1



# HEXABROMOCYCLODODECANE

DRAFT RISK PROFILE

April 20102 

### TABLE OF CONTENTS Executive summary

................................ ................................ ................................ ................................ ......... 3

1. Introduction

................................ ................................ ................................ ................................ ............. 4

1.1 Chemical identity of the proposed substa nce

................................ ................................ ................... 4

1.2 Conclusion of the Review Committee regarding Annex D information

................................ .......... 4

1.3 Data sources

................................ ................................ ................................ ................................ ..... 4

1.4 Status of the chemical under international conventions

................................ ................................ ... 5

2. Summary information relevant to the risk profile

................................ ................................ ................... 5

2.1 Sources

................................ ................................ ................................ ................................ ............. 5

2.1.1 Production, trade, stockpiles

................................ ................................ ................................ .... 5

2.1.2 Uses

................................ ................................ ................................ ................................ ......... 6

2.1.3 Releases to the environment

................................ ................................ ................................ .... 6

2.1.4 Releases from industrial point sources

................................ ................................ .................... 7

2.1.5 Releases from products during service life

................................ ................................ .............. 7

2.1.6 Releases from waste

................................ ................................ ................................ ................ 8

2.2 Environmental fate

................................ ................................ ................................ ........................... 8

2.2.1 Persistence

................................ ................................ ................................ ............................... 8

2.2.2 Bioaccumulation

................................ ................................ ................................ ...................... 9

2.2.3 Potential for long

- range environmental transport

................................ ................................ .. 11

2.3 Exposure

................................ ................................ ................................ ................................ ........ 13

2.3.1 Environmental levels and trends

................................ ................................ ............................ 13

2.3.2 Human exposure

................................ ................................ ................................ .................... 16

2.4 Hazard assessment for endpoints of concern

................................ ................................ ................. 18

2.4.1 Ecotoxicity to aquatic organisms

................................ ................................ ........................... 18

2.4.2 Toxicity in soil organisms and plants

................................ ................................ .................... 20

2.4.3 Toxicity in birds

................................ ................................ ................................ ..................... 20

2.4.4 Toxicity in terrestrial mamm als

................................ ................................ ............................. 20

2.4.5 Human toxicity

................................ ................................ ................................ ...................... 22 3. Synthesis of information

................................ ................................ ................................ ............................ 23 4. Concluding statement

................................ ................................ ................................ ................................ 24

References

................................ ................................ ................................ ................................ ..................... 253



## Executive summary

1. The commercially available brominated flame retardant, hexabromocyclododecane (HBCD), is lipophilic, has a high affinity to particles and low water solubility. Depending on the manufacturer and the production method used, technical gra de HBCD consists of 70

- 95 % γ

- HBCD and 3

- 30 % of α

- and β

- HBCD

. HBCD has attracted attention as a contaminant of concern in several regions, by international environmental forums and academia. In the EU, HBCD has been identified as a Substance of Very High Concern (SVHC) meeting the criteria of a PBT (persistent, bioaccumulative and toxic) substance pursuant to Article 57(d) in the REACH regulation. In December 2009 HBCD was considered by the Executive Body of the UNECE Convention on Long

- range Transboundar y Air Pollution (LRTAP) to meet the criteria for POPs, set out in EB decision 1998/2.

2. According to industry information from 2006 HBCD is produced in the United States of America, Japan, Israel and the Netherlands and the main share of the market volume i s used in Europe. There is information available about several suppliers situated in China, but information about amounts of HBCD imported or produced in China is not available.The demand for HBCD is increasing as are the levels in the environment. The inc rease seems to correlate with

a decrease in the use of PBDEs, in some regions. HBCD is used as an additive flame retardant in polystyrene and textile products. Its main use is in production of EPS, XPS, HIPS and textile coating. There are releases of varyi ng magnitude to the environment at all the different life stages of the HBCDs life cycle. The total releases are increasing in all regions investigated. The largest release is estimated to be to water from production of insulation boards, to water and air from textile coating and diffuse release during the life cycle of insulation boards and textiles. Because of this HBCD is found to be widespread in the global environment, with high levels in top predators in the Arctic. Several trend studies show an incre ase of HBCD in the environment and in humans from 1970/1980s until recent years

. HBCD has been found to bioaccumulate in higher trophic level biota and its increased presence in the environment is likely attributed to its increased global demand. The genera l trend is higher environmental HBCD levels near point sources and urban areas. Hot spot areas have been identified in Europe and around the East China Sea, near production sites of HBCD, manufacturing sites of products containing HBCD and waste disposal s ites including those whose processes include either recycling, landfilling or incineration .The simulation test half

- lives, together with field data on HBCD in sediments showing persistency over time, persistency in biota and levels and trends in the Arctic , indicate that HBCD is sufficiently persistent to be of global concern. HBCD has a strong potential to bioaccumulate and biomagnify. HBCD is persistent in air and is subject to long

- range transport. HBCD is found to be widespread in the Arctic environment

, where concentrations in the atmosphere are elevated.

3. HBCD is very toxic to aquatic organisms. S tudies have detected reproductive, developmental and behavioral effects in mammals, with effects being trans

- generational. Besides these effects, controlled la boratory experiments have highlighted endocrine effects of HBCD exposure such as reduced eggshell thickness, growth and survival in avian species. Recent advances also include s a better understanding of HBCD potential to interfere with the hypothalamic

pit uitary

- thyroid (HPT) axis, its potential ability to disrupt normal development, to affect the central nervous system and more thorough examinations of HBCD’s reproductive and developmental effect s.
4. HBCD is found in human tissues, including blood, plasma a nd adipose tissue

. T he main sources presently known are contaminated food and dust. For breast feeding children, mothers’ milk is the main exposure route, but HBCD is also transferred across the placenta to the foetus. Human breast milk data from the 1970s to 2000 show that the HBCD levels have increased since HBCD was commercially introduced as brominated flame retardant (BFR) in the 1980s. Though information on the human toxicity of HBCD is to a great extent lacking, and tissue concentrations found in hum ans are seemingly low

, embryos and infants could be vulnerable, particularly to the observed neuroendocrine and developmental toxicity.

5. Based on the available evidence, it is concluded that HBCD is likely, as a result of long

- range environmental transport , to lead to significant adverse environmental and/or human health effects, such that global action is wa r ranted.4



## 1 Introduction

6. On June 18 th 2008, Norway as a Party to the Stockholm Convention, submitted a proposal to list the brominated flame retardan t hexabromocyclododecane (HBCD; some authors use the abbreviation HBCDD but the shorter version is preferred here) as a possible Persistent Organic Pollutant (POP) under Annex A, of the Convention. A summary of the proposal may be found in document UNEP/PO PS/POPRC.5/4 and a copy of the proposal itself in document UNEP/POPS/POPRC.5/INF/16.



### 1.1 Chemical identity of the proposed substance

7. Commercially available hexabromocyclododecane (HBCD) is a white solid substance. Producers and importers have provided inf

ormation on this substance under two different names; Hexabromocyclododecane (EC Number 247

- 148

- 4, CAS number 25637

- 99

- 4) and 1,2,5,6,9,10

- hexabromocyclododecane (EC Number 221

- 695

- 9, CAS number 3194

- 55

- 6). The second of these names specifies the position s of the Br atoms to the cyclododecane ring , and is therefore more precise . The molecular formula of the compound is C 12 H 18 Br 6 and its molecular weight is 641 g/mol. Technical HBCD has a log K ow of 5.625

. and a water solubility of 0.066 mg/l. The solubilit y of the individual diastereoisomers ranges from 0.0024 ( γ

- HBCD ) to 0.048 mg/l ( α

- HBCD) at 20 °C .The water solubility of the technical mixture ranges from 46. 3 in saltwater to 65.6 μg/ml in freshwater at 20 oC based on the sum of the water solubilities of the individual dias t ereoisomers ( Desjardins et al. 2004

, MacGregor and Nixon 2004 ). The structural formula of HBCD is a cyclic ring structure with Br

- atoms attached (see Figure 1 POPRC/INF document xxx). 1,2,5,6,9,10

- HBCD has six stereogenic centers and, in theory, 16 stereoisomers could be formed Heeb et al. (2005). However , in commercial HBCD only three of th e stereoisomers are commonly found. Depending on the manufacturer and the production method used, technical grade HBCD consists of 70

- 95 % γ

- HBCD and 3

- 30 % of α

- and β

- HBCD ( European Commission, 200 8 ; TemaNord 2008). Each of these stereoisomers have its o wn specific CAS numbers i.e. alpha

HBCDD, CAS No: 134237

- 50

- 6, beta

- HBCD CAS No: 134237

- 51

- 7, γ

- HBCD CAS No: 134237

- 52

- 8. Two other stereoisomers, δ

- HBCD and ε

– HBCD have also been found by Heeb et al. (2005) in commercial HBCD in concentration of 0.5 % an d 0.3 %, respectively. Supplementary information pertaining to the chemical identity of HBCD is listed in the POPRC/INF document xxx.



### 1.2 Conclusion of the Review Committee regarding Annex D information

8. The POP Review Committee has evaluated Annex D infor mation for HBCD at its fifth meeting in October 2009 (UNEP/POPS/POPRC.5/10) and has concluded that the screening criteria have been fulfilled (Decision POPRC

- 5/6).



### 1.3 Data sources

9. This risk profile is elaborated using Annex E information submitted by cou ntries and observers , national reports from web sites for environment protection agencies in different countries, the bromine industry, t he Co

- operative Programme for Monitoring and Evaluation of the Long

- Range Transmission of Air Pollutants in Europe (EME P) and
t he Arctic Monitoring and Assessment Programme (AMAP). Recent relevant information from the open scientific literature is also included.

10. Several international environmental assessments of HBCD have been conducted . Three of the se have assessed
exper imental data and field data in the literature against the POP criteria in the Stockholm Convention. The se were performed by the Nordic Council (Tema Nord 2008), the Task Force on POPs under the Convention on Long

- range Transboundary Air Pollution (LRTAP) ( ECE/EB.AIR/99 ) and The European Brominated Flame Retardant Industry Panel (EBFRIP 2009 ). EBFRIP also performed a body/tissue based assessment and a total d aily intake based assessment, where estimated effect levels and no

- effect levels calculated for body/ tissue

- residue and TDI (total daily intake) are compared with estimates of exposure in the environment (EBFRIP 2009). EMEP under LRTAP has made a model assessment of potential for long

- range transboundary atmospheric transport and persistence of HBCD. The European Commission environmental risk assessment ( European Commission 2008 ), is the most extensive of the existing assessments, examining the data on environmental fate, effects and exposure levels in depth.

11. Monitoring and assessment of pollution in the Arctic is performed under the auspices of the Arctic Council which coordinates the Arctic Monitoring and Assessment Programme (AMAP). This work is import ant in identifying pollution risk s, their impact on Arctic ecosystems and in assessing the effectivenes s of international agreements on pollution control. Scientific findings obtained under AMAP have shown HBCD to be one pollutant of the Arctic

12. In the EU, HBCD has been identified as a Substance of Very High Concern (SVHC), meeting the criteria of a PBT (pe rsistent, bioaccumulative and toxic) substance pursuant to Article 57(d) in the REACH regulation. In May 2009, HBCD was included in the ECHA recommendation list of priority substances to be subject to Authorisation under REACH, based on its hazardous prope rties, the volumes used and the likelihood of exposure to humans or the environment. In June 2009 ECHA’s recommendation list was sent to the EU Commission for final decision on which substances that should be subject to Authorisation requirements at this s tage. A proposal on classification and labeling5

of HBCD as a possible reprotoxic substance is under discussion within EU ( Proposal for Harmonised Classification and Labelling, Based on the CLP Regulation (EC) No 1272/2008, Annex VI, Part 2 Substance Name: Hexabromocyclododecane Version 2, Sep. 2009)

. 13. In Canada, Australia and Japan different national assessments of HBCD are under preparation. In Ukraine the substance is registered on the hazard chemical list based on health effects. Norway has completed thei r national assessment and has included HBCD in its national action plan for Brominated Flame Retardants. USA have made an initial screening assessment and an interim evaluat ion of the risk HBCD is posing ( U.S. E nvironmental Protection Agency, US EPA).

14. An OECD SIDS Initial Assessment Profile has been compiled (OECD, 2007). The OECD SIAM 24 agreed that HBCD possesses properties indicating a hazard for human health with regard to repeated dose toxicity and possible developmental neurotoxicity and for the envi ronment with regard to acute aquatic toxicity to algae, chronic toxicity to Daphnia, and a high bioaccumulation potential.



### 1.4 Status of the chemical under international conventions

15. HBCD is included as part of the certain brominated flame retardants group in the List of Substances for Priority Action of The Convention for the Protection of the Marine Environment of the North

- East Atlantic (the OSPAR Convention)

. The OSPAR Commission is made up of representatives of the Governments of 17 Contracting Parties and the European Commission.

16. In December 2009 HBCD was considered by the Executive Body of the UNECE Convention on Long

- range Transboundary Air Pollution (LRTAP) to meet the criteria for POPs, set out in EB decision 1998/2. In 2010 the possible management options for HBCD will be assessed to give a basis for later negotiations.



## 2 Summary information relevant to the risk profile

2.1 Sources

2.1.1 Production, trade, stockpiles

17. According to a BSEF report from 2006 HBCD is known to be produced in the Unite d States of America, Japan, Israel and the Netherlands . There is information available about several suppliers situated in China, but information about amounts of HBCD imported or produced in China is not available. According to the global demand reported by the industry in 2001, the main share of the market volume was used in Europe (BSEF 2006). Total global demand for HBCD increased over 28% in 2002 to 21,447 metric tons, and rose again slightly in 2003 to 21,951 metric tons (BSEF 2006).

Table 1 Global d emand of HBCD in 2001 in different regions of the world (BSEF 2006 and ECHA 2008a ) Global demand US Europe Asia Rest of the world Total

2001 2 800 9 500 3 900 500 16 500

18. In the US EPA assessment the sum of manufactured and imported HBCD is reported to l ie between 4 540 tons to 22 900 tons in 2005 (US EPA 2008). The authorities in Japan have reported the sum of manufactured and imported HBCD to be 2 744 tons in 2008. The consumption in Japan reached 700 tons/year in the beginning of the nineties (Managaki et al. 2009), and has increased approximately 4 times since then. The demand in 2006 in EU represents an increase of 22 % from 2001. The total volume of HBCD used in the EU was estimated to be about 11 580 tons in 2006. EU was estimated to be about 11 580 tons in 2006. The demand of HBCD within EU is bigger than the production and a net import to EU was expected to have been around 6 000 tons in 2006, implying an increase in net imports of almost 50% over this period. ( ECHA 2008a ). The import of HBCD to EU is mainly from USA (OECD SIDs 2007). HBCD powder or pellets, masterbatch of HBCD, EPS beads, impregnated granules of XPS and high impact polystyrene (HIPS) pellets are often exported and imported downstream in the production chain for the manufacturing of end

products for further professional use or sales to consumers. In Canada 100,000 to 1,000,000 tons were imported into the country in the year of 2000, both as a pure compound and in products (Environment Canada 200 3 ). The Australian authorities report a n import of less than 100 tons of HBCD each year. Ukraine report an import of HBCD for use in cellular polystyrene products. The amount HBCD in impregnated granules used in production of XPS and EPS is reported to be 185 tons by the Romanian authorities. R omania does not produce HBCD

.6



### 2.1.2 Uses

19. HBCD is used as an additive flame retardant, providing fire protection during the service life of vehicles, buildings or articles, as well as protection while stored ( BSEF 200 6 ). HBCD was placed on the market in the 1960s. The use in the production of flame

- retarded polystyrene materials began in the 1980s ( ECHA 2008a ).

20. The main uses of HBCD globally are in expanded and extrueded polystyrene foam insulation but also in textile applications ( ECHA 2008a , US EPA rep ort, OECD SIDs report, , INE

- SEMARNAT 2004,

, LCSP 2006, BSEF 2006). According to the industry HBCD is used in polystyrene foam insulation boards and laminates for sheathing products, which are widely used in the building and construction industry. They ex ist in two forms, as expanded polystyrene (EPS) and extruded polystyrene (XPS) foams with HBCD concentrations averaging 0.7% ( ECHA 2008a ) . A second important application is in polymer dispersion o n cotton or cotton mixed with synthetic blends, in the back

- coating of textiles where it can be present in concentrations ranging from 2.2

– 4.3 % (Kajiwara et al. 2009).

. A further application of H BCD is in high impact polystyrene (HIPS) which is used in electrical and electronic equipment and appliances at levels ranging from 1

– 7% (IOM2008)

. HBCD may be added to latex binders, adhesives and paints to make them flame retardan t (Albemarle Corporation 2000a, Great Lakes Chemical Corporation 2005a

– for reference see Canada’s submission). The use of HBCD in EPS in pa ckaging material is believed to be minor and HBCD is not used in food packaging according to the technical report developed in EU ( ECHA 2008a ). The US EPA 2008 has reported use in crystal and high

- impact polysterene, SAN (s t yrene

- acrylonitrile) resins, adh esives and coatings. Costa Rica has reported use of HBCD in the construction sector. In Mexico HBCD has been used in EPS foams and in back

- coating of textiles since the 1980s ( Bremauntz et al. 2004 ). In EU the main use is in XPS and EPS, the use in HIPS is assumed to be less than 10 % and use in textiles is assumed to be ca 2 % ( ECHA 2008a ). In Japan 80 % of consumption of HBCD is in insulation boards (including tatami mat) and 20 % is used in textiles. One of the uses is in traditional straw matting (tatam i mat) in Japan and is taken into account in the market share for insulation boards (Managaki et al. 2009). The Japanese textile/furniture association have stated that 90 % of their industries will stop the use of HBCD between 2010 and 2013 (Managaki et al . 2009)

. In Switzerland the construction materials are by far the most important component of HBCD consumption (84%) (Morf et al., 2008)

. 21. HBCD is used in a wide range of end products ( ECHA 2008a , US EPA report, OECD SIDs report, Bremauntz et al. 2004, Mor ose et al. 2006 ). Insulation boards with EPS foam or XPS foam with HBCD are used in transport vehicles, building constructions and road and railway embankments. HBCD is used in HIPS in electric and electronic appliances, such as in audio visual equipment c abinets (video and stereo equipment), in refrigerator lining as well as in distribution boxes for electrical lines and certain wire and cable applications. HBCD is used in textile coating agents, mainly in upholstery fabrics, but also in bed mattress ticki ng, upholstery in residential and commercial furniture, upholstery seating ́s in wheicles , draperies and wall coverings, interior textiles (roller blinds) and automobile interior textiles. The current use in upholstery furniture is mainly due to the safety standards in place in some countries and regions such as United Kingdom and California that cannot be met without the addition of flame retardants, ( LCSP 2006 ).


### 2.1.3 Releases to the environment

22. There are no natural sources of HBCD. HBCD is released into t he environment during the manufacturing process, in the manufacture of products, during their use and after they have been discarded as waste. The commercial mixture of HBCD is produced and then used in the industry in manufacture of EPS, XPS, HIPS and tex tile coating. After its service, products containing HBCD will likely go to the landfill, or remain as waste in the environment or be incinerated. There are three substance flow analyses available, one from the EU ( ECHA 2008a ) and two national, from Switze rland (Morf et al. 2008) and from Japan (Managaki et al. 2009). They are based on estimations of releases from different sources and life stages of HBCD. The two national analyses are made on the basis of time series. The results differ and some of the dif ferences are caused by different local/regional use (see chapter 2.1.2). Since Switzerland does not have a production of HBCD, this life stage for HBCD is not included in the analysis. But there are also differences in ways that releases are accounted for and in the estimation factors used in the analyses. The use category ‘insulation boards’ in the substance flow analysis in Japan, for example, also takes into account the traditional tatami mat, that could have a higher potential of releases than insulatio n boards. The releases are however estimated to be the same as for insulation boards in the analysis

. 23. In Japan in 2000 the largest releases were found to be to air (571 kg/year) and water (41 kg/year) (Managaki et al., 2009). This result was in agreement w ith the substance flow analysis made in Switzerland (Morf et al. 2008). In EU the releases to water were the largest (air; 665 kg/year, waste water; 1553 kg/year, surface water; 925 kg/year) . In Switzerland one half of the releases are estimated to come fr om diffuse atmospheric emissions from installed EPS and XPS insulations boards (Morf et al. 2008). In EU the total releases from insulation boards (1627,9 kg/year ) represent more than half of the total releases (3142 kg/year) in 2006. In Japan atmospheric emissions of HBCD from textile coating in the industry accounts for more than half of the total releases from 1985 to 2011 (Managaki et al., 2009).

24. Losses to soil were considered to be minor in Japan, Switzerland and EU since waste with HBCD was disposed off in controlled landfills or incinerated. However, a survey performed by the industry (EBFRIP 2009) revealed that potential losses to land are higher than were previously understood, due to disposal practices for packaging waste.The7

industry survey (EBF RIP 2009) included a selection of HBCD producers, warehouses and first line direct users of HBCD in Europe and only included the first stages in the HBCD life cycle. Packaging waste was found to be the main contributor to potential releases to soil due to uncontrolled landfill or compost, recycling of empty paper packaging, substances going to unknown destinations and the protected storage of packaging. Annual losses to soil were estimated to be 1857 HBCD kg/year. Implementing best practices in handling red uced the total releases from 2017 kg/year in 2008 to 309 kg/year in 2009 in the survey (EBFRIP 2009).

25. Results from the substance flow analysis in Japan (Managaki et al. 2009), indicate that emissions from construction materials will continue for several d ecades and be potentially long

- term sources of HBCD leaching or volatizing to the environment, as well as representing larger releases when demolished or renovated in the future. Additionally, the increasing HBCD stock seen in the study indicates possible problems arising in the recycling of construction materials in the future, when buildings of the present period are renovated or destroyed. This is also supported by the results from the substance flow analysis in Switzerland (Morf et al. 2008). The Swiss study also high

lights the stock in waste management and landfills as long

- term sources of HBCD releases. The significance of those sources depends however on the waste management strategies chosen in the country, if the waste are incinerated, or disposed off to an uncontrolled or controlled landfill.

26. The survey by the European Brominated Flame Retardant Industry Panel (EBFRIP 2009) included a selection of HBCD producers, warehouses and first line direct users of HBCD in Europe and only included the first s tages in the HBCD life cycle. Packaging waste was found to be the main contributor to potential releases to soil due to uncontrolled landfill or compost, recycling of empty paper packaging, substances going to unknown destinations and the protected storage of packaging. Annual losses to soil were estimated to be 1857 HBCD kg/year. Implementing best practices in handling reduced the total releases from this source from 2017 kg/year in 2008 to 309 kg/year in 2009 in the survey.

27. In Switzerland the releases du ring the service life of products was found to be the largest source to releases of HBCD (Morf et al. 2008), while the releases from the industrial point sources was largest in analysis performed in Japan (Managaki et al. 2009) and in EUs technical report (industrial point sources; 2559 kg/year, releases during surface life of products; 98,9 kg/year) ( ECHA 2008a ). The total estimations for releases of HBCD from diffuse sources are probably underestimated in all analyses, since there is lacking information on releases from some products, as well as the HBCD content in imported articles.

28. According to the EU technical report ( ECHA 2008a and t able. 3 in ECHA 2008b ), total releases of HBCD from manufacture and use of insulation boards (1627,9 kg/year) and manuf acture and use of textiles (1498,2 kg/year) are of the same magnitude table . Total releases from manufacture and use of electronic devices are minor (12,6 kg/year)

. According to the substance

- flow analysis done in Switzerland (Morf et al. 2008) constructio n materials are responsible for the majority of the releases. In Japan the emissions from use in textiles represented the largest releases (Managaki et al. 2009).The total releases are increasing in the EU, in spite of the decrease in the releases from tex tile back

- coating since 2004 ( ECHA 2008a ). In Japan (Managaki et al., 2009) the releases of HBCD are increasing rapidly. In Switzerland the total releases have increased since 1980

. 29. From both industrial point sources and diffuse sources there are releases to waste water and sewage systems ( ECHA 2008a ; Morf et al. 2008; Institute Fresenius, 2000a and b; Kupper et al., 2008; Remberger et al., 2004; Sellström et al., 1999; Law et al., 2006b). The sewage sludge is either applied on agricultural land, incinerate d or land filled ( ECHA 2008a ; Morf et al., 2008). There are releases of HBCD to surface water and soil leaching from landfills (Morf et al., 2008; Morris et al.; 2004) and sewage sludge (Morf et al. 2008; Morris et al., 2004).



### 2.1.4 Releases from industria l point sources

30. There are direct emissions to air, direct discharges to waste water and to surface water from the industrial point sources. The production process of HBCD and industrial use processes are described in EUs technical report ( ECHA 2008a ). The production of HBC D is a batch

- process. Elemental bromine is added to cyclododecatriene at 20 to 70°C in the presence of a solvent in a closed system. Subsequently, a masterbatch, a concentrated mixture of HBCD encapsulated into a carrier resin such as poly styrene, is used to manufacture flame retarded end products. The manufacture of EPS, XPS and HIPS involves polymerisation and extrusion processes where HBCD is added in the process as one of the additives used. Back

- coating to textile is applied by adding a dispersion containing a polymer and HBCD among other additives as a thin coating film.



### 2.1.5 Releases from products during service life

31. HBCD is used solely as an additive in physical admixture with the host polymer, and can thus migrate within the solid matrix and volatilize from the surface of articles during their service life. There will also be particulate releases and leaching of HBCD during the service life of flame retarded end

- products. Several studies have shown the occurrence of HBCD in indoor air and house dust (Abdallah et al. 2008 a and b

, Abdallah 2009, Goosey et al. 2008, Stapleton et al. 2008, Stuart et al. 2008, Takigami et al 2009 a and b), experiments revealing emissions of HBCD from various products ( ECHA 2008a , Miyaki et al. 2009, Thom as and Stevens 2006 and Kajiwara 2009) and elevated levels in blood serum and breast milk likely related to exposure of products in use (Kakimoto et al. 2008).8

32. Estimates of releases from insulation boards during their service life have been based on the re sults of experiment measuring the loss of HBCD from a sample of foamed polystyrene, assuming a service life of 30 years ( ECHA 2008a ). Release estimates have been developed for the service life of textiles using the results of wearing and leaching tests on aged samples of treated textiles ( ECHA 2008a

- and references therein ). There are no figures on releases of HBCD from HIPS in articles.


### 2.1.6 Releases from waste

33. Insulation boards form the majority of HBCD containing waste. It is understood that most of th is material goes to landfill or incineration. The us e of HBCD in insulation boards and the HBCD built into the society in buildings and constructions is increasing

. There will be releases of HBCD in dust when buildings insulated with flame retarded insulat ion boards are demolished. R eleases from insulation boards becoming waste was estimated to be 8 655 kg HBCD per year in 2006 ( ECHA 2008a ). It is likely that those releases will be more significant in the future; particularly from about 2025 onwards, as inc reasing numbers of buildings containing HBCD will be subject to refurbishment and demolition in the future. This turn

- over will be different in different regions of the world, and range from 10

- 50 years. Electrical and electronic appliances containing HIPS treated with HBCD is sometimes recycled. Material that cannot be recycled will be disposed of to landfill or incinerated. Other HBCD containing articles will be disposed of into municipal waste, and the eventual fate of this material will also be landfill or incineration. In the EU, the overall figures of municipal waste in EU from 2006 are that 68 % goes to landfill and 32 % is incinerated ( ECHA 200 8a) In developing countries, electrical and electronic appliances containing HBCD and other toxic substances are often recycled under dangerous conditions which results in contaminated sites (Zhang et al. 2009). Open burning and dump sites are common destinations for HBCD

- containing articles and electronic wastes.

34. Disposal of used furniture containing treated t extiles with HBCD will mostly become municipal waste, and will be landfilled or incinerated. In the substance flow analysis in Switzerland (Morf et al. 2008) emissions from the recycling of vehicles, insulation panels and E&E were estimated to account for about 2% of the total releases of HBCD and the emissions from incineration was estimated to account for 0.1%. In the EU, the w aste from industrial use, including sludge from the STP, is put on landfills, disposed of by incineration or reused in the process ( ECHA 2008a ). T he situation in developing countries is not clear, though lack of waste management infrastructure often results in open burning and dump sites containing municipal and hazardous wastes.



### 2.2 Environmental fate

2.2.1 Persistence

35. To evaluate t he persistency of HBCD a compilation of data on experimentally measured half

- lives in different environmental compartments, data on half

- lives using calculations through modeling and field data have been used.

36. Early biodegradation studies using Closed Bot tle Test systems conducted in accordance with OECD Guideline 301D found no biodegradation of HBCD over a 28 day study period (Schaefer and Haberlein, 1996) . It should be noted that while the study was performed using accepted test guidelines, the concentration used was about three orders of magnitude greater than the water solubility of HBCD (mg/L vs μg/L). Results of the estimation model, BIOWIN [v4.10, EPI Suite (v4.0)], which estimates the probability for aerobic biodegradation in the presence of mixed populations of environment al microorganisms suggest that HBCD is not readily biodegradable; the expected time of primary degradation is in the order of weeks. No degradation was observed in the aerobic soil degradation simulation test of Davis et al. (2004).

37. J apanese authorit ies conducted 28

- day biodegradation Study of 1,2,5,6,9,10

- hexabromocyclododecane based on the OECD Test Guideline 301C. The test substance was a mixture which contains isomers . Two peaks were assigned to A and B in terms of retention times on HPLC. The percent age biodegradation calculated by BOD (Biochemical oxygen demand) was 1%. The percentage biodegradation of peak A and B calculated by HPLC was 5 and 6%, respectively ( Chemicals Inspection and Testing Institute, 1990).

38. Davis et al . (2005) reported on the bio degradation of technical HBCD (t

- HBCD) in freshwater sedi ments and soils (Davis et al. 2005) . Using OECD test guidelines 307 and 308, the authors demonstrated that the rate of loss of H BCD at 20 o C was appreciable faster under anoxic conditions in both media. Relative to biologically sterile controls, biotransformation of HBCD was faster in the presence of microorganisms and t 1/2 values ranged from 11 to 32 days (aerobic) and 1.1 to 1.5 d ays (anaerobic) in sediment. In soil, half

- lives under aerobic and anaerobic conditions were 63 and 6.9 days, respectively. However, in this study only the degradation of γ

- HBCD was studied since the test concentration was to low to allow detection of α

- , β

- HBCD. It was also not possible to detect transformation products.

39. Davis et al (2006) performed a study on the degradation of 14 C

- hexabromocyclododecane in activated and digester sludge, freshwater sediments and soil. The study was primarily designed for detecting transformation products. In aerobic soil no degradation of t

- HBCD was observed during 112 days of incubation at 20 °. In sediment t

- HBCD was observed to be subject to primary degradation with half

- lives of 66 and 101 days in anaerobic and aerobic sediment at 20 °C, respectively. The half lives of the individual diastereoisomers in sediment under aerobic conditions were9

estimated to 128, 92 and 72 days for the α

- , γ

- and β

- isomer, respectively . (In the EU RAR the d egradation half

- lives in aerobic sediment were calculated at 20 °C to be 113, 68 and 104 days for α

- , β

- and γ

- HBCD, respectively , based on the same data ( European Commission , 2008). T he authors note t hat the study was conducted at HBCD concentrations much greater (mg/kg) than Davis et al (2005) (μg/kg), so degradation kinetics may be limited by the mass transfer of chemical into the microbes . T he main transformation product was 1,5,9

- cyclododecatriene (CDT) which was formed via a step

- wise reductive dehalogenation of HBCDD . No CO 2 was detected during the study. However, in a study performed according to OECD guideline 301 F (Davis et al.

, 2006 ) it was shown that t,t,t

- CDT can be degrade to CO 2.
40. Degradat ion rate constants of six HBCDD stereoisomers, under anaerobic conditions in sewage sludge have been reported ( Gerecke et al.

, 2006 )

. Experiments were conducted by adding individual target compounds or mixtures to freshly collected digested sewage sludge

. The sewage sludge was amended with yeast and starch. Experiments, performed at 37 °C, with racemic mixtures of individual diastereoisomers showed that (+/

- )

- β

- HBCDD and (+/

- )

- γ

HBCDD degraded faster than (+/

- )

- α

- HBCDD by an estimated factor of 1.6 and 1.8, respectively. Based on the investigations of Davis et al. (200 4 and 2006 ) and Gerecke et al. (2006), α

- HBCD seems to be subject to a slower degradation than β

- and γ

- HBCD.

41. There are no reliable empirical data on the degradation kinetics of HBCD in water. The hydrolysis of HBCD has not been studied. Hydrolysis should howeve, not be considered as a significant route of environmental degradation for this substance due to the low water solubility, the high partitioning to organic carbon, and the lack of hydrol ysable functional groups (OECD 2007). The fact that HBCD has a high log K oc [5.07, PCKocWIN, (v1.66), EPI Suit e (v4.0)], low aqueous water solubility (2.1

- 49 μg/L) and lacks hydrolysable functional groups suggests that the hydrolysis of HBCD would be negligible. According to calculations in EMEPs report on HBCD the physical

- chemical properties of the technical mix ture and γ

- HBCDD stereoisomer gives a half

- life in water of about 5 years (EMEP 2009).

42. According to EBFRIP 2009 the half life for water and soil derived from comparing different model estimations lies in the range 8,5

– 850 days ( ca . 2 years), with a medi an of 85 days ( ca . 3 months) and confidence factor of 10 days. The half life in freshwater and marine sediments lies in the range 6

– 210 days (7 months), with a median of 35 days and confidence factor of 6 days.

43. Several studies using sediment cores show t hat HBCD congeners deposited in marine sediments in Asia and in Europe at the beginning of 1970s/1980s are still present in significant amounts ( Minh et al. 2007, Tanabe 2008, Kohler et al. 2008, Bogdal et al. 2008 ), indicating a higher persistency in sedi ments than derived from experimental studies.

44. The trophic transfer of chemicals in terrestrial or aquatic food webs can also be used to assess persistence. Chemicals that are biologically persistent, that is, slow to break

- down by biologically mediated pro cesses, will increase in concentration with increasing trophic level (TL), that is, biomagnify. The measured field data from various surveys show that HBCD biomagnifie s in some aquatic food chain s . α

- HBCD appears to be the more persistent of the isomers of HBCD, and to biomagn ify more than β

- HBCD and γ

- HBCD.

45. According to EBFRIP 2009 and TemaNord 2008 the findings in the Arctic provide additional evidence that HBCD can persist in the environment lon g enough to be transported over long distances, and that the substance therefore is sufficiently persistent to be of concern within a global scale.



### 2.2.2 Bioaccumulation

46. Several studies in laboratory, in local food webs and local ecosystems confirm the po tential for HBCDs to bioaccumulate and biomagnify. An overall BCF of 18,100 is chosen as representative from two fish studies. The field studies show a general increase of concentrations in biota with increasing trophic level in aquatic and Arctic food web s. No field studies in the terrestric environment have been identified, but two laboratory studies show that HBCD has a potential to bioaccumulate in terrestrial mammals.

47. Some studies have examined the bioaccumulation of HBCD in fish under controlled labor atory environment. The routes of exposure in two studies were via the water (Drottar and Krueger, 2000; Veith et al., 1979) while three studies us ed diet as the route of exposure (Haukås et al. 2009;Law et al., 2005 and 2006 d )

. 48. Veith et al

. (1979) estimated a steady

- state bioconcentration factor (BCF) of 18 1000 for t

- HBCD (technical HBCD) in fathead minnow ( Pimephales promeas ) in a 32

- day flow

- through test. Thirty fish were exposed in the test system and five fish were sampled and analysed on days 2, 4, 8, 16, 24 and 32. The mean test concentration of t

- HBC D was 0.0062 mg/L which was below its aqueous water solubility and the test temperature was 25 ± 0.5 °C.

49. Accumulation of HBCD was also observed in rainbow trout ( Oncorhynchus mykiss ) exposed to nominal concentrations of 0.00034 and 0.0034 mg/L in a flow

- th rough system for 35

- days (Drottar and Krueger, 2000) . The study adhered to t he OECD 305 test method and included a 35

- day depuration period following exposure. Trout exposed to the higher test concentration did not reach steady

- state tissue concentrations over the duration of testing and calculated BCFs were considered less reliab le than those determined at the lower test concentration. A steady

- state BCF value of 13 085 was calculated for the whole fish exposed to t

- HBCD at the lower concentration

.10

50. A s tudy of 1,2,5,6,9,10

- hexabromocyclododecane in carp was conducted by Japanese authorities, based on the OECD Test Guideline 305C for 14 weeks. As the supplied test substance was a mixture, it was separated by high performance liquid chromatography (HPLC) into 5 components which are referred to as components A

- E according to the orde r of the peak appearance. The components B, C and E whose molecular formula is the same as that of the test substance were analysed in this study. For component B, BCFs were 834

- 3070 and 3390

- 16100 at 24 and 2.4μg/L, respectively. For component C, BCFs wer e 816

- 1780 and 3350

- 8950 at 20.2 and 2.02μg/L, respectively. For component E, BCFs were 118

- 418 and 479

- 2030 at 144 and 14.4μg/L, respectively ( Chemicals Inspection and Testing Institute, 1995).

51. Law et al . (2005) measured the BMFs of individual isomers und er a controlled laboratory environment . By exposing juvenile rainbow trout to food intentionally fortified with each isomer the authors were able to calculate BMFs of 9. 2, 4.3 and 7.2 for α

- , β

- and γ

- isomers, respectively. The authors also noted that bioisomerization i.e.
, conversion of one isomer into another, can occur in vivo with this fish species.

52. Haukås et al . (2009) reported on the dietary exposure of juvenile r ainbow trout to HBCD. The authors noted that bioaccumulation of HBCD was significant 6 hours after the single oral exposure and concentrations peaked after 4

- 8 days. After 48 h, the rank order of the relative distribution of the isomers in the fish were li ver > muscle >> brain. The greater distribution to the liver was thought to be due to the greater blood supply to this organ from the stomach and intestine. After 21 days, the relative concentrations of the isomers decreased in the liver and brain but incr eased in muscle. It was hypothesized that the delay in elimination of the isomers from the muscle was due to the lower metabolic activity and circulation of blood to the muscle.

53. Two studies in the laboratory have examined the bioaccumulation of HBCD in mam mals (Chengelis, 2001; Yu and Atallah, 1980). In a 90

- day repeated dose ( t

- HBCD, 1000 mg/kg bw/day) toxicity study with rats, Chengelis (2 001) found that co ncentrations of the α

- isomer were much greater than that of the β

- and γ

- isomers at all sampling time points. The relative percentage of the isomers measured in the rats (α

- : 65

- 70%; β

- : 9

- 15% and γ

- : 14

- 20%) was markedly different to the proportions in th e HBCD formulation used (α

- : 8.9%; β

- : 6.6% and γ

- : 84.5%). Yu and Atallah, (1980) studied the pharmacokinetics of radiolabelled HBCD ( 14 C

- HBCD, purity > 98%) administered to rats as a single oral dose The authors found that the test substance was distribu ted throughout the body with the greatest amounts measured in fat tissue, followed by liver, kidney, lung and gonads. Rapid metabolism to polar compounds occurred in the blood, muscle, liver and kidneys, but HBCD remained mostly unchanged in the fatty tis sue. The study concluded that HBCD accumulated in fatty tissues following repeated exposure.

54. There are numerous reports showing BMFs > 1 for HBCD in aquatic ecosystems. For example, in the Lake Ontario food web, lipid normalized BMFs for bo th the α

- and γ

- isomer were greater than one for many of the feeding relationships (Tomy et al. 2004a) . In some instances, BMFs for the HBCD isomers were greater than those of other known persistent organic pollutants, for example, a BMF of 10.8 for the α

- isomer was reported for the smelt:Mysis feeding relationship and was ca . two times greater than that of p,p

- DDE and ΣPCBs ( Tomy et al., 2004 ) . In Lake Winnipeg, a freshwater lake in central Canada, BMFs of greater than one were also reported for all three HBCD isomers for many of the established predator to prey feeding relationships (Tomy et al. 2004b).

55. Similar findings have been made in the Norwegian Arctic. Sørmo et al. (2006) analyzed representative species from different trophic levels of the polar bear food chai n, using samples collected from 2002 to 2003 at Svalbard in the Norwegian Arctic. HBCD was below detection limits (minimum 0.012 ng/g lw) in the amphipod, Gammarus wilkitzkii

. Concentrations increased from polar cod ( Boreogadus saida , 1.38 to 2.87 ng/g lw) to ringed seal ( Phoca hispida , 14.6 to 34.5 ng/g lw), but decreased at the level of the top predator, polar bear ( Ursus maritimus , 5.31 to 16.51 ng/g lw). HBCD biomagnified strongly from polar cod to ringed seal (BMF of 36.4, based on whole body wet weigh t concentrations), but did not biomagnify from ringed seal to polar bear (BMF of 0.6). Lower levels in the polar bear samples were considered to indicate possible enhanced metabolic capability in the bears. In East Greenland the comparative bioaccumulation , biotransformation and/or biomagnification from East Greenland ringed seal ( Pusa hispida ) blubber to polar bear ( Ursus maritimus ) tissues (adipose, liver and brain) of HBCD and legacy POPs was investigated by Letcher et al . (2009). α

- HBCD was found to onl y bioaccumulate in the polar bear adipose tissue. The ringed seal blubber to polar bear adipose BMF for total

- (α)

- HBCD >1. The authors concluded that even if the metabolism of HBCD in polar bears was enhanced compared to other species, the high exposure of HBCD ensures biomagnification.

56. Morris et al . (2004) reported on biomagnification of HBCD in the North Sea food web (Morris et al., 2004). Although individual BMFs were not reported, the authors suggested that because concentrations of HBCD were higher in species in the top of the food chain it implied that HBCD was biomagnifying. For example, HBCD concentrations in top predators such as harbour seals and harbour porpoise were several orders of magnitude greater than those measured in the aquatic macroinver tebrates such as sea

- star and common whelk. Similarly, HBCD concentrations were high in liver samples from cormorant, a top predator bird

- species and in eggs of the common tern. Lower levels of HBCD were detected in there prey cod and yellow eel (Morris et al., 2004).11

57. According to the study by Munro Jenssen et al . 2007 HBCD was found to have a relatively high presence in the marine food chain in the North

- East Atlantic. Exposure to BFRs in animals from different trophic levels in North

- East Atlantic coas tal marine ecosystems was studied along a latitudinal gradient from southern Norway to Spitsbergen, Svalbard, in the Arctic. HBCD was found in animals from all trophic levels, except for calanoids in two localities; BMFs>1 from Atlantic cod ( Gadus morhua ) to harbor seal ( Phoca vitulina ) at the Norwegian coast and from polar cod ( Boreogadus saida ) to harbor seal ( Phoca vitulina ) at Spit s bergen.

58. Haukås et al . 2009 found the concentration ratio of the diastereoisomers of HBCD to range between 3:1:10 (α:β:γ) in the sediments to 55:1 (α:γ) in the highest trophic level species, suggesting a diastereomer

- specific bioaccumulation in the organisms. The study was conducted in a HBCD

- contaminated Norwegian fjord in a marine fo od chain, measuring levels from sediments and sediment

- dwelling organisms to sea birds. This corresponds with the results of Zhang et al. 2009 in two contaminated streams in China. In this study γ

- HBCD was found to be the dominant diastereoisomer in the se diments (63% of total HBCDs). While α

- HBCD was selectively accumulated in the biotic samples and contributed to 77%, 63% and 63% of total HBCDs in winkle, crucian carp and loach, respectively.

59. Tomy et al . (2008) investigated isomer

- specific accumulation o f HBCD at several trophic levels of an eastern Canadian Arctic marine food web. Alpha

- and γ

- HBCD were present in all species examined (beluga whale, Delphinapterus leucas ; walrus, Odobenus rosmarus ; narwhal, Monodon monoceros ; arctic cod, Boreogadus saida ; deepwater redfish, Sebastes mentella ; shrimp, Pandalus borealis and Hymenodora glacialis ; clam, Mya truncata and Serripes groenlandica ; and mixed zooplankton) with total HBCD concentrations ranging from 0.6 (geometric mean) to 3.9 ng/g lw. There was a si gnificant positive relationship of α

- HBCD with trophic level, indicative of biomagnification throughout the food web, while a significant negative relationship was observed between concentrations of γ

- HBCD and trophic level (i.e., trophic dilution). The β

- isomer was below detection limit in all samples. No clear trend was evident in the diastereomeric profile of the animals, however α

- HBCD contributed greater than 70% of the total HBCD burden in shrimp, redfish, arctic cod, narwhal and beluga, while zooplan kton, clams, and walrus contained greater than 60% γ

- HBCD. The observed differences in diastereoisomer predominance were attributed in part to differing environmental fate and behaviour of the isomers, with the least water soluble γ

- isomer more likely to d iffuse passively from the water column into zooplankton, which have proportionately high lipid content. Similarly, as benthic filter feeders, clams may be more likely to absorb a higher proportion of the γ

- isomer from sediment. The presence of higher
propo rtions of α

- HBCD, such as with the beluga and narwhal, may indicate enhanced metabolic capability based on evidence of stereoisomer

- specific biotransformation of the γ

- isomer to the α

- form (Zege rs et al. 2005; Law et al. 2006d ). This also corresponds with the findings of Tomy et al. ( 2009 ) where the α

- isomer accounted for >95% of the overall burden of HBCDs in the beluga, while the Arctic cod, the primary prey species of beluga in the western Canadian arctic marine food web, had a HBCD

- profile dominated by the γ

- isomer (>77%). In beluga the γ

- isomer contributed to less than 5% of the total

- HBCD

- body burden. The authors concluded that this was further evidence that beluga can bioprocess the γ

- to the α

- isomer.

60. There are studies on three aquatic food webs, on e from the high Canadian Arctic, and two from large freshwater lakes in Canada, all having well characterized trophic levels, that report TMFs > 1 for HBCD (Law et al., 2007; Tomy et al., 2004 ; Tomy et al., 2008)

. Tomy et al

. (2004 ) found that concentrations of ΣHBCD increased with increasing trophic level in a Lake Ontario (Canada) food web. The trophic magnification factor (TMF, defined as the slope of the regression of log concentration vs trophic level) was 6.3 (p<0.001) and greater than tha t of ΣPCBs (5.7). Law et al

. (2007) reported on a TMF of 1.8 (p<0.05) for ΣHBCD in a well

- characterized aquatic food

- web from Lake Winnipeg (Canada) ( . Tomy et al . (2008) also reported on a TMF of 7.4 ( p<0.01) for the α

- isomer in an Arctic aquatic food web from eastern Canada



### 2.2.3 Potential for long

- range environmental transport

61. HBCD is persistent in air, with an estimated half

- life of approximately two days. Studies of and modeling of environmental f

ate and environmental transport of HBCD, as well as field data provide further evidence of the potential for long

- range transport of HBCD. The detected levels in the Arctic atmosphere, biota and environment are strong indicators of the HBCD’s potential for long

- range transport.

62. The atmospheric degradation half

- life of HBCD by gas

- phase re action with hydroxyl radicals ( OH), has not been experimentally measured but can be modeled, providing an estimate (by AopWin v1.91) of 76.8 hours (3.2 days) (TemaNord, 200 8). The model estimates the degradation rate for reaction with hydroxyl radicals. The estimate was obtained by assuming a concentration of 5×10 5 OH molecules/cm 3 and that the reaction takes place 24 hours in a day (these are values used in the European Uni on risk assessments). It is noted that the model is sensitive to the chosen OH

concentration (TemaNord, 2008).

63. Bahm and Khalil (2004) derived a 24 hour global annual average [OH] of 9.2×10 5 molecules·cm

- 3 , with a value of 9.8×10 5 molecules·cm

- 3 for the nor thern hemisphere and 8.5×10 5 molecules·cm

- 3 for the southern hemisphere. These values are consistent with (Prinn et al. 1995) and (Montzka et al. 2000) who deduced [OH] from atmospheric measurements of methyl chloroform, reporting 24 hour global annual ave rage values of 9.7(±0.6)×10 5 and 1.1(±0.2)×10 6 molecules·cm

- 3 respectively. Considering the uncertainty in model estimates of k OH , the half

- life for12

photochemical degradation of HBCD ranges from 0.4 to 4 days and 0.6 to 5.4 days for the northern and southe rn hemisphere respectively ( EBFRIP 2009 ).

64. BSEF (2003) examined the long

- range transport potential (LRTP) of HBCD using four LRTP assessment models (TaPL3

- 2.10, ELPOS, Chemrange

- 2.0 and Globo

- POP), and concluded that HBCD has a low potential to reach
remote areas. LRTP indicators were expressed as the overall characteristic travel distance (CTD) for TaPL3 and ELPOS, the spatial range for Chemrange, and the Arctic contamination potential after 10 years of steady emissions (ACP10) for Globo

- POP. CTDs of 760 an d 784 km in air were predicted using TaPL3 and ELPOS, respectively, while Chemrange estimated a spatial range for HBCD in air of 11% of the earth’s circumference. An ACP10 of 2.28% was estimated from Globo

- POP. The results were comparable with those obtain ed for brominated diphenyl ether flame retardants, in particular the penta

- through decaBDE congeners that also are detected in t he Arctic (Wania and Dugani 2003). BSEF 2003 concludes, that based on the properties of HBCD

, its long

- range transport is likel y to be regulated by transport of aerosols. Overall, the low volatility of HBCD was predicted to result in significant sorption to atmospheric particulates, with the potential for subsequent removal by wet and dry deposition. The transport potential of HBC D was considered to be dependent on the long

- range transport behaviour of the atmospheric particles to which it sorbs. HBCD’s physico

– chemical properties suggest that it may experience active surface

– air exchange as a result of seasonally and diurnally flu ctuating temperatures. Subsequently, this may result in the potential for long

- range transport of the HBCD through a series of deposition/volatilization hops, otherwise known as the ‘‘grasshopper effect”, described by Gouin and Harner (2003). This assumpti on is supported by environmental data. The concentrations of HBCD in bulk samples collected in urban and remote sites in Sweden and Finland had a clear seasonal and diurnal flux rate (Remberger et al

. 2004). Precipitation samples collected from the Great L akes Basin varied with the season (Backus et al

. 2005), with a seasonal peak of HBCD in the winter time. In the study of Yu et al. (2008) in Southern China a large variable percentage of HBCD (69.1

– 97.3%) existed in the particle phase, suggesting that long

- range transport of HBCDs is governed by environmental conditions. The diastereoisomers of HBCD in the samples were both in gaseous form and particle bounded. The results of enantiomeric fractions indicated that β

- and γ

- HBCD were racemic mixtures in the tested air samples and that (

- )

- α

- HBCD was slightly enriched in the typical urban site, implying that the HBCDs in these studied sites might be a mixture of emissions from both local sources and soil

- air exchange.

65. Based on model estimates HBCD seems to have a “low” to “moderately

- low” long

- range transport potential ( EBFRIP 2009 ); however using the same modeling tools, the report by EBFRIP 2009 showed by bench

- marking that HBCD’s potential is in the range of legacy POPs. Detectable levels in remote regions suggest that long

- range transport is occurring on a larger scale than predicted by the models. The models do not include the full potential of transport by the “grass

- hopper effect” and some of the environmental co nditions typical for the wind systems of the Arctic, for example the Arctic haze events. According to the AMAP report from 2009, transport of less volatile BRFs does take place when there are large numbers of particles in Arctic air, during the Arctic haze . Therefore, periods of strong winds and no precipitation may lead to longer transport distances than the models predict for BRFs (AMAP 2009). Brown and Wania (2008) identified HBCD as a potential Arctic contaminant based on an atmospheric oxidation half

- l ife of greater than two days and structural similarities to known long range transported Arctic contaminants. This is supported by the result of modeling done by EMEP, which reach the same conclusion (EMEP 2009). This is also supported by recent field stud ies and environmental monitoring. HBCD concentrations were found to be higher in the atmosphere over the Norwegian Arctic, when the air transport came from polluted areas on the continents and lower when the air transport came from the North Pole Sea and S candinavia ( SFT 2008 ).

66. Remberger et al . (2004) analyzed air and rainfall bulk deposition samples collected in 2000 and 2001 from various locations in Sweden. Sampling sites were selected to include possible point sources (for example, a manufacturing facil ity for expanded polystyrene (XPS) with HBCD, a landfill site for construction and demolition waste, and a textile industry facility) as well as diffuse sources in an urban environment (Stockholm). In order to evaluate the potential for long

- range transpor t, air samples were collected at three sites (two in Sweden and one in the Finnish Arctic) considered remote. The depositional flux (i.e., concentration per day) measured in urban regions of Sweden were 5.5 and 366 ng/m 2 . Fluxes measured in more remote loc ations of Sweden and Finland were generally smaller and ranged from 0.02 to 13 ng/m 2 . Air concentrations at sites near potential point sources ranged from 0.013 to 1070 ng/m 3 while those at the urban stations were 0.076 to 0.61 ng/m 3 . For both groups, wint er concentrations were higher than those in summer and fall. The highest concentration, 1070 ng/m 3 , was recorded close to the exhaust for the air ventilation system of an XPS manufacturing facility. Levels at the remote sampling locations varied from below the detection limit (0.001 ng/m 3 ) to 0.28 ng/m 3

. 67. Precipitation samples collected from the Great Lakes Basin contained as much as 35 ng HBCD /L, with the highest levels occurring in the winter months (Backus et al., 2005). The researchers hypothesized that observed winter peaks resulted from increased scavenging efficiency of snow compared with rain, as well as higher concentrations in the particle phase during winter. All three major diastereomers were detected, with an average percent distribution of 77, 15 and 8% for α

- , β

- and γ

- HBCD, respectively.

68. In the study by Yu et al . (2008), air samples were collected from four sites in the city of Guangzhou, a typical fast developing metropolitan of South China, for quantifying the concentration levels, diasteroi somer profiles, gas

-13

particle distribution and enantiomeric fractions of HBCDs. The analysis on diastereoisomer profiles indicated that α

HBCD (59

– 68%) was the dominant isomer and β

- HBCD was a minor isomer in all air samples, which appeared to be different from commercial products. For gas

- particle distribution on each diastereoisomers, the similar trend in all the sites was observed that the percentage of β

- HBCD in gas phase is higher than those in particle phase (Figure 3), whereas the percentage of α and γ

- HBCD in gas phase is lower than those in particle phase. The average relative abundances of three diastereoisomers in gas

- and particle

- phases are 55/25/20 and 62/13/25, respectively. This might be caused by slightly different physicochemical properties of three diasteroeoisomers.

69. Air samples were collected at five sites (urban, semiurban, agricultural, and remote) from Lake Michigan through the U. S. Midwest to the Gulf of Mexico every 12 days during 2002

- 2003 using high

- volume samplers to study the spa tial trends of brominated flame retardants, including the HBCDs (Hoh and Hites 2005). Hexabromocyclododecanes (HBCDs) were detected only in the particle phase. The mean and median values of the HBCDs were by far the highest at the Chicago site (0.9

- 9.6 pg/ m 3 ), suggesting that urban areas are the source of this compound. However, the highest individual HBCD concentration was observed at the Arkansas site (0.2

- 11 pg/m 3 ). Tracing the air using bakward trajectories suggested that manufacturing areas in southern Arkansas could be the source region. The HBCD concentrations at the remote sites in Michigan (0.2

- 8.0 pg/m 3 ) and Louisiana (0.16

- 6.2 pg/m 3 ) were somewhat lower than the background concentrations observed in Sweden (Remberger et al. 2004). The reason for t his difference may be that HBCDs replaced the penta

- and octa

- BDE products earlier in Europe than in North America.

70. SFT 2008 reports air levels from two Norwegian background monitoring stations Birkenes in the Southern part of Norway and Zeppelinfjell, Ny

- Ålesund, Svalbard in the Nowegian Arctic. These stations are included in a Comprehensive Atmospheric Monitoring Programme a part of OSPARs s studies of land based pollution to the marine areas. Both of the stations are located in long distances from local pollution sources. The sampling was carried out by a new developed high volume sampler; the analyses of HBCD were performed both with the gas and particular phase; and 12 samples from Birkenes and 14 from Zeppelinfjell were analysed. The HBCD levels at the background stations were described as low and no significant difference between the levels from the two stations were found. The HBCD concentrations in these samples varied between 0.14 and 20.3 pg/m 3 in the Birkenes monitoring station and between 0.18 an d 23.0 pg/m 3 , in the Zeppelinfjell monitoring station. γ

- HBCD dominated in all samples. Trajectory calculations for air in the periods with highest HBCD concentration showed that the air transport went mainly from North America and Europe, and mainly from the North Atlantic Ocean and The North Pole Sea in the periods with lowest HBCD concentration. They also showed that air from polluted regions will reach the remote areas during short time.

71. Canada’s Northern Contaminants Program (NCP) has been monitoring a tmospheric concentrations of POPs in the Canadian and Russian high Arctic since 1992. Air samples are collected on a weekly basis using a high

- volume air sampler that collects ca. 13 000 m 3 of air. Composite samples are generated from the glass fibre filte r (particulate phase) and polyurethane foam (gaseous phase). HBCD air concentrations in samples collected from Alert, Tagish and Dunai from the Canadian and Russian Arctic was 1.8 pg/m 3 ( PWGSC

- INAC

- NCP, 2003 ). This is similar to ΣHBCD air concentrations in samples collected from Alert, Nunavut, in the Canadian high Arctic in 2006 and 2007, ca. 1

- 2 pg/m 3

, and similar to BDE

- 99 (Xiao et al. 2010) . Peak ΣHBCD concentrations were ca. 5 pg/m 3 and only 2 times smaller than that of peak BDE

- 47 concentrations. ΣHBCD concentrations of ca. 5

- 6 pg/cm 3 were reported by Bergander et al. (1995) in air samples collected between 1990 and 1 991 at two remote regions in Sweden: Ammarnäs (65 o 58’N, 16 o 12’E, mountainous region of Swedish Lappland and ca. 100 km south of the Arctic Circle) and at Hoburgen (the peninsula of the southern tip of Gotland) (Bergander et al. 1995)

.


### 2.3 Exposure

2.3.1 Environmental levels and trends

72. HBCD i s widespread in the global environment, with high levels in the top predators

. According to Covaci et al

. (2006) high concentrations have been measured in top predators, such as m arine mammals and birds of prey. Zegers et al. (Zegers et al. 2005) published data on HBCD concentrations in two species of marine top predators, the harbor porpoise ( Phocoena phocoena ) and the common dolphin ( Delphinus delphis ), from different European seas. The highest HBCD concentrations were measured in porpoises stranded on th e Irish and Scottish coasts of the Irish Sea (median concentration 2900 ng/g lw, maximum 9600 ng/g lw) and the northwest coast of Scotland (5100 ng/g lw). The median concentrations in porpoises from other areas were 1200 ng/g lw on the south coast of Irela nd, 1100 ng/g lw on the coasts of t he Netherlands, Belgium, and the North Sea coast of France, 770 ng/g lw for the east coast of Scotland, and 100 ng/g lw for the coast of Galicia (Spain). The median HBCD concentrations in the common dolphin, a pelagic

mar ine mammal species feeding primarily over the continental shelf and in offshore waters, were 900 ng/g lw on the west coast of Ireland, 400 ng/g lw in the English Channel coast of France, and 200 ng/g lw in Galicia (Zegers et al., 2005). Law et al . (2006 d ) studied HBCD in the blubber of harbour porpoises from the UK during the period 1994

– 2003. Eighty

- five animals were analysed for HBCD. α

- HBCD dominated over the other isomers and was detected in all samples at concentrations ranging from 10 to 19 200 μg kg

- 1 wet weight (11

– 21,300 μg kg

- 1 on a lipid basis).14

73. De Boer et al. ( 2004 ) have determined HBCDs in eggs of peregrine falcons ( Falco peregrinus ) (71 ng/g lw

– 1200 ng/g lw) and muscle in sparrowhawks ( Accipiter nisus ) (84

- 19000 ng/g lw) from the U.K., with d etection frequencies of 30% and 20%, respectively. Levels of 330

- 7100 ng/g lw were found in eggs from the common tern ( Sterna hirundo ) in the Netherlands in a study by Morris et al. (2004) and levels of 34

- 2400 ng/g lw were found in eggs of peregrine falco ns in Sweden sampled 1991

- 1999.

74. Due to their high position in the food chain and the elevated exposure in the aquatic environment, fish often exhibit high residues of contaminants. Not surprisingly, HBCDs have been detected in many studies in both freshwat er and marine biota (Covaci et al. 2006)). Concentrations of HBCDs in fish downstream of an HBCD manufacturing plant on the River Skerne (Durham, U.K.) were very high, with levels up to 10 275 ng/g lw (Allchin and Morris 2003 ). Concentrations of HBCDs were mostly between 10 and 1000 ng/g lw in urban/suburban regions of Europe, while levels in the North American Great Lakes were lower by approximately 1 order of magnitude (3

- 80 ng/g lw) (Covaci et al. 2006). The wide spread spatial occurrence of HBCD in the aquatic environment was illustrated by Ueno et al (2006) that measured HBCD in the muscle of skipjack tuna (1997

- 2001) in the Asia

- Pacific area. Levels ranged between 0,28 ng/g lw in the waters outside Brazil to 45 ng/g lw in the waters outside Japan. The study by Xian et al. 2008 in the Yangtze river in China in 2006 examined HBCD levels in the fresh water fishes examined. Levels ranged between 12 ng/g ww in the muscle of grass carp to 330 ng/g ww in the muscle of mandarine fish.

75. HBCD is ubiquitous in the Arctic environment and has found to be widespread in the Arctic food webs (de Wit et al. 2006, 2009). The Arctic top predators are at risk due to environmental changes and a high burden of persistent contaminants (AMAP 2009). A temportal trend study of HB CD and PBDEs in eggs of herring gulls ( Larus argentatus ), Atlantic puffins ( Fratercula arctica ), and black

- legged kittiwakes ( Rissa tridactyla ) in North Norway showed that levels of alpha

- HBCD increased in all species throughout the 20

- year period. Miljete ig et al . 2009) compared levels of contaminants in eggs between four Arctic colonies of ivory sea gull, one in the Norwegian Arctic (Svalbard ) and three in the Russian Arctic (Franz Josef land and Severnaya Zemlya). The contaminant levels presented are amo ng the highest reported in Arctic seabirds and were identified as an important stressor in a species already in risk due to environmental change. The population of ivory gulls ( Pagophila eburnea ) in the Arctic is decreasing and the species is on the The IU CN Red List of Threatened Species

. The levels of HBCD in the study ranged between 14,0 and 272 ng/g lw HBCD. Tomy et al. (2007) detected ΣHBCD concentrations in beluga whales ( Delphinapterus leucas ) in the Canadian Arctic, a species protected by the Convention on migratory species (the Bonn convention). The concentrations were in the range of 10

- 20 ng/g lw. Muir et al. (2006) detected level s of HBCD in polar bears ( Ursus maritimus ) in several populations in the Arctic region. The highest level detected were in the female bears from the Svalbard area (109 ng/g lw). Polar bears are listed on the IUCN Red List of vulnerable species.

76. The top pre dators in the Arctic are especially vulnerable. HBCD accumulated in the fat is released and transferered to vital organs. In the report of Knudsen et al

. ( 2007 ) glaucous gulls ( Larus hyperboreus ) and great black

backed gulls ( Larus marinus ) found dead at B jørnøya in the Norwegian Arctic between 2003

- 2005 were analysed for
contam inants,such as legacy POPs, mercury and emerging pollutants in the Arctic. The levels found for some of the contaminants, including HBCD, were higher than previously reported for fre e

- living glaucous gulls from Bjørnøya and other bird species in the Arctic and in Europe

. The α

- HBCD concentrations in the brain and liver samples of the glaucous gulls ranged from 5,1 ng/g lw to 475 ng/g lw, and from 195 ng/g lw to 15027 ng/ g lw, respectively. The levels ranged from 44

. 7

- 44

. 8 ng/g lw in the brain samples and 1881

- 3699 ng/g l w in the liver samples from the great black

- backed gulls . For a comparison the levels found in cormorant liver ( Phalacrocorax carbo ) sampled in England 1999

- 2000 were in the range 138

- 1320 ng/g lw (Morris et al. 2004). Some 40

- 45 % of the sea birds were fo und to be completely or severely emaciated. There were also observations of dying glaucous gull on Bjørnøya with apparently abnormal behavior. According to the authors this may indicate that high levels of contaminants, including high levels of HBCD may ha ve been a contributing factor to the birds’ death, directly or indirectly.

77. According to the reviews done by Covaci et al. (2006), Law et al. (2008) and Tanabe et al. (2008) the HBCD levels in the environment is generally increasing in all matrixes in the e nvironment, and seem to correlate with the increasing use of HBCD. The reviews cover over 100 published scientific studies (up to 2007) performed in North America, Europe, the Arctic, Asia and the South Pacific region. The review covers variety environment al compartments (atmosphere, indoor and outdoor air, sewage sludge’s, soils and sediments) and a variety of biological samples and food chains. In the review by de Wit et al. (2009) the few available temporal studies in the Arctic indicated an increase in biota of HBCD, no or unclear trend, depending on species and locality. A ccording to Managaki et al. 2009, the increasing trend in releases of HBCD are in good agreement with concentration data f rom sediment cores (Minh et al. 2007) and historic trends of H BCD levels in human blood in Japan (Kakimoto et al. 2008).

78. Several s ediment core analyses performed in Asia and Europe show higher levels of HBCD in the top layers and lower concentrations in the deeper layers. This findings correlate with the trend in use of HBCD. HBCD was present in three sediment cores and six surface sediment samples collected in 2002 from Tokyo Bay, Japan (Minh et al. 2007). HBCDs first appeared in the mid

- 1970s and concentrations observed in the cores have increased since then. Based on the data, Tanabe (2008) estimated concentration doubling times of 7.1 to 12 years for HBCD in the sediment. HBCD was first detected in s ediments from Lake Greifensee in the mid

- 1980s (Kohler et al., 2008) . HBCD concentrations in15

the cores then increased in an exponential manner with a peak in 2001 (2.5 ng/g, dry weight). Bogdal et al. (2008) reported on increasing HB CD concentrations up to the surface layer in two sediment cores from Lake Thun, (Bogdal et al., 2008)

. 79. A temporal trend study of HBCD and PBDEs in eggs of herring gulls ( Larus argentatus ), Atlantic puffins ( Fratercula arctica ), and black

- legged kittiwakes ( Rissa tridactyla ) in North Norway (Helgason et al. 2 009) showed that levels of alpha

- HBCD increased in all species throughout the 20

- year period . HBCD concentrations in bird egg sampled in the Northern Norway and Svalbard (Knudsen et al. 2005) exhibit a remarkably consistent increasing trend since 1983. Sta pleton et al. (Stapleton et al. 2006) have shown an exponential increase in the HBCD concentrations with a doubling time of approximately 2 years in California sea lions stranded between 1993 and 2003. Law et al (2008) have continued their analysis of HBCD in UK harbor porpoises, which now includes 223 animals spread over 13 years (1994

- 2006). The within year variation is 4

- 6 orders of magnitude, which makes any conclusions uncertain. However, the mean values indicate increasing concentrations from the mid

- nineties (30

- 70ug/kg lipid weight) with a steep and statistically significant increase between 2000 and 2001 resulting in a mean concentration of 5,450 ug/kg 2003. The steep increase was followed by a corresponding steep decrease between 2003 and 2004, res ulting in a concentration of 817 ug/kg in 2006. ( Concentrations of PBDEs and HBCDs in marine mammals from Japanese and Chinese coastal waters have drastically increased during the last 30 years (Tanabe et al. 2008). In the samples from Japan, temporal chan ges in BFR levels were associated with trends in production/use of the commercial formulations. Since the withdrawal of some PBDE products from the Japanese market in the 1990s, concentrations of HBCDs appear to exceed those of PBDEs, reflecting increasing usage of HBCDs over PBDEs. Considering its increasing importance as an alternative for PBDEs, it can be anticipated that the concentrations of HBCDs may further increase in the near future (Tanabe et al. 2008).

80. As exemplified by a number of studies, the c oncentrations in the European environment are often higher than those measured in biota in North America and the Asia

- Pacific region (Hoh and Hites, 2005; Tomy et al., 2004; Peck et al., 2005 ; Stapleton et al., 2006 ; Janák et al., 2005; Morris et al., 2004 ; Zegers et al., 2005; Yu et al., 2008; Kajiwara et al., 2006 ; Isobe et al., 200 8 ; see reviews by Tanabe et al., 2008 and Law et al., 2008.) The levels in the Asia

- Pacific region and North America is found to be in the lower range of the levels detected in sea mammals in Europe (Covaci et al. 2006). The results likely reflect the substantially higher market demand for HBCD in Europe relative to other regions of the world (Law et al., 2008; Tanabe et al., 2008). However, according to the review by Tanabe et al . (2008) HBCDs is also widespread in the Asia

- Pacific region. Tanabe et al . 2008 concluded that HBCDs were detected in all the matrices examined

- mussels, fish, marine mammals, human breast milk, house and office dust. The highest concentrations of BFRs were observed in the samples from Korea, South China, and Japan. A similar pattern emerges from other Asian studies. Assessing HBCD levels in skipjack tuna samples collected from thirteen offshore locations in the Asia

- Pacific region during 1997

– 2001 (Uen o et al., 2006) found HBCD levels that were higher in mid

- latitude areas of the Far East, as relatively high concentrations were detected in samples collected around Japan, the East China Sea and the North Pacific. In two other field studies the spatial di stribution of HBCDs in the Asia

- Pacific region were assessed by analyzing fat tissue of marine mammals from Japan and Hong Kong ( Kajiwara et al., 2006 ; Isobe et al., 200 8 ). The detected HBCD levels were higher in cetaceans from Japan than in cetaceans from Hong Kong, probably due to intensive usage in Japan in recent years. HBCD levels in mammals from Hong Kong and Japan ranged from 21 to 380 ng g

- 1 lipid wt and from 330 to 940 ng g

- 1 lipid wt, respectively. For comparison, measured levels in blubber sample d from white

- sided dolphins at the eastern coast of US between 1993

- 2004 were in the range of19

- 380 ng/g lw (14

- 280 ng/g ww) (Peck et al. 2008). Tanabe et al. (2008) concluded that the high levels of BFRs, including the HBCD levels, around the East China S ea could be due to the presence of a number of electronics manufacturing industries in this region.

81. According to Covaci et al . (2006) there is a general trend to higher environmental HBCD concentrations (air, sediment, and fish) near point sources (plants producing or processing HBCDs) and in urban areas, than in locations with no obvious sources of HBCDs. Concentrations of HBCDs are often elevated by at least one order of magnitude in the vicinity of plants either producing or using HBCDs. Several hot spo ts have been identified in Europe: the rivers Viskan (Sweden), Tees and Skerne (U.K.), Cinca (Spain), and the Western Scheldt estuary (Netherlands) (Covaci et al. 2006). All of these sites were related to present or former production facilities for HBCDs o r HBCD retarded materials. Higher HBCD concentrations are also frequently found near urban centers and industrial sites ( Minh et al ., 2007; Morris et al., 2004; Sellström et al., 1998; Eljarrat et al., 2009 ; Hoh and Hites, 2005). In particular, soil samples collected near HBCD

- processing factories are found to have high levels. More precisely Remberger et al. (2004) and Petersen et al. (200 5 ) measured HBCDs, ranging between 111 and 23 200 ng/g dry in soil samples collected outside an XPS producing plant. The levels fell with increasing distance to the point source.

82. Janák et al. (2005) measured concentrations in shrimp and several fish species collected in 2001 from five locations in the Western Scheldt estuary (The Netherlands). Two sampling sites were situat ed in the vicinity of an HBCD production plant at Terneuzen ( ICL

- IP Terneuzen formally known as, Broomchemie, 7500 tons HBCD/year), while the remaining three were further up the estuary toward Antwerp, Belgium. The α

- isomer was present in all 28 samples (2 shrimp, 26 fish) and dominated the isomeric profile. Levels of γ

- HBCD were above detection limits in 20 samples, while β

- HBCD was quantified in only three samples. Highest concentrations (1100 and 680 ng/g lw α

- isomer16

in sole, Solea solea , muscle and live r, respectively) were measured nearest to a potential source at ICL

- IP Terneuzen. Levels in the samples decreased with increasing distance from this site.

83. Arctic monitoring reveals the presence of HBCD in organisms at all trophic levels of the food chain, and suggests that HBCD is bioavailable even to organisms in remote areas of the world and at all levels of biological organization (AMAP, 2009). Moreover, Arctic data suggest that the organisms’ external living conditions can be of vital importance to the internal exposure situation. More specifically, Arctic mammals and birds go through a seasonal starvation period in the winter time where they use up their fat reserves to get enough energy to survive. However, when body lipids are mobilized contaminants accumulated in fat tissue becomes redistributed to vital organs such as liver and brain which have relatively high lipid content and blood perfusion (Knudsen et al., 2007).

84. The recent findings of Heeb et al. (2008) are also important to the issue of bioava ilability. Heeb et al . (2008) documented conversion of the γ

- isomer into α

- HBCD at temperatures exceeding 100°C. In a wider context, this finding suggests that finished products subjected to high temperatures during processing, and the releases during the service life of HBCD containing articles, as well as the releases from the industrial use of HBCD in textiles and polystyrene, may carry a higher proportion of α

- isomer than is present in the original formulation. This in turn may increase the potential fo r organism exposure to the α

- isomer, and may in part explain the predominance of α

- HBCD in biota (EU RAR 2008). Compared to α

- HBCD, γ

- and β

- isomers are commonly present at lower levels or below detection limits (EU RAR, 2008).

85. Generally, the isomeric pa ttern observed in biota varies with species. This may reflect species differences in the external exposure situation, uptake, metabolism or depuration of the three isomers. Whilst several studies show that both α

- HBCD and γ

- HBCD have a tendency to bioaccum ulate in organisms, α

- HBCD reportedly has a higher potential to biomagnify than γ

- HBCD (see section 2.2.2). The α

- isomer of HBCD therefore dominates especially at higher trophic levels in the food webs. Selective biotransformation and bioisomerization, whe reby the other stereoisomers are preferentially converted to α

- HBCD, contributes to this pattern (Law et al ., 2006 d ; Janák et al ., 2005; Zegers et al., 2005; see European Commission 2008 for overview). Another mechanism of importance can be a selective upt ake of α

HBCD and/or differences in the steroisomeric and enantiomeric profile of prey organisms. In peregrine falcons and white

- ta i led sea eagle only α

- HBCD was detected, and in terns and guillemots it was the predominant diastereoisomer (Janák et al., 20 08). This is in agreement with other studies of HBCD diastereoisomers in birds (Leonards et al., 2004; Morris et al., 2004; Knudsen et al., 2005). In the Janák et al . (2008) study there were also clear differences in enantiomer fractions of (+)

- α

- HBCD and (

- )

- α

- HBCD between the different bird species as well as between guillemot and its prey, herring, a result that also indicates species

- specific stereoselective mechanisms for uptake and metabolism of the enantiomers of α

- HBCD in birds. At the bottom of the food chain a different exposure pattern emerges. For example in a study by Tomy et al. (2008), the main isomer in bottom dwelling filter feeders and zooplankton was found to be γ

- HBCD. This is the least water soluble stereoisomer and is therefore more lik ely to diffuse passively from the water column into zooplankton, which have relatively high lipid content and may therefore be more likely to absorb a higher proportion of the γ

- isomer from sediment.


### 2.3.2 Human exposure

86. Humans, like other organisms, are exposed to HBCD via multiple sources: food, dust, air, textiles, polystyrene products and to a lesser degree electronic equipment (for overview see TemaNord 2008; European Commission 2008 ; AMAP 2009; Covaci et al., 2006). Human exposure to HBCD may be eit her dermal or oral, and may also result from inhalation of vapor and particles ( European Commission 2008 ). In the work environment direct dermal exposure and inhalation of fine HBCD dust or particles are particular concerns. In a study by Thomsen et al . (2 007) industrial workers at plants producing EPS with HBCD had elevated HBCD

- levels (i.e. 6

- 856 ng/g lw). In non

- occupationally exposed individuals indirect exposure via the environment, be it oral, dermal or by inhalation, is the main concern. In a study b y Stapleton et al. (2009) HBCD levels in dust samples from indoor environments ranged from <4.5 ng/g to a maximum of 130,200 ng/g w ith a median value of 230 ng/g. A study by Abdallah et al 2009. found HBDE in household air (median concentration 180 pg m

- 3) , household dust (median concentration 1300 ng/g), offices (median concentration 760 ng/g), and cars (medi an concentration 13,000 ng/g). Serum/ blood levels in non

- occupationally exposed individuals are typically much lower (i.e. 0.005

- 6.9 ng/ g lw) though the data indicates potentially significant sources of exposure (see Swedish Chemical Agency 2008 for overview )

. 87. Reported dietary exposure levels in humans vary g lobally and regionally (Shi et al. 2009, Roosens et al. 2009). Surveys in Europe and the US reve al dietary exposure levels for HBCD in the range of <0.01

- 5 ng/g w/w (see Roosens et al. 2009 for overview). Fatty foods of animal origin such as meat and fish and/ or from contaminated areas are likely a major source of dietary human exposure, and the exp osure situation closely depend on the consumption of those products in the population (e.g. Shi et al. 2009; Remberger et al 2004, Lind et al 2002 , Driffield et al 2008 ). Among all dietary samples, the highest HBCD concentrations (up to 9.4 ng/g w/w) are r eported for fish (Knutsen et al. 2008, Remberger et al. 2004, Allchin and Morris 2003). Accordingly in Norway and Scandinavia where fish is an important part of the diet intake of fish from a contaminated Norwegian lake and ingestion of fat

- rich marine fis h have been found to closely correlate with serum HBCD levels (Thomsen et al. 2008; Knutsen et al. 2008 ; Remberger et al. 2004; Lind et17

al. 2002). Eggs are another source of human exposure with levels in organic yolks in Canada ranging from 3.4

– 1110 ng/g lipid9. A survey of home

- grown chicken eggs sampled near contamined sites in developing countries showed significant levels including in Mexico (91 ng/g lipid), Uruguay (89 ng/g lipid), Slovakia (89 ng/g lipid), Turkey (43 ng/g lipid) and Kenya (160 ng/g lipid). However, it was shown by Driffield et al. (200 8 ), who assessed 19 different food groups representing the UK diet for 2004 for BFRs, vegetables may contain HBCD at similar concentrations as have been reported for meat and fish. The presence of HBCD in vegetables and vegetable oils and fats may arise from the presence of this substance in sewage sludge and the subsequent use of sewage sludge as a food crop fertilizer (e.g. Kupper et al. 2008, Brändli et al. 2007). Stereoisomeric patterns in food sampl es suggest both global and regional variation, as well as stereosiomeric differences depending on food type (Roosens et al., 2009; Shi et al., 2009). Although fish and meat are the major dietary sources in Europe, USA and China (Covaci et al., 2006; Schect er et al., 2008; Thomsen et al., 2008; Shi et al., 2008 ), two British studies assessing HBCD exposure in humans also highlight indoor air, and in particular dust, as important sources of exposure in both adults and toddlers (Abdallah et al. 2008a and b ). F or a toddler of 10 kg who ingests an estimated 200 mg dust/ day (HBCD contamination at the 95 th percentile) the intake via dust may exceed 10 times the levels received via diet alone (Abdallah et al. 2008). In the study of Roosens et al. ( 2009 ) , daily expo sure from food and dust was found to be approximately similar in magnitude, and HBCD concentrations in serum only correlated significantly with estimates of exposure via dust. As postulated by the authors, exposure to dust may be an important exposure rout e because exposure remains more co nstant over time than exposure from food which depends on the more periodic intake of contaminated food items (Roosens et al. 2009). As fish and meat are common food commodities in many regions, diet could cause potentiall y higher exposures than dust, denpending on consumption rates, dietary patterns and geographic distribution.

88. As a result of continuous exposure in homes, offices and cars, HBCD is found in human adipose tissue (Pulkrabová et al., 2009; Johnson

- Restrepo et al, 200 8 ; Antignac et al., 2008; Abdallah and Harrad, 2009) and blood (Weiss et al., 2004; Weiss et al., 2006; Lopez et al., 2004; Odland et al., 2005 ; Brandsma et al., 200 9 ; Thomsen et al., 2007; Meijer et al., 2008; Roosens et al., 2009). HBCD is transfe rred across the human placenta to the fetus (Mejier et al., 2008), and HBCD has been detected in breast milk in Europe (Covaci et al., 2006; Lignell et al., 2009; Eljarrat et al., 2009;, Colles et al., 2008; Polder et al., 2008a; Polder et al., 2008b; Fäng stöm et al., 2008; Antignac et al., 2008), Asia (Kakimoto et al., 2008; Shi et al., 2009; Malarvannan et al.,2009; Tue et al., 2010), Russia (Polder et al., 2008), Mexico (Lopez et al., 2004) and USA (Schecter et al., 2008). Hence exposure to HBCD occurs a t critical stages of human development, both during pregnancy and postnatally via breast milk. Reported concentrations of HBCD in breast milk range from below detection limit to 188 ng HBCD/ g lw ). H owever, according to EBFRIP 2009 the typical range of tot al HBCD concentrations in human breast milk in populations inhabiting industrialized areas appears to be <1 to 5 ng·/ g lw. Geographically, the highest HBCD levels have been found in mothers’ milk from Northern Spain (Eljarrat et al., 2009; Guerra et al., 2008). Reported levels ranged from 0.8 to

- 188 ng/ g lw, with median values of 27 and 26 ng/g lw. HBCD levels measured in Russian breast milk vary from 0.2

- 1.7 ng/g lw and fall within the detected range for HBCD in Scandinavian and Belgian women of 0.085

- 3. 2 ng/ g lipid, but notably only 3/23 and 8/14 of the samples from Arkhangelsk and Murmansk respectively had HBCD levels that were above the quantification limit (Polder et al., 2008).

89. As documented by a Japanese study (Kakimoto, 2008), HBCD levels in huma n milk appear to mirror the market consumption of HBCD; In mothers’ milk from Japanese women (age 25

– 29) HBCD levels were below the detection limit in all samples collected during the 10

- year period from 1973

- 1983, but then increased from 1988 and onwards. In the peri od 1988

- 2006, α

- HBCD was detected in all 11 pooled milk samples with levels ranging from 0.4

- 1.9 ng/ g lw. Mean total HBCD concentrations over the period 2000

– 2006 ranged from 1

- 4 ng/ g lw. The levels reported from this Japanse study are higher than valu es reported for women from Northern Norway where HBCD was detected in only 1/10 samples, at a concentration of 0.13 ng/g lw (Polder et al., 2008a). In a study from Stockholm, Sweden, temporal trends show an increase in HBCD levels in milk up to 2002 after which a leveling occur. According to calculations made in the EU risk assessment (European Commission 2008) intake of HBCD in via breast milk is 1.5 ng/ kg bw/ day for 0

- 3 month olds and 5.6 ng/ kg bw/ day in 3

- 12 month old babies. However, with the levels found in mothers’ milk from Northern Spain (A Coruña, Galicia). Eljarrat et al., 2008 calculated the intake to be 175 ng/ kg bw/ day for 1 month olds . This is 12 times higher than the estimated daily intake (EDI) for 0

- 3 months

- old infants (EU risk

assess ment, European Commission 2008) and 25

- 1458 times higher than the EDI for adults in Sweden, Netherlands, United Kingdom and N orway ( Swedish Chemical Agency 2009 ; Eljarrat et al ., 2009). Data from China, which are base d on α

- HBCD levels in the range of <LOD to 2.78 ng/ g as measured in mothers’ milk from 1237 donors from 12 different provinces, suggest an EDI of 5.84 ng / kg bw/ day when assuming a body weight of 7.8 kg and a milk consumption for 6

- month

- olds as specifi ed by the U.S. Environmental Protection Agency ( Exposure Factors Handbook US EPA ). This value is approximately 3 to 10 times lower than the calculated EDI for infants in the EU region which were proposed to be 15 and 56 ng/ kg bw/ day for 0

- 3 months and 3

- 12 month old infants, respectively (European Commission 2008). Still, the EDI for infants in China is estimated to be 14 times that of Chinese adults, where an EDI of 0.432 ng / kg bw/ day was given for a “reference” man (Shi et al., 2009).

90. Although α

- HBCD, followed by the γ

- and β

- HBCD, appears to be the predominant diastereoisomer in all biota, including humans ( European Commission 2008 ), the profiles of α

- , β

- , and γ

- HBCD isomers in human tissues are not18

consistent and differ somewhat betw een studies. In a pooled serum sample from an elderly population in Sweden, the most abundant stereoisomer was α

- HBCD (Weiss et al., 2006). On the other hand, in serum of workers from an industrial polystyrene plant in Scandinavia Thomsen et al. (2007) fou nd higher concentrations of γ

- HBCD than the other HBCD stereoisomers. In the study of Roosens et al. 2009 γ

- HBCD was the dominant isomer present in adipose tissue and accounted for 83% of the total HBCD concentrations, followed by α

- HBCD (17%). The α

- HBC D isomer appears to dominate in some breast milk samples ( Shi et al., 2009; S ch ecter et al., 2008 ) but not in others where γ

HBCD appears to dominate (Eljarrat et al., 2009; Guerra et al., 2008). These differences appear to primarily reflect the external e xposure situation (time, dose and stereosiomeric pattern). In this context it should be noted that, although technical HBCD mixtures primarily contain γ

– HBCD, thermal isomerization of HBCD can occur and may result in the enrichment of α

– HBCD both during the polymer extrusion process, and during incorporation of HBCD in textiles (Larsen and Ecker, 1986; Kajiwara et al., 2009). Similarly, in light

- exposed dust, a photolytically

- mediated shift from
γ

- HBCD to α

- HBCD may occur (Harrad et al., 2009). As illustr ated by Roosens et al. (2009), such environmental changes are reflected in the human tissue samples, but may in addition be influenced by in vivo bioisomerization of β

and γ

- HBCD to α

– HBCD and a more rapid biotransformation of β

- and γ

- HBCD than α

– HBCD (Zegers et al., 2005; Law et al., 2006 c ). In vivo studies with rats suggest that HBCD is also debrominated from HBCD to PBCDe and TBCDe. In total five different species of hydroxylated HBCD metabolites have been found by LCQ and GC

- MS; monohydroxy and dihy droxy

- HBCD, monohydroxy

- and dihydroxy

- PBCDe and monohydroxy

- TBCDe (Brandsma et al., 2009).

91. The extent of oral absorption of HBCD in humans is largely unknown, ( ECHA 2008a ). Estimations suggest that the uptake of HBCD via this exposure route ranges from 5 0

- 100 % ( ECHA 2008a

, European Commission 2008)

, HBCD is presumably readily absorbed in the gastro

- intestinal tract, and is thereafter distributed to different tissues. In rodents, the highest HBCD concentrations are reached in adipose tissue and muscle fol lowed by the liver, while lung, kidney, blood and brain have much lower concentrations Chengelis et al., 2001; van der Ven et al. 2006 ; Yu and Atallah, 1980).



### 2.4 Hazard assessment for endpoints of concern

92. The hazard potential of HBCD has been assessed i n several reports ( (European Commission 2008 , ECHA 2008b, US EPA 2008 and EBFRIP 2009). In the EU, HBCD has been identified as a substance of very high concern based on its persistency, bioaccumulation and toxicity. In US an initial screening assessment of HBCD (US EPA 2008) concluded a high concern for aquatic organisms from environmental releases, based on the bioaccumulation potential of HBCD, the high acute toxicity for aquatic plants and chronic toxicity for aquatic invertebrates, and also the potentia l for exposure and presence in remote regions. A comparison of measured concentrations in the tissues and organs of species of prey (fish) with the predicted no

- effect concentration ( PNEC) for secondary poisoning of 5 mg HBCD/ kg food for predators (mamma ls and birds) reveals that locally the concentrations in fish exceeds the PNEC. In the vicinity of point sources such as the river Skerne in UK and the river Scheldt basin in Belgium HBCD concentrations higher than 5 mg/kg wwt have been measured in eel and brown trout. The highest measured concentration in fish is 9.4 mg/kg wwt (eel in river Skerne. Also in marine mammals concentrations higher than the PNEC has been measured, the highest being 6.4 mg/kg wwt whole body weight in harbour porpoise from the UK. The potential risk of HBCD to wild life is further supported by the EBFRIP report (2009). The authors concluded from a body/tissue residue based risk assessment that freshwater fish and marine mammals from European regions with elevated back ground levels or near local point sources may exhibit some adverse effects from HBCD exposure. Levels in birds from European regions with elevated back ground levels or near local point sources were concluded to lie near the threshold levels for adverse effects. The up per third of the monitoring data used in the assessment exceeds the specific

- toxicity residue

- based PNEC for fresh water fish and for mammals, and the upper limit of the monitoring data for birds also enters this range.



### 2.4.1 Ecotoxicity to aquatic organis ms
93. Ecotoxicity testing of HBCD in aqueous media is complicated by its very low water solubility and high adsorption potential ( EBFRIP . 2009; TemaNord 2009). HBCD has a low acu te toxic ity to aquatic organisms owing in part to its limited solubility in aqueo us media (Graves and Svigert, 1997; Walsh et al., 19 8 7, CEPA 2007 and ACCBFRIP 2001 for overview ) . Regarding the long

- term toxicity of HBCD it was concluded to be very toxic to aquatic organisms in the EU Risk Assessment ((European Commission 2008). This c onclusion was based on the long

term ecotoxicity test with Daphnia magna (28d

- NOEC 3.1 μg/l; Drottar and Krueger, 1998) and on the growth inhibition test with Skeletonema costatum (72h

- EC50 52 μ g/L; Desjardins et al., 2005). In both tests calculated NOEC

a nd EC 50 values were below the water solubility of the technical mixture of HBCD (66 μ g/L).

94. HBCDD causes adverse effects to sediment organisms at exposure level relevant for the environment based on the effects in the long

- term test with Lumbriculus variega tus (Oetken et al., 2001).

95. Recent fish

- feeding studies have studied for example the interferense of HBCD with thyroid hormone regulated processes. An interference of HBCD with the HPT

- axis and liver biotransformation enzymes were reported in rainbow
trout exposed to individual HBCD diastereoisomers via food for 56 days followed by a depuration period of 112 days19

when fish were fed a reference diet (Palace et al., 2008). Lipid corrected concentrations of α

- , β

- , γ

- isomer in the food were 29.14 ± 1.95, 11.84 ± 4.62, and 22.84 ± 2.26 ng/g, respectively (means ±SEM). Liver detoxification processes (P450 CYP1A activity) were inhibited by all HBCD stereoisomers after 7 days of dosing, and also after 56 days of dosing but then only in α

- and β

- exposed fish. Thyroid follicle epithelial cell heights were significantly greater in γ

HBCD exposed fish at day 56 of the uptake phase and in fish from the α

- and γ

- HBCD exposed g roups at day 14 of the depuration phase. The link between HBCD induced disturbances in the HPT

- axis and the importance of such effects to smoltification in Atlantic salmon has also been examined (Lower and Moore 2007). Interactions of HBCD with this
proces s are relevant considering that the profound osmoregulatory adjustments required during the transition from freshwater to saltwater and that these are thought to be regulated by thyroid hormones. TH

- hormones also play a role in the olfactory imprinting res ponsible for salmon returning to t heir natal streams (Lema and Nev itt 2004). Fish were exposed to 11 ng/L of a technical HBCD mixture for 30 days during the peak smoltification period in freshwater, and was then transferred to clean seawater for 20 days (L ower and Moore 2007). At the end of the exposure, the fish were transferred to seawater for 20 days. Throughout the HBCD

- dosing and saltwater exposures, 5

- 8 fish were sampled every 7 days and gill and blood tissues were collected. In addition, electro

- olfa ctrograms were recorded in an additional 5 fish every 10 days using urine from salmon from the same stream (considered to be the cue for returning smolts) as an effector A finding was that the peak of thyroxine was shifted and occurred one week earlier in HBCD exposed fish than in controls. HBCD also attenuated the olfactory responses during the final 3 sampling periods of the freshwater HBCD exposure phase . In the Lower and Moore study, the exposure to HBCD was not observed to affect seawater adaptability

, although the peak of thyroxine was shifted and occurred one week earlier in HBCD exposed fish than in controls . A reduction in olfactory function , as evidenced by attenuated olfactory responses during early freshwater transition, was also observed. This latter effect is imporatn as it can affect successful homing, and thereby ultimately also reproductive capacity in adult salmon. In a third reported study assessing such effects in European flounders, no effects neither on the livers biotransformation cap acity or TH

- levels were reported, even though HBCD accumulated dose

- dependently (Kuiper et al. 2007). The fishes were in this instance exposed to HBCD in food (μg/g lipid) and sediment (μg/g total organic carbon (TOC)) in the following combinations; 0+0 (c ontrol); 0.3+0.08; 3+0.8; 30+8; 300+80; 3000+800; and 0+8,000 for 78 days.

96. R ecent studies with fish models suggest that HBCD may also induce oxidative stress and apoptosis. Deng et al. (2009) examined oxidative stress and the apoptosis pathway in four

- hou r post

- fertilization zebrafish ( Danio rerio ) embryos by exposing them to waterborne HBCD at concentrations (of HBCD) of 0, 0.05, 0.1, 0.5, and 1.0 mg/ L for 92 hours. HBCD reduced survival at the three middle doses equivalently, but was elevated at the hig hest dose (1mg/L). Hatching rate was only affected at the highest dose (1 mg/L) with a 10 % reduction from the controls. Malformation rates (including epiboly deformities, yolk sac and pericardial edema, tail and heart malformations, spinal curvatures and improper inflation of the swimbladder) increased dose dependently, and heart rate and body length both also decreased with exposure to HBCD

. The levels of reactive oxygen species (ROS) also increased dose dependently in fish exposed to HBCD concentrations above 0.05 mg/L.

. With regard to apoptosis, HBCD elevated expression of the pro

- apoptotic genes p53, Bax, Puma, Apaf

- 1, and caspase

- 9 and caspase

- 3 of which the response of the two latter was verified at the enzyme level. The anti

- apoptotic genes Mdm2 and Bcl

- 2 were both significantly down regulated at the highest HBCD exposure concentration. The overall results demonstrate that waterborne HBCD may produce oxidative stress in zebrafish embryos. The potential of HBCD to induc e oxidative stress in zebra fish has also been demonstrated by Hu et al. ( 2009 ) in zebrafish embryos. Here, the o xidative stress

, assessed by lipid membrane damage (ef fects at 0.5, 2.5 and 10 mg /L) in zebrafish embryos was also ac companied by delays in hatching (≤0.5 mg HBCD/ ml), dose

- independent changes in SOD enzyme activity (higher at 0.1, lower at 2.5 and 10 mg/ L ) and an elevation of heat shock proteins (Hsp70) activity (≥0.1 mg/L) , likely indicating increase d protein repair activity . Moreover, a study with Chinese rare minnows ( Gobiocyprinus rarus ) Zhang et al. (2008) observed a consistent increase in oxidative stress and cellular macromolecules in brain (ROS, carbonylation, TBARS) and erythrocytes (DNA) by water borne HBCD in the100

- 500 μg/l range (42 days). Protective enzymatic

- ( superoxide dismutase) and non

- enzymatic antioxidant glutathione were compromised even at concentrations of 10 and 1 ug /L respectively. A shorter 28 day exposure resulted in somewhat higher effect concentrations. However, since most test concentrations in these studies are above the water solubility of HBCD the relevance of these findings is uncertain.

97. Recently, Smolarz and Berger ( 2009 ) demonstrated the ability of HBCD to induce nuclear and nucleol ar abnormalities and to increase cell death (via apoptosis and necrosis) in gills of benthic clams ( Macoma balthica ) exposed to nominal HBCD concentrations of 0, 100 and 250 μg/l over a period of 50 days. Main peaks for nuclear abnormalities and malfunctio n of ribosomal genes were observed after 10 and 20

- 30 days, respectively (GLM analyses and Spearman correlation, p<0.05). According to the authors, the induction of micronuclei and other nuclear abnormalities reflect the genotoxic potential of HBCD to mari ne invertebrates. The complexity of HBCD induced effects in aquatic organisms has also been highlighted by Kling and Förlin 2009 who observed that the expression of 38 different proteins, many of which were related to protein metabolis m , cellular defence a nd cytoskeleton dynamics was altered in zebrafish ( Danio rerio ) liver cells exposed to 0.05, 0.5, 5 or 187 μM of a technical HBCD mixture f or 24 and 72 hours.20



### 2.4.2 Toxicity in soil organisms and plants

98. As reported in TemaNord 2009, long

- term toxicity of HBCD to earthworms has been assessed by Aufderheide et al. (2003), who measured survival and reproduction in Eisenia fetida (clittelate adults) following 56 day exposure to a technical HBCD mixture. HBCD was mixed dry into artific i al soil media at concentr ations of 78.5 to 500 mg/ kg dry soil weight. In this study, the NOEC for survival and reproduction were determined to 4190 and 128 mg HBCD/kg dry soi

, l respectively. The NOEC for reproduction was later recalculated to 59 mg/kg dry soil weight because the soil that was used contained a higher amoun t of organic matter than standard soil (TemaNord 2009).

99. Assessments of the toxicity of HBCD in the terrestrial ecosystems have also been conducted with plants ( Porch et al., 2002 ). This study, yielding a NOEC of > 5000 mg HBCD/ kg dry soil for the test species corn ( Zea mays ), cucumber ( Cucumis sativa ), onion ( Allium cepa ), ryegrass ( Lollium perenne ), soybean ( Glycine max) and tomato ( Lycopersicon esculentum ), was determined using a technical HBCD mixture in a see dling emergence test. For effects on soil micro

- organisms, the only conducted study reports a NOEC of ≥ 750 mg HBCD/ kg dw using nitrate production as an endpoint for assessment ( Förster et al., 2007 ).


### 2.4.3 Toxicity in birds

100. Marteinson et al (2009) studi ed reproductive effects of HBCD in American kestrels at environmentally relevant concentrations. The kestrels were exposed daily to 800 ng/g ww of a technical HBCD mixture in safflower oil from three weeks prior to pairing until two days before hatching. α

- HBCD dominated in the eggs, and was found at a concentration (163 ng/g ww) similar to concentrations found in wild birds. Only minor differences were noted in courtship behaviors both in male and females, and eggs were laid earlier after pairing. More hat chlings were noted in the treated kestrels, but they were smaller in weight and had a slower growth rate than controls. The avian reproductive toxicity of HBCD was also examined using air

- hold Japanese quails ( Coturnix coturnix japonica ) that were fed diet s containing 0, 125, 250, 500 or 1000 ppm of HBCD (a mixture of isomers : α

- , 27%; β

- , 30%; γ

- ,43%) for 6 weeks (M OEJ 2009)

. In contrast to the observed situation in American kestrels (Marteinsson et al., 2009), HBCD caused a reduction in hatchability at a ll concentrations examined. Statistically significant reduction in egg shell thickness was also observed at concentrations above 125 ppm. Decreases in egg weights and egg production rate and increase in cracked eggs were observed at 500 and 1000 ppm of HBC D. Adult mortality increased at 1000 ppm. Additional test was conducted with concentration of 0, 5, 15, 45 or 125 ppm of HBCD to confirm no

- observed

- effect concentration (NOEC) on reproductive performance. Survival of chicks hatched from eggs of HBCD fed h ens was significantly reduced at 15 ppm (2.1 mg/kg b.w./day) and more. A tendency to reduced hatchability with increasing concentration of HBCD was also observed at 15 ppm and more. NOEC for reproductive performance of quails was considered to be 5 ppm (0. 7 mg/kg b.w./day) of HBCD.

101. Chicken embryonic hepatocytes were exposed to 0.01 to 30 μM α

- HBCD or the technical mixture of HBCD (TM

- HBCD) to investigate potential effects on chicken xenobiotic

- sensing orphan nuclear receptor ( CXR), the thyroid hormone pathway, and lipid regulation using real time PCR analysis (Crump et al. 2008). The CXR receptor is related to the two mammalian xenobiotic

- activated receptors, CAR/PXR, and is involved in CYP regulation as well as phase II activity or UGT1A1. Phase I metabolizing enzymes CYP2H1 and C YP3A37 were upregulated. This result, as well the previous study by Germer et al. (2006 ), show that CYP3 induction is one of the most sensitive endpoints reported in higher animals exposed to HBCD. The avian ortholog (UGT1A9) to mammalian UGT1A1, was upreg ulated in a dose dependent manner by alpha HBCD but not by TM

- HBCD. UGT1A1 encodes for the major thyroid hormone conjugation enzyme and its specific upregulation by the predominating HBCD congener in biota provides a mechanism by which T4 may be depleted in exposed organisms (i.e. through more rapid conjugation and excretion). The gene encoding transthyretin (TTR) was downregulated by TM

- and α

- HBCD at concentrations of >1μM. Transthyretin is the primary binding protein for binding thyroid hormones in circulation of mammals and its downregulation could potentially further contribute to lower levels of T4 in blood.



### 2.4.4 Toxicity in terre strial mammals

102. Chicken embryonic hepatocytes were exposed to 0.01 to 30 μM α

- HBCD or the technical mixture of HBCD (TM

- HBCD) to investigate potential effects on chicken xenobiotic

- sensing orphan nuclear receptor ( CXR), the thyroid hormone pathway, and lipid regulation using real time PCR analysis (Crump et al. 2008). The CXR receptor is related to the two mammalian xenobiotic

- activated receptors, CAR/PXR, and is involved in CYP regulation as well as phase II activity or UGT1A1. Phase I metabolizing enzy mes CYP2H1 and CYP3A37 were upregulated. This result, as well the previous study by Germer et al. (2006 ), show that CYP3 induction is one of the most sensitive endpoints reported in higher animals exposed to HBCD. The avian ortholog (UGT1A9) to mammalian U GT1A1, was upregulated in a dose dependent manner by alpha HBCD but not by TM

- HBCD. UGT1A1 encodes for the major thyroid hormone conjugation enzyme and its specific upregulation by the predominating HBCD congener in biota provides a mechanism by which T4 may be depleted in exposed organisms (i.e. through more rapid conjugation and excretion). The gene encoding transthyretin (TTR) was downregulated by TM

- and α

- HBCD at concentrations of >1μM. Transthyretin is the primary binding protein for binding thyroid hormones in circulation of mammals and its downregulation could potentially further contribute to lower levels of T4 in blood.21

In mammals, as in aquat ic vertebrates HBCD primarily targets biotransformation processes in the liver and also affects the HPT

- axis (see e.g. TemaNord 2008; European Commission 2008

, ECHA 2008b ). Induction of oxidative stress and interference with apoptotic programmes and hormon e signaling are likely to be the initial toxic effects of HBCD exposure (e.g. Zhang et al., 2008; Reistad et al., 2006; Dingemans et al., 2009; Fery et al., 2009; Yamada

- Okabe, 2005; Hamers et al., 2006; Deng et al., 2009;Kling and Förlin, 2009; Hu et al.

. 2009). In rats daily oral HBCD exposure of 3 to 100 mg/ kg bw affects key metabolic pathways including metabolism of lipid, triacylglycerol, androstenedioene, testosterone, estrogen and cholesterol, as well as Phase I and II biotransformation (Canton et al., 2008; van der Ven et al., 2006). The fact that HBCD acts as an antagonist of important hormone receptors such as the androgen, thyroid hormone and progesterone receptors (Yamada

- Okabe, 2005; Hamers et al., 2006) has highlighted HBCD as a likely endocr ine disruptor of both the hypothalamus

- pituitary

- thyroid hormone axis and of sex

- steroid regulated processes in mammals. Of the two, the thyroid hormone effects of HBCD have received most attention and a series of studies have been undertake n. The results from in vivo subchronic tests with rats range from no observed effects to increase in thyroid

- and overall body weight, decreased serum T4 and increased serum TSH (Chengelis, 2001; van der Ven et al., 2006; Ema et al., 2008; van der Ven et al., 2009; see E U CHL for overview). M aternally transferred HBCD can a ffect viability and body weight in second

- generation rat s (Ema et al., 2008)

. 103. E ffects have been observed in both sexes but have also been limited to females only. Though the results may appear inconsist ent, there is now considerable consensus that HBCD, like other BFRs can interfere with the HPT

- axis ( Swedish Chemicals Agency 2008 , E uropean Commission 2008, TemaNord 2008). The mechanism for the thyroid effects is not clear, but a mode of action has been proposed where changes in hepatic metabolism of TH precedes changes in circulating TH levels, the pituitary, increased TSH levels and activation of thyroid with hypothyroidism and with secondary effects on lipoprotein synthesis and cholesterol

- and fatty a cid homeostasis as possible outcomes (van der Ven et al., 2006; Swedish Chemicals Agency 2009 ; Canton et al., 2008).

104. Besides their role as major regulator of body metabolism (Norris, 2001 ), thyroid hormones are required for normal development of the nervo us system, as are retinoids (Forrest et al., 2002; Maden, 2007), and disturbances in these systems may therefore result in long term neurotoxic effects in off

- spring. For HBCD, a neurotoxic potential has previously been demonstrated both i n vitro and in vi vo with rodent models (Reistad et al., 2006; Mariussen and Fonnum , 2003; Dingemans et al. , 2009; Eriksson et al., 2006; Lilienthal et al., 2009). In vitro studies have shown that HBCD may induce nerve cells and possibly also interfere with neuronal signali ng events (Reistad et al., 2006; Mariussen and Fonnum, 2003; Dingemans et al., 2009). Furthermore, in the in vivo study of Eriksson et al. (2006) neonatal direct exposure of pups to a single oral dose of HBCD (0.9 mg/ kg or 13.5 mg/ kg bw on postnatal day 10), induced alterations in spontaneous behavior with initial hypo

- reactivity, followed by impaired habituation in adult mice. This study also reported effects on spatial learning and memory as assessed in a Morris water maze test with exposed mice. In con trast, in the ir two

- generation study with rats where exposure of pups occurred indirectly via human breast milk, Ema et al., ( 2008 ), only observed transient changes in the performance of F1 males in a water

- filled T

maze test at an exposure level of 1500 p pm and higher. According to Ema et al. (2008), the discrepancy in their results could be explained by differences in exposure regime and/ or by differences in species sensitivity.

105. The in vivo neurotoxic potential of HBCD has been studied by Lilienthal et al

. 2009). In a one generation reproduction feeding study, they showed that HBCD

- induced disturbances in hearing function were paralleled by changes in dopamine dependent behavior. The behavioral effect mainly observed in female off

- spring and observed as a dose dependent reduction in latency to movement suggested HBCD to influence changes in dopaminergic activity in the striatum (Lilienthal et al., 2009). Loss of hearing function was also observed and was attributed to a cochlear effect of HBCD that result ed in increased thresholds and moderate prolongations of latencies in the lower frequency range from 0.5 to 2 kHz and after clicks. Both observed effects were dose

- dependent with lower bounds of bench mark doses (BMDL) between ≤1 and 10 mg/ kg bw. Saegusa et al . (2009) on the other hand detected weak hypothyroidism with increases in thyroid weight, thyroid follicular cell hypertrophy and serum TSH concentrations as well as a decrease in serum T3 levels in rat off

- spring exposed to 10 000 ppm HBCD in a soy

- f ree diet from gestation day 10 to day 20 after delivery. The TH changes were accompanied by a reduced density of CNPase

- positive oligodentrocytes, which is indicative of impaired oligodendroglial development. Increased thyroid weights and decreased serum T 3 concentrations were also observed in the adult stage from 1000 ppm. Though the above data suggest that TH

- signaling is linked to effects on the nerv ous system in rodents, by changes in behavior and cognition may also be impacted a decrease in apolar reti noids as observed in female rat liver s following HBCD exposure (van der Ven et al., 2006, van der Ven et al., 2009). T he interferences of HBCD with sex

- steroid hormones and their receptors should not be neglected as these hormones also exert non

- genomic ef fects on brain functions such as learning and memory, fine motor control, pain perception and mood (Boulaware and Mermelstein, 2005; Chakraborti et al., 2007; Meaney et al., 1983; Schantz and Widholm, 2001).

106. There are several studies on reproductive effec ts of HBCD. Saegusa et al (2009) performed a 1

- generation developmental toxicity study in rat, with maternal exposure via the diet to 0, 100, 1000 or 10,000 ppm HBCDD from gestation day 10 until weaning of the offspring. Thyroid effects were observed both in dams (thyroid weight increase and follicular cell hypertrophy at 10,000 ppm) and offspring (thyroid weight increase, decreased serum T3 and increased serum TSH at 1,000 and 10,000 ppm). The thyroid effects together with the impaired oligodendroglial22

dev elopment in the brain cortex (statistically significant at the high dose (

- 24%) supported by a dose

- dependent trend in the mid (

- 12%) and low (

- 8%) dose groups) and the decreased female body weight (9% in the high dose group) could indicate developmental h ypothyroidism. The LOAEL of this study is 1,000 ppm (81

- 213 mg/kg/day), and the NOAEL 100 ppm (8

- 21 mg/kg/day). The long continuous exposure study of van der Ven et al. (2009) suggest that male reproductive organs are particularly sensitive to HBCD exposur e i.e. a decreased testicular weight was observed at a BMDL of 52 μg/ g bw in F1 males. A weight reduction in other male organs prostate, the adrenals, heart and brain as well as in F1 males total weight was also observed. The observed weight loss makes it impossible to say whether any of these effects on organs weights are specific or secondary to the general body weight loss. In females the cytochrome P450 19 enzyme activity, based on group averages, showed a correlation to the internal concentration of γ

- HBCD (linear correlation coefficient of 0.90). The cytochrome P450 19 enzyme converts androgens to estrogens (Norris 2006), and is essential for differentiation and development of gonads and brains of higher vertebrates, maintenance of reproductive tissue s, and sexual behavior (Conley and Hinshelwood, 2001; Simpson et al., 2002). In females the time to vaginal opening was also delayed, but only at the top dose (BMDL 82.2 μg/ g bw at a BMR of 10 %).

107. Like the study of van der Ven et al. (2009), Ema et al. (2008) document reproductive and developmental effects (decreased pup viability, fewer primordial follicles) , and also changes in organ weights (e.g. liver and thyroid), and thyroid hormone levels

. Several effects were trans

- generational and affected both F0 parents and F1 and F2 parents and offspring. From the point of view of reproductive toxicology the general decrease in viability in F2 pups on post natal days 4 and 21 at 1500 and 15 000 ppm and the decrease in primordial follicles 1500 and 15 000 ppm H BCD exposure in F1 females were the most severe effects. A reduced number of primordial follicles suggest that reproductive potential of the female may be reduced, and is generally regarded as sensitive biomarkers for adverse reproductive effects (Parker e t al., 2006). It should be noted however that the highest dose of 15,000 ppm used by Ema et al. (2008) may be considered to be very high ( that is, 15 00 m g/kg bw /day ). However, dosing was in this study done by mixing HBCD particles into an appropriate amoun t of powdered basal diet for each dietary concentration. The absorption kinetics of HBCD likely depends on both the particle size and amount of particles administered, and is expected to be lower than for dissolved HBCD. The actual tissue doses from this s tudy are therefore presumably lower than the original dose would suggest, as may also be assumed from the findings of simil ar studies such as that of Chen g e lis et al. (2001) who only observed reversible effects at doses up to 1 000 m g/kg bw /day in their 90

- day oral exposure study.

108. In conclusion, both older and recent available literature suggest that HBCD can induce effects in mammals document and that both chronic and subchronic, high and low dose exposure to HBCD may have wide ranging and potentially seve re effects, particularly to the neuroendocrine system and to offspring during early phases of development.



### 2.4.5 Human toxicity

109. The EU risk assessment of HBCD completed in 2008 provides the most comprehensive assessment of toxic effects and risks of HBCD exposure to human health and welfare (E uropean Commission 2008). T his assessment concludes that HBCD may cause reproductive toxicity and long term toxicity whereas there is no concern for acute toxicity, irritation, sensitization, mutagenicity and carcino genicity. It moreover states that HBCD pose no risk to consumers or to workers when standard industrial hygiene measures are applied (current EU practice). These conclusions are founded on an extensive list of toxicity studies and on a comprehensive select ion of exposure and risk assessments that considers not only workers and consumers, but also indirect exposure of humans via the environment (E uropean Commission 2008). T he EU risk assessment documents that currently in the general population, human tissues show HBCD concentrations that are much below those reported to induce adverse effects in other mammals.

110. It is important to note that the e nvironmental background levels of HBCD have increased over the last decades (Law et al., 2008; Law et al., 2006 d ), an d HBCD is found in most human tissues, including serum and blood of pregnant women as well as in mothers’ milk (e.g. E uropean Commission 2008 ; TemaNord 2008; ECHA 2008b ). In some instances the reported levels in human milk have been quite high (Eljarrat et al., 2009). As demonstrated by its presence in cord serum and mothers’ milk , HBCD is transferred from mothers to their children (Covaci et al., 2008; E uropean Commission 2008). Young children may additionally ingest more HBCD via their environment than ad ults do (Abdallah et al. 2008). Prenatal exposure to HBCD may lead to subtle behavioral changes in rodents, particularly motor activity and cognition are affected (Eriksson et al., 2006). This has not been confirmed in humans (Roze et al. 2009). In the EU, there is a proposal to classify HBCD for reproductive toxicity. The proposal is under discussion (CLP; Repr. 2, the substance is suspected of damaging fertility and the unborn child (H361fd), the substance may cause harm to breast

- fed children (Lact. Effe cts H362) (Swedish Chemicals Agency 2009).

111. E arly phases of human development are tightly controlled by hormones and intracellular signaling processes such as apoptosis, of which the latter is necessary for normal embryonic

- and tissue differentiation ( Oppe nheim, 1991; Davies 2003; Barres et al., 1992). Thus, the developmental

- and neurotoxic potential of HBCD observed in animal studies give cause for concern, particularly for unborn babies and young children.23



## 3 Synthesis of information

112. HBCD is persistent in the environment and has a strong potential to bioaccumulate and biomagnify in the food chains. HBCD is widespread in the global environment and biota; elevated levels are found in top predators and in endangered species in the Arctic. The total amount of HBCD being released is increasing in all regions investigated. The increasing standing masses in construction materials are potentially long

- term sources of HBCD leaching or volatizing to the environment, as well as representing larger releases when dem olished or renovated in the future. Releases during recycling of construction materials and electronic appliances can be of importance and is likely to increase in the future.

113. HBCD is considered very toxic to aquatic organisms . In avian species, r ecent st udies report effects such as reduced eggshell thickness, growt h and survival. In mammals, HBCD ha s a potential to interfere with the hypothalamic

- pituitary

- thyroid (HPT) axis and cause reproductive and developmental effects. Many effects were trans

generat ional and affected both parents and offspring.

114. HBCD is transferred from mothers to their children. Significant l evels of HBCD in human milk and exposure through food has been reported near local sources

. In humans the main risk of HBCD exposure are possibl e neuroendocrine and developmental disturbances from exposure during the early developmental phases of the child. In the EU, there is a proposal to classify HBCD for reproductive toxicity that is under discussion. Additionally to the findings in the in viv o animal studies , there are a large number of recent in vitro studies that document how HBCD upon adsorption may act on, and possibly interfere with biological processes such as cell homeostasis, development, metabolism, intracellular signaling and neuroen docrine processes. Such studies add to the understanding that HBCD has various effects to human health and the environment, and should also be regarded when considering the toxicity of HBCD.

Table 2 POP characteristics of HBCD

Criterion Meets the

criteri on (Yes/No) Remark

Persistence Yes The measured data available from dated sediment cores indicate slow degradation rates of HBCDD. HBCD is found to be widespread in the global environment, with high levels in top predators in the Arctic

. Also the tempora lly increasing concentrations found in biota support the picture of HBCDD as a persistent substance.

The half

- life of HBCD in water is more than 60 days.

Bioaccumulation Yes Found in elevated concentrations in top predators

. Log K ow is greater than 5 (lo g K ow estimated to 5.62). BMFs > 1 for HBCD in aquatic ecosystems ( Tomy et al 2004, 2009, Sørmo et al. 2006) An overall BCF of 18,100 is chosen as represe ntative from two fish studies (Drottar and Krueger 2000 ).

Potential for Long

- Range Environmental Tr ansport Yes Found in elevated concentrations in the Arctic.

M odelling data show an estimated atmospheric half

- life of two to three days.

Adverse effects Yes HBCD exert reproductive toxicity in mammals and birds with a NOEC/NOAEL in the order of 1 mg/kg/da y. The data include:

 Decreased pup survival and fewer primordial follicles in rats at 100 mg/kg/day, NOAEL 10 mg/kg/day (Ema et al 2008).

 Decreased pup weight, as well as decreased testis and prostrate weights, impaired hearing, and decreased female bone mineral density in rat offspring at 30

- 100 mg/kg/day (van der Ven et al 2009, Lillienthal et al 2009).

 Thyroid hormone imbalance and impaired oligodendroglial development in the24 brain cortex of rat offspring at 1000 ppm (81

- 213 mg/kg/day), NOAEL 8

- 21 mg/ kg/day (Saegusa et al 2009).

 Behavioural effects in mice exposed to 13.5 mg/kg/day at day 10, NOAEL 0.9 mg/kg/day (Eriksson et al 2006).

 Bird egg/chick survival was decreased in quails exposed via the feed to 15 ppm HBCD (2.1 mg/kg/day), NOEC 5 ppm (0.7 mg/kg/day) (Ministry of the Environment, Japan, 2009 ).

 Minor differences in courtship behaviour, earlier egg

- laying, and a slower growth rate were observed in American kestrels exposed daily to 800 ng/g HBCD (Marteinson et al 2009).

 High toxicity in aquati c species with a 72h EC50 of 52 μg/l for Skeletonema
costatum and a no observed effects concentration (NOEC) of 3.1 μg/l for Daphnia magna

.



## 4 Concluding statement

115. HBCD is a synthetic substance with no known natural occurrence and that continues to be used in many countries

. HBCD is persistent in soil and bioaccumulate s and biomagnif ies in fish and mammals. Given the inherent properties of HBCD, the widespread occurrence of HBCD in biota, including in remote areas, together with environmental concentra tions that may affect ce rtain higher trophic level biota such as birds and mammals

, it is concluded that HBCD is likely, as a result of long

- range environmental transport, to cause significant adverse effects on human health or the envi r on ment, such that g lobal action is warranted.25



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