Accumulation and Soluble Binding of Cadmium, Copper, and Zinc in ...

Arch. Environ. Contam. Toxicol. 45, 468 – 478 (2003) DOI: 10.1007/s00244-003-0135-0

A R C H I V E S O F

Environmental Contamination a n d Toxicology © 2003 Springer-Verlag New York Inc.

Accumulation and Soluble Binding of Cadmium, Copper, and Zinc in the Polychaete Hediste diversicolor from Coastal Sites with Different Trace Metal Bioavailabilities B. Berthet,1,2 C. Mouneyrac,1,3 J. C. Amiard,1 C. Amiard-Triquet,1 Y. Berthelot,1 A. Le Hen,1 O. Mastain,1,3 P. S. Rainbow,4 B. D. Smith4 1

Centre National de la Recherche Scientifique (CNRS)—Groupement de Recherche 1117, Service d’ećotoxicologie, ISOMer, SMAB, 2 rue de la Houssinie`re, BP 92208, F-44322 Nantes Cedex 3, France 2 Institut Catholique d’E´tudes Supe´rieures (ICES), 17 Boulevard des Belges, 85000 La Roche-sur-Yon, France 3 Centre d’E´tude et de Recherche sur les Ećosyste`mes Aquatiques (CEREA), Universite´ Catholique de L’Ouest, 3 Place Andre´ Leroy, BP 808, 49008 Angers Cedex 1, France 4 Department of Zoology, The Natural History Museum, Cromwell Road, London SW7 5BD, United Kingdom

Received: 7 January 2003 / Accepted: 22 April 2003

Abstract. Bioaccumulation of cadmium, copper, and zinc was examined in common ragworms Hediste diversicolor from control (Bay of Somme, Blackwater) and metal-rich (Seine estuary, Boulogne harbor, Restronguet Creek) sites in France and the United Kingdom. The degree of exposure in the field was assessed by considering both total concentrations in superficial sediment and the quantities of metals which may be released in vitro at different pH levels. Among the three contaminated sites, release of the three metals was not detectable in Boulogne harbor, in correlation with limited enhancement of the metal concentrations in the common ragworms from this site. Even at those sites where zinc could be released in vitro from the sediment, zinc concentrations were not enhanced in common ragworms, in agreement with previous findings indicating that the body content of this metal is regulated in H. diversicolor. At all the studied sites, bioaccumulated zinc was mainly in cytosolic form. The distribution of cadmium and copper varied according to the origin of the common ragworms, the insoluble fraction increasing with the degree of contamination (cadmium in the Restronguet Creek, copper in the Seine estuary, and even more in Restronguet Creek). In the cytosolic fraction, metals were partly linked to cytosolic heatstable thiolic compounds (CHSTC) with molecular masses (5– 6 kDa and about 12 kDa) consistent with metallothioneinlike proteins (MTLP). Metal-binding to MTLP varied with the degree of contamination and with the metal studied. In contrast to many invertebrates, the presence of metal-binding CHSTC (MM about 2 kDa) other than MTLP seems to be a peculiar feature of H. diversicolor.

Correspondence to: J. C. Amiard; email: Jean-Claude.Amiard@ isomer.univ-nantes.fr

Estuaries are areas of high productivity— crucial in the life history of many fish, invertebrates, and birds—and the sustainability of estuarine biodiversity is vital to the ecological health of European coastal regions. European estuaries also represent habitats at risk, receiving toxic anthropogenic effluents from the great rivers of Europe, translocated from remote and nearby conurbations and industry. The sediment-dwelling worm Hediste diversicolor plays a key role in the fate of chemicals as a consequence of its bioaccumulation capacity and of its influence on metal speciation in sediment through bioturbation (François et al. 2002). It is also among the key species in estuarine and coastal ecosystems functioning as a major constituent of the benthic biomass of mudflats, as well as an important food item for crustaceans, fish, and waders (McLusky 1989). It is important, therefore, to determine whether ecologically keystone species of our estuaries are at risk from toxic contaminants, or whether different local populations are tolerant of availabilities of toxins potentially lethal elsewhere. There are records in the literature of metal-tolerant populations of invertebrates (Klerks and Weis 1987). The selection of a metal-tolerant population of the polychaete annelid H. diversicolor (as Nereis diversicolor), for example, has been demonstrated in Restronguet Creek, Cornwall, a metal-contaminated estuary (Bryan 1976a; Bryan and Gibbs 1983; Grant et al. 1989), and similar findings have been published for other annelids (Wallace et al. 1998). Tolerance may be achieved by one or more possible physiological mechanisms, for example reduction in metal uptake rates and/or enhancement of excretion rates, or storage of accumulated metals in nontoxic physicochemical forms— bound to the metal-binding protein metallothionein (MT) or incorporated into insoluble deposits or granules (Mason and Jenkins 1995). Metallothioneins, a family of low-molecular-weight, cysteine-rich metal-binding proteins have been shown to occur in most zoological taxa (Binz and Ka¨gi 1999). It is generally agreed that these proteins play a primary role in the homeosta-

Accumulation and Binding of Cd, Cu, Zn

sis of the essential metals copper (Cu) and zinc (Zn) as well as being involved in the detoxification of nonessential metals, such as cadmium (Cd) (Amiard and Cosson 1997). Metallothioneins or at least metallothionein-like-proteins (MLTP), are ubiquitous in the animal kingdom (Binz and Ka¨ gi 1999). Cdbinding MTs are present in the oligochaete Limnodrilus hoffmeisteri from a severely Cd-contaminated cove in the Hudson River, New York (Wallace et al. 1998), and in another oligochaete Lumbricus terrestris, Sturzenbaum et al. (1998) have described two isoforms of a protein showing the structural features of a Class I MT. These latter authors have demonstrated the inducibility of L. rubellus MT in specimens exposed to Cd-contaminated soils in the field or in experiments. However, in annelids, the presence of metal-binding proteins and their relationships with bioavailable metals in their environment do seem to differ greatly among the subtaxa (Dhainaut and Scaps 2001). Numerous metals—among them Cd, Cu, and Zn— can also be sequestered as electron-dense concretions, the occurrence, form, and function of which have been the subject of several reviews (Brown 1982; George 1982; Marigomez et al. 1995; Mason and Jenkins 1995). In invertebrates, these granules are present in all major phyla, including annelids and particularly H. diversicolor (Bryan 1976 a, b; Howard unpublished, in Brown 1982; Pirie et al. 1985; Fernandez and Jones 1989). When tolerant species have been able to adapt to highly metal-contaminated environments through the storage of metals in detoxified compartments, they have the potential to transfer these high accumulated concentrations along food chains, possibly with toxic effects at higher trophic levels. However, the distribution of metals in the prey species is critical for metal assimilation by the predator (Wang and Rainbow 2000 and literature cited therein). The bioavailability of metals to the predator depends on the nature and the chemical stability of the physicochemical forms of storage, and the desorption of a metal in the acidic gut as well as the metal gut passage time are critical for metal assimilation (Wang and Rainbow 2000 and literature cited therein). The storage of silver as sulphide in bivalves (Berthet et al. 1992) and of Zn in pyrophosphate granules in different invertebrates (Pullen and Rainbow 1991) all limit the risk of transfer to the consumer (Wallace and Lopez 1997). Nott and Nicolaidou (1990) have theorized such data, proposing the concept of “transfer of metal detoxication along marine food chains.” On the other hand, MT as a protein is degraded during digestion, leading most probably to an enhanced bioavailability of metals initially bound to this ligand (Wallace and Lopez 1997). In the deposit-feeding oligochaete annelid Limnodrilus hoffmeisteri experimentally exposed to Cd, Cd-resistant worms, from the Cd-contaminated cove in the Hudson River, New York, produced metallothionein-like proteins (MTLP) as well as metal-rich granules, whereas nonresistant worms from an adjacent unpolluted site produced only MT (Wallace et al. 1998). This difference in Cd subcellular distribution led to large differences in Cd bioavailability to the consumer, the omnivorous shrimp Palaemonetes pugio (Wallace et al. 1998). The absorption efficiency was in good agreement with the proportions of Cd bound to the worm’s most bioavailable fractions. The present study was designed to verify the existence of MTLP in H. diversicolor exposed to metals in their natural environments and to compare MTLP induction in common

469

ragworms from clean (Bay of Somme, France; Blackwater, UK) or metal-rich sites (Boulogne harbor and Seine estuary, France; Restronguet Creek, UK). The degree of exposure of common ragworms originating from these different sites will be determined not only by measuring the total concentration of metals in the sediment but also by modeling the potential bioavailability of sediment-bound Cd, Cu, and Zn (Kersten and Forstner 1989; Amiard 1992). The distribution of accumulated metals between soluble and insoluble components in H. diversicolor will also be examined with a view to assess their potential availability to the common ragworms’ predators.

Material and Methods Sediment and Biota Sampling Sediments and biota were sampled in winter (Bay of Somme, considered as control; Seine estuary, Boulogne harbor, France) and in spring (Blackwater estuary, considered as control; Restronguet Creek, UK) Superficial sediments (⬍1 cm deep) were scraped using a plastic blade and placed immediately in a plastic box, filled with sediment in order to eliminate air. Samples were then transported to the laboratory with icepacks in an isothermic container. Sediment to be used for the characterization of granulometry and organic matter content was stored at ⫺20°C until treatment. With a view to desorption tests on sediments from potentially impacted sites, sediments samples were stored in the dark, at 4°C and used within a week of collection. All desorption tests were carried out in triplicate. Each sediment sample (500 mg) was dispersed into one of a series of buffers at different pH levels (20 mL; prepared with 1% acetic acid and adjusted to appropriate pHs by adding suprapure ammonia). The contact time was 2 h at ambient temperature (19 –21°C). Then sediments were recovered by centrifugation (15,000g). Metal analyses were carried out on initial sediment samples as well as on sediment samples submitted to the different desorption tests. The percentage of desorption was calculated from the concentrations of metals in initial and treated sediments. Common ragworms were transported from the field to the laboratory in isothermic containers in wet seaweed or sediment from the site of origin without seawater. Common ragworms were then placed for 24 h in clean, aerated seawater from the site of origin to allow them to eliminate their gut contents, since it has been shown that the presence of sediment is responsible for an overvaluation of weight and metal content (Amiard-Triquet et al. 1984). Common ragworms were stored frozen at ⫺20°C until analysis.

Granulometry, Organic Matter Content Sediment was thawed, and a sample dried at 60°C to constant weight, then ground in a mortar. An aliquot fraction was then placed in a crucible, heated at 450°C overnight, allowed to return to room temperature, then placed at 950°C for ashing. At each stage, the sample was weighed to determine the percentage of organic matter lost at each temperature—the latter including the carbonated fraction—according to the formula: (Dry weight at 60°C) ⫺ (Dry weight after ashing) ⫻ 100 (Dry weight at 60°C) The remaining sediment was sieved through a 63-␮m mesh. The fine and coarse particles thus separated were dried at 60°C to constant weight. The granulometry was characterized using the percentage of total dry weight corresponding to particles ⬍63 ␮m.

470

Metal Compartmentation and Partial Isolation of MTLP Pooled common ragworms from each site (each about 2 g total weight) were homogenized with an Ultra-Turrax at 4°C in 20 mM TRIS, 150 mM NaCl solution adjusted to pH ⫽ 8.6 (4 mL/1 g soft tissue). In order to limit oxidation, ␤-mercaptoethanol (10⫺5 mM) was added, as well as Phenylmethanesulfonyl Fluoride (PMSF) as an anti-protease (0.1 mM). Soluble (S1) and insoluble (P1) fractions were separated by centrifugation (25,000g for 55 min at 4°C). The cytosolic heat-stable thiolic compounds (CHSTC), including MTLP (S2), were isolated by centrifugation of the soluble fraction (15,000g for 10 min at 4°C) after heat treatment (75°C for 15 min).

Gel Chromatography The cytosol before (S1) or after (S2) heat-denaturation was fractionated by gel chromatography using a Sepharose CL 6B (Pharmacia) column (920 ⫻ 16 mm) or a Sephadex G75SF (Pharmacia) column (640 ⫻ 16 mm) equilibrated with the buffer used for elution (20 mM TRIS, 150 mM NaCl, pH ⫽ 8.6). Two replicates of S1 or S2 were used for each type of chromatography. An aliquot of S1 or S2 was applied to the column, eluted with buffer (48 mL h⫺1 for CL6B, 18 mL h⫺1 for G75SF) and collected as 2.4-mL fractions. The Sepharose CL6B column was calibrated for molecular mass estimations using standard markers (ribonuclease 13.7 kDa, chymotrypsinogen 25 kDa, ferritin 440 kDa, thyroglobulin 669 kDa). The Sephadex G75SF column was calibrated using ribonuclease (13.7 kDa), chymotrypsinogen (25 kDa), ovalbumin (45 kDa), and bovine serum albumin (67 kDa). The molecular mass of the different metal-binding compounds was established using the calibration curve. Sulfhydryl groups were quantified using differential pulse polarography (DPP, see below) in the chromatography fractions obtained from the heat-denatured cytosol (S2), whereas determination of Cd, Cu, and Zn by atomic absorption spectrometry (AAS, see below) was carried out for each chromatography fraction obtained from the total cytosol (S1). This procedure has been adopted for it has been shown that heat treatment of the total cytosol can induce translocation of metals between compounds, particularly an enrichment of MTLP (Bragigand and Berthet 2003).

Polarographic Determination of Sulfhydryl Groups In chromatography fractions obtained from the heat-denatured cytosol (S2), the amount of MT was determined by differential pulse polarographic analysis (DPP), a technique based on-SH compound determination according to the Brdicka reaction (Brdicka 1933), as described by Thompson and Cosson (1984). A PAR Model 174 analyzer, a PAR/EG&G Model 303 static mercury drop electrode (SMDE), and an X-Y recorder (RE 0089) were used. The temperature of the cell was maintained at 5°C. The standard addition method was used for calibration with rabbit liver MT (Sigma Chemical Co., St. Louis, Missouri) in the absence of a marine invertebrate MT standard. The results were expressed in ␮g mL⫺1 of eluent.

Metal Analysis Nalgene bottles were used to store all reagents. All labware was soaked in 10% hydrochloric acid, rinsed three times with deionized water, and dried in a desiccator sheltered from atmospheric dust. Sediment samples were digested according to the recommendations of Charlou and Joanny (1983). In order to assess the individual variability of metal bioaccumulation in common ragworms, 20 specimens from a

B. Berthet et al.

reference (Bay of Somme) and a contaminated (Seine estuary) site were analyzed individually for total metal content in their body. These samples as well as the insoluble fractions (P1) and aliquots of soluble (S1) fractions obtained above were heated (75°C, 12 h) with suprapure concentrated nitric acid (Carlo Erba). After digestion, metal levels in these different acid solutions were determined after dilution with deionized water by flame (FAAS) (Cu, Zn) or electrothermal atomic absorption spectrophotometry (EAAS) (Cd) with the Zeeman effect (Hitachi Z 8200 spectrophotometer). Standard addition analyses were performed in an iso-medium and concentrations of each element were ⫹125, 250, 500 ng Cu and Zn. mL⫺1 for FAAS and ⫹ 0.25, 0.5, 1 ng Cd. mL⫺1 for EAAS. The results will be expressed in ␮g Cd, Cu, or Zn. g⫺1 dry weight in sediments, with the same units but additionally expressed as wet weight in common ragworms. In the latter, the total concentration was determined directly in individual common ragworms or by the addition of insoluble (P1) and soluble (S1) concentrations determined in pools of common ragworms. In fractions separated by gel chromatography, Zn levels were determined by flame AAS in the TRIS NaCl buffer after dilution in deionized water. Cd and Cu levels were determined by flameless AAS according to the standard addition mode in an isomedium (1:3 TRIS NaCl, 1:3 HNO3 suprapure, 1:3 deionized water). Results are presented as ng metal mL⫺1 eluent. The methodology for digestion and analysis of biological material has been described previously by Amiard et al. (1987a). The analytical methods have been validated by external intercalibrations in sediment (Coquery and Horvat 1996) and biological matrices (Campbell et al. 2000).

Statistical Treatment Due to the limited size of the experimental groups, nonparametric tests were preferred for the comparisons of changes in metal concentrations in the treated sediments (desorption tests) from those in original sediments. Therefore, Kruskal–Wallis tests were carried out using the statistical package Statgraphics plus 4.0. The means of SH and metal levels in the different chromatography fractions obtained from two replicates were calculated using the technique of running means over four periods using Excel.

Results Metals in Sediments Gross Concentrations in Sediments. The lowest metal concentrations were measured in the sediments from the Bay of Somme and the Blackwater estuary (Table 1). It must be noted that the strong difference between these relatively clean sites was mainly due to their physicochemical characteristics: a very low proportion of particles ⬍63 ␮m and a very low organic matter content in the sediment from the Bay of Somme. The sediment collected in the Seine estuary and Boulogne harbor had higher metal contents but the maximum contamination was found in the sediment collected in Restronguet Creek. Although there are intersite differences in the percentage fraction of small particle size and in organic matter content, these were not sufficient to explain the observed intersite differences in metal contamination (Table 1). Desorption Tests. Table 2 shows the percentage of release of Cd, Cu, and Zn from sediments collected in the three contaminated sites, namely Seine estuary, Boulogne harbor, and Re-


471

Table 1. Particle size, organic matter, and metal concentrations in sediments from clean (Bay of Somme, Blackwater) and metal-rich (Seine, Boulogne, Restronguet) sites

Fraction ⬍63 ␮m (%) Organic matter at 450°C (%) Organic matter at 900°C (%) Cd ␮g.g⫺1 d.w. Cu ␮g.g⫺1 d.w. Zn ␮g.g⫺1 d.w.

Bay of Somme

Blackwater

Seine Estuary

Boulogne Harbor

Restronguet Creek

1.3 0.9 7.7 0.05 (0.00) 0.64 (0.11) 7.3 (0.2)

97.1 13.1 18.2 0.19 (0.01) 53 (16) 141 (4)

86 6 14 1.13 (0.05) 48 (7) 219(2)

76.9 4.8 13.4 0.38 (0.01) 153 (10) 1092 (87)

94.2 8.5 11.1 2.76 (0.20) 4413 (262) 3650 (174)

Table 2. Metal release (%) from sediments sampled at metal-rich sites under different conditions of pH

Cd

Cu

Zn

Seine Boulogne Restronguet Seine Boulogne Restronguet Seine Boulogne Restronguet

pH6

pH5

pH4

pH3

pH2

19 ND 14 4 ND ND ND ND 6*

16 ND 24* 8 ND ND ND ND 12*

30 ND 32* 12 ND ND 9 ND 25*

33 ND 50* 17* ND 6 16 2 40*

35 ND 47* 21* ND 29* 14 35 52*

* Metal concentration in the treated sediment significantly lower than in original sediment, using Kruskal-Wallis tests. ND: not detectable.

stronguet Creek. There was no significant release of any of the three metals from the Boulogne sediment at any of the pHs tested, and in most cases, no difference at all was detectable between concentrations in initial and treated sediment samples. On the other hand, all three metals were released from Restronguet sediment at at least one of the pHs used. Zn was significantly released at all pHs, Cd at all but the highest pH, and Cu at the lowest pH only. In the Seine, an intermediate situation was observed, with detectable release of metals, but the differences between concentrations in treated and initial sediments were not often significant.

Metal in Hediste diversicolor Metal Distributions Among Cytosolic Compounds. The concentrations of SH compounds were determined in fractions separated by gel chromatography from the heat-denaturated cytosol (S2) for two replicate pools of common ragworms. To investigate the possible presence of SH-rich high molecular mass compounds, we used a large molecular mass (MM) determination range gel (Sepharose CL 6B, 10 – 4.103 kDa) for the chromatography of the cytosol of common ragworms from one site (Boulogne). Except for a small peak observed at 150 kDa in one replicate (5– 6% of total SH compounds present in S2), no heat-stable SH compounds were detected in high MM fractions (⬎50 kDa). The SH compounds were distributed in fractions corresponding to MM ⬍20 kDa (not shown). In order to obtain better precision over this range of MM, the heat-denatured cytosol (S2) was applied to a Sephadex G75SF chromatography column. The means of running means obtained from two replicates of SH-levels in the different fractions are shown in Figure 1 for four different sites. At each site,

an important thiol peak was observed at about 12 kDa. The maximum SH level was registered in common ragworms from the Blackwater; this level being lower and similar in both Seine and Boulogne specimens, and the lowest value being observed in common ragworms from Restronguet. In common ragworms from the Blackwater and Boulogne, a second SH peak, even higher than the first, was observed at about 5– 6 kDa. In specimens from the Seine, this peak appeared only as a shoulder, with a lower value than in both of the previous sites, whereas it was not clearly shown in common ragworms from Restronguet. A last peak was shown at about 2 kDa, the highest SH levels being determined in common ragworms from the Blackwater, followed by those from Boulogne, then Seine. This peak was practically absent in specimens from Restronguet. Cd concentrations in fractions from chromatography of total cytosol S1 on Sephadex G75SF are also shown in Figure 1 for the different sites (running means obtained from two replicates). In all cases, the most important Cd peak was determined in the void volume whereas, in the total bed volume (small MM region), quantities of this metal were negligible. Between these limits, Cd was mainly present in the fractions corresponding to the SH-peak at 5 kDa for the Blackwater common ragworms, whereas it was more abundant in the fractions corresponding to the SH-peak at 12 kDa in specimens from the Seine and Restronguet. Cd was hardly detectable in these fractions obtained from common ragworms originating from Boulogne. Cd was also associated with the SH-peak corresponding to molecular masses about 2 kDa. Intersite differences were also observed in the amounts of Cd associated with these SH-peaks; in common ragworms from both Blackwater and Restronguet, the highest concentrations were within the same order of magnitude (respectively 0.5 and 0.3 ng mL⫺1), whereas in specimens

472

B. Berthet et al.

Fig. 1. Distribution of Cd (in gel permeation fractions of the total cytosol S1) and SH compounds (in fractions of the heat-stable cytosol S2) extracted from Hediste diversicolor as a function of the molecular mass. Cd, in continuous line; SH, in dotted line. Vo ⫽ void volume. Vt ⫽ total bed volume. 10 indicates the expected elution position of proteins of 10 kDa. (A) Blackwater, (B) Restronguet Creek, (C) Boulogne harbor, (D) Seine estuary

from the Seine, a maximum concentration of 9.5 ng mL⫺1 was reached. Zn concentrations (running means obtained from two replicates) in the chromatography fractions are shown in Figure 2. In common ragworms from all of the four studied sites, Zn was mainly present in the void volume. Important quantities were also present in the total bed volume, whereas moderate amounts of Zn were linked to SH-compounds with molecular masses about 12 kDa. Intersite differences were observed in the amounts of Zn present in this peak, with a maximum at 100 ng mL⫺1 in fractions from common ragworms from Restronguet, compared to 50 –70 ng mL⫺1 in common ragworms from other sites. In the case of other SH-peaks, the pattern of Zn binding found for specimens from Blackwater differed from the patterns at other sites, with the presence of more Zn in fractions corresponding to the SH-peak at about 5– 6 kDa, and a supplementary peak of Zn associated with the SH-peak corresponding to molecular masses about 2 kDa. Cu was abundant in the void volume and was present in most of the chromatography fractions, whereas it was in low concentration or even absent in the total bed volume. The elution profiles for Cu (not shown) did not show clearly distinguished peaks, but slightly raised Cu concentrations were observed in

the fractions corresponding to the SH-peaks with molecular mass of 6 –12 kDa. Metal binding to CHSTC (cytosolic heat-stable thiolic compounds) with molecular masses of 5– 6 kDa and about 12 kDa is in agreement with the presence of MTLPs (metallothioneinlike proteins). In specimens from both control and metal-rich sites, the percentages of soluble Cu and Zn present in MTLP fractions were relatively homogeneous (respectively 25– 40% and 14 –21%) (Table 3). In the case of Cd, the low percentage (9%) bound to MTLP in common ragworms from Boulogne contrasted with higher percentages (25– 47%) in specimens from all other sites (Table 3). Metals in the Insoluble Fractions of Tissues. From metal concentrations determined in the insoluble pellets recovered after compartmentation and in the cytosol S1, it was possible to calculate the distribution of accumulated metals into soluble and insoluble components in common ragworms from the different sites. The results are shown as percentages in Figure 3. In specimens from the four sites studied, Zn was predominantly present in soluble form. The same was true for Cd in common ragworms from the reference site (Blackwater) and in common ragworms with intermediate accumulated concentra-


473

Fig. 2. Distribution of Zn in gel permeation fractions of the total cytosol S1 extracted from Hediste diversicolor as a function of molecular mass of their ligands. Dotted line, Blackwater; black continuous line, Seine estuary; broken line, Boulogne harbor; grey continuous line, Restronguet Creek. Other details as for Figure 1

Table 3. Relative importance of MTLPs in metal storage compared to other cytosolic ligands or to total incorporated metals in common ragworms Hediste diversicolor

Percentage of cytosolic metals bound to MTLPs Blackwater Seine Boulogne Restronguet Percentage of total metals bound to MTLPs Blackwater Seine Boulogne Restronguet

Cd

Cu

Zn

31 47 9 47

25 39 40 35

19 21 14 19

22 35 7 13

17 19 29 3

13 15 12 15

tions of this metal (Seine and Boulogne). On the other hand, in common ragworms from metal-contaminated Restronguet, Cd was equally distributed between soluble and insoluble forms. In the case of Cu, the percentage of body metal in insoluble form showed an increasing gradient from common ragworms with low accumulated concentrations of total Cu (reference site Blackwater and Boulogne), to moderately (Seine) then highly contaminated (Restronguet) samples (Figure 3). From these data as well as from the metal quantities present in MTLP fractions, it is possible to calculate the percentage of total metals (soluble S1 ⫹ insoluble P1) which were bound to

MTLP (Table 3). For Zn, this percentage was very similar in common ragworms from the different sites (12–15%). For Cd, a large range of percentages of MTLP-bound Cd was observed according to site (Table 3): the lowest percentage of total Cd bound to MTLPs was observed in specimens from Boulogne, in agreement with the low percentage of cytosolic Cd bound to MTLPs in these worms (Table 3, top). In the case of Cu, a striking figure was the low percentage of total Cu bound to MTLP in Restronguet common ragworms (Table 3) as a consequence of storage mainly in insoluble form (Figure 3).

Enhancement of Metal Levels in Sediments and in Hediste diversicolor Metal concentrations in the common ragworms are shown in Table 4. Since season may be a major source of variation in bioaccumulated metal concentrations in coastal invertebrates, caution needs to be exercised when comparing the results obtained for the French sites sampled in January and for the British sites sampled in spring. The ratios between metal concentrations in specimens from metal-enriched versus reference sites (Table 5) have been calculated from the data of Table 4. Concerning the French sites (all collected in January), the ratios between Seine or Boulogne versus Somme reached respectively 1.6 and 0.8 for Zn, 2.3 and 1.3 for Cu, 6.6 and 1.1 for Cd. In the case of the pair of British sites (collected in spring), this ratio reached 1.1 for Zn, 2.8 for Cd, but was as high as 89 for Cu (Table 5). The ratio between metal concentrations in contaminated versus control sites has been also calculated for sediments

474

B. Berthet et al.

Fig. 3. Metal distribution between cytosolic (half tone) and insoluble (full tone) forms in common ragworms Hediste diversicolor originating from sites in France and the UK

(Table 5). Even when the sediment is strongly metal-enriched, the concentrations in common ragworms did not increase by the same order of magnitude. The contrast is particularly striking for Cu and Zn in Boulogne, with ratios to the clean site (the Somme) as high as 239 and 150, respectively, in sediments and only 1.3 and 0.8 in common ragworms. On the other hand, common ragworms collected in spring from Restronguet are Cu-enriched in the same proportion as the sediment (Table 5), whereas Zn concentrations were poorly enhanced in the common ragworms compared to Zn-enrichment in the sediment.

For Cd, enhanced concentrations were observed in common ragworms from Restronguet and the Seine, but not in those collected from Boulogne.

Discussion The metal concentrations determined in the sediments from the five sites studied agree with the designation of the Bay of


475

Table 4. Intersite comparison of metal concentrations (␮g. g⫺1) in Hediste diversicolor

Cd Cu Zn

w.w. d.w.* w.w. d.w.* w.w. d.w.*

Bay of Somme

Blackwater

Seine

Boulogne

Restronguet

0.008 (0.004) 0.049 (0.025) 1.8 (0.7) 11 (4) 17 (4) 105 (23)

0.036 0.221 7.2 44 25 152

0.053 (0.015) 0.325 (0.092) 4.2 (1.8) 26 (11) 27 (6) 166 (37)

0.009 0.055 2.4 15 14 88

0.1 0.6 642 3940 28 173

* On an average, the dry weight represents 16.3% (SD ⫽ 1.7) of the wet weight (calculated from 10 pools of 30 – 60 common ragworms each). Means and standard deviations. Bay of Somme and Seine estuary: analyses carried out in individual common ragworms (n ⫽ 20). Other sites: pooled samples.

Table 5. Ratios between metal concentrations in sediments and common ragworms Hediste diversicolor from metal-enriched (Seine, Boulogne, Restronguet) versus reference (Bay of Somme, Blackwater) sites

Seine vs. Somme Sediment Common ragworms Boulogne vs. Somme Sediment Common ragworms Restronguet vs. Blackwater Sediment Common ragworms

Cd

Cu

Zn

22.6 6.6

75 2.3

30 1.6

7.6 1.1

239 1.3

150 0.8

14 2.8

83 89

26 1.1

Somme and the Blackwater estuary as control sites, with a gradient of contamination for the other sites: Seine estuary ⬍ Boulogne harbor ⬍⬍ Restronguet Creek. Despite the raised sediment Zn concentrations, common ragworms collected in the three metal-rich sites showed no consistent enhancement of body Zn concentrations. For Cu, enhanced concentrations were observed in the Seine estuary, but this enhancement was weak compared to the bioaccumulation of this metal in common ragworms from Restronguet Creek; no additional Cu was derived from sediment by common ragworms collected in Boulogne harbor despite the fact that this sediment was the most Cu-contaminated. For Cd, enhanced concentrations were observed in common ragworms from Restronguet Creek and the Seine estuary and not in those collected from Boulogne harbor. In the case of Zn, these observations may be attributed to the ability of certain marine organisms including Hediste diversicolor (Bryan and Hummerstone 1973a; Bryan 1976 a, b; Amiard et al. 1987b) to regulate their Zn body burden to a relatively constant level when exposed in the field or in the laboratory to this element. On the contrary, it is known that Cu and Cd are accumulated by H. diversicolor and other annelids in proportion to their concentrations in the surrounding medium (Bryan 1976a, b; Bryan et al. 1980; Amiard et al. 1987b). The enhanced Cu accumulation of Restronguet common ragworms in comparison to Blackwater common ragworms can be explained by the atypically high rate of accumulation of Cu in the Cu-tolerant Restronguet H. diversicolor, as recognized by Bryan and Hummerstone (1971). Accumulated trace metal concentrations in H. diversicolor

are good biomonitoring measures of the local bioavailabilities of Cd, Cu, and Ag (Bryan and Hummerstone 1971, 1973a, b; Bryan et al. 1980; Bryan and Gibbs 1983, 1987). The bioavailability of sediment-bound metal to H. diversicolor is not necessarily a simple direct relationship to sediment total metal concentration for metal bioavailabilities may be affected by other characteristics of the sediment that affect metal speciation. Luoma and Bryan (1982) have shown that the most important factors controling Cd and Cu accumulation in the polychaete H. diversicolor are the concentrations of these metals in surface sediments but also the partitioning of sedimentbound metals between different sediment constituents. The results of the desorption tests carried out in the present study are in agreement with these findings. Accumulated concentrations of Cd and Cu were considerably higher in Restronguet and Seine common ragworms (over control sites, Blackwater and Somme) where the remobilization of metals from the sediment was maximum. When in vitro remobilization of sediment metal was practically undetectable (Boulogne), bioaccumulation by common ragworms in the corresponding field site was comparatively negligible. Table 1 shows that the mean Cu concentration measured in our sediment sample from Restronguet Creek was 4413 ␮g g⫺1. Cu concentrations in sediment samples collected by Bryan and Hummerstone (1971) from the same region of Restronguet Creek as in this study were similarly of the order of 4000 ␮g g⫺1, indicating that the Restronguet Creek sediments continue to hold a large reservoir of Cu. Cu concentrations in H. diversicolor from these sediments reached more than 2500 ␮g g⫺1 dry weight (Bryan and Hummerstone 1971; Bryan and Gibbs 1983), comparable again to the concentrations measured in Restronguet common ragworms in this study (Table 4). It would therefore appear that there is still a large supply of bioavailable Cu in the sediment of Restronguet Creek, although mining activity has ceased, and that H. diversicolor accumulates considerable amounts of Cu from this source. In this study, we have established the presence of cytosolic, heatstable, thiolic compounds (CHSTC), binding Cd, Cu, and Zn with molecular masses of 5– 6 and about 12 kDa which are consistent with the characteristics of a monomer and a dimer of MTLP (Langston et al. 1998). This finding is in line with the fact that MT is generally recognized to be ubiquitous in living organisms (Cosson and Amiard 2000), and with the detection of MLTP in the related polychaete Eurythoe complanata (Marcano et al. 1996). CHSTC were negligible among the high molecular mass compounds isolated, whereas CHSTC with MM about 2 kDa were

476

present to different degrees according to the geographical origin of the common ragworms. In metal-tolerant H. diversicolor from Restronguet Creek, Pirie et al. (1985) considered that 16% of Cu and 31% of Zn were present in the cytosolic fraction, associated with low molecular weight components, but not with MT. Following experimental exposure of H. diversicolor to dissolved Cd, two ligands have been shown to be involved in Cd-binding: MPI with a high molecular mass, identified as an extracellular hemoglobin, and MPII with a low molecular mass, the structure of which has a strong identity with myohaemerythrin (Demuynck and Dhainaut-Courtois 1993). These authors (Demuynck and Dhainaut-Courtois 1994) did not recognize the presence of MT in these experimentally exposed common ragworms but it must be noted that the experimental doses (80 mg Cd.L⫺1) are most probably high enough (the CL50 96 h was 10 mg Cd.L⫺1, Eisler 1971) to induce toxicity interfering with the general metabolism of proteins, including metallothionein. Such an effect has been described earlier in different species exposed in the field or in the laboratory to high metal doses (George et al. 1992; Baudrimont et al. 1999; Barka et al. 2001). The MTLP compounds bound Cd, Cu, and Zn in the common ragworms from the different sites, even though the low levels of Cd in Boulogne common ragworms did not allow us to obtain a clear pattern of its cytosolic distribution. In Blackwater specimens, more Cd was associated with the compounds at 5– 6 kDa, whereas in Restronguet and Seine common ragworms, Cd binding to compounds at 12 kDa was the most important. Zn was bound to both monomeric and dimeric forms of MTLP in common ragworms from the Blackwater, whereas in specimens from all other sites, Zn was predominantly bound to the dimeric form. The variability of the relationship between metals and binding ligands in annelid subtaxa has been already underlined (Dhainaut and Scaps 2001). Mason and Jenkins (1991) found that Cd in Cd-exposed specimens of the polychaete Neanthes arenaceodentata was largely associated with high MM proteins (50 to ⬎200 kDa), with limited quantities associated with an MTLP pool (10 –20 kDa). Eriksen et al. (1990) observed differences in the relationships between biological macromolecules and metals according to the feeding habits of four marine polychaetes. These authors have shown that Cd and Cu were associated with proteins of 10 –20 kDa, in the detritivore Chaetozone setosa and the carnivorous Goniada maculate. In both species, these proteins had a high affinity for 109 Cd added in vitro. In two sediment-feeding polychaetes Pectinaria belgica and Orbinia norvegica, Cu and Zn were associated with components of either high or very low molecular mass (Eriksen et al. 1990). Quantitatively, MTLPs represent a relatively important store for Cd, Cu, and Zn in the worm Hediste diversicolor, as a proportion either of other cytosolic ligands or of total incorporated metals, with two exceptions: Cd in individuals from Boulogne harbor, the total body burden of which was the lowest among studied species (likely limiting the need for detoxification); Cu in individuals from Restronguet Creek, the total Cu body burden of which was the highest among specimens studied and in this case, Cu was mainly bound to insoluble constituents. These findings suggest that MTLPs contribute to metal detoxification in most cases. Even in the case of common ragworms highly contaminated with Cu, it remains likely that MTLPs intervene in detoxification since in other

B. Berthet et al.

invertebrates, it has been proposed that after the degradation of metallothioneins, the associated metals become stored in lysosomal vesicles or granules (Viarengo et al. 1985; Lauverjat et al. 1989). According to several authors, the different modes of detoxification in metal-tolerant prey species control trophic transfer to predators (Nott and Nicolaidou 1990; Wallace et al. 1998). Metals bound to soluble ligands such as MTLPs are easily available for predators, whereas insolubilisation reduces metal availability as demonstrated for Cd (Wallace and Lopez 1997). In the present study, Cd storage in insoluble form was clearly increased in specimens from the most highly contaminated site (Restronguet Creek). For Cu, soluble forms were predominant in common ragworms from the Blackwater estuary (a control site) or from Boulogne harbor (low metal bioavailability). Cu was equally distributed among soluble and insoluble forms in the moderately contaminated Seine estuary common ragworms, whereas less than 10% of Cu was present in the cytosol of specimens from the heavily Cu-contaminated Restronguet Creek. However, the insoluble fraction of the tissues does not only contain detoxification granules but also cellular debris such as cytoplasmic and nuclear membranes. The literature contains brief references to the presence of Cu- and Zn-rich inclusions in the cell of H. diversicolor (Bryan and Hummerstone 1971; Brown 1982; Bryan and Gibbs 1983; Pirie et al. 1985). Thus it is important to develop investigations about the ultrastructural nature of insoluble metal-rich deposits in common ragworms exposed to a range of metal contaminants (Mouneyrac et al. 2003).

Acknowledgments. Thanks are due to Jean-Claude Dauvin, Michel Desprez, and Robert Lafite for their helpful contribution to the choice of relevant clean and metal-rich sites. This research was supported in part by a grant from the CNRS France (PICS 1056).

References Amiard JC, Cosson RP (1997) Les me´ tallothione´ ines. In: Lagadic L, Caquet T, Amiard JC, Ramade F (eds) Biomarqueurs en e´ cotoxicologie—Aspects fondamentaux. Masson, Paris, pp 53– 66 Amiard JC (1992) Bioavailability of sediment-bound metals for benthic aquatic organisms. In: Vernet JP (ed) Impact of heavy metals on the environment. Elsevier, Amsterdam, pp 183–202 Amiard JC, Pineau A, Boiteau HL, Me´ tayer C, Amiard-Triquet C (1987a) Application of atomic absorption spectrophotometry using Zeeman effect to the determination of eight trace elements (Ag, Cd, Cr, Cu, Mn, Ni, Pb, and Se) in biological materials. Water Res 21: 693– 697 Amiard JC, Amiard-Triquet C, Berthet B, Me´ tayer C (1987b) Comparative study of the patterns of bioaccumulation of essential (Cu, Zn) and non-essential (Cd, Pb) trace metals in various estuarine and coastal organisms. J Exp Mar Biol Ecol 106:73– 89 Amiard-Triquet C, Me´ tayer C, Amiard JC (1984) Technical recommendations for studying the biogeochemical cycle of trace metals. Rev Intern Oce´ anogr Me´ d 73–74: 27–34 Barka S, Pavillon JF, Amiard JC (2001) Influence of different essential and non-essential metals on MTLP levels in the copepod Tigriopus brevicornis. Comp Biochem Physiol 128C:479 – 493 Baudrimont M, Andre`s S, Me´ tivaud J, Lapaquellerie Y, Ribeyre F, Maillet N, et al. (1999) Field transplantation of the freshwater


bivalves Corbicula fluminea along a polymetallic contamination gradient (river Lot, France). II Metallothionein response to metal exposure. Environ Toxicol Chem 18:2472–2477 Binz PA, Ka¨ gi JHR (1999) Metallothionein: Molecular evolution and classification. In: Klaassen CD (ed) Metallothionein IV. Birkha¨ user Verlag, Basel, pp 7–13 Berthet B, Amiard JC, Amiard-Triquet C, Martoja M, Jeantet AY, (1992) Bioaccumulation, toxicity and physico-chemical speciation of silver in bivalve molluscs: ecotoxicological and health consequences. Sci Total Environ 125:97–122 Bragigand V, Berthet B (2003) Some methodological aspects of metallothionein evaluation. Comp Biochem Physiol 134A:55– 61 Brdicka A (1933) Polarographic studies with the dropping mercury method. A new test for proteins in the presence of cobalt salts in ammoniacal solution of ammonium chloride. Collect Czech Chem Commun 5:112–128 Brown BE (1982) The form and function of metal-containing “granules” in invertebrate tissues. Biol Rev 57:621– 667 Bryan GW (1976a) Some aspects of heavy metal tolerance in aquatic organisms. In: Lockwood APM (ed) Effects of pollutants on aquatic organisms. Cambridge University Press, Cambridge, pp 7–34 Bryan GW (1976b) Heavy metal contamination in the sea. In: Johnston R (ed) Marine pollution. Academic Press, London, pp 185–302 Bryan GW, Gibbs PE (1983) Heavy metals in the Fal Estuary, Cornwall: A study of long-term contamination by mining waste and its effects on estuarine organisms. Occ Publ Mar Biol Assoc UK 2:1–112 Bryan GW, Gibbs PE (1987) Polychaete common ragworms as indicators of heavy-metal availability in marine deposits. In: Capuzzo JM, Kester DR (eds) Biological processes and wastes in the ocean Robert E. Krieger Publishing Co., Malabar, Florida, pp 37– 49 Bryan GW, Hummerstone LG (1971) Adaptation of the polychaete Nereis diversicolor to estuarine sediments containing high concentrations of heavy metals. I. General observations and adaptation to copper. J Mar Biol Assoc. UK 51:845– 863 Bryan GW, Hummerstone LG (1973a) Adaptation of the polychaete Nereis diversicolor to estuarine sediments containing high concentrations of zinc and cadmium. J Mar Biol Assoc UK 53:839 – 857 Bryan GW, Hummerstone LG (1973b) Adaptation of the polychaete Nereis diversicolor to manganese in estuarine sediments. J Mar Biol Assoc UK 53:859 – 872 Bryan GW, Langston WJ, Hummerstone LG (1980) The use of biological indicators of heavy metal contamination in estuaries. Mar Biol Assoc UK, Occ. Publ., N°1, 73 pp Campbell MJ, Radecki Z, Trinkl A, Burns KI (2000) Report on the intercomparison runs for the determination of trace and minor elements in cabbage material. IAEA-359. Rep. IAEA/AL/123, Vienna Charlou JL, Joanny M (1983) Dosage du mercure et d’autres me´ taux (Pb, Zn, Cu, Cd, Co, Mn) dans les se´ diments marins par absorption atomique. In: Amino A, Chaussepied M (eds) Manuel des analyses chimiques en milieu marin. CNEXO, Brest, pp 285–295 Coquery M, Horvat M (1996) The analytical performance study for the MED POL area: Determination of trace-elements in marine sediment SD-MEDPOL-1/TM and fish homogenate MA-MEDPOL1/TM. Rep. IAEA, Monaco Cosson RP, Amiard JC (2000) Use of Metallothioneins as biomarkers of exposure to metals. In: Lagadic L, Caquet T, Amiard JC, Ramade F (eds) Use of biomarkers for environmental quality assessment. Science Publishers Inc, Enfield, pp 79 –111 Demuynck S, Dhainaut-Courtois N (1993) Identification of extracellular haemoglobin as the major high molecular weight cadmiumbinding protein of the polychaete annelid Nereis diversicolor. Comp Biochem Physiol 106C:467– 472 Demuynck S, Dhainaut-Courtois N (1994) Metal-binding protein pat-

477

tern in the polychaete worm Nereis diversicolor during short-term acute cadmium stress. Comp Biochem Physiol 108C:59 – 64 Dhainaut A, Scaps P (2001) Immune defense and biological responses in Annelida. Can J Zool 79:233–253 Eisler L (1971) Cadmium poisoning in Fundulus heteroclitus and other marine organisms. J Fish Res Bd Canada 38:1225–1234. Eriksen KDH, Andersen T, Stenersen J, Andersen RA (1990) Cytosolic binding of Cd, Cu, Zn and Ni in four polychaete species. Comp Biochem Physiol 95C:111–115 Fernandez TV, Jones NV (1989) The distribution of zinc in the body of Nereis diversicolor. Trop Ecol 30:285–293 Franç ois F, Ge´ rino M, Stora G, Durbec JP, Poggiale JC (2002) Functional approach to sediment reworking by gallery-forming macrobenthic organisms: Modelling and application with the polychaete Nereis diversicolor. Mar Ecol Progr Ser 229:127–136 George SG (1982) Subcellular accumulation and detoxification of metals in aquatic animals. In: Vernberg WB, Calabrese A, Thurnberg FP, Vernberg FJ (eds) Physiological mechanisms of marine pollutant toxicity. Academic Press, New York, pp 3–55 George SG, Burgess D, Leaver M, Frerichs N (1992) Metallothionein induction in cultured fibroblasts and liver of a marine flatfish, the turbot Scopthalmus maximus Fish Physiol Biochem 10:43–54 Grant A, Hateley JG, Jones NV (1989) Mapping the ecological impact of heavy metals on the estuarine polychaete Nereis diversicolor using inherited metal tolerance. Mar Pollut Bull 20:235–238 Kersten M, Forstner U (1989) Speciation of trace elements in sediments. In: Batley GE (ed) Trace element speciation: Analytical methods and problems. CRC Press, Boca Raton, pp 245–317 Klerks PL, Weis JS (1987) Genetic adaptation to heavy metals in aquatic organisms. A review. Environ Pollut 45:173–205 Langston WJ, Bebianno MJ, Burt GR (1998) Metal handling strategies in molluscs. In: Langston WJ, Bebianno MJ (eds) Metal metabolism in aquatic environments. Chapman and Hall, London, pp 219 –283 Lauverjat S, Ballan-Dufranç ais C, Wegnez M (1989) Detoxification of cadmium. Ultrastructural study and electron-probe microanalysis of the midgut in a cadmium-resistant strain of Drosophila melanogaster. Biol Metals 2:97–107 Luoma SN, Bryan GW (1982) The statistical study of environmental factors controlling concentrations of heavy metals in the burrowing bivalve Scrobicularia plana and the polychaete Nereis diversicolor. Estuar Coast Shelf Sci 15:95–108 Marcano L, Nusetti O, Rodriguez-Grau J, Vilas J (1996) Uptake and depuration of copper and zinc in relation to metal-binding protein in the polychaete Eurythoe complanata. Comp Biochem Physiol 114C:179 –184 McLusky DS (1989) Primary consumers. In: The estuarine ecosystem. Blackie Academic and Professional, London, pp 77–112 Marigomez I, Soto M, Cajaraville MP (1995) Morphofunctional patterns of cell and tissue systems involved in metal handling and metabolism. In: Cajaraville MP (ed) Cell biology in environmental toxicology. University of the Basque Country Press Service, Bilbao, pp 89 –134 Mason AZ, Jenkins KD (1991) Effects of cadmium bioavailability on the cytoplasmic distribution of cadmium in Neanthes arenaceodentata. Bull Mar Sci 48:524 –529 Mason AZ, Jenkins KD (1995) Metal detoxication in aquatic organisms. In: Tessier A, Turner DR (eds) Metal speciation and bioavailability in aquatic systems. Wiley, Chichester, pp 479 – 608 Mouneyrac C, Mastain O, Amiard JC, Amiard-Triquet C, Beaunier P, Jeantet AY, et al. (2003) Trace metal detoxication and the tolerance of the estuarine worm Hediste diversicolor chronically exposed in their environment. Mar Biol (in press) Nott JA, Nicolaidou A (1990) Transfer of metal detoxication along marine food chains. J Mar Biol Assoc UK 70:905–912 Pirie B, Liu Fayi, George S (1985) Ultrastructural localisation of

478

copper and zinc in the Polychaete, Nereis diversicolor, from a highly contaminated estuary. Mar Environ Res 17:197–198 Pullen JSH, Rainbow PS (1991) The composition of pyrophosphate heavy metal detoxification granules in barnacles. J Exp Mar Biol Ecol 150:249 –266 Sturzenbaum SR, Kille P, Morgan AJ (1998) The identification, cloning and characterization of earthworm metallothionein. FEBS Letters 431:437– 442 Thompson JAJ, Cosson RP (1984) An improved electrochemical method for the quantification of metallothionein in marine organisms. Mar Environ Res 11:137–152 Thompson M, Wood R (1993) The international harmonized protocol for the proficiency testing of (chemical) analytical laboratories. IUPAC/ISO/AOAC. J Pure Appl Chem 65:2123–2144

B. Berthet et al.

Viarengo A, Moore MN, Pertica M, Mancinelli G, Zanicchi G., Pipe RK (1985) Detoxification of copper in the cells of the digestive gland of mussel: The role of lysosomes and thioneins. Sci Total Environ 44:135–145 Wallace WG, Lopez GR (1997) Bioavailability of biologically sequestered cadmium and the implications of metal detoxification. Mar Ecol Prog Ser 147:149 –157 Wallace WG, Lopez GR, Levinton JS (1998) Cadmium resistance in an oligochaete and its effect on cadmium trophic transfer to an omnivorous shrimp. Mar Ecol Prog Ser 172:225– 237 Wang WX, Rainbow PS (2000) Dietary uptake of Cd, Cr, and Zn in the barnacle Balanus trigonus: Influence of diet composition. Mar Ecol Prog Ser 204:159 –168