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Bioaccumulation and trophic transfer of polybrominated diphenyl ethers (PBDEs) in biota from the Pearl River Estuary, South China
Environment International 35 (2009) 1090–1095
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Environment International j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / e n v i n t
Bioaccumulation and trophic transfer of polybrominated diphenyl ethers (PBDEs) in biota from the Pearl River Estuary, South China Mei Yu a,b, Xiao-Jun Luo a,⁎, Jiang-Ping Wu a,b, She-Jun Chen a, Bi-Xian Mai a a b
State Key Laboratory of Organic Geochemistry, Guangzhou Institute of Geochemistry, Chinese Academy of Science, Guangzhou 510640, China Graduate School of Chinese Academy of Science, Beijing, 100039, China
a r t i c l e
i n f o
Article history: Received 16 March 2009 Accepted 12 June 2009 Available online 17 July 2009 Keywords: Polybrominated diphenyl ethers (PBDEs) Biomagnification Bioaccumulation Pearl River Estuary
a b s t r a c t Two hundred and fifty-four biota samples (four species of invertebrates and ten species of fish) were collected from the Pearl River Estuary between 2005 and 2007 and one hundred and twenty four individual or composite samples were analyzed for polybrominated diphenyl ethers (PBDEs). The concentrations of PBDEs in organisms varied from 6.2 to 208 ng/g lipid weight. This PBDE level was significantly lower than those collected in 2004, showing a decreasing trend of PBDEs in biota in the study area. Trophic magnification factors (TMFs) for nine BDE congeners were calculated with values ranging from 0.78 to 3.0. TMFs of BDE47, 66, 100, 99, 154, and 153 were statistically greater than one, indicating a biomagnifcation potential for these congeners. Significant positive correlations were also found between concentrations of the total PBDEs, BDE28, 47, 66, 100, 99, 154, and153 and lipid content in biota, indicating the that bioconcentration also played an important role in the accumulation of PBDEs. No correlation between trophic level and lipid content was found, suggesting that biomagnification was not the result of lipid content effect but indeed occurred. The concentration ratios of BDE99 to BDE100 were much lower in biota than that in water implying that potential congener-specific biotransformation of PBDEs occurred and influenced the biomagnification of BDE congeners. © 2009 Elsevier Ltd. All rights reserved.
1. Introduction Polybrominated diphenyl ethers (PBDEs) are flame retardants widely used in plastic, textile, electronic and other material. Because of their ubiquity in the environment (Law et al., 2008; Schecter et al., 2003; Vonderheide et al., 2008) and their exponentially increasing concentrations in most environment compartments with doubling times of about 4–6 years (Chen et al., 2007; Hites, 2004), the behaviors of PBDEs in environment have attracted great attention and interests from the public and environmental community. Available toxicological evidences indicate that PBDEs can disturb thyroid homeostasis, cause hepatomegaly and neurobehavioral deficits, and exhibit fetal and maternal toxicity after prolonged exposure (Darnerud et al., 2001; McDonald, 2002). PBDEs are capable of bioaccumulation in the biota own to their properties of high lipophilicity (log KOW: 5.9–10) and resistance to metabolism (Gustafsson et al., 1999). Recently, a few of papers reported that the bioaccumulation and biomagnifications of PBDEs in biota, mostly focused on the marine and fresh food webs, and the conclusions varied. For example, Wan et al. (2008) have reported that concentrations of BDE28, 47, 100 and 119 increased significantly with the increasing trophic levels in a marine food web of Bohai Bay. In the tropical– ⁎ Corresponding author. Tel.: +86 20 85290146; fax: +86 20 85290706. E-mail address: [email protected] (X.-J. Luo). 0160-4120/$ – see front matter © 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.envint.2009.06.007
subtropical waters, BDE47, 100 and 154 were found biomagnifications but BDE66, 99 and 209 were found trophic dilution in a fresh water food web from a contaminated reservoir in South China (Wu et al., 2009). While analyses of PBDEs in biota from a fresh water food web from Lake Winnipeg indicated that only BDE47 and BDE209 have TMF greater than one (Law et al., 2006). Burreau et al. (2006) found the absence of biomagnification for the highly brominated diphenyl ethers such as BDE209 and nona-BDEs and the slight biomagnification for BDE47 (TMF = 1.5, calculated from B value) in the northern Atlantic Ocean food web. On the other hand, studies conducted by Guo et al. (2008) and Wang et al. (2007) found the absence of biomagnification for PBDEs in the organisms from the Pearl River Estuary and from a river receiving effluent discharged from a large sewage treatment, respectively. These results indicated that the bioaccumulation behavior of PBDEs in biota was complex and needs more field research. The Pear River Estuary is located in the Pearl River Delta region of southern China, one of the most economically developed regions in China. Many researches have evidenced the Pear River Estuary as a significant sink for organic pollutants including PBDEs from upstream runoff, which posed risks to the organisms in the region (Mai et al., 2005; Luo et al., 2008; Wurl et al., 2006). Guo et al. (2008) have carried out a preliminary research about accumulation of PBDEs in biota in this area and found the absence of biomagnification of PBDEs in the food web. The limited samples number (50 samples for 34 species, average less than 2 samples for each species) and short food
BD (n = 6)
1091 SS, sand swimming crab (Ovalipes punctatus); SC, samoan crab (Scylla serrata); AS, ark shell (Tegillarca granosa); ON, oncomelania (Oncomelania hupensischiui); CM, common mullet (Mugil cephalus); RE, red eelgoby (Odontamblyopus rubicundus); RT, robust tonguefish (Cynoglossus robustus); SP, slimy spinefoot (Siganus canaliculatus); SL, silver sillago (Sillago sihama); PP, pompano (Psenopsis anomala); JE, Japanese eel (Anguilla japonica); FF, flathead fish (Platycephalus indicus); LY, large yellow croaker (Pseudosiaena crocea); BD, bombay duck (Harpodon nehereus). nd, not detected. a Average ± standard derive. b Median (min–max).
0.84 ± 0.34 15 ± 1.2 3.5 ± 0.31 nd (nd–0.8) 7.2 (2.3–18) 0.67 (0.2–1.1) 1.9 (1.1–10) 1.0 (0.17–4.8) 0.97 (0.38–7.4) 1.3 (0.13–16) 13 (6.3–54) 0.59 (nd–4.1) 0.7 (nd–6) nd (nd–0.66)
LY (n = 11) FF (n = 27)
0.77 ± 0.39 16 ± 1.2 3.6 ± 0.32 1.5 (nd–5.3) 38 (8.3–115) 1.3 (0.25–6.9) 9.4 (1.4–50) 0.59 (nd–18) 4.7 (0.54–30) 1.3 (nd–6.9) 59 (12–208) 0.36 (nd–5.5) nd (nd–53) nd (nd–0.31) 2.2 ± 0.91 16 ± 1.2 3.6 ± 0.31 1.1 (nd–5) 15 (5.0–39) 0.76 (0.3–3.4) 5.7 (2.2–12) 2.9 (0.62–8) 2.6 (0.99–4.7) 1.6 (0.72–4.7) 30 (14–71) 0.29 (nd–2.4) nd (nd–24) nd (nd–0.25)
JE (n = 19) PP (n = 6)
6.5 ± 1.4 17 ± 1.4 3.8 ± 0.38 1.2 (0.71–1.38) 28 (21–35) 1.3 (0.74–1.6) 5.5 (3.8–6.8) 4.7 (3.1–9) 2.2 (0.82–4.1) 1.6 (0.52–2.7) 44 (39–53) 0.09 (nd–0.58) 0.11 (nd–0.4) nd 0.87 ± 0.17 15 ± 1.2 3.4 ± 0.30 0.45 (nd–2) 13 (5.4–23) 0.45 (0.18–1.2) 1.1 (0.61–6) 0.45 (nd–1.3) 0.57 (0.36–2.9) 0.55 (0.17–4.1) 20 (7.8–37) 0.21 (nd–1) nd (nd–21) nd
SL (n = 5) SP (n = 12)
1.3 ± 0.44 16 ± 1.2 3.6 ± 0.31 0.52 (nd–2.7) 17 (6.4–47) 1.3 (0.61–3.9) 1.8 (0.73–5.4) 2.1 (0.51–14) 1.4 (0.36–2.9) 2.3 (0.94–5.7) 29 (14–73) 0.07 (nd–0.58) 4.5 (nd–69) 0.004 (nd–1.6) 0.97 ± 2.7 15 ± 1.4 3.4 ± 0.36 0.18 (nd–5.9) 17 (4.1–28) 0.71 (0.12–6.1) 4 (0.81–10) 3.5 (0.08–15) 2.9 (0.66–12) 2.4 (0.29–5) 32 (6.2–81) 1 (nd–6.7) nd (nd–0.06) nd
RT (n = 17) RE (n = 3)
0.68 ± 0.16 16 ± 0.35 3.6 ± 0.09 nd 7.3 (3.2–9.7) 0.4 (0.23–1.2) 2.1 (0.78–3.9) 0.69 (0.53–1.7) 1.4 (1.2–3) 4.1 (2.9–5.3) 19 (10–21) nd 22 (1.6–30) nd 4.9 ± 2.7 12 ± 1.4 2.6 ± 0.36 1.3 (nd–2.4) 14 (0.57–22) 0.64 (0.03–1.1) 1.6 (0.12–2.3) 0.25 (nd–1.4) 0.80 (0.08–2) 0.48 (0.22–5) 21 (1.3–29) 0.06 (nd–2.1) 0.09 (nd–31) nd (nd–1.6)
CM (n = 7) ON (n = 3) AS (n = 3) SC(n = 2) SS (n = 3)
Before the further treatment, the length and weight were recorded for each individual fish. Samples were thawed and dissected to remove skin and bones. Muscle fillets below the dorsal fin for fish and the soft part (edible part) for invertebrates were taken. The procedure for biota sample extraction and cleanup was described in detail in our previous studies (Xiang et al., 2007; Wu et al., 2008). Briefly, after being homogenized with ashed anhydrous sodium sulfate and spiked with surrogate standards (CDE99 and 13C12-PCB141), the samples were Soxhlet extracted with hexane/acetone (1:1, v/v) for 48 h. The extracts were concentrated and an aliquot of the extract was used to lipid content determination by gravimetric method, the other aliquot was subjected to gel permeation chromatography (GPC) to remove lipids. The cleaned extract was concentrated to approximately 1 mL and further purified passing through a multilayer silica/alumina column. The extracts were
2.5 ± 0.5 11 ± 0.3 2.4 ± 0.1 3.3 (3.2–5.1) 13 (12–18) 1.8 (1.7–2.3) 0.26 (0.26–0.34) 1 (1–1.3) 0.07 (0.06–0.07) 0.04 (0.03–0.07) 19 (18–27) nd nd nd
2.3. Extraction and cleanup
Lipid (%)a δ15N Trophic level BDE28b BDE47 BDE66 BDE100 BDE99 BDE154 BDE153 ∑(7 PBDEs) BDE138 BDE183 BDE209
All PBDE standards were purchased from Accustandards Inc. (New Haven, CT). CDE99 (2, 2′, 4, 4′, 5-pentachlorodiphenyl ether) was obtained from Wellington Laboratories (Ontario, Canada). 13C12PCB141 and 13C12-PCB208 were obtained from Cambridge Isotope Laboratories (Andover, MA, USA). All organic solvents were analytical reagent and redistilled using a glass system.
Species
2.2. Reagents and chemicals
Table 1 Biological parameters and PBDEs concentrations (ng/g lw) of 14 marine organisms collected from the Pearl River Estuary.
Organisms were collected using fish trawls in the Pearl River Estuary during August 2005 and August 2007. The detailed information about the sampling area has been given elsewhere (Xiang et al., 2007). All samples were stored in cool boxes and transported to the laboratory and then immediately transferred to a freezer where they were stored at −20 °C until analysis. The biota species collected in the present study included sand swimming crab (Ovalipes punctatus, three composite samples from nine individuals), samoan crab (Scylla serrata, two composite samples form six individuals), ark shell (Tegillarca granosa, three composite samples from sixty individuals), oncomelania (Oncomelania hupensischiui, three composite samples from thirty individuals), common mullet (Mugil cephalus, seven individual samples), red eelgoby (Odontamblyopus rubicundus, three composite samples from fifteen individuals), robust tonguefish (Cynoglossus robustus, seventeen individual samples), slimy spinefoot (Siganus canaliculatus, twelve individual samples), silver sillago (Sillago sihama, five individual samples), pompano (Psenopsis anomala, six composite samples from thirty individuals), Japanese eel (Anguilla japonica, nineteen individual samples), flathead fish (Platycephalus indicus, twenty seven individual samples), large yellow croaker (Pseudosiaena crocea, eleven individual samples), and bombay duck (Harpodon nehereus, six individual samples). Detailed information on the samples was provided in Table 1. Zooplankton samples were also collected using a tow net of 160 μm in August 2007, and the trawling depth ranged from 0 to 3 m. Two composite zooplankton samples were obtained and stored at −20 °C and further for nitrogen isotope measurements. A total of one hundred and twenty four samples were analyzed for PBDEs in this study.
1.7 ± 0.39 12 ± 0.05 2.5 ± 0.01 nd 2.3 (1.9–3.4) 0.13 (0.13–0.52) 2.9 (2.5–3.7) 6.5 (5.5–8.4) 2.6 (1.7–2.8) 3.1 (2.3–3.9) 17 (15–22) 1.6 (0.41–2.1) 1.5 (0.33–42) nd
2.1. Sampling
1.7 ± 0.64 15 ± 0.55 3.4 ± 0.14 0.5 (0.4–1.2) 6.89 (3.9–9.1) 0.86 (0.65–0.87 1.4 (0.79–1.4) 0.79 (0.51–1.4) 0.3 (0.07–0.48) 0.1 (0.08–0.51) 13 (6.4–13) 0.26 (nd–1.1) nd (nd–0.47) nd (nd–0.54)
2. Materials and methods
0.76 ± 0.07 16 ± 0.29 3.7 ± 0.08 1.2 (0.2–2.2) 13 (9.8–16) 0.59 (0.49–0.69 2 (0.61–3.3) 0.69 (0.63–0.76) 1.10 (17–2.1) 3.1 (0.36–5.7) 21 (14–29) 1.4 (nd–2.8) 16 (nd–32) nd
web may mask the biomagnification of PBDEs in the food web. In the present study, we analyzed concentrations of 10 BDE congeners in the food web (including four invertebrate species and ten fish species) and their relationship with δ15N and lipid content to understand their accumulation behaviors in the food web. We also compared the concentration ratios of BDE99 to BDE100 between biota and water samples to explore the potential congener-specific biotransformation of PBDEs in biota sample and its influence on the biomagnifications of BDE congeners.
0.91 ± 0.35 16 ± 0.93 3.5 ± 0.24 nd (nd–3.1) 5.9 (3.3–15) 0.30 (nd–2.6) 1 (0.49–3) 0.89 (nd–9.5) 0.35 (0.13–0.88) 0.8 (0.02–1.8) 9.8 (5–34) 0.8 (nd–3.8) nd (nd–1.8) nd (nd–1)
M. Yu et al. / Environment International 35 (2009) 1090–1095
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deviation was less than 0.6%. The isotope ratio was standardized against air according to δ15N = (Rsample /Rair − 1)× 1000‰, where R is the isotope ratio of 15N to 14N. The δ15N values were calibrated against a standard, ammonium sulfate (IAEA-N-1, International Atomatic Energy Agency Analytical Quality Control Services, Wien, Austria). The precision of the analytical method and instrument was ± 0.3‰. Trophic level (TL) was calculated for individual sample using the following equation (Post, 2002): 15
15
TL consumer = ½ðδ Nconsumer δ Nprimary
consumer Þ = 3:8
ð1Þ
+2
where 3.8 is the isotopic trophic enrichment factor. TMFs were calculated according to Tomy et al. (2004), and the references therein using following equations: Log concentration ðlipid normalizedÞ = A + B TL Fig. 1. Box plot concentrations of ∑PBDE in biota from the Pearl River Estuary among 2005 and 2007. The five lines from bottom to upper represent 5th, 25th, 50th, 75th, and 95th percentiles, respectively. Solid circles indicate average values. Asterisks above or below the boxes represent the maximum and the minimum.
concentrated, solvent exchanged to hexane, and finally concentrated to 50 μL under a gentle stream of nitrogen. A known amount of internal standard (BDE118, BDE128 and 13C12-PCB208) was added to all extracts prior to instrumental analysis. 2.4. Instrumental analysis Polybrominated diphenyl ethers (PBDEs) were measured with a Shimadzu model 2010 gas chromatograph coupled with a model QP2010 mass spectrometer (Shimadzu, Japan) using electron capture negative ionization (ECNI) in the selective ion monitoring (SIM) mode. A DB-XLB capillary column (30 m × 0.25 mm i.d. × 0.25 μm film thickness) was used to determine the tri- to hepta-BDEs (BDE28, 47, 66, 100, 99, 85, 153, 154, 138, and 183). For deca-BDEs (BDE209), a CPSil 13 CB capillary column (12.5 m × 0.25 mm i.d. × 0.20 μm film thickness) was used. Details of the GC temperature program as well as the procedure for qualification and quantification of PBDEs were given in published literature (Mai et al., 2005). The limit of detection (LOD), defined as a signal/noise ratio (S/N) = 5, ranged from 0.001 to 0.38 ng/g lipid weight. BDE85 was not quantified in all the samples because of chromatographic interferences. 2.5. Quality assurance/quality control Quality assurance was done by analyses of procedural blanks, triplicate spiked blanks, and triplicate spiked matrices. For each batch of 12 samples, a procedural blank was processed. Some of the procedural blanks (n = 15) contained traces of target chemicals, but the levels were close to the limit of quantification and they were not subtracted from those in samples. The mean recoveries of individual congeners (BDE28, 47, 66, 100, 99, 153, 154, 138, and 183) ranged from 61% to 87% with relative standard deviations (RSDs) less than 10% in triplicate spiked blanks and from 61% to 82% with RSDs less than 15% in triplicate spiked matrices, respectively. The surrogate standard recoveries of CDE99 and 13 C12-PCB141 were 83% ± 13% (ranged from 46%–125%) and 86% ± 16% (51%–118%), respectively. No surrogate corrections were made to final reported concentrations. 2.6. Nitrogen isotope measurements About 1 mg of homogenized and finely ground sample was wrapped with a tin cup, and placed in the sample container of an elemental analyzer-isotope ratio mass spectrometer (CE flash EA1112-Finnigan Delta plus XL) for stable nitrogen isotope ratio measurements. Two replicates of each sample were analyzed and the relative standard
ð2Þ
B
ð3Þ
TMF = 10
statistical significance of the regression Eq. (2) was defined at p b 0.05. 2.7. Statistical analysis For samples with contaminant concentration below LOD, zero was used for the calculations. All data were lipid-normalized. BDE85 was not quantified due to chromatographic interference in some of the samples. Simple linear regressions and correlation coefficients were used to examine correlation between the levels of PBDEs and lipid contents and between the PBDE concentrations and the trophic levels. The level of significance was set at α = 0.05 throughout the present study. 3. Results and discussion 3.1. PBDEs concentrations in organisms Of the 10 PBDE congeners measured, BDE28, 47, 66,100, 99,154, and 153 were detected in more than 70% of samples but BDE138, 183, and 209 were detected in less than 60% of samples (59%, 35%, and 18%, respectively) (Table 1). Therefore, data for ∑PBDEs only contained the congeners of BDE28, 47, 66, 100, 99, 154, and 153. The ∑PBDE concentrations in biota from the Pearl River Estuary were summarized in Table 1. Bombay duck, ark shell, and large yellow croaker had relatively low PBDE concentrations with median of 9.8, 13, and 13 ng/g lipid weight (lw), respectively. Sand swimming crab, samoan crab, oncomelania, common mullet, red eelgoby, and silver sillago had similar PBDE levels with median ranging from 17 to 21 ng/g lw, which were lower than those (median ranged from 29 to 32 ng/g lw) in robust tonguefish, slimy spinefoot and Japanese eel. The highest concentration was found in flathead fish with a median of 59 ng/g lw. The levels of PBDEs in the present study were significantly lower than those (median ranged from 67 to 194 ng/g lw) in biota samples collected in the same area in 2004, in which the same muscle tissues were analyzed in several same species (Xiang et al., 2007), indicating a decrease trend for PBDE levels in biota in the study area. In the present study, samples of robust tonguefish, Japanese eel, and flathead fish were annually collected in 2005, 2006 and 2007. The concentrations of ∑PBDEs in three years were illustrated in Fig. 1. It was clear that a decreasing trend of ∑PBDEs in biota was observed although no statistical significance was obtained, which confirmed the above conclusion. BDE47 was the most abundant congener in aquatic species and was usually chosen as a representative of PBDE contaminants. In the present study, the concentrations of BDE47 in biota ranged from 2.3 to 38 ng/g lw, which were close to the concentrations in
Table 2 Slope and p-value of regression analysis between logarithm of concentration and trophic levels, and TMFs for PBDE. Compound
Slope
R
TMF
p
BDE28 BDE47 BDE66 BDE100 BDE99 BDE154 BDE153 Sum BDEs
− 0.11 0.36 0.27 0.47 0.42 0.44 0.37 0.36
0.16 0.42 0.29 0.46 0.31 0.35 0.31 0.43
0.78 2.29 1.86 2.95 2.63 2.75 2.34 2.04
0.090 b0.001 b0.001 b0.001 b0.001 b0.001 b0.001 b0.001
M. Yu et al. / Environment International 35 (2009) 1090–1095 fishes from Lake Winnipeg, Canada (1.8–84 ng/g lw) (Law et al., 2006), from southern Greenland (7.9–41 ng/g lw) (Christensen et al., 2002), and from Canadian Arctic (2.9– 26 ng/g lw). The concentrations of BDE47 in the present study were much lower than those reported concentrations in Laizhou Bay (30–240 ng/g lw) (Jin et al., 2008), and coastal British Columbia (6.1–160 ng/g lw) (Ikonomou et al., 2002) but higher than those from Vietnam Can Tho (0.4–0.7 ng/g lw) (Minh et al., 2006), and Bohai Bay (0.03–3.8 ng/g lw) (Wan et al., 2008).
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3.2. Trophic magnification factors (TMFs) The structure of the Pearl River Estuary food web was elucidated using stable isotopes of nitrogen. Stable isotope values for the biota were shown in Table 1. The average δ15N for zooplankton was 9.7‰, and was assumed as primary consumer, i.e., TL = 2.0, because of their herbivorous feeding on primary producers (photoplankton) (Post, 2002). Trophic level differed among organisms, with two distinct groups
Fig. 2. Relationships between concentrations of PBDE congeners (ng/g lw) and trophic levels of biota.
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apparent. Sand swimming crab, oncomelania, and common mullet occupied a relative lower trophic position (range in means TL of 2.4–2.6) than all other species (mean TL of 3.4–3.8) (Table 1). Regression analysis was conducted between the lipid-normalized concentrations of PBDEs (log-transformed) and the trophic levels. Table 2 and Fig. 2 illustrate that strong positive relationships (TMF N 1, p b 0.01) between chemical concentration (log C) and trophic level were observed for all PBDE congeners except for BDE28, which showed a negative relationship. A TMF value of 2.0 was observed for ∑PBDEs concentration. TMFs for individual PBDE congeners ranged between 0.78 and 3.0 (Table 2). With exception of BDE28, all TMFs were significantly larger than one, indicating their biomagnification in the food web. This result was contrary to that reported in a previous study conducted by Guo et al. (2008) in the same study area. In that study, a total of 50 samples from 34 species (23 samples from the PRE and 27 samples from the Daya Bay) were collected. No significant correlations between concentrations of PBDEs (normalized to lipid content) and δ15N values were found. Bioconcentration, rather than biomagnification, was suggested as major accumulation mechanisms of PBDEs in fish in the Pearl River Estuary. Such discrepancy may be due to the difference in tropic levels occupied by organisms between the two studies. In Guo's study, the mean δ15N value of fish was 11‰ which was much lower than that (mean 16‰) in the present study. Additionally, relatively small number of biota samples (50 samples for 34 species, the averaged sample number for one species was less than 2) may be another reason for the absence of biomagnifications of PBDEs in Guo's investigation. The observed biomagnifications for most BDE congeners were also reported in the food webs from Bohai Bay, North China and from Baltic Sea and northern Atlantic Ocean (Wan et al., 2008; Burreau et al., 2006). The TMF values of seven individual congeners (BDE28, 71, 47, 66, 100, 119 and 99) and total PBDEs ranged from 1.56 (BDE99) to 7.24 (BDE47) in the food web from Bohai Bay, North China, and the values significantly larger than one were obtained for BDE28 (3.57, p = 0.024), BDE47 (7.24, p = 0.006), BDE100 (3.23, p = 0.049), BDE119 (2.6, p = 0.042) and total PBDEs (3.53, p = 0.046). The TMF values in the present study were lower than those in the food web from Bohai Bay, North China, which may be due to the different food web structure. The food web from Bohai Bay included zooplankton and birds as well as invertebrates and fish. Several previous studies demonstrated that TMFs of hydrophobic organics, such as HCB, HCH, p, p′-DDE and PCBs, estimated based on a food web including only poikilotherms (invertebrates, fish) were usually lower than those estimated based on a food web including both poikilotherms and homoetherms (sea-birds and mammals) (Fisk et al., 2001; Hop et al., 2002; Wan et al., 2008). In the food webs from the Baltic Sea and the northern Atlantic Ocean, Burreau et al. (2006) also found that PBDEs with up to 6–7 bromine atoms biomagnified. However, no biomagnification was found for all PBDE congeners (BDE28, 66, 99, 100, 118, 153, and 154) but for BDE47 in a Canadian arctic marine food web in a study conducted by Kelly et al. (2008). The authors suggested that biotransformation of PBDEs was the main reason for no observed biomagnification. Wu et al. (2009) found biomagnifications for BDE47, 100, and 154, but trophic dilution for BDE66, 99, 153, 183, and 209 in a highly contaminated freshwater food web from South China. The enhanced metabolic capacity induced by the high level of PBDE in organisms and the high temperature in water was responsible for this observation. Correlation between TMFs and log KOW is expected because trophic level is dietbased and biomagnification from diet correlates with KOW (Campfens and MacKay, 1997). In the present study, a line positive correlation was found between TMF and log KOW (Fig. 3), which was consistent with the expectation. However, it should be borne in mind that only seven BDE congeners were considered in the present study. The limited BDE congener numbers used in the present study might mask the realistic relationship between TMF and log KOW. A parabolic relationship between biomagnification power (B) of BDE congener and log KOW was observed in a food web from the Baltic Sea (Burreau et al., 2004). The highest biomagnification potential was found in BDE99 then biomagnification power decreased with increasing log KOW. In the present study, a
Fig. 4. Box plots of the ratios of BDE99 to BDE100 in different biota species and water (unpublished data). The five lines from bottom to upper represent 5th, 25th, 50th, 75th, and 95th percentiles, respectively. Solid circles indicate average values. Asterisks above or below the boxes represent the maximum and the minimum. SS: sand swimming crab (Ovalipes punctatus); ON: oncomelania (Oncomelania hupensischiui); CM: common mullet (Mugil cephalus); Other species contain ark shell (Tegillarca granosa), red eelgoby (Odontamblyopus rubicundus), robust tonguefish (Cynoglossus robustus), slimy spinefoot (Siganus canaliculatus), silver sillago (Sillago sihama), pompano (Psenopsis anomala), Japanese eel (Anguilla japonica), flathead fish (Platycephalus indicus), large yellow croaker (Pseudosiaena crocea), and bombay duck (Harpodon nehereus). similar trend could also be found if taking into account of BDE138 (TMF = 1.7, p = 0.15) and BDE183 (TMF = 1.8, p = 0.26), in which BDE100 had the highest TMF then TMF decreased with increasing degree of bromination. Similar parabolic relationship between TMF and log KOW has also been reported for PCBs in marine (Borgå et al., 2004; Kelly et al., 2008) and freshwater (Walters et al., 2008; Wu et al., 2009) food webs. This result suggested that the bioaccumulation behavior of PBDEs may be similar to that of PCBs.
3.3. Bioaccumulation behavior of PBDEs Congener-specific metabolism of PBDEs in the food web components undoubtedly plays a key role in the bioaccumulation of PBDEs (Kelly et al., 2008; Wu et al., 2009). Previous studies have indeed shown that biotransformation of PBDEs occurred in some fish. For example, Stapleton et al. (2004) observed in vivo debromination of BDE99 to BDE47, and BDE183 to 154 in common carp. Voorspoels et al. (2003) suggested that the ratio of BDE99 to BDE100 could be related to the difference in metabolism in aquatic organisms, which have been further confirmed by the study of Xiang et al. (2007). In the present food web, the TMF value of BDE100 was higher than that of BDE99, which was consistent with the results of previous studies in aquatic ecosystem although both of the chemicals had similar KOW(Wu et al., 2009; Wan et al., 2008; Law et al., 2006). The ratios of BDE99 to BDE100 in biota and water which supported the habitation of the organisms were investigated to take an insight to the potential influence of biotransformation on biomagnification. Water samples were collected at the same time as that of organisms. As shown in Fig. 4, the concentration ratios of BDE99 to BDE100 in water samples (median of 5.3, unpublished data) were significantly higher than those in biota samples (median of 0.5), and the ratios in low tropic level organisms such as sand swimming crab, oncomelania, and common mullet were higher than those in other high tropic level organisms. This result implied that organisms with higher tropic levels had increased metabolism ability, which was in line with the results from other studies (Wan et al., 2008; Wu et al., 2009). This result can also be used to explain why the TMF of BDE99 was usually lower than that of BDE100 in aquatic ecosystem. Thus, the extent of biotransformation in food web should be considered to elucidate the bioaccumulation behavior of PBDEs.
Table 3 Slope and p-value of regression analysis between logarithm of concentration and logarithm of lipid percent lipid contents in organisms.
Fig. 3. Relationship between TMFs and log KOW for PBDEs. Log KOW values were taken for PBDEs.
Compound
Slope
R
p
BDE28 BDE47 BDE66 BDE100 BDE99 BDE154 BDE153 Sum BDEs
1.2 0.89 0.97 0.71 1.1 0.67 0.63 0.89
0.61 0.57 0.61 0.46 0.53 0.39 0.61 0.61
b0.001 b0.001 b0.001 b0.001 b0.001 b0.001 b0.001 b0.001
M. Yu et al. / Environment International 35 (2009) 1090–1095 In cold water lakes, fish with higher concentrations of PBDEs tend to be the fattiest within a species and in a food web (Law et al., 2006). There was a significantly consistent relationship between log PBDEs (ng/g wet weight) and log percent lipid in the organisms from the Pearl River Estuary (r = 0.39–0.61, p b 0.001) (Table 3), and the slopes of the regression lines ranged from 0.63 to 1.2. These values were slightly higher than the results of the previous study in the Pearl River Estuary (The slope of the regression line for total PBDEs was 0.65.) (Guo et al., 2008). The correlations between the PBDE concentrations and lipid contents maybe resulted from either bioconcentration of the chemicals from environmental exposure and/or biomagnification via food web (Bentzen et al., 1996; Mbongwe et al., 2003). In the present study, no correlation was found between the trophic level of organisms and the percent of lipid. This implied that the biomagnification of PBDEs observed in the present study was not the result of lipid content effect. Therefore, we can conclude that biomagnifications of PBDEs via food web indeed occurred in the present food web, but the bioconcentration cannot be ruled out as another important bioaccumulation mechanism of PBDEs in biota in the studied food web.
4. Conclusions PBDE concentrations and the trophic levels were measured in fourteen species of biota from the Pearl River Estuary. The results of the present study clearly indicated that BDE congeners (BDE47, 66, 100, 99, 154, and 153) exhibited biomagnification potential in the food web. On the other hand, bioconcentration that biota accumulated PBDEs directly from the environment cannot be neglected as an important route of bioaccumulation of PBDEs in biota. The significant difference of ratios of BDE99 to BDE100 between water and biota samples implied that congener-specific biotransformation occurred and influenced the observed TMF of the food web. Acknowledgments This work was financially supported by the National Science Foundation of China (20890112, 40525012, 40821003), and the National Basic Research Program of China (No.2009CB421604). This is contribution No. is-1098 from GIGCAS. We thank Mr. T.S. Xiang for assistance in GC/MS. References Bentzen E, Lean DRS, Taylor WD, Mackay D. Role of food web structure on lipid and bioaccumulation on organic contaminants by lake trout (Salvelinus namaycush). Can J Fish Aquat Sci 1996;53:2397–407. Borgå K, Fisk AT, Hoekstra PF, Muir DCG. Biological and chemical factors of importance in the bioaccumulation and trophic transfer of persistent organochlorine contaminants in Arctic marine food webs. Environ Toxicol Chem 2004;23:2367–85. Burreau S, Zebuhr Y, Broman D, Ishaq R. Biomagnification of polychlorinated biphenyls (PCBs) and polybrominated diphenyl ethers (PBDEs) studied in pike (Esox lucius), perch (Perca fluviatilis) and roach (Rutilus rutilus) from the Baltic Sea. Chemosphere 2004;55:1043–52. Burreau S, Zebuhr Y, Broman D, Ishaq R. Biomagnification of PBDEs and PCBs in food webs from the Baltic Sea and the northern Atlantic Ocean. Sci Total Environ 2006;366:659–72. Campfens J, MacKay D. Fugacity-based model of PCB bioaccumulation in complex aquatic food webs. Environ Sci Technol 1997;31:577–83. Chen SJ, Luo XJ, Lin Z, Luo Y, Li KC, Peng XZ, et al. Time trends of polybrominated diphenyl ethers in sediment cores from the Pear River Estuary, South China. Environ Sci Technol 2007;41:5595–600. Christensen JH, Glasius M, Pecseli M, Platz J, Pritzl G. Polybrominated diphenyl ethers (PBDEs) in marine fish and blue mussels from southern Greenland. Chemosphere 2002;47:631–8. Darnerud PO, Eriksen GS, Johannesson T, Larsen PB, Viluksela M. Polybrominated diphenyl ethers: occurrence, dietary exposure, and toxicology. Environ Health Perspect 2001;109(Suppl 1):49–68. Fisk AT, Hobson KA, Norstrom RJ. Influence of chemical and biological factors on trophic transfer of persistent organic pollutants in the Northwater Polynya marine food web. Environ Sci Technol 2001;35:732–8. Guo LL, Qiu YW, Zhang G, Zheng GJ, Lam PKS, Li XD. Levels and bioaccumulation of organochlorine pesticides (OCPs) and polybrominated diphenyl ethers (PBDEs) in
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Environment International j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / e n v i n t
Bioaccumulation and trophic transfer of polybrominated diphenyl ethers (PBDEs) in biota from the Pearl River Estuary, South China Mei Yu a,b, Xiao-Jun Luo a,⁎, Jiang-Ping Wu a,b, She-Jun Chen a, Bi-Xian Mai a a b
State Key Laboratory of Organic Geochemistry, Guangzhou Institute of Geochemistry, Chinese Academy of Science, Guangzhou 510640, China Graduate School of Chinese Academy of Science, Beijing, 100039, China
a r t i c l e
i n f o
Article history: Received 16 March 2009 Accepted 12 June 2009 Available online 17 July 2009 Keywords: Polybrominated diphenyl ethers (PBDEs) Biomagnification Bioaccumulation Pearl River Estuary
a b s t r a c t Two hundred and fifty-four biota samples (four species of invertebrates and ten species of fish) were collected from the Pearl River Estuary between 2005 and 2007 and one hundred and twenty four individual or composite samples were analyzed for polybrominated diphenyl ethers (PBDEs). The concentrations of PBDEs in organisms varied from 6.2 to 208 ng/g lipid weight. This PBDE level was significantly lower than those collected in 2004, showing a decreasing trend of PBDEs in biota in the study area. Trophic magnification factors (TMFs) for nine BDE congeners were calculated with values ranging from 0.78 to 3.0. TMFs of BDE47, 66, 100, 99, 154, and 153 were statistically greater than one, indicating a biomagnifcation potential for these congeners. Significant positive correlations were also found between concentrations of the total PBDEs, BDE28, 47, 66, 100, 99, 154, and153 and lipid content in biota, indicating the that bioconcentration also played an important role in the accumulation of PBDEs. No correlation between trophic level and lipid content was found, suggesting that biomagnification was not the result of lipid content effect but indeed occurred. The concentration ratios of BDE99 to BDE100 were much lower in biota than that in water implying that potential congener-specific biotransformation of PBDEs occurred and influenced the biomagnification of BDE congeners. © 2009 Elsevier Ltd. All rights reserved.
1. Introduction Polybrominated diphenyl ethers (PBDEs) are flame retardants widely used in plastic, textile, electronic and other material. Because of their ubiquity in the environment (Law et al., 2008; Schecter et al., 2003; Vonderheide et al., 2008) and their exponentially increasing concentrations in most environment compartments with doubling times of about 4–6 years (Chen et al., 2007; Hites, 2004), the behaviors of PBDEs in environment have attracted great attention and interests from the public and environmental community. Available toxicological evidences indicate that PBDEs can disturb thyroid homeostasis, cause hepatomegaly and neurobehavioral deficits, and exhibit fetal and maternal toxicity after prolonged exposure (Darnerud et al., 2001; McDonald, 2002). PBDEs are capable of bioaccumulation in the biota own to their properties of high lipophilicity (log KOW: 5.9–10) and resistance to metabolism (Gustafsson et al., 1999). Recently, a few of papers reported that the bioaccumulation and biomagnifications of PBDEs in biota, mostly focused on the marine and fresh food webs, and the conclusions varied. For example, Wan et al. (2008) have reported that concentrations of BDE28, 47, 100 and 119 increased significantly with the increasing trophic levels in a marine food web of Bohai Bay. In the tropical– ⁎ Corresponding author. Tel.: +86 20 85290146; fax: +86 20 85290706. E-mail address: [email protected] (X.-J. Luo). 0160-4120/$ – see front matter © 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.envint.2009.06.007
subtropical waters, BDE47, 100 and 154 were found biomagnifications but BDE66, 99 and 209 were found trophic dilution in a fresh water food web from a contaminated reservoir in South China (Wu et al., 2009). While analyses of PBDEs in biota from a fresh water food web from Lake Winnipeg indicated that only BDE47 and BDE209 have TMF greater than one (Law et al., 2006). Burreau et al. (2006) found the absence of biomagnification for the highly brominated diphenyl ethers such as BDE209 and nona-BDEs and the slight biomagnification for BDE47 (TMF = 1.5, calculated from B value) in the northern Atlantic Ocean food web. On the other hand, studies conducted by Guo et al. (2008) and Wang et al. (2007) found the absence of biomagnification for PBDEs in the organisms from the Pearl River Estuary and from a river receiving effluent discharged from a large sewage treatment, respectively. These results indicated that the bioaccumulation behavior of PBDEs in biota was complex and needs more field research. The Pear River Estuary is located in the Pearl River Delta region of southern China, one of the most economically developed regions in China. Many researches have evidenced the Pear River Estuary as a significant sink for organic pollutants including PBDEs from upstream runoff, which posed risks to the organisms in the region (Mai et al., 2005; Luo et al., 2008; Wurl et al., 2006). Guo et al. (2008) have carried out a preliminary research about accumulation of PBDEs in biota in this area and found the absence of biomagnification of PBDEs in the food web. The limited samples number (50 samples for 34 species, average less than 2 samples for each species) and short food
BD (n = 6)
1091 SS, sand swimming crab (Ovalipes punctatus); SC, samoan crab (Scylla serrata); AS, ark shell (Tegillarca granosa); ON, oncomelania (Oncomelania hupensischiui); CM, common mullet (Mugil cephalus); RE, red eelgoby (Odontamblyopus rubicundus); RT, robust tonguefish (Cynoglossus robustus); SP, slimy spinefoot (Siganus canaliculatus); SL, silver sillago (Sillago sihama); PP, pompano (Psenopsis anomala); JE, Japanese eel (Anguilla japonica); FF, flathead fish (Platycephalus indicus); LY, large yellow croaker (Pseudosiaena crocea); BD, bombay duck (Harpodon nehereus). nd, not detected. a Average ± standard derive. b Median (min–max).
0.84 ± 0.34 15 ± 1.2 3.5 ± 0.31 nd (nd–0.8) 7.2 (2.3–18) 0.67 (0.2–1.1) 1.9 (1.1–10) 1.0 (0.17–4.8) 0.97 (0.38–7.4) 1.3 (0.13–16) 13 (6.3–54) 0.59 (nd–4.1) 0.7 (nd–6) nd (nd–0.66)
LY (n = 11) FF (n = 27)
0.77 ± 0.39 16 ± 1.2 3.6 ± 0.32 1.5 (nd–5.3) 38 (8.3–115) 1.3 (0.25–6.9) 9.4 (1.4–50) 0.59 (nd–18) 4.7 (0.54–30) 1.3 (nd–6.9) 59 (12–208) 0.36 (nd–5.5) nd (nd–53) nd (nd–0.31) 2.2 ± 0.91 16 ± 1.2 3.6 ± 0.31 1.1 (nd–5) 15 (5.0–39) 0.76 (0.3–3.4) 5.7 (2.2–12) 2.9 (0.62–8) 2.6 (0.99–4.7) 1.6 (0.72–4.7) 30 (14–71) 0.29 (nd–2.4) nd (nd–24) nd (nd–0.25)
JE (n = 19) PP (n = 6)
6.5 ± 1.4 17 ± 1.4 3.8 ± 0.38 1.2 (0.71–1.38) 28 (21–35) 1.3 (0.74–1.6) 5.5 (3.8–6.8) 4.7 (3.1–9) 2.2 (0.82–4.1) 1.6 (0.52–2.7) 44 (39–53) 0.09 (nd–0.58) 0.11 (nd–0.4) nd 0.87 ± 0.17 15 ± 1.2 3.4 ± 0.30 0.45 (nd–2) 13 (5.4–23) 0.45 (0.18–1.2) 1.1 (0.61–6) 0.45 (nd–1.3) 0.57 (0.36–2.9) 0.55 (0.17–4.1) 20 (7.8–37) 0.21 (nd–1) nd (nd–21) nd
SL (n = 5) SP (n = 12)
1.3 ± 0.44 16 ± 1.2 3.6 ± 0.31 0.52 (nd–2.7) 17 (6.4–47) 1.3 (0.61–3.9) 1.8 (0.73–5.4) 2.1 (0.51–14) 1.4 (0.36–2.9) 2.3 (0.94–5.7) 29 (14–73) 0.07 (nd–0.58) 4.5 (nd–69) 0.004 (nd–1.6) 0.97 ± 2.7 15 ± 1.4 3.4 ± 0.36 0.18 (nd–5.9) 17 (4.1–28) 0.71 (0.12–6.1) 4 (0.81–10) 3.5 (0.08–15) 2.9 (0.66–12) 2.4 (0.29–5) 32 (6.2–81) 1 (nd–6.7) nd (nd–0.06) nd
RT (n = 17) RE (n = 3)
0.68 ± 0.16 16 ± 0.35 3.6 ± 0.09 nd 7.3 (3.2–9.7) 0.4 (0.23–1.2) 2.1 (0.78–3.9) 0.69 (0.53–1.7) 1.4 (1.2–3) 4.1 (2.9–5.3) 19 (10–21) nd 22 (1.6–30) nd 4.9 ± 2.7 12 ± 1.4 2.6 ± 0.36 1.3 (nd–2.4) 14 (0.57–22) 0.64 (0.03–1.1) 1.6 (0.12–2.3) 0.25 (nd–1.4) 0.80 (0.08–2) 0.48 (0.22–5) 21 (1.3–29) 0.06 (nd–2.1) 0.09 (nd–31) nd (nd–1.6)
CM (n = 7) ON (n = 3) AS (n = 3) SC(n = 2) SS (n = 3)
Before the further treatment, the length and weight were recorded for each individual fish. Samples were thawed and dissected to remove skin and bones. Muscle fillets below the dorsal fin for fish and the soft part (edible part) for invertebrates were taken. The procedure for biota sample extraction and cleanup was described in detail in our previous studies (Xiang et al., 2007; Wu et al., 2008). Briefly, after being homogenized with ashed anhydrous sodium sulfate and spiked with surrogate standards (CDE99 and 13C12-PCB141), the samples were Soxhlet extracted with hexane/acetone (1:1, v/v) for 48 h. The extracts were concentrated and an aliquot of the extract was used to lipid content determination by gravimetric method, the other aliquot was subjected to gel permeation chromatography (GPC) to remove lipids. The cleaned extract was concentrated to approximately 1 mL and further purified passing through a multilayer silica/alumina column. The extracts were
2.5 ± 0.5 11 ± 0.3 2.4 ± 0.1 3.3 (3.2–5.1) 13 (12–18) 1.8 (1.7–2.3) 0.26 (0.26–0.34) 1 (1–1.3) 0.07 (0.06–0.07) 0.04 (0.03–0.07) 19 (18–27) nd nd nd
2.3. Extraction and cleanup
Lipid (%)a δ15N Trophic level BDE28b BDE47 BDE66 BDE100 BDE99 BDE154 BDE153 ∑(7 PBDEs) BDE138 BDE183 BDE209
All PBDE standards were purchased from Accustandards Inc. (New Haven, CT). CDE99 (2, 2′, 4, 4′, 5-pentachlorodiphenyl ether) was obtained from Wellington Laboratories (Ontario, Canada). 13C12PCB141 and 13C12-PCB208 were obtained from Cambridge Isotope Laboratories (Andover, MA, USA). All organic solvents were analytical reagent and redistilled using a glass system.
Species
2.2. Reagents and chemicals
Table 1 Biological parameters and PBDEs concentrations (ng/g lw) of 14 marine organisms collected from the Pearl River Estuary.
Organisms were collected using fish trawls in the Pearl River Estuary during August 2005 and August 2007. The detailed information about the sampling area has been given elsewhere (Xiang et al., 2007). All samples were stored in cool boxes and transported to the laboratory and then immediately transferred to a freezer where they were stored at −20 °C until analysis. The biota species collected in the present study included sand swimming crab (Ovalipes punctatus, three composite samples from nine individuals), samoan crab (Scylla serrata, two composite samples form six individuals), ark shell (Tegillarca granosa, three composite samples from sixty individuals), oncomelania (Oncomelania hupensischiui, three composite samples from thirty individuals), common mullet (Mugil cephalus, seven individual samples), red eelgoby (Odontamblyopus rubicundus, three composite samples from fifteen individuals), robust tonguefish (Cynoglossus robustus, seventeen individual samples), slimy spinefoot (Siganus canaliculatus, twelve individual samples), silver sillago (Sillago sihama, five individual samples), pompano (Psenopsis anomala, six composite samples from thirty individuals), Japanese eel (Anguilla japonica, nineteen individual samples), flathead fish (Platycephalus indicus, twenty seven individual samples), large yellow croaker (Pseudosiaena crocea, eleven individual samples), and bombay duck (Harpodon nehereus, six individual samples). Detailed information on the samples was provided in Table 1. Zooplankton samples were also collected using a tow net of 160 μm in August 2007, and the trawling depth ranged from 0 to 3 m. Two composite zooplankton samples were obtained and stored at −20 °C and further for nitrogen isotope measurements. A total of one hundred and twenty four samples were analyzed for PBDEs in this study.
1.7 ± 0.39 12 ± 0.05 2.5 ± 0.01 nd 2.3 (1.9–3.4) 0.13 (0.13–0.52) 2.9 (2.5–3.7) 6.5 (5.5–8.4) 2.6 (1.7–2.8) 3.1 (2.3–3.9) 17 (15–22) 1.6 (0.41–2.1) 1.5 (0.33–42) nd
2.1. Sampling
1.7 ± 0.64 15 ± 0.55 3.4 ± 0.14 0.5 (0.4–1.2) 6.89 (3.9–9.1) 0.86 (0.65–0.87 1.4 (0.79–1.4) 0.79 (0.51–1.4) 0.3 (0.07–0.48) 0.1 (0.08–0.51) 13 (6.4–13) 0.26 (nd–1.1) nd (nd–0.47) nd (nd–0.54)
2. Materials and methods
0.76 ± 0.07 16 ± 0.29 3.7 ± 0.08 1.2 (0.2–2.2) 13 (9.8–16) 0.59 (0.49–0.69 2 (0.61–3.3) 0.69 (0.63–0.76) 1.10 (17–2.1) 3.1 (0.36–5.7) 21 (14–29) 1.4 (nd–2.8) 16 (nd–32) nd
web may mask the biomagnification of PBDEs in the food web. In the present study, we analyzed concentrations of 10 BDE congeners in the food web (including four invertebrate species and ten fish species) and their relationship with δ15N and lipid content to understand their accumulation behaviors in the food web. We also compared the concentration ratios of BDE99 to BDE100 between biota and water samples to explore the potential congener-specific biotransformation of PBDEs in biota sample and its influence on the biomagnifications of BDE congeners.
0.91 ± 0.35 16 ± 0.93 3.5 ± 0.24 nd (nd–3.1) 5.9 (3.3–15) 0.30 (nd–2.6) 1 (0.49–3) 0.89 (nd–9.5) 0.35 (0.13–0.88) 0.8 (0.02–1.8) 9.8 (5–34) 0.8 (nd–3.8) nd (nd–1.8) nd (nd–1)
M. Yu et al. / Environment International 35 (2009) 1090–1095
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deviation was less than 0.6%. The isotope ratio was standardized against air according to δ15N = (Rsample /Rair − 1)× 1000‰, where R is the isotope ratio of 15N to 14N. The δ15N values were calibrated against a standard, ammonium sulfate (IAEA-N-1, International Atomatic Energy Agency Analytical Quality Control Services, Wien, Austria). The precision of the analytical method and instrument was ± 0.3‰. Trophic level (TL) was calculated for individual sample using the following equation (Post, 2002): 15
15
TL consumer = ½ðδ Nconsumer δ Nprimary
consumer Þ = 3:8
ð1Þ
+2
where 3.8 is the isotopic trophic enrichment factor. TMFs were calculated according to Tomy et al. (2004), and the references therein using following equations: Log concentration ðlipid normalizedÞ = A + B TL Fig. 1. Box plot concentrations of ∑PBDE in biota from the Pearl River Estuary among 2005 and 2007. The five lines from bottom to upper represent 5th, 25th, 50th, 75th, and 95th percentiles, respectively. Solid circles indicate average values. Asterisks above or below the boxes represent the maximum and the minimum.
concentrated, solvent exchanged to hexane, and finally concentrated to 50 μL under a gentle stream of nitrogen. A known amount of internal standard (BDE118, BDE128 and 13C12-PCB208) was added to all extracts prior to instrumental analysis. 2.4. Instrumental analysis Polybrominated diphenyl ethers (PBDEs) were measured with a Shimadzu model 2010 gas chromatograph coupled with a model QP2010 mass spectrometer (Shimadzu, Japan) using electron capture negative ionization (ECNI) in the selective ion monitoring (SIM) mode. A DB-XLB capillary column (30 m × 0.25 mm i.d. × 0.25 μm film thickness) was used to determine the tri- to hepta-BDEs (BDE28, 47, 66, 100, 99, 85, 153, 154, 138, and 183). For deca-BDEs (BDE209), a CPSil 13 CB capillary column (12.5 m × 0.25 mm i.d. × 0.20 μm film thickness) was used. Details of the GC temperature program as well as the procedure for qualification and quantification of PBDEs were given in published literature (Mai et al., 2005). The limit of detection (LOD), defined as a signal/noise ratio (S/N) = 5, ranged from 0.001 to 0.38 ng/g lipid weight. BDE85 was not quantified in all the samples because of chromatographic interferences. 2.5. Quality assurance/quality control Quality assurance was done by analyses of procedural blanks, triplicate spiked blanks, and triplicate spiked matrices. For each batch of 12 samples, a procedural blank was processed. Some of the procedural blanks (n = 15) contained traces of target chemicals, but the levels were close to the limit of quantification and they were not subtracted from those in samples. The mean recoveries of individual congeners (BDE28, 47, 66, 100, 99, 153, 154, 138, and 183) ranged from 61% to 87% with relative standard deviations (RSDs) less than 10% in triplicate spiked blanks and from 61% to 82% with RSDs less than 15% in triplicate spiked matrices, respectively. The surrogate standard recoveries of CDE99 and 13 C12-PCB141 were 83% ± 13% (ranged from 46%–125%) and 86% ± 16% (51%–118%), respectively. No surrogate corrections were made to final reported concentrations. 2.6. Nitrogen isotope measurements About 1 mg of homogenized and finely ground sample was wrapped with a tin cup, and placed in the sample container of an elemental analyzer-isotope ratio mass spectrometer (CE flash EA1112-Finnigan Delta plus XL) for stable nitrogen isotope ratio measurements. Two replicates of each sample were analyzed and the relative standard
ð2Þ
B
ð3Þ
TMF = 10
statistical significance of the regression Eq. (2) was defined at p b 0.05. 2.7. Statistical analysis For samples with contaminant concentration below LOD, zero was used for the calculations. All data were lipid-normalized. BDE85 was not quantified due to chromatographic interference in some of the samples. Simple linear regressions and correlation coefficients were used to examine correlation between the levels of PBDEs and lipid contents and between the PBDE concentrations and the trophic levels. The level of significance was set at α = 0.05 throughout the present study. 3. Results and discussion 3.1. PBDEs concentrations in organisms Of the 10 PBDE congeners measured, BDE28, 47, 66,100, 99,154, and 153 were detected in more than 70% of samples but BDE138, 183, and 209 were detected in less than 60% of samples (59%, 35%, and 18%, respectively) (Table 1). Therefore, data for ∑PBDEs only contained the congeners of BDE28, 47, 66, 100, 99, 154, and 153. The ∑PBDE concentrations in biota from the Pearl River Estuary were summarized in Table 1. Bombay duck, ark shell, and large yellow croaker had relatively low PBDE concentrations with median of 9.8, 13, and 13 ng/g lipid weight (lw), respectively. Sand swimming crab, samoan crab, oncomelania, common mullet, red eelgoby, and silver sillago had similar PBDE levels with median ranging from 17 to 21 ng/g lw, which were lower than those (median ranged from 29 to 32 ng/g lw) in robust tonguefish, slimy spinefoot and Japanese eel. The highest concentration was found in flathead fish with a median of 59 ng/g lw. The levels of PBDEs in the present study were significantly lower than those (median ranged from 67 to 194 ng/g lw) in biota samples collected in the same area in 2004, in which the same muscle tissues were analyzed in several same species (Xiang et al., 2007), indicating a decrease trend for PBDE levels in biota in the study area. In the present study, samples of robust tonguefish, Japanese eel, and flathead fish were annually collected in 2005, 2006 and 2007. The concentrations of ∑PBDEs in three years were illustrated in Fig. 1. It was clear that a decreasing trend of ∑PBDEs in biota was observed although no statistical significance was obtained, which confirmed the above conclusion. BDE47 was the most abundant congener in aquatic species and was usually chosen as a representative of PBDE contaminants. In the present study, the concentrations of BDE47 in biota ranged from 2.3 to 38 ng/g lw, which were close to the concentrations in
Table 2 Slope and p-value of regression analysis between logarithm of concentration and trophic levels, and TMFs for PBDE. Compound
Slope
R
TMF
p
BDE28 BDE47 BDE66 BDE100 BDE99 BDE154 BDE153 Sum BDEs
− 0.11 0.36 0.27 0.47 0.42 0.44 0.37 0.36
0.16 0.42 0.29 0.46 0.31 0.35 0.31 0.43
0.78 2.29 1.86 2.95 2.63 2.75 2.34 2.04
0.090 b0.001 b0.001 b0.001 b0.001 b0.001 b0.001 b0.001
M. Yu et al. / Environment International 35 (2009) 1090–1095 fishes from Lake Winnipeg, Canada (1.8–84 ng/g lw) (Law et al., 2006), from southern Greenland (7.9–41 ng/g lw) (Christensen et al., 2002), and from Canadian Arctic (2.9– 26 ng/g lw). The concentrations of BDE47 in the present study were much lower than those reported concentrations in Laizhou Bay (30–240 ng/g lw) (Jin et al., 2008), and coastal British Columbia (6.1–160 ng/g lw) (Ikonomou et al., 2002) but higher than those from Vietnam Can Tho (0.4–0.7 ng/g lw) (Minh et al., 2006), and Bohai Bay (0.03–3.8 ng/g lw) (Wan et al., 2008).
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3.2. Trophic magnification factors (TMFs) The structure of the Pearl River Estuary food web was elucidated using stable isotopes of nitrogen. Stable isotope values for the biota were shown in Table 1. The average δ15N for zooplankton was 9.7‰, and was assumed as primary consumer, i.e., TL = 2.0, because of their herbivorous feeding on primary producers (photoplankton) (Post, 2002). Trophic level differed among organisms, with two distinct groups
Fig. 2. Relationships between concentrations of PBDE congeners (ng/g lw) and trophic levels of biota.
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apparent. Sand swimming crab, oncomelania, and common mullet occupied a relative lower trophic position (range in means TL of 2.4–2.6) than all other species (mean TL of 3.4–3.8) (Table 1). Regression analysis was conducted between the lipid-normalized concentrations of PBDEs (log-transformed) and the trophic levels. Table 2 and Fig. 2 illustrate that strong positive relationships (TMF N 1, p b 0.01) between chemical concentration (log C) and trophic level were observed for all PBDE congeners except for BDE28, which showed a negative relationship. A TMF value of 2.0 was observed for ∑PBDEs concentration. TMFs for individual PBDE congeners ranged between 0.78 and 3.0 (Table 2). With exception of BDE28, all TMFs were significantly larger than one, indicating their biomagnification in the food web. This result was contrary to that reported in a previous study conducted by Guo et al. (2008) in the same study area. In that study, a total of 50 samples from 34 species (23 samples from the PRE and 27 samples from the Daya Bay) were collected. No significant correlations between concentrations of PBDEs (normalized to lipid content) and δ15N values were found. Bioconcentration, rather than biomagnification, was suggested as major accumulation mechanisms of PBDEs in fish in the Pearl River Estuary. Such discrepancy may be due to the difference in tropic levels occupied by organisms between the two studies. In Guo's study, the mean δ15N value of fish was 11‰ which was much lower than that (mean 16‰) in the present study. Additionally, relatively small number of biota samples (50 samples for 34 species, the averaged sample number for one species was less than 2) may be another reason for the absence of biomagnifications of PBDEs in Guo's investigation. The observed biomagnifications for most BDE congeners were also reported in the food webs from Bohai Bay, North China and from Baltic Sea and northern Atlantic Ocean (Wan et al., 2008; Burreau et al., 2006). The TMF values of seven individual congeners (BDE28, 71, 47, 66, 100, 119 and 99) and total PBDEs ranged from 1.56 (BDE99) to 7.24 (BDE47) in the food web from Bohai Bay, North China, and the values significantly larger than one were obtained for BDE28 (3.57, p = 0.024), BDE47 (7.24, p = 0.006), BDE100 (3.23, p = 0.049), BDE119 (2.6, p = 0.042) and total PBDEs (3.53, p = 0.046). The TMF values in the present study were lower than those in the food web from Bohai Bay, North China, which may be due to the different food web structure. The food web from Bohai Bay included zooplankton and birds as well as invertebrates and fish. Several previous studies demonstrated that TMFs of hydrophobic organics, such as HCB, HCH, p, p′-DDE and PCBs, estimated based on a food web including only poikilotherms (invertebrates, fish) were usually lower than those estimated based on a food web including both poikilotherms and homoetherms (sea-birds and mammals) (Fisk et al., 2001; Hop et al., 2002; Wan et al., 2008). In the food webs from the Baltic Sea and the northern Atlantic Ocean, Burreau et al. (2006) also found that PBDEs with up to 6–7 bromine atoms biomagnified. However, no biomagnification was found for all PBDE congeners (BDE28, 66, 99, 100, 118, 153, and 154) but for BDE47 in a Canadian arctic marine food web in a study conducted by Kelly et al. (2008). The authors suggested that biotransformation of PBDEs was the main reason for no observed biomagnification. Wu et al. (2009) found biomagnifications for BDE47, 100, and 154, but trophic dilution for BDE66, 99, 153, 183, and 209 in a highly contaminated freshwater food web from South China. The enhanced metabolic capacity induced by the high level of PBDE in organisms and the high temperature in water was responsible for this observation. Correlation between TMFs and log KOW is expected because trophic level is dietbased and biomagnification from diet correlates with KOW (Campfens and MacKay, 1997). In the present study, a line positive correlation was found between TMF and log KOW (Fig. 3), which was consistent with the expectation. However, it should be borne in mind that only seven BDE congeners were considered in the present study. The limited BDE congener numbers used in the present study might mask the realistic relationship between TMF and log KOW. A parabolic relationship between biomagnification power (B) of BDE congener and log KOW was observed in a food web from the Baltic Sea (Burreau et al., 2004). The highest biomagnification potential was found in BDE99 then biomagnification power decreased with increasing log KOW. In the present study, a
Fig. 4. Box plots of the ratios of BDE99 to BDE100 in different biota species and water (unpublished data). The five lines from bottom to upper represent 5th, 25th, 50th, 75th, and 95th percentiles, respectively. Solid circles indicate average values. Asterisks above or below the boxes represent the maximum and the minimum. SS: sand swimming crab (Ovalipes punctatus); ON: oncomelania (Oncomelania hupensischiui); CM: common mullet (Mugil cephalus); Other species contain ark shell (Tegillarca granosa), red eelgoby (Odontamblyopus rubicundus), robust tonguefish (Cynoglossus robustus), slimy spinefoot (Siganus canaliculatus), silver sillago (Sillago sihama), pompano (Psenopsis anomala), Japanese eel (Anguilla japonica), flathead fish (Platycephalus indicus), large yellow croaker (Pseudosiaena crocea), and bombay duck (Harpodon nehereus). similar trend could also be found if taking into account of BDE138 (TMF = 1.7, p = 0.15) and BDE183 (TMF = 1.8, p = 0.26), in which BDE100 had the highest TMF then TMF decreased with increasing degree of bromination. Similar parabolic relationship between TMF and log KOW has also been reported for PCBs in marine (Borgå et al., 2004; Kelly et al., 2008) and freshwater (Walters et al., 2008; Wu et al., 2009) food webs. This result suggested that the bioaccumulation behavior of PBDEs may be similar to that of PCBs.
3.3. Bioaccumulation behavior of PBDEs Congener-specific metabolism of PBDEs in the food web components undoubtedly plays a key role in the bioaccumulation of PBDEs (Kelly et al., 2008; Wu et al., 2009). Previous studies have indeed shown that biotransformation of PBDEs occurred in some fish. For example, Stapleton et al. (2004) observed in vivo debromination of BDE99 to BDE47, and BDE183 to 154 in common carp. Voorspoels et al. (2003) suggested that the ratio of BDE99 to BDE100 could be related to the difference in metabolism in aquatic organisms, which have been further confirmed by the study of Xiang et al. (2007). In the present food web, the TMF value of BDE100 was higher than that of BDE99, which was consistent with the results of previous studies in aquatic ecosystem although both of the chemicals had similar KOW(Wu et al., 2009; Wan et al., 2008; Law et al., 2006). The ratios of BDE99 to BDE100 in biota and water which supported the habitation of the organisms were investigated to take an insight to the potential influence of biotransformation on biomagnification. Water samples were collected at the same time as that of organisms. As shown in Fig. 4, the concentration ratios of BDE99 to BDE100 in water samples (median of 5.3, unpublished data) were significantly higher than those in biota samples (median of 0.5), and the ratios in low tropic level organisms such as sand swimming crab, oncomelania, and common mullet were higher than those in other high tropic level organisms. This result implied that organisms with higher tropic levels had increased metabolism ability, which was in line with the results from other studies (Wan et al., 2008; Wu et al., 2009). This result can also be used to explain why the TMF of BDE99 was usually lower than that of BDE100 in aquatic ecosystem. Thus, the extent of biotransformation in food web should be considered to elucidate the bioaccumulation behavior of PBDEs.
Table 3 Slope and p-value of regression analysis between logarithm of concentration and logarithm of lipid percent lipid contents in organisms.
Fig. 3. Relationship between TMFs and log KOW for PBDEs. Log KOW values were taken for PBDEs.
Compound
Slope
R
p
BDE28 BDE47 BDE66 BDE100 BDE99 BDE154 BDE153 Sum BDEs
1.2 0.89 0.97 0.71 1.1 0.67 0.63 0.89
0.61 0.57 0.61 0.46 0.53 0.39 0.61 0.61
b0.001 b0.001 b0.001 b0.001 b0.001 b0.001 b0.001 b0.001
M. Yu et al. / Environment International 35 (2009) 1090–1095 In cold water lakes, fish with higher concentrations of PBDEs tend to be the fattiest within a species and in a food web (Law et al., 2006). There was a significantly consistent relationship between log PBDEs (ng/g wet weight) and log percent lipid in the organisms from the Pearl River Estuary (r = 0.39–0.61, p b 0.001) (Table 3), and the slopes of the regression lines ranged from 0.63 to 1.2. These values were slightly higher than the results of the previous study in the Pearl River Estuary (The slope of the regression line for total PBDEs was 0.65.) (Guo et al., 2008). The correlations between the PBDE concentrations and lipid contents maybe resulted from either bioconcentration of the chemicals from environmental exposure and/or biomagnification via food web (Bentzen et al., 1996; Mbongwe et al., 2003). In the present study, no correlation was found between the trophic level of organisms and the percent of lipid. This implied that the biomagnification of PBDEs observed in the present study was not the result of lipid content effect. Therefore, we can conclude that biomagnifications of PBDEs via food web indeed occurred in the present food web, but the bioconcentration cannot be ruled out as another important bioaccumulation mechanism of PBDEs in biota in the studied food web.
4. Conclusions PBDE concentrations and the trophic levels were measured in fourteen species of biota from the Pearl River Estuary. The results of the present study clearly indicated that BDE congeners (BDE47, 66, 100, 99, 154, and 153) exhibited biomagnification potential in the food web. On the other hand, bioconcentration that biota accumulated PBDEs directly from the environment cannot be neglected as an important route of bioaccumulation of PBDEs in biota. The significant difference of ratios of BDE99 to BDE100 between water and biota samples implied that congener-specific biotransformation occurred and influenced the observed TMF of the food web. Acknowledgments This work was financially supported by the National Science Foundation of China (20890112, 40525012, 40821003), and the National Basic Research Program of China (No.2009CB421604). This is contribution No. is-1098 from GIGCAS. We thank Mr. T.S. Xiang for assistance in GC/MS. References Bentzen E, Lean DRS, Taylor WD, Mackay D. Role of food web structure on lipid and bioaccumulation on organic contaminants by lake trout (Salvelinus namaycush). Can J Fish Aquat Sci 1996;53:2397–407. Borgå K, Fisk AT, Hoekstra PF, Muir DCG. Biological and chemical factors of importance in the bioaccumulation and trophic transfer of persistent organochlorine contaminants in Arctic marine food webs. Environ Toxicol Chem 2004;23:2367–85. Burreau S, Zebuhr Y, Broman D, Ishaq R. Biomagnification of polychlorinated biphenyls (PCBs) and polybrominated diphenyl ethers (PBDEs) studied in pike (Esox lucius), perch (Perca fluviatilis) and roach (Rutilus rutilus) from the Baltic Sea. Chemosphere 2004;55:1043–52. Burreau S, Zebuhr Y, Broman D, Ishaq R. Biomagnification of PBDEs and PCBs in food webs from the Baltic Sea and the northern Atlantic Ocean. Sci Total Environ 2006;366:659–72. Campfens J, MacKay D. Fugacity-based model of PCB bioaccumulation in complex aquatic food webs. Environ Sci Technol 1997;31:577–83. Chen SJ, Luo XJ, Lin Z, Luo Y, Li KC, Peng XZ, et al. Time trends of polybrominated diphenyl ethers in sediment cores from the Pear River Estuary, South China. Environ Sci Technol 2007;41:5595–600. Christensen JH, Glasius M, Pecseli M, Platz J, Pritzl G. Polybrominated diphenyl ethers (PBDEs) in marine fish and blue mussels from southern Greenland. Chemosphere 2002;47:631–8. Darnerud PO, Eriksen GS, Johannesson T, Larsen PB, Viluksela M. Polybrominated diphenyl ethers: occurrence, dietary exposure, and toxicology. Environ Health Perspect 2001;109(Suppl 1):49–68. Fisk AT, Hobson KA, Norstrom RJ. Influence of chemical and biological factors on trophic transfer of persistent organic pollutants in the Northwater Polynya marine food web. Environ Sci Technol 2001;35:732–8. Guo LL, Qiu YW, Zhang G, Zheng GJ, Lam PKS, Li XD. Levels and bioaccumulation of organochlorine pesticides (OCPs) and polybrominated diphenyl ethers (PBDEs) in
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