Bioaccumulation of Several Brominated Flame Retardants and

Dec 3, 2010 - The present study is primarily designed to examine the role played by dietary sources on polybrominated diphenyl ethers (PBDE) congener ...
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Environ. Sci. Technol. 2011, 45, 400–405

Bioaccumulation of Several Brominated Flame Retardants and Dechlorane Plus in Waterbirds from an E-Waste Recycling Region in South China: Associated with Trophic Level and Diet Sources X I U - L A N Z H A N G , † X I A O - J U N L U O , * ,† HONG-YING LIU,‡ LE-HUAN YU,† SHE-JUN CHEN,† AND BI-XIAN MAI† State Key Laboratory of Organic Geochemistry, Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, China, and College of Chemistry and Chemical Engineering, Hubei University, Wuhan 200433, China

Received July 4, 2010. Revised manuscript received September 29, 2010. Accepted November 1, 2010.

The present study is primarily designed to examine the role played by dietary sources on polybrominated diphenyl ethers (PBDE) congener profiles in waterbirds collected in an e-waste recycling region in South China. Some emerging halogenated flame retardants (HFRs), such as dechlorane plus (DP), 2,3,4,5,6pentabromoethyl benzene (PBEB), pentabromotoluene (PBT), and 1,2-bis(2,4,6-tribromophenoxy)ethane (BTBPE), were also quantified. Stable isotopes (δ15N and δ13C) were analyzed to assess the trophic levels and dietary sources of the birds. PBDEs were found to be the predominant HFRs, followed by DP, PBT, PBEB, and BTBPE. The birds in which BDE209 was predominant have differential δ13C and δ15N signatures compared with other birds, suggesting that dietary source is one of the important factors in determining the PBDE congener profile in birds. The levels of ΣPBDEs, PBEB, and PBT were significantly correlated with the trophic level (δ15N) for avian species which are located in a food chain, indicating the biomagnification potential of these compounds. No correlation was found between DP concentrations and trophic level of the birds. There is a significantly negative correlation between the fraction of anti-DP and δ15N, suggesting that the metabolic capability of DP in birds increases with the trophic level of the birds.

Introduction Polybrominated diphenyl ethers (PBDEs) are the second highest product of brominated flame retardants (BFRs) in the world market (1). Two commercial PBDE products, pentaand octa-BDEs, are listed as persistent organic pollutants (2) and have been banned in the European Union (EU), China, and several U.S. states due to their persistence, long-range transport, and potential toxicity to wildlife and humans (3-5). Deca-BDE had been also banned from being added into electronic products by the EU Court in 2008 (6). With increasing * Corresponding author phone: +86-20-85290146; fax: +86-2085290706; e-mail: [email protected]. † Guangzhou Institute of Geochemistry. ‡ Hubei University. 400

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regulation and phasing out of PBDEs from production, more environmental concerns have focused on non-PBDE halogenated FRs, such as dechlorane plus (DP), 1,2-bis(2,4,6-tribromophenoxy) ethane (BTBPE), 2,3,4,5,6-pentabromoethyl benzene (PBEB), and pentabromotoluene (PBT). These “new” FRs have been reported in abiotic and biota samples (7-14). To the authors’ knowledge, there have only been two reports on the trophic transfer of DP in aquatic food webs (13, 14) and very few on PBT, PBEB, and BTBPE. In addition, no report has ever been concerned with the tissue distribution of these newly nonPBDE flame retardants. Birds exhibit species-specific PBDE congener profiles. The PBDE congener profile in aquatic birds is usually dominated by BDE47 (15, 16), while in bird feeding entirely or partially on elements of terrestrial food webs, BDE153 or BDE209 is the most abundant congener in most cases (17-22). Even in the same species of birds, buzzards for example, the PBDE congener profile in one region is different from another region (20, 23). In our previous publication (21), three different PBDE congener profiles were found in five species of waterbirds habitat in an e-waste region. The reason for this confounding PBDE congener profile is not yet fully understood. Bioavailability, biological transformation, and selective uptake/ metabolism are suggested to be responsible for the different congener profiles. Dietary exposure is also one of the important factors in determining the PBDE congener profile (24). However, there are few studies providing direct evidence for this argument. Recently, Newsome et al. (25) reported that the different PBDE congener profile in the peregrine falcon between urban and nonurban California environments can be attributed to different dietary sources. Stable isotope analysis is an efficient tool for investigating the ecotoxicology of dietary exposure and biomagnification of contaminants in wildlife (26). The stable isotope ratio of nitrogen (15N/14N or δ15N) increases by 2-5‰ per trophic level (27), thus δ15N is used to ensure the position of the organism’s trophic level. The stable isotope ratio of carbon (13C/12C or δ13C) also increases with trophic level; however, the increment is so small (about 1‰) (28, 29) that δ13C is constantly employed to analyze the dietary composition and carbon sources of organisms (30-32). Stable isotope values, which are affected by dietary exposure, provide an independent measurement of trophic level and carbon source (33). In the present study, we primarily aim to determine the influence of food sources and trophic level on the levels and congener profiles of PBDEs in waterbirds investigated in our previous study (21) by analyzing the stable carbon (δ13C) and nitrogen isotope (δ15N) composition of the birds. NonPBDE FRs (DP, PBT, PBEB, and BTBPE) in the birds were also reported in the present study. The tissue distribution of PBDEs and those “new” non-PBDE HFRs in the muscle, kidney, and liver of the birds are examined. Finally, the potential biomagnification of PBDEs and non-PBDE HFRs in these birds is assessed by the relationships between the contaminant concentrations and trophic levels defined by δ15N.

Method and Materials Sampling. Specimens (n ) 29) from five bird species, including Ardeidae (Chinese-pond heron, Ardeola bacchus; n ) 5), Rallidae (white-breasted waterhen, Amaurornis phoenicurus, n ) 11; slaty-breasted rail, Gallirallus striatus, n ) 5; ruddy-breasted crake, Porzana fusca, n ) 5), and Scolopacidae familes (common snipe, Gallinago gallinago, n ) 3), were collected between 2005 and 2007 from Qingyuan 10.1021/es102251s

 2011 American Chemical Society

Published on Web 12/03/2010

FIGURE 1. Relationship between δ15N and δ15C of wild birds from South China: a) white-breasted waterhen as a whole group and SBR was removed; b) white-breasted waterhen separated into two groups and both SBR and WBW2 did not include. Error bars represent (1 standard error. CPH: Chinese pond heron; WBW1: white-breasted waterhen (7 of 11); WBW2: white-breasted waterhen (4 of 11); CS: common snipe; SBR: slaty-breasted rail; RBC: ruddy-breasted crake. County, the second largest e-waste recycling region in the Pearl River Delta. Detailed information about the sampling site and collected bird species is provided in our previous publication (21). Pectoral muscle, liver, and kidney were excised from the aforementioned birds, which were found dead or dying from various causes (hunting, poisoning, and distress, etc). All tissues were stored at -20 °C until analysis. Analysis. The same extracts for muscle as those previously prepared for the determination of PBDEs, PCBs, and OCPs (21) are used in this work. The analytical procedure for the liver and kidney is similar to that for muscle. Detailed procedures for sample pretreatment, extraction, and cleanup are given in the Supporting Information. BTBPE, PBEB, PBT, and DP analyses were performed with a Shimadzu model 2010 gas chromatograph coupled with a model QP2010 mass spectrometer (Shimadzu, Japan) using negative chemical ionization (NCI) in the selective ion-monitoring (SIM) mode. A DB-XLB (30 m × 0.25 mm i.d., 0.25 µm film thickness) capillary column was used for the determination of PBEB and PBT, and a CP-Sil 13 CB (12.5 m × 0.25 mm i.d., 0.2 µm film thickness) capillary column was used for DP and BTBPE. Details of the GC temperature program and monitored ions are given elsewhere (34). Stable Isotopes. The subsamples of pectoral muscle for nitrogen and carbon stable isotope analysis were lyophilized and ground into ultrafine powders. Approximately 1 mg of ground samples was weighed in tin capsules and analyzed by a flash EA 112 series elemental analyzer interfaced with a Finigan MAF ConFlo 111 isotope ration mass spectrometer. Stable isotope ratios of samples were assessed against the reference standards ammonium sulfate for δ15N and carbonblack for δ13C. An isotope ratio was expressed as δX (values [‰]), with δX ) [(Rsample/Rstandard - 1) * 1000], where X is 15N or 13C and R is the corresponding ratio of 15N/14N or 13C/12C. The precision for this technique is about 0.5‰ (2 SD) for δ15N and 0.2‰ (2 SD) for δ13C. Quality Assurance and Control. A previous study conducted by Gauthier et al. (10) demonstrated that BTBPE, PBEB, PBT, and DP were extracted in a manner similar to that for PBDEs and that the recoveries of the internal standards analyzed for PBDEs were also suitable for these “new” FRs analysis. The recoveries of 2,2,4,4,5-pentachlorodiphenyl ether (CDE 99) and 13C-BDE 209, used as surrogate standards for PBDE analysis in our previous study (21), were 87%-113% and 86%-105%, respectively, and the recoveries of the spiking blanks for PBDEs ranged from 77% to 93%.

Five spiking blanks with PBEB, PBT, BTBPE, and DP and five procedural blanks were performed. The average recoveries for PBEB and PBT were 95% and 94%, respectively; the average recoveries for BTBPE, anti-DP, and syn-DP were 103%, 109%, and 102%, respectively; and the relative standard deviations (RSD) of all targets were less than 5%. The RSD of duplicates (n ) 3) were less than 15% for all contaminants. The limit of quantification (LOQ) defined as mean of the blanks plus 3 times standard deviation for compounds detected in blanks. A signal-to-noise ratio of 3 was used as LOQ for compounds which were not detected in blank. PBEB, PBT, BTBPE, anti-DP, and syn-DP were 0.12, 0.17, 0.6, 0.06, and 0.1 ng/g, respectively. The linear range of the instrument was 20 to 500 ppb (r2 > 0.99) for BTBPE and DP and 0.2 to 200 ppb for PBEB and PBT (r2 > 0.99). Data Analysis. For samples with contaminant concentrations below LOQ, zero was used for the calculations. Concentrations were expressed on a lipid weight basis (LW). To depict the tissue distribution of targets in birds, the concentration ratios of targets in muscle-versus-liver and kidney-versus-liver were calculated and compared with one by t test. A linear regression analysis between trophic levels and concentrations of target compounds were performed. Data analysis was performed using SPSS for Windows Release 11.5 (SPSS Inc.). Statistical significance was set at p < 0.05 throughout the manuscript.

Results and Discussion Stable Carbon and Nitrogen Isotopic. The stable carbon and nitrogen isotope composition of collected birds is shown in Figure 1. The relative trophic status of collected birds, defined by δ15N, increases in the following order: slatybreasted rail (7.0‰) < ruddy-breasted crake (8.4 ‰) < common snipe (9.2‰) and white-breasted waterhen (9.2‰) < Chinese-pond heron (10.7‰). This trophic status is consistent with the different feeding habits of bird species. The Chinese pond heron is a piscivorous bird feeding primarily on fish. The white-breasted waterhen eats mainly seeds, insects, and small fish, and they also often forage above ground in low bushes and small trees. The common snipe feeds mainly on aquatic insect and invertebrates in wetland areas. The ruddy-breasted crake feeds mostly on tender shoots, berries, and some aquatic insects (35). In a given food web, δ15N and δ13C are often correlated because foods enriched in δ15N are often also enriched in δ13C (26). In the present study, a positive correlation between δ15N and δ13C was found except for the slaty-breasted rail, which has a VOL. 45, NO. 2, 2011 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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10.5 (10.2-11.5)

7.3 (6.1-8.0)

8.2 (7.7-10.1)

9.4 (8.2-10.0)

-22.2 -(23.8-19.8)

-22.8 -(24.0-21.5)

-24.6 -(25.4-23.4)

-24.5 -(24.9-22.1) CS n ) 3

RBC n ) 5

SBR n ) 5

CPH n ) 5

WBW2 n ) 4

WBW1 n ) 7

a Σ PBDEs: sum of PBDE 28, 47, 99, 100, 153, 154, 183, 196, 203, 208, 207, 206, and 209; DP: sum of anti- and syn-DP. nd: contaminants concentration below LOQ; CPH: Chinese-pond heron; CS: common snipe; WBW1: white-breasted waterhen (7 of 11); WBW2: white-breasted waterhen (4 of 11); SBR: slaty-breasted rail; RBC: ruddy-breasted crake.

8.8 (6.3-9.4) -21.7 -(24.8-20.4)

PBEB (ng/g lw)

6 (3-11) 7.5 (4.5-21) 15 (4-35) 4.3 (2.6-6.8) 12 (5-31) 15 (6-19) 33 (10-110) 54 (3.3-190) 46 (1.7-120) 4.5 (1.5-5.1) 7.8 (4.3-21) 7.6 (5.2-11) 1.6 (0.7-2.2) 3.4 (1.3-6.5) 3.1 (2.2-20) 6.5 (1.7-8.5) 16 (12-22) 21 (16-28) 2 (1-7) 2 (0-21) 2 (nd-27) 1.2 (0.4-1.8) 3 (0-7) 4 (5-8) 13 (5.3-37) 15 (1.6-49) 8.6 (0.5-30) 1.0 (nd-1.6) 1.7 (nd-6.7) 2.1 (nd-5.1) 0.1 (nd-1.5) nd (nd-5.0) 2.1 (nd-6.2) 4.5 (0.3-8.0) 10 (nd-12) 16 (3.7-22)

PBT (ng/g lw) BTBPE (ng/g lw)

nd (nd-1) nd (nd-4) nd (nd-3) nd nd nd nd nd nd 1.9 (nd-20) 10 (nd-27) 11 (nd-40) 3.3 (nd-9.3) 5.3 (0.6-40) 3.4 (2.1-31) nd (nd-0.9) nd nd (nd-4) 66 (9-360) 120 (22-1100) 89 (24-886) 43 (34-94) 98 (48-260) 66 (41-72) 69 (nd-130) 190 (nd-790) 65 (nd-300) 170 (14-610) 560 (21-830) 600 (55-920) 56 (21-150) 45 (nd-180) 29 (21-120) 7.4 (2.3-600) 170 (56-2200) 220 (100-1800)

DP (ng/g lw) ΣPBDEs (ng/g lw)

500 (160-14000) 560 (210-16000) 710 (70-18000) 700 (150-1200) 1300 (530- 2300) 1400 (1000-1700) 2200 (530-2500) 6400 (89-9800) 3600 (280-3600) 820 (130-1400) 2500 (280-6400) 400 (180-1300) 37 (23-130) 30 (7-110) 94 (6-1500) 340 (270-1700) 2000 (1100-2600) 2200 (1100-2900) 9.6 (8.6-11.3)

1.5 (1.0-4.5) 2.7 (1.9-3.2) 3.0 (2.6-4.6) 1.8 (1.0-4.5) 2.8 (1.7-3.9) 3.7 (2.7-5.0) 2.1 (1.7-3.8) 2.2 (1.8-4.9) 3.4 (1.8-3.8) 1.6 (1.4-4.3) 2.7 (2.3-4.4) 3.4 (2.8-5.6) 3.3 (1.0-4.8) 2.8 (2.1-5.1) 4.7 (2.4-5.8) 2.3 (1.9-2.3) 1.8 (1.6-1.9) 2.3 (2.0-2.5) muscle kidney liver muscle kidney liver muscle kidney liver muscle kidney liver muscle kidney liver muscle kidney liver

-22.3 -(24.7-21.0)

δ15N (‰) δ13C (‰) lipid (%) tissue species

TABLE 1. Median and Range of Concentrations of Halogenated Flame Retardants in Muscle, Kidney, and Liver of Wildlife Birds from South Chinaa

relatively high δ13C but low δ15N, diverging the regression line. This indicates that this bird species has a primary food source different from that of other bird species. Levels and Profiles of Contaminants. The levels and congener profiles of PBDEs in the muscle of collected birds have been described in detail in our previously published paper (21). Table 1 presents the reported PBDE concentrations in the muscle samples of these birds, accompanying the data in liver and kidney measured in this study (Table 1). PBDE profile differences between tissues were investigated using ANOVA, which showed no differences (Figure S1). The PBDE congener profiles in the birds in the present study could be classified into three groups according to the predominant congeners (Figure S1). The Chinese pond heron and ruddy-breasted crake are in one group, in which BDE 47 is the predominant congener, similar to the previously acquired results for aquatic birds (15, 16). The common snipe and seven of the eleven white-breasted waterhens (denoted as WBW1) are in another group, in which BDE 153 is a major contribution congener to the total PBDEs, followed by BDE 183, 99, 154, and 100. The slaty-breasted rail and four other white-breasted waterhens (denoted as WBW2) are in the third group, where BDE 209 is dominant. Interestingly, the white-breasted waterhens have two completely different PBDE profiles. Four of the whitebreasted waterhens (WBW2) show similar PBDE profiles to the slaty-breasted rail, which has a completely different diet exposure from other species as revealed by the stable carbon and nitrogen isotope composition data (Figure 1a). We hypothesize that the diet source of the four white-breasted waterhens (WBW2) in which BDE 209 dominated is different from that of the other seven white-breasted waterhens (WBW1). Thus, we reanalyzed the relationship between δ15N and δ13C after dividing the white-breasted into two groups. Figure 1b shows that the composition of δ15N and δ13C in the four white-breasted waterhens (WBW2 in the Figure 1b), as in the slaty-breasted rail, obviously diverges from the regression line. The correlation coefficient increased from 0.81 to 0.96 and p value decreased from 0.19 to 0.039 after removing WBW2. Compared with WBW1, the δ13C in WBW2 increased while δ15N decreased although no significant difference exists. This composition of δ13C and δ15N is similar with that of slaty-breasted rail which also show relatively high δ13C and low δ15N. Birds in the present study were collected from two adjacent towns in which e-waste recycling activities conducted. It is very likely that some birds directly consume food items containing BDE 209 at e-waste dumping sites. These BDE209 containing materials might have different δ13C and δ15N (e.g., high δ13C but low δ15N) from other food, and they may make a minor contribution to total consuming food by bird. This is a plausible explanation for that the PBDE congener profile in WBW1 is completely different from that of WBW2, but no significant differences were found between WBW1 and WBW2 in δ13C and δ15N. Newsome et al. (25) reported that the different PBDE congeners in urban and nonurban peregrine falcons can be attributed to their different dietary compositions. Our results are consistent with their findings, confirming that diet plays a key role in determining PBDE congener profile. Both syn- and anti-DP were detectable in all birds, except for one Chinese pond heron. The DP concentrations ranged from nd-610, nd-2200, and nd-1830 ng/g lw in muscle, kidney, and liver, respectively. The highest DP level was observed in slaty-breasted rail (muscle 14-610; liver 55-920; kidney 21-830 ng/g lw). Generally, the DP concentrations in the studied birds were 1-2 orders of magnitude lower than their corresponding concentrations of ΣPBDEs, except for the ruddy-breasted crake, whose DP concentrations were comparable to those of ΣPBDEs (Table 1). DPs have been widely detected in air, dust, sediment, fish, bird, and human serum

TABLE 2. Correlation Coefficient Among Different Contaminant in Sampled Birds syn-DP BTBPE syn-DP anti-DP PBT PBEB a

a

0.92

anti-DP a

0.89 0.55a

PBT

PBEB

ΣPBDEs

0.50 0.18 -0.01

0.16 -0.02 -0.11 0.92a

0.56a 0.77a 0.07 0.19 0.11

p < 0.01.

(10, 12, 14, 36, 37). DP was detected with the median of 2.4 ng/g ww in herring gull eggs from colonies in the Laurentian Great Lakes (10). However, there has no report concerning the concentrations of DP in muscle, liver, or kidney. Thus a comparison with the same tissue is impossible in the present study. The fraction of anti-DP (fanti), defined as the concentration of anti-DP divided by the sum of concentrations of syn- and anti-DP, was 0.70 for DP in commercial products bought from the chemical market. No difference in the fanti between tissues was found in a given bird species (ANOVA). Thus, an overall fanti, not accounting for the tissues, was calculated for the five bird species. The mean fanti for the Chinese pond heron, white-breasted waterhen, common snipe, ruddybreasted crake, and slaty-breasted rail were 0.34, 0.36, 0.43, 0.46, and 0.61, respectively. These values are all lower than 0.70, indicating a preferential accumulation of syn-DP in biota samples, in line with the reports for DPs in fish samples (14). Of the three non-PBDE BFRs, PBEB was detected in all samples; PBT was detected in 79% of samples; and BTBPE was detected only in the slaty-breasted rail, ruddy-breasted crake, one white-breasted waterhen, and one common snipe with a detection frequency of 37%. Up to now, litter information was obtained on these non-PBDE BFRs in bird tissues. Gauthier et al. (10) reported that the concentrations of PBEB, PBT, and BTBPE in eggs of herring gulls from the Laurentian Great Lakes were in the range from 0.03 to 1.4, from 0.04 to 0.02, and from 0.04 to 0.26 ng/g ww, respectively. PBEB was used as a FR in the 1970s and 1980s, while PBT has been, and currently is, used as a flame retardant in textiles, polyester resins, and paint emulsions (38, 39). Information on the current production of PBT and PBEB is not publicly available, but recent studies showed that they are distributed in the environment (7, 8) and are bioavailable in organisms (10). Although BTBPE has been widely detected in air, sediment, fish, and birds (7, 40, 41), the low detection frequencies indicate that the environment of the study area might be less subject to this compound contamination. Simple linear correlation analysis revealed that DP significantly correlated with BTBPE and PBDEs, and PBT significantly correlated with PBEB (Table 2). This suggested that DP, BTBPE, and PBDEs may have the same source and/ or environmental behavior but different from those of PBT and PBEB. The syn-DP was found to significantly correlate with PBDEs, but this was not the case for anti-DP. This observation is consistent with that in fish collected from a contaminated pond in the study area, confirming that antiDP has a different behavior from syn-DP in biota (14). Tissue Distribution. The ratios of muscle concentration to liver concentration (M/L) and kidney concentration to liver concentration (K/L) were used to analyze the tissues difference of contaminants in bird species (Figure S2). Regardless of whether the concentration was expressed on a wet weight base or a lipid weight base, no significant difference (t test, p > 0.05) in the concentration of contaminants between kidney and liver was found in all species, except for common snipe, although the K/L ratio varied greatly between less than 1 to larger than 1. In common

snipe, the level of PBT was obviously higher in the liver than in the kidney. Based on wet weight, the M/L ratio was less than 1 for all contaminants in all bird species except for PBDEs which have M/L ration larger than 1 in three bird species: Chinese pond heron, slaty-breasted rail, and ruddy-breasted crake (Figure S3). The lipid contents in liver were higher than those in muscle (Table 1) which can partially explain the high concentration in liver. When the concentrations were expressed on lipid weight, the ratios of M/L for PBDEs were larger than 1 in four of the five bird species. Whereas the M/L ratios for PBT and PBEB were less than 1 for all bird species and so were the M/L ratios for syn-DP and anti-DP, except the ruddy-breasted crake. This observation suggested that PBDEs have high accumulation ability in muscle compared with other BFR. This result is consistent with laboratory exposure (42) and field (13, 43) studies, which both reported that the PBDE concentrations in muscle were higher than those in liver for birds . Correlation with Trophic Level. The relationship between trophic level of the birds and monitored BFRs and DP was examined by regressing the log-normalized concentration of the individual compounds in muscle against δ15N, with the exception of BTBPE due to its low detection frequencies. The layouts of the biplot of the δ15N and Ln-normalized concentration of the target compounds were similar to those of the biplots of the δ15N and the δ13C (Figure 2, Figure 1b, and Figure S4). The slaty-breasted rail and WBW2 diverges the regression line of δ15N and δ13C. They also show a different PBDE congener pattern from other birds. These results suggested that the sources of pollutants in these birds are different from other bird. Therefore, slaty-breasted rail and WBW2 were excluded from the analysis. A significant positive correlation between the δ15N and concentrations of ΣPBDEs, PBEB, and PBT (p < 0.05 in three cases) was found. Trophic magnification of ΣPBDEs and BDE congeners had been reported in both freshwater and marine food webs (45, 46). The biomagnification of PBDEs and BDE congeners were also observed in a small terrestrial food web composed of birds with different trophic levels (46). In the present study, the concentrations of ΣPBDEs, BDE47, BDE99, and BDE100 were found to have a significant positive linear correlation with trophic level (Figure S4). Highly brominated congeners such as BDE153, BDE154, BDE183, and BDE196 also showed an increasing trend with increasing trophic level, but the regression was not significant (p-value larger than 0.05) (Figure S4). This was due to the decrease in concentration of highly brominated congeners in the Chinese pond heron. These results indicate that most BDE congeners biomagnify with increased trophic level. To the best of our knowledge, there is no information available on the biomagnification of PBT and PBEB through the food web. The significant positive correlations between the concentrations of these two compounds and the δ15N implied that both PBT and PBEB could be biomagnified through the food web as PBDEs. Therefore, further studies on the ecotoxicology of PBT and PBEB for wildlife or humans are needed. Both syn-DP and anti-DP did not correlated well with the δ15N (p > 0.05, Figure 2), indicating that no biomagnifications occurred. Only two studies have thus far reported on the trophic transfer of DP in food webs. Biomagnification for anti-DP, but biodilution for syn-DP, was observed in the aquatic food web from Lake Winnipeg. In the food web from Lake Ontario, no statistically significant correlation between the concentrations of both isomers and trophic level was found, which is consistent with the present study (13). However, both anti-DP and syn-DP have been found to be biomagnified in the food web from a highly contaminated pool in the presently studied region after the removal of the VOL. 45, NO. 2, 2011 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 2. Relationship between δ15N and a) Ln ΣPBDEs, b) Ln PBEB, c) Ln PBT, and d) Ln DP in muscle of wild birds from South China. Error bars represent (1 standard error. CPH: Chinese pond heron; WBW1: white-breasted waterhen (7 of 11); WBW2: white-breasted waterhen (4 of 11); CS: common snipe; SBR: slaty-breasted rail; RBC: ruddy-breasted crake. of metabolizing DP increases with trophic level. A same trend for fanti was also observed for the avian species in the present study (Figure 3). The high metabolism of DP in high trophiclevel birds may mask the biomagnifications of DP. However, it should be kept in minds that no direct evidence to support the hypothesis that the fanti can reflect the metabolism of DP in biota. Selective uptake and excretion can also result in the change of fanti. Thus, more research on the biodegradation of DP is urgently needed to garner an understanding of the bioaccumulation and trophic transfer of DP.

Acknowledgments

FIGURE 3. Relationship of the anti isomer fraction (fanti) with corresponding δ15N. Error bars represent (1 standard error. CPH: Chinese-pond heron; CS: common snipe; WBW1: white-breasted waterhen (7 of 11); WBW2: white-breasted waterhen (4 of 11); SBR: slaty-breasted rail; RBC: ruddybreasted crake. highest trophic-level fish species (14). The confounding data regarding the trophic transfer of DP can be attributed to many factors, such as the composition of the food web, the contaminated levels of DPs, and the metabolism of DP in the biota. Anti-DP is more susceptible to biological attack than syn-DP due to the fact that the four interior carbons of antiDP on the cyclooctane are less blocked by chlorines than those of syn-DP (12). In this way, it has been suggested that the fanti could be related to the biota’s metabolic capacity for DP (12, 14). The decrease in fanti up the trophic ladder was found in aquatic organisms (14), indicating that the ability 404

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This work was funded by the National Science Foundation of China (Nos. 20890112, 40873074, 40773061), the National Basic Research Program of China (No. 2009CB4216604), and the Earmarked Fund of the State Key Laboratory of Organic Geochemistry (SKLOG2009A04). This is contribution No-1261 from GIG.CAS.

Supporting Information Available Detailed information for procedures of the samples pretreatment and cleanup for samples, PBDE congener profile in different birds species, tissue distribution of target compounds in different birds species, correlations between individual BDE congeners and trophic levels, chromatogram of “novel” BFR compounds and entire data set. This material is available free of charge via the Internet at http:// pubs.acs.org.

Literature Cited (1) Alaee, M.; Arias, P.; Sjodin, A.; Bergman, A. An overview of commercially used brominated flame retardants, their applications, their use patterns in different countries/regions and possible modes of release. Environ. Int. 2003, 29, 683–689.

(2) The 9 new POPs under the Stockholm Convention. http:// chm.pops.int/Programmes/NewPOPs/The9newPOPs/tabid/ 672/language/en-US/Default.aspx (accessed June 4th 2010). (3) Birnbaum, L. S.; Staskal, D. F. Brominated flame retardants: cause for concern. Environ. Health Perspect. 2004, 112, 9–17. (4) Renner, R. In US, flame retardants will be voluntarily phased out. Environ. Sci. Technol. 2004, 38, 14A. (5) Zhou, Z. M. Implement of administrative measure on the control of pollution caused by electronic information products and the exemption of deca-BDE mixture. Flame Retard. Mater. Technol. 2006, 4, 15–16, in Chinese. (6) CPA (Clean Production Action), 2008. Deca-BDE controversy in Europe. (accessed June 4, 2010). (7) Hoh, E.; Zhu, L.; Hites, R. A. Novel flame retardants, 1,2-bis(2,4,6-triphenoxy)ethane and 2,3,4,5,6-pentabromethylbenzene, in United States’ environmental samples. Environ. Sci. Technol. 2005, 39, 2472–2477. (8) Hyoe¨tylae¨inen, T.; Hartonen, K. Determination of brominated flame retardants in environmental samples. Trends Anal. Chem. 2002, 21, 13–29. (9) Kierkegaard, A.; Bjorklund, J.; Friden, U. Identification of the flameretardant decabromodiphenyl ethane in the environment. Environ. Sci. Technol. 2004, 38, 3247–3253. (10) Gauthier, L. T.; Hebert, C. E.; Weseloh, D. V. C.; Letcher, R. J. Current-use flame retardants in the eggs of herring gulls (Larus argentatus) from the Laurentian Great lakes. Environ. Sci. Technol. 2007, 41, 4561–4587. (11) Betts, K. New data on a widely used flame retardant. Environ. Sci. Technol. 2008, 42, 5–6. (12) Hoh, E.; Zhu, L. Y.; Hites, R. Dechlorane Plus, a chlorinated flame retardant, in the Great Lakes. Environ. Sci. Technol. 2006, 40, 1184–1189. (13) Tomy, G. T.; Pleskach, K.; Ismail, N.; Whittle, D. M.; Helm, P. A.; Sverko, E.; Zaruk, D.; Marvin, C. H. Isomers of dechlorane plus in Lake Winnipeg and Lake Ontario food webs. Environ. Sci. Technol. 2007, 41, 2249–2254. (14) Wu, J. P.; Zhang, Y.; Luo, X. J.; Wang, J.; Chen, S. J.; Guan, Y. T.; Mai, B. X. Isomer-specific bioaccumulation and trophic transfer of dechlorane plus in the freshwater food web from a highly contaminated site, South China. Environ. Sci. Technol. 2010, 44, 606–611. (15) Law, R. J.; Alaee, M.; Allchin, C. R.; Boon, J. P.; Lebeuf, M.; Lepom, P.; Stern, G. A. Levels and trends of polybrominated diphenyl ethers and other brominated flame retardants in wildlife. Environ. Int. 2003, 29, 757–770. (16) Elliott, J. E.; Wilson, L. K.; Wakeford, B. Polybrominated diphenyl ether trends in eggs of marine and freshwater birds from British Columbia, Canada, 1979-2002. Environ. Sci. Technol. 2005, 39, 5584–5591. (17) Voorspoels, S.; Covaci, A.; Lepom, P.; Jaspers, V. L. B.; Schepens, P. Levels and distribution of polybrominated diphenyl ethers in various tissues of birds of prey. Environ. Pollut. 2006, 144, 218–227. (18) Lindberg, P.; Sellstro¨m, U.; Ha¨ggberg, L.; de Wit, C. A. Higher brominated diphenyl ethers and hexabromocyclododecane found in eggs of peregrine falcons (Falco peregrinus) breeding in Sweden. Environ. Sci. Technol. 2004, 38, 93–96. (19) Gao, F.; Luo, X. J.; Yang, Z. F.; Wang, X. M.; Mai, B. X. Brominated Flame Retardants, Polychlorinated Biphenyls, and Organochlorine Pesticides in Bird Eggs from the Yellow River Delta, North China. Environ. Sci. Technol. 2009, 43, 6956–6962. (20) Chen, D.; Mai, B.; Song, J.; Sun, Q.; Luo, Y.; Luo, X.; Zeng, E. Y.; Hale, R. C. Polybrominated dephenyl ethers in birds of prey from northern China. Environ. Sci. Technol. 2007, 41, 1828– 1833. (21) Luo, X. J.; Zhang, X. L.; Liu, J.; Wu, J. P.; Luo, Y.; Chen, S. J.; Mai, B. X.; Yang, Z. Y. Persistent halogenated compounds in waterbirds from an e-waste recycling region in South China. Environ. Sci. Technol. 2009, 43, 306–311. (22) Chen D., Hale R. C. A global review of polybrominated diphenyl ether flame retardant contamination in birds. Environ. Int. 2010, 36, 800-811. (23) Voorspoels, S.; Covaci, A.; Lepom, P.; Jaspers, V. L. B.; Schepens, P. Levels and distribution of polybrominated diphenyl ethers in various tissues of birds of prey. Environ. Pollut. 2006, 144, 218–227. (24) Muir, D. C. G.; Backus, S.; Derocher, A. E.; Dietz, R.; Evans, T. J.; Gabrielsen, G. W.; Nagy, J.; Norstrom, R. J.; Sonne, C.; Stirling, I.; Taylor, M. K.; Letcher, R. J. Brominated flame retardants in polar bears (Ursus maritimus) from Alaska, the Canadian Arctic,

(25)

(26) (27)

(28) (29) (30) (31) (32) (33) (34)

(35)

(36) (37)

(38) (39) (40) (41)

(42)

(43)

(44)

(45)

(46)

East Greenland, and Svalbard. Environ. Sci. Technol. 2006, 40, 449–455. Newsome, S. D.; Park, J.-S.; Henry, B. W.; Holden, A.; Fogel, M. L.; Linthicum, J.; Chu, V.; Hooper, K. Polybrominated diphenyl ether (PBDE) Levels in peregrine falcon (Falco peregrinus) eggs from California Correlate with diet and human population density. Environ. Sci. Technol. 2010, 44, 5248–5255. Jardine, T. D.; Kidd, K. A.; Fisk, A. T. Applications, considerations, and sources of uncertainty when using stable isotope analysis in ecotoxicology. Environ. Sci. Technol. 2006, 40, 7501–7511. Elliott, K. H.; Cesh, L. S.; Dooley, J. A.; Letcher, R. J.; Elliott, J. E. PCBs and DDE, but not PBDEs, increase with trophic level and marine input in nestling bald eagles. Sci. Total Environ. 2009, 407, 3867–3875. Vander, Z. M. J.; Rasmussen, J. B. Variation in delta N-15 and delta C-13 trophic fractionation: Implications for aquatic food web studies. Limnol. Oceanogr. 2001, 46, 2061–2066. McCutchan, J. H.; Lewis, W. M.; Kendall, C. Variation in trophic shift for stable isotope ratios of carbon, nitrogen, and sulfur. Oikos 2003, 102, 378–390. France, R. L. Carbon-13 enrichment in benthic compared to planktonic algae: food web implications. Mar. Ecol.: Prog. Ser. 1995, 124, 307–312. Rounick, J. S.; Winterbourn, M. J. Stable carbon isotopes and carbon flow in ecosystems. BioScience 1986, 36, 171–177. O’Leary, M. H. Carbon isotopes in photosynthesis. BioScience 1988, 38, 328–336. Deniro, M. J.; Epstein, S. Influence of diet on the distribution of carbon isotopes in animals. Geochim. Cosmochim. Acta 1978, 42, 495–506. Zhang, Y.; Luo, X. J.; Wu, J. P.; Liu, J.; Wang, J.; Chen, S. J.; Mai, B. X. 2009. Persistent Halogenated Compounds in Aquatic Biota Collected from an e-Waste Recycling Region: Occurrence and Biomagnifications. Environ. Toxicol. Chem. 2010, 29, 852–859. Chang, H.; Deng, J. X.; Zhang, G. P.; Chen, W. C.; Bi, X. F.; Lai, D. X.; Lin, S. Birds of Guangdong Nanling national nature reserve. In Studies on Biodiversity of the Guangdong Nanling National Nature Reserve; Pang, X. F., Chen, J. Q., Eds.; Guangdong Science & Technology Press: Guangzhou, 2003; pp 418-444. Zhu, J.; Feng, Y. L.; Shoeib, M. Detection of dechlorane plus in residential indoor dust in the city of Ottawa, 293 Canada. Environ. Sci. Technol. 2007, 41, 7694–7698. Ren, G. F.; Yu, Z. Q.; Ma, S. T.; Li, H. R.; Peng, P. G.; Sheng, G. Y.; Fu, J. M. Determination of Dechlorane Plus in Serum from Electronics Dismantling Workers in South China. Environ. Sci. Technol. 2009, 43, 9453–9457. Environmental Protection Agency. Fed. Regist. 1988, 53, 472428. Mattsson, P. E.; Norstro¨m, Å.; Rappe, C. Identification of the flame retardant pentabromotoluene in sewage sludge. Chromatography A 1975, 111, 209–213. Qiu, X.; Marvin, C. H.; Hites, R. A. Dechlorane plus and other flame retardants in a sediment core from Lake Ontario. Environ. Sci. Technol. 2007, 41, 6014–6019. Gauthier, L. T.; Letcher, R. J. Isomers of Dechlorane Plus flame retardant in the eggs of herring gulls (Larus argentatus) from the Laurentian Great Lakes of North America: temporal changes and spatial distribution. Chemosphere 2009, 75, 115–120. Van den Steen, E.; Covaci, A.; Jaspers, V. L. B.; Dauwe, T.; Voorspoels, S.; Eens, M.; Pinxten, R. Accumulation, tissuespecific distribution and debromination of decabromodiphenyl ether (BDE 209) in European starlings (Sturnus vulgaris). Environ. Pollut. 2007, 148, 648–653. Luo, X. J.; Liu, J.; Luo, Y.; Zhang, X. L.; Wu, J. P.; Lin, Z.; Chen, S. J.; Mai, B. X.; Yang, Z. Y. Polybrominated diphenyl ethers (PBDEs) in free-range domestic fowl from an e-waste recycling site in South China: Levels, profile and human dietary exposure. Environ. Int. 2009, 35, 253–258. Wan, Y.; Hu, J. Y.; Zhang, K.; An, L. H. Trophodynamics of Polybrominated Diphenyl Ethers in the Marine Food Web of Bohai Bay, North China. Environ. Sci. Technol. 2008, 42, 1078– 1083. Yu, M.; Luo, X. J.; Wu, J. P.; Chen, S. J. Bioaccumulation and trophic transfer of polybrominated diphenyl ethers (PBDEs) in biota from the Pearl River Estuary, South China. Environ. Int. 2009, 35, 1090–1095. Voorspoels, S.; Covaci, A.; Jaspers, V. L. B.; Neels, H.; Schepens, P. Biomagnification of PBDEs in three small terrestrial food chains. Environ. Sci. Technol. 2007, 41, 411–416.

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