Association between prenatal exposure to polybrominated diphenyl ethers and young children's neurodevelopment in China

Environmental Research 142 (2015) 104–111

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Association between prenatal exposure to polybrominated diphenyl ethers and young children's neurodevelopment in China Guodong Ding a,b,1, Jing Yu c,1, Chang Cui d,1, Limei Chen e, Yu Gao e, Caifeng Wang e, Yijun Zhou e, Ying Tian a,e,n a

MOE and Shanghai Key Laboratory of Children's Environmental Health, Xinhua Hospital, Shanghai Jiao Tong University School of Medicine, Shanghai, China Department of Pediatrics, Shanghai East Hospital, Tongji University School of Medicine, Shanghai, China c Department of Endocrinology and Metabolism, Shanghai Jiao Tong University Affiliated Sixth People's Hospital, Shanghai, China d Research Base of Key Laboratory of Surveillance and Early-warning on Infectious Disease in China CDC, Shanghai Pudong New Area Center for Disease Control and Prevention, Shanghai, China e Department of Environmental Health, School of Public Health, Shanghai Jiao Tong University School of Medicine, Shanghai, China b

art ic l e i nf o

a b s t r a c t

Article history: Received 28 February 2015 Received in revised form 4 June 2015 Accepted 6 June 2015

The use of polybrominated diphenyl ethers (PBDEs) has been dramatically increasing over the last two decades in China. Animal studies suggest that prenatal exposure to PBDEs may result in neurodevelopmental deficits. Two hundred thirty-two participating mothers were recruited from a prospective birth cohort in rural northern China between September 2010 and February 2012. We analyzed 232 cord blood specimens for selected PBDE congeners and examined their association with children's developmental quotients (DQs) at 12 (n ¼ 192) and 24 (n ¼149) months of age based on the Gesell Developmental Schedules (motor, adaptive, language, and social domains). There were no substantial differences by demographic characteristics among the three time points: baseline, 12 and 24 months of age. Median cord blood levels of PBDE congeners 47, 99, 100, and 153 were 3.71, 6.70, 2.63, and 2.19 ng/g lipid, respectively. At 12 months of age, neither the individual nor total (the sum of BDEs 47, 99, 100, and 153) congener levels were associated with any of the four domain DQs. However, at 24 months of age, a 10fold increase in BDE-99 levels was associated with a 2.16-point decrease [95% confidence interval (CI):  4.52,  0.20] in language domain DQs and a 10-fold increase in BDE-47 levels was associated with a 1.89-point decrease (95% CI:  3.75,  0.03) in social domain DQs. Prenatal exposure to PBDEs was associated with lower DQs in young children. The results contribute to the growing evidence that PBDEs could act as developmental neurotoxicants,and the findings have implications for children's environmental health in China. & 2015 Elsevier Inc. All rights reserved.

Keywords: Polybrominated diphenyl ethers Children Neurodevelopment Prenatal China

1. Introduction Flame retardants are chemicals that are added to plastics, electronics, textiles, and construction material to protect against fire. Brominated flame retardants (BFRs) are the largest group on the market due to their low cost and high efficiency and account for 39% of worldwide flame retardant production (Mazdai et al., 2003). Within this group, the polybrominated diphenyl ethers (PBDEs) have been used in large quantities as flame retardant additives, and the worldwide market demand for PBDEs increased from 40,000 tons in 1992 to 67,000 tons in 2001 (Hale et al., 2011; n Corresponding author at: MOE and Shanghai Key Laboratory of Children's Environmental Health, Xinhua Hospital, Shanghai Jiao Tong University School of Medicine, 1665 Kongjiang Road, 200092 Shanghai, China. Fax: 86 21 64663944. E-mail address: [email protected] (Y. Tian). 1 Guodong Ding, Jing Yu, and Chang Cui contributed equally to this work

http://dx.doi.org/10.1016/j.envres.2015.06.008 0013-9351/& 2015 Elsevier Inc. All rights reserved.

Voorspoels et al., 2003). Because PBDEs are semivolatile and not chemically bound to substrates, they are more likely to migrate from such substrates during their lifetime. Furthermore, PBDEs are considered to be a group of persistent organic pollutants (POPs) due to similar properties with polychlorinated biphenyls in terms of lipophilicity, bioaccumulation, and persistence in the environment (Hooper and McDonald, 2000; Chen et al., 2011). Human exposure to PBDEs comes primarily from ingestion of dietary products such as fish and cow's milk. Exposure to PBDEs can also occur through dust and inhalation. Airborne contamination has been implicated, particularly in the electronics and computer industries (Mazdai et al., 2003). Over the last several decades, PBDE levels in human blood, breast milk, and adipose tissue have been dramatically increasing globally (Thomsen et al., 2002; Akutsu et al., 2003; Sjödin et al., 2004; Moon et al., 2012; Ni et al., 2013; Chen et al., 2014b), suggesting that the elevated body burden of PBDE in humans could be an important public health issue.

G. Ding et al. / Environmental Research 142 (2015) 104–111

The most pressing current concern for potential adverse health effects of PBDEs relates to their developmental neurotoxicity (Birnbaum and Staskal, 2004; Costa and Giordano, 2007). Numerous animal studies have demonstrated that prenatal or earlylife exposure to PBDEs can cause long-lasting behavioral alterations, including changes in spontaneous motor activity characterized by hyperactivity, decreased habituation, and disruption of learning and memory (Eriksson et al., 2001; Viberg et al., 2003; Kuriyama et al., 2005; Rice et al., 2007; Koenig et al., 2012). Although the association between prenatal exposure to PBDEs and adverse neurodevelopmental effects has been observed in animal models, it has not been adequately explored in human populations. Several human epidemiological studies have investigated the effects of prenatal PBDE exposures on child neurodevelopment, but these studies were mostly conducted in North America and Europe. For example, Herbstman et al. (2010) reported that prenatal BDE-47 levels were adversely associated with the 12month (Bayley Scales of Infant Development-II) Psychomotor Development Index (PDI) (n¼ 118), 24-month Mental Development Index (MDI) (n¼ 117), and 48-month full-scale and verbal IQ (the Wechsler Preschool and Primary Scale of Intelligence, Revised Edition) (n¼ 104) in children from New York. In contrast, Chen et al. (2014) reported that prenatal BDE-47 levels were not associated with PDI or MDI at ages 1–3 years (n ¼285 at 1 year, 239 at 2 years, 220 at 3 years) but negatively associated with Full-Scale IQ (the Wechsler Preschool and Primary Scale of Intelligence-III) and hyperactivity score (the Behavioral Assessment System for Children-2) at age 5 years (n ¼179) in children from Cincinnati. In a study of 62 Dutch children at age 5–6 years, Roze et al. (2009) reported that prenatal PBDE (including BDEs 47, 99, 100, 153, and 154) levels were negatively associated with fine motor coordination and sustained attention, although such levels improved coordination and visual perception and led to fewer internalizing and externalizing behaviors. In China, the nationwide production of BFRs reached 10,000 tons in 2000 and the annual demand for PBDEs has been increasing at a rate of 8%, which would inevitably result in continuous increase of PBDE levels in the environmental media (Jin et al., 2009). Penta- and octa-BDEs were removed from the European (1998) and North American (2004) marketplace; however, deca-BDEs are still being produced and used globally (Huang et al., 2014). At present, there are no legal restrictions on the production and use of penta-BDEs, octa-BDEs, and deca-BDEs in mainland China. With the increasing demand for PBDEs nationwide, concerns regarding the adverse health effects of exposure to PBDEs among some susceptible populations including pregnant women and children are increasing (Chao et al., 2007; Wu et al., 2010). However, little data have been available on prenatal exposure to PBDEs and child neurodevelopment in China. In this report, we investigated the levels of selected PBDE congeners in cord blood and evaluated the possible relationship of prenatal exposure to PBDEs with neurodevelopment as measured by the Gesell Developmental Schedules (GDS) in young children living in rural northern China. We tested the hypothesis that after adjusting for potential confounders, prenatal exposure to PBDEs would be associated with lower DQs in the motor, adaptive, language, and social domains.

105

community in the southern coastal area of Laizhou Wan (Bay) of the Bohai Sea in Shandong province, northern China (LW birth cohort). The adverse consequences of the children were assessed at delivery and during the follow-up, which lasted for a period of at least 2 years. The detailed methods used in this study have been described elsewhere (Ding et al., 2013, 2014). Pregnant women were recruited at the time they were preparing for labor and delivery in a unique county hospital located in the southern coastal area of Laizhou Wan. Eligibility criteria included a singleton pregnancy; age over 18 years old; residence in the area for at least 3 years; and no report of assisted reproduction, pregestational or gestational diabetes, chronic or pregnancy-associated hypertension, HIV infection or AIDS, and illicit drug use (Ding et al., 2013, 2014). From September 2010 to February 2012, a total of 388 women met the eligibility criteria, among whom 347 women agreed to take part in this study (response rate 89.4%). Of these women, 41 women without cord blood samples, 53 women without enough cord serum volumes, and 21 women with missing values for major confounders were excluded. Therefore, 232 mothers were included in this study. Each woman participating in the study signed an informed consent form, and the research protocol for this study was approved by the Medical Ethics Committee of Xinhua hospital, Shanghai Jiao Tong University School of Medicine. Not all mothers agreed to have their child followed after birth. Thus, among these 232 mother–infant pairs, 192 children (82.8%) completed a neurodevelopmental assessment at 12 months of age (71 week), and 149 (64.2%) children at 24 months of age (71 weeks) (Fig. 1). 2.2. Maternal interview and medical record abstraction Standardized face-to-face interviews were conducted with the women shortly after delivery by specially trained nurses in the hospital, as described previously (Ding et al., 2013, 2014). The questionnaire included demographic information (maternal age, education level, household income, and address), maternal characteristics (cigarette smoking, alcohol use, dietary habits, and employment), potential exposures to PBDEs (working in computer or electronics manufacturing, repair, or dismantling plants) and any other chemical exposures. Other relevant information such as previous pregnancy outcomes, current pregnancy complications, date of delivery, gestational age, and sex of newborn was obtained by interview and confirmed by maternal and infant medical records. Maternal pre-pregnancy body mass index (BMI) was calculated as the pre-pregnancy weight divided by the height squared. The women were then classified into three groups according to the 2000 WHO standards: underweight (BMI o18.5 /m2), normal weight (BMI 18.5 to o23.0 /m2), and overweight (BMI 4 23.0 kg/ m2). Low birth weight was defined as o2500 g. Preterm delivery was defined as birth at less than 37 completed gestational weeks. 2.3. Biological sample collection

2. Methods

Cord blood samples were collected from an umbilical vein immediately post-delivery using a syringe and two 10 mL-red topped tubes, allowed to clot, and centrifuged at 1500 rpm for 20 min; next, the serum was decanted into pre-cleaned glass vials for residue analysis. All samples were coded, frozen, and stored at  80 °C until analysis.

2.1. Participants and recruitment

2.4. PBDE exposure assessment

This study was a prospective birth cohort study which began in 2010 to determine the effects of environmental exposures on the health of pregnant women and their children living in a rural

Serum samples were stored at  80 °C until shipment on dry ice to Minzu University of China (Beijing). PBDEs extraction and gravimetric lipid determination procedures have been published

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G. Ding et al. / Environmental Research 142 (2015) 104–111

388 women met the eligibility criteria between 2010 and 2012 Exclusion for 41 women not willing to take part in this study

347 women agreed to take part in this study Exclusion for 41 women without cord blood samples

Exclusion for 53 women without enough cord serum volumes

Exclusion for 21 women with missing values for major confounders 232 women were included in this study

Exclusion for 40 children without neurodevelopmental assessment

(nanograms per gram lipid, including total cholesterol and triglycerides) to account for their lipophilic property. LOD was defined as the mean blank mass plus three standard deviations. The LODs for the eight PBDEs ranged from 0.24 to 0.48 ng/g lipids in cord blood samples. 2.5. Neurodevelopmental measures The Gesell Developmental Schedules (GDS) was selected for comparability to other studies in the Chinese population. The GDS has been adopted by the Chinese Pediatric Association and is widely used for assessing early child development in China and other countries (Tang et al., 2008; Ding et al., 2012; World Health Organization (WHO), 2004). Children in the cohort, who were 12 months and 24 months of age, were administered the GDS version for 0- to 3-year-old children adapted to the Chinese population (Beijing Mental Development Cooperative Group, 1985). Each child was assigned a developmental quotient (DQ) in each of the four specific domains: motor, adaptive, language, and social. The standardized mean (7 SD) of the DQ is 100 715. A score of 84 is the cutoff point for differentiating normal development from developmental delay (Hudon et al., 1998; Tang et al., 2008). Not all children were available for all developmental assessments, resulting in different numbers of children being tested at each age. Testing was conducted by a trained pediatrician to maximize both the reliability of the assessment and the validity of the interpretation. The tester completed a 2-week course and 1-year clinical practice at Xinhua hospital, Shanghai Jiao Tong University School of Medicine and passed standardized exams to become certified. 2.6. Statistical analysis

192 mother–infant pairs were assessed at 12 months of age

Exclusion for 43 children without neurodevelopmental assessment

149 mother–infant pairs were assessed at 24 months of age Fig. 1. Flow chart of inclusion and exclusion subject criteria.

elsewhere (Hovander et al., 2000). Samples were analyzed for eight PBDE congeners (BDEs 28, 47, 85, 99, 100, 153, 154, and 183) using a gas chromatography–mass spectrometer (Agilent Technologies, Palo Alto, CA) with negative chemical ionization (Jin et al., 2009; Cui et al., 2012). Quantification of PBDEs in the samples was determined using the internal standard method. Two masses (m/z: 79.0, 81.0) from the molecular ion cluster were selected for monitoring each of the target analytes and m/z 574.6 and 576.6 were used to monitor 13 C12-BDE-139 as a surrogate standard. The average recovery of 13 C12-BDE-139 was 102.7 710.8% in serum. The method was sufficiently robust to accommodate a satisfying quantitative analysis. To ensure the reproducibility of the measurements of PBDEs levels, one out of each 20 samples was randomly picked, and experiments were subsequently repeated, which demonstrated that the reproducibility of serum analyses was good. Procedural blanks were analyzed simultaneously with each batch of 12 samples to check for interference or contamination from solvents and glassware (Cui et al., 2012). The procedural blanks were less than the limit of detection (LOD) for all PBDE congeners in the samples. The serum PBDE levels were calculated on a lipid basis

Initial descriptive statistics were provided for characteristics of the study population, individual PBDE congeners, and DQ scores. To examine the association between PBDE levels and GDS performance, we constructed separate linear regression models for each of the four domains. We were unable to consider DQ scores as dichotomous measures, using the recommended cutoffs for defining children as “delayed” because of the small sample size and the small number of children who met these criteria (Herbstman et al., 2010). PBDE congeners frequently detected in cord serum (detected in 480% of samples) were handled as continuous variables in the statistical models (Herbstman et al., 2010). This was the case for BDEs 47, 99, 100, and 153. We used the LOD divided by the square root of two for levels below the LOD (Hornung and Reed, 1990), and this value was included in each sum. The four congeners were log10 transformed because of the right-skewed distribution. Because a large proportion (approximately one-third) of BDEs 28 and 85 had levels below the LOD, we analyzed thetwocongeners as categorical variables. For each congener, children were assigned to one of three groups: no detectable levels (reference group), detectable levels below the median of the detectable levels, and detectable levels above the median. We did not analyze BDEs 154 and 183 because only 39.2% and 37.1% of the samples, respectively, had detectable levels for these BDEs. Maternal age, education, pre-pregnancy BMI, weight gain during pregnancy, smoking during pregnancy, marital status, household monthly salary, child sex, gestational age, birthweight, and parity were initially considered as confounders. Confounders were finally selected for regression models if they were related to neurodevelopment in the literature and were associated (po 0.15) with 2 or more GDS DQs (outcomes). Potential confounders included child sex, maternal age, education, smoking during pregnancy, and household monthly salary. We included the same

G. Ding et al. / Environmental Research 142 (2015) 104–111

confounders in all regression models. Alcohol use was not included in the models because notably few women reported it. Statistical analyses were carried out using SPSS software (SPSS Inc., Chicago, IL) based on two-tailed tests, and p o0.05 indicated statistical significance.

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Table 2 Detection frequency, range, and percentile of cord serum PBDE levels (ng/g lipid) (n ¼232) PBDE Congener

Detection no. (%)

Range

Selected percentiles 25th

3. Results Table 1 describes the sociodemographic characteristics of the study population. At baseline (n ¼232), the average maternal age was 28.1 years (SD ¼ 4.8), the vast majority (97.8%) were married, two-thirds (65.5%) were primiparous, and half (50.4%) had graduated from high school or above. The majority of the women (92.7%) lived in households with a monthly income less than RMB (¥) 5000 yuan. More than half (56.5%) of the women had a normal weight before pregnancy. Almost one-third (32.8%) of the women lived with a smoker during pregnancy, although few smoked or consumed alcohol regularly. None of the mothers reported any Table 1 Demographic characteristics of the study population Characteristic

Maternal characteristic Maternal age (years) o25 25–29 30–34 Z35 Marital status Married or living as married Single Parity 0 (Primiparous) Z1(Multiparous) Education (years) r9 (Middle school) 10–12 (High school) Z13(Greater than high school or college) Household monthly salary (RMB) o000 3000–5000 45,000 Pre-pregnancy BMI (kg/m2) o18.5 18.5 to o 23.0 Z23.0 Smoking during pregnancy Yes Lived with smoker No Alcohol use during pregnancy Yes No Infant characteristic Sex Male Female Gestational age (weeks) o37 Z37 Birth weight (g) o2500 Z2500

(29.7%) (34.5%) (25.0%) (10.8%)

57 71 43 21

158 (68.1%)

BDE-47

217 (93.5%)

BDE-85

157 (67.7%)

BDE-99

217 (93.5%)

BDE-100

199 (85.8%)

BDE-153

191 (82.3%)

BDE-154

91 (39.2%)

BDE-183

86 (37.1%)

o LOD– 814.31 o LOD– 582.27 o LOD– 404.80 o LOD– 812.82 o LOD– 130.02 o LOD– 100.24 o LOD– 38.79 o LOD– 842.38

75th 95th

oLOD 2.05

4.28

27.60

2.08

3.71

5.66

49.11

oLOD 1.47

2.70

10.10

4.43

6.70

9.44

63.66

1.36

2.63

4.13

11.84

0.94

2.19

4.31

11.13

oLOD

o LOD 1.27

6.72

oLOD

o LOD 2.06

10.79

Abbreviations: BDE – brominated diphenyl ether; oLOD – below the limit of detection; PBDE – polybrominated diphenyl ether.

Baselinen ¼ 232 (%) 12 months of 24 months of age n¼149 age n¼ 192 (%) (%)

69 80 58 25

BDE-28

50th

(29.7%) (37.0%) (22.4%) (10.9%)

47 (31.5%) 53 (35.6%) 34 (22.8%) 15 (10.1%)

227 (97.8%)

189 (98.4%)

146 (98.0%)

5 (2.2%)

3 (1.6%)

3 (2.0%)

152 (65.5%) 80 (34.5%)

124 (64.6%) 68 (35.4%)

95 (63.8%) 54 (36.2%)

115 (49.6%) 64 (27.6%) 53 (22.8%)

96 (50.0%) 52 (27.1%) 44 (22.9%)

77 (51.7%) 42 (28.2%) 30 (20.1%)

152 (65.5%) 63 (27.2%) 17 (7.3%)

128 (66.7%) 52 (27.1%) 12 (6.2%)

100 (67.1%) 41 (27.5%) 8 (5.4%)

34 (14.7%) 131 (56.5%) 67 (28.9%)

30 (15.6%) 101 (52.6%) 61 (31.8%)

21 (14.1%) 82 (55.0%) 46 (30.9%)

2 (0.9%) 76 (32.8%) 154 (66.4%)

2 (1.0%) 67 (34.9%) 123 (64.1%)

1 (0.7%) 53 (35.6%) 95 (63.8%)

2 (0.9%) 230 (99.1%)

1 (0.5%) 191 (99.5%)

1 (0.7%) 148 (99.3%)

120 (51.7%) 112 (48.3%)

98 (51.0%) 94 (49.0%)

79 (53.0%) 70 (47.0%)

14 (6.0%) 218 (94.0%)

9 (4.7%) 183 (95.3%)

6 (4.0%) 143 (96.0%)

11 (4.7%) 221 (95.3%)

7 (3.6%) 185 (96.4%)

5 (3.4%) 144 (96.6%)

work-related potential for exposure to PBDEs. Overall, 51.8% of the newborn infants were male. The mean birth weight was 3347.1 g (SD ¼426.1), and the mean gestational age was 39.4 week (SD ¼1.5). A total of 4.7% (n¼ 11) of the infants had a low birth weight, and 6.0% (n ¼14) were born preterm. These infants were included in our analyses. There were no substantial differences by demographic characteristics among the three time points: baseline, 12 months of age (n ¼192), and 24 months of age (n ¼149), indicating that the studied cohort generally reflects the original one. The cord serum levels of eight selected PBDE congeners are summarized in Table 2. Four PBDE congeners (BDEs 47, 99, 100, and 153) were detected in the majority (480%) of the cord blood specimens. The congener with the highest serum level was BDE-99 (median 6.70 ng/g lipid), followed by BDE-47 (3.71 ng/g lipid), BDE-100 (2.63 ng/g lipid), and BDE-153 (2.19 ng/g lipid). These congeners are considered to be primary components of the commercial mixture known as penta-BDEs. Four other congeners (BDEs 28, 85, 154, and 183) were above the limit of detection for approximately 40–70% of the population. The median levels of BDEs 28 and 85 were 2.05 and 1.47 ng/g lipid, respectively. The median levels could not be calculated for BDEs 154 and 183 because the two congeners had a detection frequency of less than 50%. The eight congeners were moderately associated with each other (r ¼0.19–0.76, p o0.05). The distribution of GDS scores for the young children is shown in Table 3. There were 192 and 149 children with available cord PBDE measurements who also had a developmental assessment at 12 and 24 months of age, respectively. The means of the motor, adaptive, language and social scores for 1-year-old children were 98.1 (SD ¼9.4), 102.0 (SD ¼ 7.5), 99.6 (SD ¼7.6), and 101.2 (SD ¼7.5) , respectively; and for 2-year-old children were 103.1 (SD ¼6.9), 102.8 (SD ¼9.9), 97.5 (SD ¼10.0), and 103.8 (SD ¼9.3), respectively. The frequency of developmental delay ranged from 1.6% (motor) to 6.3% (language) for 12-month children and from 1.3% (motor) to 6.0% (language) for 24-month children. These infants were included in our analyses. For the 12-month and 24-month old children, all DQ domains were moderately intercorrelated (p o0.01), with r values ranging from 0.31 to 0.56. Table 4 presents the association [95% confidence intervals (CIs)] between prenatal exposure to PBDEs and neurodevelopmental scores at 12 (n ¼192) and 24 (n ¼149) months of age. At 12 months

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G. Ding et al. / Environmental Research 142 (2015) 104–111

Table 3 Score distribution of the Gesell Developmental Schedules in children at 12 and 24 months of age. Developmental quotienta Motor domain Mean 7 SD (range) Normal [n (%)] Developmental delay Adaptive domain Mean 7 SD (range) Normal [n (%)] Developmental delay Language domain Mean 7 SD (range) Normal [n (%)] Developmental delay Social domain Mean 7 SD (range) Normal [n (%)] Developmental delay a

12 months of age (n¼ 192)

24 months of age (n¼ 149)

[n (%)]

98.17 9.4 (78–129) 189 (98.4%) 3 (1.6%)

103.17 6.9 (81–125) 147 (98.7%) 2 (1.3%)

[n (%)]

102.0 7 7.5 (76–133) 188 (97.9%) 4 (2.1%)

102.8 79.9 (71–132) 142 (95.3%) 7 (4.7%)

[n (%)]

99.6 7 7.6 (68–131) 180 (93.8%) 12 (6.3%)

97.5 710.0 (69–138) 140 (94.0%) 9 (6.0%)

[n (%)]

101.2 7 7.5 (70–121) 185 (96.4%) 7 (3.6%)

103.8 79.3 (75–127) 145 (97.3%) 4 (2.7%)

Normal, 484; developmental delay, r 84.

of age, both the individual and total (the sum of BDEs 47, 99, 100, and 153) congener levels were not associated with any of the four domain DQs. However, at 24 months of age, a 10-fold (i.e., one log-

unit) increase in BDE-99 levels was associated with a 2.16-point decrease (95% CI,  4.52 to  0.20; p ¼0.04) in language domain DQs, and a 10-fold increase in BDE-47 levels was associated with a

Table 4 Association (95% CIs) between prenatal exposure to PBDEs and neurodevelopmental scores at 12 and 24 months of age. Exposure

Outcome Motor domaina

12 months of age (n¼192) Continuous variables (ng/g lipid, log10 scale) BDE-47 (n¼ 232)  0.44 (  3.04, 2.15) BDE-99 (n¼ 232)  0.59 (  3.91, 2.73) BDE-100 (n ¼232) 0.71 (  1.77, 3.18) BDE-153 (n¼ 232)  1.89 (  5.25, 1.46) Σ4PBDEs  1.42 (  4.69, 1.84) Categorical variables (ng/g lipid) BDE-28 No detectable levels (n ¼74) Referent Detectable levelsomedian  0.43 (  2.37, 1.51) (n ¼79) 0.26 (  2.18, 2.69) Detectable levelsZmedian (n ¼79) BDE-85 No detectable levels (n ¼75) Referent Detectable levelsomedian –0.34 (–2.11, 1.44) (n ¼78) Detectable levelsZmedian  0.85 (  3.67, 1.98) (n ¼79) 24 months of age (n¼ 149) Continuous variables (ng/g lipid, log10 scale) BDE-47 (n¼ 232)  0.02 (  2.29, 2.24) BDE-99 (n¼ 232)  1.24 (–4.16, 1.69) BDE-100 (n ¼232) 0.45 (  1.65, 2.55) BDE-153 (n¼ 232) 1.38 (  0.94, 3.69) Σ4PBDEs  0.28 (  3.16, 2.60) Categorical variables (ng/g lipid) BDE-28 No detectable levels (n ¼74) Referent Detectable levelsomedian 0.26 (  2.05, 2.56) (n ¼79) Detectable levelsZmedian 0.87 (  0.56, 2.30) (n ¼79) BDE-85 No detectable levels (n ¼75) Referent Detectable levelsomedian  0.77 (  2.87, 1.33) (n ¼78) Detectable levelsZmedian 0.81 (  0.66, 2.28) (n ¼79) a #

Adaptive domaina

Language domaina

Social domaina

0.89 (  1.18, 2.97)  1.11 (  3.09, 0.88) 0.35 (  1.67, 2.37) 0.23 (  2.04, 2.50) 0.82 (  1.74, 3.38)

 1.82 (  3.92, 0.28)  0.99 (  3.60, 1.62)  0.71 (  2.77, 1.34) 0.97 (  1.34, 3.28)  0.84 (  3.45, 1.78)

0.14 (  1.97, 2.24) 1.07 (  1.66, 3.80)  0.62 (  2.66, 1.42)  0.64 (  2.93, 1.66) 0.78 (  1.81, 3.38)

Referent 0.37 (  1.16, 1.89)

Referent 0.73 (  0.82, 2.28)

Referent  0.58 (  2.49, 1.33)

 0.96 (  2.78, 0.86)

0.46 (  0.71, 1.64)

 0.71 (  3.02, 1.61)

Referent –0.41 (–2.25, 1.43)

Referent –0.75 (–2.62, 1.13)

Referent 0.12 (–1.75, 1.98)

0.26 (  0.89, 1.41)

 0.61 (  2.54, 1.32)

0.37 (  0.80, 1.53)

1.11 (  2.20, 4.43) 0.73 (  3.73, 5.18)  0.94 (  3.01, 1.12)  0.96 (  4.37, 2.44)  0.51 (  4.72, 3.71)

 1.35 (  4.70, 1.99)  2.16 (  4.52,  0.20)#  1.23 (–3.32, 0.86) 0.59 (  2.87, 4.04)  2.87 (  6.12, 0.38)

 1.89 (  3.75,  0.03)#  3.44 (  7.39, 0.51)  0.81 (  3.69, 2.06) 0.88 (  2.31, 4.07)  2.55 (  6.47, 1.37)

Referent  0.86 (  3.35, 1.64)

Referent 0.36 (  3.04, 3.75)

Referent  1.07 (  3.40, 1.26)

 0.32 (  3.78, 3.14)

 0.86 (3.39, 1.67)

 1.84 (  5.08, 1.40)

Referent  0.32 (  3.67, 3.02)

Referent 0.60 (  2.52, 3.72)

Referent 0.23 (  2.65, 3.11)

 0.36 (  2.50, 1.79)

0.41 (  3.00, 3.81)

 0.28 (  2.29, 1.74)

Model included child sex, maternal age, maternal education, maternal smoking during pregnancy, and household monthly salary as covariates. p o0.05.

G. Ding et al. / Environmental Research 142 (2015) 104–111

1.89-point decrease (95% CI,  3.75 to  0.03; p¼ 0.05) in social domain DQs. At 12 months of age, we repeated the analyses using the same individuals as at the age of 24 months (n ¼149). Similar to the results from the 192 samples, we did not find any associations between PBDE exposures and DQ scores (Online Supplemental Table 1). We further examined the possible association between prenatal exposure to PBDEs and neurodevelopment stratified by child gender at 12 and 24 months of age. However, we failed to find any correlations betweenPBDE exposures and DQ scores (data not shown). Because a considerable proportion (approximately one-third) of the BDEs 28 and 85had levels below the limits of detection, we analyzed thetwo congeners ascategorical variables.After adjusting for potential confounders, linear regression models were also used to evaluate whether having detectable levels of BDEs 28 and 85 were associated with lower DQ scores, but there were no significant relationshipsof BDEs 28 and 85levels with DQ scores at both ages (Table 4).

4. Discussion In the present study, we demonstrated that prenatal exposure to BDEs 99 and 47, assessed from cord blood at delivery, was associated with lower DQs in language and social domains at 24 months of age, respectively. Although no associations were found between PBDE congeners and any of the four domain DQs at 12 months of age, these results contribute to the growing body of evidence that PBDEs may act as developmental neurotoxicants. Severalepidemiological studies have investigated the possible relationship between prenatal exposure to PBDEs and child neurodevelopment. The literature from North America and Europe, such as the New York City cohort, the HOME study, the CHAMACOS cohort, and the Menorca birth cohort (INMA project, Spain), was interesting. Herbstman et al. (2010) analyzed PBDEs in cord blood, and found that adverse associations were significant for 12month PDI (BDE-47) (n¼ 118), 24-month MDI (BDEs 47, 99, and 100) (n ¼117), 3-year MDI (BDE-100) (n ¼114), 4-year full-scale and verbal IQ (BDEs 47, 99, and 100) and performance IQ (BDE100) (n ¼104), and 5-year performance IQ (BDE-100) (n ¼96).Chen et al. (2014) reported that prenatal maternal exposure to BDE-47 was not associated with MDI or PDI at age 1–3 years (n ¼ 285 at 1 year, 239 at 2 years, 220 at 3 years), but was associated with decreased Full-Scale IQ and increased hyperactivity score at age 5 years (n ¼179). Eskenazi et al. (2013) reported that prenatal maternal PBDE exposures (sum of BDEs 47, 99, 100, and 153) were associated with poorer attention, fine motor coordination, and cognition at 5 (n ¼310) and 7 (n ¼ 323) years of age. Gascon et al. (2011) reported scores for cognitive and motor functions at 4 years of age were always lower in children prenatally exposed to PBDE47 than in referents, although none of these associations was significant (n ¼88). It should be noted that some of the observed inconsistencies may be explained by important differences in exposure scenario or measurement, neurodevelopmental assessment method, sample size, and statistical analysis. The neurodevelopmental deficits associated with prenatal exposure to PBDEs that were detected both at early ages (12–36 months of age) and at pre-school ages (4–7 years of age) may be important predictors of subsequent academic performance and even long-term behavioral changes. The reason is that these indicators can identify “delayed” or “borderline delayed” children who could benefit from early intervention programs (Kaplan, 1993; Herbstman et al., 2010). The results of our study are consistent with published toxicological experiments. Numerous animal studies have indicated

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that prenatal exposure to various PBDEs may cause long-lasting changes in spontaneous motor activity, mostly characterized as hyperactivity or decreased habituation, and disruption in performance in learning and memory tests (Costa and Giordano, 2007). Despite the exact impact of PBDEs on developmental neurotoxicity is not fully elucidated, it has been suggested that at least two modes of action may exist, one mediated by an effect on thyroid hormones, the other due to a direct effect on brain cells (Fonnum and Mariussen, 2009; Costa and Giordano, 2011; Dingemans et al., 2011). The structural similarities of PBDEs to thyroid hormone thyroxine (T4) suggest that prenatal exposure to PBDEs is of particular interest because these compounds can cross the placenta and could interfere with T4 levels (Zhou et al., 2002). Thyroid hormones are known to play an important role in fetal brain development, particularly for T4, and lower levels of T4 are associated with impaired brain development (Williams, 2008). Interestingly, many animal toxicology studies consistently show a decrease in T4 levels following exposure to PBDEs, which has led to the current hypotheses related to an enhanced metabolism and excretion of T4 or to an interaction of PBDEs with the thyroid hormone transport system (Hallgren et al., 2001; Zhou et al., 2002; Ellis-Hutchings et al., 2006). However, several epidemiological studies have examined the relationship between prenatal PBDE exposure and thyroid function. Mazdai et al. (2003) found no relationship between total PBDEs in cord blood and cord blood total T4 (TT4) and free T4 (FT4) among 9 babies, but this study lacked sufficient statistical power. In contrast, Herbstman et al., (2008) measured PBDEs, T4, and FT4 in cord blood from 297 babies, and found that BDE-100 was associated with increased odds of low TT4, and BDE-153 with increased odds of low TT4 and FT4. PBDEs can also exert direct effects on brain function and structure including oxidative stress in neurons, neuronal apoptosis, interrupted intracellular calcium (Ca2 þ ) homeostasis, disrupted signal transduction such as protein kinase C, and altered neurotransmitter release and function (Dingemans et al., 2011). A recent study has shown that in primary fetal human neural progenitor cells, BDE-47 and BDE-99 decreased migration and differentiation into neurons and oligodendrocytes. It is interesting to note that triiodothyronine (T3) antagonized the effect of PBDEs, indicating that PBDEs may affect thyroid hormone signaling in this in vitro model (Schreiber et al., 2010). Although brain development in humans extends beyond early childhood, the perinatal period, particularly the third trimester, has been proven to be very sensitive to neurotoxic effects (Rice and Barone, 2000). Nevertheless, it is currently difficult to identify which time windows of exposure are most relevant for PBDE-induced developmental neurotoxicity in humans. The literature shows that BDEs 47, 99, 100, and 153 are the dominant and most common congeners in most human samples (Hites, 2004; Herbstman et al., 2010; Eskenazi et al., 2013); and these congeners are also the primary components of commercial mixtures known as penta-BDE. However, levels of cord blood PBDEs in our population are in general lower than those reported in North America (the US and Canada) (Herbstman et al., 2007, 2010; Mazdai et al., 2003; Foster et al., 2011) but substantially higher than those reported in Asia (other domestic areas such as Guangzhou and Taiwan, South Korea, and Japan) (Chen et al., 2014b; Lin et al., 2011; Kim et al., 2009; Kawashiro et al., 2008) and Europe (Spain, Sweden, France, and Denmark) (Vizcaino et al., 2011; Guvenius et al., 2003; Antignac et al., 2009; Frederiksen et al., 2010) (Online Supplemental Table 2). The reason for the disparity between our results and those of previous studies is not well-known. Recent data suggest that PBDE levels in the European Union (EU) have dramatically decreased because the use of penta-, octa-, and deca-BDE is prohibited (Bromine Science and Environmental Forum (BSEF), 2007). However, deca-BDE is still produced

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and used in the USA, except in Maine, and the high human body burden of PBDEs may be attributable to higher standards for flammable protection and be more developed in the electronics industry (Mazdai et al., 2003). Our study population had much higher PBDE levels than those reported in other domestic areas, such as Guangdong and Taiwan (Chen et al., 2014b; Lin et al., 2011). Many of the BFR manufactories located in the eastern coastal area, particularly in the Shandong province of China, may explain this phenomenon (Jin et al., 2008). Similar to other POPs, such as PCBs and organochlorine pesticides, PBDEs can cross the placenta barrier into the fetal circulation. A previous study also found evidence that maternal and fetal blood PBDE levels were highly correlated (Mazdai et al., 2003). Measuring PBDEs in cord blood is noninvasive and may be used as a direct measure of prenatal exposure (Herbstman et al., 2007). Therefore, as with studies on the effects of PBDE exposure (Herbstman et al., 2010; Gascon et al., 2011), we used PBDE levels in cord blood as an indicator of fetal exposure during gestation. However, it should be noted that the biologic significance and bioavailability of PBDEs in fetal circulation during this period of gestation remain speculative (Mazdai et al., 2003). To the best of our knowledge, this study is the first in mainland China to assess the possible adverse effects of prenatal exposure to PBDEs on child neurodevelopment. However, our study also has several limitations. First, although several important sociodemographic factors have been adjusted as potential confounders, it is possible that unmeasured confounding cannot be ruled out, especially environmental and lifestyle factors which may affect both exposure to PBDEs and child neurodevelopment (Buttke et al., 2013; Chen et al., 2014). Second, our sample size is not large, and considerable proportion losses to follow-up may have limited our power to detect significant differences in multivariate models. Third, we did not measure some higher-brominated congeners (e.g., BDE-209) which are present in deca-BDE. Fourth, gender differences in susceptibility to endocrine-disrupting chemicals (including PBDEs) might exist. However, substantial gender differences were not clear due to the small sample size. In addition, the use of a self-questionnaire for maternal cigarette exposure may be questionable, but a previous study found that self-report smoking was highly correlated with measured cotinine levels in pregnancy (Pickett et al., 2005). In summary, prenatal exposure to PBDEs was associated with lower DQs in the LW birth cohort of young children. This study further contributes to the growing evidence that PBDEs could act as developmental neurotoxicants. Further research is required to illustrate the mechanistic pathways linking PBDE exposure and neurodevelopmental deficits and to investigate BDE‑209 for its human developmental toxicity. In the meantime, it is important to take precautions to minimize human exposure to PBDEs wherever possible.

Acknowledgments We thank the Department of Environmental Health's staff, students, hospital partners, participants and families, without whom this study would not have been possible. We specifically thank Jun Jin (Minzu University of China, Beijing) for the work on specimen management and PBDE measurement. This study was supported by the National Natural Science Foundation of China (Grant nos 81172625 and 81402645), the National Basic Research Program of China (973 Program Grant no. 2014CB943300), the Excellent Young Doctor Training Program of Shanghai Pudong New Area (PWRq2014–30), and the Excellent Young Doctor Training Program of Tongji University (2014KJ085).

The authors declare they have no actual or potential competing financial interests.

Appendix A. Supplementary material Supplementary data associated with this article can be found in the online version at http://dx.doi.org/10.1016/j.envres.2015.06. 008

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