Polybrominated Diphenyl Ethers (PBDEs) - Springer Link

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Mar 5, 2010 - Polybrominated Diphenyl Ethers (PBDEs), Polychlorinated. Biphenyles (PCBs), Hydroxylated and Methoxylated-PBDEs,.

Arch Environ Contam Toxicol (2010) 59:492–501 DOI 10.1007/s00244-010-9487-4

Polybrominated Diphenyl Ethers (PBDEs), Polychlorinated Biphenyles (PCBs), Hydroxylated and Methoxylated-PBDEs, and Methylsulfonyl-PCBs in Bird Serum from South China Juan Liu • Xiao-Jun Luo • Le-Huan Yu • Ming-Jing He • She-Jun Chen • Bi-Xian Mai

Received: 30 October 2009 / Accepted: 7 February 2010 / Published online: 5 March 2010 Ó Springer Science+Business Media, LLC 2010

Abstract Polybrominated diphenyl ethers (PBDEs), polychlorinated biphenyls (PCBs), and their derivatives, hydroxylated (OH) and methoxylated (MeO) PBDEs and methylsulfonylated (MeSO2) PCBs, were measured in sera of eight bird species collected from an e-waste recycling P region in South China. Concentrations of PCBs, ranging from 38 to 1700 ng/g lipid weight (lw), were one to two orders of magnitude higher than concentrations of P PBDEs (0.64–580 ng/g lw). The significantly positive relationship between PCB and PBDE concentrations suggested a similar pathway of exposure to these compounds. Compared with muscle in birds, serum might prefer to accumulate and/or retain less brominated/chlorinated congeners. 3-OH-BDE47 and 20 -OH-BDE68 were detected in more than 80% of the collected bird serum samples (range: not detectable (nd) to 13 and nd to 7.8 ng/g lw, respectively). The other three OH-PBDE congeners (40 OH-BDE-17, 6-OH-BDE47, and 40 -OH-BDE-49) and two MeO-PBDE congeners (3-MeO-BDE47 and 6-MeOBDE47) were occasionally detected in bird sera at concentrations ranging from nd to 2.5 ng/g lw. Both natural sources and metabolic transformation of PBDEs could contribute to the presence of these PBDE derivatives in the birds. The two MeSO2-PCB congeners (4-MeSO2-CB49 and 4-MeSO2-CB101) under investigation were detected at

J. Liu  X.-J. Luo (&)  L.-H. Yu  M.-J. He  S.-J. Chen  B.-X. Mai State Key Laboratory of Organic Geochemistry, Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, China e-mail: [email protected] J. Liu  L.-H. Yu  M.-J. He Graduate School of the Chinese Academy of Sciences, Beijing 100039, China

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respective concentration ranges of nd to 12 and nd to 0.68 ng/g lw. 4-MeSO2-CB101 exhibited the highest concentration among the nine PCB and PBDE derivatives studied, indicating that biotransformation via the mercapturic acid pathway of PCBs might have occurred in the studied bird species.

The occurrence and bioaccumulation of polychlorinated biphenyles (PCBs) and polybrominated diphenyl ethers (PBDEs) in biota have been extensively studied since their discovery as environmental pollutants several decades ago (Law et al. 2003). The adverse effects of PCB and PBDE exposure in wildlife include endocrine disfunction, reproductive failure, immunological impairment, developmental stress, and genotoxic disorders (Beineke et al. 2005; Das et al. 2006). The mechanisms of their adverse effects on biologic systems, particularly for PBDEs, are still not completely understood (Hakk and Letcher 2003). Some of their toxicities may be link to the in vivo biotransformation of PCBs and PBDEs. PCBs can be metabolized to hydroxylated PCB derivatives (OH-PCBs) by direct hydroxylation of the parent PCBs and to methylsulfonyl derivatives (MeSO2-PCBs) via the mercapturic acid pathway (Letcher et al. 2000). These metabolites have been found in both humans and animals (Houde et al. 2006; Hovander et al. 2002; Kunisue and Tanabe 2009). Recent studies have shown that the levels of MeSO2-PCBs and OH-PCBs are similar to or even higher than the levels of their parent compounds in some marine organisms (Letcher et al. 2000). Results of toxicity studies suggested that methylsulfonyl PCB exposure would reduce the thyroid hormone levels and serum thyroxine concentrations in rats (Kato et al.2000). These findings implied that, in addition to PCBs themselves, their

Arch Environ Contam Toxicol (2010) 59:492–501

metabolites may also play an important role in adverse effects for wildlife. OH-PBDEs as well as MeO-PBDEs have also been found in a number of abiotic and biotic matrices, such as water, snow (Ueno et al. 2008), alga, mussel, cyanobacteria (Malmva¨rn et al. 2008), fish, bird, mammals (Marsh et al. 2004; Verreault et al. 2005), and humans (Fa¨ngstrom et al. 2005; Weiss et al. 2006). Meanwhile, some OH-PBDEs and MeO-PBDEs have been found to be biomagnified significantly through the marine food web (Kelly et al. 2008).The mechanisms of OH- and MeO-PBDEs formation are unclarified so far. A possible mechanism of OH-PBDE formation may be via direct hydroxylation and a 1,2-shift of a bromine atom during arene oxide ring opening in vivo (Malmberg et al. 2005; Marsh et al. 2006). Some OHPBDEs, due to their structural similarity to T4, competitively bind to transthyretin (TTR; the thyroid hormone transport protein) and the estrogen receptor (Era, Erb). So it can lead to disruption and imbalance between androgens and estrogens (Meerts et al. 2001). Previous research on MeO- and OH-PBDEs suggests that several of these compounds with a MeO- or OH-group in the ortho position can be formed naturally in marine algae or by their associated microorganisms. For example, 20 -MeO-BDE68 and 6-MeO-BDE47 in the blubber of North Atlantic True’s beaked whales (Mesoplodon mirus) had been identified as originating from a natural product using radiocarbon (14C) analysis (Teuten et al. 2005). Malmva¨rn et al. (2008) recently reported that seven OHPBDEs and four MeO-PBDEs found in red algae and cynaobacteria from the Baltic Sea were all ortho MeO or OH substituted, suggesting a biogenic origin of these compounds. It is not clear whether these natural PBDE derivatives in marine environment can be found in a terrestrial environment. To date, most studies on OH-PBDEs and MeO-PBDEs were conducted on marine environment such as fish, fish-eating birds, and marine mammals. Little attention has been given to OH-PBDEs and MeO-PBDEs in terrestrial animals (McKinney et al. 2006). In the present study, we collected serum samples from inland birds inhabiting an intensive e-waste recycling site in South China, and we report measurements of PCBs and PBDEs determined in the serum samples of these birds. In our previous study, high levels of PCBs and PBDEs have been observed in muscle tissues of wild and domestic birds in this area (Luo et al. 2009a, b). The PCB and PBDE congener profiles of serum were compared with those of muscle in our previous study to investigate the serum-muscle tissue distributions within a certain species. The PBDE and PCB derivatives, OH-PBDEs, MeOPBDEs, and MeSO2-PCBs, were also identified and quantified in bird serum.

493

Materials and Methods Sample Collection Eleven white-breasted waterhens (Amaurornis phoenicurus), five pintail snipes (Gallinago stenura), one Chinese pond heron (Ardeola bacchus), two lesser coucals (Centropus bengalensis), five spotted doves (Streptopelia chinensis), three collared doves (Streptopelia decaocto), three common pheasants (Phasianus colchicus), and four common quails (Coturnix coturnix) were collected from a large e-waste recycling region, located in Qingyuan County in South China, between March and July in 2008. The common pheasant and common quail are farmed birds, while others are wild species. Small blood samples, *2 ml, were taken from the brachial or jugular vein of each bird using syringes. For samples with a blood volume \2 ml, two to five bird blood samples per species were pooled to obtain an adequate sample volume for analysis (1.5 ml serum: 1 ml for extraction and 0.5 ml for total lipid determination). The pooled blood samples were transferred into 10-ml Teflon tubes precleaned with acetone (ACE), dichloromethane (DCM), and hexane (HEX). Then they were immediately centrifuged at 3000 rpm for 15 min at 4°C. The serum (the top layer) was sucked up with precleaned capillaries and stored in glass tubes, then frozen at -20°C until analysis. Extraction and Cleanup Procedure The procedures for extraction and cleanup were developed based on Rivera-Rodriguez et al. (2007) HLB-urea extraction program. Bird serum samples were brought to room temperature and an aliquot of each sample (1 ml) was transferred to a 15-ml vial that contained a stirring bar(13 9 3 mm, cylindrical, PTFE). After spiking with surrogate standards—BDE77 for PBDEs, MeO-PBDEs, MeSO2-PCBs, PCB30, and PCB204 for PCBs, and 4-OHCB72 for OH-PBDEs—the serum proteins were denatured with a 500-mg portion of urea (*8 M), and the mixture was stirred gently for 30 min. The mixed samples were transferred to a 3-ml preconditioned Waters Oasis HLB extraction cartridge, mounted in a SPE-12G glass vacuum manifold (J. T. Baker No. 7018-00). The cartridges were conditioned by rinsing with methanol, then deionized water. After residual water was removed, the diluted serum was pulled through the cartridge at a low flow under a low vacuum (13 mm Hg). The 15-ml vial was rinsed three times with 1 ml of deionized water which was also passed through the cartridge. And then a vacuum was applied for 10 min to remove residual water. Target analytes were eluted with 1.5 ml HEX (fraction 1), 1.5 ml 7:3 (v/v) HEX/DCM (fraction 2), 2 ml 1:1 (v/v)

123

494

HEX/DCM (fraction 3), and 1 ml DCM (fraction 4), successively. Fractions 1 and 2 were pooled for analysis of neutral compounds and fractions 3 and 4 were pooled for phenolic compounds. The neutral fraction was concentrated and transferred to a 1.1-ml microvial. Then the extract was finally blown to near-dryness under gentle nitrogen and adjusted to 10 ll with 5 ll PBDE internal standards (13C-PCB208, BDE118, and BDE128) and 5 ll PCB internal standards (PCB24, -82, and -198) just before sample injection of GC/MS. The phenolic fraction was derivatized to theMeO analogues through a methylation reaction using diazomethane. The derivatized phenolic compounds were cleaned up on a silica gel column (i.d. = 0.8 cm) with 1 cm netural silica (3% deactivated) and 7 cm sulfuric acid silica (2:1 w/w), using 5 ml HEX and 5 ml HEX:DCM (1:1, v/v) as the mobile phase. After being blown to dryness, the extract was reconstituted in 10 ll with 5 ll iso-octane and 5 ll PBDE internal standards. Instruments PCBs (13 congeners) were separated and determined on a fused silica DB-5 ms column (60 m 9 0.25 mm 9 0.25-lm film thickness) in an Agilent 6890 gas chromatography (GC) coupled with a 5975B mass spectrometer (MS) system using electron impact (EI) ion source. Details of the instrument temperature programs have been described elsewhere (Wu et al. 2008). All injections were made in the splitless mode using an injection volume of 2 ll. PBDEs (10 congeners), MeO-PBDEs (3-MeO-BDE47 and 6-MeOBDE47), OH-PBDE (40 -OH-BDE-17, 3-OHBDE47, 6-OH-BDE47, 40 -OH-BDE-49, and 20 -OH-BDE68), and MeSO2-PCBs (4-MeSO2-CB49 and 4-MeSO2-CB101) were quantified by a Shimadzu 2010 GC coupled with a QP2010 MS in electron-capture negative-ionization (ECNI) mode and operated in selected ion monitoring (SIM). A DB-5 ms column (30 m 9 0.25 mm 9 0.25-lm film thickness) was used to determine PBDEs, MeO-PBDEs, and MeSO2-PCB. For OH-PBDE, DB-5HT (15 m 9 0.25 mm 9 0.1-lm film thickness) was used. The instrument temperature programs for PBDE, MeO-PBDE, and MeSO2-PCB determination were the same as in our previous study (Mai et al. 2005). The ions monitored were m/z 79 and 81 for MeO-PBDEs, m/z 389 and 404 for 4-MeSO2-CB101, and m/z 355 and 370 for 4-MeSO2-CB49, respectively. The column oven program for OH-PBDE analysis was programmed from an initial temperature of 80°C, held for 1 min, then increased at a rate of 8° min-1 to 300°C and held for 25 min. The monitored ions for OH-PBDEs were m/z 79, 81, and 161. All analytes were identified by comparing analyte retention times and ECNI or EI mass spectra to those of

123

Arch Environ Contam Toxicol (2010) 59:492–501

authentic standards under identical GC conditions. Quantification was based on internal calibration curves made from standard solutions at six concentration levels. Enzymatic Determination of Serum Lipids Concentrations of triglycerides (TGs) and total cholesterol (CHOL) in serum were determined enzymatically with an Hitachi7080 biochemical analyzer at the Haikang Policlinic of Occupational Disease Prevention and Cure Yard of Guangdong Province. The total lipid (TL) concentration in serum (g/l) was calculated by the equation: TL = 0.9 ? 1.3 (CTG ? CCHOL) (Malmberg et al. 2005). Quality Control The recoveries of added surrogates in each sample were 68–107% for CB30, 60–107% for CB204, 62–104% for BDE77, and 77–96% for 4-OH-CB72. There were three procedural blanks analyzed simultaneously with each bath of six samples in our study. During the PBDE analysis, BDE85 and BDE138 were detected systematically and background subtracted for all samples. For PCB and OHPBDE quantification, CB28, CB138, and 20 -OH-BDE68 were found, respectively, and thus deducted from the samples. In addition, three spiked blanks and three spiked matrixes were performed to test the feasibility of our experiment procedure, and the ranges of mean recoveries of targets in them were 57–87% and 50–130%, respectively. The method limit of quantification (MLOQ) was determined, based on five times the signal-to-noise ratio (S/N). The MLOQs for PBDEs, MeO-PBDEs, and MeSO2PCBs were between 0.01 and 0.3 ng/g lw. For OH-PBDEs and PCBs, they were 0.06–0.9 and 0.2–3 ng/g lw, respectively.

Results and Discussion PBDEs The total PBDE concentrations (sum of BDEs 28, 47, 66, 85, 100, 99, 138, 153, 154, and 183) ranged from 0.64 ng/g lw in white-breasted waterhen to 580 ng/g lw in lesser coucal. Farm-raised birds (common pheasants and common quail) have relatively lower levels of PBDEs compared with wild species. The PBDE concentrations in lesser coucal were significantly higher than those in other species (Table 1). The differences in habitat and dietary compositions among different species could possibly be used to explain this observation. Wild birds live in a relatively larger-scale area than farm-raised birds, which gives the wild species more chance to come into contact with

2.4

30

29

220

8.8

0.25

2.3

8.8

4.2

nd

14

14

1.5

54

100

8.5 nd

5.0

nd

22

6.6

97

16

3.4

110

5.0

61

BDE47

BDE66

BDE85

BDE99

BDE100

BDE138

BDE153

BDE154

BDE183 P PBDEs

PCB28

PCB66 PCB74

PCB99

PCB105

PCB118

PCB128

PCB138

PCB153

PCB164

PCB180

PCB187

0.96

3.9

0.22

1.41 2.0

2.0

5.7

2.5

40 -OH-BDE17

20 -OH-BDE68

6-OH-BDE47 3-OH-BDE47

40 -OH-BDE49 P OH-PBDEs

6-MeO-BDE47

1.1

nd 1.6

0.21

1.17

430

nd

PCB190 P PCBs

4.4

40

4.8

20

52

6.5

27

1.9

4.2

5.7 2.9

20

6.8

12

0.29

2.0

3.8

0.49

0.12

2.8

0.02

0.15

BDE28

7.9

4.8

Lipid (mg/ml)

nd

2.5

nd

nd 1.3

nd

1.21

57

nd

nd

2.6

nd

2.4

42

nd

3.6

nd

2.13

nd

0.56

nd

nd 0.56

nd

nd

50

nd

nd

nd

nd

1.8

39

nd

nd

nd

2.7

4.8 1.7 nd

ndb nd

0.64

nd

nd

0.10

nd

0.05

0.15

nd

0.08

0.21

0.05

9.6

4.6

3.9

0.29

0.40

0.58

0.10

0.27

1.2

nd

0.15

0.74

0.11

6.9

4

nd

2.0

nd

nd 0.99

0.14

0.83

38

nd

nd

1.6

nd

1.6

29

nd

3.1

nd

nd

nd nd

2.9

2.0

0.22

0.23

0.33

0.20

0.14

0.51

nd

0.03

0.35

0.04

9.1

5

1.1

1.7

nd

nd nd

1.66

nd

270

31

5.3

54

nd

18

68

nd

19

2.8

8.9

8.8 4.2

46

19

0.35

2.7

3.5

nd

1.7

5.7

0.10

0.24

4.3

0.14

5.7

6

0.06

2.2

nd

nd 1. 9

0.25

nd

100

6.5

1.4

8.5

1.4

5.7

48

1.8

10

0.23

4.1

3.3 1.3

8.9

8.5

0.48

0.42

0.70

0.17

0.56

3.37

0.01

0.10

2.6

0.09

7.4

7(2)

nd

2.0

nd

0.49 0.51

1.0

nd

78

nd

nd

nd

nd

nd

41

nd

nd

nd

3.5

2.3 nd

31

1.19

nd

0.22

0.12

nd

nd

0.37

nd

0.05

0.36

0.07

9.0

0.13

0.85

nd

nd 0.46

0.40

nd

82

nd

nd

4.6

2.2

5.8

25

1.2

9.7

2.9

6.2

4. 7 3.9

16

7.4

0.25

0.51

0.72

0.18

0.45

2.2

0.05

0.27

2.70

0.06

8.8

9(3)

8(2)

3(2)

1(2)a

2(2)

Pintail snipe

White-breasted waterhen

Table 1 Concentration (ng/g lipid wt) of organohalgen compounds in serum of birds

0.18

1.7

nd

0.96 0.78

nd

nd

220

22

1.0

42

nd

2.3

75.

nd

2.4

nd

3.6

4.4 nd

68

6.2

nd

1.4

0.80

nd

0.44

0.58

nd

0.11

2.8

0.12

5.6

Chinese pond heron 10

1.1

4. 8

nd

nd 4.3

0.51

nd

1700

130

19

250

39

340

200

28

340

49

76

53 50

150

580

1.1

11

37

2.1

21

56

2.3

0.48

450

0.20

14

11

1.0

13

nd

nd 13

0.70

nd

1000

48

15

93

29

190

130

21

230

42

51

60 43

80

300

0.62

9.7

15

1.8

15

19

0.83

0.67

230

0.25

7.3

12

Lesser coucal

nd

5.1

1.6

nd 1.6

1.9

nd

62

nd

nd

nd

nd

nd

50

nd

nd

nd

nd

nd nd

12

1.3

0.22

nd

0.43

nd

nd

0.17

nd

0.12

0.23

0.11

3.7

13(2)

nd

7.8

nd

nd nd

7.8

nd

68

nd

nd

nd

nd

nd

49

nd

nd

nd

3.1

nd nd

15

1.8

0.28

0.11

0.34

0.37

nd

0.34

nd

0.08

0.21

0.09

3.9

14

Common pheasant

nd

5.4

1.2

nd 2.1

0.15

1.9

59

nd

nd

nd

nd

nd

54

nd

nd

nd

1.9

nd nd

3.0

2.8

0.37

0.25

0.35

nd

0.17

0.64

0.07

0.11

0.69

0.12

5.9

15(4)

Common quail

nd

4.3

0.92

nd 1.4

0.60

1.3

130

3.2

nd

7.6

nd

6.3

69

1.9

16

2.2

2.9

2.5 3. 7

19

9.5

2.6

0.10

4.5

0.16

nd

1.32

nd

0.08

0.69

0.09

5.4

16(5)

Spotted dove

0.47

2.6

nd

nd 0.94

1.64

nd

260

32

nd

35

nd

4.5

88

nd

6.1

nd

6.2

6.5 nd

83

13

0.35

0.27

1.9

nd

0.58

4.61

nd

nd

4.8

0.21

4.9

17(3)

Collared dove

Arch Environ Contam Toxicol (2010) 59:492–501 495

123

12 Number in parentheses indicates the number of pooled birds

Under the method limit of quantification (MLOQ) b

a

nd nd

nd nd

nd nd

0.30 1.8

nd 0.68

2.8 8.9

0.20 nd

12 0.46

nd nd

6.8 0.37

nd nd

6.6 nd

nd nd

1.0 nd

nd

7.2 4-MeSO2-CB101

1.6

0.07 4-MeSO2-CB49

0.33

0.07 nd 0.08 nd nd 0.10 nd 0.23 nd nd 0.17 0.15 nd nd 0.11

6 5 4 3(2) 2(2)

0.52 0.89

9(3) 8(2) 1(2)a

7(2)

Pintail snipe White-breasted waterhen Table 1 continued

123

3-MeO-BDE47

16(5) 15(4) 14 13(2) 12 11

Spotted dove Common quail Common pheasant

Chinese pond heron 10

Lesser coucal

17(3)

Arch Environ Contam Toxicol (2010) 59:492–501 Collared dove

496

e-waste or food contaminated by e-waste in the study area, resulting in a high PBDE burden in their body. Lesser coucal feed mainly on insects, but they also eat animals such as snake, lizards, mice, and bird eggs. Therefore, the high contaminant levels in the lesser coucal could contribute, in part, to their relatively high trophic level (Hop et al. 2002). White-breasted waterhen is an insectivore/ granivore bird generally feeding on insects, worms (about 80%), and marsh plant shoots, which means it lives at a lower trophic level than lesser coucal. This could be the reason for its low concentrations. Few studies have reported the level of PBDEs in bird blood. Verreault et al. (2005) reported that the mean level of PBDEs in glaucous gull’s blood from the Norwegian Arctic was 20 ng/g wet weight (ww), which is much higher than those in the present study (\0.01–8.5 ng/g ww). The level of PBDEs in lesser coucal (8.4 and 2.2 ng/g ww) in the present study were comparable with those in bald eagle plasma (1.8–8.5 ng/g ww) from the western coast of North America (McKinney et al. 2006) and bald eaglet plasma samples (7.9 ng/g ww) from Lake Superior (Dykstra et al. 2005). Generally, PBDE congener profiles in birds of the present study could be classified into three groups according to the prominent compounds (Fig. 1). Chinese pond heron and lesser coucal clustered in one group, in which BDE47 was the predominant congener, accounting for 45 and 78% in Chinese pond heron and lesser coucal, respectively. This congener profile is in line with those in plasma of bald eaglet from the western coast of North America (McKinney et al. 2006). Breasted waterhen, pinstail snipe, common quail, and collared dove clustered in another group, in which both BDE 47 and BDE 99 make similar contributions to the total PBDEs, followed by BDE153, -154, -183, and -100. Spotted dove and common pheasant belong to the third group, where BDE153 and BDE183 were the major constituents, then BDE99 and BDE47. Different dietary compositions and metabolic capabilities might be responsible for these observed profile differences among species (Voorspoels et al. 2006). For example, fish is the main composition of the diet of Chinese pond heron. The PBDE congener pattern of Chinese pond heron is consistent with those of fish (mud carp, crucial carp, northern snakehead, and common carp) collected from the same region (Wu et al. 2008) (Fig. 2), suggesting that the PBDE profiles in piscivorous birds were influenced by their feeding habit to a large extent. Due to the limited serum samples, the factors influencing the interspecies differences in congener profiles were not investigated in the present study. The relatively large sample of white-breasted waterhen analyzed (seven composite samples from 11 individuals) made it possible to compare the PBDE congener profiles

45

30

15

Average compositions of bird and fish

60

0.8

Chinese pond heron Mud carp Crucian carp Northern snakehead Carp

0.7 0.6 0.5 0.4 0.3 0.2 0.1 0.0

BDE183

BDE154

BDE153

BDE138

BDE100

BDE99

b

White breasted waterhen Pintail snipe Common quail Collared dove

30

20

10

BDE183

BDE154

BDE153

BDE138

BDE100

BDE99

c

BDE85

BDE66

BDE47

0

Spotted dove Common pheasant

40

30

between serum and muscle tissues. A significant difference in PBDE congener profile was found between serum and muscle, investigated using ANOVA. BDE153 over BDE47 and BDE99, rather than BDE 47 and 99 over BDE153, was observed in muscle tissues of white-breasted waterhen (Luo et al. 2009b). The sum of BDE28, -47, -99, and -100 accounted for 60% in serum but for 46% in muscle. In addition, highly brominated congeners such as BDE196 and BDE207 were hardly detected in serum, but they were detectable in muscle. These results suggest that less brominated congeners might preferentially accumulate in sera of this bird species. In a study of levels and tissue distributions of PBDE in birds of prey in Belgium, Voorspoels et al. (2006) found that BDE47, BDE99, and BDE153 were the major congeners and that no differences in PBDE congener profiles existed among the various tissues (serum, muscle, liver, fat, and brain) within individuals of a certain bird species. However, in glaucous gulls, BDE 47 and -99 accounted for a greater proportion of the total PBDEs in blood than in the whole body of the gulls (Verreault et al. 2007). PCBs

20

10

BDE183

BDE154

BDE153

BDE138

BDE100

BDE99

BDE85

BDE66

BDE47

BDE28

0

BDE85

Fig. 2 Congener composition (%) of PBDEs in Chinese pond heron sera and fish. Data for fish are from Wu et al. (2008)

40

50

BDE66

BDE183

BDE154

BDE153

BDE138

BDE100

BDE99

BDE85

BDE66

50

BDE47

BDE28

0 BDE47

Average compositions of PBDE congeners (%)

497

Chinese pond heron Lesser coucal

BDE28

Average compositions of PBDE congeners (%)

75

a

BDE28

Average compositions of PBDE congeners (%)

Arch Environ Contam Toxicol (2010) 59:492–501

Fig. 1 Average congener composition (%) of PBDEs in sera of eight bird species. Error bars represent the standard error. a Chinese pond heron and lesser coucal. b White-breasted waterhen, pintail snipe, common quail, and collared dove. c Spotted dove and common pheasant

Thirteen PCB congeners, including CBs 28, 66, 74, 99, 105, 118, 128, 138, 153, 164, 180, 187, and 190, were P detected in the samples. Concentrations of PCBs ranged from 38 to 1700 ng/g lw and were one to two orders of P magnitude higher than the PBDE concentrations (Table 1). This result is in line with our previous study (Luo et al. 2009b) and most studies in other regions (Verreault et al. 2005; Gebbink et al. 2008; Verreault et al. 2006). Similarly to PBDEs, lesser coucal showed the P highest PCBs concentration among the investigated

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PBDEs concentrations (ng/g lipid weight)

species, while the two farm-raised species (common pheasantsP and common quail) had relatively low serum levels of PCBs. A significantly positive correlation was found between PCB and PBDE concentrations, suggesting that they have similar exposure pathways and/or mechanisms of accumulation (Fig. 3). The concentrations of PCBs in bird serum samples in our study (0.24–24.78 ng/g ww) were similar to those in bald eagle (Haliaeetus leucocephalus) nestling plasma collected from the western coast of North American (2.7–39.6 ng/g ww) (McKinney 400 350 300

R=0.95, p < 0.0001

250 200 150 100 50 0 0

200

400

600

800

1000

1200

1400

PCBs concentrations (ng/g lipid weight)

P P PBDE concentrations and PCB

Fig. 4 Average congener composition (%) of PCBs in sera of eight bird species. Error bars represent the standard error

Average compositions of PCB congeners in bird serum

Fig. 3 Correlation between concentrations in bird sera

0.20 0.15 0.10 0.05 0.4 0.3 0.2 0.1 0.3 0.2 0.1 0.60 0.45 0.30 0.15 0.8 0.6 0.4 0.2 0.9 0.6 0.3

Lesser coucal Pintail snipe Chinese pond heron White breasted waterhen

Common pheasant Common quail

0.45 0.30 0.15

Spotted dove

0.3 0.2 0.1 0.0

Collared dove CB190

CB187

CB180

CB164

CB153

CB138

CB128

CB118

CB105

CB99

CB74

CB66

CB28

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et al. 2006) but were 1–3 orders of magnitude lower than those in the whole blood of breeding lesser black-backed gulls from the coast of northern Norway (3.28–207.55 ng/g ww) (Bustnes et al. 2008) and in glaucous gull blood from the Norwegian Arctic (124–2655 ng/g ww) (Verreault et al. 2006, 2007). With the exception of lesser coucal, the PCB congener profiles were similar in all species; they were dominated by CB138, ranging from 32 to 91%, followed by CB28, -180, 118, and -153 (Fig. 4). This congener profile was similar to that reported in sera of humans consuming Great Lakes fish, in which CB138 was the most abundant congener, followed by CB180 and CB153. CB28, CB118, and CB170/190 also made a large contribution to the total PCBs (Humphrey et al. 2000). In lesser coucal, CB118 and CB153 made the largest contribution to the total PCBs, accounting for 20 and 19%, respectively, followed by CB138, -180, and -28. In our previous investigation of PCBs in muscle tissue of birds, PCB153, -138, -118, and 180 were the most abundant PCB congeners (Luo et al. 2009b). There was a slight difference in PCB congener profiles between muscle tissue and serum. The lowerchlorinated PCB congeners, such as CB28, contributed more to the total PCBs in serum (average 11.95%) compared to muscle (average, 0.27%). A study of tissue distributions and half-lives of individual PCBs in rat indicated that CB28 haves the lowest rate of decrease in the serum ¨ berg et al. 2002), which could explain the high (O

Arch Environ Contam Toxicol (2010) 59:492–501

concentration of BDE28 in serum samples in the present study. OH- and MeO-PBDEs and MeSO2-PCB Of the five OH-PBDE congeners monitored in serum samples, only 20 -OH-BDE68 and 3-OH-BDE47 were consistently measured in more than 80% of the samples. 40 -OHBDE17, 6-OH-BDE47, and 40 -OH-BDE49 were detected in less than 30% of the samples. Two MeO-PBDE congeners, 3-MeO-BDE47 and 6-MeO-BDE47, were detected in half of the samples (Table 1). The ortho-substituted 20 -OH-BDE68 found in sera of birds appears to be a natural product accumulated by the birds from their diet. 20 -OH-BDE68 has been detected in freshwater fish, and it was also suggested to be of natural origin (Valters et al. 2005; Kierkegaard et al. 2004). Atmospheric long-range transportation from the marine environment is a possible source of natural OH- and MeOPBDEs in inland environments (Ueno et al. 2008). On the other hand, the meta-substituted 3-OH-BDE47 in bird sera is more likely to be a metabolite of anthropogenic PBDEs, such as the biologically predominant BDE47 congener. The metabolic formation of 3-OH-BDE47 has been demonstrated in a previous study in rodents dosed with BDE47 and was considered to be related to CYP enzyme -mediated biotransformation (Marsh et al. 2006; Sanders et al. 2005). To our knowledge, 3-OH-BDE47 has not yet been confirmed as or reported t be naturally occurring in any species. Of the three less detectable OH-PBDE congeners, the para-substituted 40 -OH-BDE17 and 40 -OH-BDE49 in bird sera may be derived from metabolism of precursor BDE47, since meta- and para-substituted OH-PBDEs have not been detected in marine algae or other microorganisms (Malmva¨rn et al. 2008). 40 -OH-BDE49 has been identified as a major metabolite in plasma and feces of rodents exposed to PBDEs (Malmberg et al. 2005; Marsh et al. 2006). Metabolism of BDE47 in biota would result in the formation of 4’-OH-BDE49 via a 1,2-bromine shift analogous to OH-PCB formation from PCBs (Hakk and Letcher 2003). However, both metabolic formation and natural sources may be the origin of 6-OH-BDE47 detected in bird sera. The high detection frequency of this ortho-substituted 6-OH-BDE47 in marine algae and its associated microflora and/or microfauna indicates its natural origin (Malmva¨rn et al. 2008). But 6-OH-BDE47 is also known to be a potential metabolite of the biologically predominant BDE47 congener, as it was detected in rats exposed to BDE47 (Marsh et al. 2006). MeO-PBDE congeners have been proven to be of natural origin and there is no evidence to date that MeOPBDEs are metabolites of PBDEs (Lacorte and Ikonomou 2009; Malmva¨rn et al. 2008). A study conducted by Marsh

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et al. (2004) showed that 3-MeO-BDE47 coelutes with another ortho-substituted MeO-PBDE congener on a DB5-type GC column. So the 3-MeO-BDE47 (or 20 -MeOBDE66) detected in the present study might be 20 -MeOBDE66 or a combination of 3-MeO-BDE47 and 20 -MeOBDE66. Therefore, he two detectable MeO-PBDEs, 3MeO-BDE47 (or 20 -MeO-BDE66) and 6-MeO-BDE47, in the present study are likely the result of bioaccumulation of marine-derived natural products. Nonetheless, it has been argued that certain meta or para MeO-substituted PBDEs may be formed in vivo through methylation of OH-PBDEs (metabolically formed or accumulated from the diet) or direct methoxylation of PBDEs as a protective mechanism against toxicity (Verreault et al. 2005). So metabolic formation from BDE47 in organisms cannot be excluded as a potential source of 3-MeO-BDE47 present in the serum. The concentration of 4-MeSO2-CB101 (nd to 12 ng/g lw) was substantially higher than that of 4-MeSO2-CB49 (nd to 0.68 ng/g lw) (p \ 0.05). Its concentration is also the highest among the detected PCB and PBDE derivatives. Previous studies have suggested that CB101, as a more readily metabolized congener, tends to transform to persistent meta- and para-substituted methyl sulfones of CB101 (Fa¨ngstrom et al. 2005; Altshul et al. 2004). This might explain the higher concentration of 4-MeSO2-CB101 in our samples.

Conclusion This study has presented data on PBDEs, PCBs, and their derivatives, OH-PBDEs, MeO-PBDEs, and MeSO2-PCBs, in sera of avian species inhabiting an e-waste recycling P region in South China. Levels of PCBs were 1–2 orders P of magnitude higher than those of PBDEs. PCBs were well correlated with PBDEs, indicating that PCBs and PBDEs may have a similar exposure pathway. 20 -OHBDE68, 3-OH-BDE47, 3-MeO-BDE47, 6-MeO-BDE47, and 4-MeSO2-CB101 were detected in more than half of the collected samples. These derivatives could originate from both natural products and metabolism of parent compounds. The species-specific congener profiles for PCBs and PBDEs observed in the present study warrant that more attention be focused on the factors, such as dietary composition, habitat, trophic level, and metabolic capacity, that influence the congener profiles in the species. Further studies on the formation mechanisms of PBDE metabolic degradation in terrestrial animals are needed, and measurement of OH- and/or MeO-PBDEs in abiotic media (such as air, precipitation, and water) in an inland environment would undoubtedly provide more insight into the sources, long-range transport potential, and environmental fate of these compounds.

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500 Acknowledgments This research was supported by Grants NSFC20890112, NBRPC2009CB421604, NSFC40632012, and NSFC 40773061. This is contribution No. IS-1167 from GIGCAS. We acknowledge the assistance of Mr. T. S. Xiang in GC/MS analyses.

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