Polybrominated Diphenyl Ethers and 2,2′,4,4′,5,5′-Hexabromobiphenyl (BB-153)
Polybrominated diphenyl ethers (PBDEs) are a class of synthetic chemicals first produced commercially in the 1970s. They are added to products such as foam padding, textiles, or plastics to retard combustion. 2,2′,4,4′,5,5′-hexabromobiphenyl (BB-153) is a brominated biphenyl that was used as a flame retardant in the U.S. until the 1970s. Its use was phased out following an accidental contamination of cattle feed in the state of Michigan with the contamination extending to other animals, the environment, and into humans (Fries, 1985).
Three major commercial mixtures of PBDEs have been produced and used. These are named for the average number of bromines attached to the diphenyl ether structure, e.g., pentaBDE. The pentaBDE technical mixture contains 50-60% of PBDE congeners with five bromines, 24-38% with four bromines (tetraBDEs) and 4-8% with six bromines (hexaBDE), though reports on mixtures vary (Alaee et al., 2003; Birnbaum and Staskal, 2004; OECD, 1994). Commercial pentaBDEs are often added to polyurethane foams used in mattresses, upholstered furniture, and carpet padding. OctaBDE technical mixtures contain 10-12% of PBDE congeners with six bromines, 43-44% with seven bromines, 31-35% with eight bromines and 10-11% with nine bromines. OctaBDE mixtures are added to acrylonitrile-butadiene-styrene used in computer and appliance casings, and also to some polyolefins and nylon. DecaBDE is the most widely used PBDE globally and greater than 97% of its content includes PBDEs with ten bromines. It is added to polystyrene, polybutylene, nylon, polypropylene, and other thermoelastic polymers used in adhesives, wire insulation, casings for televisions and computers, and in some non-clothing textiles (OECD, 1994; Sjodin et al., 2003; WSDH, 2004). PBDEs are often combined with antimony trioxide to enhance the fire protection offered by the PBDEs. For example, protected polypropylene can contain 23% decaBDE and 8% antimony trioxide by weight. PBDE content in protected products varies from 3-33% (Gill et al., 2004).
PBDE production makes up about 25% of all fire retardant production. In 2000, global production was 67,000 metric tons annually with about 80% of the total being decaBDE (Birnbaum and Staskal, 2004; WSDH, 2004). Most of the pentaBDE produced has been used within the U.S. About 40% and 44% of the global production of octaBDE and decaBDE were also used in the U.S.Since PBDEs are not chemically bound to the flame-retarded material, they can enter the environment from volatilization, leaching, or degradation of PBDE-containing products (Gill et al., 2004). Also, PBDEs can enter the environment from manufacturer-related releases. When thermally decomposed, PBDEs can produce polybrominated dibenzo-p-dioxins and dibenzofurans (Watanabe and Sakai, 2003). Manufacturers of pentaBDE and octaBDE in the U.S. were to have phased out production of these chemicals by 2004 and U.S. EPA issued a rule to prevent new production (U.S. EPA, 2005). PBDEs are generally persistent in the environment and have been measured in aquatic sediments and aquatic and terrestrial animals, especially in fish where PBDEs are known to bioconcentrate. Several studies of stored biologic specimens have shown dramatic increases in PBDE concentrations over the last several decades, for example, in archived bird eggs (Norstrom et al., 2002).
Human exposure to PBDEs is thought to result from dietary sources, including fish, fatty foods, and mother’s milk. However, oral ingestion from dust and leachates may be a larger source (Sjodin et al., 2008a), particularly for children (Jones-Otazo et al., 2005; Stapleton et al., 2005). Once absorbed, PBDEs distribute into body fat. The metabolism and elimination of PBDEs in humans are not well characterized. One occupational study indicated that decaBDE has an elimination half-life of 11-18 days and the octaBDEs have half-lives ranging between 37-91 days (Thuresson et al., 2006). In animals, PBDE elimination occurs primarily through fecal excretion with decaBDE being more rapidly eliminated than the other less brominated PBDEs (Gill et al., 2004; Hardy, 2002). Some PBDEs measured in human serum by the National Biomonitoring Program (CDC, 2009) are listed in the table.
Polybrominated Diphenyl Ethers
|Polybrominated Diphenyl Ether (IUPAC number)||CAS Number|
|Serum 2,2′,4′-Tribromodiphenyl ether (BDE 17)||147217-75-2|
|Serum 2,4,4′-Tribromodiphenyl ether (BDE 28)||41318-75-6|
|Serum 2,2′,4,4′-Tetrabromodiphenyl ether (BDE 47)||5436-43-1|
|Serum 2,3′,4,4′-Tetrabromodiphenyl ether (BDE 66)||189084-61-5|
|Serum 2,2′,3,4,4′-Pentabromodiphenyl ether (BDE 85)||182346-21-0|
|Serum 2,2′,4,4′,5-Pentabromodiphenyl ether (BDE 99)||60348-60-9|
|Serum 2,2′,4,4′,6-Pentabromodiphenyl ether (BDE 100)||189084-64-8|
|Serum 2,2′,4,4′,5,5′-Hexabromodiphenyl ether (BDE 153)||68631-49-2|
|Serum 2,2′,4,4′,5,6′-Hexabromodiphenyl ether (BDE 154)||207122-15-4|
|Serum 2,2′,3,4,4′,5′,6-Heptabromodiphenyl ether (BDE 183)||207122-16-5|
Human health effects from PBDEs at low environmental doses or at biomonitored levels from low environmental exposures are unknown. In animal studies, PBDEs have low acute toxicity, but have demonstrated effects on thyroid function, neurodevelopment, hepatic enzyme induction and hepatic injury in subchronic or chronic dosing studies (Birnbaum and Staskal, 2004; Branchi et al., 2002; Gill et al., 2004; Hallgren and Darnerud, 2002; Viberg et al., 2003 and 2004; Zhou et al., 2002). Some developmental and behavioral effects may be mediated by the aforementioned effect on the thyroid, by alteration in cholinergic function (Branchi et al., 2003; Dufault et al., 2005), or by altered intracellular signaling within brain cells (Kodavanti et al., 2005). The lesser brominated PBDEs have been reported to have fetotoxic and reproductive effects, to alter expression of estrogen-regulated genes and receptors, and to have anti-androgenic effects (Ceccatelli et al., 2006; Gill et al., 2004; Kuriyama et al., 2005; Stoker et al., 2005; Talsness et al., 2005; WSDH, 2004). PentaBDE is considered more toxic than decaBDE and the most sensitive effects of pentaBDE in animal studies are neurodevelopmental and reproductive. In a study of electronics dismantlers, serum levels of PBDEs were not generally higher than in nonexposed workers and were not associated with changes in thyroid function (Julander et al., 2005). PBDEs are not considered genotoxic and are not classified by IARC and NTP with respect to human carcinogenicity. Additional information about external exposure (i.e., environmental levels) and health effects is available from ATSDR at: https://www.atsdr.cdc.gov/toxprofiles/.
Levels of PBDEs in serum reflect cumulative exposure over the recent months to years of exposure. The PBDE congeners measured for biomonitoring often include those containing three bromines (BDE-17, BDE-28), four bromines (BDE-47, BDE-66), five bromines (BDE-85, BDE-99, BDE-100), six bromines (BDE-153, BDE-154) and seven bromines (BDE-183). Analysis of the NHANES 2003-2004 subsample showed detection of BDE-47 (a tetraBDE present in commercial pentaBDEs) in nearly all participants and detection of BDE-28, BDE-99, BDE-100, and BDE-153 in greater than 60 percent of participants (Sjodin et al., 2008b). Levels of these PBDEs tended to be well-correlated with each other. Serum levels of BDE-47, BDE-99, and BDE-153 were found to decrease with increasing age from 12-19 to 20-39 to 40-59 years, and then increase slightly in the 60 years and older age group. Slight differences by gender and race/ethnicity were also observed for several PBDEs (Sjodin et al., 2008b).
From the 1970s to the late 1990s, levels of BDE-47 had increased in samples of breast milk and sera in Sweden and Norway, respectively (Meironyté et al., 1999; Thomsen et al., 2002). Also, in small samplings of residents in Japan and U.S., serum levels have been shown to increase by more than fivefold to twentyfold over the past two decades (Koizumi et al., 2005; Schecter et al., 2005; Sjodin et al., 2004). Several small studies of U.S. residents have shown increasing levels of BDE-47 during recent decades that were 3-10 times higher than contemporary European residents (Petreas et al., 2003; Sjodin et al., 2003). Serum levels of PBDEs in the NHANES 2003-2004 subsample (Sjodin et al., 2008b) also appeared generally higher than those reported for Japan, Sweden, and Norway (Koizumi et al., 2005; Thomsen et al., 2002; Thuresson et al., 2006). In most studies, BDE-47 demonstrates the highest levels of all the measured PBDEs.
Detection of BB-153 was also prevalent in the NHANES 2003-2004 subsample and increase with age (Sjodin et al., 2008b). This age trend may be due to the longer time that BB-153 stays in the body or due to greater past exposures in older people. Mexican Americans and NHANES participants born in foreign countries had lower serum concentrations of BB-153 (Sjodin et al, 2008b). Levels of BB-153 in NHANES 2003-2004 were about one-fourth to one-fortieth of the levels of BDE-47, depending on the age group. In human sera from Sweden, BB-153 was generally not detected in contrast to detectable levels a small regional sample of U.S. residents (Sjodin et al., 2001).
Finding measurable amounts of PBDEs or BB-153 in serum does not imply that the levels of these chemicals cause an adverse health effect. Biomonitoring studies of serum PBDEs and BB-153 can provide physicians and public health officials with reference values so that they can determine whether people have been exposed to higher levels of PBDEs or BB-153 than levels found in the general population. Biomonitoring data can also help scientists plan and conduct research on exposure and health effects.
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