Interaction of Cd and Zn during uptake and loss in the polychaete Capitella capitata: Whole body and subcellular perspectives

Journal of Experimental Marine Biology and Ecology 352 (2007) 65 – 77 www.elsevier.com/locate/jembe

Interaction of Cd and Zn during uptake and loss in the polychaete Capitella capitata: Whole body and subcellular perspectives Daisuke Goto a , William G. Wallace a,b,⁎ a

b

Program in Biology, Ecology, Evolutionary Biology and Behavior, Graduate School and University Center, City University of New York, 365 Fifth Avenue, New York, New York 10016, USA Biology Department, College of Staten Island, 6S-310, City University of New York, 2800 Victory Boulevard, Staten Island, New York 10314, USA Received 6 March 2006; received in revised form 19 June 2007; accepted 1 July 2007

Abstract The interaction between Cd and Zn in aquatic organisms is known to be highly variable. The purpose of this study was to use a subcellular compartmentalization approach to examine Cd and Zn interactions in the deposit-feeding polychaete Capitella capitata (sp. I). Laboratory-reared C. capitata were co-exposed to Cd (background or 50 μg Cd l− 1) and Zn (background or 86 μg Zn l− 1) with 109Cd and 65Zn as radiotracers for 1 week. After the 1-week uptake period, subsets of worms were allowed to depurate accumulated metals for an additional 1 week. Worms from both phases (uptake and loss) were then subjected to subcellular fractionation to determine the compartmentalization of metals as metal-sensitive fractions [MSF — organelles and heat-denaturable proteins (HDP)] and biologically detoxified metals [BDM — heat-stable proteins (HSP) and metal-rich granules (MRG)]. Uptake and loss of Cd and Zn in C. capitata at the whole body level were similar at bkgd-Cd/bkgd-Zn, with worms depurating the majority of accumulated metal (∼ 75% Cd and ∼ 64% Zn). When exposure of Zn or Cd was increased (bkgd-Cd/86-Zn; bkgd-Zn/50-Cd), uptake of background levels of Cd or Zn, respectively, was suppressed by ∼ 50%. These accumulated metals, however, were retained during the loss phase resulting in ∼40–50% greater Cd and Zn whole body tissue burdens than those of bkgd-Cd/bkgd-Zn worms. Beyond exhibiting similar patterns of uptake and loss at the whole body level, Cd and Zn behaved similarly at the subcellular level. Under background levels (bkgd-Cd/bkgd-Zn), after uptake, worms partitioned a majority of Cd (∼ 65%) and Zn (∼ 55%) to the HSP and organelles fractions. The HDP and MRG fractions contained less than ∼6% of both metals. Following depuration, at bkgd-Cd/bkgd-Zn, Cd and Zn were lost from all subcellular fractions; loss from HSP was the greatest contributor to whole body loss. When exposed to elevated concentrations of Zn or Cd, the suppression in uptake of bkgd-Cd or bkgd-Zn observed in whole body uptake was largely due to suppressions in the storage of Cd and Zn to HSP. These results suggest that Cd–Zn interactions reduce partitioning of both Cd and Zn to HSP, indicating that metal-binding proteins such as metallothioneins play a key role in these interactions. © 2007 Elsevier B.V. All rights reserved. Keywords: Capitella capitata; Cd; Detoxification; Interaction; Subcellular; Zn

⁎ Corresponding author. Biology Department, College of Staten Island, 6S-310, City University of New York, 2800 Victory Boulevard, Staten Island, New York 10314, USA. Tel.: +1 718 982 3876; fax: +1 718 982 3923. E-mail address: [email protected] (W.G. Wallace). 0022-0981/$ - see front matter © 2007 Elsevier B.V. All rights reserved. doi:10.1016/j.jembe.2007.07.014

1. Introduction Aquatic organisms inhabiting urban coastal areas are exposed to a variety of potentially toxic and persistent

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contaminants (Levin et al., 2001; Wiegner et al., 2003). Studies examining bioaccumulation and subsequent toxicity of contaminants have often focused on single pollutants (e.g., Deeds and Klerks, 1999; Selck et al., 1999; Pechenik et al., 2001). However, the interactions between these pollutants are widely known and need to be considered (Kraak et al., 1993; Lenihan and Oliver, 1995; Tao et al., 1999; Lock and Janssen, 2002). For example, metal–metal interactions can have antagonistic, synergistic, or non-interactive effects on both bioaccumulation and toxicity (Amiard-Triquet and Amiard, 1998; Otitoloju, 2002). Interestingly, studies on the interactions between metals have produced inconsistent results. For example, Cd interacts antagonistically with Zn in the prawn Palaemon serratus (Devineau and Amiard-Triquet, 1985), whereas synergistic interactions have been noted for the amphipod Corophium volutator (Negilski et al., 1981; Rainbow et al., 2000). Understanding the mechanism(s) underlying these interactions and their implications to bioaccumulation, toxicity, and trophic transfer is, therefore, essential from an ecosystem management standpoint (Otitoloju, 2002). Trace metals are generally categorized as being essential or non-essential. Essential metals, such as Zn and Cu, are important in many molecular and cellular functions, and are thus often regulated by homeostatic mechanisms (Depledge and Rainbow, 1990; Martinez et al., 1999). For example, Zn plays an essential role in many enzymatic activities (Adams et al., 1982; Brown et al., 1987). Non-essential metals, such as Cd and Hg, however, have no known biological functions, and can interfere with the metabolism of essential metals (Brown et al., 1990), which may lead to toxicity (Bay et al., 1990). Impacted environments are often contaminated by a suite of metals, and there is the possibility of metal– metal interactions within aquatic organisms (e.g., competition for binding sites due to similar affinities for sulphur) (Bay et al., 1990; Amiard-Triquet and Amiard, 1998; Seebaugh and Wallace, 2004). It is, therefore, important to understand how such interactions impact metal uptake, loss, and toxicity (Kraak et al., 1993). Studies examining the accumulation of metals in aquatic organisms have often focused on whole tissue concentrations, and studies on the interaction between metals are no exception (Devineau and Amiard-Triquet, 1985; Van Gestel and Hensbergen, 1997; Bat et al., 1998; Rainbow et al., 2000). Studies, however, have also shown that the interactions between metals can occur at cellular and subcellular levels such as at certain absorption sites on biological membranes or at binding sites of proteins (Hemelraad et al., 1987). Whole tissue

concentrations may, therefore, not provide sufficient information to elucidate the mechanisms underlying the interactions, as well as possible toxicological consequences thereof. As many aquatic organisms can detoxify or sequester accumulated trace metals internally into intracellular components [e.g., metallothioneins (MT) and metal-rich granules (MRG)] (Brown et al., 1990; Roesijadi, 1992; Wallace et al., 2003), interactions among these subcellular components also need to be examined (Mason and Simkiss, 1983; Amiard-Triquet and Amiard, 1998; Seebaugh and Wallace, 2004). During uptake, metals may be bound initially to metalbinding proteins such as MT with subsequent redistribution into MRG (Mason and Jenkins, 1995, Wallace et al., 2003). Metals that are not stored as detoxified forms may be bound to metal-sensitive cellular components such as organelles (e.g., mitochondria) and high-molecularweight proteins (e.g., enzymes), potentially resulting in toxic effects (Bay et al., 1990; Wallace et al., 2003). Since metals have varying affinities for these binding moieties (Brown et al., 1990; Roesijadi, 1992; Cosson, 1994a,b), it is important to understand not only metal-specific partitioning patterns, but also metal–metal interactions at the subcellular level. Subcellular partitioning of metals within organisms exposed to a mixture of metals may, therefore, provide for a more complete understanding of metal- or species-specific differences in the interactions among metals, which may not be apparent from an analysis of whole tissue burdens alone. Although some studies have investigated accumulation patterns of multiple metals at the subcellular level, few have considered the implications of the interactions between metals within subcellular components (Hemelraad et al., 1987; Brown et al., 1990; Nott and Langston, 1993; Seebaugh and Wallace, 2004). A subcellular compartmentalization approach has been developed to conceptualize the potential toxicity, detoxification, and trophic transfer of trace metals in aquatic organisms (Wallace et al., 2003; Wallace and Luoma, 2003). This approach relies on the subcellular partitioning of trace metals within aquatic organisms to various operationally-defined subcellular fractions including organelles, heat-denaturable proteins (HDP) (i.e., ‘enzymes’), heat-stable proteins (HSP) (e.g., MT), MRG, and cellular debris (e.g., cell membrane) (Wallace et al., 2003). From the standpoint of metal toxicity and detoxification, metals in detoxified forms, HSP and MRG fractions, can be defined as ‘biologically detoxified metal’ (BDM) (Wallace et al., 2003). The partitioning of metals to this compartment (BDM) may relate to the metal-detoxifying capacity of an organism and potential tolerance (Wallace et al., 2003). Furthermore, as binding of metal to organelles and HDP may

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lead to toxicity, these fractions are defined as ‘metalsensitive fractions’ (MSF) (Wallace et al., 2003). Since the interactions among metals may influence metal partitioning to these various subcellular fractions, the MSF/BDM subcellular compartmentalization approach may highlight important ecotoxicological implications of the interactions. The deposit-feeding polychaete Capitella capitata is a cosmopolitan opportunistic species that inhabits estuarine and coastal environments (Grassle and Grassle, 1976; Tsutsumi, 1990; Ward and Hutchings, 1996). Due to its rapid population growth and high tolerance to natural (e.g., salinity) as well as anthropogenic (e.g., chemical contaminants) stresses (Tsutsumi, 1990; Pechenik et al., 2001), this polychaete species is often abundant in polluted estuaries (Warren, 1977; Levin et al., 1996; Ward and Hutchings, 1996; Mendez et al., 1997). C. capitata is, therefore, a useful test specimen to investigate the impacts of pollutants, including metals (Levin et al., 1996; Selck and Forbes, 2004). The purpose of the current work was to examine how Cd and Zn interact during the uptake and loss phases in C. capitata at both the whole body and subcellular levels under varying exposure regimes. Because of the interest in understanding the implications of possible Cd–Zn interactions, the experimental design included two exposure concentrations for each metal [background (bkgd) and high], with worms being exposed to all four possible combinations (i.e., bkgd-Cd/bkgd-Zn, bkgd-Cd/ high-Zn, high-Cd/bkgd-Zn, and high-Cd/high-Zn). This experimental design, therefore, allowed for the determination of the individual and combined effects of Cd and Zn on the uptake, loss, and subcellular partitioning of these metals within C. capitata. 2. Materials and methods 2.1. Culturing and exposure of polychaetes to Cd and Zn The initial population of C. capitata (sp. I — obtained from J. Grassle, Rutgers University) was maintained in laboratory cultures with 20 ppt seawater (20 °C), continuous aeration, b 300 μm sediment (Flax Pond, NY) and a photoperiod of 12:12 (light:dark) (Linke-Gamenick et al., 1999). Polychaetes were fed biweekly on Gerber® rice cereal and seawater was changed monthly. Medium sized polychaetes (∼1 mg wet weight each) were removed from cultures and were allowed to depurate gut contents for 24 h. Polychaetes (n = 100 per treatment) were then exposed for 1 week in 20 ppt seawater at room temperature (∼ 20 °C) to one of two

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nominal Cd concentrations (background Cd or ‘high Cd’ at 50 μg l− 1) and one of two nominal Zn concentrations (background Zn or ‘high Zn’ at 86 μg l− 1), yielding a total of four possible treatments (bkgd-Cd/bkgd-Zn, 50Cd/bkgd-Zn, bkgd-Cd/86-Zn, and 50-Cd/86-Zn). Exposure concentrations were chosen based on previous studies (for Cd; Selck et al., 1998) and preliminary exposures, indicating that the combined exposure of both metals at the highest concentrations (50-Cd/86-Zn) resulted in ∼ 50% mortality. Exposures occurred within 4 l polycarbonate bottles containing 1.1 l of continuously aerated 20 ppt seawater (Instant Ocean®) maintained at room temperature. Nominal exposure concentrations of ‘high Cd’ (50 μg l− 1) or ‘high Zn’ (86 μg l− 1) were prepared by adding appropriate amounts of Cd or Zn AA standards (VWR® 1000 mg l− 1) to exposure vessels. All treatments received similar spikes of 109Cd and 65Zn (109 Cd: 80 kBq l− 1 ; 65 Zn: 181 kBq l− 1 ) Isotope Products® and 0.5 N NaOH, as needed, to neutralize HCl introduced with standards and isotope. Based on the specific activities of the spiked isotope, background metal concentrations were elevated by ∼ 1 μg l− 1 for Cd and ∼ 3 μg l− 1 for Zn. Background nominal concentrations were, therefore, set at these levels (i.e., 1 μg l− 1 for Cd and 3 μg l− 1 for Zn) as metals in Instant Ocean® were not determined. After 1 week, polychaetes were removed from the exposure solutions, rinsed three times with clean seawater (20 ppt) and enumerated. Polychaetes from each treatment were then split randomly into two groups. One group of ∼ 25 polychaetes from each treatment was further separated into replicates of 5–7 individuals (4 replicates per treatment), dabbed dry with Kimwipes®, wet weighed within tarred Eppendorf® microcentrifuge tubes, and were stored frozen (− 80 °C). The other group of polychaetes, ∼35 per treatment, were placed in 1 l plastic beakers containing 500 ml of continuously aerated 20 ppt seawater (20 °C) and a ∼1 cm layer of sediment (b 300 μm; Flax Pond, NY). Polychaetes were then allowed to depurate accumulated metals for 6 days. Polychaetes were then removed from the sediment via sieving and were allowed to depurate gut contents for an additional 24 h, therefore receiving a deputation period of 1 week. Following depuration, polychaetes (5–7 individuals per replicate, 4 replicates per treatment) were wet weighed as described above and stored frozen (−80 °C). 2.2. Subcellular fractionation Composite sets of radiolabelled polychaetes (n = 4 replicates per treatment) were used in subcellular fractionation and tissue digestion procedures modified

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Table 1 Two-way ANOVAs on Cd burdens (whole body, subcellular compartments and subcellular fractions) after uptake by Capitella capitata with nominal concentration levels of Cd and Zn (either background or high) as independent factors Cd burdens after uptake Effect A) Whole body (total) Cd level Zn level Cd level ⁎ Zn level Error B) MSF Cd level Zn level Cd level ⁎ Zn level Error C) HDP Cd level Zn level Cd level ⁎ Zn level Error D) ORG Cd level Zn level Cd level ⁎ Zn level Error

Cd burdens after uptake SS

df

MS

F

p

48.83 1.11 0.29 0.62

1 1 1 12

48.83 1.11 0.29 0.05

937.9 21.3 5.5

0.000 0.001 0.037

51.61 0.48 0.05 1.50

1 1 1 12

51.61 0.48 0.05 0.13

412.2 3.9 0.4

0.000 0.073 0.534

46.43 0.16 1.05 4.22

1 1 1 12

46.43 0.16 1.05 0.35

132.2 0.5 3

0.000 0.511 0.109

52.16 0.53 0.02 1.72

1 1 1 12

52.16 0.53 0.02 0.14

364.9 3.7 0.1

0.000 0.077 0.744

Effect E) CD Cd level Zn level Cd level ⁎ Zn level Error F) BDM Cd level Zn level Cd level ⁎ Zn level Error G) HSP Cd level Zn level Cd level ⁎ Zn level Error H) MRG Cd level Zn level Cd level ⁎ Zn level Error

SS

df

MS

F

p

46.88 0.78 0.27 3.94

1 1 1 12

46.88 0.78 0.27 0.33

142.7 2.4 0.8

0.000 0.15 0.383

43.88 0.87 0.84 1.94

1 1 1 12

43.88 0.87 0.84 0.16

271.2 5.4 5.2

0.000 0.039 0.042

43.26 0.85 0.86 2.10

1 1 1 12

43.26 0.85 0.86 0.17

247.7 4.9 4.9

0.000 0.048 0.047

62.28 0.79 0.1 6.26

1 1 1 12

62.28 0.79 0.1 0.52

119.3 1.5 0.2

0.000 0.242 0.668

Significant effects are indicated by p-value in bold (p b 0.05).

from those of Wallace et al. (2003), with an initial centrifugation step of 1000 ×g. Subcellular fractionation, subsequent to homogenization in 3 ml of cold 7.6 pH TRIS with a Polytron® homogenizer, resulted in the isolation of five operationally-defined subcellular fractions: NaOH insoluble body parts (i.e., metal-rich granules — MRG if present, and setae), NaOH soluble body parts (i.e., cellular debris — CD), organelles (100,000 ×g pellet), heat-stable proteins or HSP (e.g., MT) and heat-denatured proteins or HDP (i.e., ‘enzymes’). For the purposes of presentation, metals associated with these fractions are also grouped into compartments: metalsensitive fractions (MSF — organelles and HDP) and biologically detoxified metal (BDM — HSP and MRG).

type NaI crystal detector. Photon emissions were determined at 88 keV for 109 Cd and 1115 keV for 65Zn. Concentrations of accumulated metal were calculated based on isotopic specific activities of spiked metal (Cd: 1.2 × 10− 2 to 62.2 × 10− 2 μg kBq− 1; Zn: 1.5 × 10− 2 to 50.7 × 10− 2 μg kBq− 1) and are expressed as μg wet wt− 1. Percentage subcellular distributions of 109Cd and 65Zn within polychaetes were calculated based on radioactivity recovered after homogenization. Recovery of homogenate radioactivity subsequent to fractionation was consistently high (N 85%) (i.e., [summation of radioactivity in each of the 5 subcellular fractions] / [radioactivity following homogenization]). 2.4. Data analysis

2.3. Radioanalysis Sample radioactivity was determined with a Wallac 1480 Wizard γ-counter equipped with a 7.8 cm well-

Data were analyzed using the software Statistica version 7.1 (Statsoft, Inc®). When necessary, proportional data were arcsine-transformed, and concentration

Fig. 1. Cd concentrations (μg·g− 1 wet weight; mean ± SE, n = 4) in (A) whole body and (B–H) various subcellular fractions and compartments of Capitella capitata following 1 week of uptake (Up) at bkgd-Cd (on left in panels) or 50-Cd (on right in panels), both co-exposed to either bkgd-Zn (triangle) or 86-Zn (circle), and 1 week of loss (Ls) in “clean” water. Subcellular fractions and compartments are as follows: (B) metal-sensitive fractions (MSF); (C) heat-denatured protein (HDP); (D) organelles (ORG); (E) cellular debris (CD); (F) biologically detoxified metal (BDM); (G) heat-stable protein (HSP); (H) metal-rich granules (MRG). A significant change (2-way repeated ANOVA followed by orthogonal planned comparison, p b 0.05) in Cd body burdens (whole body and subcellular fractions/compartments) following loss is indicated by the presence of a line connecting the uptake (Up) and loss (Ls) symbols for a given treatment. A significant difference (2-way ANOVA followed by orthogonal planned comparison, p b 0.05) between Cd burdens (whole body or subcellular fractions/compartments) after uptake or loss upon varying Zn exposure (bkgdZn vs. 86-Zn) is indicated by the presence of an ⁎ above the bkgd-Zn and 86-Zn symbols.

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data were log-transformed before analyses. The normality of data was tested with Shapiro–Wilk's W test. Homogeneity of variance was assessed with Levene's test. Treatment effects on metal concentrations in whole body burdens as well as subcellular fractions after the uptake phase of exposure were analyzed using 2-way

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ANOVA, followed by orthogonal planned comparisons for specific treatment effects (Sokal and Rohlf, 1995). The effect of 1-week depuration phase (i.e., time effect) on whole body burdens and subcellular partitioning of metals was analyzed using 2-way repeated measures ANOVA, followed by orthogonal planned comparisons

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on specific treatment and time effects. There was no significant difference in the proportional data among the treatments after the uptake phase (1-way ANOVA), therefore all proportional data from treatment groups were pooled for either Cd or Zn, and analyzed for differences in proportions (between metals) using student's t-test. The critical level for these analyses was determined at p b 0.05. 3. Results 3.1. Whole body burdens and subcellular partitioning of Cd after uptake There was a significant interaction between Cd and Zn treatment effects on the uptake of Cd by C. capitata at the whole body level (Table 1A). Following a 1-week exposure to background levels of Cd (∼ 1 μg l− 1) and Zn (∼ 3 μg l− 1) (bkgd-Cd/bkgd-Zn, Fig. 1A), C. capitata accumulated a whole body Cd tissue burden of ∼0.41 μg g wet− 1. When Zn exposure was increased to 86 μg l− 1 (bkgd-Cd/86-Zn), the accumulation of Cd by C. capitata was suppressed by ∼ 56% with worms achieving a whole body Cd tissue burden of only ∼ 0.18 μg g wet− 1 (Fig. 1A). The accumulation of Cd at 50 μg l− 1, however, was unresponsive to varying levels of Zn (bkgd or 86 μg l− 1), with worms from both treatments attaining a Cd tissue burden of ∼ 9 μg g wet− 1 (Fig. 1A). There was no significant interaction between Cd and Zn treatment effects on the partitioning of Cd to the MSF compartment of C. capitata after uptake (Table 1B). After one week of exposure, worms from the bkgd-Cd/bkgd-Zn and bkgd-Cd/86-Zn treatments had similar amounts of Cd associated with the MSF compartment (Fig. 1B; ∼ 0.06 μg g wet− 1), and exhibited no differences in the partitioning of Cd to fractions thereof (Fig. 1C and D; HDP: ∼ 0.01 μg g wet− 1 and organelles: ∼ 0.05 μg g wet− 1). Worms from both the 50-Cd/bkgd-Zn and 50-Cd/86-Zn treatments were also found to contain similar amounts of Cd bound to the MSF compartment (Fig. 1B; ∼ 2.4 μg g wet− 1)

and associated fractions (Fig. 1C and D; HDP: ∼ 0.30 μg g wet− 1 and organelles: ∼ 2.5 μg g wet− 1). Within the BDM compartment, however, there was a significant interaction between Cd and Zn treatment effects (Table 1F), which was driven by the partitioning of Cd to HSP (Table 1G). The HSP fraction of bkgd-Cd/ bkgd-Zn worms contained more than twice as much Cd as that of the bkgd-Cd/86-Zn worms (Fig. 1G; ∼ 0.14 vs. 0.05 μg g wet− 1). Worms from these treatments had similar amounts of Cd bound to cellular debris and MRG (Fig. 1E and H; ∼ 0.07 and ∼ 0.004 μg g wet− 1). Worms from both the 50-Cd/bkgd-Zn and 50-Cd/86Zn treatments were found to contain similar amounts of Cd bound to the BDM compartments (Fig. 1F; ∼ 2.4 μg g wet− 1). Additionally, there was no difference between these Zn treatments in the amount of Cd bound to constituents of the BDM compartment (i.e., HSP: Fig. 1G; MRG: Fig. 1H). 3.2. Whole body burdens and subcellular partitioning of Zn after uptake The accumulation of background Zn under varying Cd exposures (i.e., bkgd-Cd/bkgd-Zn or 50-Cd/bkgdZn) followed a similar pattern as that for background Cd under varying Zn (Fig. 2). A significant interaction between Zn and Cd treatment effects on the uptake of Zn at the whole body level was exhibited by C. capitata (Table 2A). When exposed to a background level of Cd (bkgd-Cd/bkgd-Zn), C. capitata accumulated a whole body Zn tissue burden of ∼ 1.9 μg g wet− 1 (Fig. 2A). An increased exposure to Cd (50-Cd/bkgd-Zn) resulted in a whole body Zn tissue burden of only ∼ 1.0 μg g wet− 1, which was ∼ 50% lower than that of the bkgd-Cd/bkgdZn worms (Fig. 2A). Worms exposed to a high level of Zn (86 μg l− 1) accumulated a whole body Zn tissue burden of ∼ 23 μg g wet− 1 regardless of the concentration of Cd exposure, bkgd or 50 μg l− 1 (Fig. 2A). After uptake, the organelle fraction of worms exposed to a background level of Cd (bkgd-Cd/bkgd-Zn) had 1.6-fold more Zn than those exposed to an elevated level of Cd (50-Cd/bkgd-Zn) (Fig. 2D; ∼ 0.31 vs.

Fig. 2. Zn concentrations (μg·g− 1 wet weight; mean ± SE, n = 4) in (A) whole body and (B–H) various subcellular fractions and compartments of Capitella capitata following 1 week of uptake (Up) at bkgd-Zn (on left in panels) or 86-Zn (on right in panels), both co-exposed to either bkgd-Cd (triangle) or 50-Cd (circle), and 1 week of loss (Ls) in “clean” water. Subcellular fractions and compartments are as follows: (B) metal-sensitive fractions (MSF); (C) heat-denatured protein (HDP); (D) organelles (ORG); (E) cellular debris (CD); (F) biologically detoxified metal (BDM); (G) heat-stable protein (HSP); (H) metal-rich granules (MRG). A significant change (2-way repeated ANOVA followed by orthogonal planned comparison, p b 0.05) in Zn body burdens (whole body and subcellular fractions/compartments) following loss is indicated by the presence of a line connecting the uptake (Up) and loss (Ls) symbols for a given treatment. A significant difference (2-way ANOVA followed by orthogonal planned comparison, p b 0.05) between Zn burdens (whole body or subcellular fractions/compartments) after uptake or loss upon varying Cd exposure (bkgdCd vs. 50-Cd) is indicated by the presence of an ⁎ above the bkgd-Cd and 50-Cd symbols.

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0.19 μg g wet− 1), while there was no difference in the concentration of Zn associated with the HDP of these two treatments (Fig. 2C; ∼ 0.04 μg g wet− 1). Additionally, there was no significant interaction between Zn and Cd treatment effects on the partitioning of Zn to the

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entire MSF compartment after uptake (Table 2B). The greater partitioning of Zn to organelles in bkgd-Cd/ bkgd-Zn worms resulted in the MSF compartment of these worms containing more Zn than that of the 50-Cd/ bkgd-Zn worms (Fig. 2B; ∼ 0.38 vs. 0.22 μg g wet− 1).

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Table 2 Two-way ANOVAs on Zn burdens (whole body, subcellular compartments and subcellular fractions) after uptake by Capitella capitata with nominal concentration levels of Cd and Zn (either background or high) as independent factors Zn burdens after uptake Effect A) Whole body (total) Zn level Cd level Zn level ⁎ Cd level Error B) MSF Zn level Cd level Zn level ⁎ Cd level Error C) HDP Zn level Cd level Zn level ⁎ Cd level Error D) ORG Zn level Cd level Zn level ⁎ Cd level Error

Zn burdens after uptake SS

df

MS

F

p

32.18 0.41 0.54 0.37

1 1 1 12

32.18 0.41 0.54 0.03

1035 13 17

0.000 0.003 0.001

34.87 0.43 0.16 0.88

1 1 1 12

34.87 0.43 0.16 0.07

476.6 5.8 2.2

0.000 0.033 0.16

38.52 0.36 0.41 2.84

1 1 1 12

38.52 0.36 0.41 0.24

162.6 1.5 1.7

0.000 0.242 0.215

34.31 0.48 0.15 1.19

1 1 1 12

34.31 0.48 0.15 0.10

346.4 4.9 1.6

0.000 0.048 0.236

Effect E) CD Zn level Cd level Zn level ⁎ Cd level Error F) BDM Zn level Cd level Zn level ⁎ Cd level Error G) HSP Zn level Cd level Zn level ⁎ Cd level Error H) MRG Zn level Cd level Zn level ⁎ Cd level Error

SS

df

MS

F

p

34.47 0.56 0.61 1.36

1 1 1 12

34.47 0.56 0.61 0.11

303.8 4.9 5.4

0.000 0.046 0.038

34.81 1.15 0.93 1.01

1 1 1 12

34.81 1.15 0.93 0.08

413.9 13.7 11.1

0.000 0.003 0.006

34.45 1.08 0.96 1.10

1 1 1 12

34.45 1.08 0.96 0.09

375.5 11.8 10.5

0.000 0.005 0.007

41.73 3.12 0.61 2.90

1 1 1 12

41.73 3.12 0.61 0.24

172.4 12.9 2.5

0.000 0.004 0.137

Significant effects are indicated by p-value in bold ( p b 0.05).

Partitioning of Zn to both fractions of the BDM compartment was Cd-exposure dependent, which was also influenced by the interaction between Cd and Zn (Table 2F). Worms from the 50-Cd/bkgd-Zn treatment had significantly less Zn associated with HSP, as compared with bkgd-Cd/bkgd-Zn worms (Fig. 2G; ∼ 0.15 vs. ∼ 0.44 μg g wet− 1), as well as metal-rich granules (MRG) (Fig. 2H). These differences were reflected in the partitioning of Zn to the BDM comportment (Fig. 2F). The storage of Zn in cellular debris of worms from the 50-Cd/bkgd-Zn treatment was also suppressed (∼53% lower), as compared with bkgd-Cd/ bkgd-Zn worms (Fig. 2E; ∼ 0.25 vs. ∼0.52 μg g wet− 1). The concentration of Zn bound to the various subcellular compartments and fractions of worms in the bkgdCd/86-Zn and 50-Cd/86-Zn treatments was found to be similar after uptake (Fig. 2B through H). 3.3. Whole body burdens and subcellular partitioning of Cd after loss The results of 2-way repeated measures ANOVA showed that the interaction of Cd and Zn treatment effects on Cd whole body burdens in C. capitata changed over the 1-week depuration phase (Fig. 1A). Upon depuration, ∼75% of the Cd in worms exposed to the bkgd-Cd/bkgd-Zn treatment was lost, with worms

maintaining a tissue concentration of ∼ 0.10 μg g wet− 1 (Fig. 1A). The initial tissue concentration of Cd in worms exposed to the bkgd-Cd/86-Zn treatment, however, was maintained during depuration, resulting in a retained Cd tissue burden of ∼ 0.14 μg g wet− 1, which was 1.4-fold greater than that retained by worms exposed to background Zn (bkgd-Cd/bkgd-Zn) (Fig. 1A). Following 1 week of depuration, the 50-Cd/bkgd-Zn worms retained Cd, while the 50-Cd/86-Zn worms depurated ∼ 20% of the accumulated metal (Fig. 1A). This loss of Cd by 50-Cd/86-Zn worms resulted in a whole body Cd tissue burden of ∼ 6 μg g wet− 1, which was ∼ 28% lower than that retained by the 50-Cd/bkgdZn worms (Fig. 1A). The amount of Cd bound to virtually all subcellular fractions of the bkgd-Cd/bkgd-Zn worms decreased after 1 week of depuration (Fig. 1B through H), with loss from the HSP fraction accounting for ∼35% of whole body loss (Fig. 1G); only the HDP fraction retained Cd following depuration (Fig. 1C). Just with total Cd burdens, however, no subcellular fraction of worms from the bkgd-Cd/86-Zn treatment exhibited a loss of Cd (Fig. 1B through H). Retention of Cd by the bkgd-Cd/86-Zn worms following 1 week of depuration resulted in greater Cd burdens, as compared with bkgdCd/bkgd-Zn worms, being associated with organelles (Fig. 1D; ∼ 0.05 vs. ∼ 0.02 μg g wet− 1) and MRG

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(Fig. 1H; ∼ 0.0025 μg g wet− 1 vs. ∼ 0.001 μg g wet− 1). Additionally, even though the HSP fraction of bkgd-Cd/ bkgd-Zn worms initially contained ∼2.6 times more Cd than the same fraction of bkgd-Cd/86-Zn worms (Fig. 1G; ∼0.14 vs. 0.05 μg g wet− 1), after loss there was no difference between these treatments in the amount of Cd bound to HSP (Fig. 1G; ∼ 0.03 μg g wet− 1). These temporal patterns (or lack-there-of) in the concentration of Cd bound to the various subcellular fractions following 1 week of depuration were similarly reflected in the partitioning of Cd to the MSF and BDM compartments (Fig. 1B and F). No fraction of the 50-Cd/bkgd-Zn worms exhibited a loss of Cd after 1 week of depuration (Fig. 1B through H). A similar retention of Cd was exhibited by 50-Cd/86-Zn worms, though differences between the Zn treatments (50-Cd/bkgd-Zn and 50-Cd/86-Zn) after 1 week of depuration were observed for the CD fraction and the BDM compartment (Fig. 1E and F). 3.4. Whole body burdens and subcellular partitioning of Zn after loss When co-exposed to a background level of Cd (bkgdCd/bkgd-Zn), ∼ 65% of the initially accumulated whole body Zn tissue burden (∼1.9 μg g wet− 1) was lost during depuration (Fig. 2A). When co-exposed to a high level of Cd (50-Cd/bkgd-Zn), however, the whole body Zn tissue burden in C. capitata was retained during depuration, resulting in an ultimate Zn tissue burden for the 50-Cd/bkgd-Zn worms that was ∼1.5-fold greater than that retained of the bkgd-Cd/bkgd-Zn worms (Fig. 2A). No significant loss of Zn was exhibited by worms exposed to either of the high-Zn treatments, bkgd-Cd/86-Zn or 50-Cd/86-Zn, with worms from both treatments maintaining a whole body Zn tissue burden of ∼21 μg g wet− 1 (Fig. 2A). After a week of loss, worms from the bkgd-Cd/bkgdZn treatment lost Zn from all subcellular fractions (Fig. 2B through H), with loss from the HSP fraction accounting for ∼ 27% of whole body loss. In comparison, no subcellular fraction of worms from the 50-Cd/ bkgd-Zn treatment exhibited a loss of Zn (Fig. 2B through H). In fact, the partitioning of Zn to the cellular debris and MRG fractions increased during this depuration phase (Fig. 2E and H). The loss of Zn from the HSP fraction of bkgd-Cd/bkgd-Zn worms and retention in this fraction by 50-Cd/bkgd-Zn worms resulted in the HSP of 50-Cd/bkgd-Zn worms having a slightly higher Zn burden than that of bkgd-Cd/bkgdZn worms (Fig. 2G). This retention of Zn in all the subcellular fractions of the 50-Cd/bkgd-Zn worms

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resulted in these worms having a whole body Zn tissue burden, after loss, that was ∼ 1.5× greater than that of bkgd-Cd/bkgd-Zn worms, even though the bkgd-Cd/ bkgd-Zn worms originally contained ∼ 2× more Zn (Fig. 2A). A significant loss of Zn was also observed in the fractions of the BDM compartment (i.e., HSP and MRG) of 50-Cd/86-Zn worms (Fig. 2G and H), while the bkgd-Cd/86-Zn worms retained Zn in all subcellular fractions during depuration (Fig. 2B through H). The retention of Zn in the MRG of bkgd-Cd/86-Zn worms and loss from this fraction by 50-Cd/86-Zn worms resulted in the bkgd-Cd/86-Zn worms having ∼ 10× more Zn bound to this fraction (Fig. 2H). 3.5. Proportional subcellular partitioning of Cd and Zn After 1 week of uptake, there were no significant differences among treatment groups in the proportional partitioning (%) of Cd or Zn among the various subcellular fractions within C. capitata. In general, Cd and Zn partitioned equally to the HSP, ORG, and cell debris (CD) fractions, with each fraction accounting for approximately 30–40% of each metal (Fig. 3). The HDP and MRG fractions accounted for substantially lower proportions (∼ 5% and ∼ 2%) of the metals. When all treatment groups were pooled, as per metal, however, significantly more Cd was found to be associated with the HSP fraction (Fig. 3A and B; 36.3% for Cd and 28.0% for Zn), while more Zn partitioned to the cellular

Fig. 3. Proportional distribution (means of pooled data) of (A) Cd and (B) and Zn in subcellular fractions [heat-stable protein (HSP), cellular debris (CD), heat-denatured protein (HDP), organelles (ORG) and metal-rich granules (MRG)] of Capitella capitata following 1 week uptake in a variety of Cd and Zn exposures (see text for clarification/ justification of “pooled data”). Significant differences (t-test p b 0.05) between metals as per fraction are indicated by the presence of an ⁎ in the pie slice for the fraction having the higher proportion.

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debris (CD) (Fig. 3A and B; 29.5% for Cd and 39.4% for Zn). 4. Discussion 4.1. Cd and Zn interactions at the whole body level The interaction between Cd and Zn in aquatic organisms at the whole body level is known to be highly variable (Rainbow et al., 2000), with interactions being species-, population-, and exposure-dependent (Jackim et al., 1977; Amiard-Triquet and Amiard, 1998; Lange et al., 2002). For instance, some studies have shown that Zn (or Cd) suppresses Cd (or Zn) bioaccumulation by reducing uptake (i.e., antagonistic interactions) (Jackim et al., 1977; Seebaugh and Wallace, 2004), whereas other studies have shown that Zn (or Cd) enhances Cd (or Zn) bioaccumulation by increasing uptake (i.e., synergistic interaction) (Ahsanullah et al., 1981). In some aquatic organisms, however, Cd and Zn have no interaction during uptake or elimination (Kraak et al., 1993; Martinez et al., 1999; Wicklund et al., 1988). In the present study, laboratory-reared C. capitata were exposed for 1 week to Cd and Zn under varying exposure regimes. When the concentration of Zn was elevated from a background level (bkgd-Cd/bkgd-Zn) to 86 μg l− 1 (bkgd-Cd/86-Zn), Cd uptake by C. capitata was reduced by more than 50%. A similar, though not as dramatic, reduction in Cd uptake (a 9% decrease) was observed in the polychaete Nereis diversicolor, when exposed to 100 compared to 10 μg Zn l− 1 (Bryan and Hummerstone, 1973). These studies suggest that Zn may be able to suppress Cd uptake through metal–metal interactions at uptake sites, and that these interactions may result in reduced Cd toxicity in these organisms (Jenkins and Mason, 1988; Bat et al., 1998; Barata et al., 2002). Exposure to high concentrations of Zn can lead to a decreased uptake or increased elimination of Cd, which may lead to reduced Cd bioaccumulation and/or toxicity. This may result from interference or competition at external as well as internal binding sites (Oakden et al., 1984; Rainbow et al., 2000). In addition to reduced bioaccumulation, a higher retention of accumulated Cd by C. capitata was also observed when coexposed with elevated Zn. Zn-induced retention of Cd may be due to Cd being sequestered via metal-binding proteins (e.g., metallothioneins — MT) that were induced by the Zn exposure (Seebaugh and Wallace, 2004). Reciprocal interactive effects of Cd on Zn accumulation in aquatic organisms have been examined less frequently than those of Zn on Cd. In some aquatic

invertebrates, Zn uptake is not influenced by the presence of other elements including Cd, indicating some degree of internal regulation of Zn (Depledge and Rainbow, 1990; Mwangi and Alikhan, 1993; Martinez et al., 1999). For example, Zn bioaccumulation was not affected when the brine shrimp Artemia parthenogenetica was co-exposed to Cd (Martinez et al., 1999). However, a study by Devineau and Amiard-Triquet (1985) showed that the exposure to 300 and 500 μg Cd l− 1 reduces Zn uptake in larvae of the prawn Paleamon serratus. This effect was not observed in prawns exposed to 100 μg Cd l− 1 (Devineau and Amiard-Triquet, 1985). A similar impact of elevated Cd concentrations on Zn uptake was observed in Artemia franciscana (Seebaugh and Wallace, 2004). In this case, Zn uptake by A. franciscana was significantly reduced (∼50%) when co-exposed with 1 to 50 μg Cd l− 1. In the present study, when C. capitata were coexposed to a background level of Zn with an elevated level of Cd (50-Cd/bkgd–Zn), as compared with uptake upon exposure to bkgd-Cd/bkgd-Zn, Zn uptake was significantly reduced (∼ 50%) and Zn loss was suppressed. This may also be an indication of competition between Cd and Zn at uptake sites within C. capitata, thereby resulting in reduced Zn loss, as well as reduced Cd uptake (Oakden et al., 1984; Rainbow et al., 2000). Cd and Zn are known to have similar chemical structures (Rainbow et al., 2000), and this may influence the uptake patterns of these metals (Oakden et al., 1984; Wang and Fisher, 1999). Additionally, as an essential metal, Zn homeostasis in aquatic invertebrates, as well as vertebrates, is widely documented, and Zn bioaccumulation in these organisms, to some extent, is regulated (Mwangi and Alikhan, 1993; Martinez et al., 1999). The possibility of Zn regulation may explain its high retention even when co-exposed with elevated Cd, thereby allowing organisms to maintain the Zn levels required for critical cellular functions (Adams et al., 1982; Brown et al., 1987). A study by Devineau and Amiard-Triquet (1985) demonstrates that when co-exposed to Cd with various levels of Zn, the interactive effect of Zn on Cd (e.g., reduced Cd uptake) in larvae of P. serratus is only observed at 75 and 125 μg Zn l− 1. However, when exposed to a higher Zn concentration (275 μg Zn l− 1), there was no apparent interaction between these metals during uptake (Devineau and Amiard-Triquet, 1985). A similar trend was also observed in the current study. When Cd concentration was increased to 50 μg l− 1 (50-Cd/bkgd-Zn or 50-Cd/86-Zn), uptake and loss of Cd by C. capitata were unresponsive to the varying Zn levels (i.e., the similar Cd tissue burdens were retained upon loss). A similar result (i.e., no apparent effect of Cd

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on Zn uptake or loss) was observed when C. capitata were co-exposed to elevated Zn with a background or elevated level of Cd (bkgd-Cd/86-Zn or 50-Cd/86-Zn). These results suggest that Cd–Zn interactions within C. capitata may only occur within certain concentration ranges (Lange et al., 2002). 4.2. Cd and Zn interactions at the subcellular level Although Zn-induced reduction of Cd toxicity or enhancement of Cd tolerance in various aquatic organisms has been documented (Braek et al., 1980; Oakden et al., 1984; Stuhlbacher et al., 1992; Barata et al., 2002), the mechanisms and implications underlying these interactive effects (e.g., the relationship between metal bioaccumulation and toxicity) are not clearly understood. Some studies, for example, show that when organisms are co-exposed to Cd and Zn, mortality due to Cd exposure is positively correlated with whole body Cd burdens (Oakden et al., 1984; Bat et al., 1998), whereas other studies have shown the opposite result (Stuhlbacher et al., 1992). Previous studies have shown that the subcellular partitioning of metal is related to a variety of toxic effects (Jenkins and Mason, 1988; Bay et al., 1990; Wallace et al., 2000; Giguere et al., 2003). For example, following an 11-week exposure to Cd, reproductive output of the polychaete Neanthes arenaceodentata was reduced and correlated with an increased partitioning of Cd to the cytosolic fraction (Jenkins and Mason, 1988). Bay et al. (1990) demonstrated that the activity of the enzyme Cu–Zn superoxide dismutase in scorpionfish (Scorpaena guttata) intestine was negatively correlated with the partitioning of Cu to the enzyme fraction. The subcellular partitioning of metal also has a potential implication for behavioral toxicity. Wallace et al. (2000) found a positive relationship between the amount of Cd partitioned to the enzyme fraction of grass shrimp (Palaemonetes pugio) and reductions in prey capture success. Although there are a number of studies on the subcellular partitioning of metals in aquatic organisms, few have examined the potential implications of metal– metal interactions at the subcellular level from the standpoint of metal toxicity and detoxification. In the current study, the trends in whole body Cd and Zn bioaccumulation in C. capitata were mostly reflected in the subcellular partitioning, as well as the compartmentalization of the metals. Specifically, heat-stable proteins or HSP (e.g., MT) of the biologically detoxified metal (BDM) compartment appear to play key roles in the patterns of Cd and Zn bioaccumulation in C. capitata.

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When Zn was elevated (bkgd-Cd/86-Zn), Cd whole body burden in C. capitata was suppressed after 1 week of uptake. This reduced Cd uptake, as compared to worms exposed to background levels of Zn (bkgd-Cd/ bkgd-Zn), was reflected in the amount of Cd partitioned to the HSP fraction in the BDM compartment; no other fraction of the bkgd-Cd/86-Zn worms exhibited this reduced Cd burden. This may suggest that despite a higher Cd uptake, the majority of Cd in C. capitata coexposed to background levels of both Cd and Zn was stored as a detoxified form in the BDM compartment. After 1 week of the depuration, worms from the bkgdCd/bkgd-Zn exhibited a loss of Cd from the HSP fraction of the BDM compartment, corresponding to whole body loss. This suggests that Cd partitioning to metal-binding proteins (e.g., metallothioneins) is partially responsible for the high elimination rate of Cd. When C. capitata were co-exposed to an elevated level of Zn (bkgd-Cd/86-Zn), however, Cd was equally distributed between the BDM and MSF compartments; most of this Cd was retained after a week of depuration. It has been suggested that Zn can reduce Cd bioaccumulation and toxicity in many aquatic organisms through metal–metal interactions (Roesijadi, 1996; Otitoloju, 2002). Anecdotal survivorship data of C. capitata after the 1 week exposure to the various treatments in this study suggests that elevated Zn suppresses Cd toxicity (% survival: ∼ 74% for bkgd-Cd/86-Zn N ∼ 70% for bkgd-Cd/bkgd-Zn N ∼66% for 50-Cd/86-Zn N ∼52% for 50-Cd/bkgd-Zn). The results from this study indicate that there may indeed be important toxicological Cd–Zn interactions occurring within C. capitata. Additionally, this study suggests that metal-binding proteins are likely to be involved in these processes, resulting in reduced Cd bioaccumulation with possible implications for a reduction in Cd toxicity (Cain et al., 2006). 4.3. Conclusion Examining whole metal body burdens may not always be sufficient to properly elucidate the ecotoxicological significance (i.e., detoxification, toxicity and trophic transfer) of the bioaccumulation of multiple metals (Brown et al., 1990; Wallace et al., 2003; Wallace and Luoma, 2003). This current study investigated the role of subcellular partitioning in Cd–Zn interactions in C. capitata, and results show that Cd and Zn have similar impacts on the bioaccumulation pattern of the other metal. When C. capitata were co-exposed to Cd and Zn, these metals appear to compete for uptake sites within C. capitata, resulting in significant changes in uptake and loss. Furthermore, subcellular partitioning

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patterns of Cd and Zn indicate that the Cd–Zn interactions significantly influence the distributions of these metals among a variety of subcellular constituents. Cd–Zn interactions appear to reduce partitioning of both metals to heat-stable proteins (HSP) indicating that metal-binding proteins such as MT are an important site of this competition. This is in agreement with other studies (Roesijadi, 1996; Martinez et al., 1999). Altered subcellular partitioning patterns caused by Cd–Zn interaction may suggest a subsequent reduction in metal toxicity in C. capitata, and may need to be considered as a confounding factor when attempting to model the impacts of metals in aquatic ecosystems (Otitoloju, 2002). By using a subcellular compartmentalization approach, the current study provides a better understanding of the relationships between bioaccumulation patterns of metal mixtures (Cd and Zn) and potential toxic effects and/or detoxification. Acknowledgments This research was supported, in part, by PSC-CUNY Research Award #65269-0034 to W. G. Wallace. This manuscript represents contribution #0601 from the Center from Environmental Science, College of Staten Island, CUNY. The authors would like to thank J. Grassle (Rutgers University) for supplying the culture of C. capitata, A. R. Ferretti and E. C. Wallace for assistance with laboratory work, and D. R. Seebaugh and two anonymous reviewers for commenting on this manuscript. Access to radiation facilities and environmental chambers was provided by the Department of Biology at the College of Staten Island and is greatly appreciated. [SS] References Adams, E., Simkiss, K., Taylor, M., 1982. Metal ion metabolism in the moulting crayfish (Austropotamobius pallipes). Comp. Biochem. Physiol., Part A Physiol. 72, 73–76. Ahsanullah, M., Negilski, D.S., Mobley, M.C., 1981. Toxicity of zinc, cadmium and copper to the shrimp Callianassa australiensis. III. Accumulation of metals. Mar. Biol. 64, 311–316. Amiard-Triquet, C., Amiard, J.-C., 1998. Influence of ecological factors on accumulation of metal mixtures. In: Langston, W.J., Bebianno, M.J. (Eds.), Metal Metabolism in Aquatic Environments. Chapman and Hall, London, UK, pp. 351–386. Barata, C., Markich, S.J., Baird, D.J., Taylor, G., Soares, A.M.V.M., 2002. Genetic variability in sublethal tolerance to mixtures of cadmium and zinc in clones of Daphnia magna Straus. Aquat. Toxicol. 60, 85–99. Bat, L., Raffaelli, D., Marr, I.L., 1998. The accumulation of copper, zinc and cadmium by the amphipod Corophium volutator (Pallas). J. Exp. Mar. Biol. Ecol. 223, 167–184.

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