Exposure to an environmentally relevant mixture of brominated ... - Core

Toxicology 320 (2014) 56–66

Contents lists available at ScienceDirect

Toxicology journal homepage: www.elsevier.com/locate/toxicol

Exposure to an environmentally relevant mixture of brominated flame retardants affects fetal development in Sprague-Dawley rats Robert G. Berger a , Pavine L.C. Lefèvre a , Sheila R. Ernest a , Michael G. Wade c , Yi-Qian Ma a , Dorothea F.K. Rawn d , Dean W. Gaertner d , Bernard Robaire a,b , Barbara F. Hales a,∗ a

Department of Pharmacology and Therapeutics, McGill University, Montreal, QC, Canada H3G 1Y6 Department of Obstetrics and Gynecology, McGill University, Montreal, QC, Canada H3G 1Y6 c Environmental Health Science and Research Bureau, Healthy Environments and Consumer Safety Branch, Health Canada, Ottawa, ON, Canada K1A 0K9 d Food Research Division, Bureau of Chemical Safety, Health Products and Food Branch, Health Canada, Ottawa, ON, Canada K1A OL2 b

a r t i c l e

i n f o

Article history: Received 7 August 2013 Received in revised form 12 March 2014 Accepted 14 March 2014 Available online 23 March 2014 Keywords: Environmental contaminant House dust Polybrominated diphenyl ethers Mixture In utero Skeletal development

a b s t r a c t Brominated flame retardants are incorporated into a wide variety of consumer products and are known to enter into the surrounding environment, leading to human exposure. There is accumulating evidence that these compounds have adverse effects on reproduction and development in humans and animal models. Animal studies have generally characterized the outcome of exposure to a single technical mixture or congener. Here, we determined the impact of exposure of rats prior to mating and during gestation to a mixture representative of congener levels found in North American household dust. Adult female Sprague-Dawley rats were fed a diet containing 0, 0.75, 250 or 750 mg/kg of a mixture of flame retardants (polybrominated diphenyl ethers, hexabromocyclododecane) from two weeks prior to mating to gestation day 20. This formulation delivered nominal doses of 0, 0.06, 20 and 60 mg/kg body weight/day. The lowest dose approximates high human exposures based on house dust levels and the dust ingestion rates of toddlers. Litter size and resorption sites were counted and fetal development evaluated. No effects on maternal health, litter size, fetal viability, weights, crown rump lengths or sex ratios were detected. The proportion of litters with fetuses with anomalies of the digits (soft tissue syndactyly or malposition of the distal phalanges) was increased significantly in the low (0.06 mg/kg/day) dose group. Skeletal analysis revealed a decreased ossification of the sixth sternebra at all exposure levels. Thus, exposure to an environmentally relevant mixture of brominated flame retardants results in developmental abnormalities in the absence of apparent maternal toxicity. The relevance of these findings for predicting human risk is yet to be determined. © 2014 The Authors. Published by Elsevier Ireland Ltd. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/3.0/).

1. Introduction Brominated flame retardants (BFRs) are incorporated into a wide variety of household and commercial goods to reduce flame propagation and ignition rates (Camino et al., 1991). Many BFRs do not form covalent bonds with the polymer matrices in products to which they are added and leach out into the environment. Recently, a major class of BFRs, the polybrominated diphenyl ethers (PBDEs), were withdrawn from commerce in North America and Europe

Abbreviations: BFRs, brominated flame retardants; PBDEs, polybrominated diphenyl ethers; HBCD, hexabromocyclododecane. ∗ Corresponding author at: Department of Pharmacology and Therapeutics, McGill University, 3655 Promenade Sir William Osler, Montreal, QC, Canada H3G 1Y6. Tel.: +1 514 398 3610; fax: +1 514 398 7120. E-mail address: [email protected] (B.F. Hales). http://dx.doi.org/10.1016/j.tox.2014.03.005 0300-483X/© 2014 The Authors. Published by (http://creativecommons.org/licenses/by-nc-nd/3.0/).

Elsevier

Ireland

Ltd.

due to their environmental persistence and bioaccumulative properties (EC, 2011; Tullo, 2003). Despite this, products containing PBDEs, such as furniture or electronic devices, continue to be used (Stapleton et al., 2012). In spite of their being banned in North America, analyses of house dust show that BFRs are present in the majority of North American homes; PBDEs and hexabromocyclododecane (HBCD) are prominent among the BFRs that are detected (Allen et al., 2008; Stapleton et al., 2008) and have not declined appreciably since being removed from commerce (Dodson et al., 2012); recent studies demonstrate that BFRs are still found in the serum of pregnant women (Zota et al., 2013). Brominated flame retardants are detected in various human tissues, with the highest concentrations detected in North Americans (Hites, 2004; Sjödin et al., 2004; Toms et al., 2009). Primary routes of exposure have been identified as inhalation and ingestion of dust (Johnson et al., 2010; Stapleton et al., 2005; Wu et al., 2007) and dietary consumption (Fraser et al., 2009; Wu et al., 2007).

This

is

an

open

access

article

under

the

CC

BY-NC-ND

license

R.G. Berger et al. / Toxicology 320 (2014) 56–66

Recent investigations have shown that BFRs are detected in serum (Johnson et al., 2010; Meeker et al., 2009), hair (Aleksa et al., 2012), adipose tissue (Johnson-Restrepo et al., 2005), breast milk and ovarian follicular fluid (Johnson et al., 2012; Petro et al., 2012; Marchitti et al., 2013). Children have the highest serum BFR concentrations, with peak measurements up to 2–10-fold higher than in adults (Rose et al., 2010). Increased BFR levels in infants and toddlers may be a result of additional exposure through hand to mouth activities (Stapleton et al., 2008, 2012) and ingestion of contaminated breast milk (Carrizo et al., 2007; Siddique et al., 2012). Prenatal development is also vulnerable to BFR exposure through placental transfer (Mazdai et al., 2003). In the Montreal area, increasing levels of PBDEs were found in human fetal livers from 1998 through 2006, mirroring trends in adults and children (Doucet et al., 2009). Exposure to BFRs during in utero development is of significant concern. Indeed, maternal PBDE levels during pregnancy have been associated with deficits in children’s attention, fine motor coordination, and cognitive functioning (Eskenazi et al., 2013). PBDEs are structurally similar to thyroid hormone and may disrupt thyroid homeostasis (Darnerud et al., 2001; Ernest et al., 2012; Fowles et al., 1994; Hallgren et al., 2001; Blake et al., 2011). Alternatively, BFRs may act as endocrine disruptors by altering estrogen (Ceccatelli et al., 2006; Meerts et al., 2001) or androgen signaling (Stoker et al., 2005), or perturbing insulin-like growth factor 1 (IGF1) signaling (Haave et al., 2011; Suvorov et al., 2009; Suvorov and Takser, 2010). In addition to their effects on the brain and developing nervous system (Dreiem et al., 2010; Saegusa et al., 2009; van der Ven et al., 2008; Ta et al., 2011), exposure to BFRs has been associated with detrimental effects on the male and female reproductive systems (Talsness et al., 2008; van der Ven et al., 2008; Ema et al., 2008; Lilienthal et al., 2006; Zhang et al., 2013), bone (van der Ven et al., 2009; Blanco et al., 2012), the immune system (van der Ven et al., 2009; Liu et al., 2012; Bondy et al., 2013), and the liver (Blanco et al., 2012; Ema et al., 2008). The consequences of exposure to single BFR congeners, such as PBDE-47 (Talsness et al., 2008; Ta et al., 2011) or PBDE-99 (Talsness et al., 2005), or to BFR technical mixtures, such as DE-71 (Blake et al., 2011) have been reported. For example, in utero exposure to PBDE-99 delayed bone ossification and enlarged the liver (Blanco et al., 2012), decreased fertility in females (Talsness et al., 2005), and altered steroid hormone production and the onset of puberty (Lilienthal et al., 2006). Bone development was altered in animals exposed to HBCD from premating until weaning, with a decrease in trabecular bone mineral density of the tibia in females (van der Ven et al., 2009). Although humans are exposed to a complex mixture of BFRs, animal studies have focused on the effects of a single congener or technical mixture. Here, we determined the effects of an environmentally relevant BFR mixture, typically found in household dust, on female fertility, the establishment and maintenance of pregnancy, and fetal development in Sprague-Dawley rats.

2. Materials and methods 2.1. BFR mixture formulation Formulation of the BFR mixture was described previously (Ernest et al., 2012). Briefly, three technical PBDE mixtures (DE-71, DE-79 and BDE-209) and one HBCD mixture were combined to yield a ratio of PBDE congeners and HBCD comparable to the median levels observed in Boston house dust (Allen et al., 2008; Stapleton et al., 2008). This BFR mixture was incorporated into an isoflavone-free diet (Teklad Global 2019 diet; Harlan Laboratories, Madison, WI) with 4.3 g/kg corn oil. Diets were formulated to contain 0, 0.75, 250 or 750 mg of BFR mixture/kg. Diet samples from each experimental condition were collected for later analysis of BFR content via gas

57

chromatography/mass spectrometry (GC/MS), as described below. The diet formulations were intended to deliver nominal doses of 0, 0.06, 20 and 60 mg/kg body weight/day, assuming a daily food consumption of 80 g/kg body weight/day. The lowest dose was estimated to be a close approximation of maximum human exposure, based on a dust ingestion rate of 100 mg/day in children (16.5 kg body weight) and the scaling of dose from humans to rodents (1:6.9, human to rat body surface area ratio) (Allen et al., 2008; Stapleton et al., 2008). 2.2. GC/MS analysis of dietary BFR content Feed samples were thawed and ground to a fine powder, thoroughly mixed and weighed (∼0.1 g) into Erlenmeyer flasks. Each sample was fortified with surrogate standards (twelve 13 C12 labeled PBDE congeners: 15, 28, 47, 77, 99, 100, 126, 138, 153, 154, 183, 209 and 13 C12 labeled ␣-, ␤- and ␥-HBCD) (Cambridge Isotope Laboratories, Andover, MA and Wellington Laboratories, Guelph, ON, Canada). Distilled water (20 mL) was added to each sample and allowed to soak for approximately 30 min, followed by the addition of 50 mL acetone:hexane (2:1, v/v). Samples were homogenized for approximately 30 s using a Silverson homogenizer. Solid material was removed from the extract by transferring the extract to a separatory funnel through a funnel packed with glass wool. Aqueous sodium chloride (50 mL) was added to each separatory funnel, and the funnel was gently shaken and vented as necessary. Following phase separation, the aqueous layer was drained from the funnel, discarded and the extract was transferred to a round bottomed flask (rbf) after passing through a bed of anhydrous sodium chloride. Each sample was concentrated to approximately 1 mL using rotary evaporation and quantitatively transferred to a clean pre-weighed vial which was placed, uncapped, in the fume hood until a stable weight was achieved, to allow for lipid content determination. Samples were then re-dissolved in hexane (1–2 mL) with gentle swirling and added to a column containing 2% deactivated Florisil (6 g), topped with approximately 1 cm of each of glass beads and anhydrous sodium sulfate (Na2 SO4 ). Once the extract was washed onto the column and drained to within a few millimeters of the top of the Na2 SO4 , enough hexane was added to wet the entire column (∼5 mL). PBDEs were eluted with 70 mL hexane and collected in a 250 mL rbf; the flask was then removed and retained. A second rbf was placed under the Florisil column and HBCD isomers were eluted using 70 mL dichloromethane (DCM): hexane (30:70, v/v). Both fractions were reduced in volume to ∼1 mL using rotary evaporation. The PBDE extract was transferred to a v-notch vial and evaporated to dryness using a gentle stream of nitrogen, re-diluted in isooctane and mixed. This final isooctane extract was transferred to a chromatographic vial for analysis. The HBCD fraction was similarly transferred to a v-notch vial, but taken only to near dryness (∼50 ␮L) using a gentle stream of nitrogen prior to the addition of C12 2 H18 Br6 analogs of ␣-, ␤- and ␥-HBCD. The samples were then placed in a fume hood where they remained until dryness was attained. To each dry HBCD extract, 100 ␮L of methanol:water (80:20, v/v) was added, the vial was mixed and the final extract was transferred to a chromatography vial. Dilution of feed sample extracts, to which higher levels of PBDEs and HBCD had been added (e.g., 20 ppm, 60 ppm nominal concentrations), was necessary to ensure that accurate concentration measurements could be achieved. PBDE analysis was performed using an Agilent 6890 gas chromatograph coupled to a 5973N mass selective detector (Agilent, Mississauga, ON, Canada). A 15 m J&W DB5 MS (0.25 mm i.d. × 0.10 ␮m film thickness) column was used for separation of the PBDEs, following introduction of 2 ␮L sample volumes into the cool on-column injection system which was set to track the oven temperature. The initial temperature was set to 80 ◦ C and held for 1 min,

58

R.G. Berger et al. / Toxicology 320 (2014) 56–66

increased to 225 ◦ C at a rate of 32 ◦ C min−1 , followed by an increase at a rate of 3 ◦ C min−1 to 230 ◦ C and held for 1 min. The final increase was set at a rate of 40 ◦ C min−1 to 315 ◦ C and held for 7 min. Helium was used as the carrier gas set at constant flow (1.2 mL min−1 ) for all PBDE analyses and the head pressure initially was 27 kPa. The mass spectrometer was operating in electron impact ionization mode and the source temperature was 250 ◦ C. HBCD analysis was performed using an Agilent 1100 (Agilent Technologies, Mississauga, ON, Canada) high pressure liquid chromatograph coupled to a Quattro Ultima triple quadrupole MS/MS (Waters Corporation, Milford, MA) with electrospray ionization in the negative ion detection mode using a Phenomenex Kinetex C18 column (2.1 mm × 150 mm, 2.6 ␮m) (Phenomenex, Torrance, CA). Water (mobile phase A) and acetonitrile:methanol (2:1) (mobile phase B) were used to separate the HBCD isomers and the gradient was as follows: 40% mobile phase B for 1 min, 40–80% phase B for 4 min, 80–90% phase B for 13 min, then returned to 40% phase B over 0.5 min where it remained until 18 min. The flow rate was maintained at 0.22 mL min−1 and the column temperature at 30 ◦ C, to completely resolve the deuterated HBCD analogs from the 13 Clabeled and native analogs. The capillary and cone voltage were −3.0 kV and 35 V, respectively. The source temperature and desolvation temperature were 140 ◦ C and 350 ◦ C, respectively. Cone gas flow and desolvation gas flow were 77 L h−1 and 723 L h−1 , respectively. Argon was the collision gas set at 5 × 10−3 mbar and resolution was established at 90% valley at base for both quadrupole analyzers. Dwell times were set to 5 ms. All reported sample concentrations were corrected for recoveries using isotope dilution. Two reagent blanks, an aliquot of corn oil known to be free of PBDEs and HBCD fortified with known concentrations of these analytes and a reference material of fish certified to contain known amounts of PBDEs were included with each set of samples prepared for analysis. Laboratory background was accounted for by removing the average of the PBDE and HBCD concentrations observed in the two reagent blank samples for each set of samples analyzed. Average PBDE surrogate recoveries ranged from 71% to 81% (PBDE 183 and PBDE 47, respectively) and 13 C HBCD recoveries ranged from 84% to 97% (␣- and ␤-HBCD, respectively) in the diet samples. PBDE concentrations were within the expected concentration range in the reference material tested and the average oil spike recovery was 100% for PBDEs (86–109%) and 101% for HBCD (100–102%) in the fortified oil.

respective diets until euthanasia on GD 20. Throughout the premating period, dam physical examinations were undertaken and body weight and food intake were monitored once a week. Following mating, body weight and food intake were assessed on GDs 3, 8, 13, 16 and 20. 2.4. Tissue collection Dams were euthanized by CO2 asphyxiation followed by exsanguination via cardiac puncture on GD 20. The sequence in which dams from each treatment group were euthanized was random and investigators were blinded as to the treatment group. Liver, kidneys, spleen, thymus, heart, lung, uterus and ovaries were removed and weighed. Uterine horns were inspected for implantation and resorption sites. The trachea, including the thyroid gland, was removed gently and fixed in neutral buffered formalin for morphological analysis. Whole blood was transferred to a Vacutainer SST (BD Biosciences Canada, Mississauga, ON, Canada), allowed to clot for 30 min at room temperature, then held on ice for no longer than 4 h until centrifugation at 1300 × g for 20 min. Serum was aliquoted and stored at −80 ◦ C. 2.5. Thyroid gland histology Fixed thyroid glands, in situ on the trachea, were embedded in paraffin, sectioned (5 ␮m) and stained with periodic acid Schiff. The height of the thyroid gland epithelium was measured using digital analysis in sections known to be in the interior of the gland. Epithelial height measurements were taken from at least 5 separate follicles per image, selected using an overlaid, numbered grid and 5 randomly generated numbers; at least 10 images per animal were analyzed. All measurements were made by a single observer blinded to the animal treatment (n = 10/group). 2.6. Serum measures Calcium, magnesium, phosphorus, blood urea nitrogen, uric acid, albumin, total protein, creatinine kinase, creatinine, triglycerides, cholesterol, glucose, alkaline phosphatase (ALP) and alanine aminotransferase (ALT) were measured in serum with an ABX Pentra 400 clinical chemistry analyzer (Horiba ABX, Montpellier, France). Total serum thyroxine (T4) was assayed using commercial RIA kits following the manufacturer’s protocol (MP Biomedicals, Lachine, QC, Canada).

2.3. Animals and treatment

2.7. Fetal examinations

All animal studies were conducted in accordance with the procedures and principles outlined in the Guide to the Care and Use of Experimental Animals prepared by the Canadian Council on Animal Care (McGill Animal Research Centre protocol 5862). Virgin female Sprague-Dawley rats (200–250 g) were obtained from Charles River (St-Constant, QC, Canada). Animals were individually housed at the Animal Resources Centre of McGill University in rooms maintained at 20 ◦ C on a 12 h:12 h light:dark cycle. Food and water were provided ad libitum. Following one week of acclimation to the control diet, female rats were randomly assigned to one of four experimental conditions (n = 35–38/group) and fed a diet supplemented with 0, 0.75, 250 or 750 mg BFR mixture/kg for two to four weeks prior to mating. During this exposure period, estrus cyclicity was assessed daily by analyzing vaginal cytology, as previously described (Goldman et al., 2007). After this premating exposure, females in proestrus were caged with proven breeder male Sprague-Dawley rats overnight. The next morning was designated as gestation day (GD) 0 if spermatozoa were detected in the vaginal smear. Following insemination, females continued on their

The position of fetuses in the uterine horns was recorded prior to their excision from the yolk sac and separation from the placenta. Placental weights and diameters were recorded. All live fetuses were evaluated for gender, body weight, crown-rump length and anogenital distance index (AGDI), calculated by dividing the anogenital distance (mm) by the cubed root of the body mass (Gallavan et al., 1999). All fetuses were examined for external physical malformations (n = 255, 365, 285 and 283 for the control, 0.06, 20 and 60 mg/kg/day BFR doses, respectively); an average of two fetuses/litter (n = 18, 41, 26, and 38 for the control, 0.06, 20 and 60 mg/kg/day BFR doses, respectively) were prepared for the evaluation of skeletal development. Accordingly, treatment-related effects on external morphology were analyzed on the basis of the proportion of litters with affected fetuses whereas skeletal evaluations were based on the proportion of individual fetuses affected. Fetuses were euthanized by hypothermia prior to fixation and storage in 95% ethanol. The preparation and staining methods were adapted from Whitaker and Dix (1979) with minor alterations. Briefly, at the time of processing, fixed fetuses were immersed in

R.G. Berger et al. / Toxicology 320 (2014) 56–66

59

Table 1 Dietary BFR content. Congener (␮g/g)

BDE-15 BDE-17 BDE-28 BDE-37 BDE-47 BDE-66 BDE-71 BDE-75 BDE-77 BDE-85 BDE-99 BDE-100 BDE-119 BDE-126 BDE-138 BDE-153 BDE-154 BDE-160 BDE-181 BDE-183 BDE-190 BDE-205 BDE-209 HBCD-␣ HBCD-␤ HBCD-␥

 PBDE (␮g/g)  HBCD (␮g/g) 

% PBDE nominal dose Number of samples

Nominal BFR mixture dose (mg/kg/day) Control (0)

0.06

20

60

NDa ND 0.0001 ± 0.00001 ND 0.001 ± 0.0002 ND ND ND ND ND 0.002 ± 0.0003 0.001 ± 0.0001 ND ND ND 0.003 ± 0.0015 0.001 ± 0.00001 ND ND ND ND ND ND 0.0009 ± 0.0003 0.0001 ± 4 × 10−6 0.0003 ± 0.0006

ND ND 0.0009 ± 0.00003 ND 0.098 ± 0.0013 0.001 ± 0.00004 ND ND ND 0.006 ± 0.0001 0.124 ± 0.0014 0.031 ± 0.0003 ND ND 0.002 ± 0.0004 0.011 ± 0.0002 0.011 ± 0.0002 ND ND 0.001 ± 0.00004 ND ND 0.362 ± 0.0141 0.018 ± 0.0003 0.001 ± 0.00005 0.0012 ± 0.00007

ND ND 0.195 ± 0.0281 ND 32.99 ± 3.86 0.357 ± 0.0391 ND ND ND 1.331 ± 0.124 44.65 ± 4.95 11.19 ± 1.26 ND ND 0.475 ± 0.051 3.62 ± 0.388 3.78 ± 0.416 ND ND 0.506 ± 0.0798 ND ND 145.77 ± 11.47 0.304 ± 0.778 0.381 ± 0.252 0.468 ± 0.0519

ND ND 0.445 ± 0.0109 ND 100.35 ± 1.53 1.135 ± 0.027 ND ND ND 3.360 ± 0.102 136.19 ± 3.37 32.94 ± 0.86 ND ND 1.662 ± 0.048 10.66 ± 0.16 11.36 ± 0.20 ND ND 1.230 ± 0.051 ND ND 361.90 ± 6.27 19.43 ± 0.32 1.194 ± 0.013 1.486 ± 0.03

0.015 ± 0.001

0.647 ± 0.033

245.68 ± 12.14

661.23 ± 33.19

0.001 ± 0.0002

0.020 ± 0.006

7.15 ± 1.96

22.11 ± 6.03

NA 24

88.87 18

101.13 18

91.11 18

Values are expressed as mean ± SEM. a PBDE limits of detection ranged from 0.03 ng g−1 to 6.5 ng g−1 (PBDE 15 and 209, respectively) and HBCD limits of detection ranged from 8.9 to 9.1 ng g−1 (␤- and ␣-HBCD, respectively).

a water bath (50 ◦ C) for 5 min, skinned, and cervical and dorsal muscles were gently removed. Fetuses were placed in the alcian blue/alizarin red stain solution for 24 h at 37 ◦ C, followed by 1% potassium hydroxide for 48 h, with fresh 1% potassium hydroxide added at 24 h, and then 0.5% potassium hydroxide for an additional 24 h. The potassium hydroxide was then replaced with a 2:2:1 solution consisting of two parts 70% ethanol, two parts glycerine, and one part benzyl alcohol. After 24 h, stained skeletons were placed in a 1:1 solution (70% ethanol:glycerine) for evaluation and storage. The skull, sternebrae, vertebrae, ribs, pectoral and pelvic girdles, fore and hind limbs and paws were examined. 2.8. Statistical analyses Data were analyzed by one-way ANOVA followed by post hoc Dunnett’s test to compare means to control. When tests for assumptions of homogeneity of variance and normality failed, data were log transformed and retested. When homoscedasticity and normality were still not satisfied after this transformation, data were retested using Kruskal–Wallis (K–W) ANOVA on ranks. Analyses of external (n = litter) and skeletal (n = fetus) malformations were conducted using a 1-tail Chi-Square test to compare the proportions of affected fetuses. The level of significance was p < 0.05. 3. Results 3.1. BFR contents of the diets Low levels of BDE-28, -47, -99, -100, -153 and -154 and the HBCD isomers were found in the control diet (Table 1). The relative levels

of BDE-47, -99, -100, -153, -154, and -209 did not differ across the experimental groups; they were determined to be approximately 15%, 19%, 5%, 1.6% 1.7% and 57%, respectively, of the total mixture and were consistent throughout the experimental period. These congeners accounted for 99% of the PBDEs measured in the dietary mixture. Values of total PBDE were 88–101% of the nominal doses (Table 1). 3.2. Diet consumption during treatment Ingestion of the BFR mixture during the exposure period did not alter dam food consumption or weight gain (Table 2). Since the BFR dietary mixture remained constant, there was a decrease in the average BFR dose ingested with increased weight gain during gestation (Table 2). Thus, BFR ingestion was calculated to be on average 0.05, 17.4 and 47.5 mg BFR/kg/day or 74–86% of target doses for the three experimental conditions. 3.3. Effects of BFR exposure on dams Dam survival and appearance were not affected by ingestion of the BFR mixture. BFR exposure did not affect absolute tissue weights (Supplementary data Table S1); however, it did result in a significant increase in relative liver and kidney weights at the highest dose compared to controls (Table 3). Analysis of variance showed a significant effect of BFR exposure on serum T4 levels (p = 0.047) although post hoc tests did not indicate a significant difference from controls (Table 4). No significant differences in the inner thyroid epithelial follicle height were observed among the treatment groups (control: 7.9 ± 0.4 ␮m; low dose: 7.1 ± 0.2 ␮m;

60

R.G. Berger et al. / Toxicology 320 (2014) 56–66

Table 2 Consumption of the BFR diet during gestation. Parameter

Gestational day

Nominal BFR mixture dose (mg/kg/day) 0.06a

20

60

24.25 ± 0.81 (13) 27.63 ± 0.96 (13) 27.70 ± 1.16 (14) 26.27 ± 0.97 (13) 25.48 ± 1.11 (16)

22.91 ± 0.84 (19) 25.53 ± 1.04 (18) 27.76 ± 0.84 (22) 25.77 ± 1.00 (17) 25.09 ± 1.05 (22)

25.32 ± 0.96 (14) 27.95 ± 1.22 (14) 28.23 ± 0.84 (17) 24.29 ± 0.60 (13) 24.03 ± 0.76 (15)

23.32 ± 0.69 (17) 28.69 ± 1.26 (16) 25.82 ± 1.18 (17) 25.81 ± 0.89 (13) 24.33 ± 0.88 (17)

332.31 ± 5.10 (14) 358.50 ± 5.46 (14) 390.70 ± 5.88 (13) 413.77 ± 5.10 (16) 462.41 ± 7.51 (16)

330.37 ± 5.83 (19) 350.13 ± 5.87 (21) 383.55 ± 5.95 (19) 402.80 ± 6.40 (20) 447.43 ± 7.16 (21)

333.95 ± 7.57 (14) 354.06 ± 7.77 (16) 382.30 ± 8.33 (15) 399.49 ± 8.85 (12) 446.74 ± 7.49 (17)

320.73 ± 6.38 (16) 341.12 ± 6.89 (17) 370.55 ± 9.17 (15) 391.19 ± 8.14 (16) 441.81 ± 8.72 (17)

Control (0)

Food consumption (g/day)

0–3 3–8 8–13 13–16 16–20

Gestational weight (g)

3 8 13 16 20

Approximate BFR dose (mg BFR/kg BW)

3 8 13 16 20

0 0 0 0 0

0.046 0.049 0.048 0.043 0.037

19.17 19.96 18.67 15.37 13.60

49.68 57.47 47.62 45.08 37.64

Values are expressed as mean ± SEM of (n) observations per treatment. a The maximum measured content of house dust PBDEs is 192 ␮g per g dust in the studies that we chose as a basis for our BFR mixture (Allen et al., 2008). This represents a dose of 1.16 ␮g PBDE per kg/day if a toddler (16.5 kg) ingests 100 mg dust. When we apply allometric scaling (×6.9) to adjust for the difference in body weight to surface area between humans and rats, our calculated rat equivalent to human dose is 8.0 ␮g/kg/day. Thus, our low dose group represents an exposure that is approximately 5 times the estimated high exposure for toddlers. Table 3 Effects of BFR exposure on dam organ weights. Organ

n Liver Kidney Spleen Thymus Heart Lung

Nominal BFR mixture treatment dose (mg/kg/day) Control (0)

0.06

20

60

15 34.2 ± 0.9 2.15 ± 0.04 1.41 ± 0.06 0.83 ± 0.05 2.71 ± 0.06 2.71 ± 0.08

21 33.5 ± 0.7 2.23 ± 0.04 1.44 ± 0.05 0.72 ± 0.03 2.69 ± 0.03 2.66 ± 0.04

17 35.3 ± 0.8 2.30 ± 0.05 1.47 ± 0.08 0.73 ± 0.04 2.64 ± 0.04 2.82 ± 0.07

17 37.9 ± 0.7** 2.34 ± 0.04* 1.43 ± 0.04 0.79 ± 0.05 2.66 ± 0.04 2.77 ± 0.10

Values are expressed as mean tissue weight (g) normalized to body weight (kg) ± SEM. * p < 0.05 compared with control. ** p < 0.01 compared with control.

middle dose: 8.1 ± 0.3 ␮m; high dose: 9.0 ± 0.4 ␮m). BFR exposure also had minimal effects on the other serum parameters assessed in the dams (Table 4). In the high dose BFR treatment group total protein levels were significantly reduced compared to controls. Although ANOVA indicated an influence of BFR exposure on phosphorus (p = 0.029), total cholesterol (p = 0.012), and

triglycerides (p = 0.017), post hoc tests failed to reveal significant differences from controls. Neither estrous cycle length nor fertility measures were affected by BFR exposure (Table 5). Finally, BFR treatment did not affect the mean numbers of implantation and resorption sites, postimplantation loss, or the litter size.

Table 4 Effects of BFR exposure on dam serum biochemistry. Parameter

n Thyroxin (T4) (␮g/dL) Calcium (mg/dL) Magnesium (mg/dL) Phosphorus (mg/dL) Blood urea N (mg/dL) Uric acid (mg/dL) Albumin (g/dL) Total protein (g/dL) Creatinine kinase (U/L) Creatinine (mg/dL) Triglycerides (mg/dL) Cholesterol (mg/dL) Glucose (mg/dL) ALP (U/L) ALT (U/L)

Nominal BFR mixture dose (mg/kg/day) Control (0)

0.06

20

60

14 1.52 ± 0.12 12.38 ± 0.24 3.23 ± 0.07 10.40 ± 0.45 15.3 ± 0.7 3.49 ± 0.27 2.78 ± 0.04 6.52 ± 0.11 311 ± 33 0.53 ± 0.02 446 ± 52 102 ± 4 125 ± 6 89.2 ± 14.1 66.0 ± 3.2

21 1.59 ± 0.09 12.06 ± 0.21 3.19 ± 0.07 10.21 ± 0.63 14.6 ± 0.5 2.91 ± 0.21 2.72 ± 0.04 6.49 ± 0.19 319 ± 42 0.50 ± 0.02 463 ± 55 92 ± 3 119 ± 6 83.7 ± 9.1 61.3 ± 3.0

17 1.68 ± 0.13 11.92 ± 0.17 3.25 ± 0.06 8.84 ± 0.26 13.5 ± 0.6 3.18 ± 0.15 2.77 ± 0.05 6.21 ± 0.11 354 ± 35 0.52 ± 0.01 329 ± 34 95 ± 3 123 ± 3 111.4 ± 18.1 56.0 ± 2.5

17 1.30 ± 0.08 12.07 ± 0.22 3.07 ± 0.07 9.04 ± 0.20 14.1 ± 0.4 2.98 ± 0.20 2.72 ± 0.07 5.97 ± 0.15* 308 ± 31 0.52 ± 0.01 292 ± 17 109 ± 4 114 ± 5 94.8 ± 10.5 63.5 ± 3.3

n = no. of dams. Values are expressed as mean ± SEM. * p < 0.05 compared with control.

R.G. Berger et al. / Toxicology 320 (2014) 56–66

61

Table 5 Effects of BFR exposures on measures of dam fertility. Parameter

Estrous cycle length – days Mating indexa Fecundity indexb No. of implants per litter No. of resorptions per litter Postimplantation lossesc No. of dead fetuses per litter No. of live fetuses per litter

Nominal BFR mixture dose (mg/kg/day) Control (0)

0.06

20

60

4.8 ± 0.4 (31) 70.4 (27) 79.0 (15) 17.9 ± 0.4 (15) 0.9 ± 0.4 (15) 4.8 ± 1.8 (15) 0 (15) 17.0 ± 0.3 (15)

4.6 ± 0.3 (29) 85.2 (27) 91.3 (21) 18.4 ± 0.4 (21) 1.0 ± 0.3 (21) 5.4 ± 1.5 (21) 0.05 ± 0.05 (21) 17.4 ± 0.4 (21)

4.7 ± 0.2 (29) 75.0 (24) 94.4 (17) 18.4 ± 0.9 (17) 1.5 ± 0.4 (17) 8.2 ± 1.6 (17) 0.06 ± 0.06 (17) 16.8 ± 0.80 (17)

5.6 ± 0.6 (29) 70.8 (24) 100.0 (17) 17.7 ± 0.7 (17) 1.1 ± 0.3 (17) 6.8 ± 3.0 (17) 0 (17) 16.7 ± 0.9 (17)

Values are expressed as mean ± SEM with the total number of observations. a Mating index = (no. of sperm positive females/no. of mated females) × 100. b Fecundity index = (no. of pregnant females/no. of sperm positive females) × 100. c Postimplantation losses = [(no. of implants − no. of viable fetuses)/no. of implants] × 10.

3.4. Fetal developmental measures BFR exposure had no effect on placental weights (Table 6). Fetal body weights were not affected, although a significant decrease in crown rump length in female fetuses was observed in the 20 mg/kg/day BFR treatment group. Neither sex ratios nor the anogenital distance indices in male and female fetuses were affected. External fetal examinations revealed malpositioned, fused and spread digits, crooked paws and single observations of paw edema (control fetus) and a shortened tail (60 mg/kg/day fetus) in control and BFR exposed fetuses (Fig. 1A–C). Fetuses with malpositioned or fused digits were observed in 25–52% of the litters (Fig. 1D). Analysis of the proportion of litters with fetuses with malpositioned or fused digits revealed a significant increase above control only in the low dose treatment group (Fig. 1D). Among BFR exposed fetuses, 80–84% of digit defects were found in the hind paws, compared to 67% in control fetuses (data not shown). A number of skeletal variations, including offset ossification of sternebrae, unossified sternebrae and vertebrae, and incompletely fused ossified centrums, were observed (Fig. 2A–D). A significant increase in the incidence of fetuses with any skeletal variation was found in both the middle (p = 0.018) and high (p = 0.024) doses (Fig. 2E). The mean numbers of affected fetuses per litter did not differ between control and treated groups; however, affected fetuses were found in multiple litters (Supplementary data, Fig. S1A). A decrease or lack of ossification in the sternebrae was the most sensitive skeletal variation observed after BFR exposure; the incidence of fetuses with unossified sternebrae was significantly increased in all BFR treatment groups (Table 7, Supplementary data, Fig. S1B). There were no significant changes in ossification or cartilage development in other areas of the skeleton, including the skull, phalanges, carpals or metacarpals. Skeletal analysis of the paws revealed no significant changes in ossification or cartilage development (Supplementary data, Fig. S2). Supplementary Figs. SI and SII related to this article found, in the online version, at http://dx.doi.org/10.1016/j.tox.2014.03.005. 4. Discussion The current study demonstrated that exposure during pregnancy to environmentally relevant complex BFR mixture, with relative congener levels similar to those found in household dust, significantly affected fetal development. This is the first investigation to report that exposure to BFRs increased fetal digit malformations. In addition, these data are the first to demonstrate an increase in the incidence of fetal skeletal variations in the absence of any indications of maternal toxicity following exposure to an environmentally relevant BFR dose (0.06 mg/kg/day).

Effects on any parameter of maternal health were observed only in dams exposed to the high BFR dose (60 mg/kg/day); these dams had significant liver and kidney enlargement and a reduction in serum protein compared to controls. Increases in liver and kidney weights are indicative of toxicity; however, there were no changes in serum biomarkers that are typically associated with liver and kidney damage. Our findings are consistent with a previous investigation in which increased liver weight and drug metabolizing enzyme activity were observed in GD 20 dams following exposure to DE-71 (30 mg/kg/day) (Zhou et al., 2002). Adult males fed the same complex BFR mixture used here showed an increase in both liver and kidney weights following exposure to BFRs at 20 mg/kg/day (Ernest et al., 2012). Others have also reported that exposures to PBDE congeners, as well as HBCD, are associated with increased kidney (van der Ven et al., 2006) and liver (Ema et al., 2008; Stoker et al., 2004; van der Ven et al., 2006, 2008) weights. We report here that exposure of female rats prior to conception and during gestation (for approximately 35 days) to 60 mg/kg/day of our BFR mixture did not affect either serum thyroxin (T4) concentrations or thyroid epithelial cell height in the dams. In contrast, exposure of adult males to this BFR mixture (20 mg/kg/day for 70 days) induced a significant decrease in serum T4 levels (Ernest et al., 2012). Serum T4 concentrations in the male rats, independently of exposure to BFRs, were approximately five times higher than those in the pregnant female rats. This observation is consistent with previous studies showing a significant reduction of thyroid hormone levels in pregnant rats near term (Calvo et al., 1990). A number of in vivo rodent studies have reported a decrease in serum T4 concentrations after exposure to BFRs, with results varying depending on the specific BFRs, dose, length of exposure and gender or strain of animals studied (reviewed by Legler, 2008). With respect to the effects of BFR exposures during pregnancy on thyroid hormone status, the exposure of Sprague-Dawley rats to DE-71 (60 or 120 mg/kg/day) from gestation day 6 to 19 lowered plasma T4 concentrations on gestation day 20 (Ellis-Hutchings et al., 2009), while exposure to DE-71, at a dose of 60 ␮g/kg/day, from gestation day 1.5 to lactation day 20 had no effect on serum T4 or T3 in the dams (Blake et al., 2011). Positive associations between TSH and the serum concentrations of lower-brominated PBDE congeners and OH-PBDEs were reported in second trimester pregnant women from California; however, the associations with free and total T4 were generally weak, inconsistent, and not statistically significant in this study (Zota et al., 2011). The lack of an effect of BFR exposure on serum T4 concentrations or thyroid gland histomorphology in this study suggests that the effects we observed are not ascribed to this mechanism. The majority of the anomalies observed in BFR-exposed fetuses consisted of malpositioned or fused digits in the hind paws. To the best of our knowledge, no previous investigation has observed a

62

R.G. Berger et al. / Toxicology 320 (2014) 56–66

Table 6 Effects of BFR exposure on fetal developmental measures on gestation day 20. Parameter

Sex

Total number of live pups (litters) Placental weight (mg) Sex ratio (M/F)

BFR mixture dose (mg/kg/day) 0

0.06

20

60

255 (15) 477.0 ± 10.5 1.07 ± 0.14

365 (21) 470.4 ± 9.57 1.22 ± 0.11

285 (17) 464.6 ± 14.2 1.12 ± 0.17

283 (17) 494.4 ± 21.7 1.18 ± 0.12

Body weight (g)

M F

3.64 ± 0.07 3.39 ± 0.07

3.50 ± 0.05 3.32 ± 0.05

3.45 ± 0.08 3.20 ± 0.07

3.57 ± 0.05 3.35 ± 0.05

Crown rump length (mm)

M F

36.7 ± 0.2 35.7 ± 0.2

36.3 ± 0.2 35.2 ± 0.1

36.0 ± 0.3 34.8 ± 0.2*

36.5 ± 0.2 35.4 ± 0.2

Anogenital distance indexa

M F

139.9 ± 1.5 62.1 ± 0.9

140.0 ± 0.9 60.4 ± 0.6

139.1 ± 1.1 61.1 ± 0.7

136.6 ± 1.4 60.6 ± 0.8

Values are expressed as mean ± SEM using the litter as the data point. a Anogenital distance index = (anogenital distance (mm)/cubed body weight (g)) × 100. * p < 0.05 compared with control.

similar phenotype. One previous investigation reported a significant increase in tail malformations in the F2 generation following a single administration of BDE-99 on GD 6 (Talsness et al., 2005). The digit defects we observed were not associated with changes

in ossification or cartilage development in the paws of exposed fetuses, implying that only soft tissues are affected. The proportion of litters with fetuses with malpositioned or fused digits was significantly increased in the low (0.06 mg/kg/day) dose treatment

Fig. 1. Effects of gestational BFR exposure on fetal external morphology. Representative pictures of fetal hindpaws on gestation day 20. (A) Normal paw (from the 60 mg BFR/kg/body weight/day treatment group). (B) Paw with fused digits (60 mg BFR/kg/body weight/day). (C) Paw with malpositioned digits (0.06 mg BFR/kg body weight/day). (D) Scatterplot depicting the numbers of fetuses with malpositioned or fused digits in individual litters; each data point represents a single litter (n = 16, 21, 17 and 18) for the control (0), 0.06, 20 and 60 mg/kg/day BFR groups, respectively. The proportion of litters with affected fetuses was significantly increased only in the 0.06 mg/kg/day BFR treatment group (0.06 mg/kg/day, p = 0.05; 20 mg/kg/day: p = 0.09; 60 mg/kg/day: p = 0.21).

R.G. Berger et al. / Toxicology 320 (2014) 56–66

63

Fig. 2. Effects of gestational BFR exposure on fetal skeletal development. Representative pictures of fetal skeletal ossification on gestation day 20 with all six sternebrae ossified (A, 0.06 mg BFR/kg body weight), two sternebrae sites unossified (B, 60 mg BFR/kg body weight), offset ossification of the sternebrae 1–4 and incomplete ossification of sternebrae 5 and 6 (C, 0.06 mg BFR/kg body weight), and an incompletely fused ossification vertebral centrum (D, 0.06 mg BFR/kg body weight). White arrows show sites of typical ossification, yellow circles mark areas with decreased or no ossification, and red arrows show markedly offset sternebrae (C) or an incompletely fused ossification vertebral centrum (D). (E) Proportion of fetuses with any type of skeletal variation, including lack of ossification of the sternebrae, markedly offset fusion of sternebrae and incomplete or lack of ossification of the fused vertebral centrum. (F) Proportion of fetuses with no ossification of the sixth sternebra. Bars represent the proportion of affected fetuses relative to the total number of fetuses examined (n = 18, 41, 26 and 38 for the control (0), 0.06, 20 and 60 doses, respectively). *p < 0.05, **p < 0.01 compared with control. (For interpretation of the references to color in this figure legend, the reader is referred to the web version of the article.)

Table 7 Effects of BFR exposure on fetal skeletal development. Variation

Number of fetuses examined Two or more unossified sternebrae Incomplete vertebral centrum fusion All spinal skeletal variations

BFR mixture dose (mg/kg/day) 0

0.06

20

60

18 1 (5.6) 2 (11.1) 2 (11.1)

41 11(26.8)* 8 (19.5) 9 (22.0)

26 8(30.8)* 6 (23.1) 6 (23.1)

38 12(31.6)* 8 (21.1) 11 (29.0)

Values are the number of observed variations per group with the percentage of the total number of examined fetuses in brackets. * p < 0.05 compared with control.

64

R.G. Berger et al. / Toxicology 320 (2014) 56–66

groups. Digit defects were also observed in the 20 mg/kg/day and 60 mg/kg/day treatment groups but statistical significance was not attained. An increase in the incidence of fetuses with skeletal variations was observed following exposure to the BFR mixture. The most common variation was a decrease in ossification in the sternebrae, particularly the sixth sternebra. Ossification in the fetus is sensitive to maternal toxicity (Beck, 1989); however, no indications of toxicity were observed in dams exposed to either the low or middle dose BFR regimens. In utero exposures to exogenous chemicals such as ethanol or tobacco smoke alter fetal skeletal ossification; however, these insults are most often associated with endpoints that were not observed here, such as decreased fetal survival and growth retardation (Chernoff, 1977; Seller and Bnait, 1995). Ossification of the rat skeleton typically commences around GDs 15–16 and is considered an indicator of maturity. At birth typically all (95–100%) of sternebrae are ossified. Variations in the ossification of the sternum of rodents are relatively common, with incomplete ossification one day before birth ranging from 1–7% for sternebrae 1–4 and 6, and up to 35% for sternebra 5 (Fritz and Hess, 1970). In the current investigation, the sixth sternebra was unossified in 35–40% of GD20 fetuses exposed to any BFR dose, compared to 5.5% in controls. These data suggest that exposure to BFRs may induce a delay in ossification. Delays in ossification may also be induced by fetal hypothyroidism (Capelo et al., 2008). Previously, single BFR congeners have been reported to induce an increase in the incidence of variations in ossification following in utero exposure. The administration of PBDE-99 (2 mg/kg/day) to Sprague-Dawley rats from GDs 6–19 increased skeletal defects, including incomplete ossification in the skull, caudal vertebrae and ribs, but did not affect the ossification of the sternebrae (Blanco et al., 2012). An earlier investigation in rabbits reported the delayed ossification of sternebrae following exposure to polybromodiphenyl oxide at a high dose (15 mg/kg/day) on GDs 7–20 (Breslin et al., 1989). This report, to the best of our knowledge, is the first to show a significant impact on fetal bone ossification following exposure to an environmentally relevant BFR mixture. Since reduced ossification may represent a delay in developmental timing it would be useful to examine the skeletons of postnatal pups to determine whether these effects are persistent (Carney and Kimmel, 2007). The current investigation shows that BFR exposure induced developmental alterations at environmentally relevant doses that have not been reported previously. This is also the first investigation to determine the effects of exposure to an environmentally relevant complex mixture of BFRs in an animal model. Previous experiments have shown that mixtures of endocrine disrupting chemicals may produce effects that are not observed when the components are administered individually (reviewed by Kortenkamp, 2007). For instance, exposure to a mixture of thyroid hormone disrupting chemicals, including dioxins and polychlorinated biphenyls, produced effects suggesting synergistic responses and the activation of different mechanisms than those observed when the individual components were administered (Crofton et al., 2005). Since there are low concentrations of polybrominated dibenzo-␳-dioxins (PBDDs) and dibenzofurans (PBDFs) in the commercial BFR mixtures (Hanari et al., 2006), these compounds may also contribute to the effects observed. Measurements of BFRs in serum from pregnant women (Buttke et al., 2013; Woodruff et al., 2011), umbilical cord blood (Frederiksen et al., 2010), and fetal liver and placental tissues (Doucet et al., 2009) demonstrate that humans are exposed to multiple BFR congeners and technical mixtures throughout gestation. Therefore, elucidation of the effects and potential dangers of exposure to complex, environmentally relevant, mixtures is critical since the impact of exposure to environmentally relevant BFR mixtures may differ from that of individual congeners.

In conclusion, we have shown evidence that exposure to the mixture of BFRs commonly found in North American household dust alters fetal development. The phenotypes observed, including malpositioned or fused digits and delayed ossification in the sternebrae, have not been reported previously in investigations of the effects of individual BFR congeners or technical mixtures. The absence of concurrent effects on measures of toxicity and pregnancy maintenance in the dams suggests that exposure to a wide range of BFR doses may have a direct impact on fetal development. Supplementary data Supplementary data Table S1 presents the absolute values for dam organ weights. Scatterplots of the incidence of skeletal variations are provided in Supplementary data Fig. S1. Images of the skeletal staining of paws from control and BFR exposed fetuses are shown in Supplementary data Fig. S2. Conflict of interest The authors declare that there are no conflicts of interest. Transparency document The Transparency document associated with this article can be found in the online version. Acknowledgements The authors acknowledge Mr. Donald Demers, Animal Resources Division Health Canada, for food pelleting and Dr. Chunwei Huang and Ms. Lydia Goff, McGill University, for their assistance. This research was supported by grant RHF100625 from the Canadian Institutes of Health Research (CIHR) Institute for Human Development, Child and Youth Health. RGB is the recipient of an NSERC Create award from the Réseau Québécois en Reproduction and PLCL of a Fonds de la Recherche Santé du Québec fellowship. BR and BFH are James McGill Professors. Appendix A. Supplementary data Supplementary material related to this article can be found, in the online version, at http://dx.doi.org/10.1016/j.tox.2014.03.005. References Aleksa, K., Carnevale, A., Goodyer, C., Koren, G., 2012. Detection of polybrominated biphenyl ethers (PBDEs) in pediatric hair as a tool for determining in utero exposure. Forensic Sci. Int. 218, 37–43. Allen, J.G., McClean, M.D., Stapleton, H.M., Webster, T.F., 2008. Critical factors in assessing exposure to PBDEs via house dust. Environ. Int. 34, 1085–1091. Beck, S.L., 1989. Prenatal ossification as an indicator of exposure to toxic agents. Teratology 40, 365–374. Blake, C.A., McCoy, G.L., Hui, Y.Y., LaVoie, H.A., 2011. Perinatal exposure to low-dose DE-71 increases serum thyroid hormones and gonadal osteopontin gene expression. Exp. Biol. Med. (Maywood) 236 (4), 445–455. Blanco, J., Mulero, M., Domingo, J.L., Sanchez, D.J., 2012. Gestational exposure to BDE-99 produces toxicity through upregulation of CYP isoforms and ROS production in the fetal rat liver. Toxicol. Sci. 127, 296–302. Bondy, G.S., Lefebvre, D.E., Aziz, S., Cherry, W., Coady, L., Maclellan, E., Armstrong, C., Barker, M., Cooke, G., Gaertner, D., Arnold, D.L., Mehta, R., Rowsell, P.R., 2013. Toxicologic and immunologic effects of perinatal exposure to the brominated diphenyl ether (BDE) mixture DE-71 in the Sprague-Dawley rat. Environ Toxicol. 28, 215–228. Breslin, W.J., Kirk, H.D., Zimmer, M.A., 1989. Teratogenic evaluation of a polybromodiphenyl oxide mixture in New Zealand white rabbits following oral exposure. Fundam. Appl. Toxicol. 12, 151–157. Buttke, D.E., Wolkin, A., Stapleton, H.M., Miranda, M.L., 2013. Associations between serum levels of polybrominated diphenyl ether (PBDE) flame retardants and environmental and behavioral factors in pregnant women. J. Expo. Sci. Environ. Epidemiol. 23, 176–182.

R.G. Berger et al. / Toxicology 320 (2014) 56–66

65

Calvo, R., Obregon, M.J., De Ona, C.R., Ferreiro, B., Escobar Del Rey, F., Morreale De Johnson-Restrepo, B., Kannan, K., Rapaport, D.P., Rodan, B.D., 2005. Polybrominated Escobar, G., 1990. Thyroid hormone economy in pregnant rats near term: a diphenyl ethers and polychlorinated biphenyls in human adipose tissue from physiological animal model of nonthyroidal illness? Endocrinology 127, 10–16. New York. Environ. Sci. Technol. 39, 5177–5182. Camino, G., Costa, L., Luda di Cortemiglia, M.P., 1991. Overview of fire retardant Johnson, P.I., Stapleton, H.M., Sjodin, A., Meeker, J.D., 2010. Relationships between mehcanisms. Polym. Degrad. Stab. 98, 1662–1670. polybrominated diphenyl ether concentrations in house dust and serum. EnviCapelo, L.P., Beber, E.H., Huang, S.A., Zorn, T.M., Bianco, A.C., Gouveia, C.H., 2008. ron. Sci. Technol. 44, 5627–5632. Deiodinase-mediated thyroid hormone inactivation minimizes thyroid horJohnson, P.I., Altshul, L., Cramer, D.W., Missmer, S.A., Hauser, R., Meeker, J.D., 2012. Serum and follicular fluid concentrations of polybrominated diphenyl ethers mone signaling in the early development of fetal skeleton. Bone 43, 921–930. Carney, E.W., Kimmel, C.A., 2007. Interpretation of skeletal variations for human and in-vitro fertilization outcome. Environ. Int. 45, 9–14. risk assessment: delayed ossification and wavy ribs. Birth Defects Res. B: Dev. Kortenkamp, A., 2007. Ten years of mixing cocktails: a review of combination effects Reprod. Toxicol. 80, 473–496. of endocrine-disrupting chemicals. Environ. Health Perspect. 115, 98–105. Carrizo, D., Grimalt, J.O., Ribas-Fito, N., Sunyer, J., Torrent, M., 2007. Influence of Legler, J., 2008. New insights into the endocrine disrupting effects of brominated breastfeeding in the accumulation of polybromodiphenyl ethers during the first flame retardants. Chemosphere 73, 216–222. Lilienthal, H., Hack, A., Roth-Härer, A., Grande, S.W., Talsness, C.E., 2006. Effects years of child growth. Environ. Sci. Technol. 41, 4907–4912. Ceccatelli, R., Faass, O., Schlumpf, M., Lichtensteiger, W., 2006. Gene expression and of developmental exposure to 2,2 ,4,4 ,5-pentabromodiphenyl ether (PBDE-99) estrogen sensitivity in rat uterus after developmental exposure to the polyon sex steroids, sexual development, and sexually dimorphic behavior in rats. brominated diphenylether PBDE 99 and PCB. Toxicology 220, 104–116. Environ. Health Perspect. 114, 194–201. Chernoff, G.F., 1977. The fetal alcohol syndrome in mice: an animal model. TeratolLiu, X., Zhan, H., Zeng, X., Zhang, C., Chen, D., 2012. The PBDE-209 exposure during ogy 15, 223–229. pregnancy and lactation impairs immune function in rats. Mediators Inflamm Crofton, K.M., Craft, E.S., Hedge, J.M., Gennings, C., Simmons, J.E., Carchman, R.A., 212, Article ID 692467. Marchitti, S.A., LaKind, J.S., Naiman, D.Q., Berlin, C.M., Kenneke, F.J., 2013. Improving Carter Jr., W.H., DeVito, M.J., 2005. Thyroid-hormone-disrupting chemicals: evidence for dose-dependent additivity or synergism. Environ. Health Perspect. infant exposure and health risk estimates: using serum data to predict poly113, 1549–1554. brominated diphenyl ether concentrations in breast milk. Envrion. Sci. Technol. Darnerud, P.O., Eriksen, G.S., Johannesson, T., Larsen, P.B., Viluksela, M., 2001. Poly47, 4787–4795. Mazdai, A., Dodder, N.G., Abernathy, M.P., Hites, R.A., Bigsby, R.M., 2003. Polybromibrominated diphenyl ethers: occurrence, dietary exposure, and toxicology. Environ. Health Perspect. 109 (Suppl. 1), 49–68. nated diphenyl ethers in maternal and fetal blood samples. Environ. Health Dodson, R.E., Perovich, L.J., Covaci, A., Van den Eede, N., Ionas, A.C., Dirtu Brody, Perspect. 111, 1249–1252. Meeker, J.D., Johnson, P.I., Camann, D., Hauser, R., 2009. Polybrominated diphenyl J.G., Rudel, R.A., 2012. After the PBDE phase-out: a broad suite of flame retardants in repeat house dust samples from California. Environ. Sci. Technol. 46, ether (PBDE) concentrations in house dust are related to hormone levels in men. 13056–13066. Sci. Total Environ. 407, 3425–3429. Doucet, J., Tague, B., Arnold, D.L., Cooke, G.M., Hayward, S., Goodyer, C.G., 2009. PerMeerts, I., Letcher, R.J., Hoving, S., Marsh, G., Bergman, A., Lemmen, J.G., van der Burg, sistent organic pollutant residues in human fetal liver and placenta from Greater B., Brouwer, A., 2001. In vitro estrogenicity of polybrominated diphenyl ethers, Montreal, Quebec: a longitudinal study from 1998 through 2006. Environ. Health hydroxylated PDBES, and polybrominated bisphenol A compounds. Environ. Perspect. 117, 605–610. Health Perspect. 109, 399–407. Dreiem, A., Okoniewski, R.J., Brosch, K.O., Miller, V.M., Seegal, R.F., 2010. PolychloPetro, E.M., Leroy, J.L., Covaci, A., Fransen, E., De Neubourg, D., Dirtu, A.C., De Pauw, I., rinated biphenyls and polybrominated diphenyl ethers alter striatal dopamine Bols, P.E., 2012. Endocrine-disrupting chemicals in human follicular fluid impair neurochemistry in synaptosomes from developing rats in an additive manner. in vitro oocyte developmental competence. Hum. Reprod. 27, 1025–1033. Rose, M., Bennett, D.H., Bergamn, A., Fängström, B., Pessah, I.N., Hertz-Picciotto, I., Toxicol Sci. 118, 150–159. Ellis-Hutchings, R.G., Cherr, G.N., Hanna, L.A., Keen, C.L., 2009. The effects of 2010. PBDEs in 2–5 year old children from California and associations with diet marginal maternal vitamin A status on penta-brominated diphenyl ether and indoor environment. Environ. Sci. Technol. 44, 2648–2653. Saegusa, Y, Fujimoto, H., Woo, G.H., Inoue, K., Takahashi, M., Mitsumori, K., Hirose, M., mixture-induced alterations in maternal and conceptal vitamin A and fetal development in the Sprague Dawley rat. Birth Defects Res. B: Dev. Reprod. Nishikawa, A., Shibutani, M., 2009. Developmental toxicity of brominated flame Toxicol. 86, 48–57. retardants, tetrabromobisphenol A and 1,2,5,6,9,10-hexabromocyclododecane, Environment Canada. 2011. Available: www.ec.gc.ca/ceparegistry/the act/schedules 1.cfm in rat offspring after maternal exposure from mid-gestation through lactation. Ema, M., Fujii, S., Hirata-Koizumi, M., Matsumoto, M., 2008. Two-generation reproReprod Toxicol. 28, 456–467. Seller, M.J., Bnait, K.S., 1995. Effects of tobacco smoke inhalation on the developing ductive toxicity study of the flame retardant hexabromocyclododecane in rats. Reprod. Toxicol. 25, 335–351. mouse embryo and fetus. Reprod. Toxicol. 9, 449–459. Ernest, S.R., Wade, M.G., Lalancette, C., Ma, Y.Q., Berger, R.G., Robaire, B., Hales, B.F., Siddique, S., Xian, Q., Abdelouahab, N., Takser, L., Phillips, S.P., Feng, Y.L., Wang, B., 2012. Effects of chronic exposure to an environmentally relevant mixture of Zhu, J., 2012. Levels of dechlorane plus and polybrominated diphenylethers in brominated flame retardants on the reproductive and thyroid system in adult human milk in two Canadian cities. Environ. Int. 39, 50–55. Sjödin, A., Jones, R.S., Focant, J.F., Lapeza, C., Wang, R.Y., McGahee, E.E., 2004. male rats. Toxicol. Sci. 127, 496–507. Eskenazi, B., Chevrier, J., Rauch, S.A., Kogut, K., Harley, K.G., Johnson, C., Trujillo, C., Retrospective time-trend study of polybrominated diphenyl ether and polySjödin, A., Bradman, A., 2013. In utero and childhood polybrominated diphenyl brominated and polychlorinated biphenyl levels in human serum from the ether (PBDE) exposures and neurodevelopment in the CHAMACOS study. EnviUnited States. Environ. Health Perspect. 112, 654–658. ron. Health Perspect. 121 (2), 257–262. Stapleton, H.M., Dodder, N.G., Offenberg, J.H., Schantz, M.M., Wise, S.A., 2005. PolyFowles, J.R., Fairbrother, A., Baecher-Steppan, L., Kerkvliet, N.I., 1994. Immunologic brominated diphenyl ethers in house dust and clothes dryer lint. Environ. Sci. and endocrine effects of the flame-retardant pentabromodiphenyl ether (DE-71) Technol. 39, 925–931. Stapleton, H.M., Allen, J.G., Kelly, S.M., Konstantinov, A., Klosterhaus, S., Watkins, in C57BL/6J mice. Toxicology 86, 49–61. Fraser, A.J., Webster, T.F., McClean, M.D., 2009. Diet contributes significantly to the D., McClean, M.D., Webster, T.F., 2008. Alternate and new brominated flame body burden of PBDEs in the general US population. Environ. Health Perspect. retardants detected in U.S. house dust. Environ. Sci. Technol. 42, 6910–6916. 117, 1520–1525. Stapleton, H.M., Eagle, S., Sjödin, A., Webster, T.F., 2012. Serum PBDEs in a North Frederiksen, M., Thomsen, C., Frøshaug, M., Vorkamp, K., Thomsen, M., Becher, G., Carolina toddler cohort: associations with handwipes, house dust, and socioeKnudsen, L.E., 2010. Polybrominated diphenyl ethers in paired samples of materconomic variables. Environ. Health Perspect. 120, 1049–1054. Stoker, T.E., Laws, S.C., Crofton, K.M., Hedge, J.M., Ferrell, J.M., Cooper, R.L., 2004. nal and umbilical cord blood plasma and associations with house dust in a Danish Assessment of DE-71, a commercial polybrominated diphenyl ether (PBDE) mixcohort. Int. J. Hyg. Environ. Health 213, 233–242. Fritz, H., Hess, R., 1970. Ossification of the rat and mouse skeleton in the perinatal ture, in the EDSP male and female pubertal protocols. Toxicol. Sci. 78, 144–155. Stoker, T.E., Cooper, R.L., Lambright, C.S., Wilson, V.S., Furr, J., Gray, L.E., 2005. In vivo period. Teratology 3, 331–337. Gallavan Jr., R.H., Holson, J.F., Stump, D.G., Knapp, J.F., Reynolds, V.L., 1999. Interpretand in vitro anti-androgenic effects of DE-71, a commercial polybrominated ing the toxicologic significance of alterations in anogenital distance: potential diphenyl ether (PBDE) mixture. Toxicol. Appl. Pharmacol. 207, 78–88. Suvorov, A., Battista, M.C., Takser, L., 2009. Perinatal exposure to low-dose for confounding effects of progeny body weights. Reprod. Toxicol. 13, 383–390. Goldman, J.M., Murr, A.S., Cooper, R.L., 2007. The rodent estrous cycle: characteriza2,2 ,4,4 -tetrabromodiphenyl ether affects growth in rat offspring: what is the tion of vaginal cytology and its utility in toxicological studies. Birth Defects Res. role of IGF-1? Toxicology 260, 126–131. B: Dev. Reprod. Toxicol. 80, 84–97. Suvorov, A., Takser, L., 2010. Global gene expression analysis in the livers of rat Haave, M., Folven, K.I., Carrol, T., Glover, C., Heegaard, E., Brattelid, T., Hogstrand, C., offspring perinatally exposed to low doses of 2,2 ,4,4 -tetrabomodiphenyl ether. Lundebye, A.K., 2011. Cerebral gene expression and neurobehavioural developEnviron. Health Perspect. 118, 97–102. Ta, T.A., Koenig, C.M., Golub, M.S., Pessah, I.N., Qi, L., Aronov, P.A., Berman, R.F., 2011. ment after perinatal exposure to an environmentally relevant polybrominated Bioaccumulation and behavioral effects of 2,2’,4,4’-tetrabromodiphenyl ether diphenylether (BDE47). Cell Biol. Toxicol. 27, 343–361. Hallgren, S., Sinjari, T., Hakansson, H., Darnerud, P.O., 2001. Effects of polybromi(BDE-47) in perinatally exposed mice. Neurotoxicol Teratol. 33, 393–404. Talsness, C.E., Kuriyama, S.N., Sterner-Kock, A., Schnitker, P., Grande, S.W., Shakibaei, nated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) on thyroid hormone and vitamin a levels in rats and mice. Arch. Toxicol. 75, 200–208. M., Andrade, A., Grote, K., Chahoud, I., 2008. In utero and lactational exposures to Hanari, N., Kannan, K., Miyake, Y., Okazawa, T., Kodavanti, P.R.S., Aldous, K.M., low doses of polybrominated diphenyl ether-47 alter the reproductive system Yamashita, N., 2006. Occurrence of polybrominated dibenzo-rho-dioxins, and and thyroid gland of female rat offspring. Environ Health Perspect. 116, 308–314. Talsness, C.E., Shakibaei, M., Kuriyama, S.N., Grande, S.W., Sterner-Kock, A., polybrominated dibenzofurans as impurities in commercial polybrominated biphenyl ether mixtures. Environ. Sci. Technol. 40, 4400–4405. Schnitker, P., de Souza, C., Grote, K., Chahoud, I., 2005. Ultrastructural changes observed in rat ovaries following in utero and lactational exposure to low doses Hites, R.A., 2004. Polybrominated diphenyl ethers in the environment and in people: a meta-analysis of concentrations. Environ. Sci. Technol. 38, 945–956. of a polybrominated flame retardant. Toxicol. Lett. 157, 189–202.

66

R.G. Berger et al. / Toxicology 320 (2014) 56–66

Toms, L.-M.L., Hearn, L., Kennedy, K., Harden, F., Bartkow, M., Temme, C., Jochen, F., Mueller, J.F., 2009. Concentrations of polybrominated diphenyl ethers (PBDEs) in matched samples of human milk, dust and indoor air. Environ. Int. 35, 864–869. Tullo, A., 2003. Great lakes to phase out two flame retardants. Chem. Eng. News 81, 13. van der Ven, L.T., Verhoef, A., van de Kuil, T., Slob, W., Leonards, P.E., Visser, T.J., Hamers, T., Herlin, M., Håkansson, H., Olausson, H., Piersma, A.H., Vos, J.G., 2006. A 28-day oral dose toxicity study enhanced to detect endocrine effects of hexabromocyclododecane in Wistar rats. Toxicol. Sci. 94, 281–292. van der Ven, L.T., van de Kuil, T., Leonards, P.E., Slob, W., Canton, R.F., Germer, S., Visser, T.J., Litens, S., Håkansson, H., Schrenk, D., van den Berg, M., Piersma, A.H., Vos, J.G., Opperhuizen, A., 2008. A 28-day oral dose toxicity study in Wistar rats enhanced to detect endocrine effects of decabromodiphenyl ether (decaBDE). Toxicol. Lett. 179, 6–14. van der Ven, L.T., van de Kuil, T., Leonards, P.E., Slob, W., Lilienthal, H., Litens, S., Herlin, M., Håkansson, H., Cantón, R.F., van den Berg, M., Visser, T.J., van Loveren, H., Vos, J.G., Piersma, A.H., 2009. Endocrine effects of hexabromocyclododecane (HBCD) in a one-generation reproduction study in Wistar rats. Toxicol. Lett. 185, 51–62. Whitaker, J., Dix, K.M., 1979. Double staining technique for rat foetus skeletons in teratological studies. Lab. Anim. 13, 309–310.

Woodruff, T.J., Zota, A.R., Schwartz, J.M., 2011. Environmental chemicals in pregnant women in the United States: NHANES 2003–2004. Environ. Health Perspect. 119, 878–885. Wu, N., Herrmann, T., Paepke, O., Tickner, J., Hale, R., Harvey, L.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, 1584–1589. Zhang, Z., Zhang, X., Sun, Z., Dong, H., Qiu, L., Gu, J., Zhou, J., Wang, X., Wang, S.L., 2013. Cytochrome P450 3A1 mediates 2,2’,4,4’-tetrabromodiphenyl ether-induced reduction of spermatogenesis in adult rats. PLoS One. 8 (6), e66301. Zhou, T., Taylor, M.M., DeVito, M.J., Crofton, K.M., 2002. Developmental exposure to brominated diphenyl ethers results in thyroid hormone disruption. Toxicol. Sci. 66, 105–116. Zota, A.R., Linderholm, L., Park, J.S., Petreas, M., Guo, T., Privalsky, M.L., Zoeller, R.T., Woodruff, T.J., 2013. Temporal comparison of PBDEs, OH-PBDEs, PCBs, and OH-PCBs in the serum of second trimester pregnant women recruited from San Francisco General Hospital, California. Environ. Sci. Technol. (Epub ahead of print). Zota, A.R., Park, J.S., Wang, Y., Petreas, M., Zoeller, R.T., Woodruff, T.J., 2011. Polybrominated diphenyl ethers, hydroxylated polybrominated diphenyl ethers, and measures of thyroid function in second trimester pregnant women in California. Environ. Sci. Technol. 45 (18), 7896–7905.