United nations sc

1.5Environmental fate

41. 
    The environmental fate properties of BDE-209 have been assessed in various reports published by the EU, Canada and the United Kingdom (ECB 2002, 2004, ECHA 2012a, ECA 2006, 2010, UK EA 2009). Fugacity modeling predicts that most of the BDE-209 (> 96%) in the environment partitions to sediment and soil (ECA 2010, ECHA 2013a). Less than 3.4% of BDE-209 is expected to be associated with bulk air or bulk water phases. Due to its intrinsic properties i.e. an organic carbon-water partition coefficient (Koc) in the range 150,900 to 149,000,000L/kg, BDE-209 is known to adsorb strongly to organic matter in suspended particles, sewage sludge, sediment and soil (ECHA 2013a). Given its low water solubility and strong particle affinity, its mobility in soils is also likely to be low (ECHA 2013a). Consequently, transfer to other environmental compartments, by soil erosion and run-off, will depend on particle bounded transport. In the environment BDE-209 is persistent and elevated levels is found in soil and sediments.
    
42. 
    BDE-209 is also found to be a major PBDE congener in snow and ice in the Arctic (Hermanson 2010, Meyer 2012), showing that the detected air levels at lower latitudes contributes to long-range transport and pollution in remote areas. BDE-209 is also found in biota, sometimes at high levels, where it along with other PBDEs bioaccumulates and biomagnifies via the food chain (see Sections 2.2.4, 2.3.1. and 2.3.2). As further discussed in chapters 2.2.2 and 2.4.6 debromination of BDE-209 to lower brominated PBDEs in environmental matrices and biota has important implications for the risk from cdecaBDE imposed to the environment, due to the PBT, vPvB and POP properties of its metabolites.
    

1. Persistence

43. 
    Photodegradation and biodegradation are the main mechanisms for transformation of BDE-209 in the environment (ECA 2006, 2010). Due to the lack of any functional groups that are readily susceptible to hydrolysis and a very low water solubility of BDE-209, < 0.1 µg/L at 25 °C (Stenzel and Markley 1997), hydrolysis is unlikely to be a relevant degradation process in the environment (ECHA 2012a). Photodegradation might, however, be a contributing factor for BDE-209 degradation in air and topsoil (see Section 2.2.2). Yet, in the atmospheric compartment, BDE-209 will almost exclusively be adsorbed to air-particles. As air-particles protect the BDE-209 molecule, degradation in air via photolysis is not substantial (see Section 2.2.2).
    
44. 
    High persistency of BDE-209 in soil, sediment and air is demonstrated in several studies and appears to be dependent on slow biodegradation processes and the degree of exposure to light (ECHA 2012a, ECA 2010). The type of particle to which BDE-209 is bound may also influence the degradation rate. For example, studies of photolytic degradation on various solid matrices has revealed half-lifes of 36 and 44 days for BDE-209 adsorbed to montmorillonite or kaolinite, respectively, with much slower degradation occurring when sorbed on organic carbon-rich natural sediment (t1/2 =150 days) (Ahn 2006). Half-life in sand was found to be only 35-37 hours while the corresponding half-lives in sediment and soil were estimated to be 100 and 200 hours, respectively (Söderström 2004 and Tysklind 2001 as cited in ECHA 2012a). In natural waters the presence of other organic substances such as humic substances can limit photodegradation by absorbing light or by hydrophobic interactions with the BDE209 molecule (Leal 2013). Similarly, it has been shown that sand particles coated with humic acid may decrease degradation rates of BDE-209 when irradiated with UV light (Hua 2003). In addition the nature of dissolved organic matter, the amount of suspended particulates, the adsorption of BDE-209 to solid surfaces and the depth are of importance (Leal 2013). Increased adsorption to the soil or sediment matrix with ageing is an additional factor can contribute to longer environmental half-lives under natural conditions (ECHA 2013a).
    
45. 
    Under conditions (e.g. deep sea sediments) where light attenuation and matrix shielding would affect overall exposure to sunlight and potential for photodegradation, the persistency of BDE-209 appears to be high (ECHA 2012a and references therein). Estimations of half-lives in water are generally complicated by the poor water solubility of BDE-209 and are highly dependent on the experimental conditions. Yet, when correcting for the use of solvents and taking into account natural light conditions, environmental half-lives ranging from a few hours up to 660 days in water have recently been suggested (Kuivikko 2007 in Leal 2013). The longest environmental half-life is reported by Tokarz (2008), who by conducting a laboratory microcosm experiment over a period of 3.5 years at 22°C under dark conditions found the half-life of BDE-209 in sediment to range between 6 and 50 years, with an average of around 14 years. The long half-lives found by Tokarz are underpinned by monitoring in the field. Kohler (2008) investigated concentrations and temporal trends of BDE-209 in the sediments of a small lake located in an urban area in Switzerland. BDE-209 first appeared in sediment layers corresponding to the mid 1970s and the levels increased steadily to 7.4 ng/g dry weight (dw) in 2001 with a doubling time of about 9 years. No evidence for sediment-related long-term transformation processes was found in this study covering almost 30 years.
    
46. 
    Further evidence for the persistency of BDE-209 is provided by studies on sludge and soil. Liu (2011a) observed no degradation of BDE-209 after 180 days in soil samples spiked with BDE-209 in darkness. In another study on sludge-amended soil, the extrapolated primary degradation half-life under both aerobic and anaerobic conditions was found to be >360 days assuming exponential decay (Nyholm 2010, 2011, as cited in ECHA 2012a). In a controlled laboratory experiment at 37 °C under dark anaerobic conditions using digested sewage sludge spiked with BDE-209, the concentration of BDE-209 decreased by only 30% during the incubation time of 238 days (Gerecke 2005). These results are further strengthened by field studies. Eljarrat (2008) examined the fate of PBDEs in sewage sludge from five municipal WWTPs after agricultural application of sludge to the topsoil at six sludge application sites and one reference site. According to the authors, BDE-209 concentrations in soil remained high (71.7 ng/g dw) even at one site that had not received sludge applications for four years, illustrating the persistency of BDE-209 in soils. Similarly, Sellström (2005) measured PBDE levels in agricultural soil from sites that had received past sewage sludge amendments and found that levels in a farm soil between 0.015 to 22,000 ng/g dw even though contaminated sewage sludge had not been applied to the soil for many years. The highest levels were detected at a farm site that had not received amendments for 20 years.
    

1. Degradation and debromination

47. 
    In spite of BDE-209’s persistence and long environmental half-lives in sediment, soil and air, there is considerable evidence that BDE-209 is debrominated to lower brominated PBDEs in the abiotic environment, as well as in biota (ECHA 2012a,c, 2013a,b UK EA 2009, ECA 2010, POPRC 2010c, 2013a, NCP 2013). Observed debromination products detected range from mono- to nonaBDEs, and include listed POPs such as tetra-to heptaBDE and bromophenols, as well as recognized PBT/ vPvB substances such as brominated dioxins and furans (PBDD/ PBDF) and hexabromobenzene (Cristiansson 2009, UK EA 2009, ECHA 2012a,c, ECA 2010, see Tables 3.1 to 3.4 UNEP/POPS/POPRC.10/INF5). The biotransformation of BDE-209 in biota, in particular, is considered to be of concern in a number of recent reports and published studies (ACHS 2010, ECHA 2012a,c, EFSA 2011, ECA 2010, POPRC 2010a,b,c, Ross 2009, McKinney 2011a).
    
48. 
    Abiotic degradation studies have shown the formation of nona- to triBDEs (reviewed in ECHA 2012c). The most unequivocal evidence that photodebromination occur in soil, sediment, air and other matrices come from controlled laboratory studies with natural sunlight. While the identity of the degradation products is inconclusive in some studies (Örn 1997, Palm 2003, Gerecke 2006), other studies provide good evidence for the formation of hepta- and hexaBDE congeners in freshly spiked sediment, soil and sand following exposure to light under laboratory conditions (Sellström 1998a, Tysklind 2001, Söderström 2003, 2004, ECHA 2013a, Jafvert and Hua 2001a, Eriksson 2004). Ahn (2006) found that debromination of BDE-209 adsorbed to minerals was a stepwise reaction, initially forming nona-, then octa- and heptaBDE congeners after 14 days exposure to sunlight, but with increased exposure time, hexa- to triBDEs were also formed. A number of studies, while not necessarily being representative of environmental conditions, have shown that microorganisms can influence BDE209 degradation in soil and sediments as they are capable of transforming deca-, nona- and octaBDEs to at least hepta- and hexaBDEs (Robrock 2008, Lee and He 2010, Deng 2011, Qiu 2012). Photodegradation and debromination of BDE-209 have also been studied in abiotic material such as dust, plastic and textile exposed to light, and degradation products have been identified from hexa- to nonaBDE (Stapleton and Dodder, 2008, Kajiwara 2008, 2013a,b). Other degradation products such as PBDD/ PBDF, pentabromophenol and hexabromobenzene can moreover be formed from BDE-209 during processing (recycling), plastics production, photolysis, food preparation (cooking of fish) and waste disposal (Vetter 2012, Kajiwara 2008, 2013a,b, Hamm 2001, Ebert and Bahadir 2003, Weber and Kuch 2003, POPRC 2010b, Thoma and Hutzinger 1987, Christiansson 2009). Formation is strongly dependent on conditions like temperature and purity of the flame retardant.
    
49. 
    Monitoring data provide supporting evidence that degradation of BDE-209 occurs under environmental conditions (ECHA 2012c, Hermanson 2010, Xiao 2012). Results provide evidence of the formation of small amounts of nona- and octaBDEs over a period of 30 days in lake sediments (Orihel 2014 in press, see also ECHA 2012c). A few studies demonstrate degradation of BDE-209 (mainly to nona- and octaBDEs) in sewage sludge (Stiborova 2008, Gerecke 2006, ECHA 2012c), as well as in precipitation (Arinaitwe 2014). Change in the congener ratio in sludge compared to commercial formulations has also been observed (Knoth 2007). Although minimal debromination during WWT was reported in the past (Kim 2013a, Zennegg 2013), findings support the contention that BDE-209 in sewage sludge can be dehalogenated to less brominated congeners (Hale 2012). In soil, debromination of BDE-209 is assisted by the presence of plants (Du 2013, Huang 2010a, 2013, Lu 2013, Wang 2011a, 2014). The distribution pattern of lower brominated PBDEs in plant tissues was different from the soil spiked with BDE-209 suggesting that debromination of BDE-209 in the soil occurred and that further debromination within the plants may occur (Du 2013, Wang 2011a, 2014). An overview of degradation products in abiotic matrices is given in UNEP/POPS/POPRC.10/INF5 Tables 3.3 and 3.4.
    
50. 
    Debromination is also shown in studies with higher vertebrates, including birds, fish and rodents (ECHA 2012a, c, UK EA 2009, ECA 2010, POPRC 2013a). While most vertebrates appear to be able to degrade BDE-209 to lower brominated PBDEs, different species may have different ability to debrominate BDE-209, with debromination occurring more rapidly and to a greater extent in some species than in others (McKinney 2011a). However, the amount metabolised is limited by the amount absorbed and by the metabolic capacity.
    
51. 
    Several laboratory experiments and field studies on fish have shown debromination of BDE-209 after dietary- or water exposure, or after injection of BDE-209 (Kierkegaard 1999, Stapleton 2004, 2006, Kuo 2010, Munschy 2011, Vigano 2011, Noyes 2011, 2013, Zeng 2012, Wan 2013, Feng 2010, 2012, Luo 2013, Bhavsar 2008, Orihel 2014 in press or as reviewed in ECHA 2012c). A number of apparent degradation products have been detected in terms of lower brominated PBDEs ranging from mono- to octaBDE. In several studies congeners (BDE-49, BDE-126, BDE-179, BDE-188, BDE-202) not present in any technical PBDE products have been detected and reported as evidence for biotransformation of BDE-209 (Munschy 2011, Wan 2013, Vigano 2011). The concentrations of BDE209 and its degradation products varied between the different fish species which might be explained by species-specific differences in bioaccumulation capacity and metabolism between the fish species (Stapleton 2006, Luo 2013, Roberts 2011). Formation of hydroxy- and methoxyBDE degradation products have also been reported (Feng 2010, 2012, Zeng 2012).
    
52. 
    A number of studies have also revealed debromination of BDE-209 in birds or bird eggs (reviewed by Chen and Hale 2010, Park 2009, Van den Steen 2007, Letcher 2014, Holden 2009, Munoz-Arnanz 2011, Mo 2012, Crosse 2012). In American kestrels exposed to BDE-209 via diet the half-life of BDE-209 was estimated to 14 days based on plasma concentrations measured during the uptake and elimination periods (Letcher 2014). In addition, debromination products from nona- to heptaBDE were observed. As observed in fish, BDE-202 along with other unidentified congeners not present in c-decaBDE was detected in bird eggs and seen as evidence for debromination (Park 2009, Holden 2009, Mo 2012). Furthermore, the ratio between nonaBDE/BDE-209 congeners in eggs or prey fish was higher than the ratio seen in the commercial mixture indicating biotransformation of BDE-209 in birds/bird eggs (Holden 2009, Mo 2013). The bird egg congener profiles differ markedly from what have been reported in marine and aquatic biota, where lower-brominated (tetra- and penta-BDE) congeners predominate. These differences in congener profiles may be due to lower bioavailability to BDE-209 compared to lower-brominated congeners, debromination- and depletion of BDE-209 in marine and aquatic biota (McKinney 2011a, Huwe 2008). Indication of BDE-209 biotransformation in the terrestrial environment was also shown by the high amounts of BDE-208 in earthworms following exposure to BDE-209 (Sellstrøm 2005, Klosterhaus and Baker 2010). An overview of degradation products in biota are reported in UNEP/POPS/POPRC.10/INF5 Tables 3.1 and 3.2.
    
53. 
    Mammalian data indicate that debromination (nona - hepta-BDE) is the first step in biotransformation of BDE-209, followed by hydroxylation to phenols and catechols (Riu 2008, Wang 2010a, Huwe 2007), and that debromination either occurs in the intestine via metabolism by intestinal microflora or by first-pass metabolism by cytochrome P450 enzymes in the intestinal wall following uptake (Mörck 2003, Sandholm 2003).
    
54. 
    The toxicity of lower brominated congeners is well known, thus debromination of BDE-209 to lower brominated congeners contributes to the outcome of c-decaBDE toxicity (Kodavanti 2011). Environmental degradation and/or biotransformation of BDE-209 showing degradation/transformation to POP BDEs (BDE-47, 99, 153, 154 and 183) has been reported (Wan 2013, Letcher 2014, She 2013, Zhang 2014, Munschy 2011, Stapleton 2004, Feng 2010, Luo 2013, Lu 2013, Huang 2013, see UNEP/POPS/POPRC.10/INF5 Tables 3.1-3.4). Due to the wide distribution and high persistency organisms are continuously exposed, through their lifetime, to a complex mixture of BDE-209, lower brominated BDEs and other BDE-209 degradation products (ECHA 2012), increasing the likelihood for adverse effects (Ross 2009, McKinney 2011a, Kortenkamp 2014). A study by He (2011) showed that long-term chronic exposure to low doses of BDE-209 not only affects F0 growth and reproduction, but also elicits neurobehavioral alterations in F1 offspring. Bioaccumulation of BDE-209 and biotransformation to low concentrations of the congener's nona- to hexaBDE was observed and the risk of mixture toxicity was raised by the authors. Hence, uptake and bioaccumulation of BDE-209 together with its biotransformation to more bioaccumulative and toxic metabolites in organisms has the potential for significant adverse effects as a result of the combined exposures (see Section 2.4.6). Similar results have been reported by Noyes (2011) and Chen (2012a).
    

1. Bioavailability and tissue distribution

55. 
    The bioavailability of BDE-209 is low due to its high molecular weight that affects its passive diffusion through biological membranes (Frouin 2013, Mizukawa 2009) and its strong affinity for particles i.e. sediment and soil (Tian and Zhu 2011, see also Section 2.2.1). In spite of that and as confirmed by monitoring data from across the world (see Section 2.3 and Tables 5.1 and 5.2 UNEP/POPS/POPRC.10/INF5) and available laboratory studies, detectable and sometimes high BDE209 levels have been measured in a wide variety of tissues, species, food webs and top predators.
    
56. 
    As discussed in Section 1.1 the water solubility of BDE-209 is low and bioavailability via direct exposure through aquatic media is reported to be very limited (Ciparis and Hale 2005, Klosterhaus and Baker 2010). However, some evidence for the bioavailability of BDE-209 in medaka following direct water exposure is shown (Luo 2013). Due to its propensity for particle binding BDE-209 is considered to be bioavailable via food and through ingestion of particles, such as dust, sediment, soil or sand (ECHA 2012c). Uptake of BDE-209 through diet was evaluated in fish (Kierkegaard 1999, Stapleton 2004, 2006) and shown to range between 0.02% and 3.2% dependent on the species and whether debromination products of BDE-209 were taken into account when estimating the total uptake. Evidence for bioavailability of BDE-209 in terrestrial species is shown by the many studies showing uptake of BDE-209 into birds (Letcher 2014, Sagerup 2009, reviewed by Chen and Hale 2010), and is underpinned by biomonitoring data evidencing uptake also in other wildlife species, as well as in humans (Section 2.3). In rats, oral absorption is reported to range from 1-26%, inhalation absorption is estimated to be negligible (El Dareer 1987, Mörk 2003, Sandholm 2003, Riu 2008), and in an in vitro experiment dermal absorption was less than 20% (Hughes 2001). Furthermore an in vitro assessment using a human gastrointestinal tract model showed that BDE-209 was bioaccessible (14%) after exposure to indoor dust samples (Abdallah 2012). In the rat and cow the majority of BDE-209 administered is recovered in feces as the original compound (Kierkegaard 2007, Huwe 2008, Riu 2008, Biesemeier 2010).
    
57. 
    Studies have shown that BDE-209 preferentially sequesters to blood-rich tissues such as muscle, liver, intestine, gills (fish), and to lesser extent to adipose tissue (e.g. Shaw 2012, Wan 2013, EFSA 2011, ECB 2002, 2004). The sequestration to blood rich tissues may possibly be explained by BDE-209 binding to proteins (Hakk 2002, Mörck 2003). In Chinese sturgeons, lipids did not play an important role in the distribution of BDE-209 (Wan 2013). BDE-209 was detected at relatively high concentrations in organs participating in absorption, uptake and metabolism such as the liver, gills, and intestines with liver having the highest concentration of BDE-209 followed by gills. Furthermore, the estimated partition coefficients between tissues and blood were higher than those of less brominated BDEs suggesting that the low partition ratios from blood to tissues would lead to high bioaccumulation of BDE-209, especially in absorbing organs (Wan 2013). Since this organism almost stops feeding during the migration period in the Yangtze River it can be assumed that the sturgeon was in a depuration period. A similar pattern was observed in a bioaccumulation study on harbor seals (Shaw 2012). Average hepatic ΣPBDE (tri- to octa-BDE) concentrations were similar to those of the average seal blubber ΣPBDE (mono- to hexa-BDE) as reported in Shaw (2008). In contrast, BDE-209 concentrations in liver were up to five times higher than those in blubber, which is consistent with observations that BDE-209 migrates to perfused tissues such as the liver in biota. In rats, based on organ fresh weights, the highest concentrations were found in adrenals, kidney, heart, liver and ovaries (EFSA 2011, Seyer 2010, Riu 2008). In lactating cows fed naturally contaminated silage BDE-209 was the dominating congener in feed, organs, adipose tissue and feces, but not milk (Kierkegaard 2007). In dietary exposure of American kestrels, higher levels were observed in fat than liver on wet weight basis at the end of the depuration period (Letcher 2014).
    
58. 
    Human data demonstrate that BDE-209 is absorbed and distributed to fat, blood, cord blood, placenta, fetuses and breast milk (Frederiksen 2009a, Zhao 2013; UNEP/POPS/POPRC.10/INF5, Table 4.1). Maternal transfer of BDE-209 to eggs and offspring has also been reported in fish, frogs, birds, rats and reindeer (Vorkamp 2005, Lindberg 2004, Johansson 2009, Garcia-Reyero 2014, Nyholm 2008, Rui 2008, Biesemeier 2010, Cai 2011, Holma-Suutari 2014, Liu 2011c).
    

1. Bioaccumulation

59. 
    In the past, bioaccumulation of BDE-209 in biota was assumed to be low, mostly attributed to the large molecular size and extreme hydrophobicity and low bioavailability of BDE-209 (Hale 2003). However, low bioaccumulation can also be explained by factors such as low uptake and/or high ability to eliminate or metabolize BDE-209 through excretion and debromination of BDE-209 (Hale 2003, Arnot 2010). Furthermore, variable results for bioaccumulation can also be explained by analytical challenges to measure BDE-209 (Ross 2009, de Boer and Wells 2006, Covaci 2003, Kortenkamp 2014) and/or debromination to lower PBDEs. As noted by ECA (2010) a complete evaluation of a substance's bioaccumulation potential should consider the bioaccumulation potential of both the parent substance and its metabolic product(s). Environmental monitoring studies show that BDE-209 is found to be present in a variety of species as well as humans all over the world, and provide supporting evidence for bioaccumulation (see section 2.3.1 and UNEP/POPS/POPRC.10/INF, Table 5.2). The octanol-water partition coefficient (log Kow) values for BDE-209 reported in the literature are highly variable ranging from 6.27 to 12.11 depending on the measurement or estimation method used (CMABFRIP 1997, Dinn 2012, ECA 2010, Kelly 2007, Tian 2012, US EPA 2010, Watanabe and Tatsukawa 1990). While compounds with a log Kow >5 are considered bioaccumulative, chemicals (like BDE-209) with a log Kow > 7.5 are thought to be less bioaccumulative because of predicted declines in dietary absorption potential (Arnot and Gobas 2003). However, results from food web studies show that BDE-209 bioaccumulates (BMF>1) in both aquatic and terrestrial species (Yu 2011, 2013, Wu 2009a, EC 2010 and references therein).
    
60. 
    The BCF for BDE-209 in fish has been estimated to be <5000 with non-appreciable aqueous uptake predicted due to its large size and low water solubility (ECHA 2012 a, ECA 2010). However, BCF is not considered to be a good descriptor of the bioaccumulation capacity of strongly hydrophobic substances, such as BDE-209. BCFs represent processes of chemical absorption by an aquatic organism from the ambient aqueous environment through its respiratory and dermal surfaces with no dietary considerations (Arnot and Gobas 2006). According to the revised OECD guidelines for bioaccumulation studies, testing via aqueous exposure may become increasingly difficult with increasing hydrophobicity. Hence, for strongly hydrophobic substances (log KOW >5 and water solubility below ~ 0.01-0.1 mg/L) a dietary test is recommended (OECD 305, 2012).
    
61. 
    For terrestrial organisms log KOW and BCF are not good predictors of biomagnification for chemicals with log KOA 6 and log KOW>2 (Kelly 2007, 2009) and in terrestrial food chains chemicals with log KOW<5 and BCFs<5000 have been shown to biomagnify. As previously mentioned, the most important exposure route for BDE-209 in aquatic and terrestrial food webs is through the diet (Shaw 2009, Kelly 2007). The accumulated levels of BDE-209 in sediment-associated organisms and filter feeders (mussels, zoo plankton, crustacean, flat fishes, benthic invertebrates and aquatic worms) have been interpreted to be the result of uptake of particles containing adsorbed BDE209 and not as evidence of bioaccumulation, however, uptake of particles is considered to be an exposure route for higher trophic levels in aquatic food webs (Shaw 2009 and references therein). In terrestrial ecosystems, BDE-209 adsorbs strongly to atmospheric particulates (i.e. aerosols) due to its high octanol-air partition coefficient (KOA), and will be deposited to terrestrial vegetation and soil through wet and dry deposition (Christensen 2005, ECA 2010, Mizukawa 2013, Yu 2011). This provides an exposure route for terrestrial organisms that ingest soil or plants as a food source. Therefore, when considering the bioaccumulative behaviour of BDE-209, calculated or measured BAFs, biomagnification factors (BMFs) and trophic magnification factors (TMFs) are believed to give more relevant information than calculated or measured BCFs (Shaw 2009, Kelly 2007, Powell 2013).
    
62. 
    The BAF represents the bioaccumulation of chemicals in an organism by all routes of exposure, including by dietary and ambient sources. Historically, limited data have been available to allow for estimations of BAFs for BDE-209, as numerous studies measured BDE-209 concentration but did not compare these values with ambient levels. As a result, previous assessments found equivocal evidence concerning whether BDE-209 is bioaccumulative (ECHA 2012a,c, ECA 2010, US EPA 2010). Since this time, additional studies have measured BDE-209 BAFs in biota from regions captured by previous assessments, as well as from other areas. In particular in one study by Frouin (2013) the logBAF values based on lipid weight were higher than 6 for aquatic invertebrates thus exceeding logBAFs>3.7 corresponding to a BAF>5000 (He 2012, ECA 2010), fulfilling thereby the criteria on bioaccumulation in Annex D (BAF>5000). In the above mentioned studies the BDE-209 uptakes by aquatic organisms were measured and compared to water concentrations, and logBAFs were estimated.
    
63. 
    BMFs and TMFs from field data show that BDE-209 can biomagnify in several aquatic and terrestrial organisms and food webs (BMFs >1 and TMF>1; see UNEP/POPS/POPRC.10/INF5, Table 3.5 for details). BMFs range between 1-7 in terrestrial organisms and food webs as reported in the scientific literature (Yu 2011, Wu 2009a) and estimated from modeling (Kelly 2007). In a terrestrial food web study spanning several trophic levels BMFs ranging between 1.4 and 4.7 were reported (Yu 2011, 2013). In another field study on frog (influenced both by aquatic and terrestrial environments) focusing on biomagnification from insect to frog, BMFs ranged from 0.8-13.0 depending on gender (Wu 2009a). In aquatic organisms, BMFs ranged between 0.02 and 34. BMFs in seal blubber- and blood have been found to be in the range 0.03-0.06 and 8.3-20.8, respectively (Thomas 2005 as reported in ECA 2010). Other studies report BMFs between 0.2 and 2.2 for harbor seals (Jenssen 2007), 1.5 for marine biota (Baron 2013), and 1.28 for rainbow trout (Stapleton 2006 as reported in ECA, 2010). Furthermore BMFs between 0.1 and 34 were reported in an aquatic food web study by Law (2006). In a number of other aquatic studies BMFs from 1.2 to 5.1 (Mo 2012), 0.67 to 1.3 (Shaw 2009), 0.4 to 1.3 (Poma 2014), 4.8 to 12.7, although there were some uncertainties regarding the food web (Tomy 2009), and between 0.02 and 5 (Burreau 2004, 2006, reviewed in ECA 2010) have been reported. In aquatic food webs TMF values have been reported as 3.6 (Law 2006), 0.26 (Wu 2009b), 0.78 (Yu 2012, as reviewed in ECA 2010) and 0.3 (Tomy 2008). Most of the reported BMFs and TMFs for BDE-209 have all been calculated using muscle (fish, mammals, and birds), whole body (bivalves, zooplankton, and fish) or adipose tissues (fish and mammals). The differences between BMFs and TMFs reported may be species dependent and influenced by overall condition of the organism, diet, exposures and tissue analysed, metabolism, sex, and food web structure.
    
64. 
    The trophic dilution (TMF<1) observed in some studies might be due to biotransformation of BDE-209 through the food webs, since TMFs>1 have been reported for BDE-209 biotransfomation products (Wu 2009b, Poma 2014). Furthermore BMFs of known biotransformation products such as BDE-202, not present in any commercial available PBDE formulations, have been reported (Yu 2011, Mo 2012, Poma 2014). Hence, in some studies, bioaccumulation of degradation products, such as BDE-202, has been observed rather than bioaccumulation of BDE-209 itself.
    
65. 
    The biota-sediment accumulation factor (BSAF) represents the steady state concentration ratio of a contaminant between an organism and sediment, and can provide further insights into bioaccumulation and biomagnification potential. Calculated sediment BSAF values for BDE-209 suggest low biomagnification potential (BSAF<1) in a number of studies (Klosterhaus and Baker 2010, He 2012, La Guardia 2007La Guardia 2012, Sellstrøm 2005,Tian and Zhu 2011, Xiang 2007, reviewed in EC 2010). But some studies show higher sediment BSAFs > 3 suggestive of bioaccumulation potential in some shellfish species (deBruyn 2009, Wang 2009). In the study by deBruyn (2009), BDE-209 concentrations were either low (BSAF≤ 1.48) or below the limits of quantification for most samples except at one reference site where the sediment BSAFs was calculated as 3.53 (deBruyn 2009). However, in a recent study in a soil invertebrate food web BSAFs for BDE-209 ranged from 0.07 to 10.5 following land-application of biosolids. In the same study, BMFs ranged from 0.07 to 4.0, however, there was some uncertainties regarding the N isotope analysis and the authors further concluded that soil contact is likely more important that trophic status in determination PBDE accumulation in soil invertebrates inhabiting sludge-applied sites (Gaylor 2014) (see UNEP/POPS/POPRC.10/INF5, Table 3.5 for details). The interpretation of BSAF values for BDE-209 is complicated by the fact that BDE-209 metabolism varies widely across species, extremely high sediment levels in some of the field studies and problems of getting clean and sediment-free samples for sediment living organisms (ECHA 2012c, La Guardia 2012).
    
66. 
    Some studies have observed higher biomagnification or increased accumulation potential of BDE-209 in terrestrial organisms compared to aquatic organisms (Christensen 2005, Chen and Hale 2010b, Jaspers 2006, Kelly 2007, Voorspoels 2006a). This is expected given the physico-chemical properties of BDE-209 and the differences in toxicokinetics between terrestrial and aquatic organisms as defined by Kelly (2007). Kelly (2007) calculated higher BMFs for BDE-209 among terrestrial carnivores and humans (BMF=8) than for marine mammals (BMF=3) with the lowest BDE-209 BMF values in terrestrial herbivores and aquatic organisms (BMFs=1). This study concluded that BMFs for very hydrophobic chemicals like BDE-209 were higher in air-breathing animals than water-breathing animals due to slower respiratory elimination and slow elimination via urine that was reflective of the compound’s high Koa and Kow values, respectively. Conversely, other studies show that BDE-209 absorption in some teleost fish occurs at a slow rate, which may allow for lower bioaccumulation potential and greater metabolism (or debromination) and elimination of BDE-209 than seen in terrestrial species (Mörck 2003, Stapleton 2004, Kierkegaard 1999). Available data also suggests that there are additional variables influencing the bioaccumulation of BDE-209. For instance, BDE-209 partitions strongly to sediments and soils, is highly persistent, and is the dominant or one of the most predominant PBDEs detected in abiotic compartments of the global environment (see section 2.3.1). This ubiquitous abiotic contamination sometimes at high levels may allow BDE-209 to enter food webs and reach steady state levels in biota despite the large molecular size (MW=959) and high log Kow (Stapleton 2004).
    
67. 
    Little is known about the bioaccumulation of BDE-209 (or other PBDEs) in plants and in herbivores. BDE-209 was examined in a small herbivorous food chain (paddy soils, rice plants, and apple snails) from an electronic waste recycling site in South China. The BMFs for BDE-209 from the rice plant to snails ranged between 1.2 and 6.3 and clearly demonstrated that BDE-209 can bioaccumulate in the plant/herbivore food web (She 2013). On the other hand, bioaccumulation was not observed in a recent laboratory experiment where apple snails were dietary exposed to BDE-209 (reported BAF was <1) (Koch 2014). Hence from these two studies a clear conclusion of bioaccumulation in snails cannot be made. Many of the reported concentrations of BDE-209 in biota are given on a lipid-normalised basis. Although this is common practice when reporting the levels of bioaccumulative substances, in retrospect this might not be the best analytical approach for those substances that do not partition significantly to lipid (OECD 2012), as is possibly the case for BDE-209 (see section 2.2.3). Studies showing a lack of BDE-209 biomagnification in fish and mammals have been based on levels detected in muscles or adipose tissue and/or have been lipid normalised. As discussed in section 2.2.3 evidence shows that BDE-209 preferentially sequesters to blood-rich tissues, such as the liver, intestine, muscle and gills, hence some previous studies may have targeted the wrong tissue and underestimated the bioaccumulation and biomagnification potential of BDE-209 (Stapleton 2004, Voorspoels 2006a, Wan 2013). Recent experimental advances in isotopic ecology have shown that using a fixed Δ¹⁵N value is inappropriate and that escaled Δ¹⁵N value is more meaningful (Hussey 2014). Thus, correction for trophic levels introduces additional uncertainty, especially for well established food webs.
    
68. 
    Debromination of BDE-209 to lower brominated and more bioaccumulative PBDEs after uptake of BDE-209 in organisms (see section 2.2.2) adds to the concern about the use and releases of BDE209, since some PBDEs are already listed in Annex A of the Stockholm Convention for global elimination, and/or are vPvB and PBT substances in the EU (POPRC 2006, 2007, ECHA 2012a, ECA 2010). Studies of toxic effect and debromination to lower PBDEs have been paralleled by observations of bioconcentration and bioaccumulation of BDE-209 (Garcia-Reyero 2014, Noyes 2011, 2013, Kuo 2010). Furthermore the the presence of lower brominated PBDEs in field studies can be due to both debromination of BDE-209 and as a result of direct exposure from c-pentaBDE or c-octaBDE.
    

1. Potential for long-range environmental transport

69. 
    Along with other less brominated PBDEs, BDE-209 is found in various environmental compartments in the Arctic and Antarctic including air, sediment, snow, ice, soil, sediment and biota (UNEP/POPS/POPRC.10/INF, Tables 5.1 and 5.2).
    
70. 
    Several studies have reported that BDE-209 is the predominant or one of the dominating PBDEs in Arctic air (Wang 2005, Su 2007, Hermanson 2010, Hung 2010, Möller 2011b, Meyer 2012, NEA 2014, Salamova 2014). The levels of BDE-209 in the Arctic atmosphere together with studies showing a significant deposition on Arctic ice (Hermanson 2010) and snow (Meyer 2012) underlines the potential of BDE-209 to undergo long-range environmental transport to remote regions. For example, in a study assessing a total of 19 different BFRs in ice core samples from the Norwegian Arctic, BDE-209 was found to provide the second greatest share of the deposition of BFRs from air to the Arctic ice. The deposition rate for BDE-209 was found to be 320 pg cm-2 y-1 in the period 19952005, surpassed only by HBCD, and substantially higher than for other PBDEs (Hermanson 2010). The detection of BDE-209 in Antarctic air and deposition samples provide further evidence of the long-range transport of this compound over remarkable long distances (Dickhut 2012).
    
71. 
    BDE-209 deposited to the Arctic environment is bioavailable to the organisms living there and is widespread in Arctic food webs (de Wit 2006 and 2010, ECA 2010, NCP 2013). Arctic biota samples contaminated with BDE-209 include e.g. vegetation, birds of prey, seabirds and seabird eggs, marine and freshwater fish, different amphipods, zooplankton, shrimps and clams, terrestrial and marine mammals (de Wit 2006 and 2010, Letcher 2010, Tomy 2008). Typically Arctic biota is co-exposed to BDE-209 and a multitude of other PBDEs and POPs (de Wit 2006, 2010, Letcher 2010). Biomonitoring data have shown that BDE-209 contributes significantly to the total body burden of PBDEs in some Arctic species, accounting for >50% of total PBDE burden in detritus feeding ice-amphipods (Sørmo 2006), 60% in redfish and 75% in Arctic cod (Tomy 2008). BDE-209 is also the predominant congener in moss samples from remote sites in Norway (Mariussen 2008).
    
72. 
    BDE-209 is also found in air in remote areas of Asia on the Tibetan Plateau (Xiao 2012, Xu 2011). Snow pack samples in the Tartra Mountains in Slovakia showed remarkably high levels of BDE209 (Arellano 2011). Systematic monitoring at open sea from ships has also proved the abundance of BDE-209 in air samples from the Arctic, Atlantic, Indian, and Pacific oceans (Möller 2012, Möller 2011a,b, Lohmann 2013). Both oceanic and atmospheric processes contribute to the environmental transport of BDE-209 (Su 2007, Möller 2011a,b, Breivik 2006). Since BDE-209 has a very low vapor pressure, volatilization is unlikely to contribute significantly to the long-range environmental transport, rather the atmospheric long-range transport appears to be controlled by the atmospheric mobility of the particles to which it is attached (Breivik 2006, Wania and Dugani 2003). Finer particles (with a diameter around a few micrometres) might remain airborne for hours or days, provided that they are not removed by wet deposition (Wilford 2008, Meyer 2012). Furthermore particles can protect the BDE209 molecule from photolysis and lengthen its life-time in the air to >200 days (Breivik 2006, Raff and Hites 2007 as cited in de Wit 2010). In the Arctic, the deposition of airborne particles is found to be higher during the Arctic haze season (Su 2007, AMAP 2009). In tropical Asia, long-range environmental transport of PBDEs including BDE-209 associated with gas and/or particles is assisted by the monsoon (Xu 2011).
    
73. 
    Based on hydroxyl radical reaction BDE-209 has an estimated atmospheric half-life of 94 days in air according to calculations from the chemical structure using the Syracuse Research Corporation AOP program and assuming a hydroxyl radical concentration of 5x105 molecule cm-3 and a reaction rate of 1.7x10-13 cm3 molecule-1 s-1 (ECB 2002). Other applications such as EPISuite 4.1 (AOPwin module) and PBTProfiler estimate a different reaction rate (3.37x10-14 cm3 molecule-1 s-1) and therefore predict even longer half-lives of 317 days (12 h day, 1.5 x 106 OH radicals cm-3) and 470 days (24 h day, 5x105 molecules cm-3) respectively.
    
74. 
    Although local sources of releases may be present (Hale 2008, Danon-Schaeffer 2007, Li 2012c), the available data from remote regions overall shows that BDE-209 is detected in these areas as a result of long-range environmental transport.