ࡱ > * bjbjUU 4 ? ? * * (8 >8 R8 R8 R8 $ v8 v8 v8 P 8 ? v8 K fB , J ( J J J LN Q R 6 8 8 8 8 8 8 ) X 8 R8 R N N @ R R 8 R8 R8 J J < Pk Pk Pk R n R8 J R8 J 6 Pk R 6 Pk Pk J ܞM v8 U ʈ ( 0 K $ # f f # P # R8 " Q R R Pk R R R R R 8 8 Pk P R R R K R R R R # R R R R R R R R R * 6 :
DECABROMODIPHENYL ETHER
RISK PROFILE
Second Draft
(31 March 2014)
Table of Contents
To be completed.
Executive summary
Commercially available decabromodiphenyl ether (c-decaBDE) is a synthetic chemical product consisting of decabromodiphenyl ether (BDE-209, e"9 0 % ) , w i t h s m a l l a m o u n t s o f n o n a b r o m o d i p h e n y l e t h e r a n d o c t a b r o m o d i p h e n y l e t h e r . C - d e c a B D E h a s b e e n u n d e r i n v e s t i g a t i o n f o r i t s p o t e n t i a l h e a l t h a n d e n v i r o n m e n t a l i m p a c t s f o r m o r e t h a n a d e c a d e b u t i s s t i l l e x t e n s i v e l y u s e d i n m a n y g l o b a l r e g i o n s .
C - d e c a BDE is used as an additive flame retardant. It has a variety of applications including in plastics/ polymers/composites, textiles, adhesives, sealants, coating and inks. C-decaBDE containing plastics are used in housings of computers and TVs, wires and cables, pipes and carpets. It is used in commercial textiles, mainly for public buildings and transport, and in textiles for domestic furniture in countries with stringent fire safety regulations.
Emissions of c-decaBDE to the environment occur at all its life cycle stages, but are assumed to be highest during service-life and in the waste phase. Emissions from industrial point sources can also be significant. Use of c-decaDBE in the production of textiles and electronics lead to transboundary air pollution, either directly from articles or during their disposal stage. BDE-209 has low water solubility (< 0.1 g/L at 24 C), adsorbs strongly to organic matter and in the environment readily partitions to sediment and soil. It is very persistent and reported environmental half-lives in these media typically exceed 180 days.
BDE-209 is widespread and is one of the most dominant PBDEs in the global environment. When detected it is typically found along with other PBDEs originating from other commercial PBDE formulations or from debromination of c-decaBDE. Monitoring data show high concentrations of BDE-209 in sediments and soil as well as in biota worldwide. Levels are generally highest in the urban regions, near waste water discharges and in areas around electronic waste and recycling plants. In air, BDE-209 binds to particles that protect the chemical from photolytic degradation and it can be transported over long distances. The estimated atmospheric half-life is 94 days, but can exceed 200 days. Hence BDE-209 is also detected in environmental and biological samples from remote regions. Temporal trend data show increasing levels of BDE-209 in Arctic air and in some Arctic organisms.
BDE-209 has long been thought to have limited bioavailability because of its large size that constrains its ability to pass cell membranes via passive diffusion. However, biomonitoring data shows that BDE-209 is bioavailable and is taken up by humans and other organisms. BDE-209 has been found in a variety of different organisms and biological matrices including human blood serum, cord blood, placenta, fetus, breast milk and in milk of lactating cows. In some species, particularly birds and amphibians, reported levels are high and close to reported adverse effect concentrations. In rodents and birds, small amounts of BDE-209 can cross the blood-brain barrier and enter the brain. There is also evidence of transfer of BDE-209 from adult stages to eggs in fish and birds. For humans, the available intake estimates for BDE-209 also point out the importance of dust exposure, particularly for small children. Higher levels of PBDEs and BDE-209 are reported in toddlers and young children than in adults. In aquatic organisms intake via diet appears to be the most important exposure route.
The available bioaccumulation data for BDE-209 are equivocal, but several lines of evidence indicate that BDE-209 is bioaccumulative, at least in some species. The equivocation in the available bioaccumulation data largely reflects species differences in uptake, metabolism and bioaccumulation potential, as well as differences in exposure and also analytical challenges in measuring BDE-209. When considering the bioaccumulative behaviour of BDE-209, calculated or measured bioaccumulation factors (BAFs), biomagnification factors (BMFs) and trophic magnification factors (TMFs) are believed to give more relevant information than calculated or measured bioconcentration factors (BCFs). BAF>5000, BMFs and TMFs for a variety of species are established and range between 0.02 - 34 and 0.2 - 3.6, respectively, demonstrating the bioaccumulation potential of this substance. Traditionally, amounts of BDE-209 in biota are given on a lipid normalised basis. However it has recently been indicated that BDE-209 is generally associated with blood and blood-rich tissues, hence the reported lipid normalized BMFs and TMFs might be underestimated.
Debromination of BDE-209 in environmental matrices and biota to more persistent, toxic and bioaccumulative PBDEs such as already listed POP-BDEs is considered to be of high concern in a number of assessments of BDE-209. Several PBDE congeners that are not part of any commercial mixture have been identified and are considered to provide evidence for debromination of BDE-209. In addition c-decaBDE can be a source for environmental releases of toxic dioxins and furans and possibly also hexabromobenzene. Due to debromination of c-decaBDE and past releases of commercial penta- and octabromodiphenyl ether organisms are typically co-exposed to a multitude of PBDEs.
On the basis of common modes of action and common adverse outcomes, there is concern that BDE-209 and other PBDEs may act in combination, in an additive or synergistic manner and induce developmental neurotoxicity in both humans and wild organisms at environmentally relevant concentrations. BDE-209 toxicity studies provide evidence for potential adverse effects to reproductive health and output in a number of species as well as developmental and neurotoxic effects. Observed effect concentrations for increased mortality in birds and developmental effects in frogs derived from controlled laboratory studies raise concern that adverse effects may occur at environmentally realistic concentrations.
Similar to other PBDEs, BDE-209 may act as an endocrine disruptor and affect thyroid hormone homeostasis. The high persistence of BDE-209 in sediments and soils, combined with the fact that organisms are exposed to a wide range of PBDEs and that BDE-209, like other endocrine disruptors, can elicit adverse effects even at low environmental levels increase the likelihood for long-term adverse effects when organisms are continuously exposed.
Based on the available evidence it is concluded that decabromodiphenyl ether (commercial mixture, c-decaBDE) is likely, as a result of its long-range environmental transport, to lead to significant adverse human health and environmental effects, such that global action is warranted.
Introduction
On 13 May 2013, Norway as a Party to the Stockholm Convention, submitted a proposal to list decabromodiphenyl ether (commercial mixture, c-decaBDE) as a Persistent Organic Pollutant (POP) under Annex A, of the Convention. The proposal was submitted in accordance with Article 8 of the Convention and was reviewed by the POP Review Committee (POPRC) at its ninth meeting in October 2013. The proposal may be found in document UNEP/POPS/POPRC.9/2.
Chemical identity of the proposed substance
The risk profile concerns commercial decabromodiphenyl ether (c-decaBDE) and its degradation products, in accordance with para (c) of Annex E of the Convention. C-decaBDE is a chemical product that is widely used as an additive flame retardant in textiles and plastics, additional uses are in adhesives and in coatings and inks (ECHA 2013b). Commercially supp l i e d d e c a b r o m o d i p h e n y l e t h e r c o n s i s t s p r e d o m i n a n t l y o f t h e d e c a b r o m o d i p h e n y l e t h e r c o n g e n e r ( B D E - 2 0 9 ) ( e"9 7 % ) , w i t h l o w l e v e l s o f o t h e r b r o m i n a t e d d i p h e n y l e t h e r c o n g e n e r s s u c h a s n o n a b r o m o d i p h e n y l e t h e r ( 0 . 3 - 3 % ) a n d o c t a b r o m o d i p h e n y l e t h e r ( 0 - 0 . 0 4 % ) . C h e n et al. (2007a) reported that the octaBDE and nonaBDE content of two commercial decabromodiphenyl ether products from China was in the range 8.2 to 10.4 % suggesting that a higher degree of impurities may be found in some commercial mixtures. Historically a range of 77.4-98 % of decaBDE, and smaller amounts of the congeners of nonaBDE (0.3-21.8 %) and octaBDE (0-0.85%) has been reported (ECHA 2012 a, US EPA 2008, Georgalas et al. 2014, manuscript in prep.). Total tri-, tetra-, penta-, hexa- and heptaBDEs are typically present at concentrations below 0.0039 % w/w (ECB 2002, ECHA 2012 a). Trace amounts of other compounds, thought to be hydroxybrominated diphenyl compounds can also be present as impurities. In addition, polybrominated dibenzo-p-dioxins and polybrominated dibenzofurans (PBDD/Fs) as impurities in some c-decaBDE products have been reported (Ren et al. 2011).
According to available information c-decaBDE is currently available from several producers and suppliers globally (Ren et al. 2013a, Georgalas et al. 2014, manuscript in prep.) and is being marketed under different trade names (Table 1).
Chemical data on the main component of c-decaBDE, BDE-209, are presented in Figure 1 and in Tables 1 and 2 below (ECHA 2012 a). Like other PBDEs, BDE-209 shares structural similarities with PCBs. Chemical data on octa- and nonaBDE, which are minor constituents of c-decaBDE, are provided along with other supplementary information in a supporting document for the risk profile UNEP/POPS/POPRC.10/INFxx. Information on c-decaBDE degradation products and their POP-properties are described in sections 2.2.2, 2.4.6, and 3, and in UNEP/POPS/POPRC.10/INFxx.
In this document BDE-209 refers to the single fully brominated PBDE, which elsewhere sometimes is also denoted as decaBDE. The abbreviation c-decaBDE is used in this document for technical or commercial decaBDE products.
Table 1. Chemical identity of c-decaBDE and its main constituent BDE-209
CAS number:1163-19-51 CAS name:Benzene, 1,1'-oxybis[2,3,4,5,6-pentabromo-]IUPAC name:2,3,4,5,6-Pentabromo-1-(2,3,4,5,6-pentabromophenoxy)benzene EC number:214-604-9EC name:Bis(pentabromophenyl) etherMolecular formula:C12Br10OMolecular weight:959.2 g/moleSynonyms:decabromodiphenyl ether, decabromodiphenyl oxide, bis(pentabromophenyl) oxide, decabromo biphenyl oxide, decabromo phenoxybenzene, benzene 1,1 oxybis, decabromo derivative, decaBDE, DBDPE2, DBBE, DBBO, DBDPOTrade names
DE-83R, DE-83, Bromkal 82-ODE, Bromkal 70-5, Saytex 102 E, FR1210, Flamecut 110R. FR-300-BA, which was produced in the 1970s, is no longer commercially available (Environment Canada, 2010).
1In the past CAS no. 109945-70-2, 145538-74-5 and 1201677-32-8 were also used. These CAS no. have now formally been deleted, but may still be in practical use by some suppliers and manufacturers.
2DBDPE is also used as an abbreviation for Decabromodiphenyl Ethane CAS no. 84852-53-9.
Table 2. Overview of relevant physicochemical properties of c- decaBDE and its main constituent BDE-209
PropertyValueReferencePhysical state at 20C and
101.3 kPaFine, white to off-white crystalline powderECB (2002)Melting/freezing point300-310CDead Sea Bromine Group, 1993, cited in ECB (2002)Boiling pointDecomposes at >320CDead Sea Bromine Group, 1993, cited in ECB (2002)Vapour pressure4.6310-6 Pa at 21CWildlife International Ltd,
1997, cited in ECB (2002)Water solubility<0.1 g/L at 25C (column elution method)Stenzel and Markley, 1997, cited in ECB (2002)n-Octanol/water partition coefficient, Kow (log value)6.27 (measured generator column method)
9.97 (estimated using an HPLC method) MacGregor & Nixon, 1997, and Watanabe & Tatsukawa, 1990, respectively, cited in ECB (2002)Octanol-air partition coefficient Koa (log value)13.1 Kelly et al. 2007
1.2 Conclusion of the POP Review Committee regarding Annex D information
The POP Review Committee examined the proposal by Norway to list c-decaBDE in the Stockholm Convention on Persistent Organic Pollutants as well as additional scientific information provided by members and observers at its ninth meeting and concluded that the screening criteria were fulfilled (decision POPRC-9/4).
1.3 Data sources
The risk profile addresses the information requirements given in Annex E of the Convention and further elaborates on, and evaluates, the information referred to in Annex D.
The risk profile is not an exhaustive review of all available data, but rather, it presents the most critical studies and lines of evidence with relevance to the criteria in Annex E of the Convention. It centres on the main constituent of c-decaBDE, BDE-209, and its degradation products, in particular lower brominated PBDEs which are formed via abiotic and biotic degradation (described in 2.2.2). As several of the lower brominated PBDE degradation products are widely recognized as PBT/vPvB substances and/or POPs a re-assessment of their properties was considered redundant (POPRC 2006, POPRC 2007, ECHA 2013a,b, Environment Canada 2010, Table xx, UNEP/POPS/POPRC.10/INFxx). The risk profile, while mainly providing information on the less described BDE-209 congener, also, therefore, provides information on what debromination products are formed (chapter 2.2.2) and the risk for combined effects resulting from co-exposure to multiple PBDEs (chapter 2.4.6). The risk profile was developed using the Annex D information submitted by Norway in 2013 and Annex E information submitted by parties and other stakeholders including non- governmental organizations as well as industry. Sixteen countries (Bulgaria, China, Canada, Japan, Morocco, Serbia, Sweden, Denmark, the Netherlands, Germany, Austria, New Zealand, Mexico, Croatia, Argentina and USA) and two observers (the Bromine Science and Environmental Forum (BSEF) and the International POPs Elimination Network (IPEN)) submitted information under the Annex E process. All Annex E submissions are available on the Convention website (HYPERLINK "http://www.pops.int"www.pops.int).
Updated scientific literature obtained from scientific databases such as ISI Web of Science and PubMed was assessed as well as "grey" literature such as government reports, risk- and hazard assessments, industry fact sheets etc. To provide the best possible overview of the existing data/ literature which covers more than 984 reports and peer-reviewed scientific publications (Kortenkamp et al. 2014), an emphasis was put on providing excerpts of existing risk assessments and reports when such information was available as well as more detailed descriptions of newer literature.
In the past, assessments of c-decaBDE were conducted and published by the EU, Canada, the United Kingdom and the United States (ECB 2002, ECB 2004, ECHA 2012a, Environment Canada 2006, Environment Canada 2010, UK EA 2009, US EPA 2008). The EU risk assessment, which examines in depth the PBT/ vPvB properties of BDE-209, was conducted over a period of more than ten years (ECHA 2012b) and is the most up to date of these assessments.
1.4 Status of the chemical under other international conventions and forums
C-decaBDE has been under scrutiny for its potential health and environmental impacts for more than a decade. Steps to restrict the use of c-decaBDE have been taken in several countries and regions, as well as by some of the major electronic companies (for an overview: UNEP/POPS/POPRC.9/2, Ren et al. 2011).
In 1992, c-decaBDE and other brominated flame retardants were given priority in the OSPAR action plan and in 1998 BDE-209 along with the other PBDEs was included in the list of "Chemicals for Priority Action" as well as in the Joint Assessment and Monitoring Programme in OSPAR. Based on the generated monitoring data OSPAR has promoted actions in the EU on use restrictions for PBDEs, risk-reduction strategies for c-octaBDE, c-decaBDE and HBCD, and waste legislation. An OSPAR Background Paper on Certain Brominated Flame Retardants (Polybrominated Diphenylethers, Polybrominated Biphenyls, Hexabromocyclododecane) was prepared by Sweden in 2001 and updated in 2004 and 2009 (OSPAR 2009).
In 1995, OECD Member countries agreed to oversee a voluntary industry commitment (VIC) by the global manufacturers of brominated flame retardants, among them c-decaBDE, to take certain risk management actions. The VIC was implemented in the United States, Japan and Europe. Compliance with the VIC is on-going. In parallel to this work, OECD conducted an investigation of the waste management practices in member countries with respect to products containing brominated flame retardants. The results of this investigation are documented in the Report on the Incineration of Products Containing Brominated Flame Retardants (OECD 1998). A SIDS Initial Assessment Profile (SIAP) on BDE-209 was prepared under the Environment, Health and Safety (EHS) Programme of the OECD. The SIAP, prepared by France (Sponsor Country for Human Health), the United Kingdom (Sponsor Country for Environment) and the European Commission was adopted by SIAM 16 and later endorsed by the OECD Joint Meeting in 2003). The Hazard/Risk Information Sheets for decaBDE and four other Brominated Flame Retardants were updated in 2005, 2008 and 2009 (OECD 2014).
PBDEs, including BDE-209, are listed as chemicals of concern in the WHO/UNEP State of the science of endocrine disrupting chemicals, and the level of evidence for contributing to several adverse human health and wildlife effects is evaluated in the report (UNEP/ WHO 2012).
Summary information relevant to the risk profile
Sources
Production, trade, stockpiles
Global industrial c-decaBDE consumption peaked in the early 2000's (BSEF 2010 as cited in Earnshaw et al. 2013). Yet, due to limited regulatory restrictions, c-decaBDE is still extensively used worldwide. Past production data indicate that about 75 % of all the world production in PBDEs was c-decaBDE (Georgalas et al. 2014 manuscript in preparation). Total production of c-decaBDE in the period 1970-2005 was between 1.1-1.25 million tonnes, similar to the scale of production of PCBs (POPRC 2010, Breivik et al., 2002)
Globally, the total market demand for c-decaBDE differs considerably between continents. In 2001 the c-decaBDE use in America and Asia was comparable at 24,500 and 23,000 metric tonnes, while the use in Europe and the rest of the world was estimated to be 7600 and 1050 tonnes each (BSEF 2006 as cited in ACAP 2007). In recent years production and trade has declined in most western countries. In China, production of c-decaBDE increased by 200% between 2000 and 2005, from 10,000 to about 30,000 metric tons (Chen et al. 2007b).
Production facilities for brominated flame retardants (BFRs) exist in all global regions (e.g. Georgalas et al. 2014 manuscript in preparation). At present it is unknown how many of these produce c-decaBDE. Main producers of BFRs include Israel, Jordan, China, Japan, and western European countries including Austria, Belgium, France, Germany, Netherlands, and United Kingdom (ACAP 2007). Among the main BFR producing countries China, India and Japan are known to produce and export c-decaBDE (Xiang et al. 2007, Chen et al. 2007b, Xia et al. 2005, Zou et al. 2007, Annex E info Japan, IPEN, China). The overall scale of the c-decaBDE production in countries that still produce it is currently unknown.
The domestic demand for PBDEs in China was reported by Mai et al. in 2005 to have increased at an annual rate of 8%. However, whether there has been an increased demand and production in China also after 2005 is somewhat unclear. According to some sources, production in China increased significantly in the first half of the 2000's and reached between 30,000 and 41,500 tonnes in 2005 (Xiang et al. 2007, Chen et al. 2007b, Xia et al. 2005, Zou et al. 2007, Ni et al. 2013). In 2011, the reported domestic production in China was 20,500 tonnes (Ni et al. 2013). According to Annex E information from China there is at present one production facility in China which produced 21,000 tonnes of c-decaBDE in 2013. Thus, production and demand in China may have already reached its peak. Around 20 different Chinese companies claim to be suppliers of c-decaBDE (Annex E submission IPEN).
Japan produces an estimated 600 tonnes of c-decaBDE per year (Annex E submission Japan). In 2013 the c-decaBDE import to Japan was 1,000 tonnes. In 2002, eleven years earlier, the demand for c-decaBDE in Japan was 2200 tonnes/year and the stock level was about 60,000 tonnes (Sakai et al. 2006).
In Europe production of c-decaBDE ceased in 1999, but c-decaBDE is still imported in considerable quantities (ECB 2002, ECHA 2012 a, c, Georgalas et al. 2014 (manuscript in prep.)). The annual EU consumption is estimated to have ranged from 2500 to around 12,000 tonnes per year over the past decade, and is assumed to have reached its peak in the 1990's (Georgalas et al. 2014 (manuscript in prep.), Earnshaw et al. 2013, OSPAR 2009, BSEF 2006 as cited in ACAP 2007, ECB 2002). From 1970 to 2010, it is estimated a total of 185,000 to 250,000 tonnes of c-decaBDE was consumed in Europe (Earnshaw et al. 2013). According to ECHA 2012c, there is very little information on the tonnages that may be imported in mixtures (chemical formulations, also resins, polymers and other substrates) and articles (either in semi-finished articles, materials or components, or in finished products).
In the United States the two U.S. producers of c-decaBDE, and the largest U.S. importer, in 2009 voluntarily committed to end production, importation, and sales of decaBDE for most uses in the United States by December 31, 2012, and to end all uses by the end of 2013. In 2012 the national production volume which includes both domestic production and import was 18,110,826 lb/ year or around 8215 tonnes/ year.
A voluntary agreement with industry is also in place in Canada, where the three main manufacturers have committed to voluntarily phase-out all exports to Canada by 2013. Manufacture of decaBDE, octaBDE and nonaBDE was banned in Canada earlier in 2008.
C-DecaBDE is not produced in New Zealand, and no information is currently available on the amount imported and in which form(s) other than in manufactured articles (Annex E submission New Zealand). C-DecaBDE is also not produced in Morocco (Annex E submission Morocco) or in Norway where production, import, export, use and placing on the market is forbidden.
Uses
C-decaBDE is a general purpose additive flame retardant, that is physically combined with the material in which it is used to inhibit the ignition and slow the rate at which flames spread. It is compatible with a wide variety of materials. Applications include plastics/polymers/composites, textiles, adhesives, sealants, coatings and inks (e.g. ECHA 2012c, ECHA 2013a, Georgalas et al. 2014 (manuscript in prep.), Sakai et al. 2006).
End uses in plastics/polymers include e.g. housings of computers and TV sets, wires and cables, pipes and carpets (BSEF 2013, US EPA 2012b). Typically c-decaBDE is used in plastics/polymers at loadings of 10-15% by weight, though in some cases loadings as high as 20% have been reported (ECHA 2012c). In a Japanese study, c-decaBDE is reported to account for about 98 % of the bromine content found in plastic parts of older TVs (Tasaki et al. 2004). BDE-209 is also found in products made from recycled plastics, including food contact materials (Samsonek and Puype, 2013).
In textile applications, c-decaBDE is used in commercial textiles, mainly for public buildings and transport and in domestic furniture textiles in countries with stringent fire safety regulations (BSEF 2013). In the textile sector, c-decaBDE can be used to treat a wide range of synthetic, blended and natural fibres (ECHA 2013a). Main end uses are upholstery, window blinds, curtains, mattress textiles, tentage (e.g. military tents and textiles, also commercial marquees, tents and canvasses) and transportation (e.g. interior fabrics in cars, rail passenger rolling stock and aircraft). The most common method of applying flame retardants to textiles is back coating. The amount that is applied will usually be in the range 7.5 20%. Padding processes and printing processes may also be used to apply flame retardant treatments (ECHA 2012a, c).
Information from public consultations in Europe indicates that c-decaBDE can be used in adhesives in the aeronautic sector for civil and defense applications (ECHA, 2012d). Norwegian Authorities also identified uses of c-decaBDE in the adhesive layer of reflective tapes on work wear which are used as fire fighter uniforms, by the staff at oil platforms, and in the energy sector, etc (Climate and Pollution Agency Norway, 2012). The reflective tapes contained decaBDE in the range 1-5 % (by weight of reflective material). C-decaBDE applied in coatings and inks can lead to end-uses in electrical and electronic equipment, and to end-uses in the building and construction sector (Georgalas et al. 2014 (manuscript in prep.)).
According to VECAP data, textiles and plastics account for 52 % and 48 % of the volume of c-decaBDE sold in Europe, respectively (VECAP 2012). In Japan 91 % of the c-decaBDE on the market is used in plastics (Sakai et al. 2006). The remaining 9 % is applied as a flame retardant in textiles. The consumption of c-decaBDE in the United States could be broken down as follows (excluding import in articles): automotive and transportation 26%, building and construction 26%, textiles 26%, electrical and electronic equipment (EEE) 13% and others 9% (Levchick 2010).
2.1.3 Releases to the environment
As an additive flame retardant, c-decaBDE is not chemically bound to the product or the material in which it is used. It therefore has the potential to leak to the surrounding environment from products either during service-life or upon becoming waste. Emissions of c-decaBDE to the environment may however occur at all its life cycle stages, e.g. during production, formulation and other first- and second-line uses at industrial/professional sites, service life of articles, and their disposal as waste (ECHA 2012c). The release and distribution of decaBDE to the environment is confirmed by monitoring data (see Section 2.3.1-2.3.4).
Although controlled product testing has indicated low or no emission of BDE-209 from products (Kemmlein et al. 2003, Kemmlein et al. 2006), a number of available studies assessing exposure and releases under real-life conditions suggests otherwise. BDE-209 is reported to be the most prevalent PBDE congener in house dust and indoor air (e.g. Harrad et al. 2010, Fredriksen et al. 2009a, Besis and Samara 2012, Fromme et al. 2009, Coakley et al. 2013, EFSA 2011), which again is a source of releases to the environment (Bjrklund et al. 2012, Cousins et al. 2014) and to human exposure (see Section 2.3.4). Higher levels of decaBDE are typically reported in indoor environments with a higher prevalence of c-decaBDE containing products e.g. office environments, airplane interiors etc. (Bjrklund et al. 2012, Allen et al. 2013). BDE-209 in indoor environments is also a significant source to BDE-209 pollution in urban outdoor air (Brklund et al. 2012, Cousins et al., 2014). Based on measurements in sewage sludge the estimated releases of BDE-209 from the technosphere via this route, in Europe, are 16(8.6 tonnes annually and 41(22mg annually per person or 0.2% of annual c-decaBDE usage in Europe (Ricklund et al. 2008). Hence, use of c-decaDBE in the production of textiles and electronics results in environmental releases of BDE-209 and other PBDEs, either during production or directly from articles or during their disposal stage (RPA 2014, VECAP 2010) and thus contribute to environmental releases and transboundary air pollution.
Different factors may influence the release of c-decaBDE from products. First increased temperatures and thermal stress which potentially increases the volatility and release of c-decaBDE to the surrounding environment, e.g. the internal temperature of automobiles, TV's, computers and other office equipment can exceed 50(C (Earnshaw et al. 2013). Second, processes of physical abrasion, disintegration and weathering of c-decaBDE containing products during their lifetime leads to the creation of particle bound emissions. Third, sunlight exposure has been shown to contribute to photodebromination and formation of lower brominated PBDEs as well as PBDFs in flame retarded textiles and plastics (Chen et al. 2013, Kajiwara et al. 2008, Kajiwara et al. 2013 a,b).
As a general purpose flame retardant c-decaBDE is used and released at many industrial and professional sites (VECAP 2012, Li et al. 2013a, Gao et al. 2011, Marvin et al. 2013). The exact number of sites of use of c-decaBDE is unknown. The EU Annex XV report and references cited therein indicate more than 100 sites of second-line use in the EU (compounders/formulators, master batchers, injection moulders and finishers) (ECHA 2012a). Globally the number of point sources is likely much higher than the figure indicated by ECHA (Annex E submission IPEN), and also includes production sites.
In a UK assessment landfill and incineration mainly of polymers were highlighted as the main sources of c-decaBDE release, followed by releases of waste water mainly from washing of textiles and releases to air, mainly during service life of polymers (UK Environment Agency 2009). In addition, application of sludge (biosoil) as agricultural fertilizer is indicated as an important pathway for BDE-209 emissions to soil (Buser et al. 2007a, Sellstm et al. 2005, de Wit et al. 2005).
Although emissions of c-decaBDE during the service life and upon disposal of products are generally thought to be larger than those from industrial processes (UK Environment Agency 2009, ECHA 2012c, ACAP 2007, Sakai et al. 2006), ACAP data indicate substantial releases of decaBDE also from industrial point sources (ACAP 2007).
Figures generated as part of the voluntary emissions control action programme (VECAP) suggests that the potential emissions of c-decaBDE from point sources (second-line users) in Europe and North-America is low and has been significantly reduced since the introduction of the program (VECAP 2010a,b, VECAP 2012). European VECAP data indicates a reduction in estimated releases from < 4 to < 0.3 tonnes per year over a five year period from 2008 to 2012 (VECAP 2012). However, according to ECHA 2013a, environmental monitoring data do not provide evidence for a decline in emissions - despite voluntary initiatives by the European industry to reduce environmental releases of decaBDE from local point sources that have been in place since 2004. One potential explanation for this discrepancy is that VECAP data does not provide information about emissions during the service life of articles or during disposal of articles at the end of their service life.
Emission estimates for the period 1970 to 2020 calculated by Earnshaw et al. 2013 indicate that c-decaBDE atmospheric emissions in Europe increased steadily from the 1970s and reached a peak in 2004 at 10 tonnes/ year. Emissions to soil and the hydrosphere are lower but follow a similar trend of increase from the 1970s, peaking in the late 2000s and declining thereafter. Emissions to soil peaked at 4 tonnes/year in 2000 whilst those to the hydrosphere peak in the 2010s at 3.5 tonnes/ year. In Switzerland maximum emissions are predicted to have occurred in the 1990's (Buser et al. 2007a,b, Morf et al. 2007).
According to Earnshaw, comparable European estimates for production, consumption or emissions of c-decaBDE or BDE-209 are presently not available. However, emission estimates have been made for individual countries (ECB 2002, Morf et al., 2003, 2008, Palm et al., 2002, Sakai et al., 2006 as cited in Earnshaw et al. 2013). Estimated emissions to air reported in these studies range, across the different countries, between approximately 100 to > 100,000 kg/ year, while estimated emissions to water range from approximately 1,000 to >10,000 kg/ year and estimated emissions to soil from >1,000 to <100,000 kg/ year. The lowest emissions were indicated for Switzerland with a maximum of 20 kg/ year to atmosphere, 6 kg/ year to the hydrosphere and 40 kg/ year to soil (Buser et al. 2007b, Morf et al. 2007). The large difference in release estimates suggests a certain degree of uncertainty in this data and indicates that release estimates should be viewed in light of environmental monitoring data.
Further information on potential emission sources and environmental levels resulting from releases of c-decaBDE to the environment is provided in Section 2.3.1. In general, as indicated by measured environmental levels, releases to the environment are higher in industrialized and urban areas than in rural and agricultural areas where there are fewer sources (see Section 2.3.1). Environmental levels are typically lowest in remote regions, such as the Arctic.
2.2 Environmental fate
The environmental fate properties of BDE-209 have been assessed in various reports published by the EU, Canada and the United Kingdom (ECB 2002, ECB 2004, ECHA 2012a, Environment Canada 2006, Environment Canada 2010, UK EA 2009). Fugacity modeling predicts that most of the BDE-209 (> 96%) in the environment partitions to sediment and soil (Environment Canada 2010, ECHA 2013a). Less than 3.4 % of BDE-209 is expected to be associated with bulk air or bulk water phases. More precisely, 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 (< 0 . 1 g / L ) , i t s m o b i l i t y i n s o i l s i s a l s o l i k e l y t o b e l o w ( E C H A 2 0 1 3 a ) . A s a c o n s e q u e n c e , s e d i m e n t s a n d s o i l s a r e t h e p r i m a r y c o m p a r t m e n t s i n w h i c h t h e s u b s t a n c e w i l l r e s i d e a t s t e a d y s t a t e f o l l o w i n g r e l e a s e , a n d t h e s e a r e t h e m o s t i m p o r t a n t i n t e r m s o f t h e relevance of transformation to lower brominated PBDEs with PBT/vPvB and POP properties. BDE-209 is also found in biota 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 chapter 2.2.2 the debromination of BDE-209 that occurs in the environment and biota has important implications for its environmental fate.
2.2.1 Persistence
The environmental lifetime of chemicals in the environment is governed by different abiotic and biotic degradation mechanisms such as photolysis, reaction with oxidants (such as hydroxyl radicals, nitrate radicals and ozone), hydrolysis, reaction with reductants and biotransformation/biodegradation by organisms (reviewed by ECHA 2012a). Photodegradation and biodegradation are likely to be the main mechanisms for transformation of BDE-209 in the environment (Environment Canada 2006, Environment Canada 2010). Due to the lack of any functional groups that are readily susceptible to hydrolysis and a very low water solubility, < 0.1 g/L at 25 C (Stenzel and Markley, 1997), hydrolysis of BDE-209 is unlikely to be a relevant degradation process in the environment (ECHA 2012 a). In the atmospheric compartment, BDE-209 will almost exclusively be adsorbed to particles. As air-particles protect the BDE-209 molecule degradation by photolysis in air is not substantial (see Section 2.2.3).
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 exposure to light. (ECHA 2012 a, Environment Canada 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 et al., 2006).
As reported in ECHA 2012a several studies show that BDE-209 can photodegrade relatively quickly under controlled laboratory conditions, but this is not necessarily the case under natural conditions, where BDE-209 is expected to adsorb strongly to organic matter in suspended particles, sewage sludge, sediment and soil. For example the half-life for BDE-209 in sand was found to be only 35-37 hours when sand was exposed to natural sunlight (Sderstrm et al. 1998 and Tysklind et al. 2001 as cited in ECHA 2012a). The corresponding half-lives in sediment and soil were estimated to be 100 and 200 hours, respectively. However, under other 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). The longest half-life in sediment is reported by Tokarz et al. (2008) who by conducting a laboratory microcosm experiment over a period of 3.5 years at 22C under dark conditions found the half-life of BDE-209 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 et al. (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 was detected as early as 1975 and the levels increased steadily to 7.4 ng g-1 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.
Further evidence for the persistency of BDE-209 is provided by studies on sludge and soil. Liu et al. (2011a) observed no degradation of BDE-209 after 180 days in soil samples spiked with BDE-209. 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 et al. 2010 and Nyholm et al. 2011, as cited in ECHA 2012 a). 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 et al. 2005). These results are further strengthened by field studies. Eljarrat et al. (2008) examined the fate of PBDEs in sewage sludge from five municipal WWTPs after agricultural application of sludge to the top soil of the field at six sludge application sites and one reference site. The results showed that BDE-209 was the predominant PBDE congener in sewage sludge (80.6 to 1083 ng g-1 dw) and that the concentrations of BDE-209 in soils fertilized with the examined sewage sludge ranged between 14.6 to1082 ng g-1 dw at the agricultural sites. According to the authors, concentrations were found to be high (71.7 ng g-1 dw) even at one site that had not received sludge applications for four years, illustrating the persistency of BDE-209 in soils. Similarly, Sellstrm et al. (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-1 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.
2.2.2 Degradation and debromination
In spite of BDE-209s persistence and long environmental half-lives in sediment, soil and air, degradation and biotransformation of BDE-209 in environmental matrices and biota is known to occur and has been extensively reviewed (ECHA 2012, UNEP/POPS/POPRC.9/INF/19, ECHA 2013a, c, UK EA 2009, Environment Canada 2010, NCP 2013). The potential biotransformation of BDE-209 to lower brominated PBDEs with PBT/vPvB and POP properties was considered to be a high concern in the assessments of BDE-209 in a number of recent reports and published studies (ACHS 2010, ECHA 2012a,c, EFSA 2011, Environment Canada 2010, and UNEP 2010, Ross et al. 2009, McKinney et al., 2011a).
As described above (Section 2.2.1), photodegradation is likely the main factor and cause of abiotic degradation/ debromination of BDE-209. Reaction with reductants (e.g. iron-bearing minerals and sulphide ions, etc., some of which may be water-soluble) present in anaerobic conditions in both sediments and soils is a possible additional abiotic transformation route (ECHA 2012). A number of abiotic degradation studies are available (reviewed in ECHA 2012) and have shown the formation of nona- to triBDEs.
The most unequivocal evidence that photodebromination might occur in soil, sediment, air and other environmental matrices comes from controlled laboratory studies with natural sunlight. In brief, several studies have shown that BDE-209 adsorbed as a thin film on solid surfaces in water can photodegrade relatively quickly to other PBDEs (ECHA 2012c). In contrast, the evidence for phototransformation in soil and air is more limited.
While the identity of the degradation products is inconclusive in some studies (rn 1997, Palm et al. 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 (Sellstrm et al. ( ADDIN EN.CITE Sellstrm19981372(1998)1372137217Sellstrm, USderstrm, GDe Wit, CTysklind, MPhotolytic debromination of decabromodiphenyl ether (DeBDE)Organohalogen CompoundsOrganohalogen Compounds447-450351998HYPERLINK \l "_ENREF_77" \o "Sellstrm, 1998 #1372"1998), Tysklind et al. ( ADDIN EN.CITE Tysklind20011373(2001)1373137310Tysklind, MSellstrm, USderstrm, Gde Wit, CAbiotic transformation of polybrominated diphenylethers (PBDEs): photolytic debromination of decabromo diphenyl etherBrominated Flame Retardants Conference, Ontario, Canada42-452001HYPERLINK \l "_ENREF_92" \o "Tysklind, 2001 #1373"2001) Sderstrm ADDIN EN.CITE Sderstrm20031380(2003)1380138017Sderstrm, GOn the combustion and photolytic degradation products of some brominated flame retardants. Department of ChemistryEnvironmental Chemistry. Sweden: University of UmeEnvironmental Chemistry. Sweden: University of Ume2003(HYPERLINK \l "_ENREF_85" \o "Sderstrm, 2003 #1380"2003) and Sderstrm et al. ADDIN EN.CITE Sderstrm20041379(2004)1379137917Sderstrm, GunillaSellstrm, Ullade Wit, Cynthia ATysklind, MatsPhotolytic debromination of decabromodiphenyl ether (BDE 209)Environmental Science & TechnologyEnvironmental Science & Technology127-13238120040013-936X(HYPERLINK \l "_ENREF_86" \o "Sderstrm, 2004 #1379"2004) in ECHA 2013a). Formation of hexaBDEs, via abiotic debromination, was also observed by Jafvert and Hua (2001a) and Eriksson et al. (2004). Ahn et al. (2006b) 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. Photodegradation/ debromination may also take place in water (Leal et al. 2013).
The relevance of photolytic degradation of BDE-209 in the environment may however be limited. In most environmental matrices, only the surface layer is likely to be exposed to light. In aquatic environments and sediments only a very small fraction of the total BDE-209 present will be available for photodegradation due to light attenuation, shielding, etc (ECHA 2012, Kohler et al. 2008). In air, photodegradation is restricted to an extent that allows for long range transport of BDE-209 and only occurs when BDE-209 is not bound to particles (i.e. in gas phase) or when particles are suspended in air for significant periods of time (de Wit et al. 2010, ECHA 2012). In soil, depth as well as increased adsorption to ageing soil are factors that limit transformation via this pathway over longer timescales. In spite of this, monitoring data provide supporting evidence that degradation of BDE-209 occurs under environmental conditions (e.g. Orihel et al. 2011 as described in ECHA 2012, Hermanson et al. 2010, Xiao et al. 2012). Preliminary results of Canadian-government funded studies provide evidence of the formation of small amounts of nona- and octaBDEs over a period of 30 days in lake sediments (Orihel et al. 2011, see also ECHA 2012). A few studies demonstrate degradation of BDE-209 (mainly to nona- and octaBDEs) in sewage sludge (Stiborova et al 2008, Gerecke et al 2006, ECHA 2012), as well as in precipitation (Arinaitwe et al. 2014). Change in the congener ratio in sludge compared to commercial formulations has furthermore been observed (Knoth et al 2007). Although minimal debromination during the WWT process was reported by Kim et al (2013), findings support the contention that BDE-209 in sewage sludge can be debrominated to less brominated congeners (Hale et al. 2012). An overview of degradation products in abiotic matrices is given in UNEP/POPS/POPRC.10/INFxx Table 3.4.
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 et al 2008, Kajiwara et al 2013a,b). Other degradation products such as polybrominated dibenzofurans (PBDF) and dibenzo-p-dioxins (PBDD) can moreover be formed from BDE-209 during processing (recycling), plastics production, photolysis and food preparation (cooking of fish) (Bendig et al 2013a, Kajiwara et al 2008, 2013a,b, Ren et al. 2011, Hamm et al 2001, Yu et al. 2008, Zennegg et al. 2009, Weber and Kuch 2003). Formation is strongly dependent on conditions like temperature and purity of the flame retardant. Furthermore, hexabromobenzene (HBB) has also been identified as a possible degradation product of BDE-209 (Thoma and Hutzinger 1987).
As shown in a number of studies, microorganisms can influence BDE-209 degradation in soil and sediments as they are capable of transforming deca-, nona- and octaBDEs to at least hepta- and hexaBDEs ( ADDIN EN.CITE ADDIN EN.CITE.DATA HYPERLINK \l "_ENREF_72" \o "Robrock, 2008 #1406"Robrock et al. 2008, HYPERLINK \l "_ENREF_50" \o "Lee, 2010 #1405"Lee and He 2010, HYPERLINK \l "_ENREF_15" \o "Deng, 2011 #1383"Deng et al. 2011, HYPERLINK \l "_ENREF_67" \o "Qiu, 2012 #1348"Qiu et al. 2012). Although these studies are not necessarily representative of environmental conditions, they indicate that such organisms are capable of performing the transformation. Qiu et al. ADDIN EN.CITE Qiu20121348(2012)1348134817Qiu, MengdeChen, XingjuanDeng, DaiyongGuo, JunSun, GuopingMai, BixianXu, MeiyingEffects of electron donors on anaerobic microbial debromination of polybrominated diphenyl ethers (PBDEs)BiodegradationBiodegradation351-36123320120923-9820(HYPERLINK \l "_ENREF_67" \o "Qiu, 2012 #1348"2012) have demonstrated that sediment-dwelling microorganisms are capable of carrying out transformation reactions to form at least hexaBDEs under laboratory conditions over a three-month period. Similar findings have been reported by other laboratories (e.g. ADDIN EN.CITE Deng20111383(Deng et al. 2011)1383138317Deng, DaiyongGuo, JunSun, GuopingChen, XingjuanQiu, MengdeXu, MeiyingAerobic debromination of deca-BDE: Isolation and characterization of an indigenous isolate from a PBDE contaminated sedimentInternational Biodeterioration & BiodegradationInternational Biodeterioration & Biodegradation465-46965320110964-8305HYPERLINK \l "_ENREF_15" \o "Deng, 2011 #1383"Deng et al. 2011).
There is also evidence that debromination of BDE-209 in soil is assisted by the presence of plants (Du et al 2013, Huang et al 2010a, Wang et al 2011a). The distribution pattern of lower brominated PBDEs in plant tissues was different from that in 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 et al 2013, Wang et al 2011a). Notably, in soils, an average loss of BDE-209 of 20% (range 6-35%) was observed over a two-month period in a greenhouse experiment involving plants (Huang et al. 2010a). For the plants associated with the highest level of transformation, at least 122 g/kg dw of tetra- to heptaBDEs was formed (i.e. 2.7% of the total measured PBDE).
As shown in a number of studies debromination also occurs within or in the presence of organisms (ECHA 2012a,c, UK EA 2009, Environment Canada 2010, UNEP/POPS/POPRC.9/INF/19). Evidence for debromination also comes from studies with higher vertebrates, including birds, fish and rodents. 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 et al. 2011). According to Environment Canada (2010) biodegradation has been shown to be a very slow process, with half-lives on the scale of a few years to several decades, but may also be more rapid e.g. in a recent study with American kestrels exposed to BDE-209 via the diet the half-life of BDE-209 was estimated to be 14 days based on plasma concentrations over the uptake and elimination periods (Letcher et al. 2014).
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 et al 1999, Stapleton et al 2004, 2006, Kuo et al 2010, Munschy et al 2011, Noyes et al 2011, 2013, Zeng et al 2012, Wan et al 2013, Feng et al 2010, 2012, Luo et al 2013, Bhavsar et al. 2008). 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 et al 2011, Wan et al 2013, Vigano et al 2011). The concentrations of BDE-209 and its degradation products varied between the different fish species which might be explained by species-specific differences in bioaccumulation capacity and metabolic pattern between the fish species (Stapleton et al 2006, Luo et al 2013). Formation of hydroxy- and methoxyBDE degradation products have also been reported (Feng et al 2010, 2012, Zeng et al 2012).
Available mammalian data indicate that debromination is the first step in biotransformation of BDE-209, followed by hydroxylation to phenols and catechols (Riu et al. 2008). These metabolites may be conjugated via Phase II reactions and excreted via bile and feces (Health Canada 2012). Other rodent studies have demonstrated metabolism of BDE-209 into lower PBDE congeners (down to BDE-183, a heptaBDE listed in the Stockholm Convention) (Wang et al. 2010a, Huwe et al. 2007). Huwe and Smith (2007) suggested that neutral debrominated metabolites were only a fraction (1%) of the total PBDE mass balance in exposed rats. This suggests that the hydroxylation and methylation pathways and resulting metabolites are highly significant in the metabolism of BDE-209 (Environment Canada 2010). Mrk and coworkers (2003) found that 90% of the BDE-209 given to rats was excreted through feces and that the majority (65%) was identified as metabolites. Excretion via bile accounted for 10% and was almost exclusively metabolites. Based on these findings the authors suggest that BDE-209 may be actively transported into the lumen through the intestinal wall, or that BDE-209 is metabolized by intestinal microflora or by first pass metabolism by cytochrome P450 enzymes in the intestinal wall. In rodents, the half-life has been reported to be 2.5 days (Sandholm et al. 2003).
A number of studies have also revealed debromination of BDE-209 in birds or bird eggs (reviewed by Chen and Hale, 2010, Park et al 2009, Van den Steen et al 2007, Letcher et al 2014, Holden et al 2009, Munoz-Arnanz et al 2011, Mo et al.2012, Charbot-Giguere et al. 2013, Crosse et al. 2012). Interestingly, as observed in fish, BDE-202 along with other unidentified congeners not present in c-decaBDE was detected in the bird eggs in many studies and are seen as evidence that debromination occurs (Park et al 2009, Holden et al 2009, Mo et al.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 et al 2009, Mo et al 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 of BDE-209 to marine and aquatic biota. Indication of BDE-209 biotransformation in the terrestrial environment was also shown by the high amounts of BDE-208 in worms following exposure to BDE-209 spiked food and sediments (Klosterhaus and Baker, 2010). An overview of degradation products in biota are reported in UNEP/POPS/POPRC.10/INFxx Table 3.2 and 3.3.
In conclusion, there is mounting evidence that BDE-209 is debrominated to lower brominated PBDE congeners in the environment and in biota (ECHA 2012a, c, UK EA 2009, Environment Canada 2010), and that also other degradation products are formed. Observed debromination products range from mono- to nonaBDEs, and also include listed POPs as well as other recognized PBT/ vPvB substances such as PBDF, PXDD/ PXDF and HBB (UK EA 2009, ECHA 2012a,c, Environment Canada 2010, see Tables 3.2 and 3.3 UNEP/POPS/POPRC.10/INFxx). Due to the wide distribution of BDE-209 in the environment combined with its high persistence in sediment and soil, organisms are continuously exposed to a complex mixture of BDE-209 and lower brominated PBDEs as well as other BDE-209 degradation products through their lifetime (ECHA 2012), increasing the likelihood for adverse effects (Ross et al., 2009, McKinney et al., 2011, Kortenkamp et al. 2014). The substantial debromination of BDE-209 observed in biota has implications for the understanding of BDE-209 contamination in the environment as the extent of wildlife, and likely human, exposure may not be fully realized by measurement of tissue levels of BDE-209, which may be very low and highly affected by metabolism/ degradation, thus leading to underestimation of the ecosystem burden of total-BDE-209 (McKinney et al. 2011).
2.2.3 Bioavailability and tissue distribution
Due to its low water solubility and 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. The bioavailability of BDE-209 is claimed to be low due to its high molecular weight that affect its diffusion through biological membranes (Frouin et al. 2013, Mizukawa et al. 2009) but in spite of that, detectable and sometimes high levels have been detected in a variety of species and tissues and this is confirmed by the large number of laboratory studies as well as by monitoring data (see Section 2.3).
Bioavailability via direct exposure in aqueous media is reportedly low. When dosed via water as biosolids- or sediment-associated PBDEs, low bioavailability of BDE-209 was shown in several aquatic species (Ciparis and Hale, 2005, Klosterhaus and Baker, 2010). However, some evidence for the bioavailability of BDE-209 through direct water exposure in medaka is shown, despite the extremely low water solubility, by the detection of BDE-209 in muscle at all treatment groups after water exposure to BDE-209 (1 1000 ng/L BDE-209) (Luo et al 2013). In controlled laboratory experiments the oral uptake of BDE-209 through diet was evaluated in fish (Kierkegaard et al.1999, Stapleton et al 2004, 2006). Increasing concentrations of BDE-209 was observed in muscle and liver in rainbow trout fed BDE-209-spiked food. The uptake of BDE-209 in trout muscle was estimated to 0.02- 0.13 % (including lower brominated transformation products) (Kierkegaard et al., 1999). Higher absorption of BDE-209 was calculated in a feeding study using juvenile carp and juvenile rainbow trout fed BDE-209-spiked food. Limited bioavailability of BDE-209 (0.44 %) measured as lower debrominated congeners was seen for carp (Stapleton et al., 2004) while the uptake of BDE-209 in rainbow trout was estimated to 3.2 % (including lower brominated transformation products) (Stapleton et al 2006). Bioavailability of BDE-209 to zebrafish was revealed by larvae accumulating ten-fold more BDE-209 than controls after post-fertilization exposure to sediments spiked with BDE-209 (Garcia-Reyero et al 2014). Furthermore, BDE-209 was bioavailable to zebrafish (measured in whole body homogenate) and transferred to the eggs in a feeding study using fish fed a diet containing BDE-209 (high and low doses). The egg/fish ratio was > 1 for both exposure doses (Nyholm et al 2008). Studies of marine mammals also reveal bioavailability of BDE-209 by detection in blubber, liver and plasma (Tomy et al.2008; 2009, van Leeuwen et al. 2009, Thomas et al. 2005, Ikonomou et al. 2002, Villanger 2013, Shaw et al. 2009, 2012).
The bioavailability of BDE-209 is presumably higher in terrestrial than in aquatic environments (Kelly et al. 2007, see also Sections 2.2.4 and 2.3.4). Evidence for bioavailability of BDE-209 in terrestrial species is shown by the uptake of BDE-209 into birds (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 the few controlled lab studies on birds, BDE-209 was detected in plasma, liver and fat samples from American kestrels dietary exposed to BDE-209 in a controlled in vivo experiment (Letcher et al 2014). BDE-209 has also been detected in the brain of glaucous gulls in the Arctic although at low frequency and low concentrations (Sagerup et al 2009). Transfer of BDE-209 from birds to bird eggs has been reported (Vorkamp et al 2005, Lindberg et al 2004, Johansson et al 2009). Recently, a hind to foetus examination in reindeer from Finland reported a foetus/hind ratio of 2 for BDE-209 revealing both bioavailability, transfer to offspring and bioaccumulation of BDE-209 in this species (Holma-Suutari et al 2014). Earthworms living in BDE-contaminated soil were furthermore shown to accumulate BDE-209 (Sellstrm et al 2005).
Although the toxicokinetics of BDE-209 has been investigated in several studies, knowledge about uptake, distribution and elimination in various organisms is limited. However, 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 et al. 2012, Wan et al. 2013, EFSA 2011, ECB 2002, ECB 2004). Although limited evidence is available, BDE-209, like perfluorinated compounds, has also been observed to bind to proteins (Morck et al. 2003, Hakk et al. 2002). While further scientific verification is warranted, these observations indicate that decaBDE may behave more like perfluorinated compounds such as PFOS, which are proteinophilic and has a tissue distribution pattern similar to that seen for BDE-209 (Kelly et al. 2009, Martin et al. 2003, Hoff et al. 2003). Furthermore, it is possible that uptake and accumulation in tissues is not driven by lipid content but by other mechanisms such as transport via proteins (Charman 2000, Hakk et al 2002).
In fish, the toxicokinetic processes of BDE-209 were investigated in Chinese sturgeon (Wan et al 2013). Interestingly and different from the less brominated BDEs, lipids were not observed to play an important role in the distribution of BDE-209. Relatively high concentrations were detected in liver, gills, muscle and intestine compared to the adipose. The estimated partition coefficients between tissues and blood were higher than those of less brominated BDEs in various animals, suggesting that the low partition ratios from blood to tissues would lead to high bioaccumulation of BDE-209, especially in absorbing organs.
In a bioaccumulation study on PBDEs, Shaw et al. 2012 observed a similar pattern. Concentrations in liver of 56 harbor seals (6 adult males, 50 pups) were measured and compared to levels in blubber. H e p a t i c P B D E ( t r i - t o O c t a - B D E ) c o n c e n t r a t i o n s ( r a n g e 3 5 1 9 , 5 4 7 n g / g l i p i d w e i g h t , l w ) w e r e s i m i l a r t o b l u b b e r c o n c e n t r a t i o n s . I n c o n t r a s t , B D E - 2 0 9 c o n c e n t r a t i o n s i n l i v e r w e r e u p t o f i v e t i m e s h i g h e r t h a n t h o s e i n b l u b b e r , w h i c h i s c o n s i s t e n t w i t h o b s e r v ations that BDE-209 migrates to perfused tissues such as the liver in biota. Tissue distribution of PBDEs also varied significantly by age and, surprisingly, by gender among the pups.
Oral absorption in rats is reported to range from 7-26%, inhalation absorption is estimated to be negligible, and in an in vitro experiment dermal absorption was less than 20% (Hughes et al. 2001). Based on organ fresh weights, highest concentrations were found in adrenals, kidney, heart and liver in animals (EFSA 2011). In a distribution study by Seyer, highest concentration was found in the adrenal glands and the ovary (Seyer et al., 2010). Levels of BDE-209 in rat dams brains were 0.01% (Riu et al., 2008), while in mice exposed postnatal day 3 or 10 to a single dose had higher levels (0.05%) than the rat dams and compared with mice exposed at a later timepoint, postnatal day 19 (0.006%) (Viberg et al., 2003). Exposure of pregnant rats has shown that BDE-209 is distributed to the fetuses (0.5% of the total dose), but the plasma concentrations were lower compared to those of the dams (Biesemeier et al., 2010, Riu et al., 2008).
Human data demonstrate that BDE-209 is absorbed and distributed to blood, cord blood, placenta, fetuses and breast milk (Frederiksen et al. 2009a, Zhao et al. 2013; UNEP/POPS/POPRC.10/INFxx, Table 4.1). Thuresson and co-workers calculated the apparent half-life for BDE-209 in humans to be 15 days, the three nonaBDEs and four octaBDE congeners were found to have half-lives of 18-39 and 37-91 days, respectively (Thuresson et al. 2006 in EFSA 2011).
2.2.4 Bioaccumulation
This section focus on the bioaccumulation potential of BDE-209. However, as reviewed in section 2.2.2 and in other assessments (Environment Canada 2010, ECHA 2012a,c, UK EA 2009), biotransformation of BDE-209 to lower brominated PBDEs is of high concern and has to be kept in mind when discussing the bioaccumulation potential of BDE-209 as several of the lower brominated PBDEs have been identified as POPs and/or PBT/vPvB substances and are known to bioaccumulate (POPRC 2006, POPRC 2007, ECHA 2012a, Environment Canada 2010). Moreover, based on field studies is it difficult to distinguish whether the presence of lower brominated PBDEs is due to debromination of BDE-209 or as a result of direct exposure from c-pentaBDE or c-octaBDE.
In the past, low bioaccumulation of BDE-209 in biota was mostly attributed to the large molecular size, extreme hydrophobicity and low bioavailability of BDE-209 (Hale et al 2003). However, low bioavailability and bioaccumulation can be explained by factors other than molecular size, such as low uptake and/or elimination through excretion and debromination of BDE-209 (Hale et al 2003, Arnot et al., 2010), and by the fact that BDE-209 has been analytically challenging to measure (Ross et al. 2009, de Boer and Wells 2006, Covaci et al. 2003, Kortenkamp et al. 2014). Environmental monitoring studies show that BDE-209 is taken up by 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 ADDIN EN.CITE ADDIN EN.CITE.DATA (HYPERLINK \l "_ENREF_15" \o "CMABFRIP, 1997 #126"CMABFRIP 1997, HYPERLINK \l "_ENREF_19" \o "Dinn, 2012 #118"Dinn et al. 2012, HYPERLINK \l "_ENREF_21" \o "Env.Canada, 2010 #111"Environment Canada 2010, HYPERLINK \l "_ENREF_36" \o "Kelly, 2007 #141"Kelly et al. 2007, HYPERLINK \l "_ENREF_70" \o "Tian, 2012 #120"Tian et al. 2012, HYPERLINK \l "_ENREF_72" \o "U.S.EPA, 2010 #114"U.S.EPA 2010, HYPERLINK \l "_ENREF_82" \o "Watanabe, 1990 #127"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 ADDIN EN.CITE Arnot2003129(Arnot and Gobas 2003)12912917Arnot, J. A.Gobas, FapcSimon Fraser Univ, Sch Resource & Environm Management, Burnaby, BC V5A 1S6, Canada.
Gobas, FAPC (reprint author), Simon Fraser Univ, Sch Resource & Environm Management, 8888 Univ Dr, Burnaby, BC V5A 1S6, Canada.A generic QSAR for assessing the bioaccumulation potential of organic chemicals in aquatic food websQsar Comb SciQSAR Comb. Sci.Qsar & Combinatorial ScienceQSAR Comb. Sci.337-345223bioaccumulationQSARbioaccumulation factoroctanol-water partitioncoefficientPARTITION-COEFFICIENTBIOCONCENTRATION FACTORSLAKE-ONTARIOFISHMODELPOLLUTANTSELIMINATION2003May1611-020XWOS:000184648000003Article; Proceedings Paper<Go to ISI>://WOS:00018464800000310.1002/qsar.200390023English(HYPERLINK \l "_ENREF_2" \o "Arnot, 2003 #129"Arnot and Gobas 2003). However, in spite of BDE-209 having a high log Kow, there is evidence from food web studies that BDE-209 bioaccumulates in both aquatic and terrestrial species (Yu et al. 2011, Wu et al. 2009a, Tomy et al. 2009, Environment Canada 2010).
Under Annex D of the Stockholm Convention, as well as under Canadian and U.S. regulations, chemicals with bioaccumulation factors (BAFs ) o r b i o c o n c e n t r a t i o n f a c t o r s ( B C F s ) e" 5 , 0 0 0 ( l o g B A F s o r B C F s e" 3 . 7 ) a r e c o n s i d e r e d t o h a v e a h i g h p o t e n t i a l f o r b i o a c c u m u l a t i o n . B i o c o n c e n t r a t i o n a s m e a s u r e d b y B C F s r e p r e s e n t s p r o c e s s e s o f c h e m i c a l a b s o r p t i o n b y a n a q u a t i c o r g a n i s m f r o m t h e a m b i e n t a q u e ous environment through its respiratory and dermal surfaces with no dietary considerations ADDIN EN.CITE ADDIN EN.CITE.DATA (HYPERLINK \l "_ENREF_3" \o "Arnot, 2006 #132"Arnot and Gobas 2006). BCF is not a good descriptor of the biomagnification capacity of chemical substances. According to the revised OECD guidelines for strongly hydrophobic substances (log KOW > 5 and water solubility below ~ 0.01-0.1 mg/L), testing via aqueous exposure may become increasingly difficult the more hydrophobic the substance is. Accordingly, for highly hydrophobic substances a dietary test is recommended (OECD 305, 2012). In terrestrial food chains chemicals with log Kow<5 and BCFs<5000 have been shown to biomagnify. Furthermore, for air-breathing (terrestrial) organisms KOW and BCF are not good predictors of biomagnification for chemicals with log KOA (6 and KOW>2 (Kelly et al. 2007, 2009).The BCF for BDE-209 in fish has been estimated to be <5000 with non-appreciable aqueous uptake predicted due to its large size a n d l o w w a t e r s o l u b i l i t y ( <