Children’s Environmental Health in Michigan
Neurotoxicity: PBDEs
Polybrominated diphenyl ethers (PBDEs) are commonly-used flame retardants present in a multitude of consumer products. Due to the ubiquity of the compounds, human exposure to PBDEs may occur via many different pathways, and high concentrations of PBDEs have been detected in breast milk, blood, and adipose tissue of Americans (Schecter et al. 2006). PBDEs also have been detected in placenta and umbilical cord serum, indicating that prenatal exposure is likely (Gómara et al. 2007) and could continue after birth via breast milk. Although research on the health effects of PBDE exposure is limited, PBDEs have been linked to neurotoxicity, liver toxicity, reproductive toxicity, and irregular thyroid levels in animal studies (Birnbaum et al. 2004, ATSDR 2004, Zhou et al. 2002; Turyk et al. 2008). Other studies have noted adverse effects on male reproductive hormones (Meeker et al. 2009) and fertility (Akutsu et al. 2008), cryptorchidism (Main et al. 2007) and lower birth weight and length (Chao et al. 2007). This chapter will focus on neurotoxicity and PBDE’s.
There are three different commercial forms of PBDEs called penta-, octa- and deca-BDE. Penta- and octa-BDE have been voluntarily withdrawn from the marketplace although there is a reservoir of products in use that still contain these chemicals (Lunder, Jacob 2008). Deca-BDE is still widely used in electronics, wire and cable insulation, textiles, automobiles and airplanes, and other applications, however, the largest manufacturers and importers have agreed to voluntarily end most uses by 2013 (EPA, DecaBDE Phase-out Initiative, 2009). In late 2009, the EPA announced a “chemical action plan” for PBDEs to reduce the general public’s exposure to these compounds (EPA, Polybrominated Diphenyl Ethers (PBDEs) Action Plan Summary, 2009). Legislation to phase out Deca-BDE in some products has passed in four states, and is pending in others, including Michigan. This section will briefly summarize PBDE exposure and associated health effects, present leading policy practices from other states, and recommend steps to minimize childhood PBDE exposure in Michigan.
PBDEs are not the only brominated fire retardants of concern. New flame retardants being developed as alternatives may also be toxic. In addition, there are a number of chlorinated flame retardants that may pose similar hazards. In late October of 2010, 145 prominent scientists from 22 countries signed a first-ever consensus statement documenting health hazards from flame retardant chemicals including brominated and chlorinated chemicals (DiGangi, 2010). In an accompanying editorial, two signers noted: “Adequate toxicity information is lacking but data indicate that the group contains compounds that are carcinogenic, mutagenic, reproductive and developmental toxicants, neurotoxicants, and endocrine disruptors” (Birnbaum, 2010). Because the data on these compounds is less well developed, they are beyond the scope of this report.
Michigan Highlights
Exposure Data
Although Michigan data are limited, some preliminary research and evidence of exposure in wildlife suggest ubiquitous presence and widespread human PBDE exposure in the state. In 2003, the Environmental Working Group (EWG) conducted a small biomonitoring study looking for flame retardants and other industrial chemicals in breast milk, one of the study participants was from Ann Arbor, Michigan; her breast milk contained 24 of the 44 compounds tested (EWG 2003).
PBDEs are regularly detected in fish caught in the Great Lakes, indicating both widespread contamination of the environment and a potential route for human exposure. In Michigan, there are currently no advisories that provide consumers guidance on whether it is safe to eat Great Lakes fish based on PBDE levels (Alexander 2008). A study is underway to assess levels of PBDEs in breast milk of fish-eaters residing in southwest Michigan (MDEQ 2008b).
Policy
Children may be more vulnerable to exposure to certain PBDE congeners, although further study is needed to determine the extent and degree of the potential harm. However, existing evidence supports the need to take measures immediately to reduce human exposures to PBDEs to protect children’s environmental health. Some legislative and regulatory measures related to certain PBDEs exist at the state level in Michigan and in other states to protect human health and the environment.
Michigan is a national leader in PBDE policy because it was one of the first states to ban products with penta- or octa-BDEs. Subsequently, penta and octa were voluntarily withdrawn by the manufacturers. However, to better protect children’s environmental health, Michigan should adopt policy to ban deca-BDE. Michigan’s Interdepartmental Toxics Steering Group’s PBDE Sub-committee (MDEQ 2008b) recommended a ban where safer alternatives were available, as did the Michigan Department of Environmental Quality (MDEQ 2008a).
Background Information
Sources and Types of PBDEs
PBDEs are added to plastic, foam, and fabric products as flame retardants; therefore, these chemicals are mainly indoor pollutants that can migrate from homes, offices, and other buildings into the environment through dust, air, wastewater, and landfills. It can accumulate in sediment, and in the food web. Concentrations of PBDEs are generally higher in households in the U.S. when compared to European households, and indoor air and dust concentrations are roughly 5-10 times higher than outdoor concentrations (Lorber 2008). Common products containing PBDEs include computers, television sets, upholstered furniture, textiles, carpets, mattresses, appliances, cell phones, automotive electronics, and automobile seats (Betts 2008, Schecter et al. 2003). When deca-BDE is added to these products, it is not bound and can therefore be released through leaching or fracturing of the material (Siddiqi et al. 2003). One recent study found that material wear and abrasion released PBDE particles 50-200 nanometers (nm, one billionth of a meter) in diameter (Webster et al. 2009). The fate of these particles depends on the weight of the particle. Lighter PBDEs—those with fewer bromine atoms—can become airborne and attach to dust particles in the air. Heavier PBDEs, including deca-BDE, tend to stay adhered to particles of the original material. In either situation, these contaminated particles are capable of traveling long distances on the wind, spreading PBDEs into the environment (Breivik 2006, Song 2005).
The class of chemicals known as ‘polybrominated diphenyl ethers’ actually covers 209 different chemical congeners. These congeners are variants on the basic chemical structure with different numbers of bromine atoms replacing any of the ten hydrogen atoms. The congeners are ordered based on the number of bromine substitutions and their position; e.g., deca-BDE, which has all of the hydrogen replaced with bromine, is BDE-209. The prefixes refer to the number of bromine atoms in the molecule, and therefore may refer to multiple congeners; e.g. there are multiple configurations that have five bromines, so penta-BDE refers to more than one congener (Zeng 2007). Commercially, the prefixes only refer to the dominant type of congener in a mixture. For example, penta-BDE contains roughly 43% penta-BDEs as well as 37% tetra-BDEs and 6.8% hexa-BDEs (Sjodin et al. 2008). Commercial deca-BDE is almost solely BDE-209, which is the only congener containing ten bromine atoms (Sjodin et al. 2008).
Since the Michigan legislature banned penta- and octa-BDE due to their toxicity and persistence in late 2004, a growing body of evidence has suggested that deca-BDE may also be toxic. In addition, evidence has strengthened that deca-BDE can break down in the environment into less brominated forms, including penta- and octa-BDE. This process, called debromination, can occur in wastewater treatment (Steen et al. 2009), microbes (He et al. 2006), and wildlife (Law et al. 2006, Stapleton et al. 2004). Ultraviolet radiation can break down deca-BDE into lower brominated PBDEs, brominated dioxins and furans, and bromine free radicals (Schenker et al. 2008, Siddiqi et al. 2003, Steen et al. 2009). In addition to debromination, PBDEs also can be transformed into hydroxylated metabolites (OH-PBDEs) (Hakk and Letcher 2003). OH-PBDEs have been detected in humans (Stapleton et al. 2009) and fish (Valters et al. 2005) and they have greater potency as endocrine disruptors than PBDEs (Hamers et al. 2002).
It has been estimated that deca-BDE is the most widely used form of PBDE in the U.S. (Washington State Department of Health 2008), and that U.S. consumption of deca-BDE accounts for roughly 44% of the total world market (Sjodin et al. 2008). Furthermore, the Natural Resources Defense Council (NRDC) found that PBDEs have become the second largest class of additives used by the plastics industry (NRDC 2005). However, on December 17, 2009, as the result of negotiations with EPA, the two U.S. producers of decabromodiphenyl ether (deca-BDE), Albemarle Corporation and Chemtura Corporation, and the largest U.S. importer, ICL Industrial Products, Inc., announced commitments to phase out deca-BDE in the United States. The companies have committed to end production, importation, and sales of deca-BDE for most uses in the United States by December 31, 2012, and to end all uses by the end of 2013. (EPA, DecaBDE Phase-out Initiative, 2009). This agreement is voluntary, and does not eliminate every potential source of deca-BDE, but it will likely result in a dramatic reduction in deca-BDE use. Emerging brominated and chlorinated alternatives to PBDEs have already raised concerns. Assuring the safety of alternatives will be critical to protect children’s health.
On December 30, 2009, EPA Administrator Lisa Jackson released action plans describing steps EPA will take to manage concerns for polybrominated diphenyl ethers (PBDEs) in products. That action plan is being developed with the purpose of reducing exposures to PBDEs (EPA, Polybrominated Diphenyl Ethers (PBDEs) Action Plan Summary, 2009).
In October of 2010, 145 prominent scientists from 22 countries signed a first-ever consensus statement documenting health hazards from flame retardant chemicals found at high levels in home furniture, electronics, insulation, and other products. The San Antonio Statement on Brominated and Chlorinated Flame Retardants, which documents this pervasive class of chemicals, its potential to cause serious harm, its tendency to create more hazardous byproducts during combustion, and the limited fire safety benefits of these chemicals, was published in Environmental Health Perspectives (DiGangi, 2010) with an accompanying editorial. Dr. Linda Birnbaum, Director of the National Institute of Environmental Health Sciences co-penned the piece, which noted: “Adequate toxicity information is lacking but data indicate that the group contains compounds that are carcinogenic, mutagenic, reproductive and developmental toxicants, neurotoxicants, and endocrine disruptors.” (Birnbaum, 2010)
Neurotoxicity and PBDEs
Limited human data are available on health effects associated with PBDE exposure, and those studies evaluating human health effects have typically focused on occupational exposures and health effects other than neurotoxicity (Darnerud et al. 2001). However, animal studies have associated early life PBDE exposure with adverse brain development (Ericksson et al. 2001, Darnerud et al. 2001; Birnbaum et al. 2004). These studies observed both behavioral and cognitive effects of PBDE exposure. One study found that rats exposed in utero to low levels of PBDE-99, a persistent PBDE congener, led to hyperactivity in offspring (Kuriyama et al. 2005). The doses used in this study (60 and 300 µg/kg body weight) were low compared to many animal studies, but still 6-29 times higher than the highest exposure reported in human breast adipose tissue (Kuriyama et al. 2005). Recent research on mice has shown that exposure to deca-BDE during youth can result in hyperactivity, lack of habituation, and developmental neurotoxicity that worsens with age (Johansson et al. 2008).
PBDEs have been shown to disrupt the equilibrium of the thyroid hormone system. Because brain development is dependent on proper regulation of thyroid hormones, such disruption of the endocrine system may lead to impaired neurological functioning. (Darnerud et al. 2001).
A number of studies have linked prenatal PBDE exposure in mice to learning difficulties later in life. One study found an association between PBDE-99 exposure in neonatal mice and effects on learning and memory functions as well as motor behavior (Darnerud et al. 2001). Another study found that neonatal exposure to PBDE-153 led to behavioral changes and learning and memory impairment (Viberg et al. 2003). These studies led researchers to conclude in a review article that, “Available information on PBDE neurotoxicity obtained from animal studies and the possibility of neonatal exposure to PBDEs via the mother’s milk suggests that these compounds may represent a potential risk for neurobehavioral development in humans” (Branchi et al. 2003). Childhood Exposure to PBDEs
Americans, including children, are being exposed to increasing concentrations of toxic flame retardants. This increase is apparent in the reported rise in blood and breast milk PBDE levels over the past two decades (Lorber 2008; Schecter et al., 2006). In addition, the American body burden of PBDEs is roughly ten to twenty-fold higher than that of some populations in Europe (Lorber 2008; Schecter et al. 2006). PBDEs are lipophilic, persistent environmental pollutants that accumulate in fatty tissues and bioaccumulate in the food chain (Schecter et al. 2003). Exposure to PBDEs can occur through dust, and through consumption of contaminated food (e.g., fish, meats, dairy) and can be passed from a lactating mother to an infant via breast milk (Schecter et al., 2006). Because of the elevated levels present in breast milk, infants can consume up to 307 nanograms per kilogram per day (ng/kg/day) of total PBDEs, which is considerably higher than the 0.9 ng/kg/day consumed by adults. Children aged 2-5 years were estimated to consume 2 ng/kg/day of total PBDEs from food (Schecter et al. 2006).
Household dust and indoor air can become contaminated because PBDEs are released as dust or vapors over the lifetime of a product (Jones-Otazo et al, 2005). Concentrations of deca-BDE in household dust have been highly associated with levels of deca-BDE in televisions, demonstrating the connection between consumer products and PBDE exposure (Allen 2008). As consumer products are exposed to wear and abrasion, either through use or abuse, particles of the original material are liberated; PBDEs—particularly heavier species, including deca-BDE—can be attached to these particles and thus appear in dust (Webster et al. 2009).
The relative contribution of various exposure pathways is still being investigated. Several researchers estimated PBDE exposure is most significant through household dust, with diet playing a smaller role (Allen et al. 2007; Lorber 2008; Stapleton et al. 2008). One researcher estimated that roughly 80% of human exposure comes from dust and that diet accounts for roughly 20% (Lorber 2008). A more recent study found diet (consumption of red meat and poultry) to be more significant (Fraser et al. 2009). While more research is necessary to determine the significance of human exposure pathways, the continued disposal and degradation of PBDE-containing products may shift concentrations from the indoor environment to the outdoors and, consequently, increase dietary exposures (Harrad, Diamond 2006). This trend may accelerate as PBDE-containing products reach the end of their useful life and are discarded or sent to disposal facilities.
Researchers have found an association between levels of PBDEs in household dust and PBDE concentration in human breast milk (Wu 2005). Diet was also highly correlated with levels of PBDEs in breast milk. In addition to high levels of PBDE exposure from breast milk, children may also be more vulnerable to exposure from household dusts due to hand-to-mouth and other play behaviors and living lower to the floor. These factors are estimated to increase childhood exposure to PBDEs in dust by as much as a factor of 10 compared to adults (Stapleton 2008). In addition, an unborn fetus may be exposed to a mother’s PBDE body burden through the placenta and umbilical cord (ATSDR 2004). In a recent study in Spain, PBDE-47 was the predominant congener in serum samples (maternal, paternal, and umbilical cord), while PBDE-209 (deca) was predominant in placenta and breast milk samples (Gómara et al. 2007).
Materials containing PBDEs are used in a wide variety of products commonly found in every child’s indoor environment. Recent publications document high levels of these flame retardants inside automobiles, internet cafes, and offices (Mandalakis et al. 2008; Harrad et al. 2008). In fact, one study estimated that PBDE exposure from cars accounted for roughly 80% of total exposure for adults (Lorber 2008). Environments other than the home (e.g., cars, buses, schools, daycare), are thought to be major contributors to PBDE loading for children.
PBDE exposure can also occur near sources of localized contamination and emission. For example, PBDEs can leach from landfills where PBDE-containing products are disposed (Kim et al. 2006). If products containing PBDEs are incinerated, they can release other forms of PBDEs ¬¬along with brominated dioxins and furans, which can potentially pose even greater hazards (Agrell et al. 2004; Betts 2008). Children in proximity or downwind of these sites can be directly exposed to PBDEs through inhalation of emissions or contact with contaminated soils or dusts. Very few studies have analyzed PBDE concentrations in soils, although PBDEs were detected in over 90% of samples collected in a study conducted in the U.S. (Lorber 2008). More research is needed to understand the importance of soil as a PBDE exposure source.
Global and Nationwide Exposure Data
Data documenting PBDE exposure in children are currently limited, but biomonitoring studies show higher levels of PBDEs in children than in adults. Researchers in Australia have compiled what appears to be the most extensive data on PBDE levels in children (Betts 2008). The study found that children age 0-4 years had blood PBDE levels approximately 4 times higher than adults and children age 16 and up (Betts 2008).
In the United States, few studies have measured PBDE body burdens in young children, and most looked at only a handful of individuals. In a case study conducted in California in which blood PBDE concentrations were measured in one family, the children were found to have the highest levels of PBDEs (Fischer et al. 2006). The youngest child, a toddler, had the highest blood PBDE concentration. Similarly, a study by the Environmental Working Group (EWG) measuring PBDE levels in 20 toddlers and their mothers found that on average, the toddlers had levels three times as high as their mothers (Lunder, Jacob 2008). Although these results suggest that PBDE exposure is greater for younger individuals, further studies using larger cohorts are necessary to verify the findings.
There is more U.S. data on PBDE levels in children 12 years of age and older. Sjodin et al. measured PBDE concentration in serum samples collected from 2,062 U.S. residents aged 12 years and older in 2003 and 2004 (Sjodin et al. 2008). The study subjects were participants in the National Health and Nutrition Examination Survey (NHANES), a CDC study aimed to assess exposure to a variety of chemicals in a representative U.S. population sample. Penta- and hexa-BDEs, specifically the congeners BDE-47, BDE-100, and BDE-153, were detectable in over 90% of the samples. Concentrations of BDE-47, the most commonly detected PBDE congener, averaged 20.5 ng/g lipid; mean concentrations were highest for members of the population aged 12-19 years, and appeared to decrease with age (Sjodin et al. 2008). This trend was similar for BDE-100 and BDE-153, which may be attributable to different activity patterns, and therefore exposure pathways, of different age groups (Sjodin et al. 2008). Both the Fischer and Sjodin studies, along with the studies mentioned above, provide strong indications that the younger the child, the greater the exposure.
PBDE levels in human tissues around the world have increased significantly over time. In human blood, milk, and tissues, total PBDE levels have increased by a factor of approximately 100 over the last thirty years. This means that approximately every five years, concentrations in human bodies have doubled (Hites 2004). Research suggests PBDE levels in people in North America are higher compared to levels of people in Europe or Japan (Hites 2004). PBDEs have been found in the breast milk of U.S. women at levels 10 to 100 times higher than those found in Europe (Viberg et al. 2008, Schecter et al. 2003).
A comparison of PBDE levels in breast milk between the United States and Sweden is presented in Figure 1. The data from Sweden show that the average quantities of PBDEs detected in women's breast milk were increasing prior to 1997, with concentrations doubling every five years between 1972 and 1997. The decreasing response in human breast milk concentrations to Sweden’s voluntary phase-out of PBDEs by companies and branches of the government, which began as early as 1990, is noted after 1997. Total PBDE levels in Swedish women's breast milk fell about 30 percent between 1997 and 2000 (Figure 1). In contrast, regions where bans and restrictions have not been established show that PBDE concentrations in breast milk have risen far past Sweden’s 1997 peak (NRDC 2005). For example, concentrations in breast milk from U.S. women remain elevated, with levels surpassing even the peak concentrations observed in Sweden (Figure 1).
The Natural Resources Defense Council (NRDC) compiled data from three previously published studies into the graph presented below as Figure 2. The graph demonstrates mean concentrations of three commonly detected PBDE congeners (BDE-47, BDE-49, and BDE-153) in breast milk and fat tissue samples collected from North America and Europe (NRDC 2005). This data further illustrates elevated exposure to PBDEs in the United States in comparison with other countries.
Michigan Exposure Data
Although Michigan data are limited, some preliminary research and evidence of exposure in wildlife suggest ubiquitous presence and widespread human exposure in the State. In 2003, the Environmental Working Group (EWG) conducted a small biomonitoring study looking for flame retardants and other industrial chemicals in breast milk (EWG 2003). One of the study participants was from Ann Arbor, Michigan; her breast milk contained 24 of the 44 compounds tested. Her results were higher than average for deca-, nona-, hepta-, hexa-, and penta-BDE, in comparison with the twenty women in the EWG study. Decabrominated diphenyl ether, PBDE-209, was found in her breast milk at the level of 0.13 parts per billion of lipid weight, putting her in the 60th percentile of the women in the study (EWG 2003).
In 2007, the Michigan Network for Children’s Environmental Health (MNCEH) participated in a seven-state biomonitoring project coordinated by the Commonweal Biomonitoring Resource Center. Thirty-five people, including five from Michigan, were tested for three groups of chemicals. All of the Michigan participants had detectable deca-BDE and other polybrominated diphenyl ethers in their blood. The graph below (Figure 3) represents the levels of PBDEs found in the blood serum of the Michigan study participants (MNCEH 2007). It should be noted that the study population was small and therefore results are not necessarily representative of exposures throughout the state.
PBDEs are regularly detected in fish caught in the Great Lakes, indicating both widespread contamination of the environment and a potential route for human exposure. Lake Michigan salmon were found to contain PBDEs at levels above 100 parts per billion, “one of the world’s highest concentrations for salmon in open water” (Fields 2005). PBDE levels in Great Lakes walleye and lake trout rose exponentially from 1980 to 2000, doubling every 3-4 years (Zhu, Hites 2004). In 2006, 300 fish from the Kalamazoo River, Muskegon Lake, White Lake, Pentwater Lake, Saginaw Bay, and the waters near the Les Cheneaux Islands in Lake Huron were analyzed for PBDE congeners (Rediske et al. 2009). All fish collected contained detectable levels of PBDE congeners with highest concentrations found in carp and walleye. Elevated concentrations in carp were attributed to their close association with sediments and their niche as bottom feeders. The elevated levels in walleye were attributed to the bioaccumulative nature of PBDEs and the fact that walleye were the top predator in most of the locations (Rediske et al. 2009).
In Michigan, there are currently no advisories that provide consumers guidance on whether it is safe to eat Great Lakes fish based on PBDE levels (Alexander 2008). A study is underway to assess levels of PBDEs in breast milk of fish-eaters residing in southwest Michigan (MDEQ 2008b). Studies such as these may help to establish fish consumption advisory levels in order to provide consumers guidance to minimize PBDE exposure.
Policy Summary and Analysis
There is a need for additional data on human exposure to PBDEs both nationwide and in Michigan, and for a better understanding of health effects associated with these levels of exposure. Children may be more vulnerable to exposure to certain PBDE congeners, although further study is needed to determine the extent and degree of the potential harm. However, current data suggest that (1) concentrations of PBDEs in the U.S. are elevated when compared with other countries, (2) the Michigan population has the potential for additional exposure due to consumption of contaminated fish in addition to dust exposures, (3) body burden studies have indicated that children have higher concentrations of PBDEs than adults, and (4) there is evidence of an association between early-life PBDE exposure, including exposure to deca-BDE, and neurotoxicity. The evidence supports the need to take measures immediately to reduce human exposures to PBDEs to protect children’s environmental health. Some legislative and regulatory measures related to certain PBDEs exist at the state level in Michigan and in other states to protect human health and the environment. Best policy practices from other states are examined, and recommendations made to help reduce children’s exposure to PBDE’s in Michigan.
Michigan Policy Highlights
- By statute signed into law on January 3, 2005, the distribution or manufacture of items containing more than 0.1% of penta- or octa-BDE is banned, with exceptions for “manufacturer replacement service parts” and the processing of recyclables containing penta- and octa-BDEs (MCL 324.14722 & 324.14723).
- The state’s Interdepartmental Toxics Steering Group’s PBDE Sub-committee is authorized to monitor other PBDEs and make recommendations to the Department of Environmental Quality regarding the risks of other PBDEs, which are then to be reported to the legislature (MCL 324.14724). The Committee’s final report calls for a legislative ban on deca-BDE given the availability of safe alternatives, and for consideration of changes in U.S. chemical policy (MDEQ 2008b).
- On Jan. 27, 2010, the Michigan House nearly unanimously passed HB 4699 (94-6), which would phase-out the toxic flame retardant deca-BDE. HB 4699 would stop in-state sales of TVs, computers, mattresses, and residential upholstered furniture containing the toxic flame retardant by 2011. All uses of deca-BDE other than in transportation and military would be banned by 2013 while transportation and military uses would be disallowed by 2014. The Michigan Senate has not taken action on this bill.
Analysis and Policy Highlights from Other States
- Michigan was among the first states to ban the sale of products containing penta- and octa-BDEs. While many states have statutorily required monitoring of deca-BDE research, Washington, Maine, Oregon, and Vermont have passed laws to phase out deca-BDE (Chapter 65, 2007 Laws/ESHB1024 and Public Law, Chapter 296, 2007, SB 596-A, Act 61 Chapter 80: Flame Retardants). As it did for penta- and octa-BDE, the Michigan Senate should pass the House bill to phase out deca-BDE in Michigan.
Evaluation and Recommendations
Michigan is a national leader in PBDE policy because it was one of the first states to ban products with penta- or octa-BDEs. Subsequently, penta and octa were voluntarily withdrawn by the manufacturers. However, to better protect children’s environmental health, Michigan should adopt policy to ban deca-BDE, as was done in Washington, Maine, Oregon, and Vermont. The state’s Interdepartmental Toxics Steering Group’s PBDE Sub-committee (MDEQ 2008b) recommended a ban where safer alternatives were available, as did the Michigan Department of Environmental Quality (MDEQ 2008a). Although there is a voluntary national agreement to phase out deca-BDE, that agreement does not have the force of law, applies only to three deca-BDE manufacturers, and does not apply to products imported with deca-BDE already in them. Summary of Recommendations for PBDE Policy in Michigan
To reduce health hazards posed to children by exposure to PBDEs, Michigan should enact a legislative ban on deca-BDE, as seen in Washington, Maine, Oregon, and Vermont.
International Note:
In developing nations, practices of open waste disposal and burning in urban areas may increase the exposure of children to contaminated dust and soil. A recent study in Nicaragua found levels of PBDEs in serum from children to be comparable with adults in the United States (Athanasiadou et al. 2008). OH-PBDEs (hydroxylated metabolites) also were detected in the Nicaraguan children and it had been shown earlier that OH-PBDEs have greater potency as endocrine disruptors than PBDEs (Hamers et al. 2002). The detection of high levels of PBDEs and their OH-PBDE metabolites in urban children in a developing country highlights the need for a worldwide exposure assessment of emerging pollutants, not just in the developed countries. This is not only of scientific interest, but also a matter of global concern (Athanasiadou et al. 2008).
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