Subchronic organismal toxicity, cytotoxicity, genotoxicity, and feeding ...

Environmental Toxicology and Chemistry, Vol. 26, No. 10, pp. 2192–2197, 2007 䉷 2007 SETAC Printed in the USA 0730-7268/07 $12.00 ⫹ .00

SUBCHRONIC ORGANISMAL TOXICITY, CYTOTOXICITY, GENOTOXICITY, AND FEEDING RESPONSE OF PACIFIC OYSTER (CRASSOSTREA GIGAS) TO LINDANE (␥-HCH) EXPOSURE UNDER EXPERIMENTAL CONDITIONS GERARDO ANGUIANO,† RAUL LLERA-HERRERA,† EMILIO ROJAS,‡ and CELIA VAZQUEZ-BOUCARD*† †Department of Environmental Management and Conservation, Toxicogenomic and Reproduction of Aquatic Organism Laboratory (CIBNOR), P.O. 128, La Paz, Baja California Sur, Me´xico ‡Department of Genomic and Environmental Toxicology Medicine. Toxicogenomic Laboratory. IIB-UNAM. Ciudad de Me´xico, Me´xico ( Received 24 July 2006; Accepted 14 May 2007) Abstract—This study evaluated organismal toxicity, cytotoxicity, and genotoxicity and the filtration rate in response to different concentrations of subchronic lindane (gamma-hexachlorocyclohexane [␥-HCH]), exposure (12 d) in adult Pacific oysters Crassostrea gigas. Oysters were exposed in vivo in laboratory aquaria to 10 different concentrations (0.0–10.0 mg/L) of ␥-HCH. The median lethal concentration (LC50) after 12 d was calculated as 2.22 mg/L. Cytotoxic effects were observed in hemocytes, where the mean cell viability was significantly decreased at 1.0 mg/L of ␥-HCH after 12 d. Genotoxicity of ␥-HCH measured by single cell gel electrophoresis assay, in hemocytes was evident at 0.7 mg/L of ␥-HCH after 12 d. After 4 h of exposure to ␥-HCH, filtration rates were reduced compared with controls to 65.8 and 38.2% at concentrations of 0.3 and 0.7 mg/L, respectively, and after 11 d of exposure, filtration rates were reduced to 60.4 and 30.9% at concentrations of 0.1 mg/L and higher. These results show the subchronic effects of ␥-HCH at different concentrations and effect sensitivities are categorized as filtration rate ⬍ genotoxicity ⬍ cytotoxicity ⬍ mortality. The relevance of integral toxicity evaluation, considering different endpoints from molecular, cellular, and individual levels is discussed. Keywords—Lindane

Crassostrea gigas

Comet assay

Biomarkers

Stress response

are the inhibition of acetylcholinesterase activity, increased concentrations of cyclic adenosine monophosphate (cAMP), increased metallothionein levels, increased lipid content in the digestive gland, and induction of oxidative stress [12–14]. Under experimental conditions, there is no effect on glutathioneS-transferase synthesis [15]. There are no reports about biochemical or physiological effects of ␥-HCH on Crassostrea gigas. In the present study, we determined the toxic, cytotoxic, and genotoxic effects of ␥-HCH on C. gigas. This species was selected because it is widely used in environmental toxicology for the characterization of toxic effects by chemical compounds and as sentinel species in environmental monitoring programs [16]. Furthermore, C. gigas is an important species in the aquaculture and fishery international market [17].

INTRODUCTION

The use of pesticides in agriculture affects human health and produces environmental risks by biomagnification in trophic nets and bioaccumulation in organisms for human consumption. Although pesticides of low persistence and accumulation have been developed, the use of highly persistent and bioaccumulating chemicals, like the organochlorine pesticides (DDT, lindane, etc.), is still common in developing countries [1,2]. In Mexico, several coastal zones in the Gulf of Mexico and the Pacific Ocean exhibited detectable concentrations of lindane in water, sediments, and organisms [3,4]. Hexachlorocyclohexane is an organochlorine chemical with toxic effects and has eight isomers, but gamma hexachlorocyclohexane (␥-HCH) is the only isomer with an insecticidal effect. ␥-HCH (CAS 58-89-9) is highly persistent in the environment, has high bioaccumulation, and is considered a possible human carcinogen by the International Agency for Research on Cancer (IARC) [5]. The acute effects of this pesticide have been characterized by median lethal concentration (LC50) and for bioaccumulation in several aquatic vertebrates and invertebrates [6]. In fish, ␥-HCH produces hepatic lesions and causes effects in immune cells and the reproductive system [7–9]. In crustaceans, ␥-HCH affects the osmoregulation system, enzymatic activities, and acts on hormonal receptors [10,11]. However, toxic effects of ␥-HCH in bivalve mollusks are poorly documented. Some effects of exposure to ␥-HCH

METHODOLOGY

Organisms, experimental conditions, and exposure to ␥-HCH Adult Pacific oysters, C. gigas, were obtained from a commercial hatchery in Laguna San Ignacio, Baja California Sur, Mexico. Oysters were acclimated in aerated and filtered seawater for 15 d before exposure to chemical compound. The seawater conditions were 22 ⫾ 1⬚C, salinity 37 ⫾ 2‰, and pH 8.0 ⫾ 0.9. The organisms were fed daily with microalgae Isochrisys galbana, 60 ⫻ 104 cells/ml per organism. Aquaria seawater was changed daily. After 15 d of acclimatization, groups of four oysters were placed in 16-L aquaria. Exposure bioassay was performed with complete water change, feeding, and exposure to ␥-HCH every 24 h for 12 d. The application of ␥-HCH was made 20 min after feeding from a stock solution

* To whom correspondence may be addressed ([email protected]). The current address of C. Vazquez-Boucard is Centro de Investigaciones Biologicas del Noroeste (CIBNOR), S.C., Mar Bermejo 195, Colonia Playa Palo de Santa Rita, La Paz, Baja California Sur, Me´xico. 2192

Toxic effects of lindane on Crassostrea gigas

of 25 mg/ml ␥-HCH in acetone. The carrier control group was exposed to 0.4 ml/L of acetone and the seawater control group was maintained only with water changes and daily feeding. A total of 120 oysters were exposed to 0.1, 0.2, 0.3, 0.4, 0.5, 0.7, 1.0, 5.0, and 10.0 mg/L of ␥-HCH in different size groups: carrier and seawater control groups (n ⫽ 20); 0.3 and 0.7 mg/L groups (n ⫽ 16); and 0.1, 0.2, 0.4, 1.0, 5.0, and 10.0 mg/L groups (n ⫽ 8). After 12 d of exposure, hemolymph were taken from abductor muscle by needle suction. Cytotoxic and genotoxic parameters were tested individually.

Median lethal concentration test Mortality in 0.0, 0.3, 0.7, 1.0, 5.0, and 10.0 mg/L ␥-HCH groups was recorded daily and the accumulated mortality was calculated after 12 d of exposure to ␥-HCH. Median lethal concentration at 12 d was calculated by the Probit model [18].

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Table 1. Accumulated mortality and feeding activity of Crassostrea gigas by gamma-hexachlorocyclohexane (␥-HCH) treatment after 12 d of exposure ␥-HCH (mg/L)

0.0 0.1 0.2 0.3 0.4 0.5 0.7 1.0 5.0 10.0

Mortality (%)

Feeding filtration activity (%)

Feeding filtration activity (%)

12 d 0 NDa ND 12.5 ND ND 25 25 62.5 87.5

1st day 100 73.5* 96.2 65.8* 42.0* 52.1* 38.2* ND ND ND

11th day 100 60.4* 41.9* 27.5* 34.0* 45.9* 30.9* ND ND ND

* Significant differences from the control ( p ⫽ 0.05). a ND ⫽ not determined.

Cytotoxic effects on hemocytes Cytotoxicity in 3, 6, or 9 organisms of each treatment was evaluated by trypan blue [19] and dual dye with fluorescein diacetate and ethidium bromide (FDA-EtBr) [20]. Hemolymph samples were diluted in 3% sterile saline and 0.08% trypan blue (1:1, v/v). Cell concentration and viability percentage were quantified by optic microscopy (200⫻) using a Neubauer chamber (PolyLabo, Strasbourg, France). Dual dye was employed in hemolymph samples diluted in 3% sterile saline solution containing 3 ␮g/ml FDA and 8 ␮g/ml EtBr (1:1, v/ v). Mean viability was calculated from 100 hemocytes by fluorescense microscopy (200⫻) in a microscope system (Nikon Optiphot-2/Episcopic–Fluorescence attachment EF-D, Melville, NY, USA) with an excitation filter of 450 to 490 and emission filter of 510 (Nikon VA520).

Genotoxic effects on hemocytes Genotoxicity was evaluated by single cell gel electrophoresis [21]. Microscope slides were covered with three layers of 0.5% agarose containing 1 ⫻ 104 to 2 ⫻ 104 hemocytes. The slides were immersed in a high salt solution (2.5 M NaCl, 100mM ethylenediamine tetraacetic acid (EDTA), disodium salt, dihydrate. Ethylenediamine tetraacetic acid, 1 mM Tris, 10% dimethyl sulfoxide (DMSO), 1% Triton 100X (Sigma, St. Louis, MO, USA, pH 10) and stored at 4⬚C for one hour. Each sample was evaluated in triplicate or quadruplicate. Horizontal electrophoresis was run at 4⬚C with alkaline buffer (300 mM NaOH, 1 mM disodium EDTA, pH ⱖ13). Slides were immersed in buffer for 5 min, then an electric charge was applied (0.7 V/cm, 300 mA) for 10 min. The slides were neutralized (0.4 M Tris, pH 7.5) and dehydrated with absolute ethanol. Hemocytes were analyzed by fluorescence microscopy (400⫻) with an excitation filter of 510 to 560 and emission filter of 580 (Nikon G2A) using a 2 ␮g/ml solution of EtBr. The cells were grouped into five categories according to DNA damage. Each of the 100 cells per slide is assigned to a category, from 0 to 4, depending on the relative intensity of DNA fluorescence in the tail (0 ⫽ undamaged, no DNA in tail; 1 ⫽ 1–25% of DNA in tail; 2 ⫽ 26–50% of DNA in tail; 3 ⫽ 51–75% of DNA in tail; and 4 ⫽ ⬎76% of DNA in tail). The total score for the slide represents the magnitude of DNA breaks, expressed by arbitrary units (AU) [22]. From 1,200 to 3,000 hemocytes of 3, 6, or 9 organisms were evaluated for each treatment. The mean and standard deviation of arbitrary units were calculated per organism and treatment.

Filtration rate on C. gigas feeding Effects of ␥-HCH on C. gigas feeding were evaluated by filtration rate of microalgae by the oysters [23]. On the first and 11th day of exposure to ␥-HCH in the low observable effect concentration (LOEC) assay, 50-ml samples of water were taken at 0, 1, 2, 4, 8, and 24 h after exposure to ␥-HCH. The samples were fixed with 200 ␮l lugol (Iodine [I2] 1% and KI 2% in distilled water, w/v) (Sigma) and preserved at 4⬚C. Cell concentration was evaluated by optic microscopy (200⫻) in triplicate with a Neubauer chamber. The amount of microalgae filtered was calculated by subtracting the concentration at each time from the initial concentration. The rate of filtration for each treatment was computed as the slope of the curve generated by linear regression after 4 h of feeding and reported as the filtration rate compared to the control group.

Statistical analyses One-way analysis of variance (ANOVA) was performed for cytotoxic effect (Tukey test). For AU, data normality was evaluated by Shapiro–Wilk W test and Levene test was utilized for heterocedasticity analysis of cell viability and DNA damage by AU. Genotoxic effect, estimated by AU was tested by one-way ANOVA (Tukey test) (Statistica, StatSoft, Tulsa, OK, USA). Overall cellular frequencies per damage categories were evaluated by chi-square test for contingency tables [24]. Correlation between ␥-HCH concentration and DNA damage was explored by Pearson Product Moment (Statistica). Filtration rate was evaluated by slopes comparison t test [24]. All statistical test were regarded as significant at ␣ ⫽ 0.05. RESULTS

LC50 in C. gigas For the screening assay, the LC50 after 12 d of exposure to ␥-HCH calculated for adult organisms was 2.22 mg/L (y ⫽ 1.4547x ⫹ 4.4989) (r2 ⫽ 0.9648). The lowest concentration of ␥-HCH that presents mortality was 0.3 mg/L after 11 d of exposure (Table I).

Cytotoxicity Cellular viability evaluated by trypan blue (Sigma) was higher than 85% in controls and the groups lower than 1.0 mg/L ␥-HCH. In the groups of 1.0 and 5.0 mg/L ␥-HCH, cellular viability decreased to 62 and 67%, respectively. Cel-

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Fig. 2. Frequency of hemocytes of Crassostrea gigas exposed to gamma-hexachlorocyclohexane (␥-HCH) during 12 d, in five damage levels of DNA strand breaks (for 0.0, 0.3, 0.7 mg/L, n ⫽ 9 oysters; for 0.1, 0.2, 0.4, 0.5 mg/L, n ⫽ 6 oysters; for 1.0, 5.0 mg/L, n ⫽ 3 oysters). * Significant differences from the control (p ⱕ 0.05).

lular viability by FDA-EtBr was higher than 80% in controls and experimental groups, but not in the groups exposed to 1.0 and 5.0 mg/L ␥-HCH (recording values of 65 and 64% respectively), where significantly lower viability was observed (Fig. 1A).

Genotoxicity The single cell gel electrophoresis assay of hemocytes of oysters exposed to ␥-HCH indicated a significant linear correlation between ␥-HCH concentration and DNA damage by AU (r ⫽ 0.73) (Fig. 1B). Genotoxic effects by AU were significant at 0.7, 1.0, and 5.0 mg/L ␥-HCH concentrations compared with the carrier and seawater control groups (Fig. 1C). Overall cell frequencies showed significant differences at 0.7, 1.0, and 5.0 mg/L ␥-HCH compared with the control group, increasing the proportion of damaged cells (levels 1 and 2) and decreasing the proportion of undamaged cells (level 0) (Fig. 2).

Effect on feeding activity of C. gigas The filtration rate of microalgae was altered at 4 h after exposure to ␥-HCH at 0.1 mg/L and higher concentrations, except for the group exposed to 0.2 mg/L. After 11 d of exposure to ␥-HCH, the filtration rate was significantly decreased at all ␥-HCH concentrations (Table 1). During the 11 d of measurement, the filtration rate of the control group did not change significantly. DISCUSSION Fig. 1. Toxic effects of gamma-hexachlorocyclohexane (␥-HCH) on hemocytes of Crassostrea gigas during 12 d of exposure (for 0.0, 0.3, 0.7 mg/L, n ⫽ 9 oysters; for 0.1, 0.2, 0.4, 0.5 mg/L, n ⫽ 6 oysters; for 1.0, 5.0 mg/L, n ⫽ 3 oysters). (A) Cytotoxic effect. (B) Correlation of ␥-HCH treatment and DNA strand breaks, (䡬) AU mean of each organism. (C) DNA strand breaks tested by single cell gel electrophoresis assay. * Significant differences from the control (p ⱕ 0.05). Standard deviation (SD), FDA-EtBr ⫽ fluorescein diacetate and ethidium bromide.

Bivalve mollusks have particular mechanisms that enable them to resist high concentrations of toxic compounds. They can decrease their toxic exposure closing their valves and changing their feeding and metabolic activity, allowing them to survive prolonged periods without food consumption [25– 27]. These responses may be tested by several methods, one of which is the filtration rate. Reports of changes in filtration rates and clearance rates are documented following exposure to toxic compounds [28,29]. Feeding responses of C. gigas to ␥-HCH exposure were evaluated by filtration rates of microalgae. The slopes of the

Toxic effects of lindane on Crassostrea gigas

curves at 0.3 mg/L ␥-HCH treatment and higher displayed significant decreases in filtration rates, down from 65.8 to 38.2% of the control levels. After 11 d of exposure to ␥-HCH, treatment with all concentrations resulted in significant decreases in the range of 60.4 to 30.9%. These values are similar to results reported for cadmium exposure of the freshwater bivalve Lamellidens marginalis with decreases between 62 and 12% [30]. Nevertheless, the reduction in filtration rates does not necessarily mean toxic damage by ␥-HCH; however, it is likely to be a sign of stress that, if maintained over a sufficient period of time, could be detrimental to the health of the organisms. The aim of the present study was to characterize the toxic effects of ␥-HCH on the oyster C. gigas at the individual, cellular, and molecular levels of biological organization. The individual effect was evaluated by mortality during 12 d of exposure and by the generation of a dose–response curve. The LC50 of ␥-HCH in bivalves was reported as 145 mg/L in the fresh water clam Egeria radiata after 96 h of exposure and 1.5 mg/L in the mussel Mytilus galloprovincialis after 7 d of exposure [31,32]. The LC50 for C. gigas larvae is 0.17 mg/L after 9 d of exposure [33]. We report the LC50 for adult C. gigas as 2.22 mg/L after 12 d of exposure; this LC50 is within the same order of magnitude as the LC50 for M. galloprovincialis. The difference in the LC50 among the different species or developmental stages could be due to differences in the period of exposure to ␥-HCH as well as several biological factors such as different lipid proportion in each species, different metabolic rates of each species, and biological factors related to developmental stages in the same species [34,35]. The exposure to ␥-HCH in our study was 12 d; this time period was long enough to observe a constant mortality over time in experimental groups. The toxic effects of ␥-HCH in cells were evaluated in hemocytes. This cell type is a common general health indicator in bivalves [36]. Cytotoxicity was examined using two related cellular mechanisms: permeability of cellular membrane by trypan blue and membrane integrity and metabolic activity by FDA-EtBr. Concentrations lower than 1.0 mg/L exhibited viability percentages higher than 80%, similar to previously reported values for hemocytes of bivalves [37]. Hemocytes of oysters exposed to 1.0 and 5.0 mg/L of ␥-HCH decreased viability percentages to 67 and 62.33%, respectively, revealing a significant cytotoxic effect. Cytotoxicity of ␥-HCH is poorly documented in bivalves, though organochlorine and organophosphorus pesticides have been reported to exhibit cytotoxicity in several cell types in fish [38,39]. Cytotoxic effects of ␥-HCH were observed in neurons and hematic cells of mammals in vitro [40,41]. The major mode of ␥-HCH cytotoxicity is lipid peroxidation and this is potentially related to the magnitude of oxidative stress and the capacity for antioxidant responses by cells [42]. Genotoxicity of ␥-HCH was evaluated by DNA strand breaks in hemocytes using the single cell gel electrophoresis assay. Cells of the hemolymph, gill, and digestive gland of bivalves are sensitive to DNA damage by toxic compounds, although the degree of DNA damage in digestive gland and gill cells are reported to be elevated in unexposed organisms due to mechanical effects of obtaining cellular suspensions [43,44]. The evaluation of genotoxicity of several toxic compounds at polluted sites has relied on the single cell gel electrophoresis assay of bivalve hemocytes as an indicator of genotoxic effects [45,46]. Only at the concentration of 0.7


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mg/L ␥-HCH was DNA damage detectable without observable cytotoxic effects. The ␥-HCH concentrations higher than 0.7 mg/L presented genotoxic and cytotoxic effects in hemocytes of C. gigas. In mammals, ␥-HCH is considered a genotoxic agent in several cell types as evaluated by the single cell gel electrophoresis assay, although these studies were performed in vitro. In these studies, cells were exposed directly to ␥-HCH and the evaluation was focused on cellular toxicology [47]. Exposure assays in vivo, as performed in the present study, permits the direct evaluation of effects, and responses of suborganismal systems as well as in the whole organism [48]. In the present study, a positive correlation was demonstrated between concentration of ␥-HCH and DNA damage of hemocytes. This effect was corroborated by a gradual decrease in the frequency of undamaged cells and a gradual increase in the number of cells in higher damage levels. Damage to DNA by ␥-HCH was evident at 0.1 mg/L and was proportional to the concentration of ␥-HCH, with a significant increase at 0.7 mg/L as compared to the control groups. The 5.0 mg/L ␥-HCH treatment was not included in the correlation analysis because the values of cytotoxicity and DNA damage for this concentration were either lower or very close to the values measured at 1.0 mg/L (Figs. 1 and 2). These results would affect the correlation between the concentration of ␥-HCH and DNA damage (Fig. 1B). The cytotoxicity and DNA damage effects observed at 5.0 mg/L ␥-HCH could be due to the high accumulated mortality recorded for this concentration, resulting in proportionally more remaining organisms that are tolerant to ␥-HCH. Finally, in the present study we show that a toxicant can have toxic effects at different concentrations, depending on the level of effect examined in the evaluation. Molecular and cellular biomarkers were more sensitive to ␥-HCH than organism mortality, and filtration rate was more sensitive than either molecular or cellular endpoints examined in this study. These results demonstrate that it is important to analyze the effects at different biological levels to better characterize the toxicity and the possible mode of action of a chemical.

Acknowledgement—The present work was funded by the Centro de Investigaciones Biologicas del Noroeste (La Paz, Baja California Sur, Mexico) and by grants to C. Vazquez Boucard from de SEP-CONACYT (38827-B). The authors are grateful to M. Carrizosa V., F. Arcos O., M. Valverde, and A. Garcia Gasca for their collaboration. We are also grateful to American Journal Experts and reviewers of ET&C for their valuable comments and suggestions. Gerardo Alfonso Anguiano Vega was a student fellow of CIBNOR and the results presented here are part of his Doctorate of Science thesis under the direction of C. Vazquez Boucard. REFERENCES 1. Valdez SB, Garcia-Duran EI, Wiener MS. 2000. Impact of pesticides use on human health in Mexico: A review. Rev Environ Health 15:399–412. 2. Carvalho FP. 2006. Agriculture, pesticides, food security and food safety. Environ Sci Pollut 9:685–692. 3. Gonzalez-Farias F, Cisneros EX, Fuentes RC, Diaz GG, Botello AV. 2002. Pesticides distribution in sediments of a tropical coastal lagoon adjacent to an irrigation district in northwest Mexico. Environ Technol 23:1247–1256. 4. Botello AV, Diaz G, Rueda L, Villanueva SF. 2000. Persistent organochlorine pesticides (POPs) in coastal lagoons of the subtropical Mexican Pacific. Bull Environ Contam Toxicol 64:390– 397.

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