Residues of Polybrominated Diphenyl Ethers in Frogs (Rana

Michael O. Gaylor , Greg L. Mears , Ellen Harvey , Mark J. La Guardia , and Robert C. Hale. Environmental Science & Technology 2014 48 (12), 7034-7043...
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Environ. Sci. Technol. 2009, 43, 5212–5217

Residues of Polybrominated Diphenyl Ethers in Frogs (Rana limnocharis) from a Contaminated Site, South China: Tissue Distribution, Biomagnification, and Maternal Transfer J I A N G - P I N G W U , †,‡ X I A O - J U N L U O , † YING ZHANG,† SHE-JUN CHEN,† B I - X I A N M A I , * ,† Y U N - T A O G U A N , ‡ A N D ZHONG-YI YANG§ State Key Laboratory of Organic Geochemistry, Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, China, Research Center for Environmental Engineering & Management, Graduate School at Shenzhen, Tsinghua University, Shenzhen 518055, China, and School of Life Sciences, Sun Yat-Sen University, Guangzhou 510275, China

Received April 10, 2009. Revised manuscript received May 25, 2009. Accepted May 29, 2009.

Environmental pollutants are suspected to be a cause of global declines in amphibian populations, but few data are available on the bioaccumulation of polybrominated diphenyl ethers (PBDEs) in amphibians. To examine the tissue distribution, biomagnification potential, and maternal transfer of PBDEs in frogs, eighteen PBDE congeners were measured in the muscle, liver, and egg tissues of rice frogs (Rana limnocharis) and insects collected from an electronic waste (e-waste) recycling site in South China. PBDE levels in the frogs ranged from 0.63 to 11.6, 4.57 to 56.2, and 10.7 to 125 ng/g wet wt in the muscles, livers, and eggs, respectively. The frogs exhibited a unique congener profile, compared to those in aquatic and terrestrial species, with BDEs 99, 153, 183, 209, and 47 as the dominant congeners, intermediating between aquatic and terrestrial species. Most of the PBDE congeners in general showed higher affinity to liver than to muscle tissue. Except for BDEs 28, 47, 66, 138, and 206, the average biomagnification factors (BMFs) for all PBDE congeners were greater than 1.0, providing clear evidence of their biomagnification from insects to frogs. A parabolic relationship between log BMFs and bromine atom numbers or log KOW of PBDEs was observed, with the maximum BMF values for PBDEs with 6 bromine atoms (or at a log KOW of approximately 8.0). Relatively higher levels of 3-MeO-BDE 47 were found in male frogs, suggesting that male frogs in the present study might have higher metabolic capacity for PBDEs compared to female frogs. The ratio of levels in egg/female liver, indicating mother-to-egg transfer * Corresponding author phone: +86-20-85290146; fax: +86-2085290706; e-mail: [email protected]. † Chinese Academy of Sciences. ‡ Tsinghua University. § Sun Yat-Sen University. 5212

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capacity, increased with increasing bromine atom numbers up to 7 and then declined as the bromine atom numbers rose. This indicated that the physicochemical properties of the congeners (e.g., KOW, molecular sizes, and structures), resulting in different affinities to transport proteins, might impact their maternal transfer in frogs.

Introduction Amphibians are sensitive to environmental chemicals, which have been implicated as factors contributing to the worldwide decline of amphibian populations (1-3) and a cause of increased observations of morphological malformations (4). The unique characteristics of these poikilothermic animals, i.e., highly permeable skin, low metabolic rate, long life cycle, hepatic sequestration of lipids, and both aquatic and terrestrial habitats, make them vulnerable to accumulation of environmental pollutants. As a result, amphibians have been used as important environmental stress bioindicators (4). There have been growing concerns about polybrominated diphenyl ethers (PBDEs) during the past few decades due to their ubiquitous presence, environmental persistence, and potential toxicity to wildlife and humans (5). Adverse effects such as disruption of thyroid homeostasis in amphibian metamorphosis have been observed in tadpoles exposed to BDE 47 and BDE 99 (6, 7). These effects could cause reduced hind limbs, reduced body weight and body length, inhibition of tail resorption, delayed metamorphosis, and impacts on skin pigmentation. Exposure of tadpole tail tips to BDE 206 antagonizes T3-mediated tail resorption in an ex vivo bioassay, indicating possible thyroid hormone disrupting effects (8). Frogs occupy an important trophic position in the food chain between insects and vertebrates, as well as between the aquatic and terrestrial environments. However, the published information on the presence of PBDEs in frogs is limited (9, 10), and, to the authors’ knowledge, no studies on biomagnification of PBDEs in amphibian food chains have been reported. PBDE levels in wild frogs were first reported by ter Schure et al. (9) who determined BDEs 47 and 99 in the liver of common frog (Rana temporaria) across a long latitudinal gradient in Sweden. Xia et al. (10) found a different PBDE congener profile in adult bullfrogs compared to other aquatic species, but only five congeners (BDEs 47, 99, 100, 153, and 154) were analyzed in that study. Recently, more and more evidence has demonstrated the bioaccumulation of higher brominated congeners (such as BDE 209) in terrestrial wildlife (11-13). Unfortunately, no data could be acquired for PBDEs with more than seven bromine atoms in frogs, although they also possess terrestrial feeding habits. The possible thyroid hormone disrupting effects of higher brominated congeners (BDE 206) in tadpoles (8) makes it important to investigate these congeners in frogs. Exposure to chemicals during the embryonic life stages is harmful for vertebrates, especially for the oviparous/eggproducing organisms, including fish, amphibians, reptiles, and birds, because the early life stages of these organisms often exhibit the greatest toxicological sensitivity to contaminants (14-16). Studies have shown that mortality rates were higher in avian and fish embryos than in adults when exposed to organochlorine chemicals at similar concentrations (17). A recent laboratory study also implied that exposure of frogs to environmental chemicals during their embryonic stages might result in impaired reproductive success (15). Transfer of lipophilic contaminants from the 10.1021/es901103y CCC: $40.75

 2009 American Chemical Society

Published on Web 06/15/2009

TABLE 1. PBDE Concentrations (ng/g wet wt) in Muscle, Liver, and Egg Tissues of Frogs, and in Insects from the Contaminated Site as Well as in Frog Livers from the Reference Site, South China muscle female n ) 22 lipid (%) BDE 28 BDE 47 BDE 66 BDE 85 BDE 99 BDE 100 BDE138 BDE 153 BDE 154 BDE 183 BDE 196 BDE 197 BDE 203 BDE 205 BDE 206 BDE 207 BDE 208 BDE 209 ΣPBDEsc a

0.93 (0.69-1.25) 0.01 (bdlb-0.01) 0.22 (0.07-1.37) 0.01 (0.01-0.03) 0.08 (0.03-0.20) 0.36 (0.12-6.84) 0.10 (0.03-0.78) 0.07 (0.04-0.55) 0.09 (0.04-0.12) 0.17 (0.08-1.02) 0.14 (0.05-0.36) 0.04 (0.01-0.07) 0.04 (0.01-0.07) 0.02 (bdl-0.06) 0.05 (0.01-0.13) 0.04 (bdl-0.15) 0.10 (bdl-0.34) 0.05 (bdl-0.13) 0.22 (0.05-0.72) 2.31 (0.63-11.6)

Median (range).

b

a

egg

reference liver

insects

male n ) 32

female n ) 25

liver male n ) 32

n ) 22

n ) 10

n ) 4 (pool)

0.81 (0.56-1.12) 0.01 (bdl-0.01) 0.11 (0.03-0.62) 0.01 (bdl-0.02) 0.04 (0.02-0.09) 0.37 (0.19-2.67) 0.12 (0.03-0..58) 0.16 (0.05-1.05) 0.03 (bdl-0.12) 0.57 (0.12-2.19) 0.22 (0.10-1.36) 0.04 (bdl-0.37) 0.04 (bdl-0.08) 0.01 (bdl-0.05) 0.06 (0.01-0.80) 0.12 (0.07-0.54) 0.08 (0.03-0.61) 0.03 (bdl-0.16) 0.13 (0.03-0.42) 2.26 (1.41-10.1)

1.62 (1.50-2.51) 0.03 (0.02-0.04) 0.74 (0.28-6.22) 0.01 (bdl-0.05) 0.13 (0.05-0.76) 1.21 (0.67-33.4) 0.32 (0.15-3.72) 0.27 (0.21-2.50) 0.04 (bdl-0.28) 0.64 (0.43-5.08) 0.45 (0.16-1.30) 0.10 (bdl-0.28) 0.13 (bdl-0.28) 0.01 (bdl-0.14) 0.34 (0.10-0.71) 0.33 (0.09-1.72) 0.28 (bdl-0.75) 0.10 (bdl-0.48) 1.37 (0.38-2.22) 7.26 (4.57-56.2)

1.90 (0.88-3.77) 0.02 (0.01-0.04) 0.31 (0.06-2.47) 0.02 (0.01-0.04) 0.15 (0.06-0.32) 1.34 (0.62-16.8) 0.32 (0.21-3.06) 0.58 (0.29-5.13) 0.07 (bdl-0.36) 2.28 (0.81-13.2) 0.70 (0.49-3.00) 0.01 (bdl-0.68) 0.14 (0.04-1.09) 0.01 (bdl-0.55) 0.08 (bdl-3.24) 0.08 (0.01-2.10) 0.08 (0.01-1.59) 0.01 (bdl-0.63) 0.40 (bdl-3.52) 7.26 (4.21-48.6)

6.68 (5.76-7.46) 0.03 (0.01-0.04) 1.84 (0.61-14.1) 0.03 (0.02-0.15) 0.26 (0.15-3.29) 6.18 (2.07-55.6) 1.44 (0.48-12.7) 1.86 (0.87-9.71) 0.15 (0.03-1.08) 4.57 (1.68-18.1) 1.95 (1.27-5.26) 0.24 (0.08-0.43) 1.52 (0.63-5.43) 0.24 (0.09-0.81) 0.57 (0.16-1.20) 0.08 (bdl-0.37) 0.29 (0.02-0.69) 0.16 (0.03-0.39) 0.45 (0.07-0.86) 22.3 (10.7-125)

2.99 (1.85-5.34) 0.01 (bdl-0.02) 0.05 (0.03-0.07) bdl 0.02 (0.01-0.03) 0.03 (0.02-0.07) 0.02 (0.01-0.06) 0.05 (0.03-0.09) 0.10 (0.06-0.42) 0.03 (0.02-0.09) 0.07 (0.05-0.28) bdl bdl bdl bdl bdl bdl bdl 0.39 (bdl-1.41) 0.94 (0.26-2.54)

9.08 (7.86-10.5) 0.50 (0.32-3.53) 5.19 (4.80-16.7) 0.32 (0.17-3.74) 0.34 (0.17-0.44) 6.32 (2.21-7.67) 0.97 (0.45-1.31) 0.15 (0.14-0.16) 0.80 (0.46-1.33) 0.52 (0.35-1.15) 0.07 (0.05-0.07) 0.25 (0.22-0.46) 0.09 (0.04-0.15) 0.09 (0.07-0.15) 0.88 (0.72-1.41) 0.42 (0.29-0.67) 0.01 (bdl-0.07) 0.83 (0.54-1.34) 2.69 (2.41-3.86) 22.1 (17.9-38.1)

Below the detection limit. c Total PBDEs.

maternal parent is the major pathway of exposure for embryos (15). Previous studies have investigated the transfer of organic pesticides, polychlorinated dibenzo-p-dioxins, dibenzofurans, polychlorinated biphenyls (PCBs) (18), and trace elements in frogs (15) from the maternal tissues to the eggs, but no knowledge is currently available on maternal transfer of PBDEs in frogs. To fill this data gap, we collected wild frog samples from an electronic waste (e-waste) recycling site in South China and analyzed a suite of PBDE congeners. The first objective was to investigate the levels, congener profiles, and tissue distributions of PBDEs in frogs. Second, several species of insects comprising the diet of the frogs were also collected and analyzed to assess the extent of PBDE biomagnification in frogs. Finally, pairs of egg and tissue samples from the female frogs were used to evaluate maternal transfer of PBDEs in frogs.

Experimental Section Sampling. Fifty-four mature rice frogs (Rana limnocharis), 32 males and 22 females, were collected from a contaminated site located in Qingyuan County (23.6021 N, 113.0785 E), the second largest e-waste recycling region in South China, during the early phase of the spawning season (April-May) in 2006. Four pooled insect samples (about 160 individuals) were collected concurrently in the habitat of these frogs. Moreover, 10 male frogs were sampled at a location approximately 20 km northeast of the contaminated site, as reference samples. Insects were kept at -20 °C until further treatment. The frogs were anaesthetized with an overdose of MS-222 and dissected after recording the snout-vent length (SVL), body mass, and sex. The details are given in Table S1 (“S” designating Supporting Information). The hind leg muscle, liver, and egg were immediately frozen and kept at -20 °C until further analysis. Only liver was used from the reference samples. Chemical Analysis. Analysis of 18 PBDEs (BDEs 28, 47, 66, 85, 99, 100, 138, 153, 154, 183, 196, 197, 203, 205, 206, 207, 208, and 209) in insects and frog tissue samples was performed following methods described elsewhere (19), with minor modifications. Briefly, the insects and tissue samples were spiked with surrogate standards (CDE 99, 13C-PCB 141, and 13C-BDE 209), ground with anhydrous sodium sulfate,

and Soxhlet extracted with n-hexane/acetone (1/1, v/v) for 48 h. Lipids of the samples were removed by gel permeation chromatography after lipid determination (gravimetric method). Interferences were removed by fractionation on a multilayer silica column. Samples were concentrated and spiked with known amounts of internal standards (BDEs 118 and 128) before instrumental analysis. The instrumental analysis conditions were described elsewhere (19, 20). Detailed procedures of quality assurance and quality control, and the statistical analysis are given in the Supporting Information.

Results and Discussion PBDE Levels and Profiles. All of the samples contained at least 8 of the 18 congeners at concentrations above the LODs. The total PBDE concentrations ranged from 0.63 to 125 ng/g wet wt in the contaminated frogs, which were 3.5-40.7 times higher than in those from the reference site (Table 1). Only one study is available on PBDE levels in wild frogs (9). In that study, BDEs 47 and 99 were detected at concentrations of 0.003-0.123 and 0-0.178 ng/g wet wt, respectively, in livers of the frog Rana temporaria collected from Sweden, which were 2-10 times lower than our values. The results suggested that our frogs were highly contaminated by PBDEs, due to the crude e-waste recycling activities. Statistical comparison of PBDE concentrations (lipidnormalized) between females and males revealed that the median concentrations in females (249 and 408 ng/g lipid in muscles and livers, respectively) were somewhat lower than the values in males (294 and 528 ng/g lipid in muscles and livers, respectively) (Figure S1), although the difference was not significant (Mann-Whitney U test, p ) 0.15). The net transfer of contaminants by egg-laying for females was possibly responsible for the lower concentrations in female tissues. In fact, significantly lower levels of chlorinated compounds in female than in male Rana ornativentris frogs were observed in a previous report (18). The lack of significant differences of PBDE concentrations between males and females in the present study might be attributed to other factors. It could be postulated that the male frogs in the present study might have higher metabolic capacity for VOL. 43, NO. 14, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 1. PBDE congener patterns in egg, muscle, and liver tissues of frogs from a contaminated site, South China. Error bars represent ( 1 standard error. PBDEs compared to female frogs; this will be discussed in detail in the Biomagnification and Biotransformation section below. There was a significant correlation between the total PBDE concentrations and the lipid contents (r 2 ) 0.23, p < 0.005), while no statistically significant relationships were found between the PBDE concentrations and SVL or body mass of the frogs (Figure S2). This is, in part, different from the observations in the Swedish study, in which the PBDE concentrations in the frog livers were not significantly correlated with lipid content, SVL, or body mass (9). The frogs in that study were collected from a relatively large geographic area (1500-km-long latitudinal gradient). The pollutant levels might have been influenced more by the geographical variables than the biological variables. The lack of correlations between the PBDE concentrations and SVL or body mass in both studies suggested that uptake of these pollutants was not associated with the frog body sizes. This is consistent with previous results on the bioaccumulation of PCBs in frogs collected from contaminated sites in the United States and Japan (21, 22). The clear relationship between PBDE concentrations and lipid content in our frogs indicated that lipid might play a key role in the distribution of these chemicals in frogs, and might result in different toxic effects when the lipid changes during different seasons and physiological processes. For example, during the hibernation period and the metamorphism process, lipid in frogs will allocate among different tissues and the total lipid will dramatically decline, which may alter the tissue distributions of PBDEs in frogs and their potential effects. All the frog tissues exhibited similar PBDE congener profiles, with BDEs 99 and 153 being the largest contributors (Figure 1). BDE 99 and BDE 153 constituted 21-30% and 11-28% of the total PBDE concentrations, respectively, followed by BDE 183 at 6-11%, BDE 209 at 2-14%, BDE 47 at 5-10%, BDE 154 at 5-9%, and BDE 100 at 5-7%. The octa- and nona-BDEs were also detected in frogs with contributions of 0.5-6% for each congener. This PBDE congener profile was quite different from those in the aquatic and terrestrial species reported previously. BDE 47 was the dominant congener in aquatic species, followed by BDE 99, whereas the terrestrial species often preferentially accumulated BDE 153 over BDE 99 and BDE 47 (5). A predominance of BDE 209 was also observed in organisms from terrestrial sources (11-13, 23). Apparently, the PBDE congener profiles in frogs were intermediate between those of aquatic and terrestrial species. This is in agreement with their unique feeding habits that combine both the aquatic 5214

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and terrestrial environments. In their early life-stages, frogs (tadpoles) feed on planktonic species of both animals and plants from the water column, and then change to feeding on larger food organisms during metamorphosis. In the adult life-stages, frogs temporarily inhabit wetland on a daily basis and feed on a wide array of terrestrial species, including insects. Significantly different congener distribution patterns between bullfrogs and other aquatic species were also observed in biota caught from the Hudson River, New York, with higher fractions of BDE 99 than BDE 47 (10). As showed in Figure 1, the contributions of nona- to deca- BDEs in egg samples were remarkably lower than in other tissues, probably due to the less efficient transfer to the egg for the higher brominated PBDEs (see Mother to Egg Transfer section below). Liver-Muscle Tissue Distribution. When lipid-normalized concentration ratios of liver to muscle tissues (L/M ratio) were calculated, all the detected congeners except for BDEs 66 and 138 in females, and BDEs 206 and 207 in males, showed L/M ratios higher than 1.0 (mean and median; Figure S3 and Table S2). These findings suggested that this frog species may accumulate PBDEs mostly in the liver. To date, no data on tissue distributions of PBDEs in wild frogs are available, and the results on other organisms (such as fish, birds and mammals) are limited (24-28). Preferential accumulation of tri- to hepta-BDEs in the liver rather than in muscle has been reported for several fish species (25-27), as well as marine and terrestrial mammals (26, 28), due to the high affinity of these lipophilic compounds to the phase-1 biotransformation enzyme cytochrome P450 1A (25, 26). Previous studies indicated that xenobiotic contaminants such as PCBs could induce CYP 1A activity in various frog species (29). Therefore, active accumulation due to the liver’s detoxifying activities and the relatively high lipid content of the liver (0.88-3.77% in liver vs 0.56-1.25% in muscle), rather than passive distribution between the liver and other organs, could be responsible for the selective liver accumulation of PBDEs in our frogs. The deviant ratios of BDEs 66 and 138 in females are probably related to their low detection frequencies and their low concentrations in most samples. BDE 206 and 207 appear to distribute equally between muscle and liver in males as indicated by L/M ratios close to 1.0. Most octa- to deca-BDE congeners showed higher liver accumulation than tri- to hepta-BDE congeners (Figure S3). Several possible explanations may be given. First, the relatively high lipid content of liver favors the deposition of higher brominated congeners due to their high log KOW values in combination with their larger molecular size, which hinders their ability to cross membranes (30). Moreover, preferential binding to blood proteins of higher brominated congeners could also contribute to the high accumulation of these compounds in blood-rich organs such as liver. Several administration studies of BDE 209 using fish (30, 31) and rats (32, 33) have supported this explanation with measured BDE 209 levels in liver higher than in muscle and adipose tissues on a lipid-weight basis. In addition, the liver functions as a xenobiotic-metabolizing organ, where the higher brominated congeners might have been metabolized before they were transported to the muscle and other tissues due to their relatively short half-life (34). Finally, the high levels of higher brominated congeners in the liver may be caused by a recent high exposure to these contaminants. A limited number of studies in field organisms have measured the distribution of octa- to deca-BDEs in various tissues. In Belgian red fox, liver had a higher concentration of BDE 209 than adipose and muscle tissues (11). Specific accumulation of octa- to deca-BDEs in liver rather than adipose tissue was also reported in Japanese Raccoon Dogs (35). However, in a study on PBDEs in dairy cows, octa- to deca-BDEs were found at the lowest concentrations in liver among the tested tissues

including muscle. The authors explained that the cow was in a withdrawal phase resulting in mobilization of PBDEs more rapidly in well-perfused organs, e.g., the liver, than in less perfused tissues such as the adipose and muscle tissues, leading to the depletion of higher brominated congeners in the more-perfused organs (36). Biomagnification and Biotransformation. Adult Rana limnocharis are carnivorous, mostly feeding on insects. Therefore, PBDEs in the insects from the frog’s habitat were analyzed to investigate the biomagnification potential of these chemicals in frogs. Biomagnification factors (BMFs) were calculated by the ratio of the lipid-normalized concentrations in the liver of individual frog sample to the mean lipidnormalized concentration of insects. All BMFs, except for BDEs 28, 47, 66, 138, and 206, were greater than 1.0 on average, providing evidence that most PBDE congeners were biomagnified from insects to frogs (Table S3). BDEs 153, 154, 183, 203, and 205 showed the highest biomagnification potentials (BMFs > 10) in both female and male frogs, indicating that they are highly bioaccumulative compounds. As mentioned above, the relatively low BMFs for BDEs 66 and 138 may be due to their low detection frequencies and low concentrations. There were no significant differences in the BMFs between females and males, meaning a uniform biomagnification potential of PBDEs in frogs between sexes, although females might transfer large quantity of hydrophobic pollutants to the eggs (18). Increasing concentrations of PBDEs with increased trophic levels were also observed in a freshwater food web (including invertebrates and fish) from the same sampling region (37). The trophic magnification factors (TMFs) ranged from 0.53 to 2.64 for tri- to penta-BDEs (BDE 28 to BDE 100), which were on the same order of magnitude as the BMFs calculated in the present study. However, for hexa- to deca-BDEs, except for BDE 138 and BDE 208, the biomagnification extents in Rana limnocharis were 1.2-26.6 times greater than those in the freshwater food web, suggesting a greater biomagnification power for the higher brominated congeners in frogs compared with aquatic species. Additionally, BMFs of 3.3-130 for BDEs 47, 99, 100, and 153 were reported in ringed seal-polar bear (38), and 2-34 for BDEs 28, 47, 99, 100, 153, 154, and 183 in passerine-sparrow hawk, rodent-buzzard, and rodent-fox food chains (39). These BMFs were 2-40 times higher than our values found in the insect-frog food chain, possibly due to the lower trophic position of frogs (Rana limnocharis) compared to mammals and terrestrial animals, and to the influence of both aquatic and terrestrial environments on Rana limnocharis. There is a parabolic relationship between log BMF and bromine atom numbers as well as log KOW of PBDEs (limited to congeners with a log KOW available in the literature) for both female and male frogs, if the deviant values calculated from BDEs 66 and 138 are excluded (Figure 2 and Figure S4). The BMFs were maximized for PBDEs with 6 bromine atoms (or at a log KOW of approximately 8.0) and then declined with increasing degree of bromination. The rise in BMFs from trito hexa-BDEs is expected to result from their lipophilicity, while the subsequent drop in BMFs likely reflects reduced bioavailability and more elimination in feces due to high molecular size or weight, as well as faster metabolic degradation in organisms due to their short half-life. Burreau et al. (40, 41) found the same parabolic relationship between the biomagnification potential and degree of bromination of PBDEs in food webs from the Baltic Sea and the northern Atlantic Ocean. These observed parabolic relationships are also supported by ample evidence from the literature testing the biomagnification of PCBs where the maximum BMFs generally occurred at log KOW values of 6 to 8 (see ref 19 and the literature cited therein).

FIGURE 2. Relationship between log BMFs and the bromine atom numbers of PBDEs in frogs from a contaminated site in South China. Error bars represent ( 1 standard error. A significant correlation between log BMF and log KOW was observed for the tri- to hexa-BDEs (r 2 ) 0.98, p < 0.001; Figure S4). This strong relationship may suggest that these congeners are relatively recalcitrant in this frog species because biotransformation would result in deviation of the BMF values from the general trend predicted by the relationship (42). BDE 47 in male frogs had BMF values much lower than the predicted value based on the log BMF-log KOW relationship and also lower than the BMFs in females (Figure S4), possibly due to relatively high metabolism of BDE 47 in male frogs. To verify this assumption, we measured the concentrations of 3-MeO-BDE 47, known as a methoxylated metabolite of BDE 47 (43), in the muscle and liver tissues of the frogs. The results showed that the concentrations of 3-MeO-BDE 47 in males were significantly higher than the levels in females for both liver and muscle (Figure S5), which confirmed that male Rana limnocharis might have an enhanced metabolic capacity for BDE 47 compared with females. Mother to Egg Transfer. All of the congeners detected in the female liver and muscle tissues were also detected in their eggs with a median lipid-normalized concentration of 314 ng/g lipid for total PBDEs (Table 1, Figure S1), demonstrating the maternal transfer of PBDEs in frogs. It is known that in oviparous organisms, hydrophobic chemicals in females are transferred to eggs along with yolk proteins (such as lipoprotein, phosvitin, and vitellogenin), which are formed in the liver of the mother (14). Therefore, we used the lipidnormalized concentration ratios of egg to liver (E/L ratio) to assess the extent of maternal transfer of PBDEs in frogs. The mean E/L ratios ranged between a minimum of 0.29 for BDE 28 to a maximum of 1.35 for BDE 183 (Table S4). Kadokami et al. (18) reported maternal transfer ratios of 0.39-1.1 for PCBs and several organochlorine pesticides in the Japanese brown frog R. japonica, which are comparable to our results for PBDEs. No data on the maternal transfer of PBDEs in frogs are available, but the limited information on fish and birds can be used for comparison. The concentration ratio between egg and maternal tissue of the guillemot has been found to be 0.96 for ΣPBDEs (sum concentrations of BDEs 47, 99, 100, 153, and 154) (44), similar to our value (0.86). Median ratios of 1.1, 2.6, and 4.7 for BDEs 28, 183, and 209, respectively, were reported in zebrafish exposed to 10 nmol/g dry wt PBDEs (45), which are higher than our values (0.3, 1.3, and 0.1 for BDEs 28, 183, and 209, respectively). The plot of E/L ratios of PBDEs in frogs versus number of bromine atoms revealed that the maternal transfer potentials of PBDEs also followed a parabolic relationship, ignoring the deviant values of BDE 138 (Figure 3). The E/L VOL. 43, NO. 14, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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and the Earmarked Fund of the State Key Laboratory of Organic Geochemistry (SKLOG2008A05). This is contribution No. IS-1081 from GICAS.

Supporting Information Available Details on the samples collected; liver/muscle concentration ratios, BMFs and egg/liver concentration ratios; relationships between the PBDE concentrations and biological variables, between log BMF and log KOW, and between E/L ratio and log KOW. This information is available free of charge via the Internet at http://pubs.acs.org.

Literature Cited

FIGURE 3. Lipid normalized egg-to-liver concentration ratios (mean ( 1 SE) plotted versus bromination degree of PBDEs in female frogs from a contaminated site in South China. Error bars represent ( 1 standard error. ratios increased with increasing numbers of bromine atoms up to 7 and then declined as the bromine atom numbers rose. This observation is inconsistent with the fugacity mode which predicts that the lipid-normalized egg to female concentration ratio is theoretically equal to 1.0 in oviparous organisms (14). Previous studies have showed that contaminant residues in female oviparous organisms were mainly carried by several different lipoproteins in blood, including very low-density lipoproteins and high-density lipoproteins (containing vitellogenin), and then transported from the maternal liver to the eggs (14). The binding of lipophilic pollutants to these lipoproteins during the process of vitellogenesis is important to the transport. Our results clearly showed that the transport of PBDEs from the maternal tissue to the eggs in frogs appeared to be related to both lipophilicity and molecular sizes of the chemicals. For tri- to hepta-BDEs, the maternal transfer potential was likely based on their lipophilicity because chemicals with higher log KOW are expected to have higher affinity to the lipoproteins, whereas for octa- to deca-BDEs, the transfer might be inhibited by their relatively high molecular sizes. The relationships between the maternal transfer ratios and the physicochemical properties of hydrophobic contaminants have been previously described in several studies. A significantly negative correlation between transfer ratios and log KOW was observed for PCDD/Fs in the frogs R. ornativentris and R. japonica (18) and a similar behavior has been reported recently by Serrano et al. for PCBs in fish (46). However, an opposite relationship between egg/fish concentration ratios and log KOW for PBDEs was found in exposed zebrafish, which showed higher ratios for compounds with larger log KOW (45). These observed different relationships indicate that the maternal transfer mechanism of organohalogen contaminants in oviparous organisms cannot be described only by the fugacity model. Several key contaminant properties, such as the degree of halogenation, KOW, molecular sizes and structures (e.g., planarity and polarity), enzyme induction, rate of metabolism, and affinity to transport proteins, all of which are specific to each of the organohalogen groups, might have greater impacts on the behavior of maternal transfer.

Acknowledgments We thank Dr. Yong Luo and Mr. Jing Zheng from Sun YetSen University, China for assistance in samples collection. This work was supported by the National Basic Research Program of China (2009CB421604), the National Science Foundation of China (40632012, 20890112, and 40821003), 5216

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