Horacio O Gonzalez

Jianjun Hu

Kristen M Gaworecki

Jonathan A Roling

William S Baldwin

Jorge L Gardea-Torresdey

Lisa J Bain
email: lbain@clemson.edu

Dose-Responsive Gene Expression Changes in Juvenile and Adult Mummichogs (Fundulus heteroclitus) After Arsenic Exposure

Keywords: Arsenic, Fundulus heteroclitus, fish, apolipoprotein, serum amyloid precursor protein, arrays, gene

The present study investigated arsenic's effects on mummichogs (Fundulus heteroclitus), while also examining 27 what role that gender or exposure age might play. Adult male and female mummichogs were exposed to 172ppb, 28 575ppb, or 1,720ppb arsenic as sodium arsenite for 10 days immediately prior to spawning. No differences were 29 noted in the number or viability of eggs between the groups, but there was a significant increase in deformities in 30 1,720ppb arsenic exposure group. Total RNA from adult livers or 6-week old juveniles was used to probe custom 31 macroarrays for changes in gene expression. In females, 3% of the genes were commonly differentially expressed 32 in the 172 and 575ppb exposure groups compared to controls. In the males, between 1.1-3% of the differentially 33 expressed genes were in common between the exposure groups. Several genes, including apolipoprotein and 34 serum amyloid precursor were commonly expressed in either a dose-responsive manner or were dose-specific, but 35 consistent across genders. These patterns of regulation were confirmed by QPCR. These findings will provide 36 us with a better understanding of the effects of dose, gender, and exposure age on the response to arsenic. 37 38 39

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exposure to both adults and juveniles [START_REF] Roling | Hexavalent chromium reduces larval growth and alters gene expression in mummichogs (Fundulus heteroclitus)[END_REF]. Microarrays were used to identify biomarkers of 90 cadmium exposure in the European flounder (Platichthys flesus) [START_REF] Sheader | Oxidative stress response of European flounder (Platichthys flesus) to cadmium determined by a custom cDNA microarray[END_REF], and in rainbow trout 91 [START_REF] Koskinen | Response of rainbow trout transcriptome to model chemical contaminants[END_REF]. Clusters of rainbow trout genes involved in energy metabolism, protein synthesis, and 92 metal ion transport were upregulated at high exposures (0.5mg/L cadmium), while downregulated at medium or 93 low exposures (0.25mg/L, 0.05mg/L respectively). In contrast, other genes involved in stress response, receptor 5 signaling, G-protein coupled receptor, lipid biosynthesis and sulfur metabolism were upregulated at low doses and 95 repressed at high doses [START_REF] Koskinen | Response of rainbow trout transcriptome to model chemical contaminants[END_REF].

97

Only a few studies have used microarrays to examine arsenic's effects on aquatic organisms. Using 98 arrays, zebrafish exposed to 15ppm arsenic had decreased hepatic glycogen, increased incidence of cholestasis, 99 and changes in overall hepatocyte morphology. Concurrent with these findings, the authors found genes involved in carbohydrate catabolism, DNA repair, and oxidant status-related proteins to be differentially expressed [START_REF] Lam | Transcriptome kinetics of arsenic-induced adaptive response in zebrafish liver[END_REF]. Consequently, the present study used microarrays to investigate changes in the patterns of gene expression after adult and parentally-exposed juvenile mummichogs were exposed to three arsenic concentrations. These findings will provide us with a better understanding of dose-response relationships and the effects of gender and age on the response to arsenic.

Methods

Fish exposures

Mummichogs (2 males and 5 females) were randomly assigned to one of 10 different aquaria, 5 of which contained 20ppt saltwater (Coral Life, Burbank, CA) and 5 of which contained sodium arsenite in 20ppt saltwater.

Three separate studies were conducted using the same set-up, but one exposure was with 172ppb (1.32 M), the second with 575ppb (4.43 M), and the third with 1,720ppb (13.2 M) arsenic, provided as sodium arsenite. Fish were exposed for a total of 10 days, maintained on a natural photoperiod, and fed daily with AquaTox flaked fish food (Zeigler Brothers, Gardners, PA) supplemented with freeze-dried brine shrimp. All studies were conducted using a static-renewal exposure, with 80% of the water being replaced every 48 hours. Although arsenic concentrations in the control aquaria water were below the detection limits, arsenic concentrations in the food (0.03ng/g) are similar to what others have reported (Shaw et al., 2007b).

Spawning substrates were placed in each tank on the three days surrounding the full moon to collect eggs.

The total number of eggs produced was counted and their viability was determined. Eggs from each of the aquaria on each day were placed into petri dishes containing 20ppt saltwater to monitor their viability,

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6 development, and time to hatch. After completion of spawning, the parents were euthanized in 1g/L buffered MS-222 and two female fish from each group were stored at -80°C to determine arsenic body burdens. For the remaining adults, livers were removed, placed in TriReagent, and stored at -80°C. The hatchlings from each exposure group were transferred to an individual 1 gallon aquarium containing 20ppt seawater, maintained on a 16-h light/8-h dark photoperiod at 25°C, and fed recently hatched brine shrimp. The juveniles were observed over the next six weeks to determine whether additional morphological changes, developmental delays, or death would occur, but no additional phenotypes were noticed. Six weeks after hatching, the juveniles were euthanized with an overdose of buffered MS-222 and were stored at -80°C.

Chemical residue analysis.

One intact adult female fish carcass from each aquaria in 1,720ppb experiment was freeze-dried (Labconco, Kansas City, MO), microwave digested (CEM, Mathews, NC) following the USEPA 3051 method (Kingston and Jassie, 1988) and arsenic levels analyzed using an Inductively Coupled Plasma Optical Emission Spectrometer (ICP/OES) Optima 4300 DV (Perkin-Elmer Instruments, Edgewood, NM). Arsenic was detected at 189nm, which has a detection limit of 3ppb. A standard curve for sodium arsenite was used to determine body burdens in the fish, and statistical differences between the groups were determined using Student's t-test.

Morphometric analyses of the vertebrae

Ten control and ten 575ppb parentally-exposed juveniles were examined for changes in vertebral and fin bone size and number. Briefly, the fish were fixed in 10% formalin for 7 days and placed in Alcian Blue dye in 80% ethanol/20% glacial acetic acid solution for 12 hours. After neutralization with potassium hydroxide, the specimens were bleached in 3% hydrogen peroxide. The fish were then digested in a sodium borate/trypsin solution until bones and cartilage were visible, and stored in glycerol [START_REF] Taylor | Revised procedures for staining and clearing small fishes and other vertebrates for bone and cartilage study[END_REF]. A confocal microscope (Carl Zeiss, Thornwood, NY) was used to quantify the area of each vertebra, which was normalized to the weight of each fish.

Gene Expression Changes using Macroarrays

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Targeted mummichog cDNA arrays were constructed using genes from subtractive hybridization or differential display experiments from fish exposed in the laboratory or collected from field sites containing Cr(VI), Cr(III), anthracene, pyrene, and arsenic [START_REF] Bain | cDNA cloning, sequencing, and differential expression of a heart-type fatty acidbinding protein in the mummichog (Fundulus heteroclitus)[END_REF][START_REF] Maples | Trivalent chromium alters gene expression in the mummichog (Fundulus heteroclitus)[END_REF][START_REF] Peterson | Differential gene expression in anthracene-exposed mummichogs (Fundulus heteroclitus)[END_REF][START_REF] Roling | Differential gene expression in mummichog (Fundulus heteroclitus) following treatment with pyrene and fluoranthene: comparison to a creosote contaminated site[END_REF][START_REF] Roling | Hexavalent chromium reduces larval growth and alters gene expression in mummichogs (Fundulus heteroclitus)[END_REF][START_REF] Gonzalez | Physiological changes and differential gene expression in mummichogs (Fundulus heteroclitus) exposed to arsenic[END_REF]. Each membrane contained 270 duplicate clones, including 13 blanks or plasmid controls. Specific information on the array platform is available on the Gene Expression Omnibus website (GEO-GPL2535 ) [START_REF] Roling | Hexavalent chromium reduces larval growth and alters gene expression in mummichogs (Fundulus heteroclitus)[END_REF].

For each experiment, five arrays were probed using RNA samples from control groups and five were probed with samples from the arsenic-exposed groups. Array experiments were conducted for the 172, 575, and 1,720ppb males and for the 172 and 575ppb females. RNA was extracted from a pool of 7 to 10 juveniles per aquarium using TriReagent (Sigma, St. Louis, MO) and treated with DNase I. The pools of RNA were obtained from the 172, 575, and 1,720ppb exposure groups, plus their respective controls. Probes were generated by reverse transcription using 2 g of RNA, 0.1 M oligo dT primer, 1.2mM dNTPs lacking dATP, [ 33 P]dATP (3000Ci/mmol, Perkin Elmer, Boston, MA) and 200U MMLV-RT at 37 o C for 2 hours. Meanwhile, the array membranes were prehybridized for 4-6 hour in Express hybridization solution (Clontech, Palo Alto, CA) containing 100mg/ml denatured salmon sperm. Each probe was purified using Amersham's Probe Quant-50 columns (Piscataway, NJ), denatured, and incubated overnight with the array at 68 o C. Then, the arrays were washed four times with 2X SSC/0.5%SDS at 64 o C and four times with 0.5X SSC/0.5%SDS at 64 o C. The membranes were wrapped in plastic, placed on a phosphorimaging screen, and developed using a Bio-Rad Molecular Imager FX (Hercules, CA). The intensity of each spot was quantified using the Invitrogen ResGen Pathways 4-Universal Microarray Analysis Software (Carlsbad, CA). The data for this series of experiments has been deposited in NCBI's Gene Expression Omnibus (Edgar et al., 2002) and can be accessed using GEO series accession number GSE16772 (http://www.ncbi.nlm.nih.gov/geo/query/acc.cgi?acc=GSE16772).

Microarray analysis.

The CLEAR-test method was used to analyze the microarray data for identifying differentially expressed genes between normal state and different degrees of arsenic exposure for each fish group (Valls et al., 2008). This 8 method combines the z-test, which focuses on large changes, with a x 2 test to evaluate variability, and has the advantage of not reporting genes with small changes and low variances as differentially expressed. For all the duplicate spots, the raw pixel intensity values were first averaged before running CLEAR-test. Average fold expression changes for each gene were obtained for each of the three arsenic concentrations in the male, female, and juvenile groups.

Changes in RNA abundance by real-time PCR

Total RNA from the liver of 5 adult fish of each gender per exposure group was isolated using TRI-Reagent (Sigma) and then treated with RNase-free DNase. Total RNA from a pooled sample of 3 whole juveniles was also isolated after homogenization (n=5 pools per group). To prepare cDNA, total RNA (2µg) was incubated with 50ng random heblxamers, RNasin, 10mM dNTP mix, and 200U Moloney murine leukemia virus (MMLV) reverse transcriptase at 37 o C for 1 hour. Real-time PCR was performed in Bio-Rad's I-Cycler (Hercules, CA) using RT 2 SYBR Green/Fluorescein qPCR master mix (SABiosciences, Frederick, MD), along individual sets of primers for the EST 9-7 (#CO897482; forward primer 5'-GCCAACACCATGTGTGCAAGTTCT -3'; reverse primer 5'-TCTGGACACAGTCACACACAGACT-3'), apolipoprotein (8-56) (#DN474949; forward primer 5'-ATCTTGCAGCTGCTGTTCTGTTGC-3'; reverse primer 5'-GTGTCAAAGTGGGTCTTGGCCTTT-3'), and serum amyloid precursor (1-31) (#AY735160, forward primer 5'-AGGTTATCAGCGATGCCAGAGAGT-3'; reverse primer 5'-ATTTGCTTCCTGATCGGCCTCTGA-3'). 18S rRNA (#X00686; forward primer 5'-TTTCTCGATTCTGTGGGTGGTGGT-3'; reverse primer 5'-TAGTTAGCATGCCGGAGTCTCGTT-3') was used as a housekeeping gene to normalize the total number of molecules in each sample. All PCR products had a denaturing step of 95 C for 15 seconds, an annealing/extension step at 57 o C for 1 minute for a total of 40 cycles.

The cycle threshold values obtained from the real-time PCR were converted into starting number of molecules per 100ng cDNA using known concentrations of the specific gene product, which was normalized to the number of 18S rRNA molecules [START_REF] Gonzalez | Physiological changes and differential gene expression in mummichogs (Fundulus heteroclitus) exposed to arsenic[END_REF]. The standards were prepared by RT-PCR and sequenced at Clemson University to confirm their identity.

Arsenic body burdens

At the conclusion of spawning, female fish from each control and 1,720ppb arsenic aquaria were used to determine the body burdens of arsenic. In the adults, arsenic levels were 2.6-fold higher in the exposed fish than in the controls after 10 days of exposure (Figure 1). This level of uptake is consistent with other studies in Fundulus. For example, adult mummichogs exposed to 8,000ppb arsenic for 4 days had approximately 6 times more arsenic than controls while those fish exposed to 100ppb arsenic for 14 days had approximately 3 times more arsenic than the controls [START_REF] Miller | MRP2 and acquired tolerance to inorganic arsenic in the kidney of killifish (Fundulus heteroclitus)[END_REF]. Similarly, mummichogs exposed to 787 ppb for 14 days had 2.8-fold higher hepatic arsenic levels than controls [START_REF] Bears | Arsenic exposure alters hepatic arsenic species composition and stress-mediated gene expression in the common killifish (Fundulus heteroclitus)[END_REF].

Hepatic gene expression changes in adult Fundulus exposed to arsenic

Microarrays were used to determine both the dose-dependent patterns of gene expression changes in livers following arsenic exposure and whether the response to arsenic differed between genders. Replicated test array experiments were conducted for 2 different exposure concentrations in the females or 3 different exposure concentrations in the males. Of the 257 genes on the array, between 4.7-8.2% were differentially expressed after arsenic exposure amongst the 5 groups of experiments using the CLEAR test method. Analyzing the data for significantly differential changes in gene expression in a dose-responsive manner indicated that in the two female exposure groups, 172ppb and 575ppb arsenic, 16 and 19 hepatic genes were significantly changed compared to the controls, respectively. There were 7 common differentially expressed genes (Figure 2A). In comparing the livers of male fish exposed to arsenic with the controls, 21, 12, and 13 hepatic genes were differentially expressed in 172, 575, and 1,720ppb exposures. There was only 1 gene shared in common between all three male groups, which is an EST termed 10-15. Shared genes between two of the three male exposures ranged from 3 to 7 (Figure 2A). When examining the data from Table 1for the number of common genes changed after arsenic exposure based upon gender, a total of 9 differentially expressed genes were common between males and females. The differential expression was more likely to occur in the 172ppb exposure group than the 575ppb group (Figure 2B). Several genes were expressed in a dose-responsive manner. Only genes with statistically different expression are shown in Table 1. These included apolipoprotein, which was significantly downregulated in both female exposure groups and in two of the three male exposure groups. Apolipoprotein is represented by two partially overlapping clones on the array, 8-56 and Ab35, and both were changed in the same manner and the same magnitude (Table 1). Lysozyme precursor is also represented by two clones, 1-11 and 1-163, with both consistently demonstrating downregulation in the 172ppb males but upregulation in the 1,720ppb males exposed to arsenic (Table 1). Other clones show differences based upon gender, such as tributyltin (TBT)-binding protein, which is downregulated in females but upregulated in males (Table 1). A second overlapping clone for TBTbinding protein termed C3C61 (accession #BM084929) was also significantly downregulated (0.33-fold) in the 575ppb arsenic-exposed females (data not shown). Several other genes were differentially expressed in only one exposure group, but many of these genes are associated with cell stress, damage, and repair, such as heat shock proteins, glutathione S-transferases, transcription elongation factors, and ATP synthase (Table 2).

Reproductive effects, developmental effects, and gene expression changes in the juveniles

In addition to changes in hepatic gene expression, we were interested in investigating the effects of arsenic on embryonic development. After the exposure period, the fish were mated and eggs collected to determine changes in embryonic development and hatchling success. There were no differences in the number of eggs laid per tank between the control and the arsenic-exposed groups at any concentration, nor any differences in egg viability (Figure 3a). After hatching, any deformities in the embryos were recorded. In the offspring whose parents were exposed to 1,720ppb arsenic, there was a significant increase in deformities including embryos with abnormal trunk curvatures, and head and eye deformities (Figure 3b). Because of these abnormalities, it was hypothesized that arsenic could be altering the structure of bone or cartilage. However, no differences were observed in the number, orientation, or position of the vertebrae in parentally-exposed offspring compared to control offspring (Figure 3c).

Unlike the adult hepatic gene expression, we used RNA extracted from whole 6-week old juveniles.

There were very few genes in common between the three different parental exposure groups, except for two

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11

ESTs, 2-80 and 9-7 (Table 1). There was only one additional gene, complement component C8, which was shared between the juveniles and the adults. However, this gene was only significantly differentially expressed in the highest juvenile exposure group.

Confirmation of gene expression patterns

To confirm the expression patterns of several of the genes, we used quantitative PCR (QPCR) to examine expression levels of several selected genes in the livers of the adult males and females. Expression levels of the EST shared in common with the adults, 9-7, was also examined in the juveniles (Figure 4). Apolipoprotein expression on the arrays was downregulated in all adult arsenic exposed groups, regardless of gender or concentration. This was validated in four of the five QPCR reactions. The exception was the 172ppb female group, which based upon QPCR expression, showed no change in expression (fold change=1.03). The EST termed 9-7 showed a bit more variability, but the arrays and QPCR validated one another in 5 of the 7 groups (Figure 4). The expression of 9-7 in the 172ppb female group was downregulated in the arrays but upregulated in the QPCR, while the 575ppb male group was upregulated in the arrays but downregulated based upon QPCR.

The levels of 9-7 in two juvenile groups examined showed similar expression patterns, being upregulated in both the arrays and when using QPCR (Figure 4).

Serum amyloid precursor expression was downregulated in all of the male exposures and the 172ppb female exposure, but upregulated in the 575ppb female exposure group in the arrays. Again, the direction of change using QPCR matched in four of the five groups, but this time, the 172ppb male group showed no changes in expression based upon the QPCR data (fold change=1.07). For the higher exposure groups, both the directionality and magnitude of change were similar. The expression of serum amyloid precursor in the 575ppb males was downregulated 0.83-fold in the arrays and 0.67-fold in the QPCR, while the expression in the 1,720ppb males was downregulated 0.32-fold in the arrays and 0.28-fold in the QPCR. Interestingly, serum amyloid precursor expression was upregulated in the 575ppb female group, and the magnitude of change between the arrays and QPCR was also similar (2.5-fold in arrays; 3.7-fold in the QPCR). Overall, the QPCR data showed similar fold-changes and directionality, validating the array data.

Arsenic-mediated changes in gene expression

The use of fish toxicogenomics is suited for studying contaminants such as arsenic, as fish can act as both models for human diseases as well as models for the health of the aquatic environment ( Williams et al., 2003;[START_REF] Gorman | Teleosts as models for human vertebral stability and deformity[END_REF]. In the present study, adult mummichogs were exposed to three different arsenic concentrations, and the differential expression of genes was investigated by examining the effects of gender and maternal transfer to offspring. Several genes were differentially expressed as a result of the arsenic exposure, including apolipoprotein, serum amyloid precursor, tributyltin-binding protein, and cytochrome c oxidase.

Overall, arsenic appears to perturb genes involved in energy utilization.

Apolipoprotein was one gene that was downregulated in the adult livers at all exposure concentrations tested. Using QPCR, the underexpression of apolipoprotein and the magnitude of downregulation were in concordance with the array data. Apolipoproteins are lipid binding proteins important in lipid secretion, lipolytic enzyme activation, and transport and binding of lipoproteins in cells (Atkinson, 1992). Several isoforms have been identified, which bind to and act as ligands for low density or very low density lipoprotein receptors (Atkinson, 1992). In our experiment, both clones of the apolipoprotein C-1 were down-regulated similarly in females as well as males. The reduction in apolipoprotein expression is in concordance with other arsenic exposure studies. For example, arsenic downregulated apolipoprotein in the lung fluid of male mice exposed to 50ppb sodium arsenite for four weeks [START_REF] Lantz | In utero and postnatal exposure to arsenic alters pulmonary structure and function[END_REF], and was differentially expressed in zebrafish embryos exposed to arsenic [START_REF] Mattingly | Perturbation of defense pathways by low-dose arsenic exposure in zebrafish embryos[END_REF].

Another protein, serum amyloid precursor, was also downregulated in the adult males and in the 172ppb female exposure group. Its altered expression was examined by QPCR, which confirmed a dose-responsive reduction in the male 172, 575, and 1720ppb groups. Serum amyloid precursor proteins (APPs) are also a family of apolipoproteins (Zheng and Koo, 2006). Although most often associated with the development of Alzheimer's disease (reviewed in: [START_REF] Esler | A portrait of Alzheimer secretases-new features and familiar faces[END_REF][START_REF] Jaeger | Functional role of lipoprotein receptors in Alzheimer's disease[END_REF] 13 cholesterol transport, ApoE metabolism, and help maintain cholesterol homeostasis (Yao and Papadopoulos, 2002;[START_REF] Liu | Amyloid precursor protein regulates brain apolipoprotein E and cholesterol metabolism through lipoprotein receptor LRP1[END_REF]. Additionally, cellular cholesterol can modify the both the processing and transcription of APP [START_REF] Bodovitz | Cholesterol modulates alphasecretase cleavage of amyloid precursor protein[END_REF][START_REF] Fassbender | Simvastatin strongly reduces levels of Alzheimer's disease beta-amyloid peptides Abeta 42 and Abeta 40 in vitro and in vivo[END_REF][START_REF] Kiyosawa | Effect of serum cholesterol on the mRNA content of amyloid precursor protein in rat livers[END_REF]. In general, the downregulation of APP in the adult mummichogs mimics the reduction in apolipoprotein expression.

Both apolipoprotein and APPs are involved in lipid homeostasis, and in other studies, exposure to arsenic does appear to alter lipid levels. Mice exposed to arsenic via drinking water from weaning to 1 year of age, or from 8 weeks to 1 year of age had significant reductions in serum triglycerides, total cholesterol, and HDL cholesterol compared to controls [START_REF] Ahlborn | Impact of life stage and duration of exposure on arsenic-induced proliferative lesions and neoplasia in C3H mice[END_REF]. The authors hypothesized that the reduction in lipids could be due to alterations in fatty acid production, as there was a reduction in the mRNA expression of stearoyl-CoA desaturase 1, the rate limiting enzyme in monounsaturated fatty acid synthesis [START_REF] Ahlborn | Impact of life stage and duration of exposure on arsenic-induced proliferative lesions and neoplasia in C3H mice[END_REF].

Interestingly, the mice in their study exposed to arsenic only in utero did not have alterations in serum lipids [START_REF] Ahlborn | Impact of life stage and duration of exposure on arsenic-induced proliferative lesions and neoplasia in C3H mice[END_REF]. These findings are in concordance with our studies showing a marked reduction in both apolipoprotein and APP proteins in the adults, but not in the offspring. Thus, arsenic exposure may impair cholesterol homeostasis, which could have adverse physiological consequences for fish [START_REF] Heath | Water pollution and fish physiology[END_REF].

Examining lipid levels in the livers and serum of arsenic-exposed fish, and how that may impact adult health and egg quality would be worthy of future investigation.

Indeed the present study, along with many others, shows that arsenic often exhibits a complex gene expression profile. For example, arsenic exposure in zebrafish embryos resulted in the differential expression of several immune-responsive genes following a 10µg/L exposure, but no change in expression at 100µg/L [START_REF] Mattingly | Perturbation of defense pathways by low-dose arsenic exposure in zebrafish embryos[END_REF]. In additional microarray studies using cell lines or rodent models, arsenic exposure results in very few commonly differentially expressed genes between the exposure groups [START_REF] Ahlborn | Dose response evaluation of gene expression profiles in the skin of K6/ODC mice exposed to sodium arsenite[END_REF], or can alter the expression of genes in opposite directions at low versus high concentrations [START_REF] Andrew | Genomic and proteomic profiling of responses to toxic metals in human lung cells[END_REF][START_REF] Lau | Opposed arsenite-induced signaling pathways promote cell proliferation or apoptosis in cultured lung cells[END_REF][START_REF] Snow | Arsenic, mode of action at biologically plausible low doses: what are the implications for low dose cancer risk?[END_REF][START_REF] Andrew | Exposure to arsenic at levels found in U.S. drinking water modifies expression in the mouse lung[END_REF]. This may indicate that differing transcriptional

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14 networks are activated at low arsenic levels versus high arsenic levels, resulting in differing biological processes being altered [START_REF] Bodwell | Arsenic at very low concentrations alters glucocorticoid receptor (GR)-mediated gene activation but not GR-mediated gene repression: complex dose-response effects are closely correlated with levels of activated GR and require a functional GR DNA binding domain[END_REF][START_REF] Bodwell | Arsenic disruption of steroid receptor gene activation: Complex dose-response effects are shared by several steroid receptors[END_REF].

Gender differences in gene expression

We also found differing responses depending on the gender of the fish. For example, serum amyloid precursor was upregulated in the livers of the highest female exposure group, but downregulated in the livers of the males. In contrast, several ESTs were upregulated in the males by downregulated in the females. Genderbased differences in disease outcomes after arsenic exposure have been reported in the literature. For example, men have a higher incidence of pathological liver changes after arsenic exposure, while women are at greater risk of arsenic-related kidney, lung, and bladder cancers than men [START_REF] Vahter | Gender differences in the disposition and toxicity of metals[END_REF]Lindberg et al., 2008b). The same increase in bladder lesions in females has also been reported in studies with rodents [START_REF] Shen | A comparative study of the sub-chronic toxic effects of three organic arsenical compounds on the urothelium in F344 rats; gender-based differences in response[END_REF][START_REF] Waalkes | Urogenital carcinogenesis in female CD1 mice induced by in utero arsenic exposure is exacerbated by postnatal diethylstilbestrol treatment[END_REF], while increases in liver tumors after arsenic exposure are seen in male rodents ( Waalkes et al., 2003;Waalkes et al., 2004a;Waalkes et al., 2004b;[START_REF] Ahlborn | Impact of life stage and duration of exposure on arsenic-induced proliferative lesions and neoplasia in C3H mice[END_REF]. It is known that hepatic metabolism of contaminants in some organisms can vary by gender (reviewed in [START_REF] Burger | Methodologies, bioindicators, and biomarkers for assenting gender-related differences in wildlife exposed to environmental chemicals[END_REF] and it has been speculated that sex hormones affect in arsenic metabolism, altering its methylation, excretion, and thus its pathology [START_REF] Shen | A comparative study of the sub-chronic toxic effects of three organic arsenical compounds on the urothelium in F344 rats; gender-based differences in response[END_REF]Lindberg et al., 2008a;Lindberg et al., 2008b). As the fish in our study were reproductively active, hormonal changes may have influenced the effects of arsenic on gene expression.

Effect of arsenic on development

The second part of this study was to examine changes in development and gene expression in the offspring whose parents were exposed to arsenic. Viability was not different between the control and arsenicexposed groups at any exposure concentration. However, parental exposure to arsenic increased the incidence of developmental deformities, such as spinal curvatures and smaller heads, by 17.4-fold at the highest concentration as compared to controls. In a previous experiment using a concentration of 230ppb arsenic, there was a 2.8-fold increase in the incidence of spinal deformities in the hatchlings of exposed parents [START_REF] Gonzalez | Physiological changes and differential gene expression in mummichogs (Fundulus heteroclitus) exposed to arsenic[END_REF]. A study with zebrafish revealed various similar morphological abnormalities including dorsal curvature, flattened head, and pericardial edema [START_REF] Li | Developmental mechanisms of arsenite toxicity in zebrafish (Danio rerio) embryos[END_REF], and mice exposed in utero via maternal oral treatment also

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15 exhibited skeletal abnormalities [START_REF] Hill | Reproductive consequences of oral arsenate exposure during pregnancy in a mouse model[END_REF]. In humans, a decreased head and chest size was noted in children exposed to arsenic in utero [START_REF] Rahman | Arsenic exposure during pregnancy and size at birth: A prospective cohort study in Bangladesh[END_REF].

It has been well documented that arsenic can cross the maternal blood supply and be detected in the fetus [START_REF] Concha | Exposure to inorganic arsenic metabolites during early human development[END_REF][START_REF] Kubota | Placental transfer of arsenic to fetus of Dall's porpoises (Phocoenoides dalli)[END_REF][START_REF] Fängström | Impaired arsenic metabolism in children during weaning[END_REF]. A study looking at the maternal transfer of arsenic from organisms such as adult black-tailed gulls noted that 11.3% of the arsenic burden was transferred to eggs [START_REF] Kubota | Maternal transfer of arsenic to eggs of blacktailed gull (Larus crassirostris) from Rishiri Island, Japan[END_REF], fetal porpoise arsenic burden was about 36% of that measured in the mothers [START_REF] Kubota | Placental transfer of arsenic to fetus of Dall's porpoises (Phocoenoides dalli)[END_REF], and newborn mouse livers had 10% of the average arsenic concentration in maternal livers (Xie et al., 2007). Thus, even the low amounts of arsenic transferred to the offspring may impact its gene expression and subsequent development.

It is interesting to note that offspring exposed to arsenic in utero have very different pathologies based upon their gender. In mice exposed to arsenic in utero, males have increase in liver and adrenal tumors, while females have increases in urogenital tumors, but no changes in the incidence of liver or adrenal tumors (Waalkes et al., 2003;Waalkes et al., 2004a;Waalkes et al., 2004b;[START_REF] Ahlborn | Impact of life stage and duration of exposure on arsenic-induced proliferative lesions and neoplasia in C3H mice[END_REF]. We did not determine the sex of the juveniles in our study before using them in the microarray study. Given that in rodents, very different arsenic metabolite profiles and pathologies are seen, not sexing our juvenile fish might have contributed to the lack of gene expression changes and the high amount of variability associated with the offspring.

In summary, we observed changes in lipid responsive genes at environmentally-relevant concentrations, indicating that arsenic exposure may impair cholesterol homeostasis, which could have adverse effects on organismal health. These findings will provide us with a better understanding of the effects of dose, gender, and exposure age on the response to arsenic. Waalkes, M.P., Ward, J.M., Liu, J., Diwan, B.A., 2003. Transplacental carcinogenicity of inorganic arsenic in the drinking water: induction of hepatic, ovarian, pulmonary, and adrenal tumors in mice. Toxicol. Appl. Pharm. 186, 7-17. Williams, T.D., Gensberg, K., Minchin, S.D., Chipman, J.K., 2003. A DNA expression array to detect toxic stress response in European flounder (Platichthys flesus). Aquat. Toxicol. 65, 141-157. Xie, Y., Liu, J., Benbrahim-Tallaa, L., Ward, J.M., Logsdon, D., Diwan, B.A., Waalkes, M.P., 2007. Aberrant DNA methylation and gene expression in livers of newborn mice transplacentally exposed to a hepatocarcinogenesis dose of inorganic arsenic. Toxicology 236, 7-15. Yamaguchi, S., Miura, C., Ito, A., Agusa, T., Iwata, H., Tanabe, S., Tuyen, B.C., Miura, T., 2007. Effects of lead, molybdenum, rubidium, arsenic and organochlorines on spermatogenesis in fish: monitoring at Mekong Delta area and in vitro experiment. Aquat. Toxicol. 83, 43-51. Yamamoto, M., Hirano, S., Vogel, C.F., Cui, X., Matsumura, F., 2008 Representative photographs from normal and deformed offspring. C. Morphological analysis of the vertebrae.

The total size (area) for individual vertebrae were quantified using the LSM5 Pascal software. Values are the average area in square micrometers divided by the weight of the 6-week old fish. Significant differences (*) were determined using Student's t-test (p<0.05).

Fig. 4. Comparisons of gene expression changes between the arrays and quantitative real-time PCR

(QPCR). Expression of three genes from the arrays that were differentially expressed in almost all genders, ages, and exposure groups were verified by QPCR using specific primers (n=5 for each group). Array fold changes were plotted against QPCR fold changes to determine their directionality and magnitude. Genes Table 1. Gene Expression Changes in Fundulus Exposed to Arsenic. Fold-changes in gene expression were calculated and significant differences were determined using the CLEAR Test method (p<0.05). Only genes that are differentially expressed and shared between more than 1 exposure group are listed.

Fold change (Arsenic/control) 2. Genes showing unique differential expression. Fold-changes in hepatic gene expression in the males and females were calculated and significant differences were determined using the CLEAR Test method (p<0.05). Only genes that were differentially expressed in only 1 exposure group are listed. GenBank accession numbers are indicated in parentheses.

Clone Identity and GenBank Accession Number

  , APPs are thought to function to control

Fig. 1 .

 1 Figure Legend

Fig. 2 .

 2 Fig. 2. Gene expression profiles for males and female Fundulus exposed to different concentrations of

Fig. 3 .

 3 Fig. 3. Percentage viability and morphological abnormalities following arsenic exposure. Adults were

  , serum amyloid precursor, and an EST termed 9-7. The letter indicates females 631 (F), males (M), or juveniles (J), while the number (172, 575, or 1,720) indicates the exposure group.

  Figure1

  . Selective activation of NF-kappaB and E2F by low concentration of arsenite in U937 human monocytic leukemia cells. J. Biochem. Mol. Toxicol, 136-146. Yao, Z., Papadopoulos, V., 2002. Function of beta-amyloid in cholesterol transport: a lead to neurotoxicity. FASEB J. 16, 1677-1679. Yoshida, T., Yamauchi, H., Fan Sun, G., 2004. Chronic health effects in people exposed to arsenic via the drinking water: dose-response relationships in review. Toxicol. Appl. Pharmacol. 198, 243-252. Zheng, H., Koo, E., 2006. The amyloid precursor protein: beyond amyloid. Mol. Neurodegener. 1, 5. Zheng, X.H., Watts, G.S., Vaught, S., Gandolfi, A.J., 2003. Low-level arsenite induced gene expression in HEK293 cells. Toxicology 187, 39-48.

Table

	Table 1			ACCEPTED MANUSCRIPT
	Clone ID 8--56 Ab35 9--7 9--12 8-6-T7 Bb56 1--31 2--80 9--16	Gene Name apolipoprotein apolipoprotein EST TBT-binding protein EST EST serum amyloid precursor EST EST	GenBank # DN474949 CN992696 CO897482 AY725225 DN596277 DN596273 AY735160 DN474945 CO897478	Females 172ppb 575ppb 0.59 0.75 0.59 0.79 0.61 1.22 0.56 0.6 5.71 0.45 0.67 1.79 0.63 2.52 0.54 0.7 M A N U S C R I P T Males 172ppb 575ppb 1,720ppb 0.85 0.85 0.85 0.84 1.14 1.71 1.95 0.77 0.32 0.81 2.13 0.74 1.39	172ppb 4.1	Juveniles 575ppb 1.54	1,720ppb 1.45 1.29
	pa_1--15 Bb53 8--40 10--15 10--57 1--11 1--163 8--46 8--8	complement component C8 cytochrome c oxidase subunit II cytochrome c oxidase subunit II EST EST lysozyme precursor lysozyme precursor EST EST	AY521664 CV821001 CN976180 CO897479 CO897481 AY735156 AY735156 DN474948 CO897478 A C C E P T E D 0.7 1.57 0.8	0.82 0.7 0.6 1.14 1.41 0.3 0.29	1.93 1.8 1.37 1.68	0.87 0.55 2.65 2.85 0.82 0.82	1.58

expression

For example, exposure to arsenic resulted in a reduction in 11-ketotestosterone-induced spermatogenesis in

catfish (Pangasianodon hypophthalmus)(Yamaguchi et al., 2007), while zebrafish fed metal-contaminated

Acknowledgements:

This work was supported by NCRR #2G12 RR008124, NSF/EPA #EF 0830117, and CSREES/USDA,

under project number #SC-1700380. It is technical contribution No. 5647 of the Clemson University Experiment

Station. The authors would also like to thank Marian Viveros and Judi Ellzey (UTEP) for technical assistance.

A C C E P T E D ACCEPTED MANUSCRIPT

B.

Arsenic exposed offspring with trunk curvature

C.

Figure 3


Yao-Wen Qiu
email: yqiu@scsio.ac.cn

Gan Zhang

Ling-Li Guo

Gene J Zheng

Shu-Qun Cai

Bioaccumulation and Historical Deposition of Polybrominated Diphenyl Ethers (PBDEs) in Deep Bay, South China

Keywords: Polybrominated diphenyl ethers, Benthos, Cores, Bioaccumulation, Stable isotopes, Deep Bay, South China

To characterize the bioaccumulation and historical trends of polybrominated diphenyl ethers (PBDEs) in Deep Bay, an important water body between Hong Kong and mainland China with a Ramsar mangrove wetland (Maipo), marine organisms and core sediments were collected to determine their PBDEs concentrations. Sediment core dating was accomplished using the 210 Pb method. PBDEs concentrations in fish ranged from 0.17 to 4.16 ng g -1 wet wt., with a mean value of 2.00 ng g -1 wet wt.

Temporal trends of the target PBDE congeners levels in core sediment generally increased from 1948 to 2003, with the highest levels in top sediment, suggesting an ongoing PBDEs input. The average sedimentation flux of PBDEs was 0.25 ng cm -2 a -1 , and the double time of total PBDEs concentration was ca. 12.3 a. Correlations between the biota-sediment accumulation factors (BSAF) of PBDEs and their corresponding octanol-water partition coefficient (K ow ) were discussed.

Introduction

Polybrominated diphenyl ethers (PBDEs) are the additive flame retardants and the next highest production group of brominated diphenyl ethers (BFRs) currently in use.

PBDEs are widely used in manufactured products such as polymers, resins, electronic devices, building materials, textiles and polyurethane. PBDEs enter the environment via direct and indirect urban domestic/industrial sewage discharges and atmospheric deposition, and they have a tendency to accumulate in aquatic sediments and organisms due to their high octanol-water partition coefficients [START_REF] Song | Polybrominated diphenyl ethers in the sediments of the Great Lakes. 1. Lake Superior[END_REF]. Some PBDE congeners exhibit substantially long environmental half-lives, potential bioaccumulation and possible adverse effects on wildlife and humans [START_REF] De Wit | An overview of brominated flame retardants in the environment[END_REF][START_REF] Mcdonald | A perspective on the potential health risks of PBDEs[END_REF]. They can impair attention, learning, memory and behavior in laboratory animals and possible interfer with early neurodevelopment [START_REF] Fowles | Immunologic and endocrine effects of the flame retardant pentabromodiphenyl ether (DE-71) in C57BL/6J mice[END_REF][START_REF] Darnerud | Toxic effects of brominated flame retardants in man and wildlife[END_REF]Brigden et al., 2005).

PBDEs have been widely detected in air, water, sediment, organism, and humans in recent years. Their levels in biota were estimated to double every 5 years, and were reported to increase by 60-fold in human breast milk in Sweden between 1972 and 1997 [START_REF] Meironyt | Analysis of Polybrominated Diphenyl Ethers in Swedish Human Milk, 1972-1997[END_REF][START_REF] De Wit | An overview of brominated flame retardants in the environment[END_REF][START_REF] Law | Levels and trends of brominated flame retardants in the European environment[END_REF]. The estimated domestic production of predominant commercial deca-BDE mixture in China increased by 200% between 2000 and 2005, from 10 000 to about 30 000 metric tons [START_REF] Chen | Time Trends of Polybrominated Diphenyl Ethers in Sediment Cores from the Pearl River Estuary, South China[END_REF]. Furthermore, China appears to be a large dumping site of e-wastes imported from overseas legally and illegally in recent years (Brigden et al., 2005;[START_REF] Guo | Polybrominated Diphenyl Ethers in Seafood Products of South China[END_REF]. A previous study using skipjack tuna to elucidate the global distribution of PBDEs suggested that the coastal region off the East China Sea was supposedly a ''hot spot'' for PBDEs [START_REF] Ueno | Global pollution monitoring of polybrominated diphenyl ethers using skipjack tuna as a bioindicator[END_REF]. Apparently, the PBDEs contamination status in the environment of China and their effects on the ecosystem and humans remain to be fully understood.

Deep Bay, located in the eastern Lingdingyang of Pearl River Estuary, South China, is a natural, shallow estuarine bay with an average water depth of about 3 m, and a tidal range of about 1.4 m. It is 17.5 km long in SW-NE direction and 4~10 km in NW-SE direction (Fig. 1). The wetlands around Inner Deep Bay and Mai Po marshes are important to the migratory birds for many decades. The 1500 hectares of wetlands around the area was designated as a Wetland of International Importance under the Ramsar Convention in 1995. The consumption of PBDEs is believed to increase in this region because of the fast industrialization in the last two decades, which made the region the most important electronic manufacturing base in the world. [START_REF] Zheng | Concentrations of polybrominated diphenyl ethers (PBDEs) in[END_REF] reported that the PBDE concentrations in green-lipped mussels from Hong Kong waters were among the highest in the world. [START_REF] Mai | Distribution of polybrominated diphenyl ethers in sediments of the Pearl River Delta and adjacent South China Sea[END_REF] reported that although the total PBDEs concentrations in sediments of the Pearl River Delta and adjacent South China Sea were mostly <50 ng/g, within the range for riverine and coastal sediments around the world, the BDE-209 concentrations at some sites were at the high end of the worldwide figures. [START_REF] Chen | Time Trends of Polybrominated Diphenyl Ethers in Sediment Cores from the Pearl River Estuary, South China[END_REF] ). PBDEs residue levels in fish from South China had been also measured (Meng et al., 2007;[START_REF] Xiang | Polybrominated diphenyl ethers in biota and sediments of the Pearl River Estuary, South China[END_REF][START_REF] Guo | Levels, compositions and bioaccumulation of organochlorine pesticides (OCPs) and polybrominated diphenyl ethers (PBDEs) in fishes in the Pearl River estuary and Daya Bay, South China[END_REF]. Recent study showed that there was a fresh DDT input into the bay (Qiu et al., 2009a)

Materials and methods

Samples collection

Samples of benthic organisms and core sediment in seven stations from Deep Bay, as shown in Fig. 1, were collected in February 2004. A stainless steel static gravity corer (8 cm i.d.) was employed to collect the sediment which was sliced immediately at 1 cm intervals and stored in polyethene bags. The organism samples, with size ranging from 5 cm to 15 cm and lifespan approximately ranging 0.5-2 a, were collected by benthic trawling and stored in polyethene bags. The organism samples from seven stations were mixed into one to represent organisms in Deep Bay as it was not easy to collect enough samples from each station. Since most of the fishes were small benthos fish, they should be considered as local fish. All samples were transported to the laboratory at about 0 ºC and frozen at -20 ºC prior to treatments.

Chemical analysis

Samples preparation.

Analytical procedures generally followed those of [START_REF] Richardson | Chlorinated hydrocarbon contaminants in Hong Kong sediments[END_REF] and [START_REF] Zheng | Concentrations of polybrominated diphenyl ethers (PBDEs) in[END_REF]. Frozen organism samples were thawed and rinsed individually with purified water to remove possible impurities. The edible parts from about 3-5 specimens of the same specimens (17 fishes, one shrimp and one crab) collected from each location were pooled. Prior to extraction, the organism and sediment samples were freeze-dried and homogenized. Subsamples -3, 15, 28, 47, 99, 153, and 183) were added. A hexane:acetone:dichloromethane (1:1:1) solution (150 ml) was added, and heated to 65 °C to gently boil the sediment powder in solvent for 4 h. After cooling, the mixture was filtered using a pre-cleaned glass filter, and the residue was washed three times with hexane and dichloromethane (total 30 ml each solvent). For organism samples, an aliquot of one tenth was used to gravimetrically determine the lipid content and the left solution was treated with concentrated sulfuric acid (5 ml × 3) to remove the bulk of lipids. The extraction solution both organism and sediment samples was reduced in volume to 1 ml by rotary evaporation prior to column chromatography. A chromatography column consisting of 6 g of pre-activated silica gel and 1 g of activated copper powder was washed with 15 ml acetone, followed by 15 ml of dichloromethane and 20 ml of hexane. The extraction aliquot (1 ml) was then added on the column and the PBDEs were eluted with a 30 ml mixture of dichloromethane and hexane (1:1), followed by 15 ml of dichloromethane. The combined eluate (45 ml)

was reduced in volume to 300 µl for quantitation by GC-MS spectrometry.

GC-MS analysis. PBDEs ( < octa-PBDE) were analyzed in all samples using a Varian Saturn 2200 Ion Trap GC/MS/MS with a 30 m DB-1 capillary column (0.25 mm diameter and 0.25µm film thickness), with helium at a rate of 1 ml min -1 as the carrier gas. BDE197, 207, 209 were analyzed with a similar but shorter column (10 m DB-1). The injector temperature was 280 °C. The GC oven temperature was programmed as follows: 90 °C for 2 min, increasing 20 °C per minute to 230 °C, followed by 2 °C per minute to 247 °C, then 20 °C per minute to a final temperature

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of 280 °C, which was held for 25 min. The mass spectrometer was operated in the electron impact/selected ion monitoring (EI/SIM) mode. PBDE concentrations in core sediments 28,32,35,37,47,49,66,71,75,77,85,99,100,119,138,153,153,154,166,181,183 and 190) and organisms (including 15,28,47,60,85,99,100,138,153,154,183,197,207 and 209) Isotope analysis. The organism samples were analyzed for δ 15 N using an elemental analyzer EA (Flash EA 1112, CE) connected to an isotopic ratio mass spectrometer-IRMS (Delta plus XL, Finnigan). Approximately 0.8 mg of the sample was weighed in a tin cup and combusted in the elemental analyzer. Samples and standards were run in continuous flow with a standard deviation of 0.2‰ among replicate standard samples.

Sediment Dating

Sediment core dating was accomplished via 210 Pb analysis (Zhang et al., 2002;Qiu et al, 2009b). Briefly, the 210 Pb activities in sediment sub-samples were determined by

) 1 ln( 1 ξ λ - - = - t
where λ is the radioactive decay constant for 210 Pb (0.03114 a -1 ),

∫ ∫ ∞ = g dg g A dg g A 0 0 ) ( / ) ( ξ
, and dg is the thickness of each sediment interval (g cm -2 a -1 ). For sediment cores, the three undisturbed cores (A, B, E) were selected and analyzed for sediment dating. Since core A and E showed a mixing depth of ca. 7 cm and ca. 11.5 cm, respectively, only core B (total 39 cm long) was eventually selected to determine the PBDEs concentrations. In general, PBDEs concentrations in Shenzhen River estuary (site A and F) are expected to be influenced more by sewage discharge, while PBDEs concentrations in the mouth of the bay (site E) are potential to be affected more by the runoff of Pearl River. The simultaneous study showed that concentrations of hydrophobic organic compounds such as polycyclic aromatic were not significantly different (Qiu et al, 2009b). Moreover, Deep Bay is a relatively small bay and core B is located in the center of the bay. Therefore, it is reasonable to select core B as representing PBDEs distribution in the sediment of Deep Bay. Fig. S1 presents the distribution of excessive 210 Pb activity in the core-B, showing that the average sedimentation flux of the Core-B was 0.28 g cm -2 a -1 , with an apparent sedimentation rate of 0.69 cm a -1 .

Results and Discussion

PBDEs in organisms

The concentrations of 15 PBDE congeners 15,28,47,60,85,99,100,138,153,154,183,197,207 and 209) in organisms, as well as their 15 N values, are shown in Fig. 2. The mean total PBDEs concentrations in fish samples (2.00±1.14 ng g -1 wet wt.) were generally higher than those in shrimps (0.88 ng g -1 wet wt.) and those in crabs (1.44 ng g -1 wet wt.). PBDEs concentrations in wild fish of Deep Bay (mean: 2.00, 0.17-4.16 ng g -1 wet wt. or mean: 62.9, 20.9-166.4 ng g -1 lipid wt.) were higher than those wild fish of Daya Bay, South China (mean: 0.16, 0.04-0.70 ng g -1 wet wt., [START_REF] Guo | Levels, compositions and bioaccumulation of organochlorine pesticides (OCPs) and polybrominated diphenyl ethers (PBDEs) in fishes in the Pearl River estuary and Daya Bay, South China[END_REF] and those in farmed fish of the coast of South China (mean: 0.15, <0.0012-3.85 ng g -1 wet wt., Meng et al, 2007), but were lower than those wild biota in Pearl River Estuary where the median levels of the sum of BDEs 28,47,66,85,99,100,138,153,154,and 183 in large yellow croaker, silvery pomfret, flathead fish, robust tongue fish, and bombay duck were 67.3, 123.9, 114.4, 181.2, and 249.7

ng g -1 M A N U S C R I P T A C C E P T E D ACCEPTED MANUSCRIPT
11 lipid wt., respectively [START_REF] Xiang | Polybrominated diphenyl ethers in biota and sediments of the Pearl River Estuary, South China[END_REF]. This is consistent with the fact that Pearl River Estuary was more polluted than Daya Bay and Deep Bay. PBDEs concentrations in wild fish of Deep Bay were comparable to those in skipjack tunas Katsuwonus pelamis from offshore waters in Asia-Pacific regions and Indian Ocean (<0.1-53 ng g -1 lipid wt., [START_REF] Ueno | Global pollution monitoring of polybrominated diphenyl ethers using skipjack tuna as a bioindicator[END_REF]) and those in farmed Atlantic salmon Salmo salar (0.49-10.92 ng g -1 wet wt., [START_REF] Hites | Global Assessment of Polybrominated Diphenyl Ethers in Farmed and Wild Salmon[END_REF]), but were lower than those in crappie Poxomis annularis and bluegill Lepomis macrochiras form Hadley Lake in the United States (1500-1900 ng g -1 lipid wt., [START_REF] Dodder | Concentrations and spatial variations of polybrominated diphenyl ethers and Several Organochlorine Compounds in fishes from the northeastern United States[END_REF]. In the present study, relatively higher total PBDE concentrations were observed in benthic fish (e.g., Cynoglossus robustus) and carnivorous fish (e.g., Inimicus cuvieri), while relatively lower PBDEs concentration was found in planktivorous fish (e.g., Mugil cephalus) indicating a positive correlation between fish PBDEs burden and their food habit. Increases in contaminant concentrations from the prey to the predator can be successfully defined using the relative abundances of naturally occurring stable isotopes of nitrogen (i. e., δ 15 N values; [START_REF] Hobson | Determination of trophic relationships within a high Arctic marine food web using δ13C and δ15N analysis[END_REF]. The increase in δ 15 N is generally 3 to 5‰ between each trophic level [START_REF] Minigawa | Stepwise enrichment of 15N along food chains: Further evidence and the relation between δ15N and animal age[END_REF], though a consumer's diet often consists of prey organisms from different trophic levels, depending on the availability of prey organisms [START_REF] Ruus | Influence of trophic position on organochlorine concentrations and compositional patterns in a marine food web[END_REF]. In the present study, planktivorous fish Mugil cephalus showed a lower δ 15 N value (7.1‰)

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12 than carnivorous fish Inimicus cuvieri (14.4‰) in Deep Bay. As shown in Fig. 2, positive correlations between total PBDEs (normalized to lipid content) and δ 15 N values in planktivorous fish were apparently observed, whereas negative correlations between them in small carnivorous fish were also found. Relationships between log PBDEs (ng/g lipid wt.) and δ 15 N values in the organisms of Deep Bay were shown in Fig. 3-a. The same pattern was also observed for BDE-47. A weak bioaccumulation of hydrophobic chemicals in organism of the tropical-subtropical waters may be interpreted by enhanced partition of hydrophobic chemicals between water and organisms due to higher temperature, as well as by relatively quick metabolism and short resident time of the chemicals in organisms [START_REF] Guo | Levels, compositions and bioaccumulation of organochlorine pesticides (OCPs) and polybrominated diphenyl ethers (PBDEs) in fishes in the Pearl River estuary and Daya Bay, South China[END_REF]. There was a significant correlation between log PBDEs (ng/g wet wt.) and log percent lipid in the organisms from Deep Bay (R 2 =0.54; P<0.01; Fig. 3-b), confirming the fact that PBDEs are hydrophobic in nature and easily bioaccumulated in organisms. -17, 28, 32, 35, 37, 47, 49, 66, 71, 75, 77, 85, 99, 100, 119, 138, 153, 153, 154, 166, 181, 183 and 190) concentrations in the core sediment from Deep Bay ranged from 0.07 to 4.85 ng/g dry wt, with median value of 0.68 ng/g dry wt (or 65 ng/g organic carbon). In the neighboring area, i.e. the Pear River Estuarine [START_REF] Mai | Distribution of polybrominated diphenyl ethers in sediments of the Pearl River Delta and adjacent South China Sea[END_REF], the average concentration of BDE-209 in surface sediment was 18.5 ng/g (range: 0.7-111.9 ng/g), while the sum of 10 PBDE congeners concentrations in surface sediments was 3.13 ng/g (range: 0.3-21.8 ng/g), which was higher than those in Deep Bay for the same 10 congeners (2.22 ng/g). Total Σ 22 PBDEs levels in surface sediments in the present study ( 3.31 ng/g dry wt, top 2 cm in core sediment) were generally comparable to those in surface sediment samples collected in Pearl River Estuary of South China (0.11-13.03 ng/g, [START_REF] Zheng | Concentrations of polybrominated diphenyl ethers (PBDEs) in[END_REF], those in the coastal region of Korea (1.1-33.8 ng/g, [START_REF] Moon | Contaminations of polybrominated diphenyl ethers in marine sediments from the southeastern coastal areas of Korea[END_REF], and those in the river and coastal areas of Portugal (0.5-20 ng/g, [START_REF] Lacorte | Occurrence and specific congener profile of 40 polybrominated diphenyl ethers in river and coastal sediment from Portugal[END_REF], but were generally lower than those in Osaka Bay of Japan (8-352 ng/g, [START_REF] Ohta | Contamination levels of PBDEs, TBBPA, PCDDs/DFs, PBDDs/DFs, and PXDDs/Fs in the environment of Japan[END_REF] Macao [START_REF] Chen | Time Trends of Polybrominated Diphenyl Ethers in Sediment Cores from the Pearl River Estuary, South China[END_REF], Osaka Bay [START_REF] Sakai | Time trends and horizontal distribution of polybrominated diphenyl ethers (PBDEs) in sediment cores from Osaka Bay, Japan[END_REF], Western Europe [START_REF] Zegers | Levels of polybrominated diphenyl ether flame retardants in sediment core from western Europe[END_REF] and the Great Lakes [START_REF] Song | Polybrominated diphenyl ethers in the sediments of the Great Lakes. 1. Lake Superior[END_REF]. While different patterns occurred in the sediment cores of Tokyo Bay where PBDEs levels increased up to the mid-1990s and decreased afterwards [START_REF] Minh | Spatial distribution and vertical profile of polybrominated diphenyl ethers and hexabromocyclododecanes in sediment core from Tokyo Bay, Japan[END_REF]. In addition, the absence of all PBDE congeners in the older layers (before the late 1950s) in the present study indicated the absence of natural production of the target BDE congeners. This situation also occurred in sediment cores from the Drammenfjord (Norway), the western Wadden Sea (The Netherlands), and the freshwater Lake Woserin (Germany) [START_REF] Zegers | Levels of polybrominated diphenyl ether flame retardants in sediment core from western Europe[END_REF]. The apparent early arrival of PBDEs (late 1950s) in present study might actually reflect bioturbation, which can move contaminants deeper into the core.

Fig. 4. PBDE concentrations in the sediment core by year of deposition

Based on 210 Pb dating results, the current sedimentation flux of PBDEs was 0.92 ng cm -2 a -1 , with the average of 0.25 ng cm -2 a -1 . In order to understand the pollution rate, the time required for the concentration of PBDEs to double (T2) was estimated from data measured in sediment core from Deep Bay. The T2 value of PBDEs concentrations was 12.3 a (r 2 =0.80), which was longer than those in sediment cores from the adjacent Pearl River Estuary (T2=6~7 a, [START_REF] Chen | Time Trends of Polybrominated Diphenyl Ethers in Sediment Cores from the Pearl River Estuary, South China[END_REF], but was comparable to those in Lake Michigan of US (T2=11.2 a, Zhu et al., 2005). The T2 value in the present study was approximately 1-2 folds higher than the 4-6 a reported

M A N U S C R I P T A C C E P T E D ACCEPTED MANUSCRIPT

15 for biota and humans [START_REF] Hites | Polyhrominated diphenyl ethers in the environment and in people: A meta-analysis of concentrations[END_REF].

Due to their high hydrophobic properties, PBDE congeners are expected to be associated mainly with the organic carbon-rich particles in the sediment. In the present sediment core, total organic carbon (TOC) contents ranged from 0.67% to 1.41%, with an average of 1.01%. Statistical analysis showed that TOC contents explained 52% of the variability in total PBDEs concentrations, implying the role played by organic matters in the transport and redistribution of PBDEs in Deep Bay.

Compositions and Sources

Among the 15 PBDEs 15,28,47,60,85,99,100,138,153,154,183,197,207 and 209) measured in organisms in the present study, BDE-47 and BDE-100 were the predominant congeners, accounting for 46.5% and 13.5% of total PBDEs in organisms from Deep Bay, respectively (Fig. S2-a). This is consistent with the fact that BDE-47 and BDE-100 are the most bioaccumulative congeners among the commercial Penta-BDE mixtures [START_REF] Oros | Levels and distribution of polybrominated diphenyl ethers in water, surface sediments, and bivalves from the San Francisco Estuary[END_REF]. It is noteworthy that, despite the large quantity of BDE-209 used in Asia/China [START_REF] Ueno | Global pollution monitoring of polybrominated diphenyl ethers using skipjack tuna as a bioindicator[END_REF] as well as its high concentrations in some of the sediments from the Pearl River Delta [START_REF] Mai | Distribution of polybrominated diphenyl ethers in sediments of the Pearl River Delta and adjacent South China Sea[END_REF], BDE-209 comprised only <1.3% of the total PBDEs in the organisms of Deep Bay. This may imply that BDE-209 has a low bioaccumulation potential to organisms (Boon et al., 2002;[START_REF] Guo | Levels, compositions and bioaccumulation of organochlorine pesticides (OCPs) and polybrominated diphenyl ethers (PBDEs) in fishes in the Pearl River estuary and Daya Bay, South China[END_REF].

Of the 22 PBDE congeners 28,32,35,37,47,49,66,71,75,77,85,99,100,119,138,153,153,154,166,181,183 and 190) in the present study, BDE-47, -49, -99 and -183 were the abundant congeners in core sediment, accounting for

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ACCEPTED MANUSCRIPT 16 21.4%, 10.9%, 13.8% and 17.4% of total PBDEs concentrations, while in surface sediment (top 2 cm) they accounted for 30.6%, 13.9%, 14.0% and 7.2% of total PBDEs concentrations, respectively (Fig. S2-b). It should be noted that BDE-209 was the most predominant congener in sediment of the adjacent Pearl River Estuary, accounting approximately 86% of total PBDEs [START_REF] Mai | Distribution of polybrominated diphenyl ethers in sediments of the Pearl River Delta and adjacent South China Sea[END_REF]. In the present study, BDE-47 and BDE-49 in surface sediments were higher than the average values in sediment core, whereas BDE-183 was on the contrary. Recent studies showed that octa-BDE can be debrominated to hepta-BDE 183, and hexa-BDE 153, as well as the two most commonly detected PBDEs in the environment, penta-BDE 99 and tetra-BDE 47, by anaerobic dehalogenating bacteria [START_REF] Robrock | Pathways for the Anaerobic Microbial Debromination of Polybrominated Diphenyl Ethers[END_REF]. Our results suggest that some proportion of BDE-183 in sediment core may be produced in situ by debromination, but not for BDE-47 and BDE-49. Environmental debromination in sediments was not evident in Deep Bay.

The same pattern was also observed on the west coast of Canada [START_REF] Johannessen | Joined by geochemistry, divided by history: PCBs and PBDEs in Strait of Georgia sediments[END_REF]. In general, highly brominated congeners can be transformed to lower brominated congeners or metabolites, which could become an important source of low brominated BDEs in the environment [START_REF] Keum | Reductive debromination of polybrominated diphenyl ethers by zerovalent iron[END_REF].

Previous studies showed that BDE-47 and BDE-99 were the most common congeners in air, water, and sediment, as well as in people, fish, birds, and mammals [START_REF] De Wit | An overview of brominated flame retardants in the environment[END_REF][START_REF] Watanabe | Environmental release and behavior of brominated flame retardants[END_REF][START_REF] Law | Levels and trends of brominated flame retardants in the European environment[END_REF]. The pattern of the major congeners in sediment from Deep Bay remained similar to that of the industrial formulation Bromkal 70-5DE, which may be related to the widely used penta-and

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octa-BDE commercial products and is probably the original source. For example, BDE-99 is a major component in penta- and 44.8% in Bromkal 70-5DE [Chemische Fabrik Kalk], [START_REF] La-Guardia | Detailed polybrominated diphenyl ether (PBDE) congener composition of the widely used penta-, octa-, and deca-PBDE technical flame-retardant mixtures[END_REF], and in the present study, its relative abundance in the core samples was 13.8%. Compared to those in other environmental compartments of local areas, the proportion of BDE-99 in core sediment of Deep Bay was lower than those in water of Pearl River (27%, [START_REF] Guan | Riverine fluxes of polybrominated diphenyl ethers from the Pearl River Delta (China) to the coastal ocean[END_REF], air of Gunagzhou (28%, [START_REF] Chen | Concentration levels, compositional profiles, and gas-particle partitioning of polybrominated diphenyl ethers in the atmosphere of an urban city in South China[END_REF], and sediments of the Pearl River Delta and northern South China Sea (38%, [START_REF] Mai | Distribution of polybrominated diphenyl ethers in sediments of the Pearl River Delta and adjacent South China Sea[END_REF]. BDE-47 is another abundant congener in penta-BDE products (38.2% in DE-71 and 42.8% in Bromkal 70-5DE, La-Guardia et al., 2006), and was predominant in all samples, with a mean percentage content of 24%, which was generally close to those in riverine runoff (35%, [START_REF] Guan | Riverine fluxes of polybrominated diphenyl ethers from the Pearl River Delta (China) to the coastal ocean[END_REF], air (39%, [START_REF] Chen | Concentration levels, compositional profiles, and gas-particle partitioning of polybrominated diphenyl ethers in the atmosphere of an urban city in South China[END_REF], and sediments (28%, [START_REF] Mai | Distribution of polybrominated diphenyl ethers in sediments of the Pearl River Delta and adjacent South China Sea[END_REF]. BDE-183, considered as a marker compound for the octa-mix PBDE formulation (42% in DE-79;Great Lakes Chemical;[START_REF] La-Guardia | Detailed polybrominated diphenyl ether (PBDE) congener composition of the widely used penta-, octa-, and deca-PBDE technical flame-retardant mixtures[END_REF], was also found in the core sediment of Deep Bay at relatively high median concentrations, equivalent to 0.12 ng/g (15% of total PBDEs), suggesting the extensive use of technical octa-BDEs in local areas. The high proportion of BDE-183 (8%-16%) in the waterbirds of South China has also been reported [START_REF] Luo | Persistent Halogenated Compounds in Waterbirds from an e-Waste Recycling Region in South China[END_REF]. These results were completely different from those in three cores collected from Western Europe where BDE183 was systematically absent [START_REF] Zegers | Levels of polybrominated diphenyl ether flame retardants in sediment core from western Europe[END_REF]. E-waste and waste discharges in surrounding areas were reported as the potential sources [START_REF] Mai | Distribution of polybrominated diphenyl ethers in sediments of the Pearl River Delta and adjacent South China Sea[END_REF][START_REF] Wu | Bioaccumulation of polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) in wild aquatic species from an electronic waste (e-waste) recycling site in South China[END_REF].

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In core sediment, significant positive correlations were found among PBDE congeners except for BDE-181, -166 and -71 which were at very low levels (Table S1), suggesting that these congeners had similar source inputs and underwent similar deposition and decomposition processes. Principal component analysis (PCA) showed that the first two factors accounted for 75% of the total variance in the dataset (Fig. S3). Tri-to penta-BDE congeners primarily contributed to factor 1 which explained 63% of the total variance, indicating that these contaminants may originate from penta-BDE commercial products and their metabolites. Hexa-and hepta-BDEs formed another group and contributed primarily to factor 2. They might have been formed from the mixed octa-BDE products and by the degradation of deca-BDE through microorganism and photolysis [START_REF] Söderström | Photolytic Debromination of Decabromodiphenyl Ether (BDE209)[END_REF][START_REF] Gerecke | Anaerobic degradation of decabromodiphenyl ether[END_REF].

Correlations among PBDE congeners in organisms are also summarized in Table S2 and Fig. S4.

BSAF of PBDE

Hydrophobic and lipophilic organic contaminants can easily accumulate in organisms such as fish mainly via bioconcentration directly from the water environment [START_REF] Leblanc | Trophic level differences in the bioconcentration of chemicals: implications in assessing environmental biomagnification[END_REF], and/or biomagnification through trophic transfer [START_REF] Kiriluk | Using stable nitrogen and carbon isotopes to characterize the biomagnification of DDE, mirex, and PCB in Lake Ontario pelagic food web[END_REF]. Aquatic organism especially benthic organisms can absorb contaminants from sediment, and its ability can be expressed by the biota-sediment accumulation factors (BSAF), defined as a ratio of the lipid-normalized concentration of the chemical in tissue and the organic carbon-normalized concentration of the chemical in sediment. In this study, we selected the PBDE congeners measured both [START_REF] Dodder | Concentrations and spatial variations of polybrominated diphenyl ethers and Several Organochlorine Compounds in fishes from the northeastern United States[END_REF] reported that 4-6 Br PBDE in fishes from the northeastern United States had the higher ability of bioaccumulation.

It was reported that the maximum bioaccumulation for PBDE in British Columbia grizzly bears was found to be in the mid-range of PBDE log K ow [START_REF] Christensen | Persistent Organic Pollutants in British Columbia Grizzly Bears: Consequence of Divergent Diets[END_REF].

  investigated the time trend of PBDE concentration in sediment cores of Pearl River Estuary and results showed that total PBDEs (except for BDE 209) concentrations in all sediment cores increased gradually from the bottom (mid-1970s) to the middle layer (later 1980s and

  . To our knowledge, few studies on PBDEs in Deep Bay have been conducted. The present work aimed to (1) study the bioaccumulation of PBDEs in organisms of Deep Bay; (2) investigate the chronology of PBDE concentration and flux in Deep Bay as recoded in sediment core based on the 210 Pb dating; (3) analyze the compositions of PBDE congeners and the potential sources, and (4) discuss the relationships between the biota-sediment accumulation factors (BSAF) of PBDEs in Deep Bay and the corresponding octanol-water partition coefficient (K ow ).

Fig. 1 .

 1 Fig. 1. Sampling locations in Deep Bay, South China

  were dried to constant weight to determine the water contents. Dry samples were ground into powder with a stainless steel blender. Approximately 10 g of powder was placed in a round bottom flask, to which known quantities of 13 C-labeled PBDE

  α-radioactivity of its decay product 210 Po, on the assumption that the two are in equilibrium. The Po was extracted, purified, and self-plated onto silver discs at 75-80 o C in 0.5 mole/L of HCl, with 209 Po (NIST, USA) used as yield monitor and tracer in quantification. Counting was conducted by computerised multi-channel α-spectrometry with gold-silicon surface barrier detectors. Supported 210 Po was obtained by indirectly determining the α-activity of the supporting parent 226 Ra, which was carried by co-precipitated BaSO 4 . As the core did not reach the critical depth of excessive 210 Pb→0, only a Constant Initial 210 Pb Concentration (CIC) model was used, which assumes that dry mass entering the sediment pool has the same radioactivity concentration,to give an average sedimentation rate. Sedimentation age was calculated using a constant flux (CF) model (McCall et al., 1984):

  sediments of 7 sites (A, B, C, D, E, F and G) of Deep Bay

Fig. 2 .

 2 Fig. 2. The PBDEs concentrations and δ 15 N values in organisms collected from Deep

Fig. 3 .

 3 Fig. 3. The relationships between PBDEs concentrations and δ 15 N values and lipid

  and those in San Francisco Estuary of US (BDL-212 ng/g, Oros et al., 2005). PBDEs levels in surface sediment in Deep Bay were generally at the low-intermediate ranges in the world. The down-core variations of PBDE concentrations and TOC contents in the core as a function of the deposition year are shown in Fig. 4. In general, the levels of PBDE congeners started to increase from the late 1960s onward and reached the second highest values in 1990s and the highest values in 2003 (top layer). These increases coincided with the production and usage figures of penta-mixture in the world since the 1970s. The second peak PBDE concentrations occurred in the 1990s, which might indicate a critical time for the widespread use of the BFR-related substances in Pearl River Delta. There was a continuous and obvious increase of PBDEs concentrations toward the surface layer after 1982, suggesting the fresh PBDEs inputs into Deep Bay. have been also observed in the sediment cores from

  sediment (top 2 cm) to calculate the BSAF for individual PBDE congeners in each specimen. Individual BDE congeners with descending values of BSAFs (in parentheses) for all specimens in Deep Bay followed the sequence of BDE-100 (0.72) > BDE-154 (0.68) > BDE-85 (0.48) > BDE-153 (0.39) > BDE-28 (0.30) > BDE-47 (0.29) > BDE-138 (0.016) > BDE-183 (0.014) > BDE-99 (0.001). In the adjacent Pear River Estuary[START_REF] Xiang | Polybrominated diphenyl ethers in biota and sediments of the Pearl River Estuary, South China[END_REF], the median values of BSAF for biota followed the sequence of BDE-100 (17.35) > BDE-47 (11.18) > BDE-154 (4.49) > BDE-99 (3.36) > BDE-66 (2.81) > BDE-153 (2.58) > BDE-28 (1.46) > BDE-183 (0.04) > BDE-138 (0.00) > BDE-209 (0.00). In Deep Bay and Pearl River Estuary, the highest BSAF value was BDE-100 while the lower BSAF values were BDE-138 and BDE-183. The BASF values of the target PBDE congeners (<0.8) in the present study were lower than those in fish from the lakes of the northeastern United States (16-38, Dodder et al., 2002), suggesting a weaker bioaccumulation of PBDE in organisms of Deep Bay. The relatively longer age (3-4 a) of the target fish (Pomoxis annularis and Lepomis macrochirus) in the above study may partly contribute the higher BASF values. Correlations between the BSAF of PBDE and the corresponding K ow in DeepBay were shown in Fig.5. As the bromine substituted number of PBDE congener increases, the corresponding K ow and hydrophobic properties increase, leading to the higher ability for bioaccumulating contaminants. On the other hand, as the bromine substituted number of PBDE congener increases, the molecular volume and the steric effect also increase, resulting in a lower ability for bioaccumulating contaminants. In the present study, 3-6 Br PBDE except for BDE-99 and BDE-138 had the higher

Fig. 5 .Fig. 1 .

 51 Fig. 5. The biota-sediment accumulation factors (BSAF) of PBDE congeners in

Fig. 2 .

 2 Fig. 2. The PBDEs concentrations and δ 15 N values in organisms collected from

Fig. 3 .

 3 Fig. 3. The relationships between PBDEs concentrations and δ 15 N values and lipid

Fig. 4 .

 4 Fig. 4. PBDE concentrations in the sediment core by year of deposition (BDE209

Fig. 5 .

 5 Fig. 5. The biota-sediment accumulation factors (BSAF) of PBDE congeners in

Fig. 3 .Fig. 5 .

 35 Fig. 3. The relationships between PBDEs concentrations and δ 15 N values and lipid

Fig. S2 .*

 S2 Fig. S2. The percentage of PBDE congeners in organism and sediment of Deep Bay,

  were analyzed.

	QA/QC. Retention times and mass spectra fragmentation for PBDEs 3, 15, 28, 47,
	99, 153, and 183 were established by using 13 C-labeled PBDEs. Other PBDEs were
	identified based on the literature. All samples were run in duplicate. Concentrations of
	PBDEs in solvent blanks (n = 2) were low (<0.08 ng g -1 ), and the samples were
	considered valid when the concentrations exceeded three times that of the blank
	values. All reported values were corrected with a mean value (0.075 ng g -1 ) of the
	blanks. Solution spike recoveries ranged between 76% (PBDE 183) and 108% (PBDE
	47).

Table S1 .

 S1 Correlations between PBDE congeners and total organic carbon (TOC) in core sediments of Deep Bay, South China

								ACCEPTED MANUSCRIPT
		TOC(%) BDE17 BDE28 BDE32 BDE35 BDE37 BDE47 BDE49 BDE66 BDE71 BDE75 BDE77 BDE85 BDE99 BDE100 BDE119BDE138 BDE153BDE154 BDE166BDE183 BDE190 PBDEs
	TOC(%)	1	0.70	0.65	0.40	0.60	0.61	0.66	0.70	0.63	0.44	0.40	0.46	0.43	0.62	0.60	0.26	0.57	0.58	0.70	0.24	0.58	0.47	0.72
	BDE17 BDE28 BDE32 BDE35 BDE37 BDE47 BDE49 BDE66 BDE71 BDE75 BDE77 BDE85 BDE99 BDE100 BDE119 BDE138 BDE153 BDE154 BDE166 BDE183 BDE190 PBDEs * ρ<0.01 as coefficient is larger than 0.40, while ρ<0.05 as coefficient ranges 1 0.97 0.63 0.92 0.84 0.97 1.00 0.95 0.26 0.87 0.90 0.86 0.84 0.88 0.61 0.67 0.46 0.58 0.26 0.38 0.70 0.95 1 0.61 0.97 0.90 0.96 0.98 0.98 0.21 0.89 0.94 0.90 0.83 0.89 0.61 0.65 0.45 0.55 0.21 0.32 0.69 0.94 1 0.62 0.57 0.62 0.64 0.60 0.14 0.60 0.52 0.56 0.57 0.57 0.41 0.47 0.35 0.34 0.14 0.39 0.49 0.63 1 0.90 0.92 0.94 0.97 0.16 0.87 0.94 0.92 0.78 0.84 0.65 0.65 0.44 0.48 0.21 0.36 0.74 0.90 1 0.81 0.86 0.90 0.38 0.76 0.83 0.82 0.69 0.76 0.71 0.64 0.48 0.54 0.46 0.39 0.68 0.85 1 0.97 0.93 0.19 0.84 0.88 0.84 0.90 0.92 0.54 0.66 0.50 0.60 0.18 0.39 0.69 0.96 1 0.97 0.26 0.88 0.91 0.88 0.85 0.89 0.62 0.67 0.47 0.58 0.24 0.37 0.70 0.96 1 0.21 0.90 0.95 0.92 0.81 0.88 0.66 0.65 0.46 0.52 0.24 0.32 0.70 0.92 1 0.17 0.05 0.12 0.16 0.14 0.42 0.22 0.30 0.34 0.68 0.43 0.23 0.32 1 0.91 0.93 0.72 0.79 0.65 0.56 0.27 0.27 0.15 0.13 0.67 0.80 1 0.92 0.73 0.83 0.63 0.54 0.30 0.38 0.12 0.13 0.65 0.83 1 0.78 0.85 0.72 0.67 0.38 0.36 0.20 0.20 0.78 0.84 1 0.97 0.46 0.83 0.73 0.74 0.15 0.47 0.80 0.94 1 0.53 0.83 0.69 0.73 0.16 0.38 0.80 0.94 1 0.46 0.28 0.26 0.70 0.30 0.67 0.60 1 0.82 0.79 0.31 0.54 0.89 0.79 1 0.88 0.36 0.78 0.70 0.68 1 0.36 0.63 0.61 0.72 1 0.54 0.43 0.32 1 0.55 0.55 1 0.80 1 M A N U S C R I P T
	0.30~0.40 (n=38). Principal Component Analysis (PCA) was performed to get the rotated factor pattern of PBDEs in core sediment using varimax with Kaiser normalization. A C C E P T E D
												38

Table S2 .

 S2 Correlations between PBDE congeners and lipid contents in organisms ofDeep Bay, South China

		Lipid	BDE3 BDE15 BDE28 BDE47 BDE60 BDE85 BDE100 BDE99 BDE154 BDE153 BDE138 BDE183 BDE197 BDE207 BDE209 PBDEs
	Lipid	1	0.12	0.65	0.17	0.52	0.56	0.50	0.33	0.12	0.29	0.58	0.12	0.54	0.54	0.32	0.29	0.69
	BDE 3		1	-0.20	-0.14	-0.40	0.51	-0.10	0.24	1.00	-0.20	0.53	1.00	0.62	-0.16	-0.36	-0.28	-0.17
	BDE15			1	0.36	0.54	0.15	0.18	-0.02	-0.20	0.08	0.16	-0.20	0.14	0.39	0.56	0.66	0.56
	BDE28				1	0.38	-0.02	0.33	-0.03	-0.14	-0.02	0.09	-0.14	-0.13	0.07	0.37	0.45	0.54
	BDE47					1	-0.06	0.59	-0.06	-0.40	0.47	0.10	-0.40	0.03	0.56	0.68	0.60	0.85
	BDE60						1	0.12	0.72	0.51	-0.01	0.92	0.51	0.62	0.46	-0.20	-0.09	0.34
	BDE85							1	0.10	-0.10	0.74	0.34	-0.10	0.20	0.54	0.23	0.29	0.74
	BDE100								1	0.24	-0.06	0.71	0.24	0.35	0.47	-0.25	0.00	0.32
	BDE99									1	-0.20	0.53	1.00	0.62	-0.16	-0.36	-0.28	-0.17
	BDE154										1	0.16	-0.20	0.14	0.44	0.41	0.25	0.50
	BDE153											1	0.53	0.68	0.57	-0.12	0.04	0.51
	BDE138												1	0.62	-0.16	-0.36	-0.28	-0.17
	BDE183													1	0.22	-0.17	-0.08	0.24
	BDE197														1	0.24	0.36	0.74
	BDE207															1	0.71	0.54
	BDE209																1	0.60
	PBDEs																	1

Acknowledgments

This work was funded by National Natural Science Foundation of China (No.

M A N U S C R I P T A C C E P T E D ACCEPTED MANUSCRIPT

21 40821003 and No.40573061) and the commonwealth project from the State Oceanic Administration of China (No.200705026). We thank the anonymous reviewers for their kind constructive suggestions on the manuscript.

Supporting Information

Table S1. Correlations between PBDE congeners and total organic carbon (TOC) in core sediments of Deep Bay, South China

Supporting Information

AUTHORES:

Yao-Wen Qiu, Gan Zhang, Ling-Li Guo, Gene J Zheng, Shu-Qun Cai