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Accumulation, tissue-specific distribution and debromination of decabromodiphenyl ether (BDE 209) in European starlings (Sturnus vulgaris)
Environmental Pollution 148 (2007) 648e653 www.elsevier.com/locate/envpol
Accumulation, tissue-specific distribution and debromination of decabromodiphenyl ether (BDE 209) in European starlings (Sturnus vulgaris) E. Van den Steen a,*, A. Covaci b, V.L.B. Jaspers a, T. Dauwe a, S. Voorspoels b, M. Eens a, R. Pinxten a a
Department of Biology, University of Antwerp (Campus Drie Eiken), Universiteitsplein 1, 2610 Wilrijk, Belgium b Toxicological Centre, University of Antwerp, Universiteitsplein 1, 2610 Wilrijk, Belgium Received 6 July 2006; received in revised form 8 November 2006; accepted 13 November 2006
BDE 209 accumulates in the blood and tissues of a terrestrial bird species, the European starling, and can be debrominated to lower PBDE congeners. Abstract In this study we investigated the accumulation, tissue-specific distribution and possible debromination of BDE 209 in a terrestrial songbird species, the European starling, using silastic implants as a method of exposure. BDE 209 accumulated in the blood of the exposed starlings to a mean peak concentration of 16 4.1 ng/ml on day 10. After this peak, there was a decline to 3.3 0.4 ng/ml blood at the end of the exposure period of 76 days, which suggests elimination of BDE 209. In the exposed group, the muscle concentrations (461 ng/g lipid weight [lw], 430 ng/g lw) were about twofold those in liver (269 ng/g lw, 237 ng/g lw). In addition to BDE 209, other PBDE congeners, particularly octa- and nonaBDEs, were also present in the muscle and liver, suggesting bioformation from BDE 209. To our knowledge, these results are the first indications for the debromination of BDE 209 in birds. Ó 2006 Elsevier Ltd. All rights reserved. Keywords: Decabromodiphenyl ether; BDE 209; Debromination; Half-life; Bird; Silastic implant
1. Introduction Polybrominated diphenyl ethers (PBDEs) are a group of chemicals that are widely used in different materials because of their flame retarding properties. Intensive production and use have led to their ubiquitous presence in the environment and in biota, in which PBDE levels have increased rapidly (de Wit, 2002; Hites, 2004). Temporal studies have shown an exponential increase in PBDE concentrations in the eggs and tissues of different bird species (Norstrom et al., 2002). Some of the highest PBDE concentrations (maximum
* Corresponding author. Tel.: þ32 3 820 2285; fax: þ32 3 820 2271. E-mail address: [email protected] (E. Van den Steen). 0269-7491/$ - see front matter Ó 2006 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2006.11.017
64000 ng/g lipid weight) seen in wildlife to date, have been detected in the tissues of sparrowhawks (Accipiter nisus; Jaspers et al., 2006). Although information is limited, various toxic effects of PBDEs, including neurotoxicity and teratogenicity, have been reported in man and wildlife (McDonald, 2002; Darnerud, 2003). The pentaBDE and octaBDE commercial mixtures have been withdrawn from the market in Europe in 2004 (Directive EEC, 2003), because of their persistent, bioaccumulative and toxic characteristics. These products are now also facing bans in several states in the USA, and will be removed from the North American market by 2008 (California State Assembly, 2003). At present, the decaBDE commercial mixture is the only product which is still allowed for use. DecaBDE comprises approximately 80% of the world market
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demand for PBDEs, which in 2001 was reported at 56,100 metric tons (BSEF, 2003). The congener 2,20 ,3,30 ,4,40 ,5,50 ,6,60 -decabromodiphenyl ether (BDE 209) is the primary component of the decaBDE commercial mixture. Because of the high molecular weight and hydrophobicity of BDE 209, it is assumed that BDE 209 is not easily absorbed by the gut and that it consequently has a low oral bioavailability (IPCS, 1994). However, BDE 209 has been quantified in several free-living animals, including birds (Christensen et al., 2005; Jaspers et al., 2006; Lindberg et al., 2004; Voorspoels et al., 2006). Humans are also exposed, as evidenced by the detection of BDE 209 in the blood of occupationally exposed and unexposed individuals (Sjo¨din et al., 2001; Thomas et al., 2006; Thuresson et al., 2005). Dietary exposure studies with fish (Kierkegaard et al., 1999; Stapleton et al., 2004) and rats (Mo¨rck et al., 2003) have shown that BDE 209 is absorbed from the diet and that it can be debrominated by metabolic routes. Several studies have shown an inverse relationship between the potential toxicity and the number of bromine atoms among the BDE congeners (Darnerud et al., 2001; Meerts et al., 2001). Therefore, the major concern with respect to the debromination of BDE 209 is that these lower brominated congeners can be more toxic and bioaccumulative than the parent compound. Several methods of exposure have been used in toxicological studies. Each method has its own advantages and disadvantages depending on the study aims. To resemble natural conditions, the use of food as a method of exposure for hydrophobic pollutants is preferred (Andersson et al., 2001). Because of practical reasons concerning the preparation and administration of contaminated food, exposure via the food is not always possible. Alternatively, injections are often used to expose individuals to pollutants. A major drawback of this method is the fast release of the pollutants which may cause acute toxic effects. Repeated injections may cause wounds and other side effects. Another alternative method of exposure is the use of silastic implants, which have proven to be a useful method of manipulating hormone levels in birds (Ketterson et al., 1996; Pinxten et al., 2002). Although promising, this method has not often been used as a method of exposure for pollutants (Andersson et al., 2001; Harris and Osborn, 1981; Osborn and Harris, 1979; Van den Steen et al., 2007). The use of implants requires fewer manipulations (in contrast with repeated injections) and therefore normal behaviors are not disturbed. Another major advantage of the use of implants is that there is a slow release of pollutants for a prolonged period (i.e. 15 weeks, see Van den Steen et al., 2007), which make them very useful to study the effects of chronic exposure to pollutants. The aim of the present study was to investigate the accumulation and tissue-specific distribution of BDE 209 in a terrestrial songbird species, the European starling (Sturnus vulgaris), using silastic implants as a method of exposure. Attention was also given to the biotransformation of BDE 209, with respect to possible debromination to lower brominated congeners.
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2. Materials and methods 2.1. Experimental design Seven adult male starlings were housed in a large outdoor aviary (length width height ¼ 5 2 2.5 m). Food and water were provided ad libitum. We used silastic tubes (Degania silicone; length: 20 mm, inner diameter: 1.47 mm, outer diameter: 1.96 mm) to expose the starlings to an environmentally relevant concentration of BDE 209. One end of the implants was sealed with a medical adhesive and the other end was left open to allow slow diffusion of the BDE 209 solution. This method was adapted from Osborn and Harris (1979). The implants were inserted under the skin by a small incision to lie alongside the ribs. This implantation technique has previously been used to expose birds and fish to PCBs (Andersson et al., 2001; Harris and Osborn, 1981; Osborn and Harris, 1979). Silastic implants have recently also been successfully used to expose European starlings to PCB 153 (Van den Steen et al., 2007). BDE 209 (Wellington Laboratories, Guelph, ON, Canada) was dissolved in iso-octane and mixed in peanut oil (SigmaeAldrich). The iso-octane was removed by gently heating (40 C) the oil solution to a constant weight. After preparation, there were no detectable levels of other PBDE congeners in the oil solution. The exposure group (n ¼ 4) received an implantation dose of 46.8 2.2 mg BDE 209 and the control group (n ¼ 3) received an implant filled with unfortified peanut oil. At the end of the exposure period, the implants were removed and analyzed for BDE 209. Blood concentrations of BDE 209 were monitored by taking blood samples (w300 ml) every 3e7 days. After each blood sampling body mass was measured using a pesola balance. After an exposure period of 76 days, the birds were euthanized and the pectoral muscle and the liver were excised. Samples were stored at 20 C until further treatment. The Ethical Advisory Committee of the University of Antwerp approved this study following Belgian and Flemish laws concerning the protection of animal welfare.
2.2. Chemical analysis 2.2.1. Blood The method for whole blood analysis was adapted from the method described by Covaci and Voorspoels (2005) for the determination of PBDEs in serum. Approximately 300 ml of blood was spiked with internal standard (13C-BDE 209; Wellington Laboratories, Guelph, Canada), mixed with formic acid and extracted using solid-phase extraction cartridges (OasisÒ HLB, Waters, Milford, MA, USA). Clean-up was done by column chromatography on silica impregnated with concentrated sulfuric acid (48%, w/w). The cleaned extract was concentrated to 100 ml under a gentle nitrogen stream and transferred to an injection vial. 2.2.2. Muscle and liver Because of the low contamination levels, we pooled the tissue samples of all individuals of the control group. In the exposure group, the tissues were pooled per 2 individuals. The method used for the analysis of biological tissues has previously been described (Voorspoels et al., 2003) and is briefly presented below. Three to six grams of homogenized sample was dried using anhydrous Na2SO4, transferred into an extraction thimble and spiked with internal standards (BDE 77, BDE 128 and 13C-BDE 209). Extraction was carried out with 100 ml hexane/acetone (3:1, v/v) in an automated Soxhlet extractor (Bu¨chi, Flawil, Switzerland) in hot extraction mode for 2 h. The lipid content was determined gravimetrically on an aliquot of the extract, while the rest of the extract was cleaned up on a column filled with acidified silica (48%, w/w) and eluted with 15 ml hexane and 10 ml dichloromethane. The eluate was concentrated to 100 ml under a gentle nitrogen stream and transferred to an injection vial.
2.3. Instrumental analysis An Agilent 6890 gas chromatograph was connected via direct interface with a 5973 mass spectrometer (Agilent Technologies, Palo Alto, CA, USA). A 5 m 0.18 mm 0.18 mm DB-1 (J&W Scientific, Folsom, CA, USA) capillary column was used with helium as carrier gas at a constant
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flow of 0.8 ml/min. One ml was injected in the solvent vent mode (injector temperature program 90 C, kept for 0.06 min, then with 720 C/min to 300 C, vent time 0.04 min, vent flow 100 ml/min, initial pressure 2.9 psi) with the split outlet opened after 1.50 min. The temperature program of the oven was starting from 90 C, kept for 1.5 min, then with 15 C/min to 300 C, kept for 3 min. Interface, ion source and quadrupole temperatures were set at 300 C, 250 C, and 150 C, respectively. The mass spectrometer was operated in electron-capture negative ionization mode. Methane was used as a buffer gas. During the whole chromatographic run, ions m/z ¼ 484.7/486.7 and 494.7/496.7 were monitored for BDE 209 and 13 C-BDE 209, respectively, while ions m/z ¼ 79/81 were monitored for other PBDEs. The dwell time was 30 ms for each ion. Retention times and ion chromatograms were used as identification criteria.
2.4. Quality control Procedural blanks, which consisted of water instead of blood, were included with each sample batch. Procedural blanks were consistent and subtracted from the values found in the samples. Limits of quantification (LOQ) were calculated as 3 times the standard deviation of the analyte values in procedural blanks. For blood, LOQs for BDE 209 and other PBDE congeners were 0.8 and 0.5 ng/ml, respectively. For the tissues, LOQs for BDE 209 and other PBDE congeners were 5.5 and 1 ng/g lipid weight (lw), respectively. External quality control was assessed through regular successful participation (deviation from target values <20%) to the Arctic Monitoring and Assessment Programme (AMAP) and to the QUASIMEME proficiency exercises for PBDEs in environmental samples.
2.5. Statistical analysis Statistical calculations were performed using Statistica for Windows (Statsoft 1997) and GraphPad Instat 3.06 for Windows (GraphPad Software). The level of significance was set at a ¼ 0.05 throughout this study. Before statistical analysis, samples with levels below the LOQ were assigned a value of LOQ/2. Body mass and whole blood concentrations were log transformed to meet the assumptions of normality. Differences in whole blood concentrations were investigated among the treatment groups and during the exposure period using a two-way repeated-measures ANOVA. Differences in body mass were investigated using a two-way repeated-measures ANOVA on the first (until 41 days) and second half of the exposure period. We performed no statistical tests on the muscle and liver concentrations, because of the small sample sizes.
3. Results and discussion 3.1. Exposure dose At the end of the exposure period, the implants were removed and analyzed. There was 24.2 0.8 mg BDE 209 left in the implants. Thus, about 50% of the content of the implantation dose (46.8 2.2 mg) was released. Because one end of the implants was left open to allow slow diffusion of the BDE 209 solution (Osborn and Harris, 1979; Van den Steen et al., 2007), the implants were partially filled with blood and bird tissue. This may explain why the rest of the oil solution was not easily released. The use of implants with both end sealed and/or shorter implants may overcome this.
treatment groups both before and after implantation (twoway repeated-measures ANOVA: treatment: F1,5 ¼ 2.54, p ¼ 0.17) and no interaction between treatment and time (two-way repeated-measures ANOVA: treatment time: F9,45 ¼ 0.40, p ¼ 0.93). In the second part of the exposure period, body mass of the exposure group tended to be significantly lower than the control group (two-way repeatedmeasures ANOVA: treatment: F1,5 ¼ 4.84, p ¼ 0.08; Fig. 1). During this period, body mass did not change significantly (two-way repeated-measures ANOVA: time: F5,25 ¼ 0.55, p ¼ 0.73) and there was also no significant interaction between treatment and the exposure time (two-way repeated-measures ANOVA: treatment time: F5,25 ¼ 1.45, p ¼ 0.24). On the last day of the exposure period, blood concentrations tended to be negatively correlated with the body mass (Pearson’s correlation: r ¼ 0.71, p ¼ 0.07; n ¼ 7). These results suggest that exposure to BDE 209 may have a negative effect on the body mass of an individual. This effect may be caused by BDE 209 itself or by a lower brominated congener stemming from debromination of BDE 209 (see below). 3.3. BDE 209 concentrations 3.3.1. Whole blood concentrations of BDE 209 Before implantation, BDE 209 concentrations in the blood were below LOQ in both the control and the exposed group. Concentrations in the control group ranged from <0.8 ng/ml blood to 2.2 ng/ml blood throughout the exposure period (Fig. 2). In the beginning of the exposure period, BDE 209 accumulated in the blood of the exposed starlings from <0.8 ng/ml to a mean peak concentration of 16 4.1 ng/ml on day 10 (Fig. 2). After this peak, there was a decline in BDE 209 concentrations to 3.3 0.4 ng/ml blood at day 76 (Fig. 2), which suggest elimination of BDE 209. However, during the exposure period, no lower brominated BDE congeners could be detected in the blood, probably due to the low sample volume available. There was a significant interaction between treatment and the exposure time (two-way repeated-measures ANOVA: 100 Control group Exposed group
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Body mass changed significantly during the first half of the exposure period (two-way repeated-measures ANOVA: time: F9,45 ¼ 3.57, p ¼ 0.002; Fig. 1). During this period, there was no significant difference in body mass between the
0
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Days after implantation Fig. 1. Mean body mass with standard errors (g) of the control group (n ¼ 3) and the exposed group (n ¼ 4) after implantation with BDE 209.
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20
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Days after implantation Fig. 2. Mean blood concentrations of BDE 209 with standard errors (ng/ml blood) of the control group (n ¼ 3) and the exposed group (n ¼ 4) after implantation with BDE 209.
treatment time: F15,60 ¼ 9.87, p < 0.001), which confirms the usefulness of silastic implants as a method of exposure for organic pollutants (Andersson et al., 2001; Harris and Osborn, 1981; Osborn and Harris, 1979; Van den Steen et al., 2007). In the control group, BDE 209 concentrations during the exposure period did not significantly differ from the levels before implantation (paired t-tests: t < 2.0, p > 0.18) while in the exposed group, blood concentrations were significantly higher than concentrations before implantation (paired t-tests: t < 7.3, p < 0.005; significant after Hochberg correction). BDE 209 concentrations in the exposed group were significantly higher than in the control group from 3 days after implantation (one-way ANOVA: F1,5 > 24.17, p < 0.004; significant after Hochberg correction). To determine the apparent half-life of BDE 209 in the blood of the starlings, we performed a linear regression analysis (GraphPad Prism 4 for Windows) on the mean BDE 209 concentrations after the peak versus time. Based on the linear equation ( y ¼ 0.15x þ 15.68; r2 ¼ 0.90, p < 0.0001), the half-life of BDE 209 in the blood of the starlings was estimated at 13 days (95% confidence interval: 11e18 days). Several observations indicate that BDE 209 is degraded or eliminated at a much higher rate than lower brominated PBDE congeners. Sandholm et al. (2003) observed a half-life of 2.5 days in rats intravenously dosed with BDE 209. The half-life of BDE 209 assessed in gray seals, 8.5e13 days (Thomas et al., 2005), and in occupationally exposed workers, 15 days (Thuresson et al., 2006), agrees well with our estimate of 13 days. Although the calculated half-life in this study is plausible, it may actually be an overestimation, because there most likely was still release out of the implants during the apparent elimination phase (Fig. 2) since silastic implants are supposed to be responsible for a continuous release.
those in liver (muscle: 461 ng/g lw, 430 ng/g lw; liver: 269 ng/ g lw, 237 ng/g lw). These results suggest accumulation of BDE 209 in bird tissues. Recently, BDE 209 has been detected in a variety of species, mainly those living in terrestrial ecosystems (de Boer et al., 2004; Jaspers et al., 2006; Voorspoels et al., 2006). Compared to these studies, levels in the present study were higher range environmental concentrations of BDE 209. Muscle and liver concentrations were 7 and 4 times, respectively, higher than in predatory birds from Belgium (Jaspers et al., 2006). The higher concentrations in the muscle compared to the liver are probably due to a higher metabolic activity in liver than in muscle (Voet and Voet, 1995). Fig. 3 shows, in addition to BDE 209, the presence of other PBDE congeners in the muscle and liver. The PBDE profiles in muscle and liver were very similar (Fig. 3). However, the relative contribution of the nonaBDEs was greater in liver than in muscle (BDE 206: liver, 4.0%; muscle, 3.7%; BDE 207: liver, 24%; muscle, 18%; BDE 208: liver, 9.2%; muscle, 6.2%), which is probably due to the high metabolic activity of the liver (Voet and Voet, 1995). Differences between the control group and the exposed group were most pronounced for the nona- (BDE 206, BDE 207 and BDE 208) and octaBDEs (BDE 196 and BDE 197), suggesting bioformation from BDE 209 (Fig. 3). The octaBDEs, BDE 203 and BDE 205, did not differ much between both groups. A possible
(a) BDE concentrations in muscle (ng/g lipid weight)
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3.3.2. BDE 209 concentrations in muscle and liver Tissue concentrations were below the LOQ in the control group (muscle: <5.6 ng/g lw, liver: <2.9 ng/g lw). In the exposed group, the muscle concentrations were about twofold
47 99 100 153 154 183 196 197 203 205 206 207 208
PBDE congener Fig. 3. Profiles of PBDE congeners in (a) muscle and (b) liver (ng/g lw) after implantation with BDE 209.
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explanation for the differences in the formation of these octaBDEs is the position of the bromine atoms. Hexa- (BDE 153) and heptaBDEs (BDE 183) differed less between the control group and the exposed group (Fig. 3). For the other congeners, there were no notable differences between the groups (Fig. 3). Our results suggest that BDE 209 is degraded to lower brominated congeners in tissues of European starlings. Although it can not be excluded that a small fraction of the octa- and nonaBDEs found in the tissues stems from the bioaccumulation of eventual trace octa- and nonaBDEs in the oil, it is mostly plausible to assume that the presence of these congeners is the result of the debromination of BDE 209. The formation of lower brominated congeners from decaBDE has previously been reported in rats (Mo¨rck et al., 2003) and fish (Kierkegaard et al., 1999; Stapleton et al., 2004). Similar to those in fish (Kierkegaard et al., 1999; Stapleton et al., 2004), our results suggested that BDE 209 can be debrominated down to hexaBDEs. For the nonaBDEs, we observed the following order: BDE 207 > BDE208 > BDE 206. A similar pattern was found resulting from thermal decomposition of BDE 209 (Koryta´r et al., 2005), while other patterns were observed for photolytic degradation (Soderstro¨m et al., 2004) or biotransformation in fish (Mo¨rck et al., 2003). To our knowledge, these results are the first indications for the debromination of BDE 209 in birds. The present study, in accordance with previous studies (Kierkegaard et al., 1999; Mo¨rck et al., 2003; Stapleton et al., 2004), suggested that BDE 209 can be debrominated to congeners that are also present in the penta- and octaBDE commercial mixtures, which are no longer allowed for use due to their potential toxicity (Darnerud, 2003). Thus, phasing out the penta- and octaBDE commercial mixtures could be insufficient to restrict the potential risk of lower brominated congeners. These results are a great cause of concern considering the large amounts of decaBDE that are worldwide used (BSEF, 2003). More studies are needed to elucidate the toxicology and metabolism of higher brominated BDEs, especially BDE 209. The results of these studies should be taken into account for a readjusted risk assessment of the decaBDE commercial mixture. Acknowledgments This study was supported by the Fund for Scientific Research Flanders (FWO-project G.0137.04), an FWO Postdoctoral Fellowship to T.D., an FWO Research Assistantship Grant to V.J. and a Research Grant from the Research Funds of the University of Antwerp to E.V.d.S and A.C. References Andersson, P.L., Berg, A.H., Bjerselius, R., Norrgren, L., Olse´n, H., ¨ rn, S., Tysklind, M., 2001. Bioaccumulation of selected Olsson, P.E., O PCBs in zebrafish, three-spined stickleback, and Arctic char after three different routes of exposure. Archives of Environmental Contamination and Toxicology 40, 519e530. de Boer, J., Leslie, H.A., Leonards, P.E.G., Bersuder, P., Morris, S., Allchin, C.R., 2004. Screening and time trend study of decabromodiphenylether and hexabromocyclododecane in birds. Proceedings of the Third International Workshop on Brominated Flame Retardants, Toronto, Canada, 125e128.
BSEF, 2003. Major brominated flame retardants volume estimates: Total market demand by region in 2001. Bromine Science and Environmental Forum, Brussels. Available at: California State Assembly, 2003. Assembly Bill No. 302. California State Assembly, Sacramento, CA. Christensen, J.R., MacDuffee, M., Macdonald, R.W., Whiticar, M., Ross, P.S., 2005. Persistent organic pollutants in British Columbia grizzly bears: consequence of divergent diets. Environmental Science and Technology 39, 6952e6960. Covaci, A., Voorspoels, S., 2005. Optimization of the determination of polybrominated diphenyl ethers in human serum using solid-phase extraction and gas chromatography-electron capture negative ionization mass spectrometry. Journal of Chromatography B 827, 216e223. Darnerud, P.O., 2003. Toxic effects of polybrominated diphenyl ethers in man and in wildlife. Environment International 29, 841e853. Darnerud, P.O., Eriksen, G.S., Johannesson, T., Larsen, P.B., Viluksela, M., 2001. Polybrominated diphenyl ethers: occurrence, dietary exposure, and toxicology. Environmental Health Perspectives 109 (Suppl 1), 49e68. EEC, 2003. Directive 2003/11/EC of the European Parliament and of the Council of 6 February 2003 amending for the 24th time Council directive 76/769/EEC relating to restrictions on the marketing and use of certain dangerous substances and preparations (pentabromodiphenyl ether and octabromodiphenyl ether). Official Journal L 042. 15/02/2003. Harris, M.P., Osborn, D., 1981. Effect of a polychlorinated biphenyl on the survival and breeding of puffins. Journal of Applied Ecology 18, 471e479. Hites, R.A., 2004. Polybrominated diphenylethers in the environment and in people: A meta-analysis of concentrations. Environmental Science and Technology 38, 945e956. IPCS, 1994. Brominated diphenyl ethers. In: Environmental Health Criteria 162. International Programme on Chemical Safety, World Health Organization, Geneva. Jaspers, V.L.B., Covaci, A., Voorspoels, S., Dauwe, T., Eens, M., Schepens, P., 2006. Brominated flame retardants and organochlorine pollutants in aquatic and terrestrial predatory birds of Belgium: Levels, patterns, tissue distribution and condition factors. Environmental Pollution 139, 340e352. Ketterson, E.D., Nolan Jr., V., Cawthorn, M.J., Parker, P.G., Ziegenfus, C., 1996. Phenotypic engineering: using hormones to explore the mechanistic and functional base of phenotypic variation in nature. Ibis 138, 70e86. Kierkegaard, A., Balk, L., Tja¨rnlund, U., de Wit, C.A., Jansson, B., 1999. Dietary uptake and biological effects of decabromodiphenyl ether in rainbow trout (Oncorhynchus mykiss). Environmental Science and Technology 33, 1612e1617. Koryta´r, P., Covaci, A., Leonards, P.E.G., de Boer, J., Brinkman, U.A.Th., 2005. Comprehensive two-dimensional gas chromatography of polybrominated diphenyl ethers. Journal of Chromatography A 1100, 200e207. Lindberg, P., Sellstro¨m, U., Ha¨ggberg, L., de Wit, C., 2004. Higher brominated diphenyl ethers and hexabromocyclododecane found in eggs of peregrine falcons (Falco peregrinus) breeding in Sweden. Environmental Science and Technology 38, 93e96. McDonald, T.A., 2002. A perspective on the potential health risks of PBDEs. Chemosphere 46, 745e755. Meerts, I.A., Letcher, R.J., Hoving, S., Marsh, G., Bergman, A., Lemmen, J.G., van der Burg, B., Brouwer, A., 2001. In vitro estrogenicity of polybrominated diphenyl ethers, hydroxylated PBDEs, and polybrominated bisphenol A compounds. Environmental Health Perspectives 109, 399e407. ¨ rn, U., Wehler, E.K., 2003. Decabromodiphenyl ether Mo¨rck, A., Hakk, H., O in the rat: absorption, distribution, metabolism, and excretion. Drug Metabolism and Disposition 31, 900e907. Norstrom, R.J., Simon, M., Moisey, J., Wakeford, B., Weseloh, D.V., 2002. Geographical distribution and temporal trends (1981e2000) of brominated diphenyl ethers in Great Lakes herring gull eggs. Environmental Science and Technology 36, 4783e4789. Osborn, D., Harris, M.P., 1979. A procedure for implanting a slow release formulation of an environmental pollutant into a free-living animal. Environmental Pollution 19, 139e144.
E. Van den Steen et al. / Environmental Pollution 148 (2007) 648e653 Pinxten, R., De Ridder, E., Balthazart, J., Eens, M., 2002. Context-dependent effects of castration and testosterone treatment on song in male European starlings. Hormones and Behavior 42, 307e318. Sandholm, A., Emanuelsson, B.M., Wehler, E.K., 2003. Bioavailability and half-life of decabromodiphenyl ether (BDE-209) in rat. Xenobiotica 33, 1149e1158. ˚ , 2001. Brominated flame retardants Sjo¨din, A., Patterson Jr., D.G., Bergman, A in serum from U.S. blood donors. Environmental Science and Technology 35, 3830e3833. Soderstro¨m, G., Sellstro¨m, U., de Wit, C.A., Tysklind, M., 2004. Photolytic debromination of decabromodiphenyl ether (BDE 209). Environmental Science and Technology 38, 127e132. Stapleton, H.M., Alaee, M., Letcher, R.J., Baker, J.E., 2004. Debromination of the flame retardant decabromodiphenyl ether by juvenile carp (Cyprinus carpio) following dietary exposure. Environmental Science and Technology 38, 112e119. Thomas, G.O., Moss, S.E.W., Asplund, L., Hall, A.J., 2005. Absorption of decabromodiphenyl ether and other organohalogen chemicals by grey seals (Halichoerus grypus). Environmental Pollution 133, 581e586. Thomas, G.O., Wilkinson, M., Hodson, S., Jones, K.C., 2006. Organohalogen chemicals in human blood from the United Kingdom. Environmental Pollution 141, 30e41.
653
Thuresson, K., Bergman, A., Jakobsson, K., 2005. Occupational exposure to commercial decabromodiphenyl ether in workers manufacturing or handling flame-retarded rubber. Environmental Science and Technology 39, 1980e1986. ˚ ., Thuresson, K., Ho¨glund, P., Hagmar, L., Sjo¨din, A., Bergman, A Jakobsson, K., 2006. Apparent half-lives of hepta- to decabrominated diphenyl ethers in human serum as determined in occupationally exposed workers. Environmental Health Perspectives 114, 176e181. Van den Steen, E., Covaci, A., Jaspers, V.L.B., Dauwe, T., Voorspoels, S., Eens, M., Pinxten, R., 2007. Experimental evaluation of the usefulness of feathers as a non-destructive biomonitor for polychlorinated biphenyls (PCBs) using silastic implants as a novel method of exposure. Environment International 33, 257e264. Voet, D., Voet, J.G., 1995. Biochemistry, second ed. Wiley, New York. Voorspoels, S., Covaci, A., Lepom, P., Jaspers, V.L.B., Schepens, P., 2006. Levels and distribution of polybrominated diphenyl ethers in various tissues of birds of prey. Environmental Pollution 144, 218e227. Voorspoels, S., Covaci, A., Schepens, P., 2003. Polybrominated diphenyl ethers in marine species from the Belgian North Sea and Western Scheldt Estuary: Levels, profiles, and distribution. Environmental Science and Technology 37, 4348e4357. de Wit, C.A., 2002. An overview of brominated flame retardants in the environment. Chemosphere 46, 583e624.
Accumulation, tissue-specific distribution and debromination of decabromodiphenyl ether (BDE 209) in European starlings (Sturnus vulgaris) E. Van den Steen a,*, A. Covaci b, V.L.B. Jaspers a, T. Dauwe a, S. Voorspoels b, M. Eens a, R. Pinxten a a
Department of Biology, University of Antwerp (Campus Drie Eiken), Universiteitsplein 1, 2610 Wilrijk, Belgium b Toxicological Centre, University of Antwerp, Universiteitsplein 1, 2610 Wilrijk, Belgium Received 6 July 2006; received in revised form 8 November 2006; accepted 13 November 2006
BDE 209 accumulates in the blood and tissues of a terrestrial bird species, the European starling, and can be debrominated to lower PBDE congeners. Abstract In this study we investigated the accumulation, tissue-specific distribution and possible debromination of BDE 209 in a terrestrial songbird species, the European starling, using silastic implants as a method of exposure. BDE 209 accumulated in the blood of the exposed starlings to a mean peak concentration of 16 4.1 ng/ml on day 10. After this peak, there was a decline to 3.3 0.4 ng/ml blood at the end of the exposure period of 76 days, which suggests elimination of BDE 209. In the exposed group, the muscle concentrations (461 ng/g lipid weight [lw], 430 ng/g lw) were about twofold those in liver (269 ng/g lw, 237 ng/g lw). In addition to BDE 209, other PBDE congeners, particularly octa- and nonaBDEs, were also present in the muscle and liver, suggesting bioformation from BDE 209. To our knowledge, these results are the first indications for the debromination of BDE 209 in birds. Ó 2006 Elsevier Ltd. All rights reserved. Keywords: Decabromodiphenyl ether; BDE 209; Debromination; Half-life; Bird; Silastic implant
1. Introduction Polybrominated diphenyl ethers (PBDEs) are a group of chemicals that are widely used in different materials because of their flame retarding properties. Intensive production and use have led to their ubiquitous presence in the environment and in biota, in which PBDE levels have increased rapidly (de Wit, 2002; Hites, 2004). Temporal studies have shown an exponential increase in PBDE concentrations in the eggs and tissues of different bird species (Norstrom et al., 2002). Some of the highest PBDE concentrations (maximum
* Corresponding author. Tel.: þ32 3 820 2285; fax: þ32 3 820 2271. E-mail address: [email protected] (E. Van den Steen). 0269-7491/$ - see front matter Ó 2006 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2006.11.017
64000 ng/g lipid weight) seen in wildlife to date, have been detected in the tissues of sparrowhawks (Accipiter nisus; Jaspers et al., 2006). Although information is limited, various toxic effects of PBDEs, including neurotoxicity and teratogenicity, have been reported in man and wildlife (McDonald, 2002; Darnerud, 2003). The pentaBDE and octaBDE commercial mixtures have been withdrawn from the market in Europe in 2004 (Directive EEC, 2003), because of their persistent, bioaccumulative and toxic characteristics. These products are now also facing bans in several states in the USA, and will be removed from the North American market by 2008 (California State Assembly, 2003). At present, the decaBDE commercial mixture is the only product which is still allowed for use. DecaBDE comprises approximately 80% of the world market
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demand for PBDEs, which in 2001 was reported at 56,100 metric tons (BSEF, 2003). The congener 2,20 ,3,30 ,4,40 ,5,50 ,6,60 -decabromodiphenyl ether (BDE 209) is the primary component of the decaBDE commercial mixture. Because of the high molecular weight and hydrophobicity of BDE 209, it is assumed that BDE 209 is not easily absorbed by the gut and that it consequently has a low oral bioavailability (IPCS, 1994). However, BDE 209 has been quantified in several free-living animals, including birds (Christensen et al., 2005; Jaspers et al., 2006; Lindberg et al., 2004; Voorspoels et al., 2006). Humans are also exposed, as evidenced by the detection of BDE 209 in the blood of occupationally exposed and unexposed individuals (Sjo¨din et al., 2001; Thomas et al., 2006; Thuresson et al., 2005). Dietary exposure studies with fish (Kierkegaard et al., 1999; Stapleton et al., 2004) and rats (Mo¨rck et al., 2003) have shown that BDE 209 is absorbed from the diet and that it can be debrominated by metabolic routes. Several studies have shown an inverse relationship between the potential toxicity and the number of bromine atoms among the BDE congeners (Darnerud et al., 2001; Meerts et al., 2001). Therefore, the major concern with respect to the debromination of BDE 209 is that these lower brominated congeners can be more toxic and bioaccumulative than the parent compound. Several methods of exposure have been used in toxicological studies. Each method has its own advantages and disadvantages depending on the study aims. To resemble natural conditions, the use of food as a method of exposure for hydrophobic pollutants is preferred (Andersson et al., 2001). Because of practical reasons concerning the preparation and administration of contaminated food, exposure via the food is not always possible. Alternatively, injections are often used to expose individuals to pollutants. A major drawback of this method is the fast release of the pollutants which may cause acute toxic effects. Repeated injections may cause wounds and other side effects. Another alternative method of exposure is the use of silastic implants, which have proven to be a useful method of manipulating hormone levels in birds (Ketterson et al., 1996; Pinxten et al., 2002). Although promising, this method has not often been used as a method of exposure for pollutants (Andersson et al., 2001; Harris and Osborn, 1981; Osborn and Harris, 1979; Van den Steen et al., 2007). The use of implants requires fewer manipulations (in contrast with repeated injections) and therefore normal behaviors are not disturbed. Another major advantage of the use of implants is that there is a slow release of pollutants for a prolonged period (i.e. 15 weeks, see Van den Steen et al., 2007), which make them very useful to study the effects of chronic exposure to pollutants. The aim of the present study was to investigate the accumulation and tissue-specific distribution of BDE 209 in a terrestrial songbird species, the European starling (Sturnus vulgaris), using silastic implants as a method of exposure. Attention was also given to the biotransformation of BDE 209, with respect to possible debromination to lower brominated congeners.
649
2. Materials and methods 2.1. Experimental design Seven adult male starlings were housed in a large outdoor aviary (length width height ¼ 5 2 2.5 m). Food and water were provided ad libitum. We used silastic tubes (Degania silicone; length: 20 mm, inner diameter: 1.47 mm, outer diameter: 1.96 mm) to expose the starlings to an environmentally relevant concentration of BDE 209. One end of the implants was sealed with a medical adhesive and the other end was left open to allow slow diffusion of the BDE 209 solution. This method was adapted from Osborn and Harris (1979). The implants were inserted under the skin by a small incision to lie alongside the ribs. This implantation technique has previously been used to expose birds and fish to PCBs (Andersson et al., 2001; Harris and Osborn, 1981; Osborn and Harris, 1979). Silastic implants have recently also been successfully used to expose European starlings to PCB 153 (Van den Steen et al., 2007). BDE 209 (Wellington Laboratories, Guelph, ON, Canada) was dissolved in iso-octane and mixed in peanut oil (SigmaeAldrich). The iso-octane was removed by gently heating (40 C) the oil solution to a constant weight. After preparation, there were no detectable levels of other PBDE congeners in the oil solution. The exposure group (n ¼ 4) received an implantation dose of 46.8 2.2 mg BDE 209 and the control group (n ¼ 3) received an implant filled with unfortified peanut oil. At the end of the exposure period, the implants were removed and analyzed for BDE 209. Blood concentrations of BDE 209 were monitored by taking blood samples (w300 ml) every 3e7 days. After each blood sampling body mass was measured using a pesola balance. After an exposure period of 76 days, the birds were euthanized and the pectoral muscle and the liver were excised. Samples were stored at 20 C until further treatment. The Ethical Advisory Committee of the University of Antwerp approved this study following Belgian and Flemish laws concerning the protection of animal welfare.
2.2. Chemical analysis 2.2.1. Blood The method for whole blood analysis was adapted from the method described by Covaci and Voorspoels (2005) for the determination of PBDEs in serum. Approximately 300 ml of blood was spiked with internal standard (13C-BDE 209; Wellington Laboratories, Guelph, Canada), mixed with formic acid and extracted using solid-phase extraction cartridges (OasisÒ HLB, Waters, Milford, MA, USA). Clean-up was done by column chromatography on silica impregnated with concentrated sulfuric acid (48%, w/w). The cleaned extract was concentrated to 100 ml under a gentle nitrogen stream and transferred to an injection vial. 2.2.2. Muscle and liver Because of the low contamination levels, we pooled the tissue samples of all individuals of the control group. In the exposure group, the tissues were pooled per 2 individuals. The method used for the analysis of biological tissues has previously been described (Voorspoels et al., 2003) and is briefly presented below. Three to six grams of homogenized sample was dried using anhydrous Na2SO4, transferred into an extraction thimble and spiked with internal standards (BDE 77, BDE 128 and 13C-BDE 209). Extraction was carried out with 100 ml hexane/acetone (3:1, v/v) in an automated Soxhlet extractor (Bu¨chi, Flawil, Switzerland) in hot extraction mode for 2 h. The lipid content was determined gravimetrically on an aliquot of the extract, while the rest of the extract was cleaned up on a column filled with acidified silica (48%, w/w) and eluted with 15 ml hexane and 10 ml dichloromethane. The eluate was concentrated to 100 ml under a gentle nitrogen stream and transferred to an injection vial.
2.3. Instrumental analysis An Agilent 6890 gas chromatograph was connected via direct interface with a 5973 mass spectrometer (Agilent Technologies, Palo Alto, CA, USA). A 5 m 0.18 mm 0.18 mm DB-1 (J&W Scientific, Folsom, CA, USA) capillary column was used with helium as carrier gas at a constant
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flow of 0.8 ml/min. One ml was injected in the solvent vent mode (injector temperature program 90 C, kept for 0.06 min, then with 720 C/min to 300 C, vent time 0.04 min, vent flow 100 ml/min, initial pressure 2.9 psi) with the split outlet opened after 1.50 min. The temperature program of the oven was starting from 90 C, kept for 1.5 min, then with 15 C/min to 300 C, kept for 3 min. Interface, ion source and quadrupole temperatures were set at 300 C, 250 C, and 150 C, respectively. The mass spectrometer was operated in electron-capture negative ionization mode. Methane was used as a buffer gas. During the whole chromatographic run, ions m/z ¼ 484.7/486.7 and 494.7/496.7 were monitored for BDE 209 and 13 C-BDE 209, respectively, while ions m/z ¼ 79/81 were monitored for other PBDEs. The dwell time was 30 ms for each ion. Retention times and ion chromatograms were used as identification criteria.
2.4. Quality control Procedural blanks, which consisted of water instead of blood, were included with each sample batch. Procedural blanks were consistent and subtracted from the values found in the samples. Limits of quantification (LOQ) were calculated as 3 times the standard deviation of the analyte values in procedural blanks. For blood, LOQs for BDE 209 and other PBDE congeners were 0.8 and 0.5 ng/ml, respectively. For the tissues, LOQs for BDE 209 and other PBDE congeners were 5.5 and 1 ng/g lipid weight (lw), respectively. External quality control was assessed through regular successful participation (deviation from target values <20%) to the Arctic Monitoring and Assessment Programme (AMAP) and to the QUASIMEME proficiency exercises for PBDEs in environmental samples.
2.5. Statistical analysis Statistical calculations were performed using Statistica for Windows (Statsoft 1997) and GraphPad Instat 3.06 for Windows (GraphPad Software). The level of significance was set at a ¼ 0.05 throughout this study. Before statistical analysis, samples with levels below the LOQ were assigned a value of LOQ/2. Body mass and whole blood concentrations were log transformed to meet the assumptions of normality. Differences in whole blood concentrations were investigated among the treatment groups and during the exposure period using a two-way repeated-measures ANOVA. Differences in body mass were investigated using a two-way repeated-measures ANOVA on the first (until 41 days) and second half of the exposure period. We performed no statistical tests on the muscle and liver concentrations, because of the small sample sizes.
3. Results and discussion 3.1. Exposure dose At the end of the exposure period, the implants were removed and analyzed. There was 24.2 0.8 mg BDE 209 left in the implants. Thus, about 50% of the content of the implantation dose (46.8 2.2 mg) was released. Because one end of the implants was left open to allow slow diffusion of the BDE 209 solution (Osborn and Harris, 1979; Van den Steen et al., 2007), the implants were partially filled with blood and bird tissue. This may explain why the rest of the oil solution was not easily released. The use of implants with both end sealed and/or shorter implants may overcome this.
treatment groups both before and after implantation (twoway repeated-measures ANOVA: treatment: F1,5 ¼ 2.54, p ¼ 0.17) and no interaction between treatment and time (two-way repeated-measures ANOVA: treatment time: F9,45 ¼ 0.40, p ¼ 0.93). In the second part of the exposure period, body mass of the exposure group tended to be significantly lower than the control group (two-way repeatedmeasures ANOVA: treatment: F1,5 ¼ 4.84, p ¼ 0.08; Fig. 1). During this period, body mass did not change significantly (two-way repeated-measures ANOVA: time: F5,25 ¼ 0.55, p ¼ 0.73) and there was also no significant interaction between treatment and the exposure time (two-way repeated-measures ANOVA: treatment time: F5,25 ¼ 1.45, p ¼ 0.24). On the last day of the exposure period, blood concentrations tended to be negatively correlated with the body mass (Pearson’s correlation: r ¼ 0.71, p ¼ 0.07; n ¼ 7). These results suggest that exposure to BDE 209 may have a negative effect on the body mass of an individual. This effect may be caused by BDE 209 itself or by a lower brominated congener stemming from debromination of BDE 209 (see below). 3.3. BDE 209 concentrations 3.3.1. Whole blood concentrations of BDE 209 Before implantation, BDE 209 concentrations in the blood were below LOQ in both the control and the exposed group. Concentrations in the control group ranged from <0.8 ng/ml blood to 2.2 ng/ml blood throughout the exposure period (Fig. 2). In the beginning of the exposure period, BDE 209 accumulated in the blood of the exposed starlings from <0.8 ng/ml to a mean peak concentration of 16 4.1 ng/ml on day 10 (Fig. 2). After this peak, there was a decline in BDE 209 concentrations to 3.3 0.4 ng/ml blood at day 76 (Fig. 2), which suggest elimination of BDE 209. However, during the exposure period, no lower brominated BDE congeners could be detected in the blood, probably due to the low sample volume available. There was a significant interaction between treatment and the exposure time (two-way repeated-measures ANOVA: 100 Control group Exposed group
95
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Body mass changed significantly during the first half of the exposure period (two-way repeated-measures ANOVA: time: F9,45 ¼ 3.57, p ¼ 0.002; Fig. 1). During this period, there was no significant difference in body mass between the
0
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Days after implantation Fig. 1. Mean body mass with standard errors (g) of the control group (n ¼ 3) and the exposed group (n ¼ 4) after implantation with BDE 209.
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20
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treatment time: F15,60 ¼ 9.87, p < 0.001), which confirms the usefulness of silastic implants as a method of exposure for organic pollutants (Andersson et al., 2001; Harris and Osborn, 1981; Osborn and Harris, 1979; Van den Steen et al., 2007). In the control group, BDE 209 concentrations during the exposure period did not significantly differ from the levels before implantation (paired t-tests: t < 2.0, p > 0.18) while in the exposed group, blood concentrations were significantly higher than concentrations before implantation (paired t-tests: t < 7.3, p < 0.005; significant after Hochberg correction). BDE 209 concentrations in the exposed group were significantly higher than in the control group from 3 days after implantation (one-way ANOVA: F1,5 > 24.17, p < 0.004; significant after Hochberg correction). To determine the apparent half-life of BDE 209 in the blood of the starlings, we performed a linear regression analysis (GraphPad Prism 4 for Windows) on the mean BDE 209 concentrations after the peak versus time. Based on the linear equation ( y ¼ 0.15x þ 15.68; r2 ¼ 0.90, p < 0.0001), the half-life of BDE 209 in the blood of the starlings was estimated at 13 days (95% confidence interval: 11e18 days). Several observations indicate that BDE 209 is degraded or eliminated at a much higher rate than lower brominated PBDE congeners. Sandholm et al. (2003) observed a half-life of 2.5 days in rats intravenously dosed with BDE 209. The half-life of BDE 209 assessed in gray seals, 8.5e13 days (Thomas et al., 2005), and in occupationally exposed workers, 15 days (Thuresson et al., 2006), agrees well with our estimate of 13 days. Although the calculated half-life in this study is plausible, it may actually be an overestimation, because there most likely was still release out of the implants during the apparent elimination phase (Fig. 2) since silastic implants are supposed to be responsible for a continuous release.
those in liver (muscle: 461 ng/g lw, 430 ng/g lw; liver: 269 ng/ g lw, 237 ng/g lw). These results suggest accumulation of BDE 209 in bird tissues. Recently, BDE 209 has been detected in a variety of species, mainly those living in terrestrial ecosystems (de Boer et al., 2004; Jaspers et al., 2006; Voorspoels et al., 2006). Compared to these studies, levels in the present study were higher range environmental concentrations of BDE 209. Muscle and liver concentrations were 7 and 4 times, respectively, higher than in predatory birds from Belgium (Jaspers et al., 2006). The higher concentrations in the muscle compared to the liver are probably due to a higher metabolic activity in liver than in muscle (Voet and Voet, 1995). Fig. 3 shows, in addition to BDE 209, the presence of other PBDE congeners in the muscle and liver. The PBDE profiles in muscle and liver were very similar (Fig. 3). However, the relative contribution of the nonaBDEs was greater in liver than in muscle (BDE 206: liver, 4.0%; muscle, 3.7%; BDE 207: liver, 24%; muscle, 18%; BDE 208: liver, 9.2%; muscle, 6.2%), which is probably due to the high metabolic activity of the liver (Voet and Voet, 1995). Differences between the control group and the exposed group were most pronounced for the nona- (BDE 206, BDE 207 and BDE 208) and octaBDEs (BDE 196 and BDE 197), suggesting bioformation from BDE 209 (Fig. 3). The octaBDEs, BDE 203 and BDE 205, did not differ much between both groups. A possible
(a) BDE concentrations in muscle (ng/g lipid weight)
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3.3.2. BDE 209 concentrations in muscle and liver Tissue concentrations were below the LOQ in the control group (muscle: <5.6 ng/g lw, liver: <2.9 ng/g lw). In the exposed group, the muscle concentrations were about twofold
47 99 100 153 154 183 196 197 203 205 206 207 208
PBDE congener Fig. 3. Profiles of PBDE congeners in (a) muscle and (b) liver (ng/g lw) after implantation with BDE 209.
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explanation for the differences in the formation of these octaBDEs is the position of the bromine atoms. Hexa- (BDE 153) and heptaBDEs (BDE 183) differed less between the control group and the exposed group (Fig. 3). For the other congeners, there were no notable differences between the groups (Fig. 3). Our results suggest that BDE 209 is degraded to lower brominated congeners in tissues of European starlings. Although it can not be excluded that a small fraction of the octa- and nonaBDEs found in the tissues stems from the bioaccumulation of eventual trace octa- and nonaBDEs in the oil, it is mostly plausible to assume that the presence of these congeners is the result of the debromination of BDE 209. The formation of lower brominated congeners from decaBDE has previously been reported in rats (Mo¨rck et al., 2003) and fish (Kierkegaard et al., 1999; Stapleton et al., 2004). Similar to those in fish (Kierkegaard et al., 1999; Stapleton et al., 2004), our results suggested that BDE 209 can be debrominated down to hexaBDEs. For the nonaBDEs, we observed the following order: BDE 207 > BDE208 > BDE 206. A similar pattern was found resulting from thermal decomposition of BDE 209 (Koryta´r et al., 2005), while other patterns were observed for photolytic degradation (Soderstro¨m et al., 2004) or biotransformation in fish (Mo¨rck et al., 2003). To our knowledge, these results are the first indications for the debromination of BDE 209 in birds. The present study, in accordance with previous studies (Kierkegaard et al., 1999; Mo¨rck et al., 2003; Stapleton et al., 2004), suggested that BDE 209 can be debrominated to congeners that are also present in the penta- and octaBDE commercial mixtures, which are no longer allowed for use due to their potential toxicity (Darnerud, 2003). Thus, phasing out the penta- and octaBDE commercial mixtures could be insufficient to restrict the potential risk of lower brominated congeners. These results are a great cause of concern considering the large amounts of decaBDE that are worldwide used (BSEF, 2003). More studies are needed to elucidate the toxicology and metabolism of higher brominated BDEs, especially BDE 209. The results of these studies should be taken into account for a readjusted risk assessment of the decaBDE commercial mixture. Acknowledgments This study was supported by the Fund for Scientific Research Flanders (FWO-project G.0137.04), an FWO Postdoctoral Fellowship to T.D., an FWO Research Assistantship Grant to V.J. and a Research Grant from the Research Funds of the University of Antwerp to E.V.d.S and A.C. References Andersson, P.L., Berg, A.H., Bjerselius, R., Norrgren, L., Olse´n, H., ¨ rn, S., Tysklind, M., 2001. Bioaccumulation of selected Olsson, P.E., O PCBs in zebrafish, three-spined stickleback, and Arctic char after three different routes of exposure. Archives of Environmental Contamination and Toxicology 40, 519e530. de Boer, J., Leslie, H.A., Leonards, P.E.G., Bersuder, P., Morris, S., Allchin, C.R., 2004. Screening and time trend study of decabromodiphenylether and hexabromocyclododecane in birds. Proceedings of the Third International Workshop on Brominated Flame Retardants, Toronto, Canada, 125e128.
BSEF, 2003. Major brominated flame retardants volume estimates: Total market demand by region in 2001. Bromine Science and Environmental Forum, Brussels. Available at:
E. Van den Steen et al. / Environmental Pollution 148 (2007) 648e653 Pinxten, R., De Ridder, E., Balthazart, J., Eens, M., 2002. Context-dependent effects of castration and testosterone treatment on song in male European starlings. Hormones and Behavior 42, 307e318. Sandholm, A., Emanuelsson, B.M., Wehler, E.K., 2003. Bioavailability and half-life of decabromodiphenyl ether (BDE-209) in rat. Xenobiotica 33, 1149e1158. ˚ , 2001. Brominated flame retardants Sjo¨din, A., Patterson Jr., D.G., Bergman, A in serum from U.S. blood donors. Environmental Science and Technology 35, 3830e3833. Soderstro¨m, G., Sellstro¨m, U., de Wit, C.A., Tysklind, M., 2004. Photolytic debromination of decabromodiphenyl ether (BDE 209). Environmental Science and Technology 38, 127e132. Stapleton, H.M., Alaee, M., Letcher, R.J., Baker, J.E., 2004. Debromination of the flame retardant decabromodiphenyl ether by juvenile carp (Cyprinus carpio) following dietary exposure. Environmental Science and Technology 38, 112e119. Thomas, G.O., Moss, S.E.W., Asplund, L., Hall, A.J., 2005. Absorption of decabromodiphenyl ether and other organohalogen chemicals by grey seals (Halichoerus grypus). Environmental Pollution 133, 581e586. Thomas, G.O., Wilkinson, M., Hodson, S., Jones, K.C., 2006. Organohalogen chemicals in human blood from the United Kingdom. Environmental Pollution 141, 30e41.
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Thuresson, K., Bergman, A., Jakobsson, K., 2005. Occupational exposure to commercial decabromodiphenyl ether in workers manufacturing or handling flame-retarded rubber. Environmental Science and Technology 39, 1980e1986. ˚ ., Thuresson, K., Ho¨glund, P., Hagmar, L., Sjo¨din, A., Bergman, A Jakobsson, K., 2006. Apparent half-lives of hepta- to decabrominated diphenyl ethers in human serum as determined in occupationally exposed workers. Environmental Health Perspectives 114, 176e181. Van den Steen, E., Covaci, A., Jaspers, V.L.B., Dauwe, T., Voorspoels, S., Eens, M., Pinxten, R., 2007. Experimental evaluation of the usefulness of feathers as a non-destructive biomonitor for polychlorinated biphenyls (PCBs) using silastic implants as a novel method of exposure. Environment International 33, 257e264. Voet, D., Voet, J.G., 1995. Biochemistry, second ed. Wiley, New York. Voorspoels, S., Covaci, A., Lepom, P., Jaspers, V.L.B., Schepens, P., 2006. Levels and distribution of polybrominated diphenyl ethers in various tissues of birds of prey. Environmental Pollution 144, 218e227. Voorspoels, S., Covaci, A., Schepens, P., 2003. Polybrominated diphenyl ethers in marine species from the Belgian North Sea and Western Scheldt Estuary: Levels, profiles, and distribution. Environmental Science and Technology 37, 4348e4357. de Wit, C.A., 2002. An overview of brominated flame retardants in the environment. Chemosphere 46, 583e624.