Recent experimental data indicates PBDEs are developmental neurotoxicants and exposure during gestation is associated with adverse neurologic outcomes (2, 21). Understanding the predictors of exposure to PBDEs among pregnant women and women of childbearing age may help identify highly exposed subpopulations and help inform interventions to reduce exposure. This study examined body burdens of PBDEs and identified determinants of exposure among a sample of healthy pregnant women enrolled in a predominantly low-income, Hispanic cohort who provided maternal blood samples between years 2009 and 2010.
PBDE exposure is widespread in this cohort. At least one PBDE congener was detected in serum collected from each subject and 4 congeners (PBDE-47, -99, -100 and -153) were detected in nearly 100% of serum samples. Our analyses demonstrate unique demographic and dietary predictor patterns for serum PBDE-47, -99, -100 and -153. Race/ethnicity appeared predictive of PBDE body burden; levels of PBDE-153 were 20% higher among Non Hispanic Whites compared to Hispanics and 13% higher than African Americans. Though not significant, levels of PBDE-47 were 20% higher among African Americans compared to Non Hispanic Whites and Hispanics. Maternal education level is also associated with maternal PBDE body burden; levels of PBDE-47 were highest among subjects with a high school education compared to all other categories and levels of PBDE-153 were highest among college graduates compared to other categories. Consumption of solid dairy and consumption of processed meat were weakly associated with higher levels of all 4 PBDE congeners. Reporting of a greater number (≥ 10) of household electronics was significantly associated with higher levels of all 4 commonly detected PBDE congeners. More characteristics were associated with serum PBDE-153 than for any other congener.
Levels and patterns of PBDE exposure
While PBDE exposure is widespread among this cohort, levels were lower than those reported in any other U.S. based cohort of pregnant women. Within the last decade, several studies have reported levels of PBDEs in blood samples collected during pregnancy or immediately following pregnancy [26–30], for review see . Woodruff et al. (2012) reviewed Nurses Health And Nutrition Examination Survey (NHANES) data for chemicals present in samples collected from U.S. pregnant women between 2003/2004 . Levels of PBDEs measured in NHANES women were 3–4 times higher than those measured in our cohort. Similarly, a recent study of predominantly Non-Hispanic Black pregnant women living in North Carolina between years 2008–2010 reported levels of PBDEs similar to those reported in NHANES and 3–4 times higher than our levels . Levels of PBDEs measured in an ethnically diverse cohort of pregnant women living in California and sampled between 2008/2009 were among the highest reported in the country . These levels were double the national values  and up to 6 times higher than those measured in our cohort. Herbstman et al. (2010), presented levels of PBDEs measured in cord blood samples collected from infants born in New York City in 2001 . Maternal and fetal blood levels of PBDEs are highly correlated and a strong indication of PBDE exposure . Cord blood PBDE levels in the Herbstman study were nearly two times higher than our reported maternal blood levels with the exception of PBDE-153, which was only 12% higher in the 2001 cohort than our cohort.
The generally lower levels of PBDEs detected in our cohort may be due to statewide legislation and abatement policies regarding the manufacture and use of PBDEs. Due to public concerns and economic reasons, Great Lakes Chemical Corporation, the only North American producer of PBDEs, voluntarily stopped producing octa-BDE and penta-BDE mixtures in December 2004 [35, 36]. By 2006, New York State banned the manufacturing, processing, or distributing products penta- and octa-PBDEs and initiated a phase out of deca-PBDE by 2008 [37, 38]. While California also banned penta- and octa-PBDEs by 2006 , the state California maintains the nations highest flammability standards which were enacted in the 1970’s . These standards may explain the high levels of PBDEs measured in California residents.
While lower than most other U.S. cohorts, PBDE levels in this cohort were up to 8 times higher than those measured in the Netherlands and Sweden . This is consistent with estimates that North Americans have the highest global body burden of PBDEs worldwide, averaging 20 times that of Europeans . Lower levels in European countries may be due to earlier and stricter enactment of regulations in 2004 restricting the use of octa-and penta-congeners .
Though levels measured in our cohort were lower than other US cohorts and higher than international cohorts, the pattern of distribution of PBDE congeners is similar to other patterns reported for PBDEs measured in maternal and/or umbilical cord blood samples (reviewed in ). Similar to other studies, the distribution of PBDEs in our cohort is not normal, and approximately 5% of subjects have PBDE concentrations 50x times higher than the median [9, 10, 12, 42]. Also consistent with the literature, PBDE-47, -99 and -100 are more highly correlated with each other than with PBDE-153 [43, 44]. PBDE-47, -99 and -100 are all components of the penta-PBDE technical mixture.
Commercial penta-PBDE is most commonly used as a flame retardant in polyurethane foam . The origin of these congeners in the environment is likely due to off gassing from products containing the technical penta-PBDE mixture such as upholstered couches, mattresses, mattress pads, and other foam items and from degradation pathways of deca-PBDE . PBDE-47 may be the predominant congener due to its higher potential for bioaccumulation  or due to other unidentified sources of exposure . PBDE-153 is a component of both penta-PBDE and octa-PBDE technical mixtures. Octa-PBDE mixtures are used most widely in plastics and textiles . Notably, though PBDE-153 was widely detected in biological samples (reviewed in ), it contributes very little to the overall content of PBDEs in technical products; only 0.15% to 8.7% of the composition of octa-PBDE and 5.3- to 5.4% of penta-PBDE. In rodents, PBDE-153 is resistant to metabolism, which, if also applicable to humans may explain the proportionately higher PBDE-153 concentrations in human matrices compared to commercial mixtures [43, 49].
Predictors/sources of exposure
While PBDE exposure in the general population has been associated with the indoor environment and diet [10, 12, 16, 50–53], few studies consider individual sources of exposure to PBDEs during pregnancy. In one recent study of a largely Mexican immigrant population of pregnant women living in California, PBDEs concentrations increased with increasing years residing in the U.S. and with the number of pieces of stuffed furniture in the home . Herbstman (2007)  examined associations between PBDE levels in cord blood and potential predictors in 94 fetal cord serum samples. Predictors of cord blood PBDE-47 and -153 included younger maternal age and less weight gaining during pregnancy. In our study, maternal age was not associated with any PBDE congener and, while we do have information on weight gain during pregnancy, prepregnancy BMI was only weakly associated with higher levels of PBDE-153.
Two recent studies examined predictors of PBDEs among U.S. born girls [43, 54]. Both studies found that social factors influenced children’s body burden levels. Lower socioeconomic status, assessed using educational level of the mother, was associated with higher PBDE levels in children. Rose et al. (2010) found that among their largely Hispanic and White population, Hispanic children had lower levels of higher brominated congeners (sum PBDE-197 thru -209) but no significant differences in race/ethnicity were observed among the lower brominated congeners (sum PBDE-28 thru -153). In the Windham study, Black girls had significantly higher levels of PBDEs compared to whites, Hispanics had intermediate values. Authors report that the reasons for racial and socioeconomic disparity in PBDE exposure are unknown but may reflect differences in exposure pathways, specifically housing factors or household furnishings.
Among our predominantly Hispanic cohort, we observed higher levels of PBDE-47 in African Americans compared to Hispanic and non-Hispanic white mothers, though differences were not significant. PBDE-153 levels were highest among non-Hispanic whites and lowest among Hispanics. We did not observe associations between income and PBDE levels in our cohort, though education level was predictive of PBDE congeners. PBDE-47 levels were highest among those mothers reporting a high school education compared to those with a less than high school education and those with higher levels of education (college or graduate school). PBDE-153 was highest among college graduates compared with those subject with lower education and those with higher education. The different trends observed between our cohort and the previously discussed cohorts may again be associated with the time of sample collection, the catchment areas of the enrollment and the different demographic distribution of the cohorts.
Food intake is an important source of exposure to PBDEs. PBDEs accumulate in lipid rich tissues. Food items like fish from high trophic levels and lipid-rich oils have been found to contain relatively high concentrations of PBDEs [55, 56]. In addition, bioaccessibility of PBDEs is higher among fatty foods . Among food items, fish, meats, and dairy products have the highest PBDE concentrations. PBDEs are less commonly detected in fruits and vegetables , however contamination of produce is known to occur through the process of ‘hydro-cooling’ using plastic pallets coated with deca-PBDE . A recent study quantified levels of PBDEs in commonly consumed foods in the United States and showed the highest levels of PBDEs in butter and fish, and the lowest detectable levels in other liquid dairy products (milk and yogurt) and vegetables . Further, while fish tend to be more highly contaminated with PBDEs than meat and dairy, meat and dairy are more important sources of PBDE exposure in the United States [60, 61]. Previous studies have attempted to quantify PBDE concentrations and dietary intake [50, 62]. Modest associations have been reported between serum PBDEs and fish consumption among consumers of sport caught fish [63–65]. A recent study using NHANES data examined the association between food items from a 24-hour recall and a 1-year food frequency questionnaire and serum PBDE levels. Intake of poultry and red meat contributed significantly to PBDE body burden in the U.S. . Recent market basket survey show contamination highest in fish, then meat and least in dairy products , but the average American has most dietary PBDE intake from meat, then dairy and fish .
In the current study, we examined associations between PBDE congeners and dietary consumption focusing on meat, poultry, dairy products, fish and shellfish. Women with higher levels of PBDE-153 and -100 were more likely to report consumption of processed meat. Women with higher levels of PBDE-153 were also more likely to report consumption of solid (high fat) dairy products such as cheese, cottage cheese and butter. We did not observe an association between consumption of fish, including fatty fish, and any PBDE congener.
A recent review investigated human internal and external exposure to PBDEs and suggested that while diet was long considered the most important source of human exposure to PBDEs, the unintentional ingestion of dust is at least of similar importance . It is likely that PBDEs migrate out of household products including furniture and electronics and are sequestered in house dust. House dust then contributes to PBDE body burden either through oral intake of the dust or via dermal absorption . In the current study, we queried women about the number of household electric appliances in the home (i.e., televisions, computers, microwave ovens). We observed higher levels of PBDE-47, -99, -100 and -153 among women who reported at least 10 household electronic and electrical appliances (including stereo equipment, televisions, toasters, microwaves, computers and/or printers). Based on the literature, we anticipated that household electronics would predict deca-PBDEs rather than penta-PBDEs. Deca-PBDE (PBDE-209) was measured in serum samples in our cohort and detected in less than 20% of subjects. We did not observe associations between PBDE-209 and household electronics (data not shown). Interestingly, the association between PBDE-153 and household electronics was stronger among Hispanics than whites or African Americans and strongest among those subjects in the lowest income category. The association between electronics and PBDE-47 was strongest among Whites and Hispanics and consistently strong across all income categories. It is possible that the types of electronics in the home may differ by race/ethnicity and income, thus influencing the exposure profile.
Our study has several limitations. With respect to questionnaire data, several recent studies suggest PBDE body burden differs by race/ethnicity, country of origin and years living in the U.S. Rose et al. reported that children of foreign-born mothers, 52% of whom were Hispanic, had significantly lower levels of higher brominated PBDE congeners than those whose mothers were born in the U.S. . While we have a predominately Hispanic population, we do not have information on the country of birth (U.S. or foreign born) or the number of years living in the U.S.
In addition to dietary and consumer products, housing characteristics reflecting lifestyle are associated with PBDE exposure. One study found that children living in larger homes had significantly lower levels of PBDE-209; higher maternal education was correlated with larger homes . Dust concentrations of PBDEs within larger homes may be lower since increasing square footage may dilute the PBDE contamination from a household source . In the current study, we do not have information on housing quality or size.
With respect to environmental and biological monitoring data, we do not have environmental data on PBDEs concentrations in house dust or indoor air, nor do we have PBDE levels on the foods consumed by women. Further, we have no assessment of dermal absorption. Biological exposure data may not be directly correlated to source exposure data due to individual metabolism. In addition, we do not have measures of PBDE metabolites. In mice, PBDEs are metabolized in vivo to para- and ortho-hydroxylated metabolites (HO-PBDEs) . It has been suggested that the para-HO-PBDE metabolites are the most abundant and potentially the most toxic metabolites due to hormonal activity . While it is possible to measure para-hydroxylated metabolites in blood collected from pregnant women [29, 70], metabolite biomarkers were not available for this study. Finally, with respect to questionnaire items querying participants about household items likely to contain PBDEs, we did not ask the age or condition of household items. It is possible that newly purchased items are a greater source of exposure to PBDEs due to increased off gassing. Alternatively, it is possible that older items may be a greater source of exposure due to breakdown of the construction materials. Our questionnaire items querying housing characteristics did not include information on number of bedrooms, square footage, or proxies for air exchange and ventilation. As air exchange rates and size of home may affect PBDE concentrations in the home , we are unable to assess this.