Exploratory assessment of sportfish consumption and polybrominated diphenyl ether exposure in New York State anglers

ARTICLE IN PRESS Environmental Research 108 (2008) 340–347

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Exploratory assessment of sportfish consumption and polybrominated diphenyl ether exposure in New York State anglers$ Henry M. Spliethoff a,, Michael S. Bloom b, John Vena c,d, Joseph Sorce a, Kenneth M. Aldous a, George Eadon a a

Division of Environmental Disease Prevention, Wadsworth Center, New York State Department of Health, Empire State Plaza, PO Box 509, Albany, NY 12201, USA Department of Environmental Health Sciences, School of Public Health, University at Albany, State University of New York, One University Place, Rm. 153, Rensselaer, NY 12114, USA c Department of Social and Preventive Medicine, School of Public Health and Health Professions, University at Buffalo, The State University of New York, 3435 Main St., 182 Farber Hall, Buffalo, NY 14214, USA d Arnold School of Public Health, University of South Carolina, 800 Sumter Street, Columbia, SC 29208, USA b

a r t i c l e in f o

a b s t r a c t

Article history: Received 24 December 2007 Received in revised form 4 June 2008 Accepted 22 July 2008 Available online 31 August 2008

A cross-sectional study was conducted to examine the influence of sportfish consumption on body burden of nine polybrominated diphenyl ether (PBDE) congeners in 36 New York State (NYS) anglers. Participating anglers who had previously reported consuming sportfish from Lake Ontario and its tributaries were found to have significantly higher blood plasma levels of BDE-28, BDE-47, BDE-99, BDE100, and the sum of measured PBDE congeners (SPBDE), than anglers who had previously reported no consumption of sportfish from these waters. Bivariate analysis was used to evaluate potential dietary predictors of PBDE plasma levels, including indicators of consumption of sportfish, as well as commercial fish, wild waterfowl, dairy products, and beef. The number of years of reported consumption of Lake Ontario sportfish between 1980 and 1990 was found to be correlated with plasma levels of BDE-47, BDE85, BDE-99, BDE-100, BDE-153, BDE-154, and SPBDE. The number of meals, eaten in the year prior to study participation, of Lake Ontario sportfish species known to have high levels of other persistent organic pollutants (POPs) was correlated with plasma levels of BDE-28, BDE-47, BDE-85, BDE-99, BDE-100, BDE154, and SPBDE. Multiple linear regression revealed that the number of years consuming Lake Ontario sportfish between 1980 and 1990, after adjusting for plasma lipids, was a weak, but statistically significant, predictor of SPBDE plasma levels (b ¼ 0.130, 95% CI: 0.007–0.254). These results suggest that sportfish consumption can contribute measurably to PBDE body burden in NYS anglers, although there are likely to be additional, more significant, sources of exposure. & 2008 Elsevier Inc. All rights reserved.

Keywords: PBDEs Sportfish consumption Great Lakes Dietary exposure Biomonitoring

1. Introduction Polybrominated diphenyl ethers (PBDEs) have been added as flame retardants to consumer products and other materials since the 1970s. PBDEs can constitute up to 30% of the weight of foams or plastics (Hooper and She, 2003). These compounds can easily enter the environment, because they are typically bound to

$ This work was completed at the Wadsworth Center as part of the New York State Biomonitoring Program (Centers for Disease Control and Prevention (CDC), National Center for Environmental Health (NCEH) Grant U59CCU22339202) in cooperation with the New York State Angler Cohort Study at University at Buffalo, the State University of New York (Agency for Toxic Substances and Disease Registry (ATSDR), Grant H75-ATH 298338, and the Great Lakes Protection Fund, Grant H75ATH 298338). The protocol for this study was reviewed and approved by the Institutional Review Boards for Protection of Human Subjects of the New York State Department of Health and the University at Buffalo, the State University of New York (protocol #s 05-015 and SPM0711204E, respectively).  Corresponding author. Fax: +518 402 7819. E-mail address: [email protected] (H.M. Spliethoff).

0013-9351/$ - see front matter & 2008 Elsevier Inc. All rights reserved. doi:10.1016/j.envres.2008.07.009

product matrices physically, rather than chemically (Rahman et al., 2001). Like many other persistent organic pollutants (POPs), PBDEs are lipophilic (Braekevelt et al., 2003) and bioaccumulative, with relatively high concentrations measured in fish and higher trophic level wildlife (Loganathan et al., 1995; Johnson-Restrepo et al., 2005a; Muir et al., 2006). PBDEs began to attract considerable attention from the scientific and regulatory communities, after the discovery that levels in women’s breast milk in Sweden had risen exponentially between 1972 and 1997 (Meironyte et al., 1999). A growing number of studies have since reported on PBDEs in human specimens from around the world, with levels in North America generally more than an order of magnitude higher than elsewhere (Hites, 2004; Sandanger et al., 2007; Johnson-Restrepo et al., 2005b; Sjo¨din et al., 2008). Endocrine disruption, developmental effects, and neurotoxicity have been observed in exposed laboratory animals (Birnbaum and Staskal, 2004), prompting investigators to search for associations in human populations (Hagmar et al., 2001; Bloom et al., 2008).

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In response to concerns about the more persistent PBDE congeners, production and use of two commercial mixtures, referred to as the penta- and octa-bromodiphenyl ether (pentaand octa-BDE) mixtures, have been prohibited in Europe and Japan for several years, and the sole US manufacturer voluntarily ceased production in 2004. New York State (NYS) banned the use of the penta- and octa-BDE mixtures in 2006, and other states have passed comparable legislation. The deca-BDE mixture was recently banned by the European Union’s highest court, though it continues to be used in the US. Similar to other ubiquitous and bioaccumulative POPs, PBDEs can be found in animal-based human foods (Schecter et al., 2006a). Market-basket data suggest that fish have the highest levels of PBDEs (Luksemburg et al., 2004), nearly 3 times the average level in meats and 10 times that in dairy products (Schecter et al., 2006a). Still higher PBDE levels have been reported for sportfish and other non-market-basket fish (Johnson and Olson, 2001; Hale et al., 2001; Rice et al., 2002). Levels of PBDEs reported for Great Lakes sportfish are among the highest reported thus far; some SPBDE levels in lake trout exceeded 200 ng/g wet weight (ww) (Zhu and Hites, 2004). Fish PBDE levels exhibit a spatial pattern across the Great Lakes similar to that for PCBs (Carlson and Swackhamer, 2006), and levels of PBDEs and PCBs can be highly correlated in individual fish (Manchester-Neesvig et al., 2001). Although some estimates indicate that fish consumption contributes nearly half of the average daily intake of PBDEs (Gill et al., 2004), others have concluded that meat consumption, because of its higher frequency, is a greater source of exposure in the US (Schecter et al., 2006a). A more recent assessment of PBDE exposure concluded that dietary exposure was responsible for only a small fraction of the PBDE blood level in the US (Lorber, 2008). Unlike other POPs, PBDEs can have significant nonoccupational exposure sources other than food, such as household dust (Jones-Otazo et al., 2005). Despite numerous studies investigating levels of PBDEs in humans, few have found strong associations between specific exposure sources and body burden. Consumption of fatty Baltic Sea fish was highly correlated with plasma BDE-47 (Sjo¨din et al., 2000), as was fish/shellfish consumption in Japan with breast milk PBDEs (Ohta et al., 2002). High rates of consumption of highly contaminated fish from a Lake in Norway were associated with elevated PBDE serum levels (Thomsen et al., 2008). In contrast, a study of Swedish fishermen’s wives found no associations between fish consumption and serum PBDEs (Weiss et al., 2006), nor did a study of metropolitan New York City (NYC) area anglers (Morland et al., 2005). Few biomonitoring studies have collected information related to dust PBDE exposure, but recently Wu et al. (2007) reported statistically significant associations of breast milk PBDE levels with levels in house dust, as well as with dairy and meat consumption. Sportfish consumption has been considered a significant route of exposure to POPs for residents of the Great Lakes Basin, and studies have reported associations between Great Lakes sportfish consumption and human body burden (Fitzgerald et al., 2007). The current preliminary, cross-sectional study of a sample of 36 NYS anglers was intended to generate hypotheses regarding potential associations between consumption of sportfish and PBDE body burden. 2. Materials and methods 2.1. Sample selection The study sample, recruitment, and specimen collection were previously described (Bloom et al., 2006). Briefly, plasma specimens were collected between 1995 and 1997 from 38 participants sampled from a subset (n ¼ 308) of

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participants in the New York State Angler Cohort Study (NYSACS) (Bloom et al., 2005). The NYSACS (n ¼ 18,082) was a prospective investigation of health effects in consumers of Great Lakes sportfish, among licensed anglers residing in 16 NYS counties contiguous to Lakes Erie and Ontario (Vena et al., 1996). The current sample was selected to maximize the range of sportfish-consumption-related exposure to POPs, on the basis of both self-reported sportfish consumption, and previously measured blood levels of PCB-153 (the most prevalent PCB congener in Lake Ontario sportfish (Niimi and Oliver, 1989)). Participants included individuals who (i) had consumed fish from Lake Ontario, and had PCB-153 levels in the highest quartile of the concentration distribution for sera procured in 1993–1994 (‘‘high consumers’’) (n ¼ 23) and (ii) reported never having consumed sportfish from Lake Ontario, and had serum PCB-153 levels in the lowest quartile (‘‘low consumers’’) (n ¼ 15). Categorical dietary consumption frequency data were obtained from subjects by self-administered questionnaire. Reliability of sport fish consumption frequency data collected using this method had been evaluated previously by repeat interview, as well as collection of proxy data from spouses, and was determined to be acceptable (Li et al., 2005). Plasma specimens were collected following standard procedures, and archived at 70 1C at the Center for Preventive Medicine, University at Buffalo, the State University of New York (SUNY UB) until transfer on dry ice to the Wadsworth Center, NYS Department of Health (DOH) for PBDE analysis in January 2006. Specimen volume was insufficient for two subjects, who were subsequently excluded from the study. Approval was obtained for this study from the NYS DOH and SUNY UB Institutional Review Boards for Protection of Human Subjects.

2.2. Laboratory analysis Specimen preparation, cleanup, and analysis were based on a previously published method (Sjo¨din et al., 2004), with modifications to permit analysis of smaller specimen aliquots (i.e., 1 g). All glassware used for specimen preparation was rinsed with hexane and baked overnight at 250 1C. Plasma specimens were thawed and spiked with 10 ml of internal standard solution (Cambridge Isotope Laboratories, Andover, MA) with 7.5 ng/ml of each of the following 13C12-labeled congeners in methanol (MeOH): 2,4,40 -tribromodiphenyl ether (BDE-28); 2,20 , 4,40 tetraBDE (BDE-47); 2,20 , 4,40 , 5-pentaBDE (BDE-99); 2,20 , 4,40 , 6-pentaBDE (BDE100); 2,20 , 4,40 , 5,50 -hexaBDE (BDE-153); and 2, 20 4, 40 , 5,60 -hexaBDE (BDE-154). After 1 h, formic acid (1 g) was added, and specimens were vortexed, sonicated, diluted with hexane-extracted HPLC grade water (2 g), and vortexed again. Automated extraction was performed by a Zymark Rapid Traces SPE workstation. Polypropylene SPE cartridges (3 ml), packed with 1.3 g C18 sorbent (Sepra C18-E, Phenomenex; bed height 31 mm), were sequentially conditioned with 3 ml MeOH, 3 ml of 0.1 N hydrochloric acid (HCl) in 5% MeOH/water, 3 ml dichloromethane, and 3 ml 0.1 N HCl in 5% MeOH/water. Specimens and reagents (4 g total) were loaded and rinsed with 1 ml of 0.1 N HCl in 5% MeOH/water. A constant flow of nitrogen dried the cartridge for 30 min prior to elution of extract with 12 ml of 30% dichloromethane/hexane, and concentration to 1 ml at 40 1C with nitrogen using a Zymark TurboVaps LV. Coextracted lipids were removed by elution with 11 ml hexane through a second 3 ml SPE cartridge (1 g silica gel/ sulfuric acid (2:1 w/w), layer of activated silica gel (0.1 g), pre-conditioned with 10 ml hexane). Extracts were concentrated in the TurboVap, transferred to glass inserts in amber GC vials, and reduced to 40 ml, final volume. Prior to analysis, 10 ml of recovery standard spiking solution (Cambridge Isotope Laboratories, Andover, MA), with 7.5 ng/ml of 13C-labeled 3,30 ,4,40 -tetraBDE (BDE-77) and 2,20 , 3,4,40 ,6-hexaBDE (BDE-139), was added to each extract. To determine a relative response factor of a native PBDE congener vs. its analogous 13 C12-PBDE congener, a calibration standard was prepared containing: native BDEs 28, 47, 66 (2,30 ,4,4’-tetraBDE), 85 (2,20 ,3,4,40 -penta-BDE), 99, 100, 138 (2,20 ,3,4,40 ,50 hexaBDE), 153, and 154 (AccuStandard, New Haven CT); the six 13C12-PBDE congeners in the internal standard solution; and the two 13C12-PBDE congeners in the recovery standard solution. A seven-point calibration curve was run with 13C12PBDE congeners at 1.5 pg/ml and the nine native congeners at 0.05–100 pg/ml. Extracts were analyzed using a Finnigan Trace gas chromatograph Ultra coupled with an MAT95XP high-resolution mass spectrometer (R10,000; 40 eV). Two ions with the highest intensities were measured for each native and labeled congener. A J&W Scientific DB-5msMSD column (30 m length, 0.25 mm ID, 0.25 mm film thickness) and a GC PAL auto sampler (splitless injection, Tinj ¼ 270 1C, 2 ml/ injection) were used with the GC. Oven temperature was ramped from 140 1C (1.5 min) to 205 1C (0 min) at 15 1C/min, and then to 300 1C (8.3 min) at 6 1C/min. Congeners that were quantified were BDE-28, BDE-47, BDE-66, BDE-85, BDE-99, BDE-100, BDE-138, BDE-153, and BDE-154. Analysis of decabromodiphenyl ether (BDE-209) was not feasible with the method employed in this study. For every batch of five specimen extracts, three QA/QC extracts were included: 1:7 dilution newborn calf serum (Gibco, Grand Island, NY, lot # 48407) as a method blank; human serum reference material (Sigma-Aldrich, St. Louis, MO, lot # 015k8906); and human serum reference material spiked with 250 pg/g of nine native PBDE congeners. Method blank levels of PBDE congeners were o6 pg/g, with the exceptions of BDE-47 (36 pg/g) and BDE-99 (25 pg/g). Average PBDE levels in the unspiked reference material samples ranged from 1 pg/g for BDE-138

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to 169 pg/g for BDE-47, with standard deviations o7 pg/g for all congeners, except BDE-47 (19 pg/g) and BDE-99 (35 pg/g). Relative standard deviations (RSD) of PBDE levels in the spiked reference materials were o12% for all congeners. Blank injections of isooctane in every analysis batch demonstrated negligible instrumental contamination and carry-over between injections. Each accepted congener measurement met the following criteria: (i) the ratio of the two ions was 720% of the theoretical value; and (ii) the retention time was 712 s from that of a known standard. Recoveries for all congeners and samples averaged 65%742% (RSD). The limit of detection (LOD) for each congener was determined as three standard deviations of concentrations measured for seven replicate extractions of diluted (1:7) newborn calf serum spiked at 7.5 pg/g. LODs ranged from 2.3 pg/g for BDE-138 to 28.2 pg/g for BDE-47. Coefficients of variation (CV) within duplicate extractions of three specimens chosen randomly demonstrated excellent precision for eight of the nine congeners, ranging from 1% for BDE-47 to 7% for BDE-154; the CV for BDE-138 was 27%. Specimen concentrations were corrected for analytical background measured in concurrently processed method blanks. No substitutions were made for concentrations oLOD (Richardson and Ciampi, 2003; Schisterman et al., 2006). Negative values, resulting from blank subtraction, were censored as zero for the purpose of data presentation only. Plasma lipid level was determined in a separate aliquot of plasma (0.5 g) that was denatured with 0.5 ml methanol, extracted with 1:1 ether/hexane (3  5 ml), and evaporated to dryness under vacuum. Residual lipid mass was measured gravimetrically with a high-precision electrobalance. Plasma lipid levels were reported in units of mg lipid/dl plasma, assuming a density of 1.0 g/ml. PBDE levels are presented as lipid-weight (ng/g lw), although all statistical analyses were carried out using wet weight (pg/g ww) levels to eliminate bias that can result from lipid correction (Schisterman et al., 2005).

2.3. Diet characterization Questionnaire responses were used to characterize consumption of sportfish, and other potential dietary PBDE exposure sources. Long-term sportfish consumption was characterized by the number of years eating Lake Ontario sportfish from 1980 to 1990 (Bloom et al., 2005). Recent sportfish consumption was characterized with responses indicating the (i) number of NYS sportfish meals over the year prior to participation, (ii) consumption of at least one Lake Ontario sportfish meal in the prior year, and (iii) consumption of at least one NYS sportfish meal in the prior month. Because of reported correlations between levels of PBDEs and other POPs in fish (Manchester-Neesvig et al., 2001) and well-characterized levels of several other POPs across Great Lakes species, an index value was operationalized to represent high-POP Lake Ontario sportfish consumption in the prior year. The index was the sum of categorical responses for the number of meals of each of 13 species/size classes listed for Lake Ontario in the NYS DOH Chemicals in Sportfish and Game Health Advisory due to elevated levels of PCBs, mirex, and dioxins (NYSDOH, 1995). To characterize other potential exposure sources (Luksemburg et al., 2004; Kearney et al., 1999; Wu et al., 2007; Schecter et al., 2006a), the following were used as variables: consumption of wild waterfowl in the prior year, and the number of meals over the prior year for small and large ocean (commercial) fish, dairy products, and beef. Age, gender, and body mass index (BMI) were considered as potential confounders.

2.4. Statistical analysis All statistical analysis was carried out using SAS v. 9.1 (SAS Institute Inc., Cary, NC). PBDE levels were not normally distributed, and non-parametric tests were used for all bivariate analyses. Differences between high and low consumers were assessed for ordinal and continuous variates and PBDE levels with the Mann–Whitney U test, and for dichotomous variates with w2 or Fisher’s exact test. Associations between dietary variates, potential confounders, and PBDE levels were assessed with Spearman rank correlation. Statistical significance was defined as Po0.05 for a two-tailed test. No adjustments were made for potential inflation of type-1 error, so as not to impede the study goal of hypothesis generation. Dietary variates demonstrating associations with any congener or SPBDE (Po0.05) were considered potential predictors for SPBDE in a forward stepwise selection linear regression model. Prior to regression analysis, PBDE levels (offset to eliminate negative values) were natural-log transformed to stabilize the variance of the distribution, and missing covariate values were replaced by the appropriate median (i.e., for whole population, or based on consumption-status or gender). Plasma lipid levels (centered by the mean) were forced into the regression model, and other variates sequentially allowed to enter if Po0.05, and remain, if Po0.10. Potential confounders (age, BMI, gender) were sequentially entered into the model, and retained only with a X10% change in predictor coefficient(s), or a statistically significant change in model fit. Studentized residuals and graphical evaluation were employed to identify influential observations for more thorough examination and possible exclusion (Kleinbaum et al., 1998). Relative importance of contributions of individual congeners to observed associations between SPBDE

and predictor variates was evaluated with a series of linear models, one for each congener, using the SPBDE model variate(s).

3. Results High and low consumers of Lake Ontario sportfish exhibited similar plasma lipid levels, demographic characteristics, and, except for sportfish and wild waterfowl consumption, similar dietary histories (Table 1). High consumers were markedly different from low consumers with respect to sportfish consumption, reporting a median of 7–11 (maximum, 24–36) and 0 (maximum, 1–6) NYS sportfish meals, respectively, in the year prior to study participation. Sixty-eight percent of high consumers ate Lake Ontario sportfish in the prior year vs. 23% of low consumers. Thirty-six percent of high consumers ate NYS sportfish in the prior month vs. 0% of low consumers. Ten high and no low consumers ate one or more high-POP sportfish meals in the prior year (data not shown). Maximum levels of the highPOP sportfish consumption index were 8 for high consumers and 0 for low consumers; the medians of both groups were 0. Seventyseven percent of high consumers and 31% of low consumers reported eating wild waterfowl in the prior year. Differences between high consumers and low consumers were statistically significant for wild waterfowl consumption and all sportfish consumption variates. Levels of BDE-28, BDE-47, BDE-99 BDE-100, and SPBDE were statistically significantly higher for high consumers than for low consumers (Table 2). High consumers also had higher maximum levels of all congeners than low consumers, and statistically significantly greater proportions of results 4LOD for BDE-47, BDE100, and BDE-153. Despite these differences, the relative contribution of each congener to SPBDE was similar for both groups. Nearly 60% of SPBDE for both high and low consumers was contributed by BDE-47, followed by BDE-99 at nearly 20%, BDE100 at around 10%, and BDE-153 at less than 7% (Fig. 1). The number of years of reported consumption of Lake Ontario fish between 1980 and 1990 was statistically significantly correlated with SPBDE and all congeners, except BDE-28, BDE66, and BDE-138. Prior year’s consumption of high-POP Lake Ontario sportfish was statistically significantly correlated with levels of SPBDE and all congeners except BDE-138, BDE-153, and BDE-66 (Table 3). A positive correlation was found between BDE47 and the number of NYS sportfish meals in the prior year. Neither NYS sportfish consumption in the prior month nor Lake Ontario sportfish consumption in the prior year was associated with any PBDE congener level. Statistically significant correlations between PBDE levels and other non-sportfish dietary variables were not found, with the exception of BDE-47 and the number of large ocean fish meals in the prior year (Table 4). Plasma lipid levels were statistically significantly correlated with wild waterfowl consumption (rSp ¼ 0.40), and with the prior year’s number of dairy product meals (rSp ¼ 0.35). Following elimination of one highly influential observation, stepwise regression revealed the sum of years consuming Lake Ontario sportfish between 1980 and 1990, adjusted for plasma lipid levels, as the sole statistically significant predictor (b ¼ 0.130, 95% CI: 0.007–0.254; n ¼ 35) for SPBDE (Table 5, Fig. 2). There was no evidence of confounding by age, BMI, or gender. The multiple linear regression model, including years consuming Lake Ontario sportfish between 1980 and 1990, and plasma lipid levels, explained 13% of the variance in the sample, although this value was not statistically significantly different from zero (P ¼ 0.114). Individual linear regression models for each congener, using years consuming Lake Ontario sportfish between 1980 and 1990, and plasma lipid levels, suggested that BDE-47 and BDE-99 contributed most to the association observed for SPBDE.

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Table 1 Characteristics of high consumers and low consumers of Lake Ontario sportfish (New York State Angler Cohort Study (NYSACS)) Demographic and dietary characteristics

High consumersa (n ¼ 22)

Low consumersa (n ¼ 14)

P-valueb

Age—median (range) BMI (kg/m2)—median (range) Plasma lipid concentration (mg/dl)—median (range) Gender female—# (%) Small ocean fish, # meals over prior year—median (range) Large ocean fish, # meals over prior year—median (range) Dairy products, # meals over prior year—median (range) Beef, # meals over prior year—median (range) NYS sportfish, # meals over prior year—median (range) Consumed Lake Ontario fish in prior year—n (%) Consumed NYS sportfish in prior month—n (%) Consumed wild waterfowl in prior year—n (%) Years consuming Lake Ontario sportfish, # between 1980 and 1990—median (range) High-POP Lake Ontario sportfish in prior year, consumption index value—median (range)f

40 (29-44) 25.9 (18.7–34.4) 450 (280–680) 2 (9) 7–11/yr (none to 1/wk) None (none to 3/mo)d 45/wk (1–6/yr to 45/wk) 2/wk (1–6/yr to 3–4/wk) 7–11/yr (1–6/yr to 2–3/mo) 15 (68) 8 (36) 17 (77) 7 (none to 11)e 0 (0–8)

38 (31–45) 26.2 (20.8–29.3)c 425 (230–730) 4 (29) 7–11/yr (none to 3/wk)c None (none to 1–6/yr)c 45/wk (1/wk to 45/wk) c 2/wk (7–11/yr to 45/wk) None (none to 1–6/yr) 3 (23) 0 (0)c 4 (31) None (none to 2) 0

0.784 0.905 0.721 0.181 0.686 0.403 0.780 0.389 o0.001 0.010 0.015 0.007 o0.001 0.005

a High consumers reported a history of consumption of Lake Ontario sportfish and demonstrated serum PCB-153 concentrations in the highest quartile in 1993–1994; low consumers reported no consumption of Lake Ontario sportfish and demonstrated PCB-153 concentrations in the lowest quartile. b P-values for statistical comparisons between high consumers and low consumers. c n ¼ 13. d n ¼ 21. e n ¼ 19. f Consumption index: sum of category values for number of prior year’s meals of American eel, brown trout 42000 , brown trout p2000 , carp, channel catfish, chinook salmon, coho salmon 42100, coho salmon p2100, lake trout, rainbow trout 42500 , rainbow trout p2500 , white sucker, white perch (NYSDOH, 1995).

BDE congener

BDE-28 BDE-47 BDE-66 BDE-85 BDE-99 BDE-100 BDE-138 BDE-153 BDE-154 SPBDE

High consumersa (n ¼ 22)

Low consumersa (n ¼ 14)

Median Range

% 4LOD

Median Range

% 4LOD

0.6 12.8 0.2 0.4 3.4 3 0 3 0.6 23.1

37 64* 23 41 45 64* 18 64* 27

0 1.3 0 0 0 0 0 0.7 0 1.9

21 21* 14 14 14 21* 14 29* 14

0–18.6 0–1378.4 0–6.8 0–44.8 0–430.1 0–247.5 0–7.6 0–143.3 0–33.3 0–2303

0–13.6 0–624.3 0–5.4 0–22.6 0–225.4 0–105.6 0–2.6 0–51.8 0–18.7 0–1069

Pvalueb

0.050 0.009 0.068 0.162 0.025 0.012 0.832 0.064 0.098 0.015

Note that concentrations were significantly different (Po0.05) for high consumers and low consumers for the same congeners, whether wet weight PBDE concentrations or lipid weight PBDE concentrations were compared. *Po0.05, based on difference in proportion of concentrations 4LOD; LOD, limit of detection. a High consumers reported a history of consumption of Lake Ontario sportfish and demonstrated serum PCB-153 concentrations in the highest quartile in 1993–1994; low consumers reported no consumption of Lake Ontario sportfish and demonstrated PCB-153 concentrations in the lowest quartile. b Difference between unadjusted concentration (pg/g ww) distributions for high consumer and low consumer groups.

4. Discussion High consumers of Lake Ontario sportfish, when compared with low consumers, were found to have statistically significantly higher plasma levels of BDE-47, BDE-99, BDE-100, and SPBDE. Maximum concentrations were higher for all congeners and SPBDE, and the number of concentrations 4LOD were statistically significantly higher, among high consumers, for BDE-47, BDE-100, and BDE-153. The number of years consuming sportfish from Lake Ontario and the consumption of high-POP Lake Ontario sportfish in the year prior to study participation were significantly and positively correlated with PBDE levels. Multiple linear regression indicated that the number of years consuming

80 Relative contribution to ∑PBDEs (%)

Table 2 PBDE congener plasma concentrations (ng/g lw) and percentage of concentrations 4LOD for high sportfish consumers (n ¼ 22) and low sportfish consumers (n ¼ 14) (NYSACS)

70 High consumers Low consumers

60 50 40 30 20 10 0

28

47

66

85

99

100

138

153

154

PBDE congener number Fig. 1. Congener profiles of high sportfish consumers and low sportfish consumers (relative contributions (%) to mean SPBDE) (NYSACS). Error bars: standard error.

sportfish from Lake Ontario between 1980 and 1990, adjusted for plasma lipid levels, was the sole statistically significant predictor of plasma SPBDE levels. These findings suggest that duration of sportfish consumption, and possibly, the magnitude of consumption of specific, potentially high-concentration species, are important variables to consider for evaluation of impacts of sportfish consumption on human PBDE body burden. Correlations with long-term sportfish consumption, and the lack of correlations with the prior month’s consumption, are consistent with multi-year reported half-lives for these compounds (Geyer et al., 2004). The significance of highPOP sportfish consumption is consistent with reported correlations between PBDEs and other POPs in Great Lakes sportfish (Manchester-Neesvig et al., 2001). Few studies that have examined human PBDE body burden and fish consumption are available for comparison to our results.

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Table 3 Spearman correlations between PBDE congener concentrations (pg/g ww) and measures of sportfish consumption in study population (n ¼ 36) (NYSACS) BDE congener

NYS sportfish, # meals over Prior yeara

NYS sportfish, consumption in prior month (yes/no)a

Lake Ontario sportfish, consumption in prior year (yes/no)a

High-POP Lake Ontario sportfish, consumption, index for prior yeara,b

Lake Ontario sportfish, # years consuming between 1980 and 1990c

BDE-28 BDE-47 BDE-66 BDE-85 BDE-99d BDE-100 BDE-138 BDE-153e BDE-154 SPBDE

0.32 0.36* 0.23 0.12 0.28 0.28 0.14 0.19 0.23 0.32

0.12 0.18 0.06 0.05 0.12 0.16 0.13 0.08 0.14 0.18

0.15 0.18 0.23 0.13 0.03 0.13 0.18 0.12 0.07 0.11

0.41* 0.46** 0.28 0.37* 0.36* 0.43* 0.33 0.28 0.36* 0.44**

0.32 0.55** 0.2 0.47** 0.47** 0.5** 0.15 0.51** 0.43* 0.52**

*Po0.05, **Po0.01. a n ¼ 35. b Sum of category values for number of meals in prior year of American eel, brown trout 42000 , brown trout o2000 , carp, channel catfish, chinook salmon, coho salmon 42100, coho salmon o2100, lake trout, rainbow trout 42500 , rainbow trout o2500 , white sucker, white perch (NYS DOH, 1995). c n ¼ 33. d Associated with age (rSp ¼ 0.34, Po0.05). e Associated with male gender (Po0.01).

Table 4 Correlations between BDE congeners (pg/g ww) and other potential dietary exposure sources in study population (n ¼ 36) (NYSACS)

BDE-28 BDE47 BDE-66 BDE-85 BDE-99e BDE-100 BDE-138 BDE-153f BDE-154 SPBDE

Wild waterfowl, consumption (yes/no) in prior yeara,b

Large ocean fish, # meals over prior yearc

Small ocean fish, # meals over prior yeara

Dairy products, # meals over prior yeara,d

Beef, # meals over prior yeara

0.07 0.04 0.00 0.08 0.07 0.08 0.30 0.09 0.04 0.06

0.31 0.4* 0.33 0.12 0.24 0.25 0.07 0.16 0.06 0.30

0.10 0.09 0.01 0.04 0.13 0.10 0.03 0.06 0.10 0.14

0.07 0.17 0.01 0.28 0.19 0.19 0.31 0.25 0.29 0.18

0.22 0.16 0.05 0.27 0.24 0.21 0.00 0.15 0.15 0.22

2

PBDE congener

b (95% CI)

P-value

Model r (P-value)

RPBDE

0.130 0.008 0.081 0.003 0.017 0.067 0.036 0.003 0.030 0.016

0.039 0.396 0.109 0.438 0.207 0.103 0.212 0.301 0.187 0.177

0.13 0.03 0.08 0.02 0.05 0.08 0.05 0.04 0.06 0.06

BDE-28 BDE-47 BDE-66 BDE-85 BDE-99 BDE-100 BDE-138 BDE-153 BDE-154

(0.007, 0.254) (0.009, 0.023) (0.019, 0.181) (0.025, 0.036) (0.010, 0.044) (0.014, 0.148) (0.022, 0.094) (0.003, 0.010) (0.015, 0.075) (0.007, 0.039)

(0.114) (0.639) (0.270) (0.700) (0.415) (0.259) (0.441) (0.537) (0.403) (0.379)

Note: Bold font indicates statistically significant regression coefficient.

Norwegian anglers who reported high rates of consumption of fish (sum of species-specific intake rates for men of 53 g/day) from a contaminated lake had higher serum PBDE levels (men, median SPBDE 18 ng/g lw) than a referent population (median SPBDE

Plasma

Table 5 Linear regression models of log-transformed PBDE plasma concentration (pg/ g ww) on years consuming Lake Ontario sportfish between 1980 and 1990, adjusted for plasma lipids (n ¼ 35) (NYSACS)

PBDE concentration (In pg/g ww)

*Po0.05. a n ¼ 35. b Associated with plasma lipid level (rSp ¼ 0.40, P ¼ 0.016). c n ¼ 34. d Associated with plasma lipid level (rSp ¼ 0.35, P ¼ 0.039). e Associated with age (rSp ¼ 0.34, Po0.05). f Associated with male gender (Po0.01).

10

8

6

4

2

0 0

1

2

3

4

5

6

7

8

9

10

11

Years of Lake ontario sportfish consumption (1980-1990), lipid adjusted Fig. 2. Linear regression (95% confidence interval) of natural log SPBDE plasma concentration on lipid-adjusted years of Lake Ontario sportfish consumption (1980–1990) (NYSACS) (n ¼ 35).

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4 ng/g lw). Latvian men with high fish consumption (median ¼ 19 meals/month) were found to have plasma levels of BDE-47 9-fold higher (median ¼ 2.4 ng/g lw) than men with low or no fish consumption (median ¼ 0 meals/month and 0.26 ng/g lw) (Sjo¨din et al., 2000). Nursing women in Japan who consumed fish daily had statistically significantly higher breast milk levels of SPBDE (1.72 ng/g lw), than did women who consumed fish only 1–2 times per week (0.77 ng/g lw) (Ohta et al., 2002). A study of NYC anglers found subjects consuming 41 local-fish meal per week had geometric mean BDE-47 serum levels (14.37 ng/g lw) similar to those of subjects with no local-fish consumption (12.61 ng/g lw) (Morland et al., 2005). No statistically significant positive association between fish consumption and serum PBDEs was found among Swedish fishermen’s wives who consumed a median of two meals of fatty Baltic Sea fish per month (Weiss et al., 2006). For high consumers in our study, the rate of NYS sportfish consumption (median 7–11 meals in prior year) is similar to rates reported in other Great Lakes fish consumption surveys (Connelly et al., 1996), but considerably lower than rates reported in other studies that examined PBDE levels in blood or breast milk. On an annualized basis, Weiss et al. (2006) reported a median consumption rate of 24 meals of fatty Baltic Sea fish for their study population (n ¼ 50); Morland et al. (2005) reported 452 meals of local (ocean) fish for their highest-consumption category (n ¼ 54); Sjo¨din et al. (2000) reported 4144 meals of fatty Baltic Sea fish for high fish consumers (n ¼ 37); and Ohta et al. (2002) reported consumption of 365 meals of fish and shellfish for their high fish consumption category (n ¼ 5). Although Thomsen et al. (2008) did not report the number of fish meals for their participants, an average daily sportfish intake of 53 g suggests that the average number of fish meals per year may have been around one hundred or more. Despite much smaller differences in fish consumption between our high and low consumers compared to those reported for the Latvian men (Sjo¨din et al., 2000) and the Japanese women (Ohta et al., 2002), there were greater differences in PBDE levels between the two groups in the current study (e.g., median level of BDE-47 was higher in high consumers than low consumers by 11.5 ng/g lw, or a factor of 10). This disparity may be due in part to differences in levels of PBDEs in fish consumed in these studies. Concentrations of PBDEs in fish from Lake Ontario have only been reported for lake trout thus far, while reported concentrations for Lake Erie fish have been limited to walleye and, to a lesser extent, lake trout. Total PBDE concentrations in Lake Ontario lake trout increased exponentially from 2 ng/g ww in 1980 to 110 ng/g ww in 1996, while levels in Lake Erie walleye increased from 1 to 24 ng/g ww over the same time period (Zhu and Hites, 2004). Congener distributions in both fish species were generally similar over this time period, with more than half the total PBDE concentration consisting of BDE-47, while BDE-99 and BDE-100 together accounted for approximately 25% of the total. Levels of BDE-47 in lake trout sampled from Lake Ontario in 1996 averaged 60 ng/g ww (Zhu and Hites, 2005). In contrast, BDE-47 concentrations in fatty Baltic Sea fish sampled in 1995 such as those that might have been eaten by the Baltic fishermen were considerably lower, at 9.4 ng/g ww (Asplund et al., 1999). Fish and shellfish available for consumption by the Japanese study subjects (Ohta et al., 2002) were found to contain only 0.02–1.7 ng/g ww for SPBDE. However, fish consumed by the Norwegian anglers were highly contaminated (51.5–344 ng/g ww). Their fish consumption rates were also very high, and the reported elevation in participant serum levels was greater than that reported in our study. While the PBDE blood levels reported for high consumers in the current study and the NYC angler study by Morland et al. (2005) were similar, the low consumers in our study had

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substantially lower PBDE levels than the NYC low consumers. The comparatively lower PBDE levels in our low consumers could be due in part to the earlier sampling of our population (1995–1997) compared with the NYC population (2001–2003). Measured serum PBDEs in the US population in 2003–2004, shortly after the NYC specimens, were collected, which demonstrated that background levels were much higher around this time period, particularly for BDE-47 (geometric mean ¼ 19.2 ng/g lw) (Sjo¨din et al., 2008). The greater difference between PBDE levels for high consumers and low consumers observed in our study that in the NYC population could be due to the relatively higher Great Lakes sportfish PBDE levels when compared with most other reported fish concentrations. While the greater differences in median levels of PBDEs between high consumers and low consumers in our study, compared to other studies, could be due in part to the higher levels of PBDEs in Lake Ontario sportfish, they are not entirely supported by estimated contributions to PBDE body burden at the reported rates of fish consumption. Contribution to PBDE body burden at the rates reported for high consumers can be approximated using available fish concentration data, an assumption of steady-state exposure, and a simple pharmacokinetic (PK) model. For example, levels of BDE-47 measured in lake trout sampled from Lake Ontario in 1996 averaged 60 ng/g ww (Zhu and Hites, 2005). Assuming a consumption rate of 9 meals/year with an average meal size of 227 g (US EPA, 1995), and the median participant body weight (81 kg), a daily exposure of 3.8 ng/kg/day can be estimated. Using this exposure and a single-compartment model, assuming body burden is entirely in lipid (20% of body weight), and first-order elimination kinetics with a half-life of 4.9 years for males (Geyer et al., 2004), the contribution to body burden from this exposure would only be 0.63 ng/g lw. This is considerably lower than the observed difference between median high and low consumer concentrations of 11.5 ng/g lw. Consumption of commercial large ocean fish was statistically significantly correlated with a single congener, BDE-47, although the relatively low consumption rates and the lack of significance in the regression model suggest this exposure did not substantially contribute to PBDE plasma levels in this study sample. While it has been suggested that dairy and meat consumption can be significant exposures for PBDEs (Schecter et al., 2006a b), we found no positive associations. The higher relative PBDE plasma levels associated with the consumption of Lake Ontario fish may reflect other confounding exposures not accounted for in this study (e.g., to household dust). The stratified recruitment strategy for study participants may have resulted in a selection bias contributing to observed associations with sportfish consumption. However, weak or absent correlations between PCBs and PBDEs in human populations (Sandanger et al., 2007; Schecter et al., 2005) suggest that such a bias was not present in our study. Furthermore, even in an analysis restricted to the high consumers in the present study (data not shown), years of consumption of Lake Ontario sportfish was still statistically significantly correlated with SPBDE (b ¼ 0.51, P ¼ 0.025). Decabromodiphenyl ether was not measured in this study. However, this congener may not have been present in fish in the US in the 1990s at measurable concentrations (Dodder et al., 2002), suggesting that BDE-209 is not among the more important congeners for an assessment of fish consumption as a source of PBDE exposure. In summary, the results of this study suggest that long-term sportfish consumption can contribute measurably to PBDE body burden. The number of years consuming Lake Ontario sportfish, adjusted for plasma lipid levels, was a positive and statistically significant predictor of SPBDE plasma levels in our study sample. Nonetheless, our findings suggest that other, more substantial, sources of exposure exist for the study participants. Furthermore,

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the reported magnitude of the effect of sportfish consumption on PBDE levels incorporates significant uncertainty due to the small study sample size. For these reasons, the observed associations between sportfish consumption and PBDE plasma levels reported for this exploratory study would need to be confirmed through a more comprehensive investigation of a larger population sample.

Acknowledgments This research was funded in part by the Centers for Disease Control and Prevention (CDC) National Center for Environmental Health (NCEH) State Biomonitoring Implementation Program Grant U59CCU22339202, the Agency for Toxic Substances and Disease Registry (ATSDR), Grant H75-ATH 298338, and the Great Lakes Protection Fund, Grant RM 791-3021. The authors are grateful to Adriana Verschoor for her thoughtful review of this manuscript, and Craig Diederich for help with retrieval of archived specimens. References Asplund, L., Athanasiadou, M., Sjo¨din, A., Bergman, A., Borjeson, H., 1999. Organohalogen substances in muscle, egg, and blood from healthy Baltic salmon (salmo salar) and Baltic salmon that produced offspring with the M74 syndrome. Ambio 28, 67–76. Birnbaum, L.S., Staskal, D.F., 2004. Brominated flame retardants: cause for concern? Environ. Health Perspect. 112 (1), 9–17. Bloom, M., Vena, J.E., Swanson, M.K., Moysich, K.B., Olson, J.R., 2005. Profiles of ortho-polychlorinated biphenyl congeners, dichlorordiphenyldichoroethylene, hexachlorobenzene, and mirex among male Lake Ontario sportfish consumers: the New York State Angler Cohort Study. Environ. Res. 97 (2), 178–194. Bloom, M., Vena, J., Olson, J., Moysich, K., 2006. Chronic exposure to dioxin-like compounds and thyroid function among New York anglers. Environ. Toxicol. Pharmacol. 21 (3), 260–267. Bloom, M., Spliethoff, H., Vena, J., Shaver, S., Addink, R., Aldous, K., Eadon, G., 2008. Chronic exposure to PBDEs and thyroid function among New York anglers. Environ. Toxicol. Pharmacol. 25, 386–392. Braekevelt, E., Tittlemier, S.A., Tomy, G.T., 2003. Direct measurement of octanol– water partition coefficients of some environmentally relevant brominated diphenyl ether congeners. Chemosphere 51 (7), 563–567. Carlson, D.L., Swackhamer, D.L., 2006. Results from the US Great Lakes fish monitoring program and effects of lake processes on bioaccumulative contaminant concentrations. J. Great Lakes Res. 32, 370–385. Connelly, N.A., Knuth, B.A., Brown, T.L., 1996. Sportfish consumption patterns of Lake Ontario anglers and the relationship to health advisories. North Am. J. Fish. Manage. 16, 90–101. Dodder, N.G., Strandberg, B., Hites, R.A., 2002. Concentrations and spatial variations of polybrominated diphenyl ethers and several organochlorine compounds in fishes from the Northeastern United States. Environ. Sci. Technol. 36, 146–151. Fitzgerald, E.F., Hwang, S.A., Gomez, M., Bush, B., Yang, B.Z., Tarbell, A., 2007. Environmental and occupational exposures and serum PCB concentrations and patterns among Mohawk men at Akwesasne. J. Expo. Sci. Environ. Epidemiol. 17, 269–278. Geyer, H.J., Schramm, K., Darnerud, P.O., Aune, M., Feicht, E.A., Fried, K.W., et al., 2004. Terminal elimination half-lives of the brominated flame retardants TBBPA, HBCD, and lower brominated PBDEs in humans. Organohalogen Compds. 66, 3867–3871. Gill, U., Chu, I., Ryan, J.J., Feeley, M., 2004. Polybrominated diphenyl ethers: human tissue levels and toxicology. Rev. Environ. Contam. Toxicol. 183, 55–97. Hagmar, L., Bjo¨rk, J., Sjo¨din, A., Bergman, A., Erfurth, E., 2001. Plasma levels of persistent organohalogens and hormone levels in adult male humans. Arch. Environ. Health. 56 (2), 138–143. Hale, R.C., La Guardia, M.J., Harvey, E.P., Mainor, T.M., Duff, W.H., Gaylor, M.O., 2001. Polybrominated diphenyl ether flame retardants in virginia freshwater fishes (USA). Environ. Sci. Technol. 35, 4585–4591. Hites, R.A., 2004. Polybrominated diphenyl ethers in the environment and in people: a meta-analysis of concentrations. Environ. Sci. Technol. 38, 945–956. Hooper, K., She, J., 2003. Lessons from the polybrominated diphenyl ethers (PBDEs): precautionary principle, primary prevention, and the value of community-based body-burden monitoring using breast milk. Environ. Health Perspect. 111 (1), 109–114. Johnson, A., Olson, N., 2001. Analysis and occurrence of polybrominated diphenyl ethers in Washington State freshwater fish. Arch. Environ. Contam. Toxicol. 41, 339–344. Johnson-Restrepo, B., Kannan, K., Addink, R., Adams, D.H., 2005a. Polybrominated diphenylethers (PBDEs) and polychlorinated biphenyls (PCBs) in a marine foodweb of coastal Florida. Environ. Sci. Technol. 39 (21), 8243–8250.

Johnson-Restrepo, B., Kannan, K., Rapaport, D.P., Rodan, B.D., 2005b. Polybrominated diphenyl ethers and polychlorinated biphenyls in human adipose tissue from New York. Environ. Sci. Technol. 39 (14), 5177–5182. Jones-Otazo, H.A., Clarke, J.P., Diamond, M.L., Archbold, J.A., Ferguson, G., Harner, T., et al., 2005. Is house dust the missing exposure pathway for PBDEs? An analysis of the urban fate and human exposure to PBDEs. Environ. Sci. Technol. 39, 5121–5530. Kearney, J., Cole, D., Ferron, L., Weber, J.P., 1999. Blood PCB, p, p(prime)-DDE, and mirex levels in great lakes fish and waterfowl consumers in two Ontario communities. Environ. Res. 80 (2), S138. Kleinbaum, D.G., Kupper, L.L., Muller, K.E., Nizham, A., 1998. Applied Regression Analysis and Other Multivariable Methods. Duxbury Press, Pacific Grove. Li, Q., Vena, J.E., Swanson, M.K., 2005. Reliability of sport fish consumption in the New York State Angler cohort study. Environ. Res. 97 (2), 142–148. Loganathan, B.G., Kannan, K., Watanabe, I., Kawano, M., Irvine, K., Kumar, S., Sikka, H.C., 1995. Isomer-specific determination of polychlorinated biphenyls, polychlorinated/brominated dibenzo-p-dioxins, dibenzo furans, polybrominated biphenyl ethers and extractable organic halogen in carp from the Buffalo River, New York. Environ. Sci. Technol. 29, 1832–1838. Lorber, M., 2008. Exposure of Americans to polybrominated diphenyl ethers. J. Expo. Sci. Environ. Epidemiol. 18 (1), 2–19. Luksemburg, W., Wenning, R., Maier, M., Patterson, A., Braithwaite, S., 2004. Polybrominated diphenyl ethers (PBDE) and polychlorinated dibenzo-pdioxins (PCDD/F) and biphenyls (PCB) in fish, beef, and fowl purchased in food markets in Northern California (USA). Organohalogen Compds. 66, 3982–3987. Manchester-Neesvig, J.B., Valters, K., Sonzogni, W.C., 2001. Comparison of polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) in Lake Michigan salmonids. Environ. Sci. Technol. 35 (6), 1072–1077. Meironyte, D., Noren, K., Bergman, A., 1999. Analysis of polybrominated diphenyl ethers in Swedish human milk. A time-related trend study, 1972–1997. J. Toxicol. Environ. Health-Part A 58 (6), 329–341. Morland, K.B., Landrigan, P.J., Sjo¨din, A., Gobeille, A.K., Jones, R.S., McGahee, E.E., Needham, L.L., Patterson Jr., D.G., 2005. Body burdens of PBDEs among urban anglers. Environ. Health Perspect. 113, 1689–1692. Muir, D.C., Backus, S., Derocher, A.E., Dietz, R., Evans, T.J., Gabrielsen, G.W., Nagy, J., Norstrom, R.J., Sonne, C., Stirling, I., Taylor, M.K., Letcher, R.J., 2006. Brominated flame retardants in polar bears (Ursus maritimus) from Alaska, the Canadian Arctic, East Greenland, and Svalbard. Environ. Sci. Technol. 40 (2), 449–455. Niimi, A.J., Oliver, B.G., 1989. Distribution of polychlorinated biphenyl congeners and other halocarbons in whole fish and muscle among Lake Ontario salmonids. Environ. Sci. Technol. 23 (1), 83–88. NYSDOH (New York State Department of Health), 1995. Chemicals in Sportfish and Game: Health Advisory. Albany, NY. Ohta, S., Ishizuka, D., Nishimura, H., Nakao, T., Aozasa, O., Shimidzu, Y., et al., 2002. Comparison of polybrominated diphenyl ethers in fish, vegetables, and meats and levels in human milk of nursing women in Japan. Chemosphere 46, 689–696. Rahman, F., Langford, K.H., Scrimshaw, M.D., Lester, J.N., 2001. Polybrominated diphenyl ether (PBDE) flame retardants. Sci. Total Environ. 275, 1–17. Rice, C.P., Chernyak, S.M., Begnoche, L., Quintal, R., Hickey, J., 2002. Comparisons of PBDE composition and concentration in fish collected from the Detroit river, MI and Des Plaines River, IL. Chemosphere 49, 731–737. Richardson, D.B., Ciampi, A., 2003. Effects of exposure measurement error when an exposure variable is constrained by a lower limit. Am. J. Epidemiol. 157 (4), 355–363. Sandanger, T.M., Sinotte, M., Dumas, P., Marchand, M., Sandau, C.D., Pereg, D., Be´rube´, S., Brisson, J., Ayotte, P., 2007. Plasma concentrations of selected organobromine compounds and polychlorinated biphenyls in postmenopausal women of Quebec, Canada. Environ. Health Perspect. 115 (10), 1429–1434. Schecter, A., Papke, O., Tung, K.C., Harris, T.R., Dahlgren, J., 2005. Polybrominated diphenyl ether flame retardants in the US Population: current levels, temporal trends, and comparison with dioxins, dibenzofurans, and polychlorinated biphenyls. J. Occup. Environ. Med. 47, 199–211. Schecter, A., Papke, O., Harris, T.R., Tung, K.C., Musumba, A., Olson, J., et al., 2006a. Polybrominated diphenyl ether (PBDE) levels in an expanded market basket survey of US food and estimated PBDE dietary intake by age and sex. Environ. Health Perspect. 114 (10), 1515–1520. Schecter, A., Harris, T.R., Papke, O., Tung, K.C., Musumba, A., 2006b. Polybrominated diphenyl ether (PBDE) levels in the blood of pure vegetarians (vegans). Toxicol. Environ. Chem. 88, 17–112. Schisterman, E., Whitcomb, B., Louis, G., Louis, T., 2005. Lipid adjustment in the analysis of environmental contaminants and human health risks. Environ. Health Perspect. 113 (7), 853–857. Schisterman, E., Vexler, A., Whitcomb, B., Liu, A., 2006. The limitations due to exposure detection limits for regression models. Am. J. Epidemiol. 163 (4), 374–383. Sjo¨din, A., Hagmar, L., Klasson-Wehler, E., Bjo¨rk, J., Bergman, A., 2000. Influence of consumption of fatty Baltic sea fish on plasma levels of halogenated environmental contaminants in Latvian and Swedish men. Environ. Health Perspect. 108 (11), 1035–1041. Sjo¨din, A., Jones, R.S., Lapeza, C.R., Focant, J.F., McGahee III, E.E., Patterson Jr., D.G., 2004. Semiautomated high-throughput extraction and clean up method for the measurement of polybrominated diphenyl ethers, polybrominated biphenyls, and polychlorinated biphenyls in human serum. Anal. Chem. 76, 1921–1927.

ARTICLE IN PRESS H.M. Spliethoff et al. / Environmental Research 108 (2008) 340–347

Sjo¨din, A., Wong, L.Y., Jones, R.S., Park, A., Zhang, Y., Hodge, C., Dipietro, E., McClure, C., Turner, W., Needham, L.L., Patterson, D.G., 2008. Serum concentrations of polybrominated diphenyl ethers (PBDEs) and polybrominated biphenyl (PBB) in the United States population: 2003–2004. Environ. Sci. Technol. 42, 1377–1384. Thomsen, C., Knutsen, H.K., Liane, V.H., Frøshaug, M., Kvalem, H.E., Haugen, M., Meltzer, H.M., Alexander, J., Becher, G., 2008. Consumption of fish from a contaminated lake strongly affects the concentrations of polybrominated diphenyl ethers and hexabromocyclododecane in serum. Mol. Nutr. Food Res. 52 (2), 228–237. US EPA (Environmental Protection Agency), 1995. Guidance for Assessing Chemical Contaminant Data for Use in Fish Advisories. Vol. 1: Fish Sampling and Analysis, second ed. Office of Science and Technology, Office of Water, Washington, DC.

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Vena, J., Buck, G., Kostyniak, P., Mendola, P., Fitzgerald, E., Sever, L., et al., 1996. The New York Angler Cohort Study: exposure characterization and reproductive and developmental health. Toxicol. Ind. Health. 12 (3–4), 327–334. Weiss, J., Wallin, E., Axmon, A., Jonsson, B.A.G., Akesson, H., Janak, K., et al., 2006. Hydroxy-PCBs, PBDEs, and HBCDDs in serum from an elderly population of Swedish fisherman’s wives and associations with bone density. Environ. Sci. Technol. 40, 6282–6289. Wu, N., Herrmann, T., Paepke, O., Tickner, J., Hale, R., Harvey, E., La Guardia, M., McClean, M.D., Webster, T.F., 2007. Human exposure to PBDEs: associations of PBDE body burdens with food consumption and house dust concentrations. Environ. Sci. Technol. 41 (5), 1584–1589. Zhu, L.Y., Hites, R.A., 2004. Temporal trends and spatial distributions of brominated flame retardants in archived fishes from the Great Lakes. Environ. Sci. Technol. 38 (10), 2779–2784.