the radiolabelling and exposure experimentsa SPM
3. Results
3.1. Sediment geochemistry Geochemical characteristics of SPM and DS sampled from a mid-estuarine site in the Fraser River Estuary and the geochemical characteristics of the sediments used in the isotope partitioning and desorption experiment and the exposures are outlined in Table 1. For sediments used for geochemical characterization, isotope experiments and the exposures, there was no significant difference between the organic content (LOI) of the two sediments (Students t-test; P =0.1432), how-
(a) Fraser Ri6er Estuary N LOI (%) RFe (mg/kg) ERMn (mg/kg) (b) Radiolabelling and exposure experiments N LOI (%) RFe (mg/kg) ERMn (mg/kg)
DS
9 6.62 91.58 9280 9783 293 934
9 4.41 90.35 50609301 184920
5 6.5 91.29 12 320 9478 418 98
3 4.19 90.21 6545 9293 197 935
Values are mean 9S.E.; LOI (%), loss on ignition (organic matter); RFe, reducible iron (oxides of iron); ERMn, easily reducible manganese (oxides of manganese). a
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Although the partitioning of the radiolabel did not differ between the two types of sediment, greater amounts of 109Cd desorbed from DS than from SPM (Students t-test; PB 0.05) (Fig. 2). After allowing for a 96-h equilibration period, average (9 1 S.E.) dissolved activity of 109Cd for the DS exposures was 62.59 0.89 (Bq/ml), whereas for the SPM exposures, average dissolved 109 Cd activity was 39.691.17 (Bq/ml) (Fig. 2). The mass balance of the isotope indicated that only 0.7% of the added radiotracer was associated with the walls of the aquaria, the Millipore membrane and its frame. Approximately 73 and 26% were associated with the sediment and water respectively.
3.3. Exposures Fig. 3. Distribution of natural cadmium among the three sediment components in SPM and DS sampled from the Fraser River Estuary. Values are means of nine measurements9 1 S.E. ER, easily reducible; R, reducible; O, organic component of sediment.
the organic component of sediment. The partitioning of the radiolabel also remained consistent over the 7-week period (Figs. 4 and 5).
Uptake of 109Cd from SPM and DS was measured in M. balthica (Fig. 6a) and P. staminea (Fig. 6b). The concentration of isotope in both the sediment (350 kBq/g or 3.9 mg/g) and water (30 Bq/ml or 0.33 ng/ml) and the exposure conditions of 14 ppt salinity and 12°C were environmentally realistic. Activities reported for P. staminea are for the tissue of depurated organisms, and the activities reported for M. balthica
Fig. 4. Distribution of 109Cd among the three sediment components in SPM and DS at 1, 3 and 7 weeks. Values are means of three measurements91 S.E. ER, easily reducible; R, reducible; O, organic component of sediment.
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Fig. 5. Relative proportion of the radiolabel recovered from the three sediment fractions from DS and SPM at 1, 3 and 7 weeks.
are for whole depurated organisms. Two-way ANOVA indicated that the effect of sediment type was significant (N =48, F = 6.44, P = 0.015) whereas organism type and the interaction were not. For both organisms, uptake from DS was significantly greater than that from SPM, with this difference most notable for P. staminea. For M. balthica, most 109Cd uptake occurred over the first 2 days of both experiments with subsequent days showing either lower net uptake or net loss. Following the removal from the exposure vessel on day 8, 109Cd levels in M. balthica decreased. This depuration is evident in both the particulate and water, and the water only exposures. In the DS exposure, most of the uptake by P. staminea occurred over the first 2 days, with consistently low uptake on subsequent days. Decreased 109Cd levels were observed following depuration. During the SPM exposure, P. staminea showed a consistently low uptake with 109 Cd concentrations increasing slightly during depuration. In the DS exposures, M. balthica took up 86.894.0% (mean 9 S.E.) of its 109Cd from the particulate phase, while P. staminea took up 90.6 92.1% from the particulate phase. In the SPM exposures, M. balthica took up 78.9 93.8% 109 Cd from the particulate phase, while P. staminea took up 80.897.3% from the particulate
phase. Because the quantity of food consumed by these organisms was not measured, no comparisons based upon total uptake can be made, although the amount of radiotracer accumulated by the two bivalves over the 8-day period was similar.
4. Discussion
4.1. Sediment geochemistry of SPM
109
Cd in DS and
Environmental concentrations of cadmium were recovered primarily from the easily reducible followed by the reducible sediment components in both sediment types. No natural cadmium was recovered from the organic phase of sediments. Riise et al. (1994) also noted that 109Cd partitioned onto the easily reducible phase of soil (rather than sediment) supporting our field results. In contrast, 109Cd was found to associate predominantly with the reducible and to a lesser extent, the easily reducible phase with the same amount of radioactivity recovered from each sediment component for both SPM and DS. This partitioning of 109Cd was surprising given the differences in sediment geochemistry between the SPM and DS (i.e. SPM contained greater
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amounts of reducible iron and easily reducible manganese as compared to DS). Based on the behaviour of cadmium in the natural environment, more of the radiolabel should have partitioned onto the easily reducible components of sediment. This lack of a partitioning difference between the two sediments plus the equivalent amount of radioactivity recovered from each sediment component suggests that in-
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troduction of the radiotracer to sediments under laboratory conditions (i.e. addition of a relatively large amount at one time) and the subsequent partitioning of the radiolabel among the sediment components does not allow for the same processes and hence the same partitioning as that observed in nature. Unlike experimental conditions, in nature, low concentrations of the trace metal are added over a long term to an environment where
Fig. 6. 109Cd activity in M. balthica (a) and P. staminea (b) for both SPM and DS exposures at days 1, 2, 4, 8, 10 and 14. Values are means of three measurements 9 1 S.E.
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the geochemical composition of the suspended and deposited sediment are undergoing constant change. In the natural environment, both iron and manganese undergo a continuous redox cycle which leads to freshly precipitated oxides of manganese and iron, either within the oxic water column or at the sediment/water interface. The oxidation of Mn occurs primarily on the surfaces of particles via oxidizing bacteria (Santschi et al., 1990), whereas the oxidation of Fe is primarily accomplished without the mediation of oxidizing bacteria (Egeberg et al., 1988; Santschi et al., 1990). These freshly precipitated oxides present in SPM and DS will co-precipitate trace metals such as cadmium. What we recover from the natural sediment sampled from the field are these freshly precipitated oxides of Mn and Fe (as easily reducible Mn and Fe, i.e. oxides of Mn and the amorphous oxides of Fe), and cadmium which has been scavenged by these oxides. In contrast, under laboratory conditions, sediments have been taken from their dynamic environment; rather than having the trace metal co-precipitate with the freshly precipitating Mn and Fe oxides, the added radiotracer will sorb to all existing binding sites, which include both oxides of Mn (easily reducible manganese) and Fe (reducible iron). This difference in partitioning under laboratory and field situations has important implications for those studies which are attempting to assess the availability of metals from natural sediments (e.g. Gagnon and Fisher, 1997). Further studies need to focus on ways to ensure that laboratory speciation is indicative of what is actually occurring in the environment. However, although the partitioning of the radiotracer did not differ between the two types of sediment, over a period of 4 days, a greater amount of 109Cd desorbed from DS than from SPM. This difference in the desorption of 109Cd from DS versus that of SPM suggests some dissimilarities in the association of the radiolabel between the two types of sediment; specifically, a greater amount of weakly sorbed tracer present on the DS as compared to SPM. The amount of radiotracer that desorbed from the SPM and DS and the differences in the amount of radiotracer
desorbed over time was an order of magnitude less than that which was recovered via the various extracts from the three sediment components of the DS and SPM. Hence, it is possible that the simultaneous extraction procedures as applied here are not sensitive enough to detect such small differences in trace metal partitioning in the two types of sediment which appeared to have occurred as indicated by the differences in amounts of radiotracer desorbed from the sediments over time. Given the almost twofold greater concentrations of easily reducible manganese and reducible iron in the SPM versus the DS, there are most likely a greater number of available binding sites for 109Cd on the SPM versus that of the DS. This could result in greater desorption of less strongly associated 109Cd from the DS as compared to the SPM.
4.2. Uptake of 109Cd by the two bi6al6es from the two types of radiolabeled sediments The majority of isotope uptake in both organisms occurred over the first 2 days and subsequently decreased. Similar observations were made by Harvey and Luoma (1985b) for 109Cd uptake by M. balthica from bacteria Pseudomonas atlantic and by Absil et al. (1996), who note that most 64Cu uptake occurred in the first 5–10 h of exposure in M. balthica. Both test organisms took up more isotope from the particulate phase than from the water; M. balthica took up 87% of its 109 Cd from the particulate phase in the DS exposures and 79% in the SPM exposures. P. staminea took up 91% of its 109Cd from the particulate phase in the DS exposures and 81% in the SPM exposures. The proportional values of the amount of 109Cd taken up via water or sediment determined here are similar to the 94% particulate uptake estimates by Thomann et al. (1995). They suggested that this high proportion could be due to the induction of metallothionein (i.e. Roesijadi, 1992), resulting in high assimilation and low depuration rates. Indeed, this is supported by the observed high biological half time (3–7 months) of Cd (Borchardt, 1983; Viarengo et al., 1985). However, Harvey and Luoma (1985b) noted that up-
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take from particulates contributed 67% of the total metal uptake after 2 days, 50% of the total after 8 days, but only 39% of the total after 14 days. They suggested that this was due to isotope depletion in the bacteria they were using as particulate food for M. balthica. However, in their experiment, depletion of 109Cd was much greater than in the present study with only 6.7% of the original 109Cd activity being associated with bacteria after 4 days. In comparison, in the current study, 77% of the 109Cd added to SPM and 65% of the 109Cd added to DS was still associated with the sediments after 4 days. In comparing uptake from the water phase with uptake from particulate phase, it is important to note that control organisms were not exposed to sediments. The absence of the sediment could alter the filtering activity of the organisms. A difference in filtering activity will affect how much dissolved metal an organism is exposed to. If sediment were added to the exposure vessel, dissolved metal would sorb to the sediment and uptake via the particulate route would occur precluding a perfect control for this experiment. The absence or shortage of food has been observed to strongly reduce ventilation rates (Riisgard and Randlov, 1981; Hummel, 1985) and therefore lower uptake of dissolved metals. Hence, it is likely that the controls represent an underestimation of 109Cd uptake via the water route in the treatment organisms.
4.3. Comparing uptake of 109Cd by two bi6al6es; the influence of sediment geochemistry and feeding beha6iour Luoma and Jenne (1977) found that uptake of isotopes of Ag, Co and Zn from water and particulate in M. balthica was a strong function of sediment–solute distribution. They concluded that the distribution ratios and therefore metal uptake were controlled by binding intensity. Similarly, 20 years later, Gagnon and Fisher (1997) noted that the assimilation of the radiotracers, Cd, Ag and Co by M. edulis was dependent on a combination of the extent of metal desorption from the labelled sediment particles under acidic
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conditions and the retention time of the ingested food in the digestive tract. For M. balthica uptake of the radiotracer from SPM was approximately two-thirds of that taken up from DS. For P. staminea, uptake of the radiolabel from SPM was negligible compared to DS. For M. balthica, assuming that filter-feeding by the bivalve was the same under both sediment exposures, i.e. M. balthica ingesting both SPM and DS in equal amounts, differences in amounts of radiotracer desorbed from SPM versus DS could have resulted in greater uptake of the tracer from DS versus that of SPM. However, this explanation cannot be as readily applied to differences in uptake of the 109Cd from SPM versus DS for P. staminea. The feeding characteristics of the two clams differ substantially. M. balthica is a facultative filter feeder, that can ‘vacuum’ bottom sediments with its long siphon with some selectivity (Absil et al., 1996). M. balthica is also able to filter food from the overlying water column while the inhalant siphon is just at the sediment surface (Brafield and Newell, 1961; Hummel, 1985). P. staminea is an obligate filter-feeder that relies exclusively upon inhalant water currents to take up food via its short siphons (George and George, 1979; Crecelius et al., 1982). It is highly sensitive to overlying water column particulate concentrations and will stop filter-feeding if particle concentrations are too high. During the SPM exposure experiment, P. staminea accumulated only a fraction of the radiotracer as compared to M. balthica or to the DS exposures. A probable reason for this is that in the SPM exposure, and as evidenced by the almost negligible uptake of the radiotracer during the SPM exposure, under less than optimal filter-feeding conditions (i.e. a high, fine, silt load) P. staminea simply ceased to filter-feed. In the DS experiments a greater portion of the sediment settled to the bottom of the aquaria, providing more optimal filter-feeding conditions in the overlying water, which in turn, allowed P. staminea to filter in an uninterrupted manner. In contrast, throughout both exposures, M. balthica continued to ‘vacuum’ the bottom of the aquaria (personal observation) and did not rely upon inhalant cur-
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rents, hence, it was able to maintain consistent uptake in both experiments. 5. Summary and conclusions Uptake of the radiotracer 109Cd by the bivalves M. balthica and P. staminea was dependent on the degree of isotope desorption from the sediment (for M. balthica) and feeding behaviour (for P. staminea). M. balthica accumulated radiotracer from both SPM and DS with a greater amount of uptake occurring from the DS. This uptake could be explained by the higher rates of radiotracer desorption from DS versus that of SPM. P. staminea accumulated a minimal amount of radiotracer from the SPM, most likely a consequence of high sediment loads in the test aquaria which precluded optimum filter-feeding conditions. However, significant amounts of radiotracer were taken up from DS compared to water only, indicating that sediments are an important source of contaminants to this obligate filter-feeder as well. Natural cadmium was primarily associated with the easily reducible components of suspended and deposited sediment as compared to the reducible and organic component of sediment. In contrast, the radioisotope partitioned primarily on the reducible component of both types of sediment. Hence, in these experiments the geochemical behaviour of the radiotracer did not parallel that which was observed in the natural environment. This lack of agreement between natural and experimental partitioning of cadmium among the three main sediment components of SPM and DS precluded our first objective of assessing the relative importance of SPM and DS as a source of contaminants to facultative filter-feeders. Further studies need to focus on ways to spike or label natural complex sediments, such that the partitioning achieved under laboratory conditions reflects that which is observed in the field. Acknowledgements Funding support to L. Bendell-Young in the form of an NSERC operating grant is gratefully acknowledged.
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