Different expression patterns of heat shock proteins hsp 60 and hsp 70 in zebra mussels (Dreissena polymorpha) exposed to copper and tributyltin

Aquatic Toxicology 47 (2000) 213 – 226 www.elsevier.com/locate/aquatox

Different expression patterns of heat shock proteins hsp 60 and hsp 70 in zebra mussels (Dreissena polymorpha) exposed to copper and tributyltin Maureen E. Clayton 1, Roland Steinmann, Karl Fent * Swiss Federal Institute for En6ironmental Science and Technology (EAWAG) and Swiss Federal Institute of Technology (ETH), Ueberlandstrasse 133, CH-8600 Du¨bendorf, Switzerland Received 24 August 1998; received in revised form 18 February 1999; accepted 8 March 1999

Abstract To investigate the dose–response relationship of the expression of heat shock proteins hsp 60 and hsp 70 following exposure to environmentally-relevant contaminants, zebra mussels (Dreissena polymorpha) were exposed to sublethal concentrations of copper and tributyltin (TBT). Mussels were exposed to Cu (0 – 500 mg/l) or TBT (0 – 75 mg/l) for 24 h. Hsp 60 and hsp 70 expression relative to the controls was analyzed by western blotting and densitometry. Contaminant concentrations in the exposure medium and mussel tissues were measured by atomic absorption spectroscopy (AAS) or high resolution gas chromatography with flame photometric detection (HRGC-FPD) for Cu and TBT, respectively. Following copper exposure, hsp 60 showed a biphasic expression pattern, with a maximal expression of three times control levels at 22 mg Cu/g dry wt., while hsp 70 concentrations reached a plateau of approximately 2.5 times control levels after crossing an induction threshold at tissue concentrations of less than 29 mg Cu/g dry wt. In contrast, concentrations of both hsp 60 and hsp 70 were increased to approximately 2.5 – 3 times control levels in TBT-exposed mussels at all tested doses. The results of this study demonstrate that the nature of the dose–response curves depend on both the form of stress protein investigated (hsp 60 or hsp 70) and on the contaminant. The implications for the use of hsp 60 and hsp 70 as biomarkers in ecotoxicological research are discussed. © 2000 Elsevier Science B.V. All rights reserved. Keywords: Heat shock proteins; Hsp 60; Hsp 70; Copper; Tributyltin; Zebra mussels

* Corresponding author. Tel.: + 41-1-823-5332; fax: + 41-1-823-5028. E-mail address: [email protected] (K. Fent) 1 Current address. University of Northern Iowa, 2242 McCollum, Cedar Falls, IA 50614-0421, USA. 0166-445X/00/$ - see front matter © 2000 Elsevier Science B.V. All rights reserved. PII: S 0 1 6 6 - 4 4 5 X ( 9 9 ) 0 0 0 2 2 - 3

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1. Introduction Zebra mussels, Dreissena polymorpha, are an invasive species of exotic freshwater mussels which have become widely distributed in freshwater and slightly estuarine habitats throughout the world. Zebra mussels have a high filtration rate of up to more than 200 ml mussel − 1 h − 1 (Kraak et al., 1994a). As a result of high uptake and slow metabolism, they are able to bioconcentrate xenobiotic contaminants to very high levels (e.g. bioconcentration factors (BCF) for tributyltin (TBT) of approximately 12 000 – 260 000 on a wet weight basis (Fent and Hunn, 1991; Becker et al., 1992; Fent and Hunn, 1995). High tissue concentrations of these contaminants result in few or no apparent toxic effects to the mussels (Fent, 1996). D. polymorpha has also been shown to accumulate high levels of polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs) and heavy metals (Roper et al., 1996). Because of their ubiquitous distribution, high contaminant uptake, and relative lack of susceptibility to environmental contaminants, zebra mussels may be ecotoxicologically important indicator species. Zebra mussels have been used in biomonitoring studies to study contaminant levels in situ (e.g. Kraak et al., 1991; Becker et al., 1992; de Kock and Bowmer, 1993; Sta¨b et al., 1995; Mersch et al., 1996) and in studies of the sublethal effects of contaminants, primarily on filtration rates (e.g. Kraak et al., 1992, 1994a,b). Another possible use of zebra mussels in ecotoxicology is an investigation of sublethal contaminant effects utilizing molecular and biochemical biomarkers. Heat shock proteins (hsps), in particular hsp 60 and hsp 70, have been suggested as suitable biomarkers of the exposure to and effects of environmental contaminants (Sanders, 1990). The heat shock proteins are families of proteins which are classified by their molecular weight. They are also known as molecular chaperones (Ellis, 1987) for their constitutive roles in protein synthesis. Heat shock protein expression can also be induced by the presence of denatured proteins (Edington et al., 1989). Increased expression of heat shock proteins has also been called a stress response, because hsps can be increased or induced

after exposure to some environmentally-relevant stressors, including contaminants such as heavy metals (Cd, Cu, Cr, Hg, Ni, Pb and Zn), tributyltin, organophosphate and organochlorine pesticides, and other organic contaminants including benzene, 1-chloro-2,4-dinitrobenzene, 2,4-dichloroaniline, 2,4-dinitrobenzene, hexachlorobenzene, pentachlorophenol, and trichloroethylene (see review by Sanders, 1993). The common mode of action of these diverse stressors seems to be that they are all proteotoxic (Hightower, 1991), resulting in damage to proteins. Two ubiquitous environmental contaminants are copper (Cu) and tributyltin (TBT). Copper and TBT have both been used as the active ingredients in ship antifouling paints because of their high toxicities towards fouling organisms. TBT has even been described as ‘‘perhaps the most acutely toxic chemical…ever introduced to water’’ (Maguire, 1987). Due to its high toxicity to nontarget organisms, however, the use of TBT in antifouling paints has been banned or regulated in many countries. As a result, the relative use of copper-based antifouling paints is increasing. For example, as of 1995, most of the antifouling products authorized for use in Switzerland contained copper. Exposure to copper and tributyltin results in sublethal effects to aquatic organisms, particularly molluscs, including decreased growth rate, reproductive impairment, enzyme inhibition, reductions or alterations in protein and DNA synthesis, cytoskeletal alterations, and disruptions of ATP synthesis and Ca2 + homeostasis (see reviews by Viarengo, 1989; Fent, 1996; Langston, 1996). Both Cu and TBT have been shown to induce or increase hsp expression in the PLHC-1 fish cell culture line (Ryan, 1997), in the rotifer Brachionus plicatilis (Cochrane et al., 1991) and in the marine mussel Mytilus edulis (Steinert and Pickwell, 1988; Sanders et al., 1991; Steinert and Pickwell, 1993; Sanders et al., 1994). In addition to the laboratory-based studies on the induction of the stress response, attempts have been made to use hsps as biomarkers of contamination in situ (e.g. Sanders and Martin, 1993; Clayton, 1996; Ko¨hler and Eckwert, 1997). In some of these field studies, however, little or no

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relationship has been found between hsp concentration and known contaminant patterns (e.g. Pedersen et al., 1997). One possible explanation for this lack of relationship is that the expression of hsps only increases over a portion of the environmentally-relevant dose – response curve. This study investigates the dose – response relationship of the induction of the stress response (as measured by hsp 60 and hsp 70 concentrations) in mussels exposed to copper and tributyltin, environmentally-relevant contaminants in aquatic systems. Zebra mussels were exposed to Cu (0 – 500 mg/l) or TBT (0–75 mg/l) under controlled laboratory conditions. The concentrations of hsp 60 and hsp 70 in exposed mussels, relative to the controls, were related to the tissue uptake of the contaminants. The obtained dose – response curves should provide valuable information on contaminant-induced hsp expression patterns for the potential utilization of hsp 60 and hsp 70 as biomarkers.

2. Materials and methods

2.1. Reagents All reagents were of the highest quality available. CuSO4·5H2O ( \99%) and TBTCl ( \97%) were purchased from Merck (Darmstadt, Germany) and Fluka (Buchs, Switzerland), respectively, and were used without further purification. Purified recombinant human hsp 60 and hsp 70 were purchased from StressGen (Victoria, BC, Canada). A monoclonal antibody against recombinant human hsp 60 (clone LK-2) was purchased from StressGen. Various monoclonal antibodies against hsp 70 were purchased from commercial suppliers to test their crossreactivity with zebra mussel proteins. Monoclonal antibodies raised against human hsp 70 (clone C92F3A-5) and human hsp70/hsc70 (clone N27F3-4) were purchased from StressGen. Monoclonal antibodies against Drosophila hsp/hsc 70 (clone 7.10) and recombinant human hsp 70 (clones 3A3 and 5A5) were purchased from Affinity BioReagents (Golden, CO). A monoclonal antibody against purified bovine brain hsp 70 (clone BRM-22) was purchased from Sigma (St Louis, MO). Goat anti-rat

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IgG alkaline phosphatase conjugate secondary antibody was purchased from Sigma and goat antimouse IgG (H+L) alkaline phosphatase conjugate was purchased from BioRad (Hercules, CA).

2.2. Mussel collection and maintenance Zebra mussels, D. polymorpha, were collected by hand from a relatively uncontaminated (5.49 2.9 mg Cu/l, n=4 and 591 ng TBT/l, n= 3) part of the Glatt river near Du¨bendorf, Switzerland between April 1997 and February 1998. At the collection times, the river temperature varied between 4 and 23°C and pH between 7.3 and 8.0. The mussels were immediately transported to the laboratory and gradually acclimated to the experimental conditions. The mussels were maintained at 15°C for a minimum of 1 week prior to the experiments in a synthetic freshwater medium containing 1.5 mM CaCO3, 0.4 mM MgSO4·7H2O, 0.04 mM MgCl2·6H2O, 0.4 mM NaCl, 0.3 mM NaHCO3 and 0.15 mM KHCO3 in nanopure water (Sprung, 1987). The mussels were not fed during either the maintenance or experimental periods.

2.3. Experimental design All experiments were conducted in synthetic freshwater medium at 15°C. Each sample consisted of an aerated glass beaker containing 750 ml (for the tributyltin experiment) or 1 litre (for the copper experiment) medium and five mussels. Mussels were exposed for 24 h to different copper (0, 30, 40, 50, 60, 75, 100, 150, 200, 300, 400 and 500 mg Cu/l) or tributyltin (0, 1, 5, 10, 15, 20, 25, 30, 35, 40, 45, 50, 55 and 75 mg TBT/l) concentrations. The treatment stock solutions consisted of 1 g CuSO4·5H2O/l in nanopure water acidified to pH 1 with nitric acid or 0.5 mg TBTCl/ml in acetone. The controls were dosed with the carrier solvents. No effect of the carrier solvent on hsp induction was observed. For all of the Cu experiments and a subset of the TBT experiments, an aliquot of the medium was collected at the beginning and end of the experiments (T= 0 and 24 h) for chemical analy-

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sis. The medium samples were acidified to pH 1 with conc. HNO3 or to pH 2 with conc. HCl for Cu and TBT analysis, respectively, before storage for less than 1 week at 4°C. At the termination of the 24-h exposure period, the five mussels in each beaker were collected and mussel mortality was assessed. The dead mussels were discarded. The tissues of the live mussels were removed from the shells and the byssal threads were discarded. The mussel soft tissues from each beaker were pooled and weighed. An aliquot of each tissue sample was stored for less than 1 month for chemical analysis (at − 20°C for TBT and − 80°C for Cu). The remaining tissue was stored at −80°C and processed for heat shock protein analysis as soon as possible, generally within 2 days.

2.4. Chemical analysis Copper concentrations in medium and tissue samples were measured by atomic absorption spectroscopy (AAS) using a Varian AA-875 atomic absorption spectrophotometer equipped with a Varian GTA-95 graphite tube atomizer. When necessary, the medium samples were diluted with 1 M HNO3 in nanopure water prior to analysis, in order to fall within the range of the calibration curve. The tissue samples (8159260 mg) were lyophilized for 24 h before being subjected to microwave digestion (Kingston and Jassie, 1986) in the presence of 2 ml H2O, 1 ml H2O2 and 4 ml 15 M HNO3. The digested solution was diluted to 50 ml with nanopure water and analyzed by the same procedure as the medium samples. The detection limits of this analytical method were 0.5 mg Cu/l and 0.2 mg Cu/g dry wt. Mono-, di- and tributyltin compounds (MBT, DBT and TBT) were determined in medium and tissue samples as described previously (Fent and Hunn, 1995), with slight modifications. The medium samples (pH 2) were diluted as necessary to 100 ml with acidified nanopure water and spiked with tripentyl- and tripropyltin chloride as internal standards. The spiked samples were extracted twice with 0.25% tropolone in pentane/diethylether (2:3) and once with pentane. The

extracts were dried on anhydrous CaCl2 and concentrated under a gentle nitrogen stream. The extracted organotins were ethylated with a Grignard reagent (2 M ethylmagnesium bromide in tetrahydrofuran). The excess Grignard reagent was destroyed with 1 M HCl. The organic extract was then concentrated to a final volume of 2 ml. The tissue samples (1859 26 mg wet wt.) were acidified to pH 2, homogenized and spiked with internal standards before extraction and derivatization by the same procedure, with an additional final purification step of adsorption chromatography on silica gel. The tissue extracts were then concentrated to a final volume of 4 ml. Butyltin concentrations in the extracts were analyzed by a Carlo Erba HRGC 5160 gas chromatograph with split/splitless injection, a DB-5 column (0.32 mm internal diameter and 0.25 mm film) and flame photometric detection (SSD-250 detector). The results were corrected for the recovery rates of the internal standards in each sample. The average recovery rates of the internal standards were 60.59 16% (n =96) for medium samples and 70.99 31% (n= 91) for tissue samples. The detection limits of this analytical method were 0.07190.02 mg TBT/l (n=96) and 0.02390.009 mg TBT/g wet wt. (n= 91).

2.5. Heat shock protein determination The pooled mussel tissue samples from each beaker were processed for hsp analysis as described previously (Sanders et al., 1992). Briefly, the samples were homogenized in 66 mM Tris buffer (pH 7.5) containing 0.1% Nonidet, 0.1 mM PMSF and 0.1% aprotinin. The homogenates were centrifuged at 150 000×g for 90 min at 4°C and the supernatants were stored at − 80°C. The total protein concentration in the samples was determined by the Bradford protein assay method (Bradford, 1976), using bovine serum albumin as a standard. The concentrations of hsp 60 and hsp 70 in the mussel tissue samples were determined immunochemically by western blotting. Purified recombinant human hsp 60 and hsp 70 served as positive controls. Soluble proteins (200 or 10 mg total protein for hsp 60 and hsp 70, respectively) were

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separated on 12% SDS-PAGE gels (120 V for 75 min) and transferred onto 0.45-mm nitrocellulose membranes (100 V for 80 min at 4°C). The membranes were incubated for 1 h in blocking buffer (50 mM Tris, 200 mM NaCl, 0.2% Triton X-100, 3% BSA and 10% FCS). The blots were then incubated in primary antibody for 3 h at room temperature. For hsp 60 analysis, a monoclonal antibody produced against recombinant human hsp 60 (clone LK-2) was used at a 1:1000 dilution in blocking buffer. Several antibodies against hsp 70 were examined for their crossreactivity with D. polymorpha proteins with a 3-h incubation at room temperature at the suppliers’ recommended dilutions. For the results presented here, a monoclonal antibody produced against Drosophila hsp/ hsc 70 (clone 7.10) was used at a 1:2000 dilution in blocking buffer. Excess primary antibody was removed by washing three times with TBS (20 mM Tris, 0.5 M NaCl, pH 7.5), TBS/Tween (TBS+ 0.5% Tween 20) and TBS, respectively. After washing, the blots were incubated for 1 h with goat anti-mouse AP conjugate (1:1000 dilution) or for 1.5 h with goat anti-rat AP conjugate (1:10 000 dilution) as secondary antibodies for hsp 60 and hsp 70, respectively. After repetition of the washing steps, the blots were developed with 5-bromo4-chloro-3-indolyl phosphate and nitro blue tetrazolium (BCIP/NBT) for 10 or 15 min for hsp 60 and hsp 70, respectively. The blots were then rinsed well with distilled water and dried. The relative concentrations of hsps in the samples were determined by densitometry using a Molecular Dynamics Computing Densitometer equipped with Image Quant software. The gray scale values of the bands were computed. The results are expressed as a percentage of the value for the experimental control, which was also run on the same gel. The expressions of hsp 60 and hsp 70 in the controls were arbitrarily set to 100%.

tration is given in the results instead of the nominal initial medium concentration, as it is a more accurate representation of the actual Cu or TBT exposure throughout the experiment. For the chemical analysis and hsp data, the mean and standard deviations for each dose were computed from replicate samples. In a few cases in which the determined geometric mean exposure concentrations of Cu or TBT in the medium were not significantly different, the means and standard deviations were computed from all samples exposed to the same Cu or TBT geometric mean medium concentration, regardless of the nominal initial dose (Table 1). In addition, for the analysis of hsp concentration in the TBT experiments, samples in which the tissue TBT concentration was not significantly different were pooled into groups of control, low (0.3 mg/g wet wt.), medium (1.8–2.3 mg/g wet wt.), and high (4.3–4.9 mg/g wet wt.) TBT concentration, regardless of the TBT concentration in the medium. The means and standard deviations of tissue TBT concentration and hsp 60 and hsp 70 concentrations were determined for these pooled samples (Table 2). Samples were pooled to reduce the expected variability among individual mussels, as high variability appears to be a characteristic of hsp experiments (e.g. Fargnoli et al., 1990; Stringham and Candido, 1994; Yu et al., 1994; Werner and Nagel, 1997). Mean hsp 60 and hsp 70 concentrations in the mussels exposed to Cu or TBT were compared to the hsp concentrations in the controls by the nonparametric Wilcoxon signed rank test statistic, using SYSTAT 5 for Macintosh. A nonparametric test was used to avoid the assumption that the data are normally distributed. Significance was evaluated at the 0.05 level. The hsp concentrations were plotted against the tissue Cu or TBT concentrations to give the dose–response curve.

2.6. Data analysis

3. Results

Data from the medium chemical analyses at times 0 and 24 h were used to determine the geometric mean of the exposure Cu or TBT concentration. The geometric mean exposure concen-

3.1. Contaminant uptake Zebra mussels exposed to waterborne Cu and TBT for 24 h readily accumulated these contami-

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nants (Table 1). Low concentrations of Cu and TBT were found in the tissues of the control mussels (mean 9S.E.M.: 13 91 mg Cu/g dry wt., n = 14 and 0.059 0.009 mg TBT/g wet wt., n = 17). These values reflect the background levels of these contaminants in the natural Glatt river D. polymorpha population. Whole animal tissue Cu (Fig. 1a) and TBT (Fig. 1b) concentrations were highly correlated with exposure concentrations. In both cases, the relationship was non-linear, and uptake was fit with a logarithmic function (Cu: y= − 26.9+ 32 log(x); r 2 =0.84 and TBT: y = 1.06 +2.06 log(x); r 2 =0.88). Both Cu and TBT contaminant uptake leveled off at the highest doses tested; this effect was most noticeable in the TBT experiments. Mono- and dibutyltin, potential degradation products of TBT, were occasionally observed in the medium and tissue samples from the TBT experiments at levels at or near the detection limits (medium: 569 20 ng MBT/l and 63921 ng

DBT/l, n= 96; tissue: 209 9 ng MBT/g wet wt. and 229 10 ng DBT/g wet wt., n= 92). No relationship between MBT or DBT and TBT concentration was observed, however, nor was any trend in MBT or DBT concentration detected over the 24-h experimental duration. No apparent toxicity was observed as a result of contaminant exposure or uptake. The mean mortality rate for all experiments, including all treatments as well as the controls, was low (79 14%, n= 60 and 99 15%, n= 89 for the Cu and TBT experiments, respectively). Mussel mortality was not correlated with medium contaminant concentrations.

3.2. Hsp antibody crossreacti6ity The monoclonal antibody against recombinant human hsp 60 used in this study (clone LK-2) crossreacted with D. polymorpha proteins to pro-

Table 1 Copper and TBT concentrations in the medium and zebra mussel tissue samplesa Initial concentration (nominal; mg/l)

Exposure concentration (geometric mean; mg/l)

Tissue concentration (mg/g)

n

Copper 0 30, 40 50, 60, 75 100 150 200 300 400 500

0.49 0.1 18.59 2.4 31.99 2.6 79.0 9 14.1 113.29 6.9 147.4 9 2.7 188.79 5.1 267.09 4.3 345.7 9 7.0

13.2 9 1.4 16.8 9 0.4 22.2 9 1.7 28.8 9 5.5 42.1 9 14.3 32.1 9 3.3 48.9 9 4.6 46.4 9 3.0 63.3 9 7.1

14 4 9 7 5 6 5 5 5

TBT 0 1 5 10 15, 20 25, 30, 35 40, 45 50, 55 75

0.059 0.02 0.309 0.01 2.19 0.2 6.09 1.0 13.59 0.9 22.59 1.1 35.89 1.8 46.69 2.6 68.39 3.5

0.05 9 0.01 0.30 9 0.01 1.8 9 0.1 2.3 9 0.2 2.2 9 0.3 4.7 9 0.3 4.3 9 0.7 4.8 9 0.6 4.9 9 1.4

17 7 4 7 16 15 9 9 5

a The nominal initial medium concentration is expressed as mg/l Cu or TBT. The mean exposure concentration is the geometric mean of the medium Cu or TBT concentrations at the beginning and end of the experiments (T= 0 and 24 h) and is expressed as mg/l Cu or TBT (mean 9 S.E.M.). The concentration of Cu or TBT in the zebra mussel tissue samples is expressed as mg Cu/g dry wt. or mg TBT/g wet wt. (mean9 S.E.M.).

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Table 2 Concentrations of heat shock proteins hsp 60 and hsp 70 in zebra mussels exposed to copper or tributyltina Tissue concentration (mg/g)

Hsp 60 concentration (%)

Hsp 70 concentration (%)

n

Copper 0.49 0.1 16.89 0.4 22.29 1.7 28.89 5.5 42.19 14.3 32.19 3.3 48.99 4.6 46.49 3.0 63.39 7.1

1009 7 819 14 3069 76* 619 18 1139 52 979 20 1079 23 529 20* 429 14*

100 9 15 86 9 4 103 9 17 249 9 82* 251 9 87* 199 9 41* 286 9 80* 242 9 74* 265 9 69*

14 4 9 7 5 6 5 5 5

TBT 0.059 0.01 0.309 0.04 2.19 0.2 4.79 0.3

1009 9 2219 151 2629 61* 3319 74*

100 9 12 251 9 66* 239 9 38* 323 9 49**

17 7 27 38

a Tissue contaminant concentrations are expressed as mg Cu/g dry wt. or mg TBT/g wet wt. (mean 9 S.E.M.). Hsp 60 and hsp 70 concentrations are expressed as a percentage of the experimental control (mean 9 S.E.M.), with the control value set arbitrarily to 100%. Hsp concentrations which are significantly different from the control are indicated. * PB0.05. ** PB0.01.

duce one band (Fig. 2a) of slightly lower molecuar weight than the positive control (recombinant human hsp 60). In a preliminary study for crossreactivity with D. polymorpha proteins, several monoclonal antibodies against various forms of hsp 70 were examined. At antibody concentrations and incubation times at or above those recommended in the supplier’s recommended protocols, and with concentrations of up to 200 mg total protein from D. polymorpha samples, five of these antibodies (clones BRM-22, 3A3, 5A5, C92F3A-5 and N27F3-4) failed to produce distinct bands with control, heat shocked, or Cu or TBT treated mussel samples. A monoclonal antibody produced against Drosophila hsp/hsc 70 (clone 7.10), however, crossreacted with D. polymorpha proteins, producing either one band or a doublet (Fig. 2b) of slightly greater molecular weight than the positive control (recombinant human hsp 70). This antibody was employed in the experiments described in this paper.

3.3. Hsp 60 and 70 expression in Cu-exposed mussels Hsp 60 expression patterns were essentially similar when plotted against either tissue (Fig. 3a) or medium (data not shown) Cu concentration. An apparent threshold for hsp 60 induction in Cu-exposed mussels was observed at the lowest tested dose (17 mg Cu/g dry wt. or 19 mg Cu/l), where hsp 60 values were not significantly different than those of the controls (Table 2, Fig. 3a). After crossing this threshold Cu concentration, hsp 60 expression showed a significant peak at one dose, corresponding to approximately 22 mg Cu/g dry wt. or 32 mg Cu/l. At this dose, hsp 60 expression was more than three times that in control mussels. Following this hsp 60 peak, expression returned to near control levels. At the two highest medium concentrations (267 and 345 mg Cu/l or 46.4 and 63.3 mg Cu/g dry wt.), however, hsp 60 expression dropped to approximately 50% of the control value (Table 2).

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At the two lowest doses tested, hsp 70 expression in Cu-exposed mussels was not significantly different from the expression in control mussels (Fig. 3b). This threshold was larger than that observed for hsp 60 (Table 2). After this threshold, at tissue concentrations of more than 22 mg Cu/g dry wt., however, hsp 70 expression was significantly elevated over control values. Unlike the case for hsp 60, the elevated expression of hsp 70 continued up to the highest tested doses. There was, however, no significant relationship between dose and hsp 70 expression at these Cu tissue concentrations; mean hsp 70 expression in all samples (up to 63 mg Cu/g dry wt.) was approximately 2.5 times control expression. Hsp 70 expression patterns in relation to medium Cu

Fig. 1. Cu and TBT uptake in zebra mussel (D. polymorpha) tissue after 24 h of exposure. Mean (9S.E.M.) tissue Cu (a) and TBT (b) concentrations are plotted against the geometric mean ( 9 S.E.M.) of the contaminant concentrations measured in the medium at T= 0 and 24 h. The numbers of replicates are given in Table 1. Contaminant uptake was fitted with a logarithmic function (for Cu: y= − 26.9+ 32 log(x); r 2 = 0.89 and for TBT: y = 1.06 + 2.06 log(x); r 2 = 0.88).

Fig. 2. Western blots for hsp 60 (a) and hsp 70 (b). Lanes give three samples of different copper and TBT exposures including one respective control sample. Molecular weight standards (210, 111, 83, 46 kDa) are in lane 1 on the left, control samples in lane 2 and 6, samples from the Cu experiments in lanes 3 – 5, samples from the TBT experiments in lanes 7 – 9 and the positive control in lane 10 (30 ng recombinant human hsp 60 or 200 ng recombinant human hsp 70). Representative samples were chosen to include control, low, medium and high nominal exposure concentrations (30, 60, 100 mg Cu/l for hsp 60, 30, 200, 300 mg Cu/l for hsp 70; 1, 20, 50 mg TBT/l for hsp 60 and hsp 70). For hsp 60 (a), the Cu samples in lanes 2 – 5 had tissue concentrations of 12, 17, 27 and 49 mg Cu/g dry wt., respectively. The TBT samples in lanes 6 – 9 had tissue concentrations of 0.04, 0.5, 2.2 and 5.7 mg TBT/g wet wt., respectively. For hsp 70 (b), the Cu samples in lanes 2 – 5 had tissue concentrations of 13, 16, 24 and 53 mg Cu/g dry wt., respectively. The TBT samples in lanes 6 – 9 had tissue concentrations of 0, 0.3, 1.5 and 4.0 mg TBT/g wet wt., respectively.

concentration were essentially similar to those with respect to tissue Cu concentration, with the

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threshold ending at approximately 32 mg Cu/l (data not shown).

3.4. Hsp 60 and 70 expression in TBT-exposed mussels Concentrations of both hsp 60 (Fig. 4a) and hsp 70 (Fig. 4b) were increased two to three times over the expression in the control mussels at all tested TBT doses (0.3 – 68 mg/l and 0.3 – 4.9 mg/g wet wt.). This difference was statistically significant in all cases, except for hsp 60 concentration at the lowest tested dose (Table 2). In addition to the increased expression compared to the controls, there was a trend towards an increase in expression with increasing dose, as measured by a

Fig. 4. Hsp 60 (a) and hsp 70 (b) expression in TBT-exposed mussels. Hsp 60 and hsp 70 concentrations were increased relative to the experimental control at all doses. Mean ( 9 S.E.M.) hsp expression (as a percentage of the experimental control) is plotted against the tissue TBT concentration in mg/g wet wt. (mean 9 S.E.M.). The numbers of replicates are given in Table 2. Significant elevations in hsp expression compared to the controls are noted (*PB0.05, **PB 0.01). A trend towards increasing hsp concentration with increasing dose was observed after linear regression analysis on the treated samples, which produced equations of y = 212 + 25.3x (r 2 =0.997) for hsp 60 and y = 230 + 17.5x (r 2 =0.72) for hsp 70.

Fig. 3. Hsp 60 (a) and hsp 70 (b) expression in Cu-exposed mussels. Mean ( 9S.E.M.) hsp expression (as a percentage of the experimental control) is plotted with respect to tissue Cu concentration in mg/g dry wt. (mean9 S.E.M.). The numbers of replicates are given in Table 2. Hsp concentrations which were significantly different from the controls are noted (*PB 0.05).

least squares linear regression analysis. The regression equations were y=212+ 25.3x (r 2 = 0.997) for hsp 60 and y= 230+ 17.5x (r 2 = 0.72) for hsp 70. Essentially similar relationships were seen when the hsp concentrations were plotted against the TBT concentration in the medium (data not shown). When hsp expression was plotted against medium TBT concentration in mg/l, the regression equations were y=230+2.75x for hsp 60 (r 2 = 0.96) and y=238+2.18x for hsp 70 (r 2 = 0.91).

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4. Discussion The tissue concentrations of hsp 60 and hsp 70 are altered in Cu- and TBT-exposed mussels. While there is relatively high variability among individuals, significant differences from the controls were detected when replicate samples were pooled. The shape of the dose – response curve is crucial for understanding the regulation of heat shock protein expression after exposure to environmental contaminants. For the most part, increasing tissue concentrations of Cu or TBT result in increasing concentrations of hsp 60 and hsp 70 in the exposed mussels. The nature of the dose – response curve, however, depends on both the form of hsp (hsp 60 or hsp 70) and the nature of the stressor (Cu or TBT).

4.1. Dose–response cur6es Three important features of any dose – response curve are: (1) the presence or absence of a threshold, or dose below which no effects are seen, (2) whether the observed response increases monotonically, or if the curve is biphasic, and (3) the presence or absence of a plateau effect at the highest doses, in which the response is maintained at a constant level regardless of the dose applied. The features of the dose – response curves for hsp 60 and hsp 70 in Cu- and TBT-exposed D. polymorpha are discussed in this section. With respect to feature (1), a threshold for increased expression of hsps seems to occur for mussels exposed to copper, but not to TBT, at least within the range of TBT concentrations examined in this study. The threshold seems to be larger for hsp 70 expression than for hsp 60 expression. Contaminant thresholds below which no increase in hsp expression is detectable have been observed in previous studies of hsp expression (e.g. Hatayama et al., 1991; Sanders et al., 1991; Ryan and Hightower, 1994; Sanders et al., 1994; Zanger et al., 1996; Eckwert et al., 1997), including experiments in which copper was the stressor. The absence of a threshold after TBT exposure was also observed in Mytilus edulis (Steinert and Pickwell, 1993).

With respect to feature 2, monotonic increases of hsp 60 and hsp 70 with increasing TBT concentrations were observed in the current study. Hsp 70 concentrations in Cu-exposed mussels also increased before reaching a maximum. The concentration of hsp 60 in Cu-exposed mussels, however, peaked at approximately 22 mg Cu/g dry wt., and then returned to control or lower levels with increasing tissue concentrations. A similar effect, without the pronounced decrease to control levels, was observed for hsp 60 concentrations in rotifers exposed to Cu (Cochrane et al., 1991), while this effect was not observed in copper-exposed M. edulis (Sanders et al., 1991, 1994). A biphasic hsp 60 dose–response curve has also been detected in the amphipods Hyalella azteca and Rhepoxynius abronius after dieldrin exposure and in Ampelisca abdita and R. abronius after fluoranthene exposure (Werner and Nagel, 1997) and has been proposed as a general model of hsp expression, at least in isopods exposed to heavy metals (Eckwert et al., 1997). In the first case, the phenomenon was linked to hormesis, or the induction of beneficial effects by low doses of contaminants which are otherwise harmful (Werner and Nagel, 1997). In the second case, the decrease in hsps at high metal concentrations was attributed to pathological damage and inhibition of protein synthesis in the cell (Eckwert et al., 1997). Neither of these hypotheses are fully consistent with our results for hsp 60 expression in Cutreated mussels, however, as a hormetic response would theoretically be observed for both hsp 60 and hsp 70, and inhibition of protein synthesis at high tissue Cu concentrations is not likely as hsp 70 concentrations remain elevated above control levels. Further research is necessary to verify the exact shape of the hsp 60 expression curve in the critical range of Cu concentrations, and to explain why this curve differs from that of hsp 70 expression after Cu exposure or hsp 60 and hsp 70 expression patterns induced by TBT exposure. As hsp 60 and hsp 70 are found in different cellular compartments (primarily mitochondrial and cytosolic, respectively), and have different physiological roles, we might expect them to have different expression patterns. Hsp 60 apparently stabilizes proteins during intermediate folding

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stages and provides a surface on which newly synthesized proteins can fold (Martin et al., 1991). During cellular stress, hsp 70 appears to bind to exposed hydrophobic surfaces on denatured proteins, preventing aggregation (Pelham, 1986) and is involved in targeting proteins for lysosomal degradation (Chiang et al., 1989). Furthermore, members of the hsp 70 family appear to be essential for the refolding of denatured proteins (Gaitanaris et al., 1990; Skowyra et al., 1990). It is unclear, however, why the observed hsp 60 and hsp 70 expression patterns differ after Cu exposure, but not after TBT exposure, as both contaminants appear to share some similar cellular and subcellular effects (see reviews by Viarengo, 1989; Stohs and Bagchi, 1995; Fent, 1996; Langston, 1996). With respect to feature (3), the dose – response curves for hsp 60 and hsp 70 in Cu-exposed mussels both showed a plateau (hsp 70) or lack of relationship to dose (hsp 60). The hsp 70 levels remained elevated at the highest doses while the hsp 60 levels returned to concentrations at or below the control. In contrast, there was a slight trend with TBT exposure towards a continuous increase in levels of hsp 60 and hsp 70, even at the highest tested doses. Our results do not, however, indicate whether this trend would be maintained at TBT concentrations greater than those tested or whether a plateau would be reached as for hsp 70 after Cu exposure. In practice, however, such large concentrations of Cu or TBT are likely to be found only at severely contaminated sites, so that this portion of the dose – response curve is outside the general range of environmentally-relevant Cu and TBT concentrations.

4.2. En6ironmental rele6ance of exposure concentrations The medium concentrations of Cu and TBT to which mussels were exposed in this experiment were quite high, and the 24-h exposure period only allowed for the determination of the effects of acute exposures to these contaminants. The tissue concentrations of these contaminants, however, were within the range of environmentally-occurring concentrations in zebra mussels, especially

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in polluted habitats. For example, concentrations of Cu in tissues of D. polymorpha have been found to be 30–240 mg/g dry wt. in contaminated reservoirs in France (Mersch et al., 1996) and 13–170 mg/g dry wt. at various locations along the Seine river (Chevreuil et al., 1996). In this study, Cu-exposed mussels accumulated 17–63 mg/g dry wt. Specimens of D. polymorpha from harbors in Swiss lakes have been found to contain 0.2–9.35 mg TBT/g wet wt. (Fent and Hunn, 1991; Becker et al., 1992; Fent and Hunn, 1995), while in the present study tissue TBT concentrations in the treated mussels ranged from 0.3 to 4.9 mg/g wet wt. Further experiments, however, will be required to determine whether the effects observed in this study of rapid contaminant uptake from concentrated solutions are applicable to field conditions where mussels have accumulated similar levels of contaminants from chronic exposure to lower water contaminant concentrations.

4.3. Implications for use as biomarkers The results of this study have some potential implications for the utilization of hsp 60 and hsp 70 as biomarkers in ecotoxicological research. First, the problem of pronounced individual variability in hsp expression can be overcome by the statistical analysis of the means of replicate samples, and by relating the hsp concentration to tissue contaminant concentrations. Secondly, the dose–response relationship of hsp expression vs. tissue contaminant concentrations is dependent upon both the nature of the stressor and the specific hsp investigated, suggesting that dose–response curves will have to be determined separately for each stressor and hsp of interest. Hsp 60 expression is not expected to be a good biomarker in Cu-treated D. polymorpha due to the biphasic nature of the dose–response curve. Furthermore, Cu-exposed mussels have altered hsp expressions only after a contaminant threshold has been crossed, suggesting that significant differences in hsp expression may only be observed in populations from highly contaminated sites. The expression of hsp 70 in Cu-exposed mussels and hsp 60 and hsp 70 after TBT treatment are best described as providing a yes or no response,

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indicating whether or not the mussels have been exposed to contaminant concentrations with the potential to induce a stress response. This hsp expression pattern can, therefore, serve as a biomarker of contaminant effect, as hsp concentrations are believed to increase in response to the accumulation of damaged proteins in the cell (Ananthan et al., 1986; Parsell and Sauer, 1989). Further research is needed, however, to determine whether the hsp expression patterns observed in D. polymorpha extend to other ecologically relevant species, as the relative insensitivity of zebra mussels to contaminant effects might be related to a species-specific protective increase in heat shock protein expression, in addition to other factors.

5. Conclusion Hsp 60 and hsp 70 expression patterns in Cuand TBT-exposed zebra mussels are hsp familyand contaminant-specific. An induction threshold for the stress response was observed following Cu, but not TBT, exposure. The dose – response curve for hsp 60 expression in Cu-treated mussels is biphasic, with a return to control or lower levels after a maximum expression of three times control levels. In contrast, hsp 60 and hsp 70 levels were elevated at all TBT concentrations, and hsp 70 concentration increased in Cu-exposed mussels after the induction threshold. To the best of our knowledge, this is the first report of an induction of the stress response in D. polymorpha. While further research is required to determine if the results of this study are also applicable to other species and after chronic exposure to lower contaminant concentrations resulting in similar tissue contaminant concentrations, some implications for the use of hsp 60 and hsp 70 as biomarkers in ecotoxicological research are suggested. It appears that hsp 60 and hsp 70 may have different expression patterns, and that the dose – response relationship will have to be determined for each environmentally-relevant contaminant, as relevant features of the relationship, including the presence or absence of thresholds and plateaus, as well as the general shape of the curve, are not consistent

among contaminants. Furthermore, contaminantspecific expression patterns should be considered in the design of experiments to study the effects of contaminant mixtures on hsp expression. Rigorous analysis of the hsp dose–response relationships of environmentally-relevant contaminants should precede the use of hsps as biomarkers in field studies.

Acknowledgements We would like to thank Peter Looser for assistance with the organotin analyses and mussel collections and David Kistler for assistance with the copper analyses.

References Ananthan, J., Goldberg, A.L., Voellmy, R., 1986. Abnormal proteins serve as eukaryotic stress signals and trigger the activation of heat shock genes. Science 232, 522 – 524. Becker, K., Merlini, L., de Bertrand, N., de Alencastro, L.F., Tarradellas, J., 1992. Elevated levels of organotins in Lake Geneva: Bivalves as sentinel organism. Bull. Environ. Contam. Toxicol. 48, 37 – 44. Bradford, M., 1976. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Anal. Biochem. 72, 248 – 254. Chevreuil, M., Blanchard, M., Teil, M.-J., Carru, A.-M., Testard, P., Chesterikoff, A., 1996. Evaluation of the pollution by organochlorinated compounds (polychlorinated biphenyls and pesticides) and metals (Cd, Cr, Cu and Pb) in the water and in the zebra mussel (D. polymorpha Pallas) of the river Seine. Water Air Soil Pollut. 88, 371 – 381. Chiang, H., Terlecky, S.R., Plant, C.P., Dice, J.F., 1989. A role for a 70-kilodalton heat shock protein in lysosomal degradation of intracellular proteins. Science 246, 382 – 385. Clayton, M.E., 1996. Lipoproteins and Heat Shock Proteins as Measures of Reproductive Physiology in the Soft Shell Clam Mya arenaria. Ph.D. Dissertation, MIT/WHOI. Cochrane, B.J., Irby, R.B., Snell, T.W., 1991. Effects of copper and tributyltin on stress protein abundance in the rotifer Brachionus plicatus. Comp. Biochem. Physiol. 98C, 385 – 390. de Kock, W.C., Bowmer, C.T., 1993. Bioaccumulation, biological effects, and food chain transfer of contaminants in the zebra mussel (Dreissena polymorpha). In: Nalepa, T.F., Schloesser, D.W. (Eds.), Zebra Mussels: Biology, Impacts, and Control. Lewis, Boca Raton, FL, pp. 503 – 533.

M.E. Clayton et al. / Aquatic Toxicology 47 (2000) 213–226 Eckwert, H., Alberti, G., Ko¨hler, H.-R., 1997. The induction of stress proteins (hsp) in Oniscus asellus (Isopoda) as a molecular marker of multiple heavy metal exposure: I. Principles and toxicological assessment. Ecotoxicology 6, 249 – 262. Edington, B.V., Whelan, S.A., Hightower, L.E., 1989. Inhibition of heat shock (stress) protein induction by deuterium oxide and glycerol: additional support for the abnormal protein hypothesis of induction. J. Cell. Physiol. 139, 219– 228. Ellis, R.J., 1987. Proteins as molecular chaperones. Nature 328, 378 – 379. Fargnoli, J., Kunisada, T., Fornace, A.J., Schneider, E.L., Holbrook, N.J., 1990. Decreased expression of heat shock protein 70 mRNA and protein after heat treatment in cells of aged rats. Proc. Natl. Acad. Sci. 87, 846–850. Fent, K., 1996. Ecotoxicology of organotin compounds. Crit. Rev. Toxicol. 26, 1 – 117. Fent, K., Hunn, J., 1991. Phenyltins in water, sediment, and biota of freshwater marinas. Environ. Sci. Technol. 25, 956 – 963. Fent, K., Hunn, J., 1995. Organotins in freshwater harbors and rivers: temporal distribution, annual trends and fate. Environ. Toxicol. Chem. 14, 1123–1132. Gaitanaris, G.A., Papavassiliou, A.G., Rubock, P., Silverstein, S.J., Gottesman, M.E., 1990. Renaturation of denatured lambda repressor requires heat shock proteins. Cell 61, 1013 – 1020. Hatayama, T., Tsukimi, Y., Wakatsuki, T., Kitamura, T., Imahara, H., 1991. Different induction of 70,000-Da heat shock protein and metallothionein in HeLa cells by copper. J. Biochem. 110, 726 –731. Hightower, L.E., 1991. Heat shock, stress proteins, chaperones, and proteotoxicity. Cell 66, 191–197. Kingston, H.M., Jassie, L.B., 1986. Microwave energy for acid decomposition at elevated temperatures and pressures using biological and botanical samples. Anal. Chem. 58, 2534 – 2541. Ko¨hler, H.-R., Eckwert, H., 1997. The induction of stress proteins (hsp) in Oniscus asellus (Isopoda) as a molecular marker of multiple heavy metal exposure. II: Joint toxicity and transfer to field situations. Ecotoxicology 6, 263–274. Kraak, M.H.S., Scholten, M.C.T., Peeters, W.H.M., de Kock, W.C., 1991. Biomonitoring of heavy metals in the western European rivers Rhine and Meuse using the freshwater mussel Dreissena polymorpha. Environ. Pollut. 74, 101– 114. Kraak, M.H.S., Lavy, D., Peeters, W.H.M., Davids, C., 1992. Chronic ecotoxicity of copper and cadmium to the zebra mussel Dreissena polymorpha. Arch. Environ. Contam. Toxicol. 23, 363 – 369. Kraak, M.H.S., Toussaint, M., Lavy, D., Davids, C., 1994a. Short-term effects of metals on the filtration rate of the zebra mussel Dreissena polymorpha. Environ. Pollut. 84, 139 – 143. Kraak, M.H.S., Lavy, D., Schoon, H., Toussaint, M., Peeters, W.H.M., van Straalen, N.M., 1994b. Ecotoxicity of mix-

225

tures of metals to the zebra mussel Dreissena polymorpha. Environ. Toxicol. Chem. 13, 109 – 114. Langston, W.J., 1996. Recent developments in TBT ecotoxicology. Toxicol. Ecotox. News 3, 179 – 187. Maguire, R.J., 1987. Environmental aspects of tributyltin. Appl. Organomet. Chem. 1, 475 – 498. Martin, J., Langer, T., Boteva, R., Schramel, A., Horwich, A.L., Hartl, F.-U., 1991. Chaperonin-mediated protein folding at the surface of groEL through a ‘molten globule’like intermediate. Nature 352, 36 – 42. Mersch, J., Wagner, P., Pihan, J.-C., 1996. Copper in indigenous and transplanted zebra mussels in relation to changing water concentrations and body weight. Environ. Toxicol. Chem. 15, 886 – 893. Parsell, D.A., Sauer, R.T., 1989. Induction of a heat shocklike response by unfolded protein in Escherichia coli: dependence on protein level not protein degradation. Gene. Dev. 3, 1226 – 1232. Pedersen, S.N., Lundebye, A.-K., Depledge, M.H., 1997. Field application of metallothionein and stress protein biomarkers in the shore crab (Carcinus maenas) exposed to trace metals. Aquat. Toxicol. 37, 183 – 200. Pelham, H.R.B., 1986. Speculations on the functions of the major heat shock and glucose-regulated proteins. Cell 46, 959 – 961. Roper, J.M., Cherry, D.S., Simmers, J.W., Tatem, H.E., 1996. Bioaccumulation of toxicants in the zebra mussel, Dreissena polymorpha, at the Times Beach Confined Disposal Facility, Buffalo, New York. Environ. Pollut. 94, 117 – 129. Ryan, J.A., 1997. Fish stress proteins as in 6itro biomarkers. Ph.D. Dissertation, University of Connecticut. Ryan, J.A., Hightower, L.E., 1994. Evaluation of heavy-metal ion toxicity in fish cells using a combined stress protein and cytotoxicity assay. Environ. Toxicol. Chem. 13, 1231 – 1240. Sanders, B.M., 1990. Stress proteins: potential as multitiered biomarkers. In: McCarthy, J.F., Shugart, L.R. (Eds.), Biomarkers of Environmental Contamination. Lewis, Boca Raton, FL, pp. 165 – 191. Sanders, B.M., 1993. Stress proteins in aquatic organisms: an environmental perspective. Crit. Rev. Toxicol. 23, 49 – 75. Sanders, B.M., Martin, L.S., 1993. Stress proteins as biomarkers of contaminant exposure in archived environmental samples. Sci. Total Environ. 139/140, 459 – 470. Sanders, B.M., Martin, L.S., Nelson, W.G., Phelps, D.K., Welch, W.J., 1991. Relationships between accumulation of a 60 kDa stress protein and scope-for-growth in Mytilus edulis exposed to a range of copper concentrations. Mar. Environ. Res. 31, 81 – 97. Sanders, B.M., Pascoe, V.M., Nakagawa, P.A., Martin, L.S., 1992. Persistence of the heat-shock response over time in a common Mytilus mussel. Mol. Mar. Biol. Biotech. 1, 147 – 154. Sanders, B.M., Martin, L.S., Howe, S.R., Nelson, W.G., Hegre, E.S., Phelps, D.K., 1994. Tissue-specific differences in accumulation of stress proteins in Mytilus edulis exposed to a range of copper concentrations. Toxicol. Appl. Pharmacol. 125, 206 – 213.

226

M.E. Clayton et al. / Aquatic Toxicology 47 (2000) 213–226

Skowyra, D., Georgopoulos, C., Zylicz, M., 1990. The E. coli dnaK gene product, the hsp70 homolog, can reactivate heat-inactivated RNA polymerase in an ATP hydrolysis-dependent manner. Cell 62, 939–944. Sprung, M., 1987. Ecological requirements of developing Dreissena polymorpha eggs. Arch. Hydrobiol. Suppl. 79, 69–86. Sta¨b, J.A., Frenay, M., Freriks, I.L., Brinkman, U.A.T., Cofino, W.P., 1995. Survey of nine organotin compounds in the Netherlands using the zebra mussel (Dreissena polymorpha) as biomonitor. Environ. Toxicol. Chem. 14, 2023–2032. Steinert, S.A., Pickwell, G.V., 1988. Expression of heat shock proteins and metallothionein in mussels exposed to heat stress and metal ion challenge. Mar. Environ. Res. 24, 211 – 214. Steinert, S.A., Pickwell, G.V., 1993. Induction of HSP70 proteins in mussels by ingestion of tributyltin. Mar. Environ. Res. 35, 89 – 93. Stohs, S.J., Bagchi, D., 1995. Oxidative mechanisms in the toxicity of metal ions. Free Radic. Biol. Med. 18, 321–336.

.

Stringham, E.G., Candido, E.P.M., 1994. Transgenic hsp 16lacZ strains of the soil nematode Caenorhabditis elegans as biological markers of environmental stress. Environ. Toxicol. Chem. 13, 1211 – 1220. Viarengo, A., 1989. Heavy metals in marine invertebrates: mechanisms of regulation and toxicity at the cellular level. Rev. Aquat. Sci. 1, 295 – 317. Werner, I., Nagel, R., 1997. Stress proteins HSP60 and HSP70 in three species of amphipods exposed to cadmium, diazinon, dieldrin and fluoranthene. Environ. Toxicol. Chem. 16, 2393 – 2403. Yu, Z., Magee, W.E., Spotila, J.R., 1994. Monoclonal antibody ELISA test indicates that large amounts of constitutive hsp-70 are present in salamanders, turtle and fish. J. Therm. Biol. 19, 41 – 53. Zanger, M., Alberti, G., Kuhn, M., Ko¨hler, H.-R., 1996. The stress-70 protein family in diplopods: induction and characterization. J. Comp. Physiol. B 165, 622 – 627.