We found evidence suggesting that children who had higher cord blood concentrations of BDEs 47, 99, and 100 scored lower on tests of mental and physical development at ages 12–48 and 72 months. These associations were significant for 12-month PDI (BDE-47); 24-month MDI (BDEs 47, 99, and 100); 48-month Full IQ (BDEs 47, 100, and 153); Verbal IQ (BDEs 47 and 100) and Performance IQ (BDEs 100 and 153); and 72-month full and Performance IQ (BDEs 100 and 153). Children who were in the highest 20% of cord blood concentrations of BDEs 47, 99, or 100 had significantly lower developmental scores compared with children who were in the lower 80% of the exposure distributions for these chemicals. These differences were particularly evident at 48 months of age.
Adverse neurodevelopmental effects associated with prenatal PBDE exposure can be detected both at early ages (12–36 months) and as the children age (48 and 72 months). Neurodevelopmental deficits documented by the WPPSI during the preschool period are an important predictor of subsequent academic performance (Kaplan 1993
). Documenting the first appearance of potentially longer-term adverse effects at early ages is also important, because these indicators may identify children who could benefit from early intervention programs. The identification of later deficits may indicate the persistence of early effects and/or an increase in the magnitude of effect with age, as has been shown in some animal studies (e.g., Viberg et al. 2003a
Our results are consistent across congeners and over time. This may be predictable because the PBDE congeners are highly correlated, and for individuals, repeated developmental scores are also correlated. Although the number of participants lost to follow-up between 12 and 72 months was relatively low (81% of subjects available for analyses at 12 months were also assessed at 72 months) and losses are independent of exposure, our overall sample is relatively small. Therefore, even small losses to follow-up may limit our power to detect significant differences in multivariate models. The resulting small sample size precluded the analysis of exposure effects on developmental delay, and we were unable to look at interactions. However, the developmental deficits of the magnitude we observed in this study are likely to have the largest functional impact on those who score at the lower end of the population distribution.
The only other epidemiologic study reporting the neurodevelopmental effects of prenatal exposure to PBDEs was published recently (Roze et al. 2009
). In this study of 62 Dutch children, the authors present correlations between exposure to PBDEs (measured during the 35th week of pregnancy) and > 20 indices of child development and behavior at age 60–72 months. The authors report that prenatal PBDE exposure was associated with some adverse effects on development (reduced fine manipulative abilities and increased attentional deficits) as well as some beneficial effects (better coordination, better visual perception, and better behavior). The authors evaluated, but did not find statistically significant, correlations between prenatal PBDE exposures and any of the WPPSI-R domains. Our results are not consistent with these findings. However, important differences in exposure (median exposure in our population was 4 times higher for BDEs 47 and 99; 2.3 times higher for BDE-100; and one-fifth their concentration for BDE-153), sample size, and statistical analyses performed may account for some of the observed inconsistencies.
Our results are consistent with published toxicologic experiments [reviewed by Costa and Giordano (2007)
]. For example, studies evaluating the neurodevelopmental effects of neonatal exposure to PBDEs in mice during critical developmental periods have reported altered habituation patterns (Viberg et al. 2003a
), hyperactivity (Gee and Moser 2008
), and learning and memory deficits (Dufault et al. 2005
; Viberg et al. 2003a
). There is some evidence suggesting that BDE-99 is more potent than BDE-47 (Viberg et al. 2003a
) and also that effects worsen (or are more apparent) with age (Viberg et al. 2003a
). In general, we observed the largest associations with prenatal exposure to BDE-100, and the associations with prenatal exposures were still apparent, albeit not consistently significant, in our smaller sample examined at 72 months of age.
A number of potential mechanisms have been proposed to explain the cognitive and locomotive deficits observed in animals after PBDE exposure during critical developmental periods, including direct neurotoxic effects on neuronal and glial cells (Costa et al. 2008
) resulting from changes in the quantity of cholinergic nicotinic receptors in the hippocampus (Viberg et al. 2003a
) and induction of apoptotic cerebellar granule cell death (Reistad et al. 2006
). In addition, there is compelling experimental and epidemiologic evidence suggesting that PBDEs can interfere with thyroid hormone pathways (Legler 2008
). Because thyroid hormones are critical for normal brain development, this provides an attractive explanation for observed neurodevelopmental effects after neonatal PBDE exposure (Bigsby et al. 1999
; Porterfield 2000
). Toxicologic evidence corroborating this theory includes potentially causal associations between neonatal exposure to BDE-47, BDE-99, or commercial PBDE mixtures (DE-71 and Bromokal 70-5 DE) and reduced thyroxine (T4
) concentrations in experimental murine models (Fowles et al. 1994
; Hallgren et al. 2001
; Kuriyama et al. 2007
Although only limited human epidemiologic data are available, increased levels of BDEs 47, 99, and 100 in dust in the homes of adult human males recruited through a U.S. infertility clinic were associated with altered hormone levels. PBDEs were inversely associated with free androgen index and with luteinizing and follicle-stimulating hormones and were positively associated with inhibin B, sex hormone–binding globulin, and free T4
(Meeker et al. 2009
). In another study of adult males, increased serum concentrations of PBDEs were positively related to T4
and inversely related to total triiodothyronine (T3
) and thyroid-stimulating hormone (TSH) (Turyk et al. 2008
). The positive associations between PBDEs and T4
levels demonstrated in these human studies are not consistent with the results from experimental animal models, raising the possibility that the underlying mechanism of the effect of PBDEs on thyroid disruption may differ among species. However, it is difficult to extrapolate findings from studies evaluating exposure effects in adults to prenatal exposures, because PBDEs may exhibit differential effects on thyroid hormone levels at different stages of the life span. A recent study of PBDEs measured in human cord blood of infants born to a cross-section of women delivering in Baltimore, Maryland, showed a consistent nonsignificant negative association with both total and free T4
in infants (Herbstman et al. 2008
). More research is necessary to fully characterize the association of human prenatal exposure to PBDEs with thyroid hormone levels.
The exact mechanism of thyroid disruption by PBDEs in humans has not yet been elucidated, but two potential pathways through which PBDE exposure could lead to thyroid disruption have been proposed [reviewed by Zhang et al. (2008)
]. The structural similarities of T4
to polyhalogenated aromatic hydrocarbons suggest that hydroxylated PBDE metabolites could displace thyroid hormones from thyroid transport proteins (i.e., transthyretin), altering free thyroid hormone levels (Turyk et al. 2008
). Alternatively (or in addition), PBDEs might affect hormone levels by influencing thyroid hormone synthesis and/or stimulating thyroid hormone metabolism (Szabo et al. 2009
; Turyk et al. 2008
). Brain development in the fetus is contingent on the precise timing of thyroid hormone levels, particularly for T4
, and deviations above or below the normal levels can lead to developmental deficits (Williams 2008
). The fetus originally derives all thyroid hormone from the mother, but over the course of the pregnancy, its thyroid gland develops, and hormones produced within the fetus gradually replace the maternal source. The surge in maternal T4
in the first trimester, coupled with TSH inhibition, is thought to provide a supply of hormone during this critical developmental period, and alteration of T4
levels by PBDEs at this time could alter neurodevelopment (Williams 2008
). Although low serum T4
from maternal hypothyroidism during gestation (e.g., iodine deficiency) is known to cause mental retardation in children, elevated levels of T4
have been associated with increased rates of miscarriage (Anselmo et al. 2004
) and could potentially be linked to neurodevelopmental problems.
Because of their similar chemical structures, PBDEs and polychlorinated biphenyls (PCBs) have been compared in terms of their potential health effects. Although PCBs were banned in most industrialized countries > 25 years ago, they are still measurable in human and environmental samples because of their long half-lives in the environment and in humans (Talsness 2008
). Prenatal exposure to PCBs has been shown in several cohort studies to significantly reduce cognitive function during childhood [reviewed by Schantz et al. (2003)
] and has also been associated in some studies with altered thyroid hormone levels (Chevrier et al. 2007
; Herbstman et al. 2008
). Because of the structural similarity of PCBs and PBDEs, it has been postulated that they exert biological effects through similar processes.
This study population is unique in that participants were initially recruited to measure the extent and the effects of prenatal exposure to contaminants (including PBDEs) that were potentially released by the destruction of the WTC towers. Studies examining environmental samples collected pre- and post-9/11 near the WTC site found indications of higher concentrations of PBDEs after the attacks (Litten et al. 2003
) and nearer to the WTC disaster site (Butt et al. 2004
). These trends may be attributable to debris containing office equipment known to be treated with PBDEs (de Wit 2002
). In our study population, cord plasma levels of PBDEs were not significantly related to residential distance from the WTC site. There is some evidence suggesting that PBDE exposure may be related to the WTC attack based on the gestational age on 9/11, such that women who were in the second half of their pregnancy on 9/11 had children with higher cord concentrations of PBDEs (unpublished data). It is also possible that just after 9/11, some women had elevated levels of PBDEs but that these levels declined with the passage of time between the peak exposure and delivery, resulting in lower observed levels. In either scenario, it is not clear how much this apparent association between gestational age on 9/11 and exposure concentration contributes to the body burden, and it is certain that sources other than the WTC are also accountable. In this report, our interest is in the association between the integrated prenatal PBDE exposure from multiple sources and neurodevelopment. It is also possible that there are other unknown factors associated with PBDEs that may confound the observed relationships between prenatal PBDE exposure and adverse neurodevelopment.
Levels of cord blood PBDEs in our population are consistent with those reported in other U.S. populations (Herbstman et al. 2007
; Mazdai et al. 2003
; Wu et al. 2007
). Compared with cord blood measurements in an inner-city population in Baltimore, Maryland, our study population had slightly lower median concentrations (i.e., 11.2 ng/g lipid vs. 13.6 ng/g lipid for BDE-47) (Herbstman et al. 2007
). In the Baltimore cohort as well as in this New York City cohort, higher cord PBDE concentrations were associated with mothers’ African American or non-Asian race/ethnicity, although a higher proportion of the Baltimore population was African American (70% vs. 15%) and a lower proportion was Asian (8% vs. 30%). Increasing maternal age was associated with lower PBDE concentrations in the Baltimore cohort but not in New York City; however the median maternal age was also lower in Baltimore (25 years vs. 30 years) (Herbstman et al. 2007
). The demographic differences between these two populations may explain the small differences in blood levels observed at the population level.
Although dietary ingestion was once thought to be the largest route of PBDE exposure in humans, the similarity of PBDE levels in foods in Europe, Asia, and North America fails to adequately explain the high blood levels in the U.S. population (Frederiksen et al. 2009
). Dust inhalation may be a more important exposure route to PBDEs, particularly BDEs 47, 99, and 100. In a review of median PBDE levels in dust and air samples, measured BDE-47 dust levels in Europe and North America were 32 and 429 ng/g of dust, respectively. Similar disparities were observed for BDE-99 and BDE-100 levels (Frederiksen et al. 2009
; Sjodin et al. 2008a
). Particular attention should be given to this exposure route in young children, who are more likely to encounter dust because of their proximity to the floor. Dust is estimated to contribute from 80 to 93% of PBDE exposure in toddlers, and their small body size compounds the effect of their exposures (Costa and Giordano 2007
). In this study, we were not able to control for postnatal dust exposure.
In the general population, infants and toddlers have the highest body burden of PBDEs, and along with dust exposure, exposure via breast milk is thought to be a major contributor to this burden (Costa et al. 2008
; Toms et al. 2009
). Breast-fed infants are estimated to be exposed to 306 ng/kg body weight/day PDBE compared with 1 ng/kg body weight/day in adults, with the most prominent congeners being BDEs 47, 99, and 153 (Costa et al. 2008
). In our study, breast-feeding rates were higher in children with higher cord PBDE levels, indicating that PBDEs measured in cord blood may underestimate the exposure of breast-fed children. Breast-feeding was, as expected, associated with higher scores on neurodevelopmental indices, making it an important potential confounder to include in multivariate statistical models.