Stapleton et al. (21.5 ng/g lipid), and non-Hispanic blacks (24.3 ng/g lipid) were higher than the geo- metric mean level measured for our cohort (16.5 ng/g lipid). Thus, our results are con- sistent with decreasing PBDE levels over time within race and age groups. S e v e r a l p h e n o l i c m e t a b o l i t e s w e r e also identified in a subset of these samples (Table 2). Previous studies have observed these phenolic compounds in both human tissues and in laboratory animals exposed to PBDE mixtures, suggesting they are in fact metabolites of PBDEs (Athanasiadou et al. 2008; Hakk and Letcher 2003; Qiu et al. 2007, 2009; Wan et al. 2010). Qiu et al. (2009) reported mean levels of 0.8, 0.3, and 0.3 ng/g lipid for 246 TBP, 4´-OH-BDE-49, and 6-OH-BDE-47, respectively, in mater- nal blood samples (n = 4) collected in 2003– 2004. These levels are similar to the median levels we report here of < 2.5, 0.11, and 0.17 ng/g lipid for 246 TBP, 4´-OH-BDE-49, and 6-OH-BDE-47, respectively. Although 246 TBP is a known metabolite of PBDEs, it is also produced and used as a flame retardant in phenolic and epoxy resin itself (Andersson et al. 2006). Thus, there can be more than one source of exposure to this compound in human tissues. All PBDEs and OH-BDEs were highly c o r r e l a t e d , e x c e p t 6 - O H - B D E - 4 7 a n d BDE-99. This weaker relationship may be attributable to the fact that 6-OH-BDE-47 is not a likely metabolite of BDE-99 and/or may have marine natural sources (Malmvarn et al. 2005). In addition, BDE-99 may be metabolized much more rapidly than BDE-47 and BDE-153 (Chen et al. 2006; Hakk et al. 2002; Lupton et al. 2009; Stapleton et al. 2004b, 2009). Positive associations between PBDEs and FT4 and TT4 levels have been observed in nonpregnancy cohorts in previous stud- ies (Meeker et al. 2009; Turyk et al. 2008). Unfortunately, we were not able to control for potential confounders [e.g., polychlorinated biphenyls (PCBs), organochlorine pesticides] in this study. However, both Turyk et al. (2008) and Chevrier et al. (2010) were able to control for these confounders, and they dem- onstrated that associations between PBDEs and thyroid hormones remained. In addition, PCBs have been associated with decreases in T4, not increases as observed in our study (Brouwer et al. 1998; Wang et al. 2005). Positive associations between PBDEs and T4 observed are opposite to those typically observed in laboratory animal exposure studies, where exposed animals experience decreases in circulating T4 and T3 levels (Fernie et al. 2005; Tomy et al. 2004; Zhou et al. 2001). A previous study examining PBDEs in preg- nant women found no significant associations between PBDEs and T4; however, the authors 1458 did observe an inverse association between PBDEs and TSH (Chevrier et al. 2010). Low TSH levels are generally associated with higher T4 levels via feedback mechanisms regulated by the pituitary (Hulbert 2000); thus, the results observed here and by Chevrier et al. (2010) suggest a similar trend with respect to the influ- ence of PBDEs on thyroid hormone regulation during pregnancy. Differences in TSH asso- ciations between the two studies may be related to the fact that African-American women in general have lower TSH levels (Walker et al. 2005). Both cohorts displayed very similar PBDE levels, with geometric mean BDE-47 levels of 15.3 and 16.9 for their cohort and our cohort, respectively. However, as noted by Chevrier et al. (2010), different methods for measuring thyroid hormones may also be a fac- tor in different trends observed. Here we used a chemiluminescent immunoassay for measuring unbound (i.e., free) thyroid hormones (which may be a limitation), whereas Chevrier et al. (2010) used an equilibrium dialysis method. Diurnal variations in thyroid hormone secre- tions may also be a factor. Our findings suggest an inverse association between 4´OH-BDE-49 and TT3, but our analysis was limited by the small sample num- ber (n = 56). Most animal exposure studies have observed significant decreases in T4 with exposure to PBDEs, but not T3. However, a large proportion of this cohort (63%) did have TT3 levels that were above normal levels for an average adult, which is likely linked to their late stage of pregnancy. Previous in vitro studies have demonstrated that hydroxylation of the PBDE congener results in significantly increased binding affinities for thyroid hor- mone transporters found in serum (Marchesini et al. 2008; Meerts et al. 2000). Specifically, 4´-OH-BDE-49 has been shown to have a high binding affinity to the thyroid hormone serum transporter transthyretin (Ucan-Marin et al. 2009). Thus, it may be possible that PBDE metabolites are competing for space on the thyroid hormone transporters, resulting in more free (i.e., unbound) T4. However, TT4 was also positively associated with PBDEs. An alternate explanation may be related to the effect of PBDEs and/or their metabolites on deiodinase activity. Previous in vivo and in vitro studies in fish models have suggested that deio- dinase enzymes (DIs) may be metabolizing PBDEs via a dehalogenation pathway (Noyes et al. 2010; Stapleton et al. 2004a). DIs serve to metabolize T4 to T3 in peripheral tissues, supplying the main source of T3 that binds to thyroid nuclear receptors to activate transcrip- tional processes (Kohrle 1999). If DI activity was inhibited by PBDEs and/or their metabo- lites, it might result in an increase in circulating T4 levels and a decrease in T3 levels. Here we have reported on the levels of PBDEs in a pregnancy cohort consisting primarily of African-American women and report on the associations between OH-BDEs and thyroid hormones. Although there are some limitations to this study (e.g., analy- sis during the third trimester of pregnancy, method used for measuring thyroid hor- mones), the results do support previous stud- ies that have observed associations between PBDEs and thyroid hormones. PBDEs have a very long half-life in the human body, and PBDE levels measured in the third trimester are likely to reflect levels throughout preg- nancy. Consequently, it is possible that PBDEs may be affecting thyroid hormone regulation throughout pregnancy, including the first tri- mester when the fetus relies solely on mater- nal thyroid hormone supply. More research is needed to determine the mechanisms by which PBDEs affect thyroid hormone regula- tion during pregnancy and how this, in turn, affects the development of the fetus. Ref Rencs Alaee M, Arias P, Sjodin A, Bergman A. 2003. An overview of commercially used brominated flame retardants, their applications, their use patterns in different coun‑ tries/regions and possible modes of release. Environ Int 29(6):683–689. Andersson PL, Oberg K, Orn U. 2006. Chemical characteriza‑ tion of brominated flame retardants and identification of structurally representative compounds. Environ Toxicol Chem 25(5):1275–1282. Athanasiadou M, Cuadra SN, Marsh G, Bergman A, Jakobsson K. 2008. Polybrominated diphenyl ethers (PBDEs) and bioaccumulative hydroxylated PBDE metabo‑ lites in young humans from Managua, Nicaragua. Environ Health Perspect 116:400–408. Brouwer A, Morse DC, Lans MC, Schuur AG, Murk AJ, Klasson‑ Wehler E, et al. 1998. Interactions of persistent environ‑ mental organohalogens with the thyroid hormone system: mechanisms and possible consequences for animal and human health. Toxicol Ind Health 14(1–2):59–84. Burgette LF, Reiter JP. 2010. Multiple imputation for miss‑ ing data via sequential regression trees. Am J Epidemiol 172(9):1070–1076. Chao HR, Wang SL, Lee WJ, Wang YF, Papke O. 2007. Levels of polybrominated diphenyl ethers (PBDEs) in breast milk from central Taiwan and their relation to infant birth outcome and maternal menstruation effects. Environ Int 33(2):239–245. Chen LJ, Lebetkin EH, Sanders JM, Burka LT. 2006. Metabolism and disposition of 2,2’,4,4’,5‑pentabromodiphenyl ether (BDE99) following a single or repeated administration to rats or mice. Xenobiotica 36(6):515–534. Chevrier J, Harley KG, Bradman A, Gharbi M, Sjodin A, Eskenazi B. 2010. Polybrominated diphenylether (PBDE) flame retardants and thyroid hormone during pregnancy. Environ Health Perspect 118:1444–1449. Covaci A, Voorspoels S, Thomsen C, van Bavel B, Neels H. 2006. Evaluation of total lipids using enzymatic methods for the normalization of persistent organic pollutant levels in serum. Sci Tot Environ 366(1):361–366. de Wit CA. 2002. An overview of brominated flame retardants in the environment. Chemosphere 46(5):583–624. Fernie KJ, Shutt JL, Mayne G, Hoffman D, Letcher RJ, Drouillard KG, et al. 2005. Exposure to polybrominated diphenyl ethers (PBDEs): changes in thyroid, vitamin A, glutathione homeostasis, and oxidative stress in American kestrels (Falco sparverius). Toxicol Sci 88(2):375–383. Hakk H, Larsen G, Klasson‑ Wehler E. 2002. Tissue disposition, excretion and metabolism of 2,2’,4,4’,5‑pentabromodi ‑ phenyl ether (BDE‑ 99) in the male Sprague‑ Dawley rat. Xenobiotica 32(5):369–382. Hakk H, Letcher RJ. 2003. Metabolism in the toxicokinetics and fate of brominated flame retardants—a review. Environ Int 29(6):801–828. ovlume 119 | number 10 | October 2011 • Environmental Health Perspectives