United nations sc

1.6Exposure

1. Environmental levels and trends

75. 
    BDE-209 is widely dispersed in the global environment and is found in biotic and abiotic matrices worldwide. An overview of environmental levels is reported in several reviews (de Wit 2006, 2010, ECA 2010, Letcher 2010, Law 2014) and in UNEP/POPS/POPRC.10/INF5 Table 5.1 and 5.2. In most environmental matrices BDE-209 co-exists with other PBDEs and is the main or one of the dominating PBDEs detected.
    
76. 
    BDE-209 is detected in air in urban, rural and remote regions (UNEP/POPS/POPRC.10/INF5 Table 5.1), as well as in precipitation (Ma 2013, Robson 2013, Arinaitwe 2014 and references therein). In urban and rural environments detected levels range between 4.1 and 60 pg m-3 (as reviewed by Syed 2013) while concentrations in Arctic air ranges from non-detectable to 41 pg m-3 (reviewed by de Wit 2010). Levels in background locations outside the Arctic have been reported to range from nondetectable to 29 pg m−3, i.e. higher than the levels found in the Arctic and lower than the levels found in urban and rural environments (Xiao 2012, Möller 2011a,b, 2012). Recently, however, Lohman (2013) reported particle-bound and gas-phase BDE-209 concentrations in tropical Atlantic Ocean air as high as 43.89 and 260 pg m−3, respectively. Based on these measurements Lohman (2013) calculated the total deposition of BDE-209 to the Atlantic Ocean from air to be approximately 27.5 tonnes annually, 20 and 7.5 tonnes each for the gas- and particle phase, respectively. The findings indicate that air-levels and deposition over the global oceans may be higher than previously thought. As described above, both in the Arctic and the tropics, air-transport of BDE-209 is influenced by seasonal weather phenomena (Xu 2011, Su 2007, AMAP 2009). Air mass back trajectory analysis indicates that major potential source regions of BDE-209 are widely distributed in industrialized and urbanized areas in tropical Asia (Xu 2011).
    
77. 
    Most available data reporting BDE-209 levels in soil are from affected areas. Reported levels in soil worldwide range from non-detectible up to 8600 pg g-1 dw soil in polluted areas, but may possibly be even higher (reviewed by Wang 2010b). BDE-209 was detected in soil at landfill sites in Arctic Canada (Danon-Schaeffer 2007, Li 2012c), but PBDE soil levels outside the landfills were similar to levels measured in soil at background locations elsewhere in the Arctic (de Wit 2006, 2010), suggesting that emissions of BDE-209 and other PBDEs from these sources to the Arctic environment at present are small. Compared to remote sites, BDE-levels in urban and rural areas are significantly higher. In particular, the levels of BDE-209 in soil at e-waste sites such as recycling plants, dumping- and industrial sites in China are very high (Wang 2011b, reviewed by Wang 2010b, Gao 2011, Li 2012a). Sewage sludge from several countries is reported to contain BDE-209 and when soil is amended with sludge BDE-209 is transferred to soil and biota (de Wit 2005, NEA 2012, NERI 2003, Annex E Denmark, Ricklund 2008a,b, Earnshaw 2013). As shown by Sellstrøm (2005) and de Wit (2005) levels of BDE-209 were 100-1000 fold higher at sites fertilized with sewage sludge compared to reference sites. In this study, BDE-209 was the dominant congener in soil and earthworms, with higher levels reported in the worms than in the soil (Sellstrøm 2005).
    
78. 
    Reported BDE-209 levels in sediments worldwide range from non-detectable to 16,000 ng/g dw i.e. slightly higher than in soil (see Wang 2010b, Eljarrat 2007, Sellstrøm 1998b; POPRC/ INF Table 5.1). High concentrations in sediment are typically found in the vicinity of industrial sites (Wang 2010b, Eljarrat 2004, 2005, 2007, Sellstrøm 1998b). Similar to findings in soil, BDE-209 is the predominant congener reported in sediments contributing almost 100% to the total PBDE measured in some studies (Wang 2010b, Eljarrat 2005, Marvin 2013). Levels of BDE-209 in soil and sediment in remote regions are low (de Wit 2006, 2010, CPAN 2010, 2012b, Boitsov and Klungsøyr 2013, SFT 2008a,b), but have been found to be elevated at a few sites affected by local contamination such as at landfills and in the vicinity waste water outfalls (Hale 2008, Danon-Schaeffer 2007, Li 2012c). The doubling time for BDE-209 in these sediments range between 5.3 and 8.4 years (Chen 2007b, Kwan 2014, Zhu and Hites, 2005, Zegers 2003; UNEP/POPS/POPRC.10/INF5 Table 5.1).
    
79. 
    BDE-209 is also found in a variety of terrestrial and aquatic species globally (de Wit 2006, 2010, Letcher 2010, ECA 2010, Chen and Hale 2010, NCP 2013; see UNEP/POPS/POPRC.10/INF5 Table 5.2). Measurements include plants, seabirds such as eider, guillemot, glacous- and herring gulls, different birds of prey, fish and marine invertebrates, marine crustaceans, insects and frogs, as well as marine and terrestrial mammals. BDE-209 is found in a variety of tissues in adults, as well as in eggs of oviparous organisms.
    
80. 
    In the Arctic, reported biota levels range from non-detectable to 250 ng/g lw (UNEP/POPS/POPRC.10/INF/5 Table 5.2). Extrapolation of lab-based effects data to field measurements suggests that the health of fish from the Canadian Arctic is likely being negatively impacted by current decaBDE exposure levels (Noyes 2013, Tomy 2008, 2009). The same applies for fish from south China where measured decaBDE concentrations exceed levels where effects have been observed (Noyes 2013, Mo 2012). In areas more affected by antrophogenic activities levels up to 12000 ng/g lw have been reported. High BDE-209 levels have often been shown in terrestrial environments, and likely reflect the low volatility of BDE-209 and its high affinity to organic matter in dust and soils (Chen and Hale 2010, Chen 2012d). In particular, BDE-209 levels in birds and bird eggs are extensively studied, and have in some instances reported to be very high. Common kestrels from China contain very high BDE-209 levels with concentrations of 2150 and 2870 ng/g lw, in muscle and liver, respectively (Chen 2007a). One specimen contained as much as 6220 ng/g lw in the muscle and 12 200 ng/g lw in the liver. These levels are among the highest BDE-209 levels reported in wildlife (Chen 2007a), and much higher than reported elsewhere (Bustnes 2007, Fliedner 2012, Johansson 2011, Sørmo 2011, Vorspooels 2006b, Gentes 2012, Chabot-Giguère 2013, Mo 2013, Chen 2010a, Chen and Hale 2010). High levels were also reported in kingfishers, Eurasian tree sparrows and common kestrels in China (Chen 2007a, Yu 2011, Mo 2012) and peregrine falcons in Sweden (Johansson 2011). BDE209 is also detected in several species of Arctic birds (de Wit 2006, 2010). Red foxes in urban, rural areas in Belgium had liver levels opp to 760 ng/g lw, where BDE-209 typically contributed to ~70 % of the total PBDE load (Voorpoels 2006a). High levels were also found in Eurasian otters from the UK, with concentrations up to 6800 ng/g lw in the liver (Pountney 2014).
    
81. 
    Limited data on temporal changes in environmental BDE-209 levels is available. In the few studies reporting temporal trend data from abiota in the Arctic, BDE-209 levels in Canadian Arctic air were found to increase from 2002-2005 (Su 2007, Hung 2010), while the same was not observed from 2007-2009 (NCP, 2013). For the period 2002-2005 doubling times in the range of 3.5-6.2 years were reported (Su 2007, Hung 2010). No temporal trends of BDE-209 in air for 2007-2013 can be seen at the Norwegian Arctic sites (i.e. Zeppelin and Andøya observatory) (NEA 2014). Instead the concentrations fluctuate from year to year. In contrast, levels in Antarctic ice are reported as stable and unchanged in the period 2001-2007 (Dickhut 2012). In urban and rural air and precipitation where levels are influenced by both diffuse and point sources, the pattern is more complex and suggests either no significant change (Ma 2013), an increase (Arinaitwe 2014) or a decline (Robson 2013) in BDE-209 levels over time. It should be noted that in most studies on air or precipitation no clear temporal or spatial trends in is reported. Although BDE-209 is stabilized by binding to air particles (de Wit 2010), the lack of any observable time-trends may in some cases possibly result from photolysis/debromination of BDE-209 to lower brominated PBDEs (Wang 2005, Xiao 2012, see also Meyer 2012, Robson 2013, Arinaitwe 2014). With regards to geographical trends, geographical comparisons of air monitoring data are generally hampered by the fact that most studies report episodic monitoring data.
    
82. 
    Concentration- and time trends of BDE-209 in sediment have also been reported. In sediment cores from a remote lake in Switzerland BDE-209 levels increased steadily in the period from 1990s to 2001 with a doubling time of about 9 years (Kohler 2008). Sediment cores from urban/polluted areas show doubling times for BDE-209 between 5.3 and 8.4 years (Chen 2007b, Kwan 2014, Zhu and Hites, 2005, Zegers 2003; UNEP/POPS/POPRC.10/INF5 Table 5.1). Based on a study in the south of China, the time at which BDE-209 started to increase in sediments in China seems to be 10-20 years later than those in North America and Europe, which might reflect differences in historical patterns of production and use of BDE-209 in these continents (Chen 2007b). In dated sediment cores taken from lakes in Ontario, Quebec and northern New York State along a latitudinal transect in North America, BDE-209 was generally detected only in recent sediment horizons, and sedimentation fluxes were found to decline exponentially with latitude (Breivik 2006).
    
83. 
    Paralleling the observation of increasing levels of BDE-209 in Arctic air, Vorkamp (2005) observed a significantly increasing temporal trend in BDE-209 concentration in eggs of peregrine falcon from southwestern Greenland collected from 1986 to 2003. The measured concentrations in this study ranged from 3.8 to 250 ng/g lw, with a median of 11 ng/ g lw. An increace in BDE-209 was also observed in peregrine falcon eggs from Sweden collected from 1974 to 2007 from <4 to 190 ng/g lw (Johansson 2011). BDE-209 levels in UK peregrine falcon eggs were found to increase from 1975 to 1995, and thereafter to decrease from 1995 to 2001 (Leslie 2011). Increasing temporal trends in BDE209 concentrations were also reported in herring gull eggs from the Laurentian Great Lakes in North-America (Gauthier 2008). From 1982 to 2006 BDE-209 doubling times ranged from 2.1 to 3 years. In contrast, steady state levels were reported in eggs of coastal herring gulls in Germany (Fliedner 2012). In the same study, no significant temporal trend was observed between 1973 and 2001 for BDE209 concentrations in sparrowhawk muscle tissue, although some samples with higher concentrations were seen in later years. In Norway, geographical trends in BDE-209 levels are observed in moss (Mariussen 2008). Following a transect from south to north BDE-209 levels in moss is observed to decline, indicating that BDE-209 originating in source regions south of Norway are transported towards the Arctic via atmospheric processes and deposited along the way, resulting in the observed decreasing geographical gradient. BDE-209 levels in moss in Norway also appear to have increased over time likely reflecting similar changes in air levels (SFT 2002, Mariussen 2008, CPAN 2012b).
    

1. Human exposure

84. 
    Dust, indoor air and to a lesser extent food, are considered to be the most important sources and pathways for human exposure to PBDEs (US EPA 2010). In this assessment household consumer products were identified as the main source for the PBDEs in house dust. On the contrary a Canadian assessment identified food and dust as main sources for exposure in adults (HCA 2012). Detected BDE209 concentrations in the indoor air range from
85. 
    BDE-209 is widely present in food and is reported in concentrations ranging from ~2 to >50,000 pg/g ww as reviewed by Frederiksen (2009a). The highest concentrations of lower brominated PBDE were generally measured in fish, and shellfish, while BDE-209 was found in sausage and dairy products, but also food wrapping may contribute (Schecter 2011, EFSA 2011 and Riviere 2014). Contributions from drinking water and outdoor air to indirect BDE-209 exposure are low compared to intakes from food and often considered negligible.
    
86. 
    The internal dose, e.g. assessed using human biomonitoring, reflects an integrated exposure over time comprising various sources and pathways. BDE-209 has been measured in placental samples in concentrations ranging from 0.05 to 8.4 ng/g lw in a Danish and Spanish study, the median were 1.14 and 1.0 ng/g lw, respectively (Frederiksen 2009b, Gomara 2007). Both studies reported BDE-209 to be the dominating PBDE, representing around 50% of the total PBDEs. A similar congener pattern was observed in a recent study from China, where prenatal placental concentrations were in the range of 1.33 to 8.84 ng/g lw (median 2.64 ng/ g lw) (Zhao 2013). Biomonitoring studies on cord blood showed BDE-209 median concentrations to be in the range <1.2 to 27.1 ng/g lw (UNEP/POPS/POPRC.10/INF5, Table 4.1). BDE-209 was in general the largest contributor to the sum of the PBDEs. Exposures to BDE-209 continue in early infancy due to its presence in breast milk. The extensive review by Frederiksen (2009a) covered studies published until 2007 and showed that BDE209 was reported in the concentration range of 0.1 to 2.9 ng/g lw. More recent studies report similar median concentrations (UNEP/POPS/POPRC.10/INF5, Table 4.1), while maximum values vary considerably within and between geographical regions. BDE-209 concentrations in serum or plasma of adult populations with no known occupational exposure were shown to range from 1 to 18.5 ng/g lw (Frederiksen 2009a). More recent studies show similar levels (UNEP/POPS/POPRC.10/INF5, 4.1), except from the strikingly high levels (mean 220 ng/g lw) reported from Laizhou in China, a previous production area of halogenated flame retardants (He 2013). One study from Sweden has assessed the concentration of BDE-209 in serum from first time mothers living in Uppsala sampled from 1996 to 2010 (Lignell 2011). The mean of the 36 serum pools was 1.3 ng/g lw and no significant temporal trend was seen. This is in accordance with the lack of time trend seen for breast milk collected at the Faroe Islands in 1987, 1994-5 and 1999 (Fängström 2005a). In summary, the biomonitoring data show widespread and ongoing exposures to BDE-209 throughout the world, and confirm fetal exposure and absorption in adults.
    
87. 
    Studies on occupational exposure are mostly from Scandinavia and Asia, where high-exposure occupational groups like electronic dismantlers have been the main focus. In Sweden, the median BDE209 blood level in electronic dismantling workers and computer technicians was reported to be 4.8 and 1.53 ng/g lw, respectively (Sjödin 1999, Jakobsson 2002), while a median of 35 ng/g lw were reported among rubber workers (Thuresson 2005). The widespread recycling and dismantling of e-waste under primitive conditions in China has received increasing attention. Median BDE-209 concentrations in Guiyu (Bi 2007) were 50-200 times higher than previously reported in the occupationally exposed populations in Sweden. The highest concentration of BDE-209 in human serum ever reported has been observed in a study by Qu (2007), i.e. 3,436 ng/g lw which is about 3,000 times higher than usually observed in general populations. In contrast, a recent study (Yang 2013) did not find significant difference between the residents in an e-waste recycling area and the reference group.
    
88. 
    The estimated mean dietary intake of BDE-209 for average consumers in Europe ranged from 0.35 (minimum lower bound (LB)¹ to 2.82 ng/kg bw (maximum upper bound (UB)) per day (EFSA 2011). Based on a daily intake of 50 mg dust and a bw of 70 kg, EFSA estimated the exposure of adults to be 0.045 to 7 ng/kg bw per day. Lorber (2008) reviewed exposure to PBDEs in the US and showed that for BDE-209, soil/dust ingestion with 104.8 ng/day made the largest contribution to the exposure, followed by soil/dust through dermal contact (25.2 ng/day). The total exposure was estimated to 147.9 ng/day of which food and drinking water contributed only 16.3 and 0.09 ng/day, respectively. The total exposure corresponds to 2.11 ng/kg bw per day given a body weight of 70 kg as used by EFSA. Health Canada estimated the upper-bound total daily intake of BDE-209 to be 9.3 ng/kg bw for Canadian adults (20-59 years) (HCA 2012). Food and indoor dust were the dominant sources of exposure, contributing 51 and 45% to the total intake, respectively.
    
89. 
    Breast milk concentrations measured in Europe, China/Taiwan, Ghana and India were recently used to estimate mean daily intakes for breastfed infants ≤3 months. The intakes were similar, ranging from 1.0 (LB) to 13.8 (UB) ng/kg bw/day (Kortenkamp 2013). Health Canada (HCA 2012) estimated the total intake of breast fed infants up to 6 months to be between 50-187 ng/kg bw per day, with dust contributing 40 ng/kg bw per day. A study from New Zealand estimated that BDE-209 intakes for infants aged 3 to 6 months was 11.7 ng/kg bw/day, while that in 6 to 12 months old children was estimated to 8.2 ng/kg bw/day (Coakley, 2013). Children 1 to 2 years old had the highest estimated intakes of BDE-209 with 13.2 ng/kg bw/day which likely reflects the high dust ingestion rate (60 mg/day) for this group. The daily intake of decaBDE and other PBDEs from dust and breast milk measured in this study was below US EPA Reference Dose values (7 g/kg bw d, US EPA 2008).
    
90. 
    Several studies show that toddlers and young children have higher levels of PBDEs than adults (Frederiksen 2009a), which was also seen for BDE-209 (Fischer 2006, Lunder 2010, Sahlström 2014). Small children, as a result of their behavior receive considerable PBDE doses from house dust. Assuming a daily ingestion of 100 mg dust the exposure for 1-3 year old children in Europe were estimated to range from 0.53 to 83 ng/kg bw per day, which is higher than the corresponding calculated median dietary intake ranging between 2.59 and 6.4 ng/kg bw (EFSA 2011). Health Canada estimated the daily BDE-209 intake for the age group 0.5 to 4 years to be 89 ng/kg bw of which diet and dust contributed 24 and 64 ng/g kg bw, respectively. Children’s toys, specifically hard plastic toys, have been identified as a potential source of exposure of young children to c-decaBDE (Chen 2009). This exposure was modelled in the assessment of oral intake of BDE-209 of Canadian children for the 0.5- to 4-year age group (HCA 2012). The upper-bound estimate was 120 ng/kg bw per day, which was twice the exposure estimate from soil (dust) for this age group. Congener-to-congener correlations within the mother or toddler cohorts in a Swedish study suggested diet as an important exposure pathway for tetra to nonaBDEs for mothers (Sahlström 2014). For infants breastfeeding was the predominant exposure pathway for tetra-to hexaBDEs and dust the most important exposure pathway for octa- to decaBDEs for toddlers. Despite some geographic differences, all available intake estimates for BDE209 point out the importance of dust exposure, particularly for small children.
    

1.7Hazard assessment for endpoints of concern

91. 
    National and regional assessments conducted by the EU, the United Kingdom, Canada and US have evaluated the potential for c-decaBDE/BDE-209 to induce adverse effects in wildlife and humans (e.g. ECB 2002, 2004, 2007, UK EA 2009, ECHA 2012a, HCA 2006, 2012, US EPA 2008, EFSA 2011). In addition, the toxicity of BDE-209 and other PBDEs has been the topic of several scientific papers and reviews (see e.g. Dingemans 2011, Chen and Hale 2010, Costa and Giordano 2011, Kortenkamp 2014). In the literature toxic effects are reported for soil organisms, plants, birds, fish, frog, rat, mice and humans. Reported effects of BDE-209 range from changes at biochemical- and cellular level to effects which may have more direct implications at higher-levels of biological organization including survival, growth, behavior, immune function, reproduction, development, nervous system and endocrine modulating effects. In vertebrates, the liver, the thyroid hormone (TH) axis and the nervous system appear to be the main targets for BDE-209 toxicity (for review see Costa and Giordano 2011). In both wildlife and humans, early developmental stages appear more vulnerable to BDE-209 exposure than adults. In addition, debromination of PBDEs to more toxic PBDEs is a reason for concern in several assessments (UK EA 2009, ECHA 2012a,c, ECA 2010, Kortenkamp 2014). While some studies either report no effects or effects only at high doses, other studies suggest that BDE-209 may induce adverse effects at low and/or environmentally relevant concentrations.
    

1. Toxicity to aquatic organisms

92. 
    C-decaBDE and its main constituent BDE-209 has limited water solubility, and early hazard assessments suggested that significant acute or chronic toxic effects was not likely to occur in aquatic organisms at concentrations below water solubility (e.g. ECB 2002, 2004, 2007, UK EA 2009). However, the most recent EU assessment of BDE-209 raised a concern for adverse effects also to aquatic organisms based on new studies documenting effects on important biological endpoints including reproduction, development, nervous system, endocrine system, growth and fitness (ECHA 2012 a).
    
93. 
    Aquatic toxicity studies have revealed a number of effects on aquatic organisms, mostly fish and amphibians. Through their influence on the TH system, PBDEs including nona- and BDE-209, was shown to have the potential to affect development and metamorphosis in amphibians (Schriks 2006, 2007, Balch 2006, Qin 2010). According to the available studies BDE-209 and BDE-206, which is one of the congeners present in c-decaBDE and a possible degradation product of BDE-209, can delay metamorphosis in African Clawed Frog tadpoles. In a study by Shricks (2006) a significantly reduced tail tip regression was observed following BDE-206 exposure of tails ex vivo. In a more recent in vivo study a c-decaBDE (DE-83R) consisting of 98.5% w/w BDE-209 was reported to affect metamorphosis in African Clawed Frog tadpoles by delaying the time to forelimb emergence (Qin 2010). The delayed forelimb emergence was accompanied by histological changes in the thyroid gland and reduced expression of the thyroid receptor in tail tissue. Based on this study an aquatic NOEC of around 0.001 mg/L (1 μg/L) for delayed metamorphosis in African Clawed Frog tadpoles was indicated (ECHA 2012a). Studies have also demonstrated that BDE-209, following flow-through exposure to 0 ppb, 0.1 ppb, 10 ppb, and 100 ppb for 12 weeks, can alter the anatomy and function of the African Clawed Frog vocal system by affecting the laryngeal motor neurons when animals are exposed during the androgen sensitive critical period of vocal system development and during adulthood when the tissues are utilizing androgens to vocalize (Ganser 2009). In this study BDE-209 also inhibited maletypical vocalization, a critical aspect of mating behavior by reducing the number of calls elicited as well as the average call amplitude. The data suggest that BDE-209 can alter anatomy and function, mediated through pathways that include blocking the androgens necessary for proper vocal system. These findings may be of concern given that wild frogs are exposed to BDE-209 already at the egg stage and that BDE-209 in frogs also is transferred to brain and testis (Liu 2011c, Wu 2009a).
    
94. 
    In fish, controlled feeding studies with fathead minnows conducted at environmentally relevant concentrations have shown that BDE-209 either alone and/or in combination with its debromination products may interfere with the TH system in adult and juvenile fathead minnow (Noyes 2011, 2013). In the latter study, adult fish dietary exposed to a low dose of ~3 ng/g BDE-209 bw per day for 28 days showed a 53% and 46% decline in circulating total thyroxine (TT4) and 3,5,3'-triiodothyronine (TT3), respectively, compared to controls (Noyes 2013). In fish exposed to a high dose of 300 ng BDE-209/g bw, the levels of TT4 and TT3 were lowered to 62 and 59%, respectively. Both in high and low-dose exposed fish, TH levels remained supressed after a 14-day depuration period. Both doses also reduced brain deiodinase activity (T4-ORD) with 65% compared to control. Similarly, the study by Chen (2012a) indicated that BDE-209 has the potential to cause adverse effects in zebrafish at early life stages with impacts on T3 and T4 concentrations. Li (2011) observed changes in expression of TH-associated genes in rare minnow larvae and adults following exposure to 0.01- 10 µg/L BDE-209 via water for 21 days. In contrast to these findings, Thienpont (2011) and Garcia-Reyero (2014) report no visible effect on thyroid function in exposed fish embryos. However, it should be noted that Thienpont (2011), who exposed 48 hours post fertilization embryos to 960 μg/L BDE-209 for three days conclude that the assay used, a T4 immunofluoresence quantitative disruption test, was not suitable for detecting effects of chemical pollutants such as BDE-209 that indirectly disrupt thyroid gland function. Garcia-Reyero (2014) speculated that the absence of effects on the TH-system in their study may be explained by shorter exposure and/or lower doses than those used by Noyes (2011) and Chen (2012a). Potential TH disruption in fish by several PBDEs was also investigated in vitro by Morgado (2007) with negative results. In this study neither BDE-209 nor BDE-206 showed any binding to sea bream transthyretin (TTR), a TH binding protein in the blood. The result suggests that BDE-209 likely does not interfere with binding of TH to TTR.
    
95. 
    Other effects, both chronic and acute, have also been observed in fish following exposure to BDE-209. In the above dietary study from 2013, Noyes observed a significant increase in percent cumulative mortality as well as a decline in gonadal-somatic index. Chen (2012a) observed significant decreases in body weight and survival rate of zebrafish larvae exposed to 1.92 mg/L BDE-209 via water for 14-days. Significant changes were not observed at any of the lower exposure doses tested (0, 0.08, 0.38 mg/L).
    
96. 
    Based on measurements of otolith increment widths in juvenile lake whitefish (~5 months old) fed BDE-209-spiked diets (control, 0.1, 1, and 2 µg/g-diet) there were indications that BDE-209 may affect growth rates in fish at environmentally relevant levels of BDE-209 found in sediment (Kuo 2010, de Wit 2002).
    
97. 
    He (2011) documented effects on overall fitness, reproductive parameters and behavior as well as motor neuron and skeletal muscle development in a low dose chronic toxicity study with zebrafish. Several of the effects reported by He (2011) were trans-generational i.e. they were observed in offspring of exposed parents and are according to the authors likely explained by maternal transfer of BDE-209. In male fish, indicators of sperm quality were significantly affected even at the lowest exposure dose (0.001 µM or 0.96 µg/L).
    
98. 
    Potential reproductive toxicity of BDE-209 was also demonstrated in rare minnow (Li 2011). In this study, reduction of spermatocytes and inhibition of spermatogenesis was demonstrated in adult rare minnow exposed to 10 µg BDE-209/L via water. Changes in the expression of TH and spermatogenesis associated genes in rare minnow larve and adults were observed following exposure to 0.1-10 µg BDE209/L. In addition effects on body length and gonadosomatic index of adult females were observed at 10 µg/L, but no significant histological changes were found in the ovary at any of the concentrations tested. Furthermore, no change in mortality or body length of larvae and adult males was observed.
    
99. 
    In the above mentioned study (Garcia-Reyero 2014) BDE-209 impacted expression of neurological pathways and altered the behavior of zebrafish larvae, although it had no visible effects on TH function or motor neuron and neuromast development. In this study fish were exposed to BDE-209 spiked sediment, at a concentration of 12.5 mg/kg. Concentrations in exposed larvae and solvent control measured after 8 days were 69.69.8 ng/g ww and 6.70.5 ng/ g ww, respectively.
    
100. 
     Besides the other effect reported above, BDE-209 was shown to induce oxidative stress in the liver of goldfish. A reduction in glutathion level and in the activity of antioxidant enzymes, (glutathione peroxidase, superoxide and catalase) was observed from 7-30 days after a single intraperitoneal injection of 10 mg/kg (Feng 2013a, b).
     
101. 
     In several of the above fish studies BDE-209 was reported to debrominate to lower brominated PBDEs (Noyes 2011, 2013 Chen 2012c, Kuo 2010, He. 2011), thus it is possible that other PBDE congeners besides BDE-209 contributed to the effects reported in these studies. Reported debromination products included nona-, octa-, hepta-, hexa- and pentaBDEs.
     
102. 
     In summary, the lowest aquatic NOEC for exposure via water reported appears to be below 0.001 mg/L (1 μg/L) and was observed for delayed metamorphosis in amphibians. Based on Noyes (2013) a LOEL of ~3 ng/g BDE-209 bw/day or 0.41 ng/g ww food can be derived for TH disruptive effects and mortality in fish. Overall the aquatic toxicity data suggest that BDE-209 can have adverse effects on critical endpoints such as survival, growth, fitness, reproduction, development, somatic maintenance, thyroid hormone homeostasis and neurological function. The data, moreover, add to the concern regarding the bioaccumulation potential of BDE-209 and debromination in organisms in the environment, since they show that the accumulation of BDE-209 can lead to adverse effects in vulnerable life stages of mammals, fish and amphibians (Chen 2012a, He 2011, Noyes 2011). The levels used in some of the experiments were comparable to levels in more polluted areas (Zhang 2010a, Wang 2011b).
     

1. Toxicity in soil organisms and plants

103. 
     Toxicity data are available for soil microorganisms, plants and earthworms. Most of the published data is new and was not reviewed in any of the previous risk assessments and evaluations (e.g. ECHA 2012a, UK EPA 2009, ECA 2006). Based on a plant toxicity study by (Porch and Krueger, 2001) and two 28- and 56-day toxicity studies with earthworms, ECB (2002) reported that no effects were seen on plants at concentrations up to 5,349 mg/kg dry weight. and that a NOEC ≥ 4,910 could be derived for earthworms. Based on these results and using an assessment factor of 50 PNEC values for soil of 98 mg/kg dry and 87 mg/kg wet weight were estimated.
     
104. 
     Xie (2011) observed a significant increase in hydroxyl radical levels in earthworms at 0.0110 mg/kg of BDE-209, which is within the range of environmental levels reported in soil (Syed 2013). The effect was paralleled by oxidative damage to protein and lipids and a reduction in antioxidant capacity. In this study oxidative stress and oxidative lipid damages were observed at concentrations as low as 0.01 mg/kg (Xie 2011). In a more recent acute earthworm study by the same authors, effects on behavior, survival, growth and reproductive parameters were investigated following exposure to 0.1-100 mg/kg BDE-209 for 48 hours and 28 days. Except for a significant decrease in the number of juveniles per hatched cocoon and non-signifcant changes in avoidance response at 1000 mg/kg BDE-209, no other effects were reported suggesting that adult earthwoms have a strong tolerance for BDE-209 in soils, but that a potential toxicity exist for earthworm embryos or juveniles (Xie 2013b).
     
105. 
     In ryegrass seedlings exposed to 100 mg/kg BDE-209 Xie (2013a) observed a 35% inhibition of root growth and 30% decrease of the chlorophyll b and carotenoid contents of leaves. No other visual signs of toxicity were observed, but BDE-209 exposure induced oxidative stress and damage, altered the activity of several antioxidant enzymes and reduced the non-enzymatic antioxidant capacity at concentrations starting from 1 mg/kg. Sverdrup (2006) did not observe effects on nitrifying bacteria, red clover seedling emergence or survival and reproduction of soil invertebrates at concentrations up to 1,000 mg BDE-209/kg spiked soil and speculated that the absence of toxicity could be due to the low water solubility of BDE-209.
     
106. 
     In summary, BDE-209 appears not to be acutely toxic to plants and soil organisms and adverse effects are generally observed at high doses (ECB 2002, Sverdrup 2006, Xie 2013 a,b). However, new data suggests that toxic effects of BDE-209 in some instances may occur at lower doses (0.01-1 mg BDE-209/kg) than previously shown (Zhu 2010, Liu 2011a, Zhang 2012, 2013c, Xie 2011, 2013a).
     

1. Toxicity in birds

107. 
     As highlighted by Chen and Hale (2010), birds exhibit some of the highest concentrations of BDE-209 reported in wildlife and may be at risk for experiencing adverse effects (ECHA 2012a, see also UNEP/POPS/POPRC.10/INF5, Table 5.2). However, a limited number of studies examining adverse effects of BDE-209 exposure to birds are available.
     
108. 
     In a study on swallows nesting at a WWTP, a positive relationship between egg size and BDE209 levels were found, however, no significant correlation were found for reproductive parameters (Gilchrist 2014). BDE-209 concentrations were not reported.
     
109. 
     Sifleet (2009) observed a mortality of up to 98% in embryos of captive chicken injected with a single dose of 80 μg BDE-209 /egg and exposed for 20-days via the yolk sac. The reported LD50 from this study was 44 μg/egg (740 μg kg ww). An assessment undertaken by the EU, revealed that the BDE209 concentrations typically found in wild bird eggs are around 2-10 times lower than the concentrations that according to Sifleet (2009) induce mortality (ECHA 2012 a). Reported concentrations in bird eggs typically range between 1-100 μg/kg ww, but up to 420 g/kg ww have been reported (ECHA 2012a). In spite of important study limitations, the EU risk assessment indicated that the margin between exposure levels in wild birds and observed effect levels is not high, especially considering that Sifleet (2009) did not take into account potential sub-lethal effects, and that additional BDE-209 would likely have been assimilated following hatching and desorption of the remaining yolk thereby further increasing exposure.
     
110. 
     A reduction in body mass was observed in European starlings exposed to BDE-209 by silica implants (van den Steen 2007).
     
111. 
     Birds are reported to metabolize BDE-209 to lower brominated PBDEs, including some POPBDEs (BDE-183) (Letcher 2014 ) and exposure to lower brominated PBDEs have been associated with immunomodulatory changes, developmental toxicity, altered reproductive behavior, reduced fertility and reproductive success (for overview see Chen and Hale 2010, Glichrist 2014, POPRC 2007). In a study on captive American kestrels exposed to DE-71, a commercial penta-PBDE mixture, at environmentally relevant levels in ovo, the low BDE-209 levels present (<2.5 %) was found to be associated with an increase in flight behavior of male kestrels both in the courtship period as well as during brood rearing later in life (Marteinsson 2010). The BDE-209 concentrations measured in this study were not reported. These findings suggest that BDE-209 like other PBDEs may affect behavior in birds and is consistent with research on laboratory rodents where some studies report that BDE-209 causes changes in spontaneous behaviour. For a comprehensive discussion on behavioral effects in rodents see chapter 2.4.4 below.
     

1. Toxicity in terrestrial mammals

112. 
     The toxicity of c-decaBDE to terrestrial mammals has mainly been investigated in rodents. Although several effects are reported including reproductive toxicity, data in particular point to neurodevelopmental toxicity and effects on the TH-system. In addition available scientific evidence suggests that BDE-209 either alone or in concert with other PBDEs could act as a developmental neurotoxicant in terrestrial mammals and humans (Dingemans 2011, Messer 2010, Kicinski 2012, Costa and Giordano 2011, HCA 2006, 2012, Gascon 2012, Chao 2011, Kortenkamp 2014).
     
113. 
     Developmental neurotoxicity is the reported critical endpoint of several PBDEs (Blanco 2013, Branchi 2002, Eriksson 2001, Kuriyama 2005, Rice 2007, 2009, Suvorov 2009, Viberg 2003, 2004, 2007, Xing 2009, Zhang 2013a, UNEP/POPS/POPRC.10/INF5, Table 6.1). Several mechanisms for developmental neurotox effects are proposed, for instance impaired thyroid homeostasis, direct toxicity to neuronal and stem cells, and disturbing neurotransmitter systems (Costa 2014). Developmental neurotoxicity has also been reported for BDE-209 in some studies (Johansson 2008, Viberg 2003, 2007, Rice 2007, 2009, Fujimoto 2011, Heredia 2012, Chen 2014, Reverte 2013, 2014, Buratovic 2014, Mariani 2014) but not by others (Biesemeier 2011). Mariani (2014) reports neurodevelopmental effects of BDE-209 in mice at dose levels relevant for pregnant women. Neurobehavioral effects of BDE-209 in rodents during juvenile development or adulthood have also been reported more recently (Buratovic 2014, Heredia 2012, Chen 2014, Reverte 2013, 2014). For example, long-lasting effects in spatial learning and memory were observed in transgenic mice after postnatal exposure to BDE-209 and reduction in anxiety levels and delayed learning in spatial memory tasks were found in wild type mice (Reverte 2013, Heredia 2012). In another study a single dose of BDE-209, administered orally at post natal day 10, was also observed to cause long-lasting effects on emotional learning and TH-levels in mice carrying two variants of apolipoprotein E, apoE2 and E3 (Reverte 2014). Moreover, Chen (2014) reported that prenatal BDE-209 exposure in rats impaired learning acquisition in a dose dependent manner, and in vitro data suggested that this impairment in rat learning acquisition may be linked to effects on brain neurogenesis.
     
114. 
     The majority of developmental studies with BDE-209 used oral administration, but only a few were designed according to the OECD 426 guideline “Developmental Neurotoxicity Studies” (OECD 2007). In mice and rats administered a single dose of BDE-209 during the “brain growth spurt” period consistent and persistent alterations in behavior, habituation and memory were observed by Viberg (2003, 2007) and Johansson (2008). Other researchers (e.g. Hardy 2008, 2009, Goodman 2009, William and DeSesso 2010) have noted limitations with the former studies, in particular for not using the litter as a basis for the statistical evaluation. Despite this, the US EPA used the studies from Eriksson and Viberg in their derivation of oral reference doses for BDE-209 (as reviewed in US EPA 2008). A study conducted by Rice (2007) did not show a consistent depression in motor activity over time in mice, however their follow-up study showed neurobehavioral long-lasting deficits when tested at 16 months (Rice 2009). Similar to the findings of Viberg (2003, 2007) and Johansson (2008), behavioral effects from developmental BDE-209 exposure appeared to get worse with age. Additional evidence for neurodevelopmental effects of BDE-209 come from several publications that indicate that PBDEs affect the cholinergic system in both mouse and rat brain which could lead to disturbed cognition (learning and memory) (Fischer 2008a, Viberg 2003, 2007, Liang 2010, Buratovic 2014). In further support of findings indicating that BDE-209 can act as a neurotoxicant in mammals, Fujimoto (2011) showed that BDE-209 exposure resulted in reductions in the neural connections between the left and right brain hemispheres (the corpus callosum area) and that it caused irreversible white matter hypoplasia targeting oligodendrocytes in rats. This effect was accompanied by developmental hypothyroidism. In contrast, no clinical signs, or any neurobehavioral changes, effects on startle response, or learning behavior were reported at any dose level by Biesemeier (2011), where motor activity and behavior of BDE-209 exposed rats was assessed at two, four, and six months of age. The Biesemeier study has, however, since been critically evaluated by Shibutani (2011) who noted the omission of measurement of thyroid-related effects, histopathological parameters on neuronal migration, oligodendroglial development, discussions of the significant decreases in the hemisphere height and decrease in the pons and cortex vertical thicknesses. The Biesemeier study has also been discussed in the Health Canada (HCA 2012) report, where lower LOAEL and NOAEL values have been suggested instead of the value reported in the original study.
     
115. 
     In line with the findings of Chen ( 2014), other studies show that BDE-209 can exert direct toxic effects on neuronal cells (reviewed by Dingemans 2011, UNEP/POPS/POPRC.10/INF5 Table 6.2) and interfere with neuronal signalling, neuronal development and induce oxidative stress and apoptosis (Chen 2010b, Huang 2010b, Al-Mousa and Michelangeli 2012, Hendriks 2012, Liang 2010, Xing 2010, Mariani 2014) effects that may lead to neurotoxicity and interfere with learning and memory by affecting long-term potentiation as shown by Viberg (2008) and Xing (2009). BDE-209 is further shown to cause changes in gene expression, intracellular protein levels, and disturbance of synaptogenesis and cell differentiation (Pacyniak 2007, Viberg 2008, 2009, Zhang 2010b, Song 2013, Mariani 2014).
     
116. 
     In addition to neurotoxic effects available data point to BDE-209 and lower brominated PBDEs as potential endocrine disruptors. PBDEs structurally resemble THs, and as indicated earlier, effects on the TH system (TH: T4 and T3), along with the above mentioned and more direct toxic effects to neuronal cells is suggested as underlying mechanisms of BDE-209 and PBDE neurotoxicity (Ahmed 2008, Gilbert 2012, Dingemans 2011). In vitro (Hamers 2006, Ibhazehiebo 2011, Ren 2013b) and in vivo studies assessing TH/TSH effects due to BDE-209 administration (UNEP/POPS/POPRC.10/INF5 Table 6.3) show that BDE-209 and other PBDEs interfere with the TH-system, but the results on BDE209 or c-decaBDE mixtures are not consistent in terms of what effects are observed. For instance, whereas most animal studies report decreased T3 levels following high BDE-209 exposures (Lee 2010b, Chi 2011, Fujimoto 2011), also no change (Wang 2010a, Zhou 2001), and increase in T3 levels has been reported (Van der Ven 2008, Wang 2011c). For T4, animal studies report both decreases in T4 levels at high dose (Rice 2007, Kim 2009, Chi 2011, Fujimoto 2011) as well as no change in T4 level (Tseng 2008, Van der Ven 2008, Wang 2010a,2011c, Lee 2010b, Zhou 2001). For TSH, two animal studies performed with BDE-209 (Kim 2009, Lee 2010b) both report increased TSH levels at the highest BDE-209 exposures while no effects were reported in adult rats dosed with commercial cdecaBDE mixture DE-83R at doses of 0.3-300 mg/kg/day for four days. Repeated dietary administration of BDE-209 (at a high dose) induced thyroid follicular cell hyperplasia in male mice but not in female mice or in either sex of rats (NTP 1986). Studies reporting significant changes in TH/TSH levels in rats and mice have often administered BDE-209 at doses that are orders of magnitude higher than human exposures. However, studies on rodent offsprings have indicated that low doses of BDE209 may adversely affect the developing thyroid organ (Kim 2009, Lee 2010b, Fujimoto 2011). The recent WHO/UNEP report (2013) concluded that endocrine disruptors can cause adverse effects at low environmental levels, may display non-monotonic dose-responses, and that the timing of exposure can be more ciritical than the level of exposure. Thus the observed inconsistencies in reported TH/THS effects may possibly, at least in part, be explained by differences in the experimental conditions used in these studies.
     
117. 
     Studies suggest that in utero exposure to BDE-209 at high parental doses may cause reproductive toxicity and lead to developmental abnormalities such as decreased anogenital distance and testicular histopathological changes, sperm-head abnormality, and sperm chromatin DNA damage (Tseng 2006, 2013, van der Ven 2008). Effects on testicular development has also been reported following exposure at post natal days 1-5 at low doses (Miyaso 2012). Reported low-dose effects (0.025 mg/kg, subcutaneously) included reduction in testicular weight, sperm count, elongated spermatid and sertoli cell numbers as well as changes in protein expression and phosphorylation status.Also possible modulation of sex steroids in the male (van der Ven 2008) and female (Hamers 2006, Gregoraszczuk 2008) genital system cannot be entirely ruled out. In contrast, no reproductive toxicity was observed in Sprague-Dawley pregnant female rats exposed to BDE-209 from gestational day 019 (Hardy 2002). Similarly, Ernest (2012) reported that a mixture composed of three commercial BDEs (52.1% DE-71, 0.4% DE-79, and 44.2% decaBDE-209) affected liver and thyroid physiology but not male reproductive parameters in exposed rats. Yet, in female mice adrenals, decreased activity in the dehydroepiandrosteron synthesis assay was observed indicating reduced CYP17 enzyme activity and potential effects on steroid hormone production (van der Ven 2008). Further, BDE-209 can inhibit estradiol-sulfotransferase in vitro (Hamers 2006), which could implicate a (local) increase of endogenous estradiol in vivo. In another in vitro study Gregoraszczuk (2008) found that BDE-209 exposure led to increased testosterone-, progesterone- and estradiol secretion in porcupine ovary cells, a finding that suggests that BDE-209 can induce preterm luteinization in antral follicles followed by the disruption of ovulation.
     
118. 
     Oxidative stress and impaired glucose homeostasis has been reported in rats exposed to BDE209. Dose-related fasting hyperglycemia was observed in adult rats exposed to BDE-209 (0.05 mg/kg) for 8 weeks (Zhang 2013e). Reduced insulin levels and increased levels of tumor necrosis factor- (TNF-alpha) were observed in plasma followed by reduction in the oxidative stress markers glutathione and superoxide dismutase. Dose-dependent morphological changes such as blurring boundaries among pancreatic islet cells were observed (Zhang 2013e). Van der Ven (2008) also observed insulitis in male rats in a 28-days exposure study however, no differences were observed between the exposure groups. Similar to the reported effects on the steroid and TH systems the observed effects on glucose homeostasis/ insulin levels are suggestive of the endocrine disruptive potential of BDE-209.
     
119. 
     Immunotoxic effects of BDE-209 have been reported in some studies (Teshima 2008, Watanabe 2008, 2010, Zeng 2014), although immunotoxicity is not regarded as a critical toxic endpoint of PBDEs in general. In the most recent of the studies showing that BDE-209 can act as an immunotoxicant, reduced qualitative and quantitative CD8 T-cell response was observed in mice after long-term BDE209 exposure (Zeng 2014). In contrast to these studies, van der Ven (2008) reported no immunotoxic effects on the T cells in rats.
     
120. 
     Gene mutations are suggested not to occur after exposure to BDE-209 or other PBDEs (Anderson 1990, EFSA 2011, HCA 2012, JETOC 2000, Kirkland 2005, NTP 1986), although recent studies have indicated that BDE-209 may cause DNA damage through the induction of oxidative stress in vitro (Ji 2011, Tseng 2011). There is limited evidence for carcinogenicity of BDE-209 in experimental animals (EFSA 2011, HCA 2012). According to the NTP report (1986) there is some evidence at high dose levels for an increase in liver adenoma in rats and liver adenoma and carcinoma in mice, but this may be related to a secondary mode of action (EFSA 2011).
     

1. Human toxicity

121. 
     A number of studies have assessed the risk of BDE-209 and other PBDEs to humans. The primary focus has been on assessing the risk for developmental neurotoxicity, which is generally considered as the most critical effect in mammals.
     
122. 
     The observation, as outlined in section 2.3.4, that exposure takes place already during the early phases of human development i.e. in utero via placental transfer and postnatal via mothers milk (e.g. Gómara 2007, Kawashiro 2008, Wu 2010, Miller 2012, Mannetje 2013, Coakley 2013), support the notion that the developmental neurotoxicity observed in mammalian models could have implications also for humans. The risk for implications to human health is further underpinned by epidemiological data. Although having a limited number of individuals, studies have shown an association between BDE-209 levels in cholostrum and lower mental development scores in children 12-18 months of age (Gascon 2012), and that human prenatal or postnatal exposure to BDE-209 delays cognitions and potentially affects neurological development (Chao 2011). Furthermore, several epidemiological studies support that exposure to PBDEs may result in human neurodevelopmental toxicity (Harley 2011, Hoffman 2012, Herbstman 2008, Chevrier 2010, 2011, Gascon 2011, Roze 2009, Eskenazi 2013, Schreiber 2010). Some human studies also observed associations between TH/TSH levels and exposure to BDE-209 or other high congeners such as BDE≥183 (Huang 2014, Zota 2011, Wang 2010c, see UNEP/POPS/POPRC.10/INF5 Table 6.4).
     
123. 
     A risk characterization and a hazard and dose-response assesment of BDE-209 suggested that the daily intake of BDE-209 in the USA and Canada was not likely to result in neurodevelopmental toxicity for infants (Health Canada, 2012, US EPA 2008, 2010). EFSA also concluded that current dietary exposure or the intake of BDE-209 by breast-fed infants does not constitute a health concern in the EU (EFSA 2011). Among the four PBDEs (BDE-47, BDE-99, BDE-153 and BDE-209) investigated by EFSA, a potential health concern with respect to current dietary exposure was only identified for BDE-99 (EFSA 2011). A recent PBDE risk assessment based on oral, dermal, and inhalation exposure of infants 0-5 years of age, indicates no risk for adverse health effects in infants that are restrained in a car seat (Fowles and Margott 2013). However, these assessments do not consider the possibility that several PBDEs could act in concert, inducing additive or synergistic effects as suggested by the available in vitro data, or that there may be multiple sources of exposure (e.g. Pellacani 2012, Tagliaferri 2010, Llabjani 2010, Karpeta and Gregoraszczuk 2010, Hallgren and Darnerud 2002, He 2009).
     

1. Mixture toxicity and combined effects of multiple stressors

124. 
     In the environment, the exposure and response to toxic compounds as well as the likelihood for adverse effects is influenced by a number of factors besides the inherent properties of the compound. Such effects include environmental temperature, salinity and pH, the physiological status of the organisms, toxicokinetic processes, food web or trophic structure, environmental transport, partitioning, transfer mechanisms and deposition (for overview see Letcher 2010, Schiedek 2007, AMAP 2011, POPRC 2013b). Climate change impact on ecosystems may also have an effect on several of these factors and hazardous chemicals can affect the ability of organisms to adapt to climate changes and endure their physical environments (AMAP 2003, POPRC 2013b, UNEP/AMAP 2011, NCP 2013). In addition, wildlife and humans are typically not only exposed to BDE-209 alone but rather to a complex mixture of multiple PBDEs as well as other POPs (de Wit 2006, 2010, Kortenkamp 2014, EFSA 2011, NCP 2013). Thus when considering the likelihood for adverse effects to humans and wildlife all these factors need to be considered and may provide additional reasons for concern.
     
125. 
     While the mixture toxicity of BDE-209 and other PBDEs have not been studied experimentally to a large extent, a combination of BDE-47 and -99 was observed to induce synergistic cytotoxic effects in neuronal cells (Tagliaferri 2010). Furthermore, a mixture of PBDE congeners (BDE-47, -99, -100 and -209) at levels detected in human blood had irreversible effect on hormone secretion in ovarian follicles (Karpeta 2010). The results from this study suggest that combined effects of PBDEs may be much larger than indicated by the sum of the effects of the individual congeners. The presence of other POPs may also affect the toxicy of PBDEs. In an in vitro study with binary mixtures of PCBs and PBDEs (BDE-47, -153, -183, or 209), Ljabljani (2010) found that PCB-126 and PBDEs could mutually inhibit each other while PCB-153 and PBDEs jointly could exacerbate the observed biochemical alterations. PBDEs are considered to be potential endocrine disruptors which may act additively at low concentrations UNEP/WHO (2013).
     
126. 
     Further indications for possible mixture toxicity between PBDEs is provided by Kortenkamp (2014), who evaluated the likelihood and type of combined effects between BDE-209 and other PBDEs to humans and wildlife based on concentration addition using the hazard index approach. The common modes of action are not yet fully established; however, common adverse outcomes are established. The study finds that it can be expected that BDE-209 and other PBDEs may produce combined developmental neurotoxicity both in humans and wildlife (Kortenkamp 2014). For humans, the study shows that by taking into account the combined exposure to PBDEs the tolerable combined exposures are exceeded for all age groups, particularly for small children. Though the authors indicate that further research is necessary, the mixture risk assessment nonetheless indicates that combined exposures to BDE-209 and other PBDE likely pose significant health concerns, especially for young children of age 6 months to 3 years which bear the highest PBDE exposures of all age groups. The study also indicates a risk from combined PBDE exposure to wildlife, including Arctic top predators such as the polar bear. Relative to the other PBDEs BDE-209 was found not to make a significant contribution to the overall risk to wildlife. Overall, the study shows that a consideration of BDE-209 in isolation, without taking into account co-exposure to other BDEs, would underestimate the risk. A similar concern was also indicated by Villanger (2011a,b, 2013) who demonstrated that organohalogen contaminant mixtures including several PBDEs (BDE28, -47,-99,-100 and -153) may influence the thyroid homeostasis in Arctic marine mammals. Though the impact of BDE-209 was not assessed in these studies and only correlations were reported, the study, similar to the findings of Kortenkamp (2014), raises concerns that PBDEs due to a similar mode of action may act in concert and induce adverse toxic effects and thereby pose a threat to Arctic marine top predators and other wildlife.
     
127. 
     Available studies also suggest a risk to birds arising from exposure to a combination of different PBDEs and other environmental pollutants. In a field study Plourde (2013) observed that concentrations of the hexa-, hepta-, octa- and BDE-209 congeners (BDE-154, -183, -201 and -209) in liver and BDE209 in plasma of male ring-billed gulls breeding in the urbanized Montreal region were negatively correlated with trabecular and cortical bone mineral density of the tarsus. The finding suggests that the PBDEs at the levels reported in these birds (i.e. liver BDE-209 2.74-283 ng /g ww and PBDE 26.2680 ng/g ww, plasma BDE-209 0.70-19.1 ng/g ww and PBDE 3.55-89.2 ng/g ww) can negatively affect bone tissue structure and metabolism in birds. In another study, the combined effects of several organochlorine pesticides, PCBs and PBDEs including BDE-209 and several nonaBDEs were postulated to have contributed to the death of weakened individuals of glaucous gull found in the breeding seasons 2003-2005 on Bjørnøya in the Barents Sea (Sagerup 2009). However, BDE-209 was only detected at very low concentrations in liver and brain (
128. 
     Additional concerns relate to multiple stressor effects i.e. possible combined effects between toxic chemicals and other factors. Iodine deficiency, a common condition worldwide (reviewed by Walker 2007), can increase the sensitivity to adverse effects from thyroid-disrupting chemicals such as BDE-209 (see Dingemans 2011). Second, exposures to thyroid-disrupting chemicals, including BDE209 and other PBDEs, may also impair the ability of vertebrates to adequately respond to the climate change impact on their environment (Hooper 2013, POPRC 2013b). Third, climate change and elevated temperatures may increase degradation, and long-range environmental transport of BDE-209 (POPRC 2013b, IPCC 2007, NCP 2013, Xu 2011, Christensen 2014).
     

3.