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Chronic exposure of 2,2′,4,4′-tetrabromodiphenyl ether (PBDE-47) alters locomotion behavior in juvenile zebrafish (Danio rerio)
Aquatic Toxicology 98 (2010) 388–395
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Chronic exposure of 2,2 ,4,4 -tetrabromodiphenyl ether (PBDE-47) alters locomotion behavior in juvenile zebrafish (Danio rerio) Chun-Ting Chou a , Yu-Chen Hsiao a , Fung-Chi Ko a,b , Jing-O. Cheng b,c , Ying-Ming Cheng b , Te-Hao Chen a,b,∗ a b c
Institute of Marine Biodiversity and Evolutionary Biology, National Dong Hwa University, 2 Houwan Rd., Checheng, Pingtung, 944, Taiwan, ROC National Museum of Marine Biology and Aquarium, 2 Houwan Rd., Checheng, Pingtung, 944, Taiwan, ROC Department of Marine Environment and Engineering, National Sun Yat-Sen University, 70 Lienhai Rd., Kaohsiung, 804, Taiwan, ROC
a r t i c l e
i n f o
Article history: Received 11 December 2009 Received in revised form 2 March 2010 Accepted 17 March 2010 Keywords: PBDE-47 Zebrafish Behavior Histology
a b s t r a c t In the present study, we used zebrafish (Danio rerio) as a model to address possible effects of chronic exposure of polybrominated diphenyl ether (PBDE) flame retardants on locomotion behavior, body size, and gonad development in fish. Zebrafish were fed food dosed with PBDE-47 (control, solvent control, low, medium, and high dose groups) from 21 days post hatch (dph) to 90 dph. Fish locomotion parameters, including maximum swimming speed, total distance moved, and percent time active, were assessed using a video-based animal movement analysis system. At the end of the exposure, all fish were euthanized for length and weight measurement, and then subjected to either whole fish histological analysis or tissue PBDE-47 measurement. Survival, body size, and gonad histology were similar between the five groups. However, both total swimming distance and percent time active were negatively correlated with tissue PBDE-47 concentration and were significantly lower in the high dose group. Tissue levels of PBDE-47 in the exposed fish were comparable to that reported in previous field studies. In summary, this study showed that developmental exposure of PBDE-47 at an ecologically relevant level altered locomotion behavior without affecting body size or gonad development of zebrafish. © 2010 Elsevier B.V. All rights reserved.
1. Introduction Polybrominated diphenyl ethers (PBDEs) are widely used as flame retardants in many consumer products, such as textiles, upholstery foam and electronic devices (Chernyak et al., 2005). These compounds are lipophilic, environmentally stable, and structurally similar to polychlorinated biphenyls (PCBs) (de Wit, 2002; Talsness, 2008). Increasing concentrations of PBDEs have been detected in samples of sewage, surface water and sediments (Darnerud et al., 2001; Talsness, 2008). PBDEs are also detected in biotic samples, especially in the aquatic environments. PBDE-47 (2,2 ,4,4 -tetrabromodiphenyl ether), a congener with four bromine atoms, is typically the most abundant congener found in animal tissues (Hites, 2004; Peng et al., 2007). Exponential increases of tissue PBDE concentrations in recent decades have been reported in fish, such as lake trout (Salvelinus namaycush), walleye (Stizostedion vitreum), and rainbow smelt (Osmerus mordax) from the Great Lakes (Zhu and Hites, 2004; Chernyak et al., 2005), and in human tissues
∗ Corresponding author at: Institute of Marine Biodiversity and Evolutionary Biology, National Dong Hwa University, 2 Houwan Rd., Checheng, Pingtung, 944, Taiwan, ROC. Tel.: +886 8 8825001x8055; fax: +886 8 8825066. E-mail address: [email protected] (T.-H. Chen). 0166-445X/$ – see front matter © 2010 Elsevier B.V. All rights reserved. doi:10.1016/j.aquatox.2010.03.012
(Alaee, 2003; Hites, 2004). Indeed, such rapid increase of PBDEs in biotic samples has raised concern about their potential toxicities in wildlife and human (Siddiqi et al., 2003). PBDEs have been shown to disrupt the thyroid and sex steroid endocrine systems. PBDEs have the potential to interrupt the thyroid system by direct binding to thyroid hormone (TH) receptors, competition with thyroxine (T4 ) for transport protein transthyretin, and/or increased metabolism of T4 (Schriks et al., 2007; Talsness, 2008). Compared to other organohalogen compounds, PBDEs are likely to be more potent thyroid disruptors (Darnerud et al., 2001). Studies have shown that PBDEs can decrease plasma T4 levels in rodents and fish (Hallgren et al., 2001; Lema et al., 2008). PBDEs may also cause negative effects on animal reproduction systems. Previous in vitro study showed that PBDEs have estrogenic potencies by acting as estrogen receptor (ER) agonists (Meerts et al., 2001), suggesting possible effects on gonad development or sex differentiation. In vivo studies involving rodents have shown anti-androgenic effects, such as reduced sperm production and suppressed growth of androgen-dependent tissues in males (Talsness, 2008). Inhibited breeding and sperm production were observed in fathead minnows (Pimephales promelas) exposed to PBDE-47 via diet (Muirhead et al., 2006). Neurobehavioral effect is another great concern for PBDE toxicity. Thyroid hormones play a crucial role during brain development,
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and many studies involving rodents have shown subtle changes in motor and cognitive function following pre- or postnatal exposure of PBDEs (Branchi et al., 2003; Costa and Giordano, 2007). Indeed, changes in behavior have been shown to be a sensitive endpoint to PBDE exposure in many studies involving rodents (Talsness, 2008). In contrast, very few studies have employed behavioral approaches to study the neurobehavioral effects of PBDEs on fish (TimmeLaragy et al., 2006). In the area of fish ecotoxicology, common endpoints usually include hatchability, survival, morphological deformities, and sometimes histopathology. Behavioral endpoints, however, are not commonly included in the toxicity bioassays despite their ecological relevance. Nevertheless, behavioral endpoints of fish are valuable tools to address effects of environmental stress because they integrate endogenous and exogenous factors that can link biochemical and physiological processes, and can infer the effects of environmental contaminants at higher levels of biological organization (e.g., individual, population, and community levels) (Kane et al., 2004). Quantifiable behavioral changes in fish associated with toxicant exposure thus provide novel information that cannot be gained from traditional toxicological methods (Little and Finger, 1990; Bridges, 1997; Kane et al., 2004). Unfortunately, behavioral endpoints have been underutilized in ecotoxicology for a long time, partly due to the lack of interdisciplinary integration between ethology and ecotoxicology (Newman, 2001; Clotfelter et al., 2004). With the recent development of computer technology, video-based movement tracking systems have been greatly improved and used extensively in quantification of animal locomotion behavior in areas such as pharmacology and psychology (e.g., Shinba et al., 1996; Gentry et al., 2004). When combined with other lethal and sublethal endpoints, behavioral studies could be very effective to assess the impact of contaminants on fish (Beauvais et al., 2001; Newman, 2001; Kane et al., 2005; Jakka et al., 2007). Zebrafish (Danio rerio) are now considered as a good fish model for investigating endocrine disruption and neurobehavioral effects of environmental contaminants (Creton, 2009; Segner, 2009). In this study, zebrafish were exposed to environmentally relevant levels of PBDE-47 via diet from the larval stage to adult stage for 70 days. In zebrafish, this window of time is very sensitive to chemical exposure because morphological changes, hormonal alterations, and gonad differentiation all take place during this period (Brown, 1997; Maack and Segner, 2003). The objectives of this study were to assess the effect of environmental levels of dietary PBDE-47 exposure on zebrafish growth and gonad histopathology. In addition, a video-based animal tracking system was employed to evaluate possible neurobehavioral effect of PBDE-47 on zebrafish. 2. Materials and methods 2.1. Animal model Adult AB strain zebrafish for broodstock were obtained from G. fish Animal Model Co. (Taipei, Taiwan). Male and female adult fish were maintained separately in dechlorinated tap water at 28.5 ◦ C under a 14L:10D photoperiod, with a diet of commercial fish flakes and live Artemia nauplii. The fish had been maintained in our laboratory for one month prior to breeding. Zebrafish embryos were obtained from spawning adults in groups containing six females and three males overnight in spawning tanks. Eggs were collected on the next day and allowed to hatch. Hatched fry were maintained at 28.5 ◦ C in a 10 L tank with gentle aeration. From four day post hatch (dph) to 20 dph (right before the exposure experiment began), larval fish were fed twice a day with commercial fish powder, and the tank was cleaned daily by siphoning the bottom. The exposure was not started until 20 dph because prior to that time the larval survivorship was not stable even without any treatment and may mask any possible effect of PBDE-47 exposure.
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The experimental protocol involving handling and treatment of zebrafish was in accordance with the Taiwan national guidelines for animal welfare and was approved by the IACUC (Institutional Animal Care and Usage Committee) of National Museum of Marine Biology and Aquarium, Taiwan. 2.2. Experimental diets PBDE-47 standard (purity 98.6%, GC/MS certified) was obtained from ChemService (West Chester, PA, USA) and used to prepare a stock solution (1 mg/L) in methanol (J.T. Baker, Phillipsburg, NJ, USA). Fish powder (Chuankuan, Kaohsiung, Taiwan) or roughly ground TetraMin® MiniGranules fish pellets (Tetra, Melle, Germany) were used to prepare the experimental diets. During the 70 days of exposure period, the fish were gradually acclimated from fish powder to TetraMin (the feeding regime will be detailed later). For the PBDE treatment diets (nominal concentration = 10, 100, and 100 ng/g food wet weight for the low, medium, and high dose groups, respectively), an appropriate volume of the stock solution was added to a mixture of methanol and fish food (either fish powder or ground TetraMin; 1 mL methanol: 1 g fish food). The solvent control diet was prepared with similar procedures without adding any stock solution in the mixture. The mixtures were thoroughly blended and the methanol was completely evaporated in a hood at room temperature. For the control diet, neither methanol nor the stock solution was added. The experimental diets were stored at −20 ◦ C throughout the experiment. 2.3. Experimental design At 21 dph, 30 larval zebrafish were randomly attributed to each of 15 glass aquariums (30 cm × 45 cm × 30 cm with 40 L dechlorinated tap water). These aquariums were randomly divided to five experimental groups (i.e., control, solvent control, low, medium, and high dose groups). The aquariums were equipped with aerated power water filters and were maintained at 28 ± 1 ◦ C under a 14L:10D photoperiod throughout the experiment. The fish in each group were fed their respective diet twice a day from 21 to 90 dph. During the exposure period, the amount and type of food were adjusted with fish development based on our previous experience and fish size. Before reaching 40 dph, the zebrafish were fed the fish powder diet (0.04 g/tank/day). From 40 to 70 dph, the ration was adjusted to 0.06 g/tank/day of a mix of fish powder and TetraMin (fish powder:TetraMin = 2:1). After 70 dph, the fish were still fed 0.06 g/tank/day, but the fish powder: TetraMin ratio was changed to 1:2. Because of the very small particle sizes, the food immediately dispersed in the water after dispensing and the food consumption by individual fish was therefore not controlled during each feeding. Nevertheless, we assumed that the fish were fed to satiation because there was always a little food leftover on the tank bottom after the feeding. For each aquarium, 5 L of water was changed and the tank bottom was siphoned three times a week throughout the exposure period. At about one month after exposure began (54 dph), six fish from each tank were sampled for behavioral testing (detailed below). After the behavioral trials, the fish were returned to their respective tanks. At 90 dph when the fish were supposed to reach sexual maturation (Maack and Segner, 2003), all fish were euthanized in MS222 solution (ethyl 3-aminobenzoate methanesulfonate salt, Sigma, St. Louis, MO, USA). The fish in each tank were counted and their standard length (from the tip of the snout to the posterior end of the last vertebra) and wet weight were measured. The condition factor (K-factor) was calculated for each fish according to the formula: K-factor = (weight (g)/length (cm)3 ) × 100 (Brion et al., 2004). From each tank, three to five fish were wrapped in aluminum foil and
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stored at −20 ◦ C for tissue PBDE-47 measurement, and 5–14 fish were fixed in Dietrich’s fixative for histological analysis.
stage was counted using ImageJ software. Proportion of oocytes at each stage was calculated by dividing the oocyte count at respective stages by total oocyte count.
2.4. Behavioral testing The behavioral testing took place between 8 am and 12 noon at 54 dph. The fish were fast for 12 h prior the behavioral testing. From each tank, six fish were sampled and individually placed in each well (with 10 mL of dechlorinated water) of a 6-well microplate. The microplate was placed on the plastic holding stage, specifically designed to fit microplates (from 6- to 96-well plates), with infrared illumination from the bottom, and a CCD video camera was centrally positioned 110 cm above the holding stage. The holding stage and the video camera were surrounded by a black curtain during each trial to avoid any possible disturbance from the researcher. After 2 min of acclimation, the movement of each fish in the microplate was recorded with the video camera for 6 min. To stimulate predator-avoiding response of the fish, the holding stage was slightly knocked with a plastic hammer every 30 s during the third and forth minute. The video output from the camera was fed to a computer running an animal movement tracking software (Ethovison® XT, Noldus Information Technology, Wageningen, Netherlands). Fish locomotion parameters, including maximum velocity, total distance moved, and percent time active, were quantified from video with a tracking rate of 25 frames/s. Because fish swimming performance may be affected by body size, we measured standard length of the fish from the video image clips using ImageJ software (National Institute of Health, USA). Statistical analysis showed no significant difference of fish length between groups, indicating that body size was not a confounding factor in fish locomotion analysis. The fish behaviors were analyzed at 54 dph instead of 90 dph because otherwise the fish sizes will be too big for the 6-well microplates. 2.5. Histological preparation and gametogenesis analysis The number of fish processed for histological analysis was 27, 25, 20, 29, and 29 in the control, solvent control, low, medium, and high dose group, respectively. After routine tissue processing and paraffin embedding, whole fish were sectioned coronally into 5 m sections. For gonad visualization, sectioning was performed at a level simultaneously showing the gonads, the liver and the swim bladder (van der Ven et al., 2003). Sections were mounted on slides and stained with hematoxylin/eosin (H & E). Histology of gonad tissues was examined and fish gender was determined under an Olympus CX-31 light microscope equipped with an Olympus DE-71 digital camera (Olympus, Tokyo, Japan). The stage of testicular development was assessed in slides of fish that were histologically discerned to be male. At 40× magnification on the light microscope, six digital images of testicular area were randomly taken from each specimen, and four images were further randomly selected for image analysis. The size of testicular area, which containing spermatogonia, primary or secondary spermatocytes, and spermatids (mature sperm) (Weber et al., 2003), in each image was measured using ImageJ software. Relative area of spermatids was estimated by dividing spermatid area by total testicular area. The calculated values of four images from each specimen were averaged for statistical analysis. The stage of ovarian development was assessed in slides of fish that were histologically discerned to be female. For each specimen, a field of view covering the whole ovarian area was digitally imaged using an Olympus DE-71 digital camera connected to an Olympus ZX-61 dissecting microscope. Only oocytes at perinucleolar, cortical alveolus, vitellogenic, and mature stages were analyzed, thus excluding the smallest oocytes (oogonia) (Maack and Segner, 2003; Weber et al., 2003; Koc et al., 2008). The number of oocytes at each
2.6. Quantification of PBDE-47 in food and fish tissue PBDEs in food (2 g, n = 4) and fish tissues (exposed for 70 days; pooled samples of 3–5 individuals from each tank, n = 3) were extracted with an accelerated solvent extraction system (ASE 300 system, Dionex, Sunnyvale, CA, U.S.) using dichloromethane (DCM) as the solvent. One-tenth of the extract was designated for gravimetric lipid determination after solvent evaporation (Lauenstein et al., 2002). The clean-up steps of the extracts involved gelpermeation cleanup followed by Florisil columns (8 g, contain 2.5% water). The eluate was evaporated and dissolved in about 5 mL of n-hexane, then to a final volume of 1 mL with a gentle stream of nitrogen. PCB 30 and PCB 204 were added to each sample as internal standards. One blank vial spiked with 20 ng PBDE-47 standard was prepared as recovery standard. The sample analyses were preformed on a Varian 3800 gas chromatograph equipped with an autosampler, splitless injector, ion trap mass detector Varian Saturn 4000 (Varian Chromatography Systems, Walnut Creek, CA, USA). Recovery of PBDE-47 standard was 80.88 ± 6.41% (mean ± standard error). 2.7. Statistics Statistical analyses were performed with SYSTAT® (V. 12, Systat Software, Richmond, CA, USA). Shapiro–Wilk’s test was used to test the assumption of normality. Except the sex ratio data, which was analyzed with a chi-square test, differences between treatments were compared by analyses of variance (ANOVA), using “tank” as the experimental unit, i.e., mean values from each tank were used for the statistical analyses. Proportion data were adjusted with an arcsine square-root transformation. When the overall ANOVA was significant, Tukey’s test was performed for post-hoc pairwise comparisons. Correlations between tissue PBDE-47 concentrations (log10 transformed) and distance moved and proportion time active were further tested by simple linear regression analyses. Statistical significance was accepted for p < 0.05. Data were presented as mean ± standard error (SE). 3. Results 3.1. PBDE-47 concentrations in the diets and fish tissue Table 1 lists the levels of PBDE-47 measured in the experimental diets and fish whole-body tissue homogenates. For the diets, measured concentrations were lower than nominal concentrations, probably due to adsorption on the inner wall of the beakers during diet preparation. Tissue concentrations of PBDE-47 varied among groups as expected by the increased PBDE-47 levels in the dosed diets. 3.2. Survival and body size At 90 dph, survival in each group ranged from 64 to 83% and was not significantly different between groups (p = 0.21) (Fig. 1A). There was no statistical difference between groups for the standard length (p = 0.23) and wet weight (p = 0.12) (Fig. 1B–C). For those fish sampled for histological analysis (gender was thus determined), condition factors were similar between groups in both males (ranged from 1.66 to 1.81, p = 0.35) and females (ranged from 1.82 to 2.17, p = 0.56).
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Table 1 Measured PBDE-47 concentration (mean ± standard error) in experimental diets and fish tissue. Treatmenta
Dietb (ng/g wet weight)
Control Solvent control Low dose (10 ng/g) Medium dose (100 ng/g) High dose (1000 ng/g)
0.68 ± 0.28 1.34 ± 0.61 12.46 ± 2.03 81.28 ± 9.11 643.60 ± 57.68
a b c
Fish tissuec (ng/g wet weight)
(ng/g lipid weight)
<0.003 <0.003 2.70 ± 0.77 31.92 ± 4.34 409.68 ± 82.01
<0.02 <0.02 11.20 ± 2.48 189.04 ± 2.33 1924.16 ± 289.36
Values in parentheses are nominal concentrations of PBDE-47 in the diets. Sample size n = 4. Sample size n = 3.
3.3. Behavioral testing Fig. 2 shows the maximum velocity, total distance moved and percent time active of fish during the behavioral testing. Exposure to PBDE-47 did not cause a significant effect on maximum velocity (p = 0.42) (Fig. 2A). However, total distance moved and percent time active were both significantly affected (p = 0.04 and 0.01, respectively) (Fig. 2B–C). Tukey’s post-hoc pairwise comparison showed that total distance moved in high dose group (360.33 ± 20.84 cm) was significantly lower than that in control group (582.84 ± 57.27 cm) (p = 0.03) though not significantly different from the solvent control group (506.17 ± 34.76 cm) (p = 0.19). Data of percent time active showed a similar pattern; it was significantly lower in the high dose group (47.81 ± 3.40%) than that in the control group (76.17 ± 4.05%) (p = 0.01) but did not differ from the solvent control group (64.18 ± 5.52%) (p = 0.2). Simple linear regression analyses showed a significant dose-dependent decline in both total distance moved (r = −0.67, p = 0.007) and percent time active (r = −0.68, p = 0.005) with increasing tissue PBDE-47 concentrations (Fig. 3). 3.4. Gametogenesis analysis Two underdeveloped (standard length <10 mm) individuals in the high dose group had undifferentiated gonads, and we were thus not able to determine their gender histologically. Therefore, these two individuals were excluded from the statistical analysis. Sex ratio was not significantly different between groups (p = 0.9) (Fig. 4). No ovotestis was observed from the histological analysis. Relative area of spermatids in testicular tissue of male fish ranged from 37.9 to 46.2% and did not differ between groups (p = 0.75). Proportion of oocytes at each stage in females was not affected by PBDE-47 exposure, either (p > 0.5) (Fig. 5). 4. Discussion PBDE-47 tissue concentrations measured in the present study were comparable to that in wild fish reported in the literature. Tissue concentrations of PBDE-47 in the low and medium dose groups were 11 and 189 ng/g lipid, respectively, which were similar to that in fish collected from six rivers and three estuaries in Taiwan (PBDE47 = 17–176 ng/g lipid; total PBDEs = 25–281 ng/g lipid) (Peng et al., 2007). Tissue concentrations of PBDE-47 in the high dose group was 1924 ng/g lipid, which was close to that measured in fish from San Francisco Bay, USA (PBDE-47 = 122–1286 ng/g lipid; total PBDEs = 306–2235 ng/g lipid) (Holden et al., 2003). The tissue PBDE47 concentrations measured in our study were also much lower than that reported in previous bioassay studies (at the g/g level) using small laboratory fish such as fathead minnow and zebrafish (Lema et al., 2007, 2008; Kuiper et al., 2008). To our knowledge, the tissue PBDE-47 levels reported in this current study probably are the lowest compared with other fish bioassay studies in the literature. It should be noted that the tissue PBDE-47 concentrations
were analyzed at 90 dph, while the behavioral endpoints were measured at 54 dph. In a previous study with fathead minnow orally exposed to PBDE-47, tissue PBDE-47 increased rapidly and reached a plateau at around 10 days post exposure (Muirhead et al., 2006). We assumed that tissue residues at 90 dph were similar to that at 54 dph and correlated the chemical analysis data with the behavioral measurements (distance moved and time active) in statistical analyses. Very limited information is available so far concerning the effect of PBDEs on animal growth. PBDEs have been shown to suppress growth in rodent, amphibian, and fish (IPCS, 1994; Balch et al., 2006; Lema et al., 2007). However, increased growth has also been observed in PBDE-exposed bird and rat (Fernie et al., 2006; Suvorov et al., 2009), partially due to increased appetite (Fernie et al., 2006). Regarding bony fish, thyroid hormones are critical to the growth and development before fish reaching juvenile stage (Power et al., 2001). Recent studies showed that PBDE-47 exposure caused depressed plasma T4 levels in adult fathead minnows (Lema et al., 2008) and reduced growth in embryonic zebrafish (Lema et al., 2007). The thyroid gland of zebrafish becomes functional at the onset of feeding and then increases in activity throughout the larval and transitional periods (∼21 dph) to the juvenile stage (∼35 dph); exposure to goitrogens during this period of time can inhibit growth and development of larval zebrafish (Brown, 1997). In the present study, zebrafish body size parameters (standard length, weight, and condition factor) measured at 90 dph were not affected by PBDE-47 exposure. It is possible that the exposure levels of PBDE-47 in this study were not high enough to elicit an observable effect on fish size. It is also possible that the thyroid system, as well as growth, were inhibited during larval-juvenile stages, but with time the fish continued to grow and thus compensated for their reduced growth. Though thyroid hormones are critical to larval development and metamorphosis, fish growth after juvenile stage is more dependent on other factors, such as growth hormone, IFG-1, and insulin (Mommsen, 2001). In one of our recent studies, zebrafish orally exposed to PBDE-47 (74.9 ng/g food) from 20 to 60 dph were significantly smaller at 38 dph but not different in size from the control at 60 dph (Chen et al., 2010). Similar growth compensation has also been observed in zebrafish exposed to goitrogens (Brown, 1997). Histopathology is a valuable tool for assessing endocrinedisrupting effects on fish. Histopathological changes, especially those in the gonads, are indicative of dysfunctional reproduction and thus the negative effect on the population level can be deduced (van der Ven et al., 2003). Developmental exposure (from 2 to 60 dph in water) to estrogenic compounds, such as 4-nonylphenol (NP) and 17␣-ethinylestrodiol (EE), in zebrafish has been shown to cause suppression of gametogenesis in both male and female fish (Weber et al., 2003). Sex ratio is another crucial endpoint to determine feminization and masculinization in fish exposed to EDCs (Orn et al., 2006b). For instance, developmental exposure of zebrafish to EE and effluent water from a Swedish pulp mill from 1 to 60 dph caused 100% female and male-biased sex ratios, respectively (Orn et al., 2006a,b). Estrogenic or anti-androgenic activity
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Fig. 1. Standard length (A), wet weight (B), and survival rate (C) at 90 dph of zebrafish exposed to different levels of PBDE-47 via diet (C: control; S: solvent control; L: low dose; M: medium dose; H: high dose). Data are presented as mean ± standard error. There was no statistical difference between groups regarding standard length (p = 0.23), wet weight (p = 0.12), and survival rate (p = 0.21).
Fig. 2. Maximum swimming speed (A), total distance moved (B), and percent time active (C) measured at 54 dph of zebrafish exposed to different levels of PBDE-47 via diet (C: control; S: solvent control; L: low dose; M: medium dose; H: high dose). Data are presented as mean ± standard error. Different letters indicate significant difference between groups (p < 0.05).
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Fig. 4. Sex ratio of zebrafish exposed to different levels of PBDE-47 via diet (C: control; S: solvent control; L: low dose; M: medium dose; H: high dose). Data are presented as mean ± standard error. Sex ratio was similar between groups (p = 0.9).
exposure may not pose significant impact on gonad development in fish. Another great concern for the potential health effects of PBDEs comes from their developmental neurobehavioral toxicity (Birnbaum and Staskal, 2004). Developmental neurobehavioral toxicity of PBDEs has been extensively studied in rodents. Available studies involving rodents indicate pre- or postnatal exposure to various PBDEs (including PBDE-47) cause alterations in motor and cognitive functions, such as spontaneous behavior, learning, and memory (Branchi et al., 2003; Costa and Giordano, 2007).
Fig. 3. Correlation between tissue PBDE-47 concentrations (log10 transformed) and total distance moved (A) or proportion time active (arcsine square-root transformed) (B). The solid lines are the least-squares regression lines and the regression equations are given in the plot.
of PBDEs have been observed in both in vitro and in vivo studies (Talsness, 2008). Previous studies using histological examination revealed significantly reduced mature sperm in male fathead minnows orally exposed to PBDE-47 (Muirhead et al., 2006; Lema et al., 2008). In another study, adult zebrafish were exposed to analytically cleaned commercial penta-BDE mixture (DE-71, 5–500 g/L) for 30 days, but no major histopathological changes were observed in the gonads (Kuiper et al., 2008). In this present study, neither sex ratio nor gametogenesis was affected by PBDE-47 exposure. Zebrafish is a juvenile hermaphrodite, with all individuals having immature ovary – like gonads during early life and the bisexual gonad differentiation taking place only after about 7 weeks (49 days) post-fertilization (Maack and Segner, 2003). Therefore, our dietary exposure duration (from 20 to 90 dph) should be able to cover the period of gonad differentiation and maturation. The lack of observable effect of PBDE-47 on zebrafish gonad histology was probably due to our relatively low (but environmentally relevant) levels of exposure. Our results suggest that environmental PBDE-47
Fig. 5. Percentage of oocytes at each stage (Po: perinucleolar oocyte stage; Ca: cortical alveolus stage; Vo: vitellogenic oocyte stage; M: mature oocyte stage) in the ovaries of female zebrafish exposed to different levels of PBDE-47 via diet (C: control; S: solvent control; L: low dose; M: medium dose; H: high dose). Data are presented as mean ± standard error. Proportion of oocytes at each stage was similar between groups (p > 0.05).
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The mechanism underlying the neurobehavioral effects of PBDEs is as yet unclear. It has been proposed that disruption of thyroid hormones, which play a key role in brain development, and/or interactions with neurotransmitter systems may explain the neurotoxicity of PBDEs (Costa and Giordano, 2007; Gee and Moser, 2008). In contrast to the abundant literature regarding rodents, few studies have addressed neurobehavioral effects of PBDEs on fish. In the estuarine minnow (Fundulus heteroclitus) exposed to the commercial PBDE mixture DE-71 via water (0.001 and 0.01 g/L) during 0–7 day post-fertilization (dpf), hypoactivity and impaired fright response were observed at 4 dpf (Timme-Laragy et al., 2006). The authors also observed poorer predation performance in the hypoactive fish tested at 11 dpf (Timme-Laragy et al., 2006). In another study, embryonic exposure of zebrafish to waterborne PBDE47 (5000 g/L) resulted in reduced movement of cerebrospinal fluid in the brain ventricles and neural tubes, and the authors suggested that this may lead to the development of neural deficiencies later in life (Lema et al., 2007). In our current study, fright response (measured as maximum burst speed) was not affected by PBDE-47 exposure, but hypoactivity (i.e., decreased total distance moved and percent time active) did occur in the PBDE-47 treated fish. From an ecological point of view, capability of swimming is crucial for fish survival within the aquatic environment (Wolter and Arlinghaus, 2003). Aberrant locomotion behavior and activity level could have significant impact on predator–prey interactions, reproductive behavior, migration, and dispersal and thus decrease fitness of fish (Little and Finger, 1990; Vieira et al., 2009). For example, hypoactivity may result in less chance of encountering prey by reducing search areas and decrease the amount of energy for growth (Little and Finger, 1990). Previous literature showed that toxicants can disrupt startle responses, swimming activity, and ability to escape from a predator of prey fish (Scott and Sloman, 2004). Indeed, swimming behavior has long been proposed to be incorporated in test protocols to expend the sensitivity of standard toxicity test (Little and Finger, 1990). In this study, the behavioral endpoints are more sensitive than histological analysis to detect sublethal effects of PBDE-47. With modern electronic equipments and software, swimming behavior measurements can be rigorously monitored in the laboratory (Little and Finger, 1990; Kane et al., 2004). Our laboratory is currently testing the potential of this behavioral analysis system for high throughout measurement of behavioral responses of larval fish exposed to other contaminants. Because of their sensitivity and biological relevance, we suggest that quantitative behavioral measurements such as swimming activity should be more broadly included in standard aquatic toxicity assessment (Little and Finger, 1990; Kane et al., 2004; Scott and Sloman, 2004).
5. Conclusion Environmental dietary exposure of PBDE-47 to zebrafish during larval-juvenile stage affected fish locomotion behavior but not fish size and gonad development. Tissue PBDE-47 concentrations measured in this study were comparable to that reported in previous field studies, suggesting that wild fish with similar exposure level may be at risk of neurobehavioral toxicity. This study also demonstrated the utility of fish behavioral parameters in ecotoxicological assays. The video-based movement analysis system was shown to be useful for rapid, quantifiable, and high throughput assessment of neurobehavioral toxicity in zebrafish (Kane et al., 2004; Creton, 2009). We suggest that behavioral indicators such as swimming capability should be more widely included in fish bioassays for environmental contaminants (Scott and Sloman, 2004).
Acknowledgements This study was supported by the Thematic Research Grant from NMMBA (Grant No. 981003022) and a grant from the National Science Council of Taiwan, Republic of China (Grant NSC97-2313B291-001). We appreciate Dr. Chung-Der Hsiao for his advice for zebrafish breeding and histology and Yi-Ru Chen for the chemical analysis. References Alaee, M., 2003. Recommendations for monitoring of polybrominated diphenyl ethers in the Canadian environment. Environ. Monit. Assess. 88, 327–341. Balch, G.C., Velez-Espino, L.A., Sweet, C., Alaee, M., Metcalfe, C.D., 2006. Inhibition of metamorphosis in tadpoles of Xenopus laevis exposed to polybrominated diphenyl ethers (PBDEs). Chemosphere 64, 328–338. Beauvais, S.L., Jones, S.B., Parris, J.T., Brewer, S.K., Little, E.E., 2001. Cholinergic and behavioral neurotoxicity of carbaryl and cadmium to larval rainbow trout (Oncorhynchus mykiss). Ecotox. Environ. Safe. 49, 84–90. Birnbaum, L.S., Staskal, D.F., 2004. 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Contents lists available at ScienceDirect
Aquatic Toxicology journal homepage: www.elsevier.com/locate/aquatox
Chronic exposure of 2,2 ,4,4 -tetrabromodiphenyl ether (PBDE-47) alters locomotion behavior in juvenile zebrafish (Danio rerio) Chun-Ting Chou a , Yu-Chen Hsiao a , Fung-Chi Ko a,b , Jing-O. Cheng b,c , Ying-Ming Cheng b , Te-Hao Chen a,b,∗ a b c
Institute of Marine Biodiversity and Evolutionary Biology, National Dong Hwa University, 2 Houwan Rd., Checheng, Pingtung, 944, Taiwan, ROC National Museum of Marine Biology and Aquarium, 2 Houwan Rd., Checheng, Pingtung, 944, Taiwan, ROC Department of Marine Environment and Engineering, National Sun Yat-Sen University, 70 Lienhai Rd., Kaohsiung, 804, Taiwan, ROC
a r t i c l e
i n f o
Article history: Received 11 December 2009 Received in revised form 2 March 2010 Accepted 17 March 2010 Keywords: PBDE-47 Zebrafish Behavior Histology
a b s t r a c t In the present study, we used zebrafish (Danio rerio) as a model to address possible effects of chronic exposure of polybrominated diphenyl ether (PBDE) flame retardants on locomotion behavior, body size, and gonad development in fish. Zebrafish were fed food dosed with PBDE-47 (control, solvent control, low, medium, and high dose groups) from 21 days post hatch (dph) to 90 dph. Fish locomotion parameters, including maximum swimming speed, total distance moved, and percent time active, were assessed using a video-based animal movement analysis system. At the end of the exposure, all fish were euthanized for length and weight measurement, and then subjected to either whole fish histological analysis or tissue PBDE-47 measurement. Survival, body size, and gonad histology were similar between the five groups. However, both total swimming distance and percent time active were negatively correlated with tissue PBDE-47 concentration and were significantly lower in the high dose group. Tissue levels of PBDE-47 in the exposed fish were comparable to that reported in previous field studies. In summary, this study showed that developmental exposure of PBDE-47 at an ecologically relevant level altered locomotion behavior without affecting body size or gonad development of zebrafish. © 2010 Elsevier B.V. All rights reserved.
1. Introduction Polybrominated diphenyl ethers (PBDEs) are widely used as flame retardants in many consumer products, such as textiles, upholstery foam and electronic devices (Chernyak et al., 2005). These compounds are lipophilic, environmentally stable, and structurally similar to polychlorinated biphenyls (PCBs) (de Wit, 2002; Talsness, 2008). Increasing concentrations of PBDEs have been detected in samples of sewage, surface water and sediments (Darnerud et al., 2001; Talsness, 2008). PBDEs are also detected in biotic samples, especially in the aquatic environments. PBDE-47 (2,2 ,4,4 -tetrabromodiphenyl ether), a congener with four bromine atoms, is typically the most abundant congener found in animal tissues (Hites, 2004; Peng et al., 2007). Exponential increases of tissue PBDE concentrations in recent decades have been reported in fish, such as lake trout (Salvelinus namaycush), walleye (Stizostedion vitreum), and rainbow smelt (Osmerus mordax) from the Great Lakes (Zhu and Hites, 2004; Chernyak et al., 2005), and in human tissues
∗ Corresponding author at: Institute of Marine Biodiversity and Evolutionary Biology, National Dong Hwa University, 2 Houwan Rd., Checheng, Pingtung, 944, Taiwan, ROC. Tel.: +886 8 8825001x8055; fax: +886 8 8825066. E-mail address: [email protected] (T.-H. Chen). 0166-445X/$ – see front matter © 2010 Elsevier B.V. All rights reserved. doi:10.1016/j.aquatox.2010.03.012
(Alaee, 2003; Hites, 2004). Indeed, such rapid increase of PBDEs in biotic samples has raised concern about their potential toxicities in wildlife and human (Siddiqi et al., 2003). PBDEs have been shown to disrupt the thyroid and sex steroid endocrine systems. PBDEs have the potential to interrupt the thyroid system by direct binding to thyroid hormone (TH) receptors, competition with thyroxine (T4 ) for transport protein transthyretin, and/or increased metabolism of T4 (Schriks et al., 2007; Talsness, 2008). Compared to other organohalogen compounds, PBDEs are likely to be more potent thyroid disruptors (Darnerud et al., 2001). Studies have shown that PBDEs can decrease plasma T4 levels in rodents and fish (Hallgren et al., 2001; Lema et al., 2008). PBDEs may also cause negative effects on animal reproduction systems. Previous in vitro study showed that PBDEs have estrogenic potencies by acting as estrogen receptor (ER) agonists (Meerts et al., 2001), suggesting possible effects on gonad development or sex differentiation. In vivo studies involving rodents have shown anti-androgenic effects, such as reduced sperm production and suppressed growth of androgen-dependent tissues in males (Talsness, 2008). Inhibited breeding and sperm production were observed in fathead minnows (Pimephales promelas) exposed to PBDE-47 via diet (Muirhead et al., 2006). Neurobehavioral effect is another great concern for PBDE toxicity. Thyroid hormones play a crucial role during brain development,
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and many studies involving rodents have shown subtle changes in motor and cognitive function following pre- or postnatal exposure of PBDEs (Branchi et al., 2003; Costa and Giordano, 2007). Indeed, changes in behavior have been shown to be a sensitive endpoint to PBDE exposure in many studies involving rodents (Talsness, 2008). In contrast, very few studies have employed behavioral approaches to study the neurobehavioral effects of PBDEs on fish (TimmeLaragy et al., 2006). In the area of fish ecotoxicology, common endpoints usually include hatchability, survival, morphological deformities, and sometimes histopathology. Behavioral endpoints, however, are not commonly included in the toxicity bioassays despite their ecological relevance. Nevertheless, behavioral endpoints of fish are valuable tools to address effects of environmental stress because they integrate endogenous and exogenous factors that can link biochemical and physiological processes, and can infer the effects of environmental contaminants at higher levels of biological organization (e.g., individual, population, and community levels) (Kane et al., 2004). Quantifiable behavioral changes in fish associated with toxicant exposure thus provide novel information that cannot be gained from traditional toxicological methods (Little and Finger, 1990; Bridges, 1997; Kane et al., 2004). Unfortunately, behavioral endpoints have been underutilized in ecotoxicology for a long time, partly due to the lack of interdisciplinary integration between ethology and ecotoxicology (Newman, 2001; Clotfelter et al., 2004). With the recent development of computer technology, video-based movement tracking systems have been greatly improved and used extensively in quantification of animal locomotion behavior in areas such as pharmacology and psychology (e.g., Shinba et al., 1996; Gentry et al., 2004). When combined with other lethal and sublethal endpoints, behavioral studies could be very effective to assess the impact of contaminants on fish (Beauvais et al., 2001; Newman, 2001; Kane et al., 2005; Jakka et al., 2007). Zebrafish (Danio rerio) are now considered as a good fish model for investigating endocrine disruption and neurobehavioral effects of environmental contaminants (Creton, 2009; Segner, 2009). In this study, zebrafish were exposed to environmentally relevant levels of PBDE-47 via diet from the larval stage to adult stage for 70 days. In zebrafish, this window of time is very sensitive to chemical exposure because morphological changes, hormonal alterations, and gonad differentiation all take place during this period (Brown, 1997; Maack and Segner, 2003). The objectives of this study were to assess the effect of environmental levels of dietary PBDE-47 exposure on zebrafish growth and gonad histopathology. In addition, a video-based animal tracking system was employed to evaluate possible neurobehavioral effect of PBDE-47 on zebrafish. 2. Materials and methods 2.1. Animal model Adult AB strain zebrafish for broodstock were obtained from G. fish Animal Model Co. (Taipei, Taiwan). Male and female adult fish were maintained separately in dechlorinated tap water at 28.5 ◦ C under a 14L:10D photoperiod, with a diet of commercial fish flakes and live Artemia nauplii. The fish had been maintained in our laboratory for one month prior to breeding. Zebrafish embryos were obtained from spawning adults in groups containing six females and three males overnight in spawning tanks. Eggs were collected on the next day and allowed to hatch. Hatched fry were maintained at 28.5 ◦ C in a 10 L tank with gentle aeration. From four day post hatch (dph) to 20 dph (right before the exposure experiment began), larval fish were fed twice a day with commercial fish powder, and the tank was cleaned daily by siphoning the bottom. The exposure was not started until 20 dph because prior to that time the larval survivorship was not stable even without any treatment and may mask any possible effect of PBDE-47 exposure.
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The experimental protocol involving handling and treatment of zebrafish was in accordance with the Taiwan national guidelines for animal welfare and was approved by the IACUC (Institutional Animal Care and Usage Committee) of National Museum of Marine Biology and Aquarium, Taiwan. 2.2. Experimental diets PBDE-47 standard (purity 98.6%, GC/MS certified) was obtained from ChemService (West Chester, PA, USA) and used to prepare a stock solution (1 mg/L) in methanol (J.T. Baker, Phillipsburg, NJ, USA). Fish powder (Chuankuan, Kaohsiung, Taiwan) or roughly ground TetraMin® MiniGranules fish pellets (Tetra, Melle, Germany) were used to prepare the experimental diets. During the 70 days of exposure period, the fish were gradually acclimated from fish powder to TetraMin (the feeding regime will be detailed later). For the PBDE treatment diets (nominal concentration = 10, 100, and 100 ng/g food wet weight for the low, medium, and high dose groups, respectively), an appropriate volume of the stock solution was added to a mixture of methanol and fish food (either fish powder or ground TetraMin; 1 mL methanol: 1 g fish food). The solvent control diet was prepared with similar procedures without adding any stock solution in the mixture. The mixtures were thoroughly blended and the methanol was completely evaporated in a hood at room temperature. For the control diet, neither methanol nor the stock solution was added. The experimental diets were stored at −20 ◦ C throughout the experiment. 2.3. Experimental design At 21 dph, 30 larval zebrafish were randomly attributed to each of 15 glass aquariums (30 cm × 45 cm × 30 cm with 40 L dechlorinated tap water). These aquariums were randomly divided to five experimental groups (i.e., control, solvent control, low, medium, and high dose groups). The aquariums were equipped with aerated power water filters and were maintained at 28 ± 1 ◦ C under a 14L:10D photoperiod throughout the experiment. The fish in each group were fed their respective diet twice a day from 21 to 90 dph. During the exposure period, the amount and type of food were adjusted with fish development based on our previous experience and fish size. Before reaching 40 dph, the zebrafish were fed the fish powder diet (0.04 g/tank/day). From 40 to 70 dph, the ration was adjusted to 0.06 g/tank/day of a mix of fish powder and TetraMin (fish powder:TetraMin = 2:1). After 70 dph, the fish were still fed 0.06 g/tank/day, but the fish powder: TetraMin ratio was changed to 1:2. Because of the very small particle sizes, the food immediately dispersed in the water after dispensing and the food consumption by individual fish was therefore not controlled during each feeding. Nevertheless, we assumed that the fish were fed to satiation because there was always a little food leftover on the tank bottom after the feeding. For each aquarium, 5 L of water was changed and the tank bottom was siphoned three times a week throughout the exposure period. At about one month after exposure began (54 dph), six fish from each tank were sampled for behavioral testing (detailed below). After the behavioral trials, the fish were returned to their respective tanks. At 90 dph when the fish were supposed to reach sexual maturation (Maack and Segner, 2003), all fish were euthanized in MS222 solution (ethyl 3-aminobenzoate methanesulfonate salt, Sigma, St. Louis, MO, USA). The fish in each tank were counted and their standard length (from the tip of the snout to the posterior end of the last vertebra) and wet weight were measured. The condition factor (K-factor) was calculated for each fish according to the formula: K-factor = (weight (g)/length (cm)3 ) × 100 (Brion et al., 2004). From each tank, three to five fish were wrapped in aluminum foil and
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stored at −20 ◦ C for tissue PBDE-47 measurement, and 5–14 fish were fixed in Dietrich’s fixative for histological analysis.
stage was counted using ImageJ software. Proportion of oocytes at each stage was calculated by dividing the oocyte count at respective stages by total oocyte count.
2.4. Behavioral testing The behavioral testing took place between 8 am and 12 noon at 54 dph. The fish were fast for 12 h prior the behavioral testing. From each tank, six fish were sampled and individually placed in each well (with 10 mL of dechlorinated water) of a 6-well microplate. The microplate was placed on the plastic holding stage, specifically designed to fit microplates (from 6- to 96-well plates), with infrared illumination from the bottom, and a CCD video camera was centrally positioned 110 cm above the holding stage. The holding stage and the video camera were surrounded by a black curtain during each trial to avoid any possible disturbance from the researcher. After 2 min of acclimation, the movement of each fish in the microplate was recorded with the video camera for 6 min. To stimulate predator-avoiding response of the fish, the holding stage was slightly knocked with a plastic hammer every 30 s during the third and forth minute. The video output from the camera was fed to a computer running an animal movement tracking software (Ethovison® XT, Noldus Information Technology, Wageningen, Netherlands). Fish locomotion parameters, including maximum velocity, total distance moved, and percent time active, were quantified from video with a tracking rate of 25 frames/s. Because fish swimming performance may be affected by body size, we measured standard length of the fish from the video image clips using ImageJ software (National Institute of Health, USA). Statistical analysis showed no significant difference of fish length between groups, indicating that body size was not a confounding factor in fish locomotion analysis. The fish behaviors were analyzed at 54 dph instead of 90 dph because otherwise the fish sizes will be too big for the 6-well microplates. 2.5. Histological preparation and gametogenesis analysis The number of fish processed for histological analysis was 27, 25, 20, 29, and 29 in the control, solvent control, low, medium, and high dose group, respectively. After routine tissue processing and paraffin embedding, whole fish were sectioned coronally into 5 m sections. For gonad visualization, sectioning was performed at a level simultaneously showing the gonads, the liver and the swim bladder (van der Ven et al., 2003). Sections were mounted on slides and stained with hematoxylin/eosin (H & E). Histology of gonad tissues was examined and fish gender was determined under an Olympus CX-31 light microscope equipped with an Olympus DE-71 digital camera (Olympus, Tokyo, Japan). The stage of testicular development was assessed in slides of fish that were histologically discerned to be male. At 40× magnification on the light microscope, six digital images of testicular area were randomly taken from each specimen, and four images were further randomly selected for image analysis. The size of testicular area, which containing spermatogonia, primary or secondary spermatocytes, and spermatids (mature sperm) (Weber et al., 2003), in each image was measured using ImageJ software. Relative area of spermatids was estimated by dividing spermatid area by total testicular area. The calculated values of four images from each specimen were averaged for statistical analysis. The stage of ovarian development was assessed in slides of fish that were histologically discerned to be female. For each specimen, a field of view covering the whole ovarian area was digitally imaged using an Olympus DE-71 digital camera connected to an Olympus ZX-61 dissecting microscope. Only oocytes at perinucleolar, cortical alveolus, vitellogenic, and mature stages were analyzed, thus excluding the smallest oocytes (oogonia) (Maack and Segner, 2003; Weber et al., 2003; Koc et al., 2008). The number of oocytes at each
2.6. Quantification of PBDE-47 in food and fish tissue PBDEs in food (2 g, n = 4) and fish tissues (exposed for 70 days; pooled samples of 3–5 individuals from each tank, n = 3) were extracted with an accelerated solvent extraction system (ASE 300 system, Dionex, Sunnyvale, CA, U.S.) using dichloromethane (DCM) as the solvent. One-tenth of the extract was designated for gravimetric lipid determination after solvent evaporation (Lauenstein et al., 2002). The clean-up steps of the extracts involved gelpermeation cleanup followed by Florisil columns (8 g, contain 2.5% water). The eluate was evaporated and dissolved in about 5 mL of n-hexane, then to a final volume of 1 mL with a gentle stream of nitrogen. PCB 30 and PCB 204 were added to each sample as internal standards. One blank vial spiked with 20 ng PBDE-47 standard was prepared as recovery standard. The sample analyses were preformed on a Varian 3800 gas chromatograph equipped with an autosampler, splitless injector, ion trap mass detector Varian Saturn 4000 (Varian Chromatography Systems, Walnut Creek, CA, USA). Recovery of PBDE-47 standard was 80.88 ± 6.41% (mean ± standard error). 2.7. Statistics Statistical analyses were performed with SYSTAT® (V. 12, Systat Software, Richmond, CA, USA). Shapiro–Wilk’s test was used to test the assumption of normality. Except the sex ratio data, which was analyzed with a chi-square test, differences between treatments were compared by analyses of variance (ANOVA), using “tank” as the experimental unit, i.e., mean values from each tank were used for the statistical analyses. Proportion data were adjusted with an arcsine square-root transformation. When the overall ANOVA was significant, Tukey’s test was performed for post-hoc pairwise comparisons. Correlations between tissue PBDE-47 concentrations (log10 transformed) and distance moved and proportion time active were further tested by simple linear regression analyses. Statistical significance was accepted for p < 0.05. Data were presented as mean ± standard error (SE). 3. Results 3.1. PBDE-47 concentrations in the diets and fish tissue Table 1 lists the levels of PBDE-47 measured in the experimental diets and fish whole-body tissue homogenates. For the diets, measured concentrations were lower than nominal concentrations, probably due to adsorption on the inner wall of the beakers during diet preparation. Tissue concentrations of PBDE-47 varied among groups as expected by the increased PBDE-47 levels in the dosed diets. 3.2. Survival and body size At 90 dph, survival in each group ranged from 64 to 83% and was not significantly different between groups (p = 0.21) (Fig. 1A). There was no statistical difference between groups for the standard length (p = 0.23) and wet weight (p = 0.12) (Fig. 1B–C). For those fish sampled for histological analysis (gender was thus determined), condition factors were similar between groups in both males (ranged from 1.66 to 1.81, p = 0.35) and females (ranged from 1.82 to 2.17, p = 0.56).
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Table 1 Measured PBDE-47 concentration (mean ± standard error) in experimental diets and fish tissue. Treatmenta
Dietb (ng/g wet weight)
Control Solvent control Low dose (10 ng/g) Medium dose (100 ng/g) High dose (1000 ng/g)
0.68 ± 0.28 1.34 ± 0.61 12.46 ± 2.03 81.28 ± 9.11 643.60 ± 57.68
a b c
Fish tissuec (ng/g wet weight)
(ng/g lipid weight)
<0.003 <0.003 2.70 ± 0.77 31.92 ± 4.34 409.68 ± 82.01
<0.02 <0.02 11.20 ± 2.48 189.04 ± 2.33 1924.16 ± 289.36
Values in parentheses are nominal concentrations of PBDE-47 in the diets. Sample size n = 4. Sample size n = 3.
3.3. Behavioral testing Fig. 2 shows the maximum velocity, total distance moved and percent time active of fish during the behavioral testing. Exposure to PBDE-47 did not cause a significant effect on maximum velocity (p = 0.42) (Fig. 2A). However, total distance moved and percent time active were both significantly affected (p = 0.04 and 0.01, respectively) (Fig. 2B–C). Tukey’s post-hoc pairwise comparison showed that total distance moved in high dose group (360.33 ± 20.84 cm) was significantly lower than that in control group (582.84 ± 57.27 cm) (p = 0.03) though not significantly different from the solvent control group (506.17 ± 34.76 cm) (p = 0.19). Data of percent time active showed a similar pattern; it was significantly lower in the high dose group (47.81 ± 3.40%) than that in the control group (76.17 ± 4.05%) (p = 0.01) but did not differ from the solvent control group (64.18 ± 5.52%) (p = 0.2). Simple linear regression analyses showed a significant dose-dependent decline in both total distance moved (r = −0.67, p = 0.007) and percent time active (r = −0.68, p = 0.005) with increasing tissue PBDE-47 concentrations (Fig. 3). 3.4. Gametogenesis analysis Two underdeveloped (standard length <10 mm) individuals in the high dose group had undifferentiated gonads, and we were thus not able to determine their gender histologically. Therefore, these two individuals were excluded from the statistical analysis. Sex ratio was not significantly different between groups (p = 0.9) (Fig. 4). No ovotestis was observed from the histological analysis. Relative area of spermatids in testicular tissue of male fish ranged from 37.9 to 46.2% and did not differ between groups (p = 0.75). Proportion of oocytes at each stage in females was not affected by PBDE-47 exposure, either (p > 0.5) (Fig. 5). 4. Discussion PBDE-47 tissue concentrations measured in the present study were comparable to that in wild fish reported in the literature. Tissue concentrations of PBDE-47 in the low and medium dose groups were 11 and 189 ng/g lipid, respectively, which were similar to that in fish collected from six rivers and three estuaries in Taiwan (PBDE47 = 17–176 ng/g lipid; total PBDEs = 25–281 ng/g lipid) (Peng et al., 2007). Tissue concentrations of PBDE-47 in the high dose group was 1924 ng/g lipid, which was close to that measured in fish from San Francisco Bay, USA (PBDE-47 = 122–1286 ng/g lipid; total PBDEs = 306–2235 ng/g lipid) (Holden et al., 2003). The tissue PBDE47 concentrations measured in our study were also much lower than that reported in previous bioassay studies (at the g/g level) using small laboratory fish such as fathead minnow and zebrafish (Lema et al., 2007, 2008; Kuiper et al., 2008). To our knowledge, the tissue PBDE-47 levels reported in this current study probably are the lowest compared with other fish bioassay studies in the literature. It should be noted that the tissue PBDE-47 concentrations
were analyzed at 90 dph, while the behavioral endpoints were measured at 54 dph. In a previous study with fathead minnow orally exposed to PBDE-47, tissue PBDE-47 increased rapidly and reached a plateau at around 10 days post exposure (Muirhead et al., 2006). We assumed that tissue residues at 90 dph were similar to that at 54 dph and correlated the chemical analysis data with the behavioral measurements (distance moved and time active) in statistical analyses. Very limited information is available so far concerning the effect of PBDEs on animal growth. PBDEs have been shown to suppress growth in rodent, amphibian, and fish (IPCS, 1994; Balch et al., 2006; Lema et al., 2007). However, increased growth has also been observed in PBDE-exposed bird and rat (Fernie et al., 2006; Suvorov et al., 2009), partially due to increased appetite (Fernie et al., 2006). Regarding bony fish, thyroid hormones are critical to the growth and development before fish reaching juvenile stage (Power et al., 2001). Recent studies showed that PBDE-47 exposure caused depressed plasma T4 levels in adult fathead minnows (Lema et al., 2008) and reduced growth in embryonic zebrafish (Lema et al., 2007). The thyroid gland of zebrafish becomes functional at the onset of feeding and then increases in activity throughout the larval and transitional periods (∼21 dph) to the juvenile stage (∼35 dph); exposure to goitrogens during this period of time can inhibit growth and development of larval zebrafish (Brown, 1997). In the present study, zebrafish body size parameters (standard length, weight, and condition factor) measured at 90 dph were not affected by PBDE-47 exposure. It is possible that the exposure levels of PBDE-47 in this study were not high enough to elicit an observable effect on fish size. It is also possible that the thyroid system, as well as growth, were inhibited during larval-juvenile stages, but with time the fish continued to grow and thus compensated for their reduced growth. Though thyroid hormones are critical to larval development and metamorphosis, fish growth after juvenile stage is more dependent on other factors, such as growth hormone, IFG-1, and insulin (Mommsen, 2001). In one of our recent studies, zebrafish orally exposed to PBDE-47 (74.9 ng/g food) from 20 to 60 dph were significantly smaller at 38 dph but not different in size from the control at 60 dph (Chen et al., 2010). Similar growth compensation has also been observed in zebrafish exposed to goitrogens (Brown, 1997). Histopathology is a valuable tool for assessing endocrinedisrupting effects on fish. Histopathological changes, especially those in the gonads, are indicative of dysfunctional reproduction and thus the negative effect on the population level can be deduced (van der Ven et al., 2003). Developmental exposure (from 2 to 60 dph in water) to estrogenic compounds, such as 4-nonylphenol (NP) and 17␣-ethinylestrodiol (EE), in zebrafish has been shown to cause suppression of gametogenesis in both male and female fish (Weber et al., 2003). Sex ratio is another crucial endpoint to determine feminization and masculinization in fish exposed to EDCs (Orn et al., 2006b). For instance, developmental exposure of zebrafish to EE and effluent water from a Swedish pulp mill from 1 to 60 dph caused 100% female and male-biased sex ratios, respectively (Orn et al., 2006a,b). Estrogenic or anti-androgenic activity
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Fig. 1. Standard length (A), wet weight (B), and survival rate (C) at 90 dph of zebrafish exposed to different levels of PBDE-47 via diet (C: control; S: solvent control; L: low dose; M: medium dose; H: high dose). Data are presented as mean ± standard error. There was no statistical difference between groups regarding standard length (p = 0.23), wet weight (p = 0.12), and survival rate (p = 0.21).
Fig. 2. Maximum swimming speed (A), total distance moved (B), and percent time active (C) measured at 54 dph of zebrafish exposed to different levels of PBDE-47 via diet (C: control; S: solvent control; L: low dose; M: medium dose; H: high dose). Data are presented as mean ± standard error. Different letters indicate significant difference between groups (p < 0.05).
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Fig. 4. Sex ratio of zebrafish exposed to different levels of PBDE-47 via diet (C: control; S: solvent control; L: low dose; M: medium dose; H: high dose). Data are presented as mean ± standard error. Sex ratio was similar between groups (p = 0.9).
exposure may not pose significant impact on gonad development in fish. Another great concern for the potential health effects of PBDEs comes from their developmental neurobehavioral toxicity (Birnbaum and Staskal, 2004). Developmental neurobehavioral toxicity of PBDEs has been extensively studied in rodents. Available studies involving rodents indicate pre- or postnatal exposure to various PBDEs (including PBDE-47) cause alterations in motor and cognitive functions, such as spontaneous behavior, learning, and memory (Branchi et al., 2003; Costa and Giordano, 2007).
Fig. 3. Correlation between tissue PBDE-47 concentrations (log10 transformed) and total distance moved (A) or proportion time active (arcsine square-root transformed) (B). The solid lines are the least-squares regression lines and the regression equations are given in the plot.
of PBDEs have been observed in both in vitro and in vivo studies (Talsness, 2008). Previous studies using histological examination revealed significantly reduced mature sperm in male fathead minnows orally exposed to PBDE-47 (Muirhead et al., 2006; Lema et al., 2008). In another study, adult zebrafish were exposed to analytically cleaned commercial penta-BDE mixture (DE-71, 5–500 g/L) for 30 days, but no major histopathological changes were observed in the gonads (Kuiper et al., 2008). In this present study, neither sex ratio nor gametogenesis was affected by PBDE-47 exposure. Zebrafish is a juvenile hermaphrodite, with all individuals having immature ovary – like gonads during early life and the bisexual gonad differentiation taking place only after about 7 weeks (49 days) post-fertilization (Maack and Segner, 2003). Therefore, our dietary exposure duration (from 20 to 90 dph) should be able to cover the period of gonad differentiation and maturation. The lack of observable effect of PBDE-47 on zebrafish gonad histology was probably due to our relatively low (but environmentally relevant) levels of exposure. Our results suggest that environmental PBDE-47
Fig. 5. Percentage of oocytes at each stage (Po: perinucleolar oocyte stage; Ca: cortical alveolus stage; Vo: vitellogenic oocyte stage; M: mature oocyte stage) in the ovaries of female zebrafish exposed to different levels of PBDE-47 via diet (C: control; S: solvent control; L: low dose; M: medium dose; H: high dose). Data are presented as mean ± standard error. Proportion of oocytes at each stage was similar between groups (p > 0.05).
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The mechanism underlying the neurobehavioral effects of PBDEs is as yet unclear. It has been proposed that disruption of thyroid hormones, which play a key role in brain development, and/or interactions with neurotransmitter systems may explain the neurotoxicity of PBDEs (Costa and Giordano, 2007; Gee and Moser, 2008). In contrast to the abundant literature regarding rodents, few studies have addressed neurobehavioral effects of PBDEs on fish. In the estuarine minnow (Fundulus heteroclitus) exposed to the commercial PBDE mixture DE-71 via water (0.001 and 0.01 g/L) during 0–7 day post-fertilization (dpf), hypoactivity and impaired fright response were observed at 4 dpf (Timme-Laragy et al., 2006). The authors also observed poorer predation performance in the hypoactive fish tested at 11 dpf (Timme-Laragy et al., 2006). In another study, embryonic exposure of zebrafish to waterborne PBDE47 (5000 g/L) resulted in reduced movement of cerebrospinal fluid in the brain ventricles and neural tubes, and the authors suggested that this may lead to the development of neural deficiencies later in life (Lema et al., 2007). In our current study, fright response (measured as maximum burst speed) was not affected by PBDE-47 exposure, but hypoactivity (i.e., decreased total distance moved and percent time active) did occur in the PBDE-47 treated fish. From an ecological point of view, capability of swimming is crucial for fish survival within the aquatic environment (Wolter and Arlinghaus, 2003). Aberrant locomotion behavior and activity level could have significant impact on predator–prey interactions, reproductive behavior, migration, and dispersal and thus decrease fitness of fish (Little and Finger, 1990; Vieira et al., 2009). For example, hypoactivity may result in less chance of encountering prey by reducing search areas and decrease the amount of energy for growth (Little and Finger, 1990). Previous literature showed that toxicants can disrupt startle responses, swimming activity, and ability to escape from a predator of prey fish (Scott and Sloman, 2004). Indeed, swimming behavior has long been proposed to be incorporated in test protocols to expend the sensitivity of standard toxicity test (Little and Finger, 1990). In this study, the behavioral endpoints are more sensitive than histological analysis to detect sublethal effects of PBDE-47. With modern electronic equipments and software, swimming behavior measurements can be rigorously monitored in the laboratory (Little and Finger, 1990; Kane et al., 2004). Our laboratory is currently testing the potential of this behavioral analysis system for high throughout measurement of behavioral responses of larval fish exposed to other contaminants. Because of their sensitivity and biological relevance, we suggest that quantitative behavioral measurements such as swimming activity should be more broadly included in standard aquatic toxicity assessment (Little and Finger, 1990; Kane et al., 2004; Scott and Sloman, 2004).
5. Conclusion Environmental dietary exposure of PBDE-47 to zebrafish during larval-juvenile stage affected fish locomotion behavior but not fish size and gonad development. Tissue PBDE-47 concentrations measured in this study were comparable to that reported in previous field studies, suggesting that wild fish with similar exposure level may be at risk of neurobehavioral toxicity. This study also demonstrated the utility of fish behavioral parameters in ecotoxicological assays. The video-based movement analysis system was shown to be useful for rapid, quantifiable, and high throughput assessment of neurobehavioral toxicity in zebrafish (Kane et al., 2004; Creton, 2009). We suggest that behavioral indicators such as swimming capability should be more widely included in fish bioassays for environmental contaminants (Scott and Sloman, 2004).
Acknowledgements This study was supported by the Thematic Research Grant from NMMBA (Grant No. 981003022) and a grant from the National Science Council of Taiwan, Republic of China (Grant NSC97-2313B291-001). We appreciate Dr. Chung-Der Hsiao for his advice for zebrafish breeding and histology and Yi-Ru Chen for the chemical analysis. References Alaee, M., 2003. Recommendations for monitoring of polybrominated diphenyl ethers in the Canadian environment. Environ. Monit. Assess. 88, 327–341. Balch, G.C., Velez-Espino, L.A., Sweet, C., Alaee, M., Metcalfe, C.D., 2006. Inhibition of metamorphosis in tadpoles of Xenopus laevis exposed to polybrominated diphenyl ethers (PBDEs). Chemosphere 64, 328–338. Beauvais, S.L., Jones, S.B., Parris, J.T., Brewer, S.K., Little, E.E., 2001. Cholinergic and behavioral neurotoxicity of carbaryl and cadmium to larval rainbow trout (Oncorhynchus mykiss). Ecotox. Environ. Safe. 49, 84–90. Birnbaum, L.S., Staskal, D.F., 2004. 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