Is decabromodiphenyl ether (BDE-209) a developmental neurotoxicant?

NeuroToxicology 32 (2011) 9–24

Contents lists available at ScienceDirect

NeuroToxicology

Review

Is decabromodiphenyl ether (BDE-209) a developmental neurotoxicant? Lucio G. Costa a,b,*, Gennaro Giordano a a b

Department of Environmental and Occupational Health Sciences, University of Washington, Seattle, WA, USA Department of Human Anatomy, Pharmacology and Forensic Science, University of Parma Medical School, Parma, Italy

A R T I C L E I N F O

A B S T R A C T

Article history: Received 4 November 2010 Accepted 13 December 2010 Available online 21 December 2010

Polybrominated diphenyl ether (PBDE) flame retardants have become ubiquitous environmental pollutants. The relatively higher body burden in toddlers and children has raised concern for their potential developmental neurotoxicity, which has been suggested by animal studies, in vitro experiments, and recent human epidemiological evidence. While lower brominated PBDEs have been banned in several countries, the fully brominated decaBDE (BDE-209) is still utilized, though manufacturers will discontinue production in the U.S.A. in 2013. The recent decision by the U.S. Environmental Protection Agency to base the reference dose (RfD) for BDE-209 on a developmental neurotoxicity study has generated some controversy. Because of its bulky configuration, BDE-209 is poorly absorbed and does not easily penetrate the cell wall. Its acute and chronic toxicities are relatively low, with the liver and the thyroid as the primary targets, though there is some evidence of carcinogenicity. A few animal studies have indicated that BDE209 may cause developmental neurotoxicity, affecting motor and cognitive domains, as seen for other PBDEs. Limited in vivo and in vitro studies have also evidenced effects of BDE-209 on thyroid hormone homeostasis and direct effects on nervous cells, again similar to what found with other lower brominated PBDEs. In contrast, a recent developmental neurotoxicity study, carried out according to international guidelines, has provided no evidence of adverse effects on neurodevelopment, and this should be considered in a future re-evaluation of BDE-209. While estimated exposure to BDE-209 in children is believed to be several orders of magnitude below the most conservative RfD proposed by the USEPA, questions remain on the extent and relevance of BDE-209 metabolism to lower brominated PBDEs in the environment and in humans. ß 2010 Elsevier Inc. All rights reserved.

Keywords: Polybrominated diphenyl ethers DecaBDE BDE-209 Developmental neurotoxicity Reference dose

Contents 1. 2. 3. 4.

5. 6. 7.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Developmental neurotoxicity of tetra-, penta-, and hexa-BDEs. . . . . . . . . . . . . . . . . . . . . . 2.1. Possible mechanisms of developmental neurotoxicity of lower brominated PBDEs General toxicology of BDE-209 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. Toxicokinetic considerations for BDE-209 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Evidence of developmental neurotoxicity of BDE-209 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1. Animal studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2. In vitro studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Occurrence and body burden of BDE-209. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Derivation of an RfD for BDE-209 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions and research needs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

. . . . . . . . . . . . .

.. 9 . 10 . 10 . 10 . 11 . 12 . 12 . 16 . 16 . 18 . 20 . 21 . 21

1. Introduction * Corresponding author at: Department of Environmental and Occupational Health Sciences, University of Washington, 4225 Roosevelt Way NE, Suite 100, Seattle, WA 98105, USA. Tel.: +1 206 543 2831; fax: +1 206 685 4696. E-mail address: [email protected] (L.G. Costa). 0161-813X/$ – see front matter ß 2010 Elsevier Inc. All rights reserved. doi:10.1016/j.neuro.2010.12.010

Polybrominated diphenyl ethers (PBDEs) are a class of brominated flame retardants consisting of 209 congeners, which have been commercialized as penta-, octa-, and deca-brominated

10

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

mixtures (Alaee et al., 2003). PentaBDE and octaBDE mixtures are no longer produced and commercialized in the European Union and the United States, though they are still used in other parts of the world. In contrast, decaBDE continues to be widely used, though some European countries (e.g. Norway) or states in the U.S.A. (e.g. Maine) have recently banned their use. The manufacturers have agreed to discontinue production of decaBDE in the U.S.A. as of 2013 (USEPA, 2010). PBDEs are additive flame retardants, i.e. they are not bound to the polymer like other flame retardants, and can leach from the products and find their way into the environment. In the past decade, PBDEs have become widespread environmental pollutants, having been detected in outdoor and indoor air, dust, sediments, soil, sludge, birds, fish, marine mammals and other mammals (deWit, 2002; Hites et al., 2004; Law et al., 2006; Chen and Hale, 2010; USEPA, 2010). Exposure of humans has also been documented by several studies (Gill et al., 2004; McDonald, 2005). Exposure occurs primarily through dust and through the diet. Body burden is higher, by at least an order of magnitude, in the U.S. population compared to Japan and the E.U. (Thomsen et al., 2002; Schecter et al., 2005a; Lorber, 2007; USEPA, 2010), though more recent data indicate high levels of exposures in fast developing nations such as India and China (J. Wang et al., 2010a). The highest body burden is found in infants and children, because of exposure through breast milk and household dust (Furst, 2006; Schecter et al., 2003; Fischer et al., 2006; Harrad et al., 2008a), and in occupationally exposed workers (Sjodin et al., 1999; Qu et al., 2007). PBDEs have also been found to cross the placenta, and levels in cord blood are comparable to those found in the mothers (Gomara et al., 2007). A recent document, the ‘‘San Antonio statement of brominated and chlorinated flame retardants’’, voices concern for the widespread contamination of these compounds, including PBDEs (DiGangi et al., 2010). 2. Developmental neurotoxicity of tetra-, penta-, and hexaBDEs The high body burden during early development has raised concerns about possible developmental toxicity and neurotoxicity of PBDEs. Several studies have indicated that PBDEs cause developmental neurotoxicity, and recent reviews have summarized the findings of animal and in vitro studies (Branchi et al., 2003; Birnbaum and Staskal, 2004; McDonald, 2005; Costa and Giordano, 2007; Costa et al., 2008; Williams and DeSesso, 2010). In vivo exposure to various PBDEs [BDE-47, BDE-99, BDE-153, DE-71 (a pentaBDE mixture)] given prenatally and/or postnatally by various routes of exposures (oral, i.p., s.c.) for different periods (from a single injection to repeated exposures) has been shown to cause a variety of neurobehavioral effects in developing and adult animals. Most affected were locomotor functions, with hyperactivity and decreased habituation, and cognitive behavior (measured in the Morris water maze and other tests). A complete discussion (and references) of most studies can be found in Costa and Giordano (2007) and Costa et al. (2008). More recent animal studies with BDE-47 (Suvorov et al., 2009; Abdelouahab et al., 2009), BDE-99 (Cheng et al., 2009), and DE-71 (Driscoll et al., 2009; Kodavanti et al., 2010) have confirmed the earlier findings. Limited evidence is also emerging of an association between PBDE exposure and neurodevelopmental toxicity in humans. A first study determined the presence of PBDEs in cord blood in a group of 297 newborns in Baltimore, MD; PBDEs were detected in 94% of the samples, and BDE-47 was found to be the dominant congener (median = 13.6 ng/g serum lipids, 51.2% of total PBDEs) (Herbstman et al., 2007). These results confirm those previously seen by other investigators (see Table 7 in Costa and Giordano, 2007), indicating that PBDEs can cross the placenta and reach the fetus. In the same cohort, a decrease in serum T4 levels was also reported

(Herbstman et al., 2008), differently from a previous report (Mazdai et al., 2003). However, the strongest correlations were between cord blood levels of T4 and PCBs, rather than with levels of PBDEs (Herbstman et al., 2008). In another study in the Netherlands, an association was found between blood PBDE levels in the mother at the 35th week of pregnancy and neurobehavioral development of the child up to age six (Roze et al., 2009). An association was found between PBDEs and motor function, cognition and behavior, but not with thyroid hormone levels (Roze et al., 2009). In an additional cohort in New York City, NY, consisting of 329 pregnancies, prenatal PBDE exposure (as indicated by cord blood PBDE levels) was associated with lower scores on tests of mental and physical development at the ages of 1–4 and 6 years (Herbstman et al., 2010). It has also been suggested that PBDEs may represent a potential risk factor for autism (Messer, 2010). However, results of a study presented in abstract form, indicate that children (2–5 years of age) with autism had actually lower blood levels of PBDEs than controls (Hertz-Picciotto, 2008). Thus, though information on potential developmental neurotoxic effects of PBDEs in humans is still very limited, initial evidence appears to suggest potential risks for neurobehavioral development. 2.1. Possible mechanisms of developmental neurotoxicity of lower brominated PBDEs Though the exact mechanisms of PBDEs’ developmental neurotoxicity are not known, it has been suggested that at least two, not mutually exclusive, modes of action, may exist, one mediated by an effect on thyroid hormones, the other due to a direct effect of PBDEs on brain cells (Costa and Giordano, 2007; Costa et al., 2008, 2010; Fonnum and Mariussen, 2009). PBDEs have been reported to decrease levels of thyroid hormones following developmental exposures (Zhou et al., 2001, 2002; Ellis-Hutchings et al., 2006; Branchi et al., 2002; Kuriyama et al., 2007). Given that thyroid hormones play a relevant role in brain development (Chan and Rovet, 2003), and that hypothyroidism results in various neuroanatomical and neurobehavioral alterations (Schalock et al., 1977; Zoeller and Crofton, 2005), such effect on thyroid hormones may contribute to PBDEs’ developmental neurotoxicity. However, behavioral and neurochemical alterations have been found upon administration of BDE-47, in the absence of any effect on thyroid hormones (Gee et al., 2008; Giordano and Costa, unpublished observations). PBDEs can also exert direct effects on neuronal and glial cells (see Costa and Giordano, 2007; Costa et al., 2008, for a complete discussion). In particular, several PBDEs have been shown to cause oxidative stress in neurons, leading to apoptotic neuronal death (He et al., 2008; Giordano et al., 2008; Huang et al., 2010; Costa et al., 2010; Tagliaferri et al., 2010). In addition, PBDEs can disrupt signal transduction mechanisms such as protein kinase C (Madia et al., 2004; Kodavanti and Ward, 2005), and calcium homeostasis, particularly through their hydroxylated metabolites (Coburn et al., 2008; Dingemans et al., 2008, 2010). A recent study (Schreiber et al., 2010) reported that in primary fetal human neural progenitor cells, cultured as neurospheres, BDE-47 and BDE-99 decreased migration and differentiation into neurons and oligodendrocytes. Of interest is that T3 antagonized the effect of PBDEs, suggesting that PBDEs may affect thyroid hormone signaling in this in vitro model. 3. General toxicology of BDE-209 BDE-209 (3,30 ,4,40 ,5,50 ,6,60 -decabromodiphenyl ether) is a fully brominated PBDE. It is sold as a decaBDE mixture which contains primarily (>97%) BDE-209, minor amounts of nonaBDEs (BDE-206, BDE-207, BDE-208), and traces of octaBDEs (Goodman, 2009).

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

Among all PBDEs, BDE-209 has the most extensive toxicological database. Standard toxicology studies (e.g. acute, sub-chronic, chronic, reproductive/developmental toxicity) are available, as well as a carcinogenicity study. These studies have been described in various reviews (Norris et al., 1975; Hardy, 2002; Hardy et al., 2002, 2009; Illinois EPA, 2006; WHO, 2006; USEPA, 2008) and are only very briefly summarized here. BDE-209 has a low acute toxicity when given by the oral route. No indications of toxicity were reported in rats after oral administration of FR-300-BA (a commercial product containing 77.4% BDE-209, 21.8% nonaBDEs, and 0.8% octaBDEs) at doses as high as 2000 mg/kg bw (Norris et al., 1975). A short-term acute toxicity study conducted by the National Toxicology Program (NTP) found no effects in rats or mice fed BDE-209 (99% pure) up to 100,000 ppm (approximately equivalent to 10,000 mg/kg/day in rats, and 20,000 mg/kg/day in mice; USEPA, 2008) in the diet for two weeks (NTP, 1986). Acute inhalation studies indicated no effects in rats upon 1–2 h exposure to commercial BDE-209 products at concentrations of 48 mg/L or 200 mg/L (reported in Hardy et al., 2009). No toxicity was also observed in rabbits upon dermal exposure of BDE-209 (unknown purity) at doses up to 8000 mg/kg (reported in Hardy et al., 2009). Early subchronic toxicity studies (30 days) with FR-300-BA in rats evidenced a NOEL of 80 mg/kg/day (0.1% in the diet) for liver changes, and of 8 mg/kg/day (0.01% in the diet) for thyroid effects (hyperplasia) (Norris et al., 1975). NTP (1986) reported NOEL values of 3000–4000 mg/kg/day in rats, and 10,000 mg/kg/day in mice, both the highest dose level tested, in a 14-weeks feeding study with BDE-209 (97–99% pure). A more recent study in which BDE-209 (purity >97%) was dissolved in toluene, then added to an emulsion of phospholipon and Lutrol F127, to increase absorption (see Section 3.1), and given by gavage for 28 days, reported increases in some metabolizing enzymes [e.g. cytochromes P450 (CYP)] at the dose of 1 mg/kg/day (Van der Ven et al., 2008). A two-fold induction of CYP1A and CYP2B by BDE-209 (>98% pure) at the doses of 10–1000 mg/kg/day was also reported, though these effects were not dose-dependent (Bruchajzer et al., 2010). Such induction of CYPs may be due to activation of the pregnane X receptor (PXR) by BDE-209 (Pacyniak et al., 2007). In another recent study BDE-209 (99% pure) was given for 90 days to rats, orally in corn oil, at the single dose level of 100 mg/kg/ day (F. Wang et al., 2010). The main observed effect was an induction of CYP2B1. BDE-209 (98% pure) was also given by gavage in corn oil–Tween 80 to rats from PND 10 to PND 42 (Lee et al., 2010). Major effects were an induction of various hepatic cytochromes P450 (CYP2B1, CYP3A1, CYP1A2), an increase of PXR and of CAR (constitutive androstane receptor) in the liver, decreased levels of T3, and histological alterations in the thyroid and the liver at the doses of 300 and 600 mg/kg/day (Lee et al., 2010). It should be noted that a study by Wahl et al. (2008) indicated that CYP induction by BDE-47 was due to contamination with brominated furans; hence, the possibility that the effects on CYPs observed with BDE-209 may be due to similar contaminants should be considered. Chronic toxicity/carcinogenicity studies were carried out with a low purity BDE-209 formulation (FR-300-BA) (Norris et al., 1975; Kociba et al., 1975), and with 94–97% pure BDE-209 (NTP, 1986). In the earlier study, no evidence of toxicity or neoplastic lesions was reported in rats exposed to BDE-209 in the diet at doses up to 1 mg/ kg/day (Kociba et al., 1975; Hardy et al., 2009). In the NTP study, BDE-209 was given in the diet at doses up to 2240–2550 mg/kg/ day in rats, and 6650–7780 mg/kg/day in mice (males and females, respectively) (NTP, 1986; USEPA, 2008). No signs of toxicity were observed. Non-neoplastic changes were found in the liver of male rats, and increases in neoplastic nodules were observed in rats of both sexes. Lymphoid hyperplasia of the mandibular lymph nodes

11

was also observed in male rats. In male mice, increased incidence of hepatic granulomas, thyroid follicular cell hyperplasia, hepatocellular adenomas/carcinomas were found. Stomach ulcers were found in female mice. NTP indicated a NOEL of 1120 and 2550 mg/ kg/day for male and female rats, respectively, and of 3200 mg/kg/ day for male mice (NTP, 1986). A NOEL of 3760 mg/kg/day was suggested for female mice (USEPA, 2008). NTP concluded there was ‘‘some, equivocal’’ evidence of carcinogenicity, while it has been rebutted that high dose exposure may induce proliferative lesions that may progress to neoplasia (Hardy et al., 2009). BDE-209 does not appear to be genotoxic (Hardy et al., 2009), and is classified by IARC in Group III (not classifiable as to its carcinogenicity to humans) (IARC, 1990). Exposure of rats to FR-300-BA by gavage (10, 10, 1000 mg/kg/ day) on gestational days (GD) 6–15 did not provide any evidence of toxicity to the dams, nor of teratogenicity (Norris et al., 1975). A one-generation reproductive study conducted in rats with the same formulation at doses of 3, 30 and 100 mg/kg/day given in the diet prior to mating, and during mating, gestation and lactation, did not provide any indication of adverse effects (Norris et al., 1975). In a more recent study, carried out in compliance with current USEPA and OECD guidelines, BDE-209 (97% pure) was administered by gavage to pregnant rats on GD 0–19 at the doses of 100, 300, and 1000 mg/kg/day (Hardy et al., 2002, 2009). No effects were found in the dams and fetuses, leading to a NOEL of 1000 mg/ kg/day (Hardy et al., 2002, 2009). In summary, BDE-209 appears to have very low acute toxicity when given by the oral, inhalation and dermal route. Upon subchronic/chronic exposure, target organs appear to be the liver and the thyroid gland. BDE-209 does not seem to be genotoxic, and evidence of carcinogenicity (in male rodents) is ‘‘equivocal’’. BDE209 is not teratogenic, and does not appear to cause developmental toxicity. NOEL values derived from all these studies are quite high, ranging from a few hundreds to several thousands mg/kg/day. 3.1. Toxicokinetic considerations for BDE-209 BDE-209 is a large, bulky molecule, with a molecular weight of 959. It has very low solubility in water, and a limited solubility in organic solvents; however, despite being lipophilic, because of its size it cannot be absorbed through the intestinal tract by passive diffusion (Morck et al., 2003). Absorption of lipophilic compounds with high molecular weight may be facilitated by carrier proteins such as P-glycoprotein (Charman, 2000). Earlier studies of oral exposure to BDE-209 (dissolved in corn oil) indicated absorption of less than 1% of the administered dose, with excretion almost exclusively through the feces (Norris et al., 1975; El Dareer et al., 1987). More than 99% of the dose was recovered in the feces by 48 h after administration, indicating a lack of accumulation in tissues (Norris et al., 1975). Using 14C-BDE (97% pure), NTP estimated an absorption following oral administration in the diet in rats of 0.3–1.5% (NTP, 1986; the higher number is from Hardy et al., 2009). More recent studies, however, have reported higher absorption rates, which may be dependent upon the solvent utilized to administer BDE-209. Sandholm et al. (2003) dissolved BDE-209 (>98% pure) in dimethylamide/polyethylene/glycol/ water and administered it to rats by gavage or intravenously at the dose of 1.9 mg/kg. Based on the comparison to plasma levels following i.v. injection, oral bioavailability was calculated to be 26%. In another gavage study by Morck et al. (2003), 14C-BDE-209 (>98% pure) was dissolved in toluene, sonicated, and suspended in Lutrol F127/soy phospholipon/water, after which the toluene was evaporated. About 90% of the dose was excreted in the feces within 3 days, mostly (65%) as metabolites, while only trace amounts (<0.1% of the dose) were excreted in the urine. Bile excretion represented 9.5% of fecal excretion (similar to the 7% previously

12

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

reported; NTP, 1986), and consisted entirely of metabolites, indicating absorption of at least 10%. The high percentage of metabolites in the feces may be explained by metabolism occurring at the levels of the gastrointestinal tract. Both of these studies (Sandholm et al., 2003; Morck et al., 2003) are discussed in detail by Hardy et al. (2009). Huwe and Smith (2007) examined the absorption, distribution, and excretion of DE-83R (a commercial product containing 98.5% BDE-209) in male rats upon dietary administration of a low dose (0.08 mg/kg). About 58% of the dose was recovered in the feces, while none was found in the urine. The remaining dose was found associated with tissues. The toxicokinetics of 14C-BDE-209 (99.8% pure, dissolved in peanut oil) in pregnant rats upon oral exposure on GD 16–19 (2/mg/kg/day) was investigated by Riu et al. (2008). Approximately 72% of the dose was found in the feces and the digestive tract content, while only 0.1% was excreted in the urine. The remaining of the dose was distributed in various tissues (e.g. 6.5% was in the liver). The fetuses (whole litter) contained only 0.43% of the dose (Riu et al., 2008). A study by Hughes et al. (2001) examined dermal absorption of 14C-BDE (>98% pure) in vitro, in skin of hairless mice. Results indicate a very poor absorption of 0.07–0.34%, depending on the concentration of BDE-209. Placental transfer of BDE-209 has been investigated in a human placenta perfusion system (Frederiksen et al., 2010), and found to be very low compared to lower brominated PBDEs. This may explain the very low levels or absence of BDE-209 in fetal or umbilical cord blood (see Section 5 and Table 4), though placental BDE-209 levels may be relatively higher than other PBDEs in relationship to maternal blood levels (Frederiksen et al., 2009). A recent study reported that administration of BDE-209 (98% pure) to rats from GD 7 to lactational day (LD) 4, at the oral dose of 5 mmol/kg/day, resulted in 10-fold lower levels in fetuses and pups than in the dams (Cai et al., in press). In another recent experimental study, BDE-209 (97.5% pure) was given by gavage in corn oil to rats from GD 6 to LD 4 at the doses of 100, 300, and 1000 mg/kg/day (Biesemeier et al., 2010). Plasma levels in GD 20 fetuses were 2.5–5-fold lower than in the dams, and in either dams or fetuses there was a lack of dose–response. Levels of BDE-209 in milk were lower than in the mother’s plasma (Biesemeier et al., 2010). In summary, BDE-209 differs from other lower brominated PBDEs which were found to be absorbed to a significant degree and to have long half-lives (Hakk and Letcher, 2003). BDE-209 absorption upon oral administration is limited (10–25% at most), and it is rapidly excreted through the feces, with little accumulation in tissues, and a short half-life. However, no animal studies are available to assess BDE-209 absorption, distribution and excretion upon inhalation and dermal exposures, which may have relevance for humans. Results from a human study in BDE-209-exposed workers has indicated a half life of serum BDE-209 of 15 days, compared to half-lives of 18–39 days for nonaBDEs, 37–91 days for octaBDEs (Thuresson et al., 2006), and even longer half-lives for lower brominated PBDEs (Geyer et al., 2004). Similar half-lives for BDE-209 have been found in grey seals (8.5–13 days) and in European starlings (13 days) (Thomas et al., 2005; Van den Steen et al., 2007). A continuing issue of debate is whether BDE-209 is metabolized, to what extent it is metabolized, where metabolism occurs, and what metabolites are formed (Hakk and Letcher, 2003). In the environment, BDE-209 has been shown to undergo photolytic debromination with the formation of nona- to tetraBDEs (Soderstrom et al., 2004; Ahn et al., 2006; Christiansson et al., 2009). During such photodegradation of BDE-209, free radicals may be formed (Suh et al., 2009). Microbial debromination has also been reported, with the formation of hepta- and octaBDEs (He et al., 2006), and hexa- to nonaBDEs (Tokarz et al., 2008). Octa- and

nonaBDEs were also found when house dust containing BDE-209 was exposed to natural sunlight (Stapleton and Dodder, 2008). Debromination of BDE-209 has been observed in juvenile rainbow trout and in common carp exposed to food containing 939 ng/g wet weight of BDE-209 (Stapleton et al., 2006). Hexa-, hepta-, octa- and nonaBDEs were detected in fish tissues, with the highest metabolite being BDE-202, an octaBDE (Stapleton et al., 2006). Based on the total body burden of PBDE congeners, the uptake of BDE-209 was estimated at 3.2% (Stapleton et al., 2006). The same PBDE congeners were produced when liver microsomes from either fish were incubated in vitro with BDE-209 (Stapleton et al., 2006). Biota (sunfish, crab chub, and crayfish) from a wastewater receiving stream were also shown to debrominate BDE-209 to hepta- and octaBDEs (La Guardia et al., 2007). Debromination of BDE-209 to hepta-, octa- and nonaBDEs in lactating cows has also been suggested (Kierkegaard et al., 2007). In rats fed 100 mg/kg/day BDE-209 for three months, octa- and nona-brominated PBDEs were found in liver and kidney (F. Wang et al., 2010). In contrast, human hepatocytes in vitro have provided evidence of low or no metabolism, possibly because of low entry of BDE-209 into cells under the experimental conditions utilized (Stapleton et al., 2009). 4. Evidence of developmental neurotoxicity of BDE-209 4.1. Animal studies There are only a limited number of published studies investigating the developmental neurotoxicity of BDE-209, and a summary is given in Table 1. A first series of studies was carried out at Uppsala University in Sweden by Viberg, Eriksson, and their colleagues (Viberg et al., 2003a, 2007, 2008; Johansson et al., 2008; Viberg, 2009a), and these have been critically discussed in some recent publications (Goodman, 2009; Hardy et al., 2009). The same paradigm of exposure, similar to that utilized by these investigators with other PBDEs, was used in all studies: BDE-209 (purity >98%) was dissolved in a mixture of egg lecithin and peanut oil, which was sonicated with water to yield a 20% fat emulsion. BDE209 was administered by gavage (10 ml/kg bw) with a metal gastric tube to NMRI (Naval Medical Research Institute) mice on postnatal day (PND) 3. Doses of BDE-209 ranged from 1.34 to 20.1 mg/kg bw (1.4–21.0 mmol/kg bw). In an initial study (Viberg et al., 2003a) BDE-209 was also administered to mice at PND 10 (1.34, 13.4 and 20.1 mg/kg bw) and PND 19 (2.22 and 20.1 mg/kg bw). The PND 10 exposure is the same used by these investigators with other PBDEs and with other chemicals, and had been found in these other cases to produce developmental neurotoxicity (see e.g. Eriksson et al., 2001; Viberg et al., 2002, 2003b). In contrast, no behavioral effects were seen when BDE-209 was administered on PND 10, which led the investigators to administer the compound on PND 3 or PND 19 instead. Viberg et al. (2003a) evaluated the mice ‘‘habituation profile’’ upon a 60 min test period, divided into three 20 min periods. In control animals, habituation is observed, defined as a decrease of locomotion, rearing and total activity over the 60 min test period. In two month-old animals exposed to BDE-209 on PND 3, there was a decrease of locomotor activity in the first 20 min of observation, and an increase in the last 20 min, interpreted by the investigators as suggestive of a decreased habituation. However, an alternative interpretation of this finding is that the decreased activity at the beginning of the session is a possible sign of increased anxiety or arousal, and the elevated activity in the last part of the test period simply reflects the fact that the animals finally begin to explore their environment (Costa and Giordano, 2007). These changes were observed mostly at the high dose of BDE-209 (20.1 mg/kg), while only changes in locomotion were

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

13

Table 1 Developmental neurotoxicity of BDE-209. Species

Treatment

Effects

Reference

NMRI mice

2.22 and 20.1 mg/kg, oral, in el/po, on PND 3 or 10 or 19 6.7 and 20.1 mg/kg, oral, in el/ po, on PND 3 1.34, 2.22, 13.4, 20.1 mg/kg, oral, in el/po, on PND 3

Decreased habituation at 2, 4, 6 months at the high dose, only if exposed on PND 3 Decreased habituation at 2 months at both doses; altered response to nicotine at 2 months (high dose only) Decreased habituation at 2 and 4 months at all but the lowest dose; worsening with age; altered response to nicotine at 4 months Alteration in CaMKII (increase), GAP43 (increase/decrease), and BDNF (decrease) expression in brain areas at PND 10 Alterations in synaptophysin (increase) expression in hippocampus at PND 10 Some developmental delays; hyperactivity in males on PND 70 (high dose only); no effect at 1 year; decreased T4 in males at PND 21 Alterations in FI, FR1 and visual discrimination task at 16 months, not at PND 70 (high dose only) Decreased T3 levels; changes in drug metabolizing enzymes

Viberg et al., 2003a

Sprague–Dawley rats NMRI mice

NMRI mice NMRI mice C57BL6/J mice

C57BL6/J mice CD-1 mice Sprague–Dawley rats

Wistar rats

Wistar rats Wistar rats

CD(SD) IGS rats Sprague–Dawley rats

20.1 mg/kg, oral, in el/po, on PND 3 20.1 mg/kg, oral, in el/po, on PND 3 6 and 20 mg/kg/day, oral, in el/po, on PND 2–15 6 and 20 mg/kg/day, oral, in el/po, on PND 2–15 10, 500, 1500 mg/kg/day, oral, from GD 0 to GD 17 5, 40, 320 mg/kg/day, oral, in co/Tween-80, from GD 6 to GD 18 100, 300, 600, 1200 mg/kg/ day, oral in po, during pregnancy 100, 300, 600, 1200 mg/kg/ day during pregnancy 20 mg/kg/day, oral, in el/po on GD 1–21, LD 1–21, PND 3–21, PND 21–41, GD 1–LD 21 + PND 22–41 10, 100, 1000 ppm in the diet from GD 10 to LD 20 1, 10, 100, 1000 mg/kg/day, oral, in co, from GD 6 to LD 21

Viberg et al., 2007 Johansson et al., 2008

Viberg et al., 2008 Viberg, 2009a Rice et al., 2007

Rice et al., 2009 Tseng et al., 2008

Decreased T4 levels (females only, high dose only). No histological changes in hippocampus

Kim et al., 2009

Decreased levels of GAP43 (high dose only) and of BDNF (all but the lowest dose) in hippocampus

Jiang et al., 2008

Alterations in Morris maze and in hippocampus (600 and 1200 mg/kg only) Changes in hippocampal synaptic plasticity (LTP) particularly with direct exposure of pups

Wu et al., 2008

Histopathological changes in thyroid; impaired oligodendroglia development (100, 1000 ppm) No effects on locomotors activity, learning and memory (Biel maze swimming test), brain histopathology

Xing et al., 2009

Fujimoto et al., in press Jacobi et al., 2009; Silberberg et al., 2009

El/po = egg lecithin/peanut oil; co = corn oil.

observed at the low dose (2.22 mg/kg). Similar changes were also observed at 4 and 6 months (Viberg et al., 2003a). From all these data, a habituation ratio was derived, which suggested that behavioral effects of BDE-209 increased with age. As said, no changes in habituation were seen when mice were dosed on PND 10 or PND 19. To explore possible reasons explaining the lack of effect of BDE-209 on PND 10, as had been observed for several other PBDEs (Eriksson et al., 2001; Viberg et al., 2002, 2003b), levels of radioactivity were measured in mouse brain upon oral administration of 14C-BDE-209 on PND 3, PND 10 or PND 19. After administration on PND 3, levels of radioactivity were 4.8% of the dose at 24 h, and increased to 7.4% at 7 days. Similar results were found after administration on PND 10, while upon administration on PND 19, radioactivity levels were only 0.6% of the dose, and did not increase with time (Viberg et al., 2003a). The authors calculated that levels of radioactivity were similar 7 days after administration on PND 3 and 24 h after administration on PND 10; since the latter treatment did not induce any behavioral effect, they concluded that metabolites of BDE-209, accumulating in brain upon exposure, may be responsible for the neurotoxic effect. The presence of such putative metabolites in brain on PND 10 would thus be of most relevance, as possible formation of these metabolites, even at higher levels, on PND 17 (7 days after exposure on PND 10) was devoid of effects (Viberg et al., 2003a). This study has been criticized on many levels. In a letter to the editor, Vijverberg and van den Berg (2004) questioned the levels of BDE-209 (and/or its metabolites) present in brain which they calculated as 3.9 mM (26.6 mg/g lipid), which is about three orders of magnitude higher than the levels found in human tissues (see also Section 5 and Table 4). They also wondered that lower brominated PBDEs would exert the same neurotoxic effects at similar dose levels and brain concentrations, when usually higher

brominated compounds present lower biological and toxicological activity. The authors responded that though the dose may be high, their findings, similar to those found with other PBDE congeners and with other compounds (e.g. PCBs), may have value for risk assessment, as they point out to a developmental period of potential great susceptibility to external insult (Eriksson and Viberg, 2004). An additional criticism, which has been addressed at different times to several other studies carried out by this Swedish group (see for example, Hardy and Stedeford, 2008), was related to the so called litter-effect (Goodman, 2009; Hardy et al., 2009). In developmental toxicity and neurotoxicity studies, when the dam is exposed to the chemical, it is generally accepted that the statistical unit should be the litter, not the individual pup (Holson et al., 2008). It has been also argued that the same approach should be used for direct dosing of preweaning pups (Moser et al., 2005). Viberg et al. (2003a) selected 10 mice from 3 to 5 litters; thus the n value should have been 3 or 5 instead of 10. Viberg, Eriksson and colleagues have responded to this criticism in a number of occasions (Eriksson et al., 2005; Eriksson, 2008). They presented data, obtained with BDE-99, indicating that the same statistically significant effects were found when nine pups were treated with BDE-99 on PND 10 and assessed for locomotor activity at 2 months of age, regardless of whether they derived from three litters (3/ litter) or from nine litters (1/litter). Furthermore, it should be noted that another group of investigators utilized the same exposure protocol, and reported locomotor activity changes upon a single PND 10 administration of BDE-47 utilizing the litter as the statistical unit (Gee and Moser, 2008). In a follow-up study in mice, four dose levels of BDE-209 were used (1.34, 2.22, 13.4 and 20.1 mg/kg bw), given on PND 3 by gavage (Johansson et al., 2008). The results of the spontaneous behavior and habituation test, at 2 months of age, were similar to

14

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

the previous study, in that a dose-dependent decrease in locomotion, rearing and total activity was observed in the first 20 min of the test, while an increased activity (suggesting a decreased habituation) was seen in the last 20 min. The effects were also present at 4 months of age, and appeared to worsen with age. Somewhat differently from the earlier study (Viberg et al., 2003a), doses as low as 2.22 mg/kg bw were shown to cause such behavioral alterations. At the age of 4 months, mice were challenged with an acute dose of nicotine (80 mg/kg, s.c.). In control mice this dose of nicotine induced hyperactivity; such hyperactivity was less pronounced in mice given the two lower doses of BDE-209 (1.34 and 2.22 mg/kg) on PND 3, while mice given the two higher doses of BDE-209 (13.4 and 20.1 mg/kg) exhibited hypoactivity (Johansson et al., 2008). A similar response (hypoactivity following exposure to nicotine) had been previously shown in mice exposed to BDE-99 on PND 10 (Viberg et al., 2002). The authors interpreted this finding as indication of a possible long-lasting modification of the cholinergic system caused by BDE209 or/and its metabolites, but the mechanisms and significance of these observations with nicotine remain obscure. Four month-old mice, exposed to BDE-209 on PND 3 were also tested in the elevated plus maze; no significant effects were found with regard to entries or time spent in the open arms. This is somewhat surprising, as the elevated plus maze can often reflect alterations in locomotor activity (Vitalone et al., 2008). Criticisms of this study focused particularly on the lack of control for the litter effect (Goodman, 2009; Hardy et al., 2009). The effects of neonatal exposure to BDE-209 (6.7 or 20.1 mg/kg bw by gavage on PND 3) were also investigated in Sprague–Dawley rats (Viberg et al., 2007). The results were essentially the same as those found in mice. Both doses of BDE-209 caused alterations in spontaneous behavior and habituation at 2 months of age. In addition, rats treated with the high dose of BDE-209 displayed hypoactivity rather than hyperactivity after a single dose of nicotine (80 mg/kg bw, s.c.). Also in this study with rats, there was no consideration of the litter effect. Two studies by the same group examined biochemical alterations in mice on PND 10 following exposure to BDE-209 on PND 3 (20.1 mg/kg bw) (Viberg et al., 2008; Viberg, 2009a). A significant 38% increase of CaMKII (calcium/calmodulin dependent protein kinase II) protein was found in the hippocampus, but not in cerebral cortex. Levels of GAP-43 (growth-associated protein-43) were significantly increased (by 13.4%) in the hippocampus, but were decreased (by 13%) in the cerebral cortex. BDNF (brainderived neurotrophic factor) levels were decreased in the hippocampus and unchanged in the cerebral cortex (Viberg et al., 2008). Levels of synaptophysin were also increased (by 41%) in the hippocampus, but not in the cerebral cortex, while levels of tau were unchanged in either brain region (Viberg, 2009a). All these proteins are known to play important role in brain processes such as neurite outgrowth, synaptogenesis, synaptic plasticity, and neuronal survival. Though the significance of the observed modifications remains obscure, a study by Jiang et al. (2008) published in a Chinese journal and available only in abstract form, reported similar alterations in Wistar rats exposed during pregnancy (exact period not specified) by gavage to BDE-209 (purity not specified) dissolved in peanut oil at the doses of 100, 300, 600 and 1200 mg/kg bw. Levels of GAP-43 in the hippocampus of offspring rats (age not specified) were decreased in the two high dose groups, while BDNF levels were decreased in all but the lowest dose group. These results thus appear to confirm those obtained by Viberg et al. (2008), though the paucity of information available does not allow a full evaluation of the study by Jiang et al. (2008). Two additional studies by the Swedish group (Viberg et al., 2006; Viberg, 2009b) examined the behavioral and biochemical

effects of exposure to two other highly brominated PBDEs, BDE203 (2,20 ,3,4,40 ,5,50 ,6-octaBDE; 16.8 mg/kg bw) and BDE-206 (2,20 ,3,30 ,4,40 ,5,50 ,6-nonaBDE; 18.5 mg/kg bw). Mice were treated by gavage with either compound (purity >98%; dissolved as BDE209) on PND 10. At two months of age, mice exhibited the same alterations in spontaneous behavior and habituation as shown for BDE-209, as well as an altered response to nicotine (Viberg et al., 2006). In addition, 24 h after administration (i.e. on PND 11) both compounds caused increases in CaMKII and synaptophysin levels in the hippocampus, but not in cerebral cortex, and no changes in GAP-43 and tau levels in either brain region (Viberg, 2009b). The authors considered these findings as confirmation of their hypothesis that BDE-209 exerts its developmental neurotoxicity through the accumulation of debrominated metabolites in the brain, as these putative metabolites (BDE-203 and BDE-206), when administered on PND 10, caused the same effects of BDE-209 given on PND 3. In support of this hypothesis, they report that BDE-206 caused developmental effects only when given on PND 10, but not on PND 3, while BDE-203 caused behavioral effects on spontaneous behavior and habituation when given at both time points, but effects were significantly greater upon the PND 10 administration (Viberg et al., 2006). Furthermore, subtle but statistically significant effects in the Morris water maze were seen in 3 month-old mice upon administration of BDE-203 on PND 10 (Viberg et al., 2006). In this study, a heptaBDE (BDE-183) was also tested and found to exert behavioral effects on locomotor activity/habituation when given on PND 3 but not on PND 10. Thus, overall, results are not as clear-cut, and not so easy to interpret. BDE-203 appeared the most potent of this group of highly brominated PBDEs. BDE-183, BDE-203 and BDE-206 have been shown to be metabolites of BDE209 upon oral administration to rats (Morck et al., 2003; Sandholm et al., 2003). However, whether any of these metabolites is responsible for the observed effects of BDE-209 remains to be determined. A study by Tseng et al. (2008) in CD-1 mice investigated the reproductive toxicity of BDE-209. BDE-209 (98%) was dissolved in corn oil and administered by gavage to dams daily from GD 0 to GD 17 at the doses of 10, 500 and 1500 mg/kg bw. No effects on reproductive parameters (e.g. gestational length, litter size etc.), or on developmental landmarks (e.g. pinna detachment, incisor eruption, eyes opening etc.) were found. These findings confirmed previous results obtained by the same investigators (Tseng et al., 2006). At PND 71, three male offspring per litter were randomly selected. Serum T3 levels were equally decreased (by 20%) in mice exposed in utero to 10 or 1500 mg/kg BDE-209, while serum levels of T4 were unaffected. Minor histological changes were found in the thyroid of the 1500 mg/kg group. Hepatic UDPGT (uridinediphosphate-glucuronosyltransferase) activity was unchanged, while there was a significant increase in EROD (7-ethoxyresorufin O-deethylase) in the high dose group (Tseng et al., 2008). These findings differ somewhat from those obtained with other PBDEs, where a decrease of T4 was reported, together (though not always), with an increase in UDPGT activity (Zhou et al., 2002; Richardson et al., 2008). A similar study was carried out by Kim et al. (2009). Sprague– Dawley rats were exposed from GD 6 to GD 18, by gavage to BDE209 (98% pure, dissolved in corn oil and Tween-80) at the doses of 5, 40 or 320 mg/kg bw. Decreases in body weight of dams in the 40 mg/kg group and in offspring of the 320 mg/kg group were observed; however, there were no changes in reproductive parameters. On PND 42, in male offspring, there was a non dose-dependent increase in thyroid weight, while in females a decreased weight of the adrenals (40 and 320 mg/kg) and the uterus (high dose only) was observed. Thyroid hyperplasia had been also observed upon exposure to decaBDE in adult rats (Norris et al., 1975). Serum levels of T4 were decreased in female rats,

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

while TSH levels were increase in male and female rats, only in the 320 mg/kg group. No differences with controls were found with regard to BrdU-positive cells in the hippocampus, hippocampal neuronal density or shape or glial markers (Kim et al., 2009). Xing et al. (2009) carried out a study with BDE-209 (99%, dissolved in a mixture of egg lecithin and peanut oil (1:10 w/w) and then sonicated with water to yield a 20% (w/w) fat emulsion) in Wistar rats. BDE-209 was administered daily by gavage via a metal gastric tube at the dose of 20 mmol/kg (approx. 20 mg/kg bw) according to the following pre- and post-natal schedules: Group I, GD 1–21; Group II, LD 1–21; Group III, PND 3–21; Group IV, PND 21–41; Group V, GD 1–LD 21 and PND 22–41. At the age of 60 days, animals from three to four litters, both males and females, underwent a series of electrophysiological tests to assess input/ output functions, paired-pulse reactions, and LTP (long term potentiation) in the hippocampus. The results indicated that developmental BDE-209 exposure had significant effects of synaptic plasticity. However, such effects were not seen in all groups. Indeed, prenatal exposure (Group I) did not cause any effect, while lactational exposure (Group II) caused a significant alteration of LTP only. All parameters measured were altered in the other three groups which included direct dosing of the pups. Hippocampal BDE-209 levels were highest in the most affected groups; indeed upon gestational exposure, BDE-209 concentration was about 10 ng/g (approximate value, derived from Fig. 6 in Xing et al., 2009, corresponding to 10 nM), it increased to about 40 ng/ g in Group II, after lactational exposure, but was over 160 ng/g upon direct exposure to pups on PND 1–21, and more than 200 ng/ g (200 nM) in Group V, which received BDE-209 daily from GD 1 to PND 42. Thus, exposure during gestation yielded the lowest brain levels of BDE-209 and was devoid of effects, while lactational exposure, which impaired LTP, resulted in brain levels approximately 4-fold higher, indicating that exposure via milk may have a more significant effect on body burden and on possible ensuing effects. Direct exposure of pups led to high brain BDE-209 levels, and significant alterations in all parameters examined (Xing et al., 2009). Criticisms of the study reflect the lack of consideration for the litter effect, which would reduce the n value from 5–13 to 1–4 (G.T. Johnson et al., 2010). Rice et al. (2007) investigated the effects of developmental exposure to BDE-209 in C57BL6/J mice. BDE-209 (99.5% pure) was dissolved in a 1:10 egg lecithin:peanut oil mixture that was sonicated and hand-shaken to a 20% emulsion in sterile water, and was given to mouse pups using a micropipette daily from PND 2 to PND 15, at the doses of 6 or 20 mg/kg bw. Exposure to BDE-209 did not affect developmental milestones or body weights. A Functional Observational Battery (FOB), adapted for mice pups (Rice et al., 2007), conducted (in a blind fashion) every other day between PND 2 and PND 20, revealed a developmental delay in a number of functions and reflexes (palpebral reflex, forelimb grip, struggling behavior during handling). Mice were tested for locomotor activity on PND 70; a significant hyperactivity was observed in male, but not in female animals, at the higher dose. Naı¨ve animals (exposed to BDE-209, but never handled or tested), were examined for effects on locomotor activity at one year of age; no difference between treated and control mice were found (Rice et al., 2007). Serum levels of T4 were decreased in males only at PND 21. In all experiments the litter was considered as the statistical unit. In a follow-up study, Rice et al. (2009) utilized littermates of mice of the previous study, to investigate the effects of developmental exposure to BDE-209 (6 or 20 mg/kg, by gavage from PND 2 to PND 15) on three behavioral tasks: fixed-ratio (FR) schedule of food reinforcement, fixed-interval (FI) schedule, and light–dark visual discrimination. Animals underwent behavioral training and testing at the ages of 70 days or 16 months. Male and female mice were used, and the statistical unit was represented by

15

the litter. No effects on FI performance were found in young adult mice; however, older littermates exposed to the high dose of BDE209 performed less efficiently than controls. The higher response rate in aging mice may represent increased impulsivity, as suggested by other studies with PCBs or lead (Stewart et al., 2005; Rice, 2006). Similarly, developmental BDE-209 exposure (only at the 20 mg/kg dose) altered performance on the FR1 schedule in aging mice. Furthermore, older BDE-209 exposed mice (again, high dose only) performed poorly in the light–dark visual discrimination task, while younger mice performed only minimally differently from controls. Rice et al. (2009) noted an increase in perseverative errors in the latter testing procedure, suggestive of an inability to respond appropriately to the consequences of previous choices, in turn an indication of increased impulsivity. Effects were seen in both male and female mice. Thus, while hyperactivity was present in young (PND 70) but not older (one year-old) mice (Rice et al., 2007), the opposite seemed to be true for the behavioral tasks examined in the most recent study (Rice et al., 2009). Indeed in the latter, aging appears to unmask behavioral effects not seen in younger mice. Viberg et al. (2003a) had similarly reported that behavioral effects of developmental BDE-209 exposure appeared to worsen with age. Similar worsening of behavioral effects as the animal ages had been also observed with other PBDEs (e.g. BDE-153, BDE-99; Viberg et al., 2003b, 2004) and methylmercury (Newland and Rasmussen, 2000), but not with other compounds (e.g. PCB 126; Vitalone et al., 2010). An additional study, published in a Chinese journal and available only in abstract form, describes the effects of maternal exposure to BDE-209 on learning and memory in the offspring (Wu et al., 2008). Wistar rats were gavaged during pregnancy (exact period not specified) with BDE-209 (purity not specified) dissolved in peanut oil at the doses of 100, 300, 600 and 1200 mg/kg bw. Animals exposed to the three highest doses of BDE-209 showed prolonged escape latency in the Morris water maze (age of testing not specified). Animals exposed in utero to 600 and 1200 mg/kg BDE-209 also presented histological alterations in the hippocampus. The lack of details does not allow a full evaluation of this study. In another recent study, BDE-209 (purity >98%) was given to pregnant rats in the diet at levels of 10, 100 and 1000 ppm, from GD 10 to LD 20 (Fujimoto et al., in press). No reproductive effects were observed, in agreement with previous findings by Hardy et al. (2002). Body weights in male pups were slightly increased at the two lowest dose levels, while thyroid hormones were decreased at the highest dosage, with hypertrophy of thyroid follicular cells. In brain, a decrease of CNPase (20 ,30 -cyclic nucleotide 30 -phosphodiesterase, a marker of oligodendrocytes) was found in the cingulated deep cortex at postnatal week 11 at the two highest dose levels, indicating white matter hypoplasia (Fujimoto et al., in press). In the amphibian Xenopus laevis, exposure of tadpoles to the technical decabromodiphenyl ether mixture DE-83R (containing 96–99% BDE-209, as well as other PBDEs) in water, from stage 46/ 47 (free swimming larvae) to stage 62, at the concentrations of 1– 1000 ng/L, did not cause any malformation or abnormal behavior. However, histological alterations of follicular epithelial cells were found at 100 and 1000 ng/L, with a decrease in thyroid receptor (TR bA) mRNA levels observed at all DE-83R concentrations (Qin et al., 2010). In contrast to all these studies, a recent study, described by Hardy et al. (2009), and presented so far only in abstract form (Silberberg et al., 2009; Jacobi et al., 2009), did not find any evidence of developmental neurotoxicity of BDE-209. This developmental neurotoxicity study was carried out according to OECD 426 (2007) and USEPA 870.6300 (1998) guidelines. BDE-209 (97.5% pure), dissolved in corn oil, was administered by gavage to Sprague–Dawley rats, daily from GD 6 to LD 21, at the doses of 1,

16

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

10, 100 or 1000 mg/kg bw. For all evaluations, the litter served as the statistical unit. No treatment-related effects were found with regard to offspring survival and development, locomotor activity (PND 13, 17, 21, 61, 120, 180), startle response (PND 20, 60), learning and memory (PND 22, 62; Biel maze swimming test), brain weight and neuropathological examination on PND 21 and PND 72 (Silberberg et al., 2009; Jacobi et al., 2009). The NOAEL from this study was thus considered to be 1000 mg/kg/day (Hardy et al., 2009; Silberberg et al., 2009; Jacobi et al., 2009). In summary, a limited number of studies, carried out using different exposure protocols, indicate subtle developmental effects of BDE-209, particularly in the domains of locomotor activity and of cognitive behavior. These types of behavioral alterations were also reported upon exposure to other PBDEs. In contrast to these studies, a developmental neurotoxicity study carried out according to established guidelines, failed to provide evidence of developmental effects. 4.2. In vitro studies While there are several publications on various effects of lower brominated PBDEs in vitro (see Section 2 and references in Costa and Giordano, 2007; Costa et al., 2008), there is only limited information for BDE-209. Chen et al. (2010) investigated the effects of BDE-209 (purity >98%), dissolved in DMSO (dimethylsulfoxide) in rat hippocampal neurons. Concentrations of 10, 30 and 50 mg/ml (equivalent to approximately 10.4, 31.2 and 52 mM) were utilized. BDE-209 caused a concentration-dependent decrease of cell viability, an increase in apoptotic cell death, and increased phosphorylation of p38 MAPK at all concentrations tested. Levels of ROS (reactive oxygen species), malonyldialdehyde (MDA), and nitric oxide (all indexes of oxidative stress), and of intracellular calcium, were also increased, while levels of superoxide dismutase (an antioxidant enzyme) were decreased. Global DNA methylation was reduced at the two lower concentrations, but not altered by the highest concentration (50 mg/ml) (Chen et al., 2010). Another study in rat hippocampal neurons confirmed that BDE-209 (10, 30 and 50 mM; purity not specified) caused a decrease in cell viability, apoptotic cell death, increase in intracellular calcium and ROS levels, which were antagonized by the antioxidant N-acetylcysteine (Zhang et al., 2010a). A study by Huang et al. (2010) compared the cytotoxicity and intracellular accumulation of five PBDE congeners, including BDE209, in mouse cerebellar granule neurons. Overall, this study showed that all congeners tested were capable of inducing apoptotic cell death, most likely mediated by the induction of oxidative stress, and that neurotoxicity was related to the ability of PBDE congeners to accumulate in cells, particularly in mitochondria (Huang et al., 2010). BDE-209 (99.9% pure) was the least potent of the five congeners tested (BDE-47, BDE-99, BDE-100, BDE-153). In cytotoxicity assays, cell viability was decreased (by about 20%) by BDE-209 only at 50 mM, while no apoptosis was detected in a range of concentrations between 1 and 50 mM. Small increases in ROS (measured at 60 min) and MDA (measured at 12 h) were induced only by 50 mM BDE-209. Among the congeners tested, BDE-209 showed the least intracellular accumulation upon exposure for 24 h to different concentrations (0.01, 0.1, 1.0, 10.0 mM). Magnification at 1 mM was about 30-fold, compared for example to 300-fold for BDE-47. Subcellular fractionation indicated that the fraction of each PBDE congener associated with mitochondria was also lowest for BDE-209 (11%, compared for example with 43% of BDE-100). Most BDE-209 was associated with the microsomal fraction (a collection of endoplasmic reticulum, Golgi apparatus, intracellular vesicles, and plasma membrane), with lower levels in the cytosol (approx. 10%), and very low levels in the nucleus (0.01% of the total BDE recovered) (Huang et al.,

2010). This study indicates that BDE-209 displayed the least biological activity in vitro and the least cellular accumulation, most likely because of its bulky configuration (Huang et al., 2010). Induction of apoptosis and of oxidative stress by BDE-209 were also investigated in human hepatoma cells HepG2 (Hu et al., 2007) and in RTG-2 cells, a fibroblast-like cell line derived from trout gonadal tissue (Jin et al., 2010). In hepatoma cells, BDE-209 (>99.5% pure) inhibited cell proliferation, induced apoptotic cell death, and increased levels of ROS in a concentration-dependent manner between 10 and 100 mM (Hu et al., 2007). Similar findings, with concentrations of BDE-209 (98% pure) ranging from 12.5 to 100 mM were reported in RTG-2 cells (Jin et al., 2010). It should be noted that, given the low solubility of BDE-209, results obtained with concentrations above 40–50 mM may be compromised, and difficult to interpret. In contrast, another in vitro study reported that DE-83 (containing 96.9% BDE-209 and 3.1% nonaBDEs) did not alter the expression of antioxidant genes such as the thioredoxins in human umbilical vein endothelial cells, at concentrations up to 40 mM, while a penta- and an octaBDE mixture did (Kawashiro et al., 2009). These latter results appear to be in agreement with those found with individual congeners (Huang et al., 2010). In one study which addressed end-points other than apoptotic cell death and oxidative stress, Xing et al. (2010) investigated the interaction of BDE-209 with voltage-gated sodium channels in rat hippocampal neurons. Using a whole-cell patch-clamp approach they reported that BDE-209 (99% pure) at concentrations of 0.1– 2.0 mM decreased sodium channel currents. A partial antagonism of such effects was afforded by ascorbic acid and by Vitamin E, leading the authors to conclude that oxidative stress may play a role in sodium channel inactivation by BDE-209. The limited description of experimental methodologies, and the lack of direct testing of this hypothesis, however, does not allow any conclusion in this regard. Another study examined the effects of BDE-209 (10, 30, 50 mM, purity not specified) on differentiation of rat neural stem cells (NSC) in vitro (Zhang et al., 2010b). BDE-209 was found to inhibit neurite outgrowth and the differentiation of NSC into neurons in a concentration-dependent manner, while enhancing the ratio of differentiation of NSC into glial cells (Zhang et al., 2010b). In another recent study, Ibhazehiebo et al. (in press) reported that BDE-209 (purity >99%) disrupted thyroid hormonemediated transcription in a reporter gene assay in fibroblastderived CV-1 cells, at concentrations as low as 0.01 nM. Such effect appeared to be due to interference with the thyroid receptor-DNA binding domain, rather than to competitive inhibition of T3 binding to the thyroid receptor or to alterations of thyroid receptor-cofactor binding. At similar low concentrations (0.1 nM), BDE-209 was found to inhibit thyroid hormone-induced dendrite arborization of Purkinje cells in vitro (Ibhazehiebo et al., in press). In summary, the few in vitro studies with BDE-209 have evidenced similar cellular effects (e.g. oxidative tress, apoptosis) as observed with other PBDEs. However, with few exceptions, these effects were generally observed at higher concentrations (approx. 50 mM), possibly because of the difficulty of BDE-209 to enter the cells given its bulky configuration. 5. Occurrence and body burden of BDE-209 The analysis of BDE-209 is considered a challenging task, due to its high molecular weight, its peculiar chromatographic characteristics, its thermolability and photosensitivity, and the possibility of contamination from flame-retarded electronic equipment (Thuresson et al., 2005). Thus, it is not surprising that much less information exists on levels of BDE-209 in environmental media, biota and humans, compared to other lower-brominated PBDEs. Furthermore, the indicated issues may lead to under- or over-estimation of actual BDE-209 levels. As for all PBDEs, the main

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

sources of exposure to BDE-209 are thought to be the diet, household dust, and the occupational setting, with different relative contributions of each route depending on the age of the individual (Johnson-Restrepo and Kannan, 2009). A 2003–2004 survey of 62 food items purchased in three supermarkets in Dallas, Texas, indicated that BDE-209 was present in all foods except margarine, with concentrations ranging from 10.3 ppt in eggs to 288 ppt in chicken liver (Schecter et al., 2006). A more recent survey in the same geographical area identified BDE209 in most of 31 distinct food types (310 total samples of meats, fish and dairy products); the highest levels of BDE-209 were found in butter (5190 ppt) (Schecter et al., 2010a). Similar findings were reported for foodstuff purchased in supermarkets in Los Angeles, California, and Albany, New York, indicating an overall lack of regional differences (Schecter et al., 2010b). A survey carried out in Spain identified BDE-209 as the predominant PBDE present in oil and eggs, and present in all foodstuffs examined (Gomara et al., 2006). In a survey of seafood products (shellfish) of South China, BDE-209 was found to represent about 50% of the total PBDE levels, with highest levels of 962 ppt found in crab (Guo et al., 2007). Levels of BDE-209 in honey samples from various countries ranged from non detectable to 9260 ppt, with the highest levels found in samples from developing countries, where it accounted for 65% of all PBDE content (J. Wang et al., 2010b). PBDEs have also been found in human breast milk, and levels in North America (U.S.A. and Canada) are higher than those reported in Europe, Asia or Australia (see Costa and Giordano, 2007). However, this does not appear to be the case for BDE-209. Table 2 shows the levels of BDE-209 in breast milk in different countries. All milk samples were collected in the past decade, which had P already seen a significant increase of PBDE levels from the 1970s (Costa and Giordano, 2007). Data from the Faroe Islands confirm that there was an increase of BDE-209 levels in breast milk over time (Table 2; Fangstrom et al., 2005a, 2005b). With few exceptions (e.g. Gomara et al., 2007; Jin et al., 2009), BDE-209 levels in milk were in the range of 0.5–3.0 ng/g of lipid, suggesting that this congener is not a preponderant congener among the

17

PBDEs found in milk, as most often BDE-47, BDE-99, and BDE-153 are found (Costa and Giordano, 2007). For example, in the U.S., P breast milk levels of PBDE were 20–50 ng/g lipid (Table 2 in Costa and Giordano, 2007), while BDE-209 levels were <1 ng/g lipid (Table 2). Unfortunately, no data on breast milk levels of BDE209 were found for the U.K., where dust levels of this congener are the highest (see Table 3). Evidence that occupational exposure may lead to increased breast milk levels of BDE-209 is provided by the limited data in Vietnam (Table 2; Tue et al., 2010). It should be also noted that studies in cows and in rats suggest that highly brominated PBDEs such as BDE-209 do not transfer to a large degree in milk (Kierkegaard et al., 2007; Huwe et al., 2008). A number of studies have examined the levels of BDE-209 in indoor air and household dust. Levels in air are usually low (Karlsson et al., 2007), though significant levels (up to 174 pg/m3) have been reported (Allen et al., 2007). Overall, air is considered to be a minor contributor of total BDE-209 exposure, though in certain circumstances it may contribute to up to 20% of intake (Karlsson et al., 2007; Allen et al., 2007). In contrast, dust may be an important source of exposure to BDE-209. Levels of BDE-209 in dust reported in different studies are summarized in Table 3. In the few cases when it was reported, BDE-209 represented a high percentage (36 to >95%) of all PBDEs in dust. Overall, levels of BDE209 in dust appear to vary by several orders of magnitude; they appear to be particularly high in dust collected in homes in the United Kingdom, possibly because of fire laws in that country (Harrad et al., 2008a). Furthermore, high BDE-209 levels were found in dust collected in cars, both in the U.S. and the U.K. (Lagalante et al., 2009; Batterman et al., 2009; Harrad et al., 2008b). It has been estimated that dust may contribute to 36–93% of total BDE-209 intake in and adult (with mean and high dust ingestion, respectively), and 88–97% intake in a child (Allen et al., 2007). Clothes dryer lint has also be reported to contain BDE-209. In two studies in the United States levels were 58–2890 ng/g (n = 5) (Stapleton et al., 2005) and 56–2149 ng/g (n = 12) (Schecter et al., 2009). Comparable levels (17–2035 ng/g; n = 7) were also found in Germany (Schecter et al., 2009).

Table 2 BDE-209 in breast milk (ng/g of lipid). Country

Yeara

N

Mean

USA USA Norway Norway Faroe Islands Faroe Islands Faroe Islands France Spain Spain Germany Russia Russia China China Taiwan Taiwan Philippines Vietnam Vietnam Indonesia Japan Japan Australia

2002 2003 2000–2002 2003–2008 1987 1994–1995 1999 2004–2006 2003–2004 2003–2004 2001–2003 2003–2004 2002 2004 2007 2000–2001 2007–2008 2004 2007 2007 2001–2003

47 40 10 46 10 10 10 77 30 22 62 10 23 19 15 0.27 46 33 28b 5c 30 89 8 10

0.92 0.80 0.22 0.61 0.6 0.5 1.2 1.9 2.8 2.9 0.17
2005–2006 2007–2008

LOD = limit of detection; LOQ = limit of quantification. a Year of milk collection. b Means of samples from 4 different geographical areas. c Milk from e-waste recycling workers.

0.47 1.7
Median 0.43 0.13

1.6

0.1 0.19 2.6

0.45 1.7

<1.3

Range

Reference


Schecter et al., 2003 She et al., 2007 Polder et al., 2008a Thomsen et al., 2010 Fangstrom et al., 2005a Fangstrom et al., 2005a Fangstrom et al., 2005a Antignac et al., 2009 Gomara et al., 2007 Gomara et al., 2007 Vieth et al., 2004 Tsydenova et al., 2007 Polder et al., 2008b Sudaryanto et al., 2008a Jin et al., 2009 Chao et al., 2007 Chao et al., 2010 Malarvannan et al., 2009 Tue et al., 2010 Tue et al., 2010 Sudaryanto et al., 2008b Inoue et al., 2006 Kawashiro et al., 2008 Toms et al., 2009

<0.5–3.4 0.4–6.8 <0.16–33 <0.16–52
<1.3–1.9
L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

18 Table 3 Levels of BDE-209 (ng/g) in household dust. Country

n

USA USA USA USA USA USA USA USA USA USA USA USA USA

9 17 50 10 17 20 20 18 9 16 60 12 10 50 68 7 5 2 1 1 1 2

Mean

Median 2000 8567 2090

1600

11,000 200,000 15,000,000 272,119 2810 6930 1909 1100 670 470

665 1350 2000 1300 4502 1703 190 7 3100 48,100 903 1

Range

%

143–65,777 162–8750
Notes Houses in Dallas 36

Houses in Texas Living rooms Bedrooms Houses Garages Cars Used cars

4380–3,570,000 327–9210

USA USA Canada Canada Sweden Denmark Denmark Finland Finland Germany Germany Germany UK UK UK UK UK UK UK UK

10 10 9 10 16 30 18 20 43

Austria Belgium Netherlands Italy Kuwait Singapore China

2 19 4 2 17 31 39

129 2200 5672

83 1000 988

68–13,000 105–140,000

e-Waste area

China

27

6298

4039

498–40,500

Urban area

Australia Global

10 20

23–13,000
Aircraft cabins

630 560

Offices in Michigan


100

260 1100 100

69 78 290–1500

Parliament

980 9820

63 7100

45,000 260,000 30,000 410,000 8500

10,000 2800 8100 6200 100,000 5000

13.8

11.1

730 18,000

<6–410 3800–19,900 1401–54,795 910–54,000 120–520,000
94

Homes Offices Used cars Schools and daycare centers Parliament 45–93 Parliament Parliament 86

Reference Sjodin et al., 2004 Schecter et al., 2005b Stapleton et al., 2005 Wu et al., 2007 Sjodin et al., 2008 Harrad et al., 2008a Allen et al., 2008 Allen et al., 2008 Batterman et al., 2009 Batterman et al., 2009 Batterman et al., 2009 Lagalante et al., 2009 Johnson-Restrepo and Kannan, 2009 Batterman et al., 2010 P.I. Johnson et al., 2010 Wilford et al., 2005 Harrad et al., 2008a Karlsson et al., 2007 Santillo et al., 2001 Santillo et al., 2003 Santillo et al., 2001 Santillo et al., 2003 Santillo et al., 2001 Knoth et al., 2003 Sjodin et al., 2008 Santillo et al., 2003 Pless-Mulloli et al., 2006 Sjodin et al., 2008 Harrad et al., 2008a Harrad et al., 2008b Harrad et al., 2008b Harrad et al., 2008b Harrad et al., 2010 Santillo et al., 2001 Roosens et al., 2009 Santillo et al., 2001 Santillo et al., 2001 Gevao et al., 2006 Tan et al., 2007 J. Wang et al., 2010a, 2010b; F. Wang et al., 2010 J. Wang et al., 2010a, 2010b; F. Wang et al., 2010 Sjodin et al., 2008 Christiansson et al., 2008

LOD = limit of detection.

P Body burden of PBDE in the general population of the United States in the range of 30–100 ng/g lipid, with BDE-47 being the most prominent congener (50%), while levels outside the U.S. are generally around 10 ng/g lw (USEPA, 2010). Higher levels of PBDEs are consistently found in infants and toddlers, possibly due to exposure through breast milk and house dust. Table 4 shows a compilation of blood BDE-209 levels from studies around the world. With few exceptions, BDE-209 blood levels appear to be low (1–5 ng/g lipid), and higher levels in the U.S. compared to other countries are not apparent. Few available data indicate that blood BDE-209 levels may be higher in toddlers and children, most likely because of exposure to dust. Fetal exposure appears to be low, as suggested by low BDE-209 levels in umbilical cord blood (Table 4). When maternal and umbilical cord BDE-209 levels were pairmeasured, levels in the latter were equal or lower to those found in the mothers (Gomara et al., 2007; Kawashiro et al., 2008; Wu et al., 2010).

6. Derivation of an RfD for BDE-209 Over the years, various agencies and organizations have proposed reference doses (RfD) for BDE-209 (Table 5). The National Research Council of the National Academy of Sciences (NAS, 2000) based its RfD of 4.0 mg/kg/day on the chronic National Toxicology Program (NTP) study (NTP, 1986), by applying a 300 uncertainty factor (UF; 10 for interspecies differences, 10 for intraspecies differences, 3 for database deficiencies) to the NOAEL (no-observed adverse effect level) of 1120 mg/kg/day for liver damage. Confidence in this RfD was considered medium to low (NAS, 2000). Hardy et al. (2009), by applying benchmark dose calculation to the results of the same study, proposed an identical RfD of 4.0 mg/kg/day. The BMDL10 for liver damage was 419 mg/kg/day, to which a 100 UF was applied (Hardy et al., 2009). ATSDR (Agency for Toxic Substances and Disease Registry) utilized the developmental study of Hardy

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

19

Table 4 Body burden of BDE-209 (blood levels, ng/g of lipid). Country

Year

Age (sex)

USA USA USA USA USA USA USA

2003 2003 2004 2004 2004 2004 2006–2007

28 10 1 1 1 1 20

3.5

USA

2006–2007

Adults (M/F) Adults (M/F) Toddler (M) Child (F) Adult (M) Adult (F) Children 1.5–4 y.o. Adult (F)

20

USA

2003–2005

94

USA Mexico

2006

Nicaragua Nicaragua

2002 2002

Children 2–5 y.o. (M/F) Adult (M/F) Children 6–13 y.o. (M/F) Children Children

Nicaragua Sweden Sweden Sweden

2002

Sweden Sweden Sweden

2000 2000, 2002 2000, 2002

Sweden Denmark Denmark Denmark Faroe Islands Faroe Islands France France France Belgium Belgium Belgium UK Spain Spain Spain Spain Spain Spain Spain Japan Japan Japan Japan Japan Japan China China

Adult Adult Adult Adult

(F) (F) (F) (M/F)

N

Mean

Median

Range

Notes

Reference

Mississippi New York City

1.7


1.7

1.5

ND–3.2

4.4

2.6

Schecter et al., 2005a Schecter et al., 2005a Fischer et al., 2006 Fischer et al., 2006 Fischer et al., 2006 Fischer et al., 2006 EWG, 2008; Lunder et al., 2010 EWG, 2008; Lunder et al., 2010 Rose et al., 2010

24 173

4.7–16 4.1–9.3

<0.7 <0.7 4.8

Adult (M) Adult (M) Adult (M)

17 11 19

2.5

2006 2007 2007 2007 2001–2002 1994–1995

Adults (M/F) Placenta Adult (F) Umbilical cord Children Adult (F)

5 50 51 40 42 57

10 1.25 1.80
2004–2006 2004–2006 2004–2006 1999–2004 1999–2004 2007 2003 2003–2004 2003–2004 2003–2004 2003–2004 2003–2004 2003–2004 2003–2004

Adult (F) Cord serum Adult (F) Cord blood Adult (F) Young adults (M/F) Adult (M/F) Adult (F) Adult (M) Umbilical cord Adult (F) Adult (M) Umbilical cord Placenta Adult Adult (M) Adult (F) Cord blood Adult (F) Adult (M/F) Adult (M/F) Adult (M/F)

27 54 7 4 (pools) 11 19 11 52 53 48 61 51 44 30 89 18 10 8 16 72 20 15

9.5 78.1 0.8

Adult (M/F)

20

China

China China China China China

2007 2007 2007 2007 2007

Adult (F) Adult (M) Umbilical cord Umbilical cord Adult (F)

China

2007

Adult (F)

6 (pools) 6 (pools) 102 51

Mothers


20 20 19

2003–2004 2003–2004 2005–2006 2005–2006 2007–2008

Father Mother

2.5–4.5 <0.3–3.9 <0.3–8.0 <0.3–9.9 0.92–9.7 1.2–250 6.9–280

1.14 1.71


1 0.77


5.8 27.1 0.7

0.8–37.4 3.4–363.3 0.1–4.4 22.5–44.5 3.6–33.1

<15 1.1 1.1 1.4 1.1 1.1 2.2 1.0

<15–49 <1.1–31 <1.1–91 <1.1–24 <1.1–20 <1.1–59 <1.1–11 <0.05–8.4

Urban Working at e-waste recycling Hospital cleaners Clerks Workers at e-dismantling plant Rubber workers Cable manufacturing workers

Mothers 7 y.o. Mothers (during pregnancy)

Adipose tissue


1.2 1.9 0.6 7 1.0

1.2 0.7 6.5 0.9 5.7 18.5

83.5

620 186 4.2 2.5 2.0 6.9

<0.5–12 <0.5–1.7

206–1640 106–316
Getafe district; mother Getafe district; father Getafe district Valleces district; mother Valleces district; father Valleces district Valleces district Adipose tissue Adipose tissue Mothers Referents Residents near e-waste dismantling facility Workers at e-waste dismantling facility

Guinyu Province Chaonan Province Mothers (not involved in e-waste recycling) Mothers (involved in e-waste recycling)

P.I. Johnson et al., 2010 Perez-Maldonado et al., 2009 Athanasiadou et al., 2008 Athanasiadou et al., 2008 Athanasiadou et al., 2008 Sjodin et al., 1999 Sjodin et al., 1999 Sjodin et al., 1999 Thuresson et al., 2005 Thuresson et al., 2005 Thuresson et al., 2005 Karlsson et al., 2007 Frederiksen et al., 2009 Fredriksen et al., 2010 Fredriksen et al., 2010 Fangstrom et al., 2005b Fangstrom et al., 2005b Antignac et al., 2009 Antignac et al., 2009 Antignac et al., 2009 Covaci and Voorspoels, 2005 Covaci and Voorspoels, 2005 Roosens et al., 2009 Thomas et al., 2006 Gomara et al., 2007 Gomara et al., 2007 Gomara et al., 2007 Gomara et al., 2007 Gomara et al., 2007 Gomara et al., 2007 Gomara et al., 2007 Inoue et al., 2006 Kunisue et al., 2007 Kunisue et al., 2007 Kawashiro et al., 2008 Kawashiro et al., 2008 Uemura et al., 2010 Qu et al., 2007 Qu et al., 2007

Qu et al., 2007

Jin et al., 2009 Jin et al., 2009 Wu et al., 2010 Wu et al., 2010 Wu et al., 2010 Wu et al., 2010

LOQ = limit of quantification.

et al. (2002), which provided a NOAEL of 1000 mg/kg/day, to derive a MRL (intermediate oral Minimal Risk Level) of 10 mg/ kg/day (ATSDR, 2004). The total UF was 100 (for inter- and intraspecies differences).

In contrast, the USEPA had initially proposed an RfD of 0.01 mg/ kg, which was based on a NOAEL of 1.0 mg/kg (the highest dose tested) from a chronic study in rats by Kociba et al. (1975), with a 100 UF. The study was carried out with three dose levels (0.01,

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

20 Table 5 Suggested reference doses for BDE-209. RfD (mg/kg/day)

NOAEL (mg/kg)

UF

Notes

Reference

0.01 4.0 10 0.007 4.0 ?

1.0 1120 1000 2.22 1120 1000

100 300 100 300 30 ?

Based on Kociba et al., 1975 Based on NTP, 1986 MRL; based on Hardy et al., 2002 Based on Viberg et al., 2003a Based on BMDL from NTP, 1986

USEPA, 1989 NAS, 2000 ATSDR, 2004 USEPA, 2008 Hardy et al., 2009 Jacobi et al., 2009; Silberberg et al., 2009

0.1, 1.0 mg/kg/day, for 2 years) of a commercial decaBDE consisting of 77.4% decabromodiphenyl oxide, 21.8% nonaBDE, and 0.8% octaBDE (Kociba et al., 1975), and examined the following parameters: hematology, clinical chemistry, food consumption, organ weight, body weight, incidence of histopathological lesions. The confidence in this RfD was considered low, because of the low doses, and low number of animals (25/sex/dose) (USEPA, 1989). More recently, the USEPA has lowered its RfD for BDE-209 to 0.007 mg/kg/day (USEPA, 2008). This value is derived from the NOAEL of 2.22 mg/kg, given as a single dose to 10 day-old mice (Viberg et al., 2003a). To this value, a 300 UF was applied to take into account interspecies differences (10), intraspecies differences (10), and acute to chronic extrapolation (3). Confidence in this value was deemed low (USEPA, 2008). This latter value (0.007 mg/kg/day) is 571–1428-fold lower than other proposed RfDs, but only 1.43-fold lower than the previous RfD proposed by USEPA 20 years earlier (USEPA, 1989). As said, the PoD (point of departure) for the latest USEPA assessment was the NOAEL of 2.22 found in the Viberg et al. (2003a) study. It should be noted, however, that the same investigators later published a similar study in which the NOAEL was 1.34 mg/kg (Johansson et al., 2008). If such study (not available at the time of the USEPA (2008) determination), were considered, the resulting RfD (applying the same 300 UF) would be even lower (0.004 mg/kg/day). Either value is about three orders of magnitude lower than those derived by others, though all proposed RfD values were deemed to provide a low degree of confidence. While the earlier USEPA RfD did not appear to generate overt controversy, this latter one did (Goodman, 2009; Hardy et al., 2009; Williams and DeSesso, 2010). The focus of the controversy is related to the notion of whether there is sufficient evidence to indicate that BDE-209 is a potential developmental neurotoxicant, and on the fact that the Viberg et al. (2003a) study may not be ideal for use in such determination (see Section 4.1 for a full discussion). Presumptions of developmental neurotoxicity would call for an additional default 10 UF; if applied, for example, to the 4.0 mg/kg/day RfD (NAS, 2000; Hardy et al., 2009), this would lead to an RfD of 0.4 mg/kg/day. However, a recent developmental neurotoxicity study, carried out according to OECD and USEPA guidelines, though only available so far in abstract form, appears to provide no evidence of developmental neurotoxicity. In this case, from a NOAEL of 1000 mg/kg/day, after applying a 100 UF, the resulting RfD would be 10 mg/kg/day. 7. Conclusions and research needs As other flame retardants, PBDEs have contributed to a significant reduction of mortality, morbidity, and property damage due to fires in the past several decades. Yet, their propensity to leach from the products and to become widespread environmental pollutants, have raised concerns in recent years (DiGangi et al., 2010). This has led to the ban of penta- and octa-BDEs in several countries, and restrictions on decaBDE (BDE-209) in others. One of the driving arguments for concern on PBDEs has been their relatively higher levels in toddlers and infants, because of exposure

through breast milk and household dust, and a number of findings in animal and in in vitro studies (and most recently also in humans; Roze et al., 2009; Herbstman et al., 2010) for potential developmental neurotoxicity (Branchi et al., 2003; Birnbaum and Staskal, 2004; McDonald, 2005; Costa and Giordano, 2007). BDE-209 has also been studied in this regard, though to a lesser extent than other lower brominated PBDEs, and evidence so far suggests that it may cause similar effects as other PBDEs, as indicated by limited in vivo and in vitro studies (see Section 4). Since the RfD for BDE-209 recently proposed by USEPA (USEPA, 2008) was based on a less than ideal developmental neurotoxicity study, the notion that BDE-209 may be a developmental neurotoxicant has been challenged (Goodman, 2009; Hardy et al., 2009; Williams and DeSesso, 2010). In support of this, a developmental neurotoxicity study carried out according to OECD and USEPA guidelines, and available so far only in abstract form (Jacobi et al., 2009; Silberberg et al., 2009), provided no evidence of adverse effects at all doses tested (1, 10, 100, 1000 mg/kg/day). Developmental neurotoxicity testing guidelines developed both in the U.S.A. (USEPA, 1998) and in Europe (OECD, 2007) involve exposure of the mother to the test chemicals from GD 6 to PND 10 or 21, thus ensuring exposure of the offspring in utero and through maternal milk. Tests involve measurements of developmental landmarks and reflexes, motor activity, auditory startle test, learning and memory tests, and neuropathology (USEPA, 1998; OECD, 2007). Developmental neurotoxicity testing has been proven to be useful and effective in identifying compounds with developmental neurotoxicity potential (Makris et al., 2009). However, this is not to say that current developmental neurotoxicity testing guidelines cannot be improved; indeed it has been pointed out that they may be overly sensitive and produce a high rate of false positives (Claudio et al., 2000), or, in contrast, that they may be too insensitive and not enough comprehensive (CorySlechta et al., 2001). Thus, albeit important, a negative guideline developmental neurotoxicity study may be judged, in a weight of evidence approach, together with all available information. As new studies become available, further evidence would shift the pendulum toward a final conclusion on whether BDE-209 is a developmental neurotoxicant. From the point of view of protecting human health, it is important that BDE-209 appears to be absorbed to a limited degree, to have a relatively short half-life, and to penetrate cells with some difficulty. It is also reassuring that exposure children may be orders of magnitude below the most conservative RfD, which would in itself make any discussion mood. Indeed, the exposure of children to P PBDEs, primarily through dust and diet, has been estimated as approximately 400–700 ng/day (8–50 ng/kg/day, depending on body weight) (USEPA, 2010). Of this, about 28% is represented by BDE-209, for an intake of 2.2–14 ng/kg/day. This value is still several order of magnitude lower than the most conservative, and contested, RfD proposed by USEPA (7000 ng/kg/day). However, other aspects of BDE-209 still warrant further investigations; in particular, the possibility that BDE-209 is transformed in the environment or in the human body to lower

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

brominated PBDE of potentially greater human toxicity (and developmental neurotoxicity) needs to be further investigated. Conflict of interest statement The authors declare that they have no conflicts of interest. Acknowledgment Research by the authors has been supported in part by a grant from NIEHS (P30ES07033). References Abdelouahab N, Suvorov A, Pasquier JC, Langlois MF, Praud JP, Takser L. Thyroid disruption by low-dose BDE-47 in prenatally exposed lambs. Neonatology 2009;96:120–4. Ahn M, Filley TR, Jafvert CT, Nies L, Hua I, Bezares-Cruz B. Photodegradation of decabromodiphenyl ether adsorbed onto clay minerals, metal oxides, and sediment. Environ Sci Technol 2006;40(22):798–804. Alaee M, Arias P, Sjodin A, Bergman A. An overview of commercially used brominated flame retardants, their applications, their use pattern in different countries/ regions, and possible modes of release. Environ Int 2003;29:683–9. Allen JG, McClean MD, Stapleton HM, Nelson JW, Webster TF. Personal exposure to polybrominated dipheyl ethers (PBDEs) in residential indoor air. Environ Sci Technol 2007;41:4574–9. Allen JG, McClean MD, Stapleton HM, Webster TF. Critical factors in assessing exposure to PBDEs via house dust. Environ Int 2008;34:1085–91. Antignac JP, Cariou R, Zalko D, Berrebi A, Cravedi JP, Maume D, et al. Exposure assessment of French women and their newborn to brominated flame retardants: determination of tri- to deca-polybromodiphenylethers (PBDE) in maternal adipose tissue, serum breast milk and cord serum. Environ Pollut 2009;157:164–73. Athanasiadou M, Cuadra SN, Marsh G, Bergman A, Jakobsson K. Polybrominated diphenyl ethers (PBDEs) and bioaccumulative hydroxylated PBDE metabolites in young humans from Managua, Nicaragua. Environ Health Perspect 2008;116:400–8. ATSDR (Agency for Toxic Substances and Disease Registry). Toxicological profile for polybrominated biphenyls and polybrominated diphenyl ethers. Atlanta, GA: Public Health Service, U.S. Department of Health and Human Services; 2004. Batterman SA, Chernyak S, Jia C, Gowin C, Charles S. Concentrations and emissions of polybrominated diphenyl ethers from U.S. houses and garages. Environ Sci Technol 2009;43:2693–700. Batterman SA, Godwin C, Chernyak S, Jia C, Charles S. Brominated flame retardants in offices in Michigan, U.S.A.. Environ Int 2010;36:548–56. Biesemeier JA, Beck MJ, Silberberg H, Myers NR, Ariano JM, Bodle ES, et al. Effects of dose, administration route an/or vehicle on decabromodiphenyl ether (decaBDE) concentrations in plasma of maternal, fetal and neonatal rats and in milk of maternal rats. Drug Metab Disp 2010;38:1648–54. Birnbaum LS, Staskal DF. Brominated flame retardants: cause for concern? Environ Health Perspect 2004;112:9–17. Branchi I, Alleva E, Costa LG. Effects of perinatal exposure to a polybrominated diphenyl ether (PBDE 99) on mouse neurobehavioral development. Neurotoxicology 2002;23:375–84. Branchi I, Capone F, Alleva E, Costa LG. Polybrominated diphenyl ethers: neurobehavioral effects following developmental exposure. Neurotoxicology 2003;24:449– 62. Bruchajzer E, Frydrych B, Sporny S, Szymanska JA. Toxicity of penta- and decabromodiphenyl ethers after repeated administration to rats: a comparative study. Arch Toxicol 2010;84:287–99. Cai Y, Zhang W, Hu J, Sheng G, Chen D, Fu J. Characterization of maternal transfer of decabromodiphenyl ether (BDE-209) administered to pregnant Sprague–Dawley rats. Reprod Toxicol; in press, doi:10.1016/j/reprotox.2010.08.05. Chan S, Rovet J. Thyroid hormones in the fetal central nervous system development. Fetal Matern Med Rev 2003;14:177–208. Chao HA, Chen SCC, Chang CM, Koh TW, Chang-Chien GP, Ouyang E, et al. Concentrations of polybrominated diphenyl ethers in breast milk correlated to maternal age, education level, and occupational exposure. J Hazard Mater 2010;175:492–500. Chao HR, Wang SL, Lee WJ, Wang YF, Papke O. Levels of polybrominated diphenyl ethers (PBDEs) in breast milk from central Taiwan and their relation to infant birth outcome and maternal menstruation effects. Environ Int 2007;33:239–45. Charman WN. Lipids, lipophilic drugs and oral delivery—some emerging concepts. J Pharm Sci 2000;89:967–78. Chen D, Hale RC. A global review of polybrominated flame retardant contamination in birds. Environ Int 2010;36:800–11. Chen J, Liufu C, Sun W, Sun X, Chen D. Assessment of the neurotoxic mechanisms of decabrominated diphenyl ether (PBDE-209) in primary cultured neonatal rat hippocampal neurons includes alterations in second messenger signaling and oxidative stress. Toxicol Lett 2010;192:431–9. Cheng J, Gu J, Ma J, Chen X, Zhang M, Wang W. Neurobehavioral effects, redox responses and tissue distribution in rat offspring developmental exposure to BDE-99. Chemosphere 2009;75:963–8.

21

Christiansson A, Hovander L, Atahnassiadis I, Jakobsson K, Bergman A. Polybrominated diphenyl ethers in aircraft cabins – a source of human exposure? Chemosphere 2008;73:1654–60. Christiansson A, Eriksson J, Teclechiel D, Bergman A. Identification and quantification of products formed via photolysis of decabromodiphenyl ether. Environ Sci Pollut Res Int 2009;16:312–21. Claudio L, Kwa WC, Russell AL, Wallinga D. Testing methods for developmental neurotoxicity of environmental chemicals. Toxicol Appl Pharmacol 2000;164:1– 14. Coburn CG, Curraz-Collazo MC, Kodavanti PRS. In vitro effects of environmentally relevant polybrominated diphenyl ether (PBDE) congeners on calcium buffering mechanisms in the rat. Neurochem Res 2008;3:355–64. Cory-Slechta DA, Crofton KM, Foran JA, Ross JF, Sheets LP, Weiss B, et al. Methods to identify and characterize developmental neurotoxicity for human health risk assessment, I: behavioral effects. Environ Health Perspect 2001;109(Suppl. 1):79–91. Costa LG, Giordano G. Developmental neurotoxicity of polybrominated diphenyl ether (PBDE) flame retardants. Neurotoxicology 2007;28:1047–67. Costa LG, Giordano G, Tagliaferri S, Caglieri A, Mutti A. Polybrominated diphenyl ether (PBDE) flame retardants: environmental contamination, human body burden and potential adverse effects. Acta Biomed 2008;79:172–83. Costa LG, Huang S, Giordano G. Oxidative stress as a potential mechanism for developmental neurotoxicity of polybrominated diphenyl ether (PBDE) flame retardants. In: Merlani PB, editor. Flame retardants: functions, properties and safety. Happanauge, NY: Nova Publishers; 2010, pp. 101–15. Covaci A, Voorspoels S. Optimization of the determination of polybrominated diphenyl ethers in human serum using solid-phase extraction and gas chromatography– electron capture negative ionization mass spectrometry. J Chromatogr B 2005;827:216–23. deWit CA. An overview of brominated flame retardants in the environment. Chemosphere 2002;46:583–624. DiGangi J, Blum A, Bergman A, deWit C, Lucas D, Mortimer D, et al. San Antonio statement on brominated and chlorinated flame retardants. Environ Health Perspect 2010;118:A516–8. Dingemans MML, de Groot A, van Kleef RGDM, Bergman A, van der Berg M, Vijverberg PM, et al. Hydroxylation increases the neurotoxic potential of BDE-47 to affect exocytosis ad calcium homeostasis in PC12 cells. Environ Health Perspect 2008;116:637–43. Dingemans MML, Heusinkveld HJ, Bergman A, van den Berg M, Westerink RHS. Bromination pattern of hydroxylated metabolites of BDE-47 affects their potency to release calcium from intracellular stores in PC12 cells. Environ Health Perspect 2010;118:519–25. Driscoll LL, Gibson AM, Hieb A. Chronic postnatal DE-71 exposure: effects on learning, attention and tyroxine levels. Neurotoxicol Teratol 2009;31:76–84. El Dareer SM, Kalin JR, Tillery KF, Hill DL. Disposition of decabromobiphenyl ether in rats dosed intravenously or by feeding. J Toxicol Environ Health 1987;22:405–15. Ellis-Hutchings RG, Cherr GN, Hanna LA, Keen CL. Polybrominated diphenyl ether (PBDE)-induced alterations in vitamin A and thyroid hormone concentrations in the rat during lactation and early postnatal development. Toxicol Appl Pharmacol 2006;215:135–45. Eriksson P. Response to: use of the pup as the statistical unit in developmental neurotoxicity studies; overlooked model or poor research design? Toxicol Sci 2008;103:411–3. Eriksson P, Viberg H. Letter to the Editor. Toxicol Sci 2004;79:205–6. Eriksson P, Jakobson E, Fredriksson A. Brominated flame retardants: a novel class of developmental neurotoxicants in our environment? Environ Health Perspect 2001;109:903–8. Eriksson P, von Rosen D, Viberg H, Fredriksson A. Developmental toxicology in the neonatal mouse: the use of randomly selected individuals as statistical unit compared to the litter in mice neonatally exposed to PBDE 99. Toxicol Sci 2005;84(Suppl. 1):219. EWG (Environmental Working Group). Fire retardants in toddlers and their mothers., 2008 p. 28 In: http://www.ewg.org. Fangstrom B, Strid A, Grandjean P, Weihe P, Bergman A. A retrospective study of PBDEs and PCBs in human milk from the Faroe Islands. Environ Health 2005a;4:12. Fangstrom B, Hovander L, Bignert A, Athanassiadis I, Linderholm L, Grandjean P, et al. Concentrations of polybrominated diphenyl ethers, polybrominated biphenyls, and polychlorobiphenols in serum from pregnant Faroese women and their children 7 years later. Environ Sci Technol 2005b;39:9547–63. Fischer D, Hooper K, Athanasiadou M, Athanassiadis I, Bergman A. Children show highest levels of polybrominated diphenyl ethers in a California family of four: a case study. Environ Health Perspect 2006;114:1581–4. Fonnum F, Mariussen E. Mechanisms involved in the neurotoxic effects of environmental toxicants such as polychlorinated biphenyls and brominated flame retardants. J Neurochem 2009;111:1327–47. Frederiksen M, Thomsen M, Vorkamp K, Knudsen LE. Patterns and concentration levels of polybrominated diphenyl ethers (PBDEs) in placental tissue of women in Denmark. Chemosphere 2009;76:1464–9. Frederiksen M, Vorkamp K, Mathiesen L, Mose T, Knudsen LE. Placental transfer of the polybrominated diphenyl ethers BDE-47, BDE-99 and BDE-209 in a human placenta perfusion system: an experimental study. Environ Health 2010;9:32. Fredriksen M, Thomsen C, Froshaug M, Vorkamp K, Thomsen M, Becher G, et al. Polybrominated diphenyl ethers in paired samples of maternal and umbilical cord blood plasma and associations with house dust in a Danish cohort. Int J Hyg Environ Health 2010;213:233–42.

22

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

Fujimoto H, Woo GH, Inoue K, Takahashi M, Hirose M, Nishikawa A, et al. Impaired oligodendroglial development by decabromodiphenyl ether in rat offspring after maternal exposure from mid-gestation through lactation. Reprod Toxicol; in press, doi:10.1016/j/reprotox.2010.09.003. Furst P. Dioxins polychlorinated biphenyls and other organohalogen compounds in human milk. Mol Nutr Food Res 2006;50:922–33. Gee JR, Moser VC. Acute postnatal exposure to brominated diphenylether 47 delays neuromotor ontogeny and alters motor activity in mice. Neurotoxicol Teratol 2008;30:79–87. Gee JR, Hedge JM, Moser VC. Lack of alterations in thyroid hormones following exposure to polybrominated diphenylether 47 during a period of rapid brain development in mice. Drug Chem Toxicol 2008;31:245–54. Geyer HJ, Schramm KW, Darnerud PO, Aune M, Feicht A, Fried KW, et al. Terminal elimination and half-lives of the brominated flame retardants TBBPA, HBCD, and lower brominated PBDEs in humans. Organohalog Compd 2004;66:3867–72. Gevao B, Al-Bahloul M, Al-Ghadban AN, Al-Omair A, Ali L, Zafar J, et al. House dust as a source of human exposure to polybrominated diphenyl ethers in Kuwait. Chemosphere 2006;64:603–8. Gill U, Chu I, Ryan JJ, Feelry M. Polybrominated diphenyl ethers: human tissue levels and toxicology. Rev Environ Contamin Toxicol 2004;183:55–97. Giordano G, Kavanagh TJ, Costa LG. Neurotoxicity of a polybrominated diphenyl ether mixture (DE-71) in mouse neurons and astrocytes is modulated by intracellular glutathione levels. Toxicol Appl Pharmacol 2008;232:161–8. Gomara B, Herrero L, Gonzalez MJ. Survey of polybrominated diphenyl ether levels in Spanish commercial foodstuffs. Environ Sci Technol 2006;40:7541–7. Gomara B, Herrero L, Ramos JJ, Mateo JR, Fernandez MA, Garcia JF, et al. Distribution of polybrominated diphenyl ethers in human umbilical cord serum, paternal serum, maternal serum, placentas, and breast milk from Madrid population, Spain. Environ Sci Technol 2007;41:6961–8. Goodman JE. Neurodevelopmental effects of decabromodiphenyl ether (BDE-209) and implications for the reference dose. Regul Toxicol Pharmacol 2009;54:91–104. Guo JY, Wu FC, Mai BX, Luo XJ, Zeng EY. Polybrominated diphenyl ethers in seafood products of South China. J Agric Food Chem 2007;55:9152–8. Hakk H, Letcher RJ. Metabolism and toxicokinetics and fate of brominated flame retardants—a review. Environ Int 2003;29:801–28. Hardy ML. The toxicology of the three commercial polybrominated diphenyl oxide (ether) flame retardants. Chemosphere 2002;46:757–77. Hardy ML, Schroeder R, Biesemeier J, Manor O. Prenatal oral (gavage) developmental toxicity study of decabromodiphenyl oxide in rats. Int J Toxicol 2002;21:83–91. Hardy ML, Stedeford T. Use of the pup as the statistical unit in developmental neurotoxicity studies: overlooked model or poor research design? Toxicol Sci 2008;103:409–10. Hardy ML, Banasik M, Stedeford T. Toxicology and human health assessment of decabromodiphenyl ether. Crit Rev Toxicol 2009;39(S3):1–44. Harrad S, Ibarra C, Diamond M, Melymuk L, Robson M, Douwes J, et al. Polybrominated diphenyl ethers in domestic indoor dust from Canada, New Zealand, United Kingdom and United States. Environ Int 2008a;34:232–8. Harrad S, Ibarra C, Abdallah MAE, Boon R, Neels H, Covaci A. Concentration of brominated flame retardants in dust from United Kingdom cars, homes, and offices: causes of variability and implications for human exposure. Environ Int 2008b;34:1170–5. Harrad S, Goosey E, Desborough J, Abdallah MAE, Roosens L, Covaci A. Dust from U.K. primary school classrooms and daycare centers: the significance of dust as a pathway of exposure of young U.K. children to brominated flame retardants and polychlorinated biphenyls. Environ Sci Technol 2010;44:4198–202. He J, Robrock KR, Alvarez-Cohen L. Microbial reductive debromination of polybrominated diphenyl ethers (PBDEs). Environ Sci Technol 2006;40:4429–34. He P, He W, Wang A, Xia T, Xu B, Zhang M, et al. PBDE-47-induced oxidative stress, DNA damage and apoptosis in primary cultured rat hippocampal neurons. Neurotoxicology 2008;29:124–9. Herbstman JB, Sjodin A, Apelberg BJ, Witter FR, Patterson DG, Halden RU, et al. Determinants of prenatal exposure to polychlorinated biphenyls (PCBs) and polybrominated diphenyl ethers (PBDEs) in an urban population. Environ Health Perspect 2007;115:1794–800. Herbstman JB, Sjodin A, Aperlberg BJ, Witter FR, Halden RU, Patterson DG, et al. Birth delivery mode modifies the association between prenatal polychlorinated biphenyl (PCB) and polybrominated diphenyl ether (PBDE) and neonatal thyroid hormone levels. Environ Health Perspect 2008;116:1376–82. Herbstman JB, Sjodin A, Kurzon M, Lederman SA, Jones RS, Rauh V, et al. Prenatal exposure to PBDEs and neurodevelopment. Environ Health Perspect 2010;118:712–9. Hertz-Picciotto I. A case control investigation of PBDEs and autism in children. Epidemiology 2008;19:S75. Hites RA, Foran JA, Schwager SJ, Knuth BA, Hamilton MC, Carpenter DO. Global assessment of polybrominated diphenyl ethers in farmed and wild salmon. Environ Sci Technol 2004;38:4945–9. Holson RR, Freshwater L, Maurissen JPJ, Moser VC, Phang W. Statistical issues and techniques appropriate for developmental neurotoxicity testing. A report from the ILSI Research Foundation/Risk Science Institute expert working group on neurodevelopmental endpoints. Neurotoxicol Teratol 2008;30:326–48. Hu XZ, Xu Y, Hu Dc. Hui Y, Yang FX. Apoptosis induction on human hepatoma cells HepG2 of decabrominated diphenyl ether (PBDE-209). Toxicol Lett 2007;171:19– 28. Huang S, Giordano G, Costa LG. Comparative cytotoxicity and intra-cellular accumulation of five polybrominated diphenyl ether (PBDE) congeners in mouse cerebellar granule neurons. Toxicol Sci 2010;114:124–32.

Hughes MF, Edwards BC, Mitchell CT, Bhooshan B. In vitro dermal absorption of flame retardant chemicals. Food Chem Toxicol 2001;39:1263–70. Huwe JK, Smith DJ. Accumulation, whole-body depletion and debromination of decabromodiphenyl ether in male Sprague–Dawley rats following dietary exposure. Environ Sci Technol 2007;41:2371–7. Huwe JK, Hakk H, Birnbaum LS. Tissue distribution of polybrominated diphenyl ethers in male rats and implications for biomonitoring. Environ Sci Technol 2008;42:7018–24. IARC (International Agency for Research on Cancer). Some flame retardants and textile chemicals, and exposures in the textile manufacturing industry. Summary of data reported and evaluation. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans 1990;48:73–84. Ibhazehiebo K, Iwasaki T, Kimura-Koroda J, Miyazaki W, Shimokawa N, Koibuchi N. Disruption of thyroid hormone receptor-mediated transcription and thyroid hormone-induced Purkinje cell dendrite arborization by polybrominated diphenylethers. Environ Health Perspect; in press, doi:10.1289/ehp.1002065. Illinois EPA (Environmental Protection Agency). A Report to the General Assembly and the Governor in Response to Public Act 94-100. ‘‘DecaBDE study: a review of available scientific research’’. Illinois EPA; January 2006 p. 49. Inoue K, Harada K, Takenaka K, Uehara S, Kono M, Shimizu T, et al. Levels and concentration ratios of polychlorinated biphenyls and polybrominated diphenyl ethers in serum and breast milk in Japanese mothers. Environ Health Perspect 2006;114:1179–85. Jacobi S, Silberberg H, Stedeford T, Biesemeier J, Adriano J, Beck M. An oral (gavage) developmental neurotoxicity study of decabromodiphenyl oxide (DeBDPO) in rats. Toxicol Lett (Suppl. 1):2009;S260. Jiang HP, Yu YH, Chen DJ, Wu Y, Xu B. Effect of maternal BDE-209 exposure on the expression of GAP-43 and BDNF in the hippocampus of the offspring rats. Nan Fang Yi Ke Da Xue Xue Bao 2008;28:1319–22. Jin J, Wang Y, Yang C, Hu J, Liu W, Cui J, et al. Polybrominated diphenyl ethers in the serum and breast milk of the resident population from production area, China. Environ Int 2009;35:1048–52. Jin S, Yang F, Hui Y, Xu Y, Lu Y, Liu J. Cytotoxicity and apoptosis induction on RTG-2 cells of 2,20 ,4,40 -tetrabromodiphenyl ether (BDE-47) and decabrominated diphenyl ether (BDE-209). Toxicol In Vitro 2010;24:1190–6. Johansson N, Viberg H, Fredriksson A, Erksson P. Neonatal exposure to deca-brominated diphenyl ether (BDE 209) causes dose–response changes in spontaneous behavior and cholinergic susceptibility in adult mice. Neurotoxicology 2008;29:911–9. Johnson GT, Zhu Y, Harbison RD. Comment on: effects of decabrominated diphenyl ether (PBDE 209) exposure at different developmental periods on synaptic plasticity in the dentate gyrus of adult rats in vivo. Toxicol Sci 2010;114:387–8. Johnson PI, Stapleton HM, Sjodin A, Meeker JD. Relationship between polybrominated diphenyl ether concentrations in house dust and serum. Environ Sci Technol 2010;44:5627–32. Johnson-Restrepo B, Kannan K. An assessment of sources and pathways of human exposure to polybrominated diphenyl ethers in the United States. Chemosphere 2009;76:542–8. Karlsson M, Julander A, van Bavel B, Hardell L. Levels of brominated flame retardants in blood in relation to levels in household air and dust. Environ Int 2007;33:62–9. Kawashiro Y, Fukata H, Omori-Inoue M, Kubonoya K, Jotaki T, Takigami H, et al. Perinatal exposure to brominated flame retardants and polychlorinated biphenyls in Japan. Endocr J 2008;55:1071–84. Kawashiro Y, Fukata H, Sato K, Aburatani H, Takigami H, Mori C. Polybrominated diphenyl ethers cause oxidative stress in human umbilical vein endothelial cells. Hum Exp Toxicol 2009;28:703–13. Kierkegaard A, Asplund L, de Wit CA, McLachlan MS. Fate of higher brominated PBDEs in lactating cows. Environ Sci Technol 2007;41:417–23. Kim TH, Lee YJ, Lee E, Kim MS, Kwack SJ, Kim KB, et al. Effects of gestational exposure to decabromodiphenyl ether on reproductive parameters, thyroid hormone levels, and neuronal development in Sprague Dawley rats offspring. J Toxicol Environ Health A 2009;72:1296–303. Knoth W, Mann W, Meyer R, Nebhuth J. Brominated diphenyl ether in indoor dust. Organohalog Compd 2003;61:207–10. Kociba RJ, Frauson LO, Humiston GG, Norris JM, Wade CE, Lisowe RW, et al. Results of a two-year dietary feeding study with decabromodiphenyl oxide (DBDPO) in rats. JFF/Combust Toxicol 1975;2:267–85. Kodavanti PR, Ward TR. Differential effects of commercial polybrominated diphenyl ethers and polychlorinated biphenyl mixtures on intracellular signaling in rat brain in vitro. Toxicol Sci 2005;85:952–62. Kodavanti PR, Coburn CG, Moser VC, MacPhail RC, Fenton SE, Stoker TE, et al. Developmental exposure to a commercial PBDE mixture, DE-71: neurobehavioral, hormonal, and reproductive effects. Toxicol Sci 2010;116:297–312. Kunisue T, Takayanagi N, Isobe T, Takahashi S, Nose M, Yamada T, et al. Polybrominated diphenyl ethers and persistent organochlorine in Japanese human adipose tissues. Environ Int 2007;33:1048–56. Kuriyama SN, Wanner A, Fidalgo-Neto AA, Taslness CE, Koerner W, Chahoud I. Developmental exposure to low-dose PBDE-99: tissue distribution and thyroid hormone levels. Toxicology 2007;242:80–90. Lagalante AF, Oswald TD, Calvosa FC. Polybrominated diphenyl ether (PBDE) levels in dust from previously owned automobiles at United States dealerships. Environ Int 2009;35:539–44. La Guardia MJ, Hale RC, Harvey E. Evidence of debromination of decabromodiphenyl ether (BDE-209) in biota from a wastewater receiving stream. Environ Sci Technol 2007;41:6663–70.

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24 Law RJ, Allchin CR, deBoer J, Covaci A, Herzke D, Lepom P, et al. Levels and trends of brominated flame retardants in the European environment. Chemosphere 2006;64:187–208. Lee E, Kim TH, Choi JS, Nabanata P, Kim NY, Ahn MY, et al. Evaluation of liver and thyroid toxicity in Sprague–Dawley rats after exposure to polybrominated diphenyl ether BDE-209. J Toxicol Sci 2010;35:535–45. Lorber M. Exposure of Americans to polybrominated diphenyl ethers. J Expos Sci Environ Epidemiol 2007;1–8. Lunder S, Hovander L, Athanassiadis I, Bergman A. Significantly higher polybrominated diphenyl ether levels in young U.S. children than in their mothers. Environ Sci Technol 2010;44:5256–62. Madia F, Giordano G, Fattori V, Vitalone A, Branchi I, Capone F, et al. Differential in vitro neurotoxicity of the flame retardant PBDE-99 and of the PCB Aroclor 1254 in human astrocytoma cells. Toxicol Lett 2004;154:11–21. Makris SL, Raffaele K, Allen S, Bowers WJ, Hass U, Alleva E, et al. A retrospective performance assessment of the developmental neurotoxicity study in support of OECD test guideline 426. Environ Health Perspect 2009;117:17–25. Malarvannan G, Kunisue T, Isobe T, Sudaryanto A, Takahashi S, Prudente M, et al. Organohalogen compounds in human breast milk from mothers living in Payatas and Malate, the Philippines: levels, accumulation kinetics and infant health risks. Environ Pollut 2009;157:1924–32. Mazdai A, Dodder NG, Abernathy MP, Hites RA, Bigsby RM. Polybrominated diphenyl ethers in maternal and fetal blood samples. Environ Health Perspect 2003;11:1249–52. McDonald TA. Polybrominated diphenylether levels among United States residents: daily intake and risk of harm to the developing brain and reproductive organs. Integr Environ Assess Manag 2005;1:343–54. Messer A. Mini-review: polybrominated diphenyl ether (PBDE) flame retardants as potential autism risk factors. Physiol Behav 2010;100:245–9. Morck A, Hakk H, Orn U, Klasson Wehler E. Decabromodiphenyl ether in the rat: absorption, distribution, metabolism, and excretion. Drug Metab Dispos 2003;31:900–7. Moser VC, Walls I, Zoetis T. Direct dosing of preweaning rodents in toxicity testing and research: deliberations of an ILSI RSI expert working group. Int J Toxicol 2005;24:87–94. NAS (National Academy of Sciences). Toxicological risks of selected flame retardant chemicals. National Research Council, Commission on Life Sciences, Board of Environmental Studies and Technology, Committee on Toxicology, Sub-committee on Flame-retardant Chemicals. Washington, DC: National Academy Press; 2000. Newland MC, Rasmussen EB. Aging unmasks adverse effects of gestational exposure to methylmercury in rats. Neurotoxicol Teratol 2000;22:819–28. Norris JM, Kociba RJ, Schwetz BA, Rose JQ, Humiston CG, Jewett GL, et al. Toxicology of octabromobiphenyl and decabromodiphenyl oxide. Environ Health Perspect 1975;11:153–61. NTP (National Toxicology Program). Toxicology and carcinogenicity studies of decabromodiphenyl oxide (CAS No. 1163-19-5) in F344/N rats and B6C3F mice (feed studies). Research Triangle Park: NTP; 1986 Technical Report Series No. 309. OECD (Organization of Economic Co-operation Development). OECD guidelines for the testing of chemicals. Developmental neurotoxicity study. OECD/OCDE 426. 2007. Pacyniak EK, Cheng X, Cunningham ML, Crofton K, Klaassen CD, Guo GL. The flame retardants, polybrominated diphenyl ethers, are pregnane X receptor activators. Toxicol Sci 2007;97:94–102. Perez-Maldonado IN, Ramirez-Jimenez MDR, Martinez-Arevalo LP, Lopez-Guzman OD, Athanasiadou M, Bergman A, et al. Exposure assessment of polybrominated diphenyl ethers (PBDEs) in Mexican children. Chemosphere 2009;75:1215–20. Pless-Mulloli T, Schecter A, Schilling B, Papke O. Levels of PBDE in household dust and lint in the UK, Germany and the USA. Organohalog Compd 2006;68:495–8. Polder A, Thomsen C, Lindstrom G, Loken KB, Skaare JU. Levels and temporal trends of chlorinated pesticides, polychlorinated biphenyls and brominated flame retardants in individual human breast milk samples from Northern and Southern Norway. Chemosphere 2008a;73:14–23. Polder A, Gabrielsen GW, Odland JO, Savinova TN, Tkachev A, Loken KB, et al. Spatial and temporal changes of chlorinated pesticides. PCBs, dioxins (PCDDs/PCDFs) and brominated flame retardants in human breast milk from Northern Russia. Sci Total Environ 2008b;391:41–54. Qin X, Xia X, Yang Z, Yan S, Zhao Y, Wei R, et al. Thyroid disruption by technical decabromodiphenyl ether (DE-83R) at low concentrations in Xenopus laevis. J Environ Sci 2010;22:744–51. Qu W, Bi X, Sheng G, Lu S, Fu J, Yuan J, et al. Exposure to polybrominated diphenyl ethers among workers at an electronic waste dismantling region in Guangdong, China. Environ Int 2007;33:1029–34. Rice DC. Animal models of cognitive impairment produced by developmental lead exposure. In: Levin ED, Buccafusco JJ, editors. Animal models of cognitive impairment. Boca Raton, FL: CRC Press; 2006, pp. 73–100. Rice DC, Reeve EA, Herlihy A, Zoeller RT, Thompson WD, Markowski VP. Developmental delays and locomotor activity in the C57BL6/J mouse following neonatal exposure to the fully-brominated PBDE, decabromodiphenyl ether. Neurotoxicol Teratol 2007;29:511–20. Rice DC, Thompson WD, Reeve EA, Onos KD, Assadollahzahed M, Markowski VP. Behavioral changes in aging but not young mice after neonatal exposure to the polybrominated flame retardant decaBDE. Environ Health Perspect 2009;117:1903–11. Richardson VM, Staskal DF, Ross DG, Diliberto JJ, DeVito MJ, Birnbaum LS. Possible mechanisms of thyroid disruption in mice by BDE 47, a major polybrominated diphenyl ether congener. Toxicol Appl Pharmacol 2008;226:244–50.

23

Riu A, Cravedi JP, Debrauwer L, Garcia A, Canlet C, Jouanin I, et al. Disposition and metabolic profiling of [14C]-decabromodiphenyl ether in pregnant Wistar rats. Environ Int 2008;34:318–29. Roosens L, Abdallah MA, Harrad S, Neels H, Covaci A. Factors influencing concentrations of polybrominated diphenyl ethers (PBDEs) in students from Antwerp, Belgium. Environ Sci Technol 2009;43:3535–41. Rose M, Bennet DH, Bergman A, Fangstrom B, Pessah IN, Hertz-Picciotto I. PBDEs in 2–5 year-old children from California and association with diet and indoor environment. Environ Sci Technol 2010;44:2648–53. Roze E, Meijer L, Bakker A, Van Braeckel KNJA, Sauer PJJ, Bos AF. Prenatal exposure to organohalogens, including brominated flame retardants, influence motor, cognitive, and behavioral performance at school age. Environ Health Perspect 2009;117:1953–8. Sandholm A, Emanuelsson BM, Klasson Wehler E. Bioavailability and half-life of decabromodiphenyl ether (BDE-209) in rat. Xenobiotica 2003;33:1149–58. Santillo D, Johnston P, Bridgen K. The presence of brominated flame retardants and organotin compounds in dust collected from parliament buildings from eight countries. Greenpeace Research Laboratories Report No. GRL-TN-03-2001. UK: Exeter; 2001 p. 24. Santillo D, Labunska I, Davidson H, Johnston P, Strutt M, Knowles O. Consuming chemicals: hazardous chemicals in house dust as an indicator of chemical exposure in the home. Greenpeace Research Laboratories Report No. GRL-TN-01-2003. UK: Exeter; 2003 p. 74. Schalock RL, Brown WJ, Smith RL. Neonatal hypothyroidism: behavioral, thyroid hormonal and neuroanatomical effects. Physiol Behav 1977;19:489–91. Schecter A, Pavuk M, Papke O, Ryan JJ, Birnbaum L, Rosen R. Polybrominated diphenyl ethers (PBDEs) in U.S. mothers’ milk. Environ Health Perspect 2003;111:1723–9. Schecter A, Papke O, Tung KC, Joseph J, Harris TR, Dahlgren J. Polybrominated diphenyl ether flame retardants in the U.S. population: current levels, temporal trends, and comparison with dioxins, dibenzofurans, and polychlorinated biphenyls. J Occup Environ Med 2005a;47:199–211. Schecter A, Papke O, Joseph JE, Tung KC. Polybrominated diphenyl ethers (PBDEs) in U.S. computers and domestic carper vacuuming: possible sources of human exposure. J Toxicol Environ Health A 2005b;68:501–13. Schecter A, Papke O, Harris TR, Tung KC, Musumba A, Olson J, et al. Polybrominated diphenyl ether (PBDE) levels in an expanded market basket survey of U.S. food and estimated PBDE dietary intake by age and sex. Environ Health Perspect 2006;114:1515–20. Schecter A, Shah N, Colacino JA, Brummit SI, Ramakrishnan V, Harris TR, et al. PBDEs in US and German clothes dryer lint: a potential source of indoor contamination and exposure. Chemosphere 2009;75:623–8. Schecter A, Haffner D, Colacino J, Patel K, Opel M, Birnbaum L. Polybrominated diphenyl ethers (PBDEs) and hexabromocyclodecane (HBCD) in composite U.S. food samples. Environ Health Perspect 2010a;118:357–62. Schecter A, Colacino J, Patel K, Kannan K, Yun SH, Haffner D, et al. Polybrominated diphenyl ether levels in foodstuffs collected from three locations from the United States. Toxicol Appl Pharmacol 2010b;243:217–24. Schreiber T, Gassman K, Gotz C, Hubenthal U, Moors M, Krause G, et al. Polybrominated diphenyl ethers induce developmental neurotoxicity in a human in vitro model: evidence for endocrine disruption. Environ Health Perspect 2010;118:572–8. She J, Holden A, Sharp M, Tanner M, Williams-Derry C, Hooper K. Polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) in breast milk from the Pacific Northwest. Chemosphere 2007;207:S307–17. Silberberg H, Jacobi S, Stedeford T, Biesemeier J, Ariano J, Beck M. An oral developmental neurotoxicity study of decabromodiphenyl ether (decaBDE) in rats. In: BFR2009. May 19–20, Ottawa, Canada; 2009. Sjodin A, Hagmar L, KLasson-Wehler E, Kronholm-Diab K, Jakobsson E, Bergman A. Flame retardant exposure: polybrominated diphenyl ethers in blood from Swedish workers. Environ Health Perspect 1999;107:643–8. Sjodin A, Papke O, McGahee E, Jones RS, Focant JF, Pless-Mulloli T, et al. Concentration of polybrominated diphenyl ethers (PBDEs) in household dust from various countries—inhalation a potential route of human exposure. Organohalog Compd 2004;66:3817–22. Sjodin A, Papke O, McGahee E, Focant JF, Jones RS, Pless-Mulloli T, et al. Concentration of polybrominated diphenyl ethers (PBDEs) in household dust from various countries. Chemosphere 2008;73(Suppl):S131–6. Soderstrom G, Sellstrom U, de Wit C, Tysklind M. Photolytic debromination of decabromodiphenyl ether (BDE 209). Environ Sci Technol 2004;38:127–32. Stapleton HM, Dodder NG, Offenberg JH, Schantz M, Wise SA. Polybrominated diphenyl ethers in house dust and clothes dryer lint. Environ Sci Technol 2005;39:925–31. Stapleton HM, Brazil B, Holbrook D, Mitchelmore CL, Benedict R, Kostantinov A, et al. In vivo and in vitro debromination of decabromodiphenyl ether (BDE 209) by juvenile rainbow trout and common carp. Environ Sci Technol 2006;40:4653–8. Stapleton HM, Dodder NG. Photodegradation of decabromodiphenyl ether in house dust by natural sunlight. Environ Toxicol Chem 2008;27:306–12. Stapleton HM, Kelly SM, Pei R, Letcher RJ, Gunsch C. Metabolism of polybrominated diphenyl ethers (PBDEs) by human hepatocytes in vitro. Environ Health Perspect 2009;117:197–202. Stewart P, Reihman J, Gump B, Lonky E, Darvill T, Pagano J. Response inhibition at 8 and 9 years of age in children prenatally exposed to PCBS. Neurotoxicol Teratol 2005;27:771–80. Sudaryanto A, Kajiwara N, Tsydenova OV, Isobe T, Yu H, Takahashi S, et al. Levels and congener specific profiles of PBDEs in human breast milk from China: implication on exposure sources and pathways. Chemosphere 2008a;73:1661–8.

24

L.G. Costa, G. Giordano / NeuroToxicology 32 (2011) 9–24

Sudaryanto A, Kajiwara N, Takahashi S, Muawanah. Tanabe S. Geographical distribution and accumulation features of PBDEs in human breast milk from Indonesia. Environ Pollut 2008b;151:130–8. Suh YW, Buettner GR, Venkataraman S, Treimer SE, Robertson LW, Ludewig G. UVA/Binduced formation of free radicals from decabromodiphenyl ether. Environ Sci Technol 2009;43:2581–8. Suvorov A, Girard S, Lachapelle S, Abdelouahab N, Sebire G, Takser L. Perinatal exposure to low-dose BDE-47, an emergent environmental contaminant, causes hyperactivity in rat offspring. Neonatology 2009;95:203–9. Tagliaferri S, Caglieri A, Goldoni M, Pinelli S, Alinovi R, Poli D, et al. Low concentrations of the brominated flame retardants BDE-47 and BDE-99 induce synergistic oxidative stress-mediated neurotoxicity in human neuroblastoma cells. Toxicol In Vitro 2010;24:116–22. Tan J, Cheng SM, Loganath A, Chong YS, Obbard JP. Polybrominated diphenyl ethers in house dust in Singapore. Chemosphere 2007;66:985–92. Thomas GO, Moss SEW, Asplund L, Hall AJ. Absorption of decabromodiphenyl ether and other organohalogenated chemicals by grey seals (Helichoerus grypus). Environ Pollut 2005;133:581–6. Thomas GO, Wilkinson M, Hodson S, Jones KC. Organohalogen chemicals in human blood from the United Kingdom. Environ Pollut 2006;141:30–41. Thomsen C, Lundanes E, Becher G. Brominated flame retardants in archived serum samples from Norway: a study on temporal trends and the role of age. Environ Sci Technol 2002;36:1414–8. Thomsen C, Stigum H, Froshaug M, Broadwell SL, Becher G, Eggesbo M. Determinants of brominated flame retardants in breast milk from a large scale Norwegian study. Environ Int 2010;36:68–74. Thuresson K, Bergman A, Jakobsson K. Occupational exposure to commercial decabromodiphenyl ether in workers manufacturing or handling flame retarded rubber. Environ Sci Technol 2005;39:1980–6. Thuresson K, Hoglund P, Hagmar L, Sjodin A, Bergman A, Jakobsson K. Apparent halflives of hepta-to decabrominated diphenyl ethers in human serum as determined in occupationally exposed workers. Environ Health Perspect 2006;114:176–81. Tokarz JA, Ahn MY, Leng J, Filley TR, Nies L. Reductive debromination of polybrominated diphenyl ethers in anaerobic sediment and a biometric system. Environ Sci Technol 2008;42:1157–64. Toms LML, Hearn L, Kennedy K, Harden F, Bartkow M, Temme C, et al. Concentrations of polybrominated diphenyl ethers (PBDEs) in matched samples of human milk, dust and indoor air. Environ Int 2009;35:864–9. Tseng LH, Lee CW, Pan MH, Tsai SS, Li MH, Chen JR, et al. Postnatal exposure of the male mouse to 2,20 ,3,30 ,4,40 ,5,50 ,6,60 -decabrominated diphenyl ether: decreased epididymal sperm functions without alterations in DNA content and histology in testis. Toxicology 2006;224:33–43. Tseng LH, Li MH, Tsai SS, Lee CW, Pan MH, Yao WJ, et al. Developmental exposure to decabromodiphenyl ether (PBDE 209): effects on thyroid hormone and hepatic enzyme activity in male mouse offspring. Chemosphere 2008;70:640–7. Tsydenova OV, Sudaryanto A, Kajiwara N, Kunisue T, Batoev VB, Tanabe S. Organohalogen compounds in human breast milk from Republic of Buryatia, Russia. Environ Pollut 2007;146:225–32. Tue NM, Sudaryanto A, Minh TB, Isobe T, Takahashi S, Viet PH, et al. Accumulation of polychlorinated biphenyls and brominated flame retardants in breast milk from women living in Vietnamese e-waste recycling sites. Sci Total Environ 2010;408:2155–62. Uemura H, Arisawa K, Hiyoshi M, Dakeshita S, Kitayama A, Takami H, et al. Congenerspecific body burden levels and possible determinants of polybrominated diphenyl ethers in the general Japanese population. Chemosphere 2010;79:706–12. USEPA (United States Environmental Protection Agency). Decabromodiphenyl ether (CASRN 1163-19-5). Integrated Risk Information System (IRIS). Washington, DC: National Center for Environmental Assessment; 1989. USEPA (United States Environmental Protection Agency). Developmental neurotoxicity study. OPPTS 870. 6300. EPA 712-C-98-239. 1998. USEPA (United States Environmental Protection Agency). Toxicological review of decabromodiphenyl ether (BDE-209). Washington, DC: USEPA; June 2008 p. 72. USEPA (United States Environmental Protection Agency). An exposure assessment of polybrominated diphenyl ethers. Washington, DC: USEPA; May 2010 p. 378. Van den Steen E, Covaci A, Jaspers VLB, Dauwe T, Voorspoels S, Eens M, et al. Accumulation, tissue-specific distribution and debromination of decabromodiphenyl ether (BDE-209) in European starling (Sturnus vulgaris). Environ Pollut 2007;148:648–53. Van der Ven LTM, van de Kuil T, Leonards PEG, Slob W, Canton RF, Germer S, et al. A 28day oral dose toxicity study in Wistar rats enhanced to detect endocrine effects of decabromodiphenyl ether (decaBDE). Toxicol Lett 2008;179:6–14. Viberg H. Neonatal ontogeny and neurotoxic effect of decabrominated diphenyl ether (PBDE 209) on levels of synaptophysin and tau. Int J Dev Neurosci 2009a;27:423–9. Viberg H. Exposure to polybrominated diphenyl ethers 203 and 206 during the neonatal brain growth spurt affects proteins important for normal neurodevelopment in mice. Toxicol Sci 2009b;109:306–11. Viberg H, Fredriksson A, Eriksson P. Neonatal exposure to the brominated flame retardant 2,20 ,4,40 ,5-pentabromodiphenyl ether causes altered susceptibility in the cholinergic transmitter system in the adult mouse. Toxicol Sci 2002;67:104–7.

Viberg H, Fredriksson A, Jakobsson E, Orn U, Eriksson P. Neurobehavioral derangements in adult mice receiving decabrominated diphenyl ether (PBDE 209) during a defined period of neonatal brain development. Toxicol Sci 2003a;76:112–20. Viberg H, Fredriksson A, Eriksson P. Neonatal exposure to polybrominated diphenyl ether (PBDE 153) disrupts spontaneous behavior, impairs learning and memory, and decreases hippocampal cholinergic receptors in adult mice. Toxicol Appl Pharmacol 2003b;192:95–106. Viberg H, Fredriksson A, Eriksson P. Investigations of strain and gender differences in developmental neurotoxic effects of polybrominated diphenyl ethers in mice. Toxicol Sci 2004;81:344–53. Viberg H, Johansson N, Fredriksson A, Eriksson J, Marsh G, Eriksson P. Neonatal exposure to higher brominated diphenyl ethers, hepta-, octa-, and nonabromodiphenyl ether, impairs spontaneous behavior and learning and memory functions of adult mice. Toxicol Sci 2006;211–8. Viberg H, Frederiksson A, Eriksson P. Changes in spontaneous behavior and altered response to nicotine in the adult rat, after neonatal exposure to the brominated flame retardant, decabrominated diphenyl ether (PBDE 209). Neurotoxicology 2007;28:136–42. Viberg H, Mundy W, Eriksson P. Neonatal exposure to decabrominated diphenyl ether (PBDE 209) results in changes in BDNF, CaMKII and GAP-43, biochemical substrates of neuronal survival, growth, and synaptogenesis. Neurotoxicology 2008;29:152– 9. Vieth B, Herrmann T, Mielke H, Ostermann B, Papke O, Rudiger T. PBDE levels in human milk: the situation in Germany and potential influencing factors—a controlled study. Organohalog Compd 2004;66:2643–8. Vijverberg HPM, van den Berg M. Letter to the Editor. Toxicol Sci 2004;79:205–6. Vitalone A, Catalani A, Chiodi V, Cinque C, Fattori V, Goldoni M, et al. Neurobehavioral assessment of rats exposed to low doses of PCB 126 and methyl mercury during development. Environ Toxicol Pharmacol 2008;25:103–13. Vitalone A, Catalani A, Cinque C, Fattori V, Matteucci P, Zuena AR, et al. Long-term effects of developmental exposure to low doses of PCB 126 and methylmercury during development. Toxicol Lett 2010;197:38–45. Wahl M, Lahni B, Guenther R, Kuch B, Yang L, Straehle U, et al. A technical mixture of 2,20 ,4,40 -tetrabromo diphenyl ether (BDE47) and brominated furans triggers aryl hydrocarbon receptor (AhR) mediated gene expression and toxicity. Chemosphere 2008;73:209–15. Wang F, Wang J, Dai J, Hu G, Wang J, Luo X, et al. Comparative tissue distribution, biotransformation and associated biological effects by decabromodiphenyl ethane and decabrominated diphenyl ether in male rats after a 90-day oral exposure study. Environ Sci Technol 2010;44:5655–60. Wang J, Kliks MM, Jun S, Li QX. Residues of polybrominated diphenyl ethers in honeys from different geographical regions. J Agric Food Chem 2010b;58:3495–501. Wang J, Ma YJ, Chen SJ, Tian M, Luo XJ, Mai BX. Brominated flame retardants in house dust from e-waste recycling and urban areas in South China: implications on human exposure. Environ Int 2010a;36:535–41. World Health Organization. Safety evaluation of certain contaminants in food. Polybrominated diphenyl ethers. WHO Food Additives Series 2006;55:351–561. Wilford BH, Shoeib M, Harner T, Zhu J, Jones KC. Polybrominated diphenyl ethers in indoor dust in Ottawa, Canada: implications for sources and exposure. Environ Sci Technol 2005;39:7027–35. Williams AL, DeSesso JM. The potential of selected brominated flame retardants to affect neurological development. J Toxicol Environ Health B 2010;13:411–48. Wu K, Xu X, Liu J, Guo Y, Li Y, Huo X. Polybrominated diphenyl ethers in umbilical cord blood and relevant factors in neonates from Guiyu, China. Environ Sci Technol 2010;44:813–9. Wu N, Herrmann T, Papke O, Tickner J, Hale R, Harvey E, et al. Human exposure to PBDEs: associations of BDE body burdens with food consumption and house dust concentrations. Environ Sci Technol 2007;41:1584–9. Wu Y, Yu YH, Chen DJ, Jiang HP. Effect of maternal BDE-209 exposure on the learning and memory ability of offspring rats and the dose–effect relation. Nan Fang Yi Ke Da Xue Xue Bao 2008;28:976–8. Xing T, Chen L, Tao Y, Wang M, Chen J, Ruan DY. Effects of decabrominated diphenyl ether (PBDE 209) exposure at different developmental periods of synaptic plasticity in the dentate gyrus of adult rats in vivo. Toxicol Sci 2009;110:401–10. Xing TR, Yong W, Chen L, Tang ML, Wang M, Chen JT, et al. Effects of decabrominated diphenyl ether (PBDE 209) on voltage-gated sodium channels in primary cultured rat hippocampal neurons. Environ Toxicol 2010;25:400–8. Zhang C, Liu F, Liu X, Chen D. Protective effect of N-acetylcysteine against BDE-209induced neurotoxicity in primary cultured neonatal rat hippocampal neurons in vitro. Int J Dev Neurosci 2010a;28:521–8. Zhang C, Liu X, Chen D. Role of brominated diphenyl ether-209 in the differentiation of neural stem cells in vitro. Int J Dev Neurosci 2010b;28:497–502. Zhou T, Ross DG, DeVito MJ, Crofton KM. Effects of short-term in vivo exposure to polybrominated diphenyl ethers on thyroid hormones and hepatic enzyme activities in weanling rats. Toxicol Sci 2001;61:76–82. Zhou T, Taylor MM, DeVito MJ, Crofton KM. Developmental exposure to brominated diphenyl ethers results in thyroid hormone disruption. Toxicol Sci 2002;66:105–16. Zoeller RT, Crofton KM. Mode of action: developmental thyroid hormone insufficiency – neurological abnormalities resulting from exposure to propylthiouracil. Crit Rev Toxicol 2005;35:771–81.