## Technical Review of Hexabromocyclododecan e (HBCD)

CAS Registry Numbers 3194 - 55 - 6 and 25637 - 99 - 4

Office of Environmental Information Environmental Analysis Division Analytical Support Branch January 25 , 201 6Purpose of this Hazard Assessment for Hexabromocyclododecane

Before a chemical can be proposed for addition to or deletion from section 313 of the Emergency Planning and Community Right

- To

- Know Act (EPCRA) of 1986, EPA performs a hazard assessment on the chemical. The statutory chemical listing/delisting criteria of EPCRA section 313(d)(2) are generally based on human health and environmental hazards. The analysis focuses on identifying and characterizing a chemical’s potential for toxicity to humans and the environment.
This t echnical report provides an overall assessment and review of the toxicity of hexabromocyclododecane, which includes a review and consideration of both human health and environmental hazards. It includes an assessment of the severity of the effect(s) and th e dose or concentration levels at which the effect(s) were observed. This report contains several chapters including chemistry and environmental fate, human health hazards, and ecological toxicity. An “Executive Summary” is included to provide the user wit h a brief review of the key findings presented in the f

ollowing chapters of the report.3 TABLE OF CONTENTS 1. EXECUTIVE SUMMARY 4 2. CHEMISTRY AND ENVIRONMENTAL FATE 7 2.1. SUMMARY 7 2.2. SYNONYMS/IDENTIFIERS 11 2.3. PHYSICAL AND CHEMICAL PROPERTIES 11 2.4. PRODUCTION METHODS 15 2.5. USE PROPERTIES 17 2.6. ENVIRONMENTAL FATE 17 2.7. REFERENCES 33 ATTACHMENT A 40 ATTACHMENT B 43 3. HUMAN HEALTH HAZARD ASSESSMENT 44 3.1 SUMMARY 44 3.2. KINETICS AND METABOLISM 47 3.3. EFFECTS OF ACUTE EXPOSURE 55 3.4 EFFECTS OF REPEATED EXPOSURE 61 3.5. DEVELOPMENTAL AND REPRODUCTIVE TOXICITY 68 3.6. GENOTOXICITY 85 3.7. OTHER CHRONIC EFFECTS 87 3.8. POTENTIAL MODES OF ACTION 88 3.9. CONCLUSION 88 3.10. REFERENCES 90 4. ECOLOGICAL HAZARD ASSESSMENT 96 4.1 SUMMARY 96 4.2. ECOLOGICAL TOXICITY 98 4.3. BIOACCUMULATION POTENTIAL 134 4.4 CONCLUSIONS 144 4.5 REFERENCES 145 ATTACHMENT A 155 ATTACHMENT B 1584 1. EXECUTIVE SUMMARY Hexabromocyclododecane (HBCD) is a brominated flame retardant consisting of a 12

- membered carbon ring with 6 bromine atoms attached at the 1, 2, 5, 6, 9, and 10 positions (ChemIDPlus, 2011). HBCD has 16 possible stereoisomers. α

- , β

- , and γ

- HBCD are the pr edominant diastereomers found in technical HBCD. HBCD is widely used as an additive to impart flame

- retardant properties to plastics and textiles (Covaci et al., 2006; ECB, 2008). The widespread production and use of HBCD as a brominated flame retardant ha s resulted in its release to the environment, wastewater, surface water, air and soil (ECB, 2008; U.S. EPA, 2010b). In the environment, HBCD is expected to be immobile in soil and to bind strongly to sediments and suspended solids in water. It is expecte d to volatilize slowly from moist soils and water surfaces (Great Lakes Chemical Corp., 1990 as cited in HSDB, 2011; Lyman et al., 1990; Stenzel and Nixon, 1997). HBCD is considered to be persistent in the environment (ECB, 2008) and appears to be slowly d egraded through abiotic reactions ( IUCLID, 2005 ; Davis et al., 2006; ECB, 2008) . Bioconcentration factor (BCF) values for HBCD in fathead minnows and rainbow trout indicate that it has a very high potential to bioconcentrate in aquatic organisms (ECB, 2008; Veith et al., 1979). HBCD is absorbed via the gastrointestinal tract and metabolized in rodents. Once absorbed, HBCD is distributed to a number of tissues, including fatty tissue, muscle, and the liver. Elimination of HBCD is predominantly via feces (as the parent compound), but it is also eliminated in urine (as secondary metabolites). HBCD has been detected in human milk and
serum.5

HBCD was not found to be highly toxic in acute oral, inhalation, and dermal tests in rodents. In short

- term and subch ronic studies, HBCD demonstrated effects on the thyroid and liver (van der Ven et al., 2006; Chengelis, 1997 as cited by U.S. EPA, 2001; Chengelis, 2001; 2002). Adequate s tudies evaluating the carcinogenic potential of HBCD in humans or animals were not a vailable so no determination of carcinogenicity could be made. HBCD, however, does not appear to be genotoxic. The developmental and reproductive toxicity of HBCD have bee n investigated in several rodent studies , and there is evidence of developmental an d reproductive effects at moderately low to low doses. Developmental effects on the liver, thyroid, and bone were observed in several studies (van der Ven et al., 2009; Saegusa et al., 2009; Ema et al., 2008). Neurodevelopmental effects on behavior and learning have also been observed (Saegusa et al., 2009; Eriksson et al., 2009). Ema et al. (2008) reported a decrease in the number of primordial follicles in first generation female rats, but no other significant reproductive toxicity was reported in the literature (van der Ven et al. 2009; Saegusa et al. 2009). The available evidence indicates that HBCD has the potential to cause developmental and reproductive toxicity at moderately low to low doses. Based on the available data, The Environmental Protection Agency ( EPA ) concludes that there is sufficient evidence to support a concern for moderately high to high toxicity of HBCD. HBCD presents a high concern for adverse effects on the environment. This conclusion is based on demonstrated high potential for bioaccumulation and biomagnification, high acute and chronic aquatic toxicity, and high toxicity to terrestrial wildlife (Drottar and Krueger, 1998; Du et al., 2012b; Fernie et al., 2011; Haukås et al., 2010a; Law et al., 2006; Marteinson et al., 2011,6

2012; Shaw et al., 2012; MOEJ, 2009 as cited in UNEP, 2010; Walsh et al., 1987; Zhang et al., 2014b). Assessment of HBCD aquatic toxicity is complicated by the very low water solubility of the chemical and differences in the solubility of the three main HBCD isomers , which makes testing difficult and interpretation uncertain
. In some acute and chronic aquatic toxicity studies conducted using me thods, test species, and endpoints recommended by the EPA , no effects were reported at or near the limit of water solubility. However, based on the potential of HBCD to bioaccumulate, biomagnify, and persist in the environment, water solubility is not cons idered a limiting factor for hazard determination for aquatic species.7 2. CHEMISTRY AND ENVIRONMENTAL FATE

2.1. SUMMARY Hexabromocyclododecane (HBCD) is a brominated flame retardant consisting of a 12

- membered carbon ring with 6 bromine atoms attached. HBCD is widely used as an additive to impart flame

- retardant properties to plastics and textiles (Covaci et al., 2006; ECB, 2008). HBCD is typically marketed as either a non

- specific mixture of HBCD isome rs (CASRN 25637

- 99

- 4) or as HBCD with bromine atoms at the 1, 2, 5, 6, 9, and 10 positions (CASRN 3194

- 55

- 6). There are 16 possible HBCD stereoisomers. Technical HBCD predominantly consists of three diastereomers, α

- , ß

- and γ

- HBCD with (

- ) and (+)

- enanti omers, typically in >95% concentration (Heeb et al., 2007). Interconversion among HBCD diastereomers has been observed by biological, thermal, and photolytic mechanisms (Peled et al., 1995; Law et al., 2006; Harrad et al., 2009a). Results from studies dete rmining physical properties of HBCD, such as water solubility and melting point, demonstrate the unique properties of the individual HBCD diastereomers. HBCD is produced through the bromination of trans,trans,cis

- 1,5,9

- cyclododecatriene (Becher, 2005; Da gani et al., 2005; ECB, 2008; Heeb et al., 2005). A range of technical products with various isomer compositions are formed depending on the purity of the starting material and the industrial process. According to industrial reports provided to the EPA und er the Inventory Update Rule (U.S. EPA, 2010a, 2011a), HBCD had a production volume of greater than 1 million to 10 million pounds per year during reporting years 1986 and 1990 and greater than8

10 million to 50 million pounds during reporting years 1994, 1 998, 2002, and 2006. HBCD is not currently listed in the Toxics Release Inventory (TRI) (U.S. EPA, 2009); therefore, companies are not required to report releases of this substance. The widespread production and use of HBCD as a brominated flame retardant has resulted in its release to the environment (ECB, 2008). U.S. EPA (2010b) reported that 3,100 kg/year of HBCD were released in Europe during 2008, with 50% released to wastewater, 29% to surface water, and 21% to air. Landfill disposal of HBCD

- containin g articles , such as scrap building materials or old electronics , is expected to result in some release of this substance to soil (ECB, 2008; U.S. EPA, 2010b). HBCD is a white, odorless solid with low water solubility (8.6 μg/L at 25 °C; reported for a technical

- grade HBCD sample) (Stenzel and Markley, 1997 as cited in IUCLID, 2005 ). In the environment, HBCD is expected to be immobile in soil and to bind strongly to sediments and suspended solids in water based on a K oc value of 7.6×10 4 , determined from a log K ow of 5.625 (MacGregor and Nixon, 1997 as cited in IUCLID, 2005 ; U.S. EPA, 2011b). HBCD is expected to volatilize slowly from moist soils and water surfaces based on its Henry’s Law constant of 4.6×10 - 5 atm
- m 3 /mol (Great Lakes Chemical Corp., 1990 as cited in HSDB, 2011; Lyman et al., 1990; Stenzel and Nixon, 1997). Estimated volatilization half

- lives are 34.8 hours from a model river and 24.7 days from a mod el lake, assuming no adsorption to suspended solids or sediment (U.S. EPA, 2011b). Volatilization is expected to be attenuated by adsorption. Therefore, the rate of HBCD volatilization from water will decrease with increasing amount of suspended particulat es. The half

- life for the volatilization of HBCD from a model pond has been estimated to be months to years if adsorption is considered (U.S. EPA, 2005; HSDB, 2011). In the atmosphere, HBCD is expected to exist primarily in the particulate phase based on its vapor pressure of 4.7×10 - 7 mm Hg at 21 °C (Bidleman, 1988; Stenzel and Nixon, 1997;9

Covaci et al., 2006). HBCD in the particulate phase is expected to be removed from the atmosphere through wet or dry deposition (HSDB, 2011); however, this substance a lso appears to undergo long

- range atmospheric transport based on its detection in Arctic air and biota samples (Covaci et al., 2006; Ueno et al., 2006; U.S. EPA, 2010b). Bioconcentration factor (BCF) values of 18,100 for HBCD in fathead minnows and 8,974

– 1 3,085 in rainbow trout indicate that this substance has a very high potential to bioconcentrate in aquatic organisms (Veith et al., 1979; ECB, 2008). Differing HBCD diastereomer ratios have been identified in samples collected in different environmental co mpartments, as reported in numerous monitoring studies. Typically, γ

- HBCD is the predominant diastereomer found in sediments and air, while α

HBCD is the predominant diastereomer found in biota (Harrad et al., 2009b; de Wit et al., 2010; Marvin et al., 201 1; NICNAS, 2012). HBCD is considered to be persistent in the environment because it has been detected in sediments several years after its release at those locations had ended (ECB, 2008). HBCD appears to be slowly degraded through abiotic reductive dehal ogenation, with observed degradation proceeding more rapidly under anaerobic conditions than under aerobic conditions; however, the mechanisms for this are unclear ( IUCLID, 2005 ; Davis et al., 2006; ECB, 2008; NICNAS, 2012) . Limited biodegradation data were located for this substance; however, available information indicates that this is a slow process ( IUCLID, 2005 ; ECB, 2008)
. No biodegradation was observed (0% of the theoretical oxygen demand) over 28 days in a screening

- level test for ready biodegrad ability ( Schaefer and Haberlein 1996 as cited in ECB, 2008; IUCLID, 2005 ). Observed abiotic and viable degradation of α

- , ß

- , and γ

- HBCD during simulation tests were approximately 33 and 61%, respectively, in anaerobic freshwater sediment, 15 and 44%, resp ectively, in aerobic freshwater sediment, and 6 and 10%, respectively, in10

aerobic soil after 112

– 113 days ( IUCLID, 2005 ; Davis et al., 2006; ECB, 2008). These studies suggest that HBCD may degrade slowly in the environment, although faster in sediment than in soil, faster under anaerobic conditions than aerobic conditions, and faster with microbial action than without microbial action. Other abiotic degradation processes, such as hydrolysis and direct photolysis, are not expected to be important for HBCD (E CB, 2008; HSDB, 2011). Although HBCD is expected to exist mostly in the particulate phase in the atmosphere, a small percentage may exist in the vapor phase as well, based on its vapor pressure of 4.7×10 - 7 mm Hg at 21 °C (Bidleman, 1988; Stenzel and Nixon, 1997). A half

- life of 26 hours has been estimated for the reaction of HBCD in the vapor phase with photochemically produced hydroxyl radicals in the atmosphere (U.S. EPA, 2011b). The chemistry and environmental fate of HBCD are discussed in this chapter . Synonym and physical

- chemical property information was obtained during a search of the on

- line source ChemIDPlus Advanced, handbook references, the SRC PHYSPROP database, and the SRC Environmental Fate Data Base (EFDB). Manufacturing, use, and environmen tal fate data were obtained through a search of the literature compiled in EFDB, the Hazardous Substances Data Bank (HSDB) record for HBCD (available on

- line), the European Chemicals Bureau Risk Assessment for Hexabromocyclododecane (ECB, 2008), the U.S. E PA Hexabromocyclo

dodecane Action Plan (U.S. EPA, 2010b), and estimations performed using EPISuite TM v4.10 software (U.S. EPA, 2011b). Primary sources were consulted when available; however, data that originated with individual manufacturers and industry c onsortia were typically only available from secondary sources.11

2.2. SYNONYMS/IDENTIFIERS (ACS, 2011; ChemIDPlus, 2011) Cyclododecane, hexabromo

1,2,5,6,9,10

- Hexabromocyclododecane Cyclododecane, 1,2,5,6,9,10

- hexabromo

Bromkal 73

- 6D FR 1206 Pyroguard SR 104M 88A CAS Registry Numbers associated with HBCD (ChemIDPlus, 2011; ECB, 2008): CAS Name 3194

- 55

- 6 HBCD (mixture of diastereomers) 25637

- 99

- 4 HBCD (mixture of diastereomers) 134237

- 50

- 6 α - HBCD 134237

- 51

- 7 β

- HBCD 134237

- 52

- 8 γ - HBCD 2.3. PHYSICAL AND CHEMICAL PROPERTIES Table 2

- 1 provides selected chemical and physical properties for HBCD. For many endpoints, such as water solubility and melting point, multiple value ranges are reported in the scientific literature. The differences in rep orted values are due to differences among technical mixtures of HBCD, which contain varying ratios of HBCD diastereomers. In addition to data reported for unspecified mixtures, the unique property values for α

- , β

- , and γ

- HBCD diastereomers were reported, when available. For example, there is a >10

- fold difference in water solubility between α

- and γ

- HBCD. It has been hypothesized that the measurable differences in properties are a result of differences in the three

- dimensional physical structures of HBCD. The physical structures, shown in Figure 2

- 1, were calculated using a semi

- empirical method with molecular modeling using the global minimum energy conformer, 1 H nuclear magnetic resonance (NMR) and 13 C NMR data (Arsenault et al., 2007).12

In the absence of experimental data, EPISuite TM v.4.10 was used to estimate values for physical

- chemical property and environmental fate endpoints (Attachment A). Experimental values for log K ow (5.625), vapor pressure (4.70×10 - 7 mm Hg), and water solubility (8.6 μg/L), were used in the EPISuite TM v4.10 calculations. An average melting point (190 °C) was selected from the available experimental data from isomeric mixtures of HBCD (ranging from 170 to 209 °C) and was also used in the EPISuite TM v.4.10 calculations (U.S. EP A, 2011b). Table 2 - 1. Physical and chemical properties of HBCD Property Value a Reference Molecular formula C 12 H 18 Br 6 Molecular weight 641.695 Lide, 2008 Melting point 172

- 184 °C (crude product)

201

- 205 °C (highest

- melting version) 179

- 181 °C (α

- HBCD) 170

- 172 °C (β

- HBCD) 207

- 209 °C (γ

- HBCD) 190±20 °C (average) Smith et al., 2005

Boiling point Not applicable; thermally induced

isomerization occurs at 160 to 220 °C and decomposition occurs at >220 °C Peled et al., 1995; Heeb et al., 2008 Vapor pressure 4.70×10 - 7 mm Hg (6.27×10 - 5 Pa) at 21 °C

(technical product) Stenzel and Nixon, 1997 Water solubility 8.6 μg/L at 25 °C ( 14 C

- labeled technical

product) Great Lakes Chemical Corp., 1990 as cited in HSDB, 2011 3.4±2.3 μg/L at 25 °C (based on γ

- HBCD in technical product reported with coefficient of variation) Stenzel and Markley, 1997 as cited in IUCLID, 2005 , ECB, 2008 and Marvin et al., 2011 48.8±1.9 μg/L (α

- HBCD) 14.7±0.5 μg/L (β

- HBCD) 2.1±0.2 μg/L (γ

- HBCD) ƩHBCD 65.6 μg/L MacGregor and Nixon, 2004 as cited in IUCLID, 2005 , ECB, 2008 and Marvin et al., 2011 Log K ow 5.625 at 25±0.05°C (technical product) Hayward et al., 2006; MacGregor and Nixon, 1997 as cited in IUCLID, 2005 , ECB, 2008 and Marvin et al., 201113

Table 2 - 1. Physical and chemical properties of HBCD Property Value a Reference

5.07±0.09 (α

- HBCD) 5.12±0.09 (β

- HBCD) 5.47±0.10 (γ

- HBCD) (estimated) MacGregor and Nixon, 1997 as cited in IUCLID, 2005 , ECB, 2008, and Marvin et al., 2011 K oc 7.6×10 4 (estimated) U.S. EPA, 2011b (EPISuite TM v4.10) Henry’s Law constant 4.6×10 - 5 atm
- m 3 /mol (calculated)

Derived from a vapor pressure of 4.70×10 - 7 mm Hg and water solubility of 8.6 μg/L VP: Stenzel and Nixon, 1997 WS: Great Lakes Chemical Corp., 1990 as cited in HSDB, 2011 Density 2.24 g/cm 3

2.38 g/cm 3 (technical product) Great Lakes Chemical Corporation, 1994 and Albemarle Corporation, 1994 as cited in ECB, 2008 Flash point Not applicable

a Values are experimental unless otherwise indicated.

Source: Arsenault et al., 2007. Chemosphere by PERGAMON. Reproduced with permission of PERGAMON in the format reuse in a government report via Copyright Clearance Center.

Figure 2 - 1. Structures and dipole moments for α - , β - and γ - HBCD
In the absence of experimental data, EPISuite TM v.4.10 was used to estimate values for physical

- chemical property and environmental fate endpoints (Attachment A). Experimental values for log K ow (5.625), vapor pressure (4.70×10 - 7 mm Hg), and water solubility (8.6 μg/L),14

were used in the EPISuite TM v4.1 0 calculations. An average melting point (190 °C) was selected from the available experimental data from isomeric mixtures of HBCD (ranging from 170 to 209 °C) and was also used in the EPISuite TM v.4.10 calculations (U.S. EPA, 2011b). HBCD will decompose at temperatures >220 °C (Peled et al., 1995). HBCD commercial product exhibited 50% weight loss at 269 °C and 90% weight loss at 274 °C in a thermogravimetric analysis study (Albemarle Corporation, 2000 as cited in IUCLID, 2005 ). Thermally induced isomeri zation of HBCD at temperatures of 160

– 200 °C resulted in a thermal equilibrium ratio of 78:13:9 α

- , β

- , and γ

- HBCD after 2

– 3 hours (Peled et al., 1995). At 190 °C, the thermal rearrangement of HBCD diastereomers occurred, but decomposition was reported to be minimal (<5%) (Peled et al., 1995). Zhao et al. (2010) determined that  - HBCD is theoretically the most thermodynamically stable diastereomer using density functional theory (DFT) and time

- dependent DFT. The 

- and γ

- HBCD diastereomers undergo isomerization to α

HBCD by thermal, metabolic, and photolytic isomerization mechanisms (Peled et al., 1995; Law et al., 2006; Harrad et al., 2009a). It has been suggested that under high temperatures or in the presence of light, degradation of HBCD to pent abromocyclododecene (PBCD) occurs by a radical mechanism as shown in Figure 2

- 2 (Kroschwith, 1993 as cited in ECB, 2008; Barontini et al., 2001).

Figure 2 - 2. Dehalogenation of HBCD by a radical mechanism15

2.4. PRODUCTION M ETHODS HBCD is produced through the bromination of trans,trans,cis

- 1,5,9

- cyclododecatriene (Becher, 2005; Dagani et al., 2005; Heeb et al., 2005; ECB, 2008). Bromine atoms are added to the double

- bonded carbons at the 1, 2, 5, 6, 9, and 10 positions on the ring.

Industrial production is a batch process carried out in the presence of a solvent at temperatures ranging from 20 to 70 °C in a closed system (ECB, 2008). The resulting suspension is filtered and the product is dried. Use of nonpolar solvents leads to decreased yields and lower quality product due to free

- radical substitution reactions (Dagani et al., 2005). Purities and yields are improved by using mixed solvent systems such as benzene

- tert
- butyl alcohol or perc hloroethylene

- ethanol or by adding the reagents separately to a polar solvent such as ethanol (Dagani et al., 2005). Feeding bromine and cyclododecatriene (3:1 molar ratio) separately to 96% ethanol in a stirred reactor resulted in a yield of 92% (Dagani e t al., 2005). A 1

– 5% excess of bromine was used and a temperature of 20

– 30 °C was maintained over 2.5

– 3 hours. The mother liquor was recovered and recycled into subsequent reactions (Dagani et al., 2005). The addition of bromine to the cis

- and trans
- dou ble bonds of cyclododecatriene results in stereoisomerism in the HBCD product (Becher, 2005; Covaci et al., 2006; ECB, 2008; Heeb et al., 2005). Although 16 stereoisomers are theoretically possible, technical mixtures of HBCD consist primarily of the three racemic diastereomers: (  ) γ

- HBCD (75

– 89%), (  ) α

- HBCD (10

– 13%), and (  ) β

- HBCD (<0.5

– 12%) (Becher, 2005). Although β

- HBCD can constitute over 10%16

of some technical mixtures, it is inconsistently found in environmental compartments in monitoring studies; the reason for this is not apparent in the available literature (NICNAS, 2012). Stereochemical structures for these three diastereomers are shown in Figure 2

- 3.

Figure 2 - 3. Stereochemical structures for α - , β - , and γ - HBCD

Small amounts of δ

- HBCD (0.5%) and ε

- HBCD (0.3%) have also been reported and may be due to impurities in the starting material (Heeb et al., 2005). δ

- and ε

- HBCD are tentatively regarded as meso forms, as chirality has not been detected for these st ereoisomers. Although δ

- and ε

- HBCD are rarely detected in environmental compartments, one monitoring study reported detection of δ

- HBCD in 13 fish samples from English lakes, but not the associated water or sediment samples (Harrad et al., 2009b). Consis tent with this finding, δ

- HBCD has been suggested to be a biotransformation product in fish. Trout S9 fraction assays detected the production of δ

- HBCD in samples containing only α

- , β

- , or γ - HBCD (Abdallah et al., 2014).17

2.5. USE PROPERTIES HBCD is used as an additive to impart flame retardant properties to plastics and textiles. The major use of HBCD (approximately 90%) is in the manufacture of expandable polystyrene (EPS) and extruded polystyrene (XPS) for rigid insulation panels that are used in b uilding construction (Covaci et al., 2006; ECB, 2008). XPS undergoes thermal isomerization during processing and production, and therefore, XPS has a different diastereomer pattern than EPS. While EPS contains α

- , β

- and γ

- HBCD in a typical ratio of 6:13:8 0, XPS contains α

- , β

- and γ - HBCD in a ratio of 43:36:21 (Heeb et al., 2010). Smaller amounts of HBCD (approximately 2%) are used in the manufacture of high impact polystyrene (HIPS), which is used in housings for electrical components and devices, in dist ribution boxes for electrical lines, and in electronic appliances (Covaci et al., 2006; ECB, 2008). HBCD is also added to polymer dispersions that are back

- coated onto the surface of textiles used for furniture upholstery, draperies, wall coverings, bed ma ttress ticking, roller blinds, and automobile interiors (Covaci et al., 2006; ECB, 2008).

2.6. ENVIRONMENTAL FATE 2.6.1. Environmental Transport. Based on a classification scheme that relates K oc to soil adsorption and mobility (Swann et al., 1983), an estimated K oc value of 7.6×10 4 , calculated based on a log K ow of 5.625, indicates that HBCD will be immobile in soil and will bind strongly to soil, sediment, and suspended organic matter (MacGregor and Nixon, 1997 as cited in IUCLID, 2005 ; U.S. EPA, 2011b). A Henry’s Law constant of 4.6×10 - 5 atm
- m 3 /mol at 25 °C, calculated based on a vapor pressure of 4.70×10 - 7 mm Hg at 21 °C and a water solubility of 8.6 μg/L at 25 °C, indicates that HBCD may volatilize slowly from moist soil and water surfaces (Great18

Lakes Chemical Corp., 1990 as cited in HSDB, 2011; Lyman et al., 1990; Stenzel and Nixon, 1997). Using this Hen ry’s Law constant and the EPIWIN estimation software (U.S. EPA, 2011b), the volatilization half

- lives of HBCD are calculated to be 34.8 hours for a model river 1 and 24.7 days for a model lake 2 . However, adsorption to suspended solids and sediment will redu ce the rate of volatilization from water. The half

- life for the volatilization of HBCD from a model pond has been estimated to be months to years if adsorption is considered (U.S. EPA, 2005; HSDB, 2011). In an aqueous phase partitioning study, 47±4.7% HBCD (sum of diastereomers) detected from freshwater lakes was in the freely

- dissolved phase, defined as capable of passing through a 1 μm pore size glass fiber filter. The remainder, 53±5.4% HBCD, was associated with the particulate phase (Harrad et al., 2009 b). The measured vapor pressure suggests that HBCD will not volatilize from dry soil surfaces (Stenzel and Nixon, 1997; HSDB, 2011). In the atmosphere, HBCD is expected to exist primarily in the particulate phase, based on its vapor pressure (Bidleman, 1 988; Stenzel and Nixon, 1997; Covaci et al., 2006; Marvin et al., 2011). HBCD in the particulate phase is expected to be removed from the atmosphere through wet or dry deposition (HSDB, 2011); however, the widespread detection of HBCD in air samples and bi ota from remote locations such as the Arctic indicates that long

- range atmospheric transport can occur (Covaci et al., 2006; Ueno et al., 2006; U.S. EPA, 2010b; Marvin et al., 2011). Information regarding specific mechanisms of long

- range atmospheric trans port was not located. Bioconcentration in aquatic organisms is an important environmental process for HBCD. In flow

- through bioconcentration tests using 1.8 and 0.18 μg/L, BCF values of 8,974 and 13,085,

1 Parameters for the model river are 1 m deep, flowing 1 m/sec ond , wind velocity of 5 m/sec ond
. 2 Parameters for the model lake are 1 m deep, flowing 0.05 m/sec ond , wind velocity of 0.5 m/sec ond
.19

respectively, were determined for HBCD in rainbow t rout (Drottar and Kruger, 2000 as cited in ECB, 2008). Veith et al. (1979) measured a BCF of 18,100 for HBCD in fathead minnows. These BCF values indicate that HBCD has a very high potential to bioconcentrate in fish. Widespread detection of this substance in aquatic organisms is further evidence that HBCD bioconcentrates (Covaci et al., 2006; ECB, 2008; Marvin et al., 2011). Based on measurements of HBCD in invertebrates, fish, birds, and marine mammals, biomagnification of HBCD in the aquatic food web is evident, with the highest levels of HBCD measured in seals and porpoises (de Boer et al., 2 002; Covaci et al., 2006; ECB, 2008; U.S. EPA, 2010b; Letcher e t al., 2009; Shaw et al., 2012 ). Higher concentrations of HBCD were measured in eggs from wild peregri ne falcons feeding on birds from terrestrial food webs in southwestern Sweden than in the eggs of captive falcons feeding on domestic chickens, indicating that HBCD bioaccumulation in terrestrial food chains may also be important (Lindberg et al., 2004; U. S. EPA, 2010b). The Ecological Hazard Assessment portion of this document contains additional information on bioconcentration, bioaccumulation, and biomagnification of HBCD.

2.6.1.1. Diastereomers and isomerization. As reported in Section 2.4, γ

- HBCD is the dominant stereoisomer present in most commercial HBCD formulations (>75%) (Becher, 2005). Similar proportionality is observed with HBCD stereoisomers detected in abiotic environmental media (Covaci et al., 2006; M arvin et al., 2011). However, in living organisms, α

- HBCD is the predominant stereoisomer identified, accounting for 70

– 90% of all HBCD detected (Covaci et al., 2006; Harrad et al., 2009b; de Wit et al., 2010; Marvin et al., 2011). For example, t he majorit y of the HBCD detected in human milk is α

- HBCD. In milk samples from Birmingham, United Kingdom, 62

– 95% of the total HBCD was α

- HBCD (Abdallah and Harrad, 2011). In20

milk samples from Boston, Massachusetts, HBCD was detected in all 43 samples and α

- HBCD
con stituted 40

– 84% of the total HBCD (Carignan et al., 2012). The reasons for the difference in HBCD composition found in living organisms compared to that found in the abiotic environment have not been determined. Possible explanations include different rat es of bioconcentration for the α

- , β

- , and γ

- stereoisomers and diastereomer interconversion and metabolism within organisms (Covaci et al., 2006; Law et al., 2006; ECB, 2008; Marvin et al., 2011; NICNAS, 2012). In the Western Scheldt (Netherlands) estuari ne food web, γ

- HBCD concentrations decreased and α

- HBCD concentrations increased with an increase in trophic level (Brandsma et al., 2015). This finding suggests differential bioaccumulation/ biomagnification of diastereomers. A possible example of diaster eomer interconversion is detection of α

- HBCD in rainbow trout fed for 168 days on food containing only β

- or γ - HBCD in statistically significant amounts (Law et al., 2006). It is not yet clear to what extent the levels of specific HBCD diastereomers found in the environment and biota arise from interconversion processes. The rates and mechanisms of conversion between the HBCD diastereomers and the importance of these processes to environmental fate is an active area of current research.
2.6.2. Abiotic Deg radation. A summary of the environmental fate and persistence characteristics of HBCD is presented in Table 2

- 2. Table 2 - 2. Environmental Degradation of HBC D Property Value a Reference Air Photodegra

dation Photo

- induced isomerization of γ

- HBCD to α

- HBCD in indoor dust

with a measured decrease in HBCD concentration concurrent with an increase of PBCDs in indoor dust Harrad et al., 2009a21 Table 2 - 2. Environmental Degradation of HBC D Property Value a Reference

Indirect photolysis half

- life: 26 hours AOPWIN v1.92 (estimated) U.S. EPA, 2011b Water Hydrolysis Not expected due to lack of functional groups that hydrolyze under

environmental conditions and low water solubility (estimated) ECB, 2008 Sediment Aerobic conditions No biodegradation observed in 28

- day closed

- bottle test Organisation

for Economic Co

- operation and Development (OECD) Guideline 301D, EPA OTS 796.3200 Schaefer and Haberlein, 1996 as cited in ECB, 2008; IUCLID, 2005 Half

- life: 128, 92, and 72 days for α

- , γ - , and 

- HBCD, respectively (estimated), based on a 44% decrease in total initial radioactivity in viable freshwater sediment of 14 C

- labeled HBCD ( 4.67 mg/kg dry weight) after 112 days; method based on OECD 308 Davi s et al., 2006

Half

- life: >120 days b (estimated), based on a 15% decrease in total initial radioactivity in abiotic freshwater sediment of 14 C

- labeled HBCD (4.67 mg/kg dry weight) after 112 days; method based on OECD 308 Half

- life: 11 and 32 days (estimated) in viable sediment collected from Schuylkill River and Neshaminy creek, respectively, using nominal HBCD concentrations of 0.034

– 0.089 mg/kg; method based on OECD 308 Davis et al., 2005

Half

- life: 190 and 30 days (estimated) in abiotic sediment collected from Schuylkill River and Neshaminy creek, respectively, using nominal HBCD concentrations of 0.034

– 0.089 mg/kg; method based on OECD 308 Anaerobic conditions Half

- life: 92 days (estimated), based on a 61% decrease in total initial radioactivity in viable freshwater sediment of 14 C

- labeled HBCD (4.31 mg/kg dry weight) after 113 days; method based on OECD 308 Davis et al., 2006

Half

- life: >120 days b (estimated), based on a 33% decrease in total initial radioactivity in abiotic freshwater sediment of 14 C

- labeled HBCD (4.31 mg/kg dry weight) after 113 days; method based on OECD 308 Half

- life: 1.5 and 1.1 days (estimated) in viable sediment collected from Schuylkill River and Neshaminy creek, respectively, using nominal HBCD concentrations of 0.063

– 0.089 mg/kg; method based on OECD 308 Davis et al., 2005

Half

- life: 10 and 9.9 days (estimated) in abiotic sediment collected from Schuylkill River and Neshaminy creek, respectively, using nominal HBCD concentrations of 0.063

– 0.089 mg/kg; method based on OECD 308 Soil Aerobic conditions Half

- life: >120 days b (estimated), based on a 10% decrease in total initial radioactivity in viable soil of 14 C

- labeled HBCD after 113 days; method based on OECD 307 using HBCD at 3.04 mg/kg dry weight Davis et al., 200622 Table 2 - 2. Environmental Degradation of HBC D Property Value a Reference

Half

- life: >120 days b (estimated), based on a 6% decrease in total initial radioactivity in abiotic soil of 14 C

- labeled HBCD after 113 days; method based on OECD 307 using HBCD at 3.04 mg/kg dry weight Half

- life: 63 days (estimated) in viable soil amended with activated sludge using a nominal HBCD concentration of 0.025 mg/kg dry weight ; method based on OECD 307 Davis et al ., 2005

Half

- life: >120 days b (estimated) in abiotic soil using a nominal HBCD concentration of 0.025 mg/kg dry weight; method based on OECD 307 Anaerobic conditions Half

- life: 6.9 days (estimated) in viable soil amended with activated sludge using a nominal HBCD concentration of 0.025 mg/kg dry weight ; method based on OECD 307 Davis et al., 2005

Half

- life: 82 days (estimated) in abiotic soil using a nominal HBCD concentration of 0.025 mg/kg dry weight ; method based on OECD 307

a Values are experimental unless otherwise indicated. b A default value of > 120 days was used to represent minimal degradation. HBCD is not expected to undergo significant direct photolysis since it does not absorb radiation in the environmentally available region of the electromagnetic spectrum that has the potential to cause molecular degradation (HSDB, 2011) . Although HBCD is expected to exist primarily in the particulate phase in the atmosphere, a small percentage may also exist in the vapor phase based on its vapor pressure (Bidleman, 1988; Stenzel and Nixon, 1997; Covaci et al., 2006; Marvin et al., 2011). HBCD in the vapor phase will be degraded by reaction with photochemically produced hydroxy l radicals in the atmosphere. An estimated rate constant of 5.01×10 - 12 cm 3 /molecules

- second at 25 °C for this reaction corresponds to a half

- life of 26 hours, assuming an atmospheric hydroxyl radical concentration of 1.5×10 6 molecules/cm 3 and a 12

- hour day (U.S. EPA, 1993, 2011b). No significant debromination or isomerization of HBCD was found when technical HBCD incorporated into textiles was exposed to natural sunlight (Kajiwara et al., 2013).23

Photolytic isomerization of HBCD has been described in both indoor dust samples and in samples of HBCD standards dissolved in methanol using artificial light (Harrad et al., 2009a). After 1 week in the presence of light, indoor dust containing predominantly γ

- HBCD was found to decrease in γ

- HBCD and increase in α

- H BCD concentration. There was a measured decrease in HBCD concentration concurrent with an increase in PBCDs in the indoor dust exposed to artificial light. The three diastereomerically pure HBCD standards (α

- , 

- , and γ

- HBCD) that were dissolved in methano l also began to interconvert within 1 week, resulting in a decrease in γ - HBCD concentration and an increase in α

- HBCD concentration. Under simulated sunlight, HBCD may also degrade in the presence of Fe(III)

- oxalate, Fe(III)

- citrate and Fe(III)

- carboxylate complexes with H 2 O 2 through a hydroxyl radical mechanism and pseudo

- first

- order rate constants (Zhou et al., 2014). HBCD photodegradation increased with increasing concentrations of Fe(III) complexes in solution and varied based on pH of the individual
so lutions. HBCD is not expected to undergo hydrolysis in environmental waters on the basis of its low water solubility (ECB, 2008). Observed abiotic degradation of HBCD during simulation tests based on OECD 307 and 308 was approximately 33% in anaerobic f reshwater sediment, 15% in aerobic freshwater sediment, and 6% in aerobic soil after 112

– 113 days (Davis et al., 2006; ECB, 2008). The results from these studies correspond to estimated h alf

- lives >120 days in soil and sediment due to minimal degradation b eing observed. Initial concentrations of 14 C radiolabeled HBCD (α

- , β

- , and γ

14 C

- HBCD in a ratio of 7.74:7.84:81.5) were 3.0

– 4.7 mg/kg dry weight in the sediment and soil systems. Additional results evaluating degradation under viable conditions, in the presence of microbes, from this study are described in Section 2.6.3. HBCD degradation observed under abiotic conditions was attributed to abiotic reductive dehalogenation ( IUCLID,24

2005 ; Davis et al., 2006; ECB, 2008). Degradation proceeded through a stepwise process to form tetrabromocyclododecene, dibromocyclododecadiene (DBCD), and 1,5,9

- cyclodo deca triene (see Figure 2

- 4).

Figure 2 - 4. Stepwise reductive dehalogenation of HBCD Further degradation of 1,5,9

- cyclod odecatriene was not observed. In this study, HBCD degradation occurred faster in sediment than in soil and faster under anaerobic conditions compared to aerobic conditions (Davis et al., 2006; ECB, 2008). Previous OECD 308 and 307

- based simulation tests from the same authors presented results suggesting faster abiotic degradation, particularly in sediment under anaerobic conditions, but were performed at much lower HBCD concentrations and measured only γ

- HBCD (Davis et al., 2005; IUCLID, 2005 ; ECB, 2008; Marvin et al., 2011; NICNAS, 2012) . In this study, abiotic degradation half

- lives in freshwater sediments were 30

– 190 days under aerobic conditions and 9.9

– 10 days under anaerobic conditions
. Estimated half

- lives in abiotic soil were >120 days under aerob ic conditions and 82 days under anaerobic conditions . Additional results evaluating degradation under viable conditions, in the presence of microbes, from this study are described in Section 2.6.3. This study evaluated γ

- HBCD only and did not address inter conversion of HBCD isomers or α

- and 

- HBCD degradation. The i nitial concentrations of HBCD were 0.025

– 0.08925

mg/kg dry weight in the sediment and soil systems, nearly 100 times less than the HBCD concentrations used in the subsequent Davis et al. (2006) st udy. Higher concentrations of HBCD ( 3.0

– 4.7 mg/kg dry weight) in the Davis et al. (2006) study allowed for quantification of individual isomers, metabolite identification and mass balance evaluation (Davis et al., 2006; NICNAS, 2012). Additionally, the Dav is et al. (2005) study was considered to be of uncertain reliability for quantifying HBCD persistence because of concerns regarding potential contamination of sediment samples, an interfering peak corresponding to γ

- HBCD in the liquid chromatography/mass s pectrometry ( LC/MS) chromatograms, and poor extraction of HBCD leading to HBCD r ecoveries of 33

– 125% (ECB, 2008; NICNAS, 2012).
2.6.3. Biotic Degradation. A few studies on the biodegradation of HBCD were located. A closed bottle screening

- level test for ready biodegradability (OECD Guideline 301D, EPA OTS 796.3200 ) was performed using an initial HBCD concentration of 7.7 mg/L and an activated domestic sludge i noculum (Schaefer and Haberlein, 1996 as cited in ECB, 2008; IUCLID, 2005 ). No biodegradation was observed (0% of the theoretical oxygen demand) over the test period of 28 days under the stringent guideline conditions of this test. Degradation of HBCD dur ing simulation tests with viable microbes, based on OECD 307 and 308, was approximately 61 % in anaerobic freshwater sediment, 44% in aerobic freshwater sediment, and 10% in aerobic soil after 112

– 113 days (Davis et al., 2006; ECB, 2008). The results from t his study correspond to estimated HBCD half

- lives of 92 days in anaerobic freshwater sediment, 128, 92, and 72 days for α

- , γ - , and 

- HBCD, respectively in aerobic freshwater sediment, and >120 days in aerobic soil. An initial total 14 C

- HBCD concentration of 3.0

– 4.7 mg/kg dry weight in the sediment and soil systems was used , allowing for quantification26

of individual isomers, metabolite identification, and mass balance evaluation (Davis et al., 2006; NICNAS, 2012). Although very high spiking rates can be tox ic to microorganisms in biodegradation studies and lead to unrealistically long estimated half

- lives, the results of this study did not suggest toxicity to microorganisms. Tests with viable microbes demonstrated increased HBCD degradation compared to the b iologically inhibited control studies (described in Section 2.6.2). In combination, these studies suggest that HBCD will degrade slowly in the environment, although faster in sediment than in soil, faster under anaerobic conditions than aerobic conditions , faster with microbial action than without microbial action, and at different rates for individual HBCD diastereomers (slower for α - HBCD than for the γ

- and β

stereoisomers ). The same researchers previously conducted a water

- sediment simulation test for commercial HBCD based on OECD guideline 308 using nominal HBCD concentrations of 0.034

– 0.089 mg/kg dry weight (Davis et al., 2003b, 2005; IUCLID, 2005 ; ECB, 2008). Aerobic and anaerobic microcosms were pre

- incubated at 20 °C for 49 days and at 23 °C for 43

– 44 days, respectively. HBCD was then added to 14

– 37 g dry weight freshwater sediment samples in 250 m l serum bottles (water:sediment ratio of 1.6

– 2.9) and the microcosms were sealed and incubated in the dark at 20 °C for up to 119 days. For the aerobic microcosms, the headspace oxygen concentration was kept above 10

– 15%. This study evaluated only γ

- HBCD and did not address interconversion of HBCD isomers or α

- and 

- HBCD degradation. Disappearance half

lives of HBCD with sediment collected from Schuylkil l River and Neshaminy creek were 11 and 32 days in viable aerobic sediments, respectively (compared to 190 and 30 days in abiotic aerobic controls , respectively), and 1.5 and 1.1 days in viable anaerobic sediments, respectively ( compared to 10 and 9.9 days in abiotic anaerobic controls).27

Data from these tests suggest that anaerobic degradation is faster than aerobic degradation of HBCD in viable and abiotic sediments and that degradation is faster in viable conditions than abiotic conditions. While these findings are consistent with Davis et al. (2006), the actual degradation rates in this study are much faster. However, results from this study do not provide a reliable indication of HBCD persistence. A mass balance could not be established because only γ

- HBCD was used to quantify HBCD concentrations, 14 C

- radiolabelled HBCD was not used, and degradation products were not identified; therefore, apparent disappearance of HBCD in this study may not reflect biodegradation. In addition, there were concerns that contaminated sediment may have been used, H BCD extraction was incomplete (HBCD recovery varied from 33 to 125 %), and an interfering peak was observed in the LC/MS chromatograms corresponding to γ

- HBCD (ECB, 2008; NICNAS, 2012). Similarly, a soil simulation test was conducted based on OECD guideline 307 for commercial HBCD using 50 g dry weight sandy loam soil samples added to 250 m l serum bottles (Davis et al., 2003a, 2005; IUCLID, 2005 ; ECB, 2008 ). The moisture content was 20% by weight. Aerobic and anaerobic microcosms were pre

- incubated at 20 °C for 35 days and at 23 °C for 43 days, respectively. Activated sludge was added to the soil at 5 mg/g, and HBCD was added to the soil to achieve a nominal concentration of 0.025 mg/kg dry weight. The micr ocosms were then incubated in the dark at 20 °C for up to 120 days. The disappearance half

lives were 63 days in viable aerobic soil (compared to >120 days in abiotic aerobic controls) and 6.9 days in viable anaerobic soil (compared to 82 days in abiotic a naerobic controls). As in the sediment studies, HBCD degradation in soil occurred faster under anaerobic conditions compared to aerobic conditions, and faster in viable conditions than abiotic conditions . The disappearance half

- lives in soil were slower th an those in sediment.28

Biological processes were suggested to be responsible for the increased degradation of HBCD in this study using viable conditions, relative to abiotic conditions; however, degradation was not adequately demonstrated in soil because no degradation products were detected and only γ - HBCD was used to quantify HBCD concentrations, making it impossible to calculate a mass balance. HBCD r ecoveries on day 0 of the experiment were well below (0.011

– 0.018 mg/kg dry weight) the nominal test concentrations (0.025 mg/kg dry weight), suggesting rapid adsorption of HBCD to soil and poor extraction methods (ECB, 2008; NICNAS, 2012)
. In studies using 0.025

– 0.089 mg/kg HBCD (Davis et al., 2005), the estimated half

- life values were shorter than stud ies using 3.0

– 4.7 mg/kg HBCD (Davis et al., 2006) by approximately one order of magnitude for aerobic, viable sediment (11

– 32 days compared to72

– 128 days) and anaerobic viable sediment (1.1

– 1.5 days compared to 92 days). The viable aerobic soil half

- life u sing lower concentrations of HBCD (Davis et al., 2005) was less than half of the half

- life based on the higher HBCD concentration (63 days compared to >120 days) (Davis et al., 2006). Both Davis et al. (2005, 2006) studies suggest that HBCD degrades faster in sediment than in soil, faster under anaerobic conditions than aerobic conditions, and faster with microbial action than without microbial action. HBCD is poorly soluble, and it was suggested that at higher concentrations of HBCD, degradation is limited by mass transfer of HBCD into microbes. However, results from the Davis et al. (2005) study likely overestimate the rate of HBCD biodegradation, for the reasons noted above (primarily, failure to use 14 C

- radiolabelled HBCD, quantify isomers other than γ

- H BCD, identify degradation products, or establish a mass balance, but also procedural problems with contamination of sediment, incomplete H BCD extraction, and occurrence of an interfering peak in the LC/MS chromatograms corresponding to γ

- HBCD).29

Furthermor e, the rapid biodegradation rates from Davis et al. (2005) are not consistent with environmental observations. HBCD has been detected over large areas and in remote locations in environmental monitoring studies (U.S. EPA, 2010b). Dated sediment core sampl es indicate slow environmental degradation rates (Davis et al., 2005; ECB, 2008; Marvin et al., 2011; NICNAS, 2012). For example, HBCD was found at concentrations ranging from 112 to 70,085 μg/kg dry weight in sediment samples collected at locations near a production site in Aycliffe, United Kingdom 2 years after the facility was closed down (ECB, 2008). Monitoring data do not provide a complete, quantitative determination of persistence because HBCD emission sources, rates, and quantities are typically unk nown, and all environmental compartments are not considered. However, the monitoring data do provide evidence in support of environmental persistence. Rapid HBCD biodegradation has been demonstrated under laboratory conditions not representative of typica l environmental conditions. A study designed to elucidate HBCD degradation mechanisms and optimize biodegradation capability reported an HBCD degradation half

- life of only 0.66 days in a naerobic digested sewage sludge amended with yeast and starch at 37 °C . In this test, α - HBCD had lower susceptibility to degradation than β

- or γ - HBCD (Gerecke et al., 2006) . The authors noted that these results are specific to the anaerobic conditions established by the experiment, and that the degradation rate constants ar e expected to vary based on redox conditions of each specific anaerobic environment. Two pure cultures isolated from a continuous anaerobic reactor degraded HBCD to 2

- dodecene, TBCD, DBCD, and 1,5,9

- cyclododecatriene (Peng et al., 2015). HBCD biodegraded 9 0% after 8 days of treatment under optimized conditions of 30 °C, pH 7 and an initial HBCD concentration of 500 μg/L.30

Enzyme studies using the haloalkane dehalogenase LinB from Sphingobium indicum (Heeb et al., 2012, 2013) and the dehydrochlorinase LinA2 from S. indicum (Heeb et al., 2014) have also demonstrated degradation of HBCD. LinB converts HBCD to pentabromocyclo

dodecanols (PBCD

- OHs) and tetrabromocyclododecadiols (TBCD

- OHs) and LinA2 conver ts only (

- ) 

- HBCD to pentabromocyclododecenes (PBCDs) . Transformation rates for the conversion of (

- ) 

- HBCD by 10 mg/L LinA2 and 100 mg/L LinB are 24 and 2.3 L mol - 1 s - 1 , respectively. Both LinA2 and LinB demonstrate faster conversion of 

- HBCD compared to
α - HBCDs (Heeb et al., 2015). Although these studies are not indicative of environmental persistence of HBCD, investigation into these systems provides improved understanding of HBCD degradation mechanisms and potential for bioremediation.

2.6.4. Environm ental Distribution Assessment. U.S. EPA used the level III fugacity model within the EPISuite TM estimation software (U.S. EPA, 2011b) to estimate HBCD’s distribution in the environment if released at 7,900 kg/hour to water, 2,100 kg/hour to air, and 0 kg/h our to soil. This release scenario is based on 79% release of this substance to water and 21% release to air, reflecting HBCD release patterns observed in Europe (U.S. EPA, 2010b). Measured physical

- chemical properties (melting point, vapor pressure, water solubility, and octanol

- water partition coefficient) were used as model input parameters. The half

- life input value for air was based on the estimated photooxidation half

- life of 26 hours. O bserved viable and abiotic degradation of HBCD during the most re liable simulation tests (Davis et al., 2006) was approximately 33

– 61% in anaerobic freshwater sediment, 15

– 44% in aerobic freshwater sediment, and 6

– 10% in aerobic soil after 112

– 113 days ; therefore, a half

- life input value of 100 days (2,400 hours) was used for sediment, and values for water and soil were chosen to be31

arbitrarily high (240,000 hours) since available data indicate that degradation in these compartments is very slow ( Davis et al., 2006; ECB, 2008) . The advective flow rate was set to zero by selecting very large advection times (1×10 20 hours) for air, water, and sediment. The model results indicate that, for this release scenario, HBCD will partition primarily to soil, with lower amounts partitioning to sediment and water. The model partitioning results, half

- life inputs, and emissions inputs for each environmental compartment are shown in Table 2

- 3. See Attachment A for EPISuite TM model outputs. Table 2 - 3. Environmental distribution of H BCD based on release to air (21%) and water (79%)
Compartment Emissions (kg/hour) Half - life Values (hours) Mass Amount (percent) Air 2,100 26 0.188 Water 7,900 240,000 9.18 Soil 0 240,000 72.1 Sediment 0 2,400 18.6 Source: U.S. EPA (2011b). As mentioned in Section 2.6.1, the detection of HBCD in air and biota in remote locations indicates that this substance has the potential for long

- range atmospheric transport (Covaci et al., 2006; Ueno et al., 2006; U.S. EPA, 2010b; Marvin et al., 2011); howe ver, a discussion of possible mechanisms for this process was not located.

2.6.5. Toxic Transformation Products. Studies evaluating HBCD degradation products must take into account impurities formed during the synthesis of technical HBCD that are also tra nsformation products of HBCD. For example, some technical

- HBCD samples may contain PBCD

- OHs and isobutoxypentabromocyclododecanes (iBPBCDs) as synthetic byproducts (Heeb et al., 2010, 2012). However, PBCD

- OHs are also potential environmental transformation32

products and metabolites of biological HBCD transformations. Evaluation of the starting material is required to adequately identify the source of HBCD degradation products. During a study of 14 C radiolabeled HBCD degradation in a digester sludge experiment, HBCD transformation products from reductive dehalogenation included TBCD, DBCD, and 1,5,9

- cyclododecatriene (Davis et al., 2006; ECB, 2008). During a thermal degradation study using a heating rate of 10 °C/minute that reached temperatures of 50 0 °C in 100% nitrogen gas or air (21 mol % oxygen), hydrogen bromide was the major degradation product. Other high

molecular weight brominated and non

- brominated organic compounds, such as PBCDs, TBCD, naphthalene, and biphenyl were detected under experime ntal conditions using Fourier transform infrared spectroscopy (FTIR) and gas

- chromatographic (GC)/MS analysis (Barontini et al., 2001). D ebromination and hydroxylation are in vivo metabolic pathways of HBCD. In rats dosed orally with 30 or 100 mg/kg body weight/day for 30 days, debrominated structures (PBCDs, TBCD) and corresponding mono

- and di

- hydroxy structures (HBCD

- OH, PBCD

- OH, TBCD

OH, di

- OH

- HBCD, di

- OH

- PBCD, and di

- OH

- TBCD) were identified in adipose, liver, lung, and muscle tissues ( Brandsma et al ., 2009) . C57BL/6 mice dosed once with 3 mg/kg body weight
α - [ 14 C]HBCD orally were found to have detectable HBCD

- OH (in liver, brain, and adipose tissues), PBCD

- OH (in liver and adipose tissues), tri

- and tetrabromohexene glutathione, hexabromododecanedioi c acid, tribromononenoic acid, and methylmercapturate of TBCDs (in urine) and PBCD

- OH, di

- OH

- PBCDs, and methylmercapturate of tribromononeneoic acid (in feces) ( Hakk et al., 2012). In vitro metabolism studies using rat liver S9 fractions identified PBCD, TBCD, HBCD

- OH, di

- OH

- HBCD, PBCD

- OHs, and TBCD

- OHs (Abdallah et al., 2014). These33

metabolites were also identified after treatment with trout liver S9 fractions in this study. The estimated rate of metabolism was the fastest for 

- HBCD (half

- lives of 6.43 and 38.10 minutes) and slowest for α

- HBCD (half

- lives of 17.14 and 134.91 minutes) in both the rat and trout liver S9 fractions, respectively. Esslinger et al. (2011) used rat liver microsomes to evaluate stereospecificity of HBCD metabolism. Estimated m etabolic half

- lives reported for (˗)

- β, (+)

- β, (˗)

- γ, (+)

- α, (˗)

- α, and (+)

- γ HBCD were 6.32, 6.35, 11.6, 14.1, 24.4, and 32.3 minutes, respectively. TBCD and PBCD were not detected in this experiment, but mono

- and di

- hydroxylated α - and γ

- HBCD metabolite s were found. In general, γ

- HBCD is the dominant stereoisomer in commercial formulations and
α - HBCD is dominant in biota and higher trophic level animals (Covaci et al., 2006; Marvin et al., 2011; NICNAS, 2012). Although the reasons for the difference in H BCD composition found in the environment compared to that found in living organisms have not been determined, bioisomerisation of α

- HBCD from γ

- HBCD, preferential α

- HBCD uptake, and the higher water solubility of α

- HBCD have been suggested as possible sour ces of enrichment (Covaci et al., 2006; Law et al., 2006; ECB, 2008; Marvin et al., 2011; NICNAS, 2012).

2 .7. REFERENCES Abdallah, MA; Harrad, S. (2011) Tetrabromobisphenol

- A, hexabromocyclododecane and its degradation products in UK human milk: relationship to external exposure. Environ Int 37:443

– 448. Abdallah, MA; Uchea, C; Chipman, JK; et al. (2014) Enantioselective biotransformation of hexabromocyclododecane by in vitro rat and trout hepatic sub

- cellular fractions. Environ Sci Technol 48:2 732

– 2740. ACS (American Chemical Society). (2011) CAS presents "common chemistry." CAS registry number: 3194

- 55

- 6. American Chemical Society. Available online at http://www.c ommonchemistry.org/ChemicalDetail.aspx?ref=3194

- 55

- 6 (accessed June 2, 2011).34

Albemarle Corporation. (1994) Baton Rouge, LA [as cited in ECB, 2008]. Albemarle Corporation. (2000) Saytex HP

- 900 Flame Retardant. Technical Data Sheet. BC1033 (1 l/00) [as cited in IUCLID, 2005 ]. Arsenault, G; Chittim, B; McAlees, A; et al. (2007). Nuclear magnetic resonance spectral characterization and semi

- empirical ca lculations of conformations of α

- and γ

- 1,2,5,6,9,10

- hexabromocyclododecane. Chemosphere 67(9):1684

– 1694. Barontini, F; Cozzani, V; Petarca, L. (2001). Thermal stability and decomposition products of hexabromocyclododecane. Ind Eng Chem Res 40(15): 3270

– 3280. Becher, G. (2005) The stereochemistry of 1,2,5,6,9,10

- hexabromocyclododecane and its graphic representation. Chemosphere 58:989

– 991. Bidleman, TF. (1988) Atmospheric processes. Environ Sci Technol 22(4):361

– 367. Brandsma, SH; Van der Ven, LT; D e Boer, J; et al. (2009) Identification of hydroxylated metabolites of hexabromocyclododecane in wildlife and 28

- days exposed Wistar rats. Environ Sci Technol 43:6058

– 6063. Carignan, CC; Abdallah, MA; Wu, N; et al. (2012) Predictors of tetrabromobisphenol

- A (TBBP

- A) and hexabromocyclododecanes (HBCD) in milk from Boston mothers. Environ Sci Technol 46:12146

– 12153. ChemIDPlus. (2011) 1,2,5,6,9,10

- Hexabromocyclododecane. Bethesda, MD: National Library of Medicine. Available online at http://chem.sis.nlm.nih.gov/chemidplus/ (accessed May 20, 2015). Covaci, A; Gerecke, AC; Law, RJ; et al. (2006) Hexabromocyclododecanes (HBCDs) in the environment and humans: a review. Environ Sci Technol 40(12):3679

– 368 8. Dagani, MJ; Barda, HJ; Benya, TJ; et al. (2005) Bromine compounds. In: Ullmann's encyclopedia of industrial chemistry. Available online at http://onlinelibrary.wiley.com/d oi/10.1002/14356007.a04_405/pdf
. Davis, JW; Gonsior, SJ; Marty, GT. (2003a) Evaluation of aerobic and anaerobic transformation of hexabromocyclododecane in soil. Project Study ID 021082, 61 pp. DOW Chemical Company: Midland, MI, USA. Submitted under TS CA Section FYI; EPA Document No. 84040000010; FYI

- 1103

- 01472, pg 379 . Available online at http://java.epa.gov/oppt_chemical_search/proxy?filename=2003

- 11

- FYI

- 03

- 01472A_fyi_1103_01472a.pdf
. Davis, JW; Gonsior, SJ; Marty, GT. (2003b) Evaluation of aerobic and anaerobic transformation of hexabromocyclododecane in aquatic sediment systems. Proje ct Study ID 021081, 87 pp. DOW Chemical Company: Midland, MI, USA. Submitted under TSCA Section FYI; EPA Document No. 84040000010; FYI

- 1103

- 01472, pg 440 . Available online at35

http://java.epa.gov/oppt_chemical_search/proxy?filename=2003

- 11

- FYI

- 03

- 01472A_fyi_1103_01472a.pdf
. Davis, JW; Gonsior, SJ; Marty, GT; et al. (2005) The transformation of hexabromocyclododecane in aerobic and anaerobic so ils and aquatic sediments. Water Res 39:1075

– 1084. Davis, JW; Gonsior, SJ; Markham, DA; et al. (2006) Biodegradation and product identification of [ 14 C]hexabromocyclododecane in wastewater sludge and freshwater aquatic sediment. Environ Sci Technol 40:539 5

– 5401. Including supporting information document. de Boer, J; Allchin, C; Zegers, B; et al. (2002) HBCD and TBBP

- A in sewage sludge, sediments and iota, including interlaboratory study. RIVO Report No. C033/02, RIVO, Netherlands Institute for Fisheries Research, Wageningen, The Netherlands. Available onli ne at http://www.epa.gov/opptintr/tsca8e/pubs/8ehq/2002/oct02/8ehq_1002_15166b.pdf
. de Wit, CA; Herzke, D; Vorkamp, K. (2010) Brominated flame retardants in the Arctic environment

-- trends and new candidates. Sci Total Environ 408:2885

– 2918. Drottar, KR; Krueger, HO. (2000) Hexabromocyclododecane (HBCD): A flow

- through bioconcentration test with the rainbow trout (Oncorhynchus mykiss). Final report. 439A

- 111, p p 1

- 137. Wildlife International Ltd, Easton, Maryland, USA [as cited in ECB, 2008]. ECB (European Chemicals Bureau). (2008) Risk assessment hexabromocyclododecane. CAS

No.: 25637

- 99

- 4. EINECS

- No.: 247

- 148

- 4. Final report. European Union. Available online at http://echa.europa.eu/documents/10162/661bff17

- dc0a

- 4475

- 9758

- 40bdd6198f82
. Esslinger, S; Becker, R; Maul, R; et al. (2011) Hexabromocyclododecane enantiomers: microsomal degradation and patterns of hydroxylated metabolites. Environ Sci Technol 45:3938

– 3944. Great Lakes Chemical Corp. (1990) Water solubility of several industri al chemicals flame retardants and a herbicide vel

- 3510 laboratory report with test data and cover letter. Submitted under TSCA Section 8D. OTS0523252 [as cited in HSDB, 2011. Great Lakes Chemical Corporation. (1994) West Lafayette, IN [as cited in ECB, 20 08]. Gerecke, AC; Giger, W; Hartmann PC; et al. (2006) Anaerobic degradation of brominated flame retardants in sewage sludge. Chemosphere 64:311

– 317. Hakk, H; Szabo, DT; Huwe, J; et al. (2012) Novel and distinct metabolites identified following a single oral dose of alpha

- or gamma - hexabromocyclododecane in mice. Environ Sci Technol 46:13494

– 13503.36

Harrad, S; Abdallah, MA; Covaci, A. (2009a) Causes of variability in concentrations and diastereomer patterns of Hexabromocyclododecanes in indoor dust. Env ironment International 35:573

– 579. Harrad, S; Abdallah, MA; Rose, NL; et al. (2009b) Current

- Use Brominated Flame Retardants in Water, Sediment, and Fish from English Lakes. Environ Sci Technol 43:9077

– 9083. Hayward, SJ; Lei, YD; Wania, F. (2006) Compara tive evaluation of three high

- performance liquid chromatography

- based K ow estimation methods for highly hydrophobic organic compounds: polybrominated diphenyl ethers and hexabromocyclododecane. Environ Toxicol Chem 25(8):2018

– 2027. Heeb, NV; Schweizefr, WB; Kohler, M; et al. (2005) Structure elucidation of hexabromocyclododecanes

- a class of compounds with a complex stereochemistry. Chemosphere 61:65

– 73. Heeb, NV; Schweizer, WB; Mattrel, P; et al. (2007) Solid

- state conformations and absolute configuratio ns of (+) and (

– ) α - , β

- , and γ

- hexabromocyclododecanes (HBCDs). Chemosphere 68:940

– 950. Heeb, NV; Schweizer, WB; Mattrel, P; et al. (2008) Regio

- and stereoselective isomerization of hexabromocyclododecanes (HBCDs): Kinetics and mechanism of γ - to α - HBCD isomerization. Chemosphere 73:1201

– 1210. Heeb, NV; Heidi, G; Schweizer, BW; et al. (2010) Thermally

- induced transformation of hexabromocyclododecanes and isobutoxypentabromocyclododecanes in flame

- proofed polystyrene materials. Chemosphere 80 :701

– 70 8. Heeb, NV; Zindel, D; Geueke, B; et al. (2012) Biotransformation of Hexabromocyclododecanes (HBCDs) with LinB

-- an HCH

- converting bacterial enzyme. Environ Sci Technol 46:6566

– 6574. Heeb, NV; Wyss, SA; Geueke, B; et al. (2015) Stereochemistry of enzyma tic transformations of (+)beta

- and (

- )beta

- HBCD with LinA2

-- a HCH

- degrading bacterial enzyme of Sphingobium indicum B90A. Chemosphere 122:70

– 78. Heeb, NV; Wyss, SA; Geueke, B; et al. (2014) LinA2, a HCH

- converting bacterial enzyme that dehydrohalogenat es HBCDs. Chemosphere 107:194

– 202. Heeb, NV; Zindel, D; Graf, H; et al. (2013) Stereochemistry of LinB

- catalyzed biotransformation of delta

- HBCD to 1R,2R,5S,6R,9R,10S

- pentabromocyclododecanol. Chemosphere 90:1911

– 1919. HSDB (Hazardous Substance Data B ank). (2011) 1,2,5,6,9,10

- Hexabromocyclododecane. Hazardous Substances Data Bank. Part of the National Library of Medicine's Toxicology Data37

IUCLID. (2005) International Uniform ChemicaL Information Database (IUCLID) Hexabromocyclododecane dataset. Submit ted to U.S. EPA’s High Production Volume (HPV) Chemical Program by Albemarle Corporation. Available online at http://www.epa.gov/chemrtk/pubs/summaries/cyclodod/c13459rr.pdf (accessed June 17, 2015). Network (TOXNET 7 ). Bethesda, MD. Available online at http://toxnet.nlm.nih.gov/cgi

bin/sis/htmlgen?HSDB (accessed May 31, 2011). Kajiwara, N; Desborough, J; Harrad, S; et al. (2013) Photolysis of brominated flame retardants in textiles exposed to natural sunlight. Environ Sci Process Impacts 15:653

– 660. Kroschwith, JI. (1993) Kirk

- Othmer. Encyclopedia of chemical technology. Fourth edition. 10, pp 954

– 976, 998

– 1022. New York: John Wiley & Sons [as cited in ECB 2008]. Law, K; Palac, VP; Halldorson, T; et al. (2006) Dietary accumulation of Hexabromocyclododecane diastereoisomers in juvenile rainbow trout (Oncorhynchus mykiss) I: bioaccumulation parameters and evidence of bioisomerization. Environ Toxicol Chem 25:1757e1761. Letcher, R J; Gebbink, WA; Sonne, C; et al. (2009) Bioaccumulation and biotransformation of brominated and chlorinated contaminants and their metabolites in ringed seals (Pusa hispida) and polar bears (Ursus maritimus) from East Greenland. Environ Int 35:1118

– 1124. Lide, DR. (2008) CRC handbook of chemistry and physics. Boca Raton: Taylor and Francis Group, LLC, 3

– 276. Lindberg, P; Sellström, U; Häggberg, L; et al. (2004) Higher brominated diphenyl ethers and hexabromocyclododecane found in eggs of Peregrine falco ns ( Falco peregrines ) breeding in Sweden. Environ Sci Technol 38:93

– 96. Lyman, WJ; Reehl, WF; Rosenblatt, DH. (1990) Handbook of chemical properties estimation methods. Washington, DC: American Chemical Society, 15

- 15

– 15

- 17. MacGregor, JA; Nixon, WB. ( 1997) Hexabromocyclododecane (HBCD): determination of n

octanol/water partition coefficient with cover letter dated 06/27/1997. Chemical Manufacturers Association. Submitted under TSCA Section 8D. OTS0573665 [as cited in IUCLID, 2005 ]. MacGregor, JA; Nix on, WB. (2004) Determination of water solubility of Hexabromocyclododecane (HBCD) using a generator column method. Project No. 439C

- 138, Wildlife International Ltd, Easton, MD [as cited in IUCLID, 2005 ]. Marvin, CH; Tomy, GT; Armitage, JM; et al. (2011) Hexabromocyclododecane: current understanding of chemistry, environmental fate and toxicology and implications for global management. Environ Sci Technol 45:8613

– 8623. Including supporting information document.38

NICNAS (National Industrial Chemicals Noti fication and Assessment Scheme). (2012) Hexabromocyclododecane. Priority existing chemical assessment report. Volume 34. Commonwealth of Australia: Australia. National Industrial Chemicals Notification and Assessment Scheme. PEC34. Peled, M; Scharia, R; Sondack, D. (1995) Thermal rearrangement of hexabromocyclododecane (HBCD). In: Desmurs, JR; Gérard, B; Goldstein, MJ, eds. Industrial Chemistry Library. Advances in organobromine chemistry II. Peng, X; Huang, X; Jing, F; et al. (2015) Study of novel p ure culture HBCD

- 1, effectively degrading Hexabromocyclododecane, isolated from an anaerobic reactor. Bioresour Technol 185:218

– 224. Schaefer, EC; Haberlein, D. (1996) Hexabromocyclododecane (HBCD): Closed bottle test. 439E

- 102, Wildlife International Ltd, Easton, MD, USA [as cited in ECB, 2008]. Shaw, SD; Berger, ML; Weijs, L; et al. (2012). Tissue

- specific accumulation of polybrominated diphenyl ethers (PBDEs) including Deca

- BDE and hexabromocyclododecanes (HBCDs) in harbor seals from the northwest A tlantic. Environ Int 44:1

– 6. Smith, K; Liu, C; El

- Hiti, GA; et al. (2005) An extensive study of bromination of cis, trans,trans
- 1,5,9

- cyclododecatriene: product structures and conformations. Org Biomol Chem 3:1880

– 1892. Stenzel, JI; Nixon, WB. (1997) He xabromocyclododecane (HBCD): determination of the vapor pressure using a spinning rotor gauge with cover letter dated 08/15/1997. Chemical Manufacturers Association. Submitted under TSCA Section 8D. OTS0573702. Stenzel, J; Markley, B. (1997) Final Report. Hexabromocyclododecane (HBCD): Determination of the Water Solubility. Wildlife International Project Number: 439C

- 105. Wildlife International, Ltd. Easton, MD [as cited in IUCLID, 2005 ]. Swann, RL; Laskowski, DA; McCall, PJ; et al. (1983) A rapid method for the estimation of the environmental parameters octanol/water partition coefficient, soil sorption constant, water to air ratio, and water solubility. Residue Rev 85:17

– 28. Ueno, D; Al aee, M; Marvin, C; et al. (2006) Distribution and transportability of hexabromocyclododecane (HBCD) in the Asis

- Pacific region using skipjack tuna as a bioindicator. Environ Pollut 144(1): 238

– 247. U.S. EPA (U.S. Environmental Protection Agency). (1993) Determination of rates of reaction in the gas

- phase in the troposphere. 5. Rate of indirect photoreaction: Evaluation of the atmospheric oxidation computer program of Syracuse Research Corporation for estimating the second

- order rate constant for the react ion of an organic chemical with hydroxyl radicals. Washington, DC: U.S. Environmental Protection Agency. EPA744R93001.39

U.S. EPA (U.S. Environmental Protection Agency). (2005) EXAMS results for CAS 003194

- 55

- 6. Download EXAMS v2.98.04.06. U.S. Environmental Protection Agency. Available online at http://www2.epa.gov/exposure

- assessment
- models/exams

- version

- index (accessed May 21, 2015). U.S. EPA (U.S. Environ mental Protection Agency). (2009) TRI explorer. Version 5.1. U.S. Environmental Protection Agency. Available online at http://www.epa.gov/triexplorer/ (accessed June 21, 2011). U.S. EPA (U.S. Environmental P rotection Agency). (2010a) Cyclododecane, 1,2,5,6,9,10

- hexabromo

- . CAS No. 3194

- 55

- 6. Inventory update reporting (IUR). Non

- confidential 2006 IUR records by chemical, including manufacturing, processing and use information. U.S. Environmental Protection Ag ency. Available online at http://www.epa.gov/iur/index.html (accessed June 1, 2011). U.S. EPA (U.S. Environmental Protection Agency). (2010b) Hexabromocyclododecane (HBCD) action plan. U.S. Environmental P rotection Agency. Available online at http://www.epa.gov/oppt/existingchemicals/pubs/actionplans/RIN2070

- AZ10_HBCD%20action%20 plan_Final_2010

- 08

- 09.pdf (accessed June 29, 2011). U.S. EPA (U.S. Environmental Protection Agency). (2011a) CAS No. 3194

- 55

- 6. Non

confidential Production Volume Information submitted by companies under the 1986, 1990, 1994, 1998, and 2002 IUR. U.S. Env ironmental Protection Agency. Available online at http://www.epa.gov/iur/tools/data/2002

- vol.html (accessed June 2, 2011). U.S. EPA (U.S. Environmental Protection Agency). (2011b) EPI Suite r esults for CAS 003194

- 55

- 6. Download EPI Suite TM v4.0. U.S. Environmental Protection Agency. Available online at http://www.epa.gov/opptintr/exposure/pubs/episuitedl.htm (accessed Ju ne 30, 2011). Veith, GD; DeFoe, DL; Bergstedt, BV. (1979) Measuring and estimating the bioconcentration factor of chemicals in fish. J Fish Res Board Can 36:1040

– 1048. Zhao, YY; Zhang, XH; Sojinu, OS. (2010) Thermodynamics and photochemical properties o f alpha, beta, and gamma

- hexabromocyclododecanes: a theoretical study. Chemosphere 80:150

– 156. Zhou, D; Wu, Y; Feng, X; et al. (2014) Photodegradation of hexabromocyclododecane (HBCD) by Fe(III) complexes/H2O 2 under simulated sunlight. Environ Sci Poll ut Res 21:6228

– 6233.40

ATTACHMENT A : EPI Suite Output for HBCD , CAS Registry Number 3194 - 55 - 6 SMILES : BrC(C(Br)CCC(Br)C(Br)CCC(Br)C(Br)C1)C1 CHEM : 1,2,5,6,9,10

- HEXABROMOCYCLODODECANE MOL FOR: C12 H18 Br6 MOL WT : 641.70

------------------------------ EPI SUMMARY (v4.10)

-------------------------- Physical Property Inputs: Log Kow (octanol

- water): 5.63 Boiling Point (deg C) :

------ Melting Point (deg C) : 190.00 Vapor Pressure (mm Hg) : 4.7E

- 007 Water Solubilit y (mg/L): 0.0086 Henry LC (atm

- m3/mole) :

------ Log Octanol

- Water Partition Coef (SRC): Log Kow (KOWWIN v1.68 estimate) = 7.74 Boiling Pt, Melting Pt, Vapor Pressure Estimations (MPBPVP v1.43): Boiling Pt (deg C): 462.03 (Adapted Stein & Brown method) Melting Pt (deg C): 180.03 (Mean or Weighted MP) VP(mm Hg,25 deg C): 1.29E

- 008 (Modified Grain method) VP (Pa, 25 deg C) : 1.72E

- 006 (Modified Grain method) VP (exp database): 4.70E

- 07 mm Hg (6.27E

- 005 Pa) at 21 deg C Subcooled liquid VP: 2.01E

- 005 mm Hg (21 deg C, user

- entered VP ) : 0.00268 Pa (21 deg C, user

- entered VP ) Water Solubility Estimate from Log Kow (WSKOW v1.42): Water Solubility at 25 deg C (mg/L): 0.003653 log Kow used: 5.63 (user entered) melt pt used: 190.00 deg C Water Sol (Exper. database match) = 0.0086 mg/L (25 deg C) Exper. Ref: TSCATS; OTS0523262 Water Sol Estimate from Fragments: Wat Sol (v1.01 est) = 0.0030722 mg/L ECOSAR Class Program (ECOSAR v1.00): Class(es) found: Neutral Organics Henrys Law Constant (25 deg C) [HENRYW IN v3.20]: Bond Method : 1.72E

- 006 atm

- m3/mole (1.74E

- 001 Pa

- m3/mole) Group Method: 6.43E

- 011 atm

- m3/mole (6.52E

- 006 Pa

- m3/mole) Exper Database: 4.60E

- 05 atm

- m3/mole (4.66E+000 Pa

- m3/mole) For Henry LC Comparison Purposes: User

- Entered Henry LC: not ente red Henrys LC [via VP/WSol estimate using User

- Entered or Estimated values]: HLC: 4.614E

- 005 atm

- m3/mole (4.676E+000 Pa

- m3/mole)41

VP: 4.7E

- 007 mm Hg (source: User

- Entered) WS: 0.0086 mg/L (source: User

- Entered) Log Octanol

- Air Partition Coefficient (25 deg C) [KOAWIN v1.10]: Log Kow used: 5.63 (user entered) Log Kaw used:

- 2.726 (exp database) Log Koa (KOAWIN v1.10 estimate): 8.356 Log Koa (experimental database): None Probability of Rapid Biodegradation (BIOWIN v4.10): Biowin1 (Linear Model) : 0.1650 Biowin2 (Non

- Linear Model) : 0.0000 Expert Survey Biodegradation Results: Biowin3 (Ultimate Survey Model): 1.9548 (months ) Biowin4 (Primary Survey Model) : 3.1342 (weeks ) MITI Biodegradation Probability: Biowin5 (MITI Linear Model) :

- 0 .4234 Biowin6 (MITI Non

- Linear Model): 0.0000 Anaerobic Biodegradation Probability: Biowin7 (Anaerobic Linear Model): 2.5072 Ready Biodegradability Prediction: NO Hydrocarbon Biodegradation (BioHCwin v1.01): Structure incompatible with current estimatio n method! Sorption to aerosols (25 Dec C)[AEROWIN v1.00]: Vapor pressure (liquid/subcooled): 0.00268 Pa (2.01E

- 005 mm Hg) Log Koa (Koawin est ): 8.356 Kp (particle/gas partition coef. (m3/ug)): Mackay model : 0.00112 Octanol/air (Koa) model: 5.57E

- 00 5 Fraction sorbed to airborne particulates (phi): Junge

- Pankow model : 0.0389 Mackay model : 0.0822 Octanol/air (Koa) model: 0.00444 Atmospheric Oxidation (25 deg C) [AopWin v1.92]: Hydroxyl Radicals Reaction: OVERALL OH Rate Constant = 5.0148 E

- 12 cm3/molecule

- sec Half

- Life = 2.133 Days (12

- hr day; 1.5E6 OH/cm3) Half

- Life = 25.595 Hrs Ozone Reaction: No Ozone Reaction Estimation Fraction sorbed to airborne particulates (phi): 0.0605 (Junge

- Pankow, Mackay avg) 0.00444 (Koa method) Note: the s orbed fraction may be resistant to atmospheric oxidation Soil Adsorption Coefficient (KOCWIN v2.00):42

Koc : 9.678E+004 L/kg (MCI method) Log Koc: 4.986 (MCI method) Koc : 7.689E+004 L/kg (Kow method) Log Koc: 4.886 (Kow method) Aqueous Base/Acid

- C atalyzed Hydrolysis (25 deg C) [HYDROWIN v2.00]: Total Kb for pH > 8 at 25 deg C : 1.174E

- 007 L/mol

- sec Kb Half

- Life at pH 8: 1.871E+005 years Kb Half

- Life at pH 7: 1.871E+006 years (Total Kb applies only to esters, carbmates, alkyl halides) Bioaccumulati on Estimates (BCFBAF v3.01): Log BCF from regression

- based method = 3.382 (BCF = 2408 L/kg wet

- wt) Log Biotransformation Half

- life (HL) = 0.8413 days (HL = 6.939 days) Log BCF Arnot

- Gobas method (upper trophic) = 3.372 (BCF = 2354) Log BAF Arnot

- Gobas method (upper trophic) = 3.643 (BAF = 4391) log Kow used: 5.63 (user entered) Volatilization from Water: Henry LC: 4.6E

- 005 atm

- m3/mole (Henry experimental database) Half

- Life from Model River: 34.83 hours (1.451 days) Half

- Life from Model Lake : 592 .3 hours (24.68 days) Removal In Wastewater Treatment (recommended maximum 95%): Total removal: 95.69 percent Total biodegradation: 26.32 percent Total sludge adsorption: 69.32 percent Total to Air: 0.06 percent (using Biowin/EPA draft m ethod) Level III Fugacity Model: Mass Amount Half

- Life Emissions (percent) (hr) (kg/hr) Air 0.188 26 2100 Water 9.18 2.4e+005 7900 Soil 72.1 2.4e+005 0 Sediment 18.6 2.4e+003 0 Persistence Time: 9.42e+003 hr43

ATTACHMENT B : ECOSAR Output44 3. HUMAN HEALTH HAZARD ASSESSMENT
3.1. SUMMARY HBCD is absorbed via the gastrointestinal tract and metabolized in rodents (Arita et al., 1983; translated from Japanese; Yu and Atallah, 1980; Szabo et al., 2010; Szabo et al., 2011b). Once absorbed, HBCD is distributed to a number of tissues, including f atty tissue, muscle, and the liver (Reistad et al., 2006; van der Ven et al., 2006; 2009; Brandsma et al., 2009; Hakk et al., 2012; Sanders et al., 2013). Elimination of HBCD is predominantly via feces (as the parent compound), but it is also eliminated in urine (as secondary metabolites) (Arita et al., 1983; translated from Japanese; Yu and Atallah, 1980; Szabo et al., 2010). HBCD has been detected in human milk and serum (Antignac et al., 2008; Weiss et al., 2006; Kakimoto et al. 2008; Rawn et al., 2014; Abdallah and Harrad 2011; Meijer et al., 2008; Thomsen et al., 2007; Fangstrom et al., 2005; 2008; Covaci et al., 2006; Johnson

- Restrepo, 2008). Acute oral exposure to HBCD resulted in mortality with a median lethal dose (LD 50 ) of 1258 mg/kg in male Charl es River CD rats, 680 mg/kg in female Charles River CD rats, and >10,000 mg/kg in Wistar albino rats (IRDC, 1978a; Gulf South Research Institute, 1988). Acute inhalation exposure to HBCD resulted in some minor symptoms (such as eye squint, slight dyspnea, salivation, lacrimation, and nasal discharge), and Gulf South Research Institute identified a median lethal concentration (LC 50 ) of >200 mg/L (IRDC, 1977; Lewis and Palanker, 1978; IRDC, 1978b; Gulf South Research Institute, 1988). An LD 50 for dermal toxicity in male and female New Zealand white rabbits was estimated to be 3,969 mg/kg (IRDC, 1978a). The available literature indicates that HBCD is not a dermal irritant at concentrations up to 0.5 m l , but one study found HBCD to be a mild skin allergen ( IRDC, 1978a; Gulf South Research Institute, 1988; Momma et al., 1993). Studies have found HBCD to be an eye irritant (IRDC, 1978a; Gulf South Research Institute, 1988).45

In short

- term and subchronic studies, HBCD demonstrated effects on the thyroid and live r (van der Ven et al., 2006; Chengelis, 1997 as cited by U.S. EPA, 2001; Chengelis, 2001; 2002). In two studies, however, most of these effects were not significant after a recovery period of 14 days and 28 days, respectively (Chengelis, 1997 as cited by U .S. EPA, 2001; Chengelis, 2001; 2002). Older short

- term studies , using an HBCD product that is no longer manufactured , concluded that observed liver findings were adaptive rather than adverse (Zeller and Kirsch, 1969 as cited by U.S. EPA, 2001; Zeller and Kirsch, 1970 as cited by U.S. EPA, 2001). The developmental and reproductive toxicity of HBCD have been investigated in a single

- generation study (van der Ven et al., 2009), a study incorporating gestational and post

natal exposure (Saegusa et al., 2009), a two
- generation study (Ema et al., 2008), a gestational exposure study ( Murai et al., 1985; Japanese with English abstract), and a neurotoxicity study of adult mice neonatally exposed to HBCD (Eriksson et al., 2006)
. A few reproductive effects were observed in these studies. Ema et al. (2008) reported a decrease in the number of primordial follicles in first generation (F1) female Crl:CD(SD) rats at 1,500 ppm (approximately 138 mg/kg

- day) and a significant increase in the number of litters lost in the F1 generation at 15,000 ppm (approximately 1,363 mg/kg

- day). These authors reported no other significant treatment

- related effects in any generation for indicators of reproductive health, including: estrous cyclicity, sper m count and morphology, copulation index, fertility index, gestation index, delivery index, gestation length, number of pups delivered, number of litters, or sex ratios. Neither van der Ven et al. (2009) nor Saegusa et al. (2009) observed reproductive effe cts at the doses tested (up to 100 mg/kg

- day in Wistar rats and up to approximately 1,504.8 mg/kg

- day in Sprague

- Dawley rats, respectively). Developmental effects were observed in Wistar rats by van der Ven et al. (2009), including immune system effects, indications of liver toxicity, and decreases in bone mineral density at very low doses (i.e., <1.3 mg/kg

- day). The authors noted, however, that the vehicle46

used (corn oil) may have affected observations at higher doses, including: increased mortality duri ng lactation, decreased liver weight in males, decreased adrenal weight in females, decreased plasma cholesterol in females, and other immunological markers of toxicity. Saegusa et al. (2009) observed an increased relative thyroid weight and decreased triiodothyronine (T3) levels in F1 male Sprague

- Dawley rats at postnatal week (PNW) 11 following dietary exposure to 1,000 ppm (approximately 146.3 mg/kg

- day) HBCD. Saegusa et al. (2009) also reported a significant reduction in the number of CNPase

- positive oligodendrocytes at 10,000 ppm (approximately 1,504.8 mg/kg

- day). Ema et al. (2008) reported a reduced viability index on Day 4 and Day 21 of lactation among second generatio n (F2) offspring at 15,000 ppm (approximately 1,363 mg/kg

- day). Ema et al. (2008) observed additional developmental effects at doses as low as 1,500 ppm (approximately 115 and 138 mg/kg

- day for F1 males and females, respectively), including: an increase in dihydrotestosterone (DHT) in F1 males and an increased incidence of animals with decreased thyroid follicle size in both sexes and generations. These authors reported no effects on sexual development indicated by anogenital distance, vaginal opening, or p reputial separation among F1 or F2 generations. Developmental toxicity was not observed in F1 Wistar rats following dietary exposure to HBCD during gestation (Murai et al., 1985; Japanese with English abstract). Eriksson et al. (2006) observed effects on s pontaneous motor behavior, learning, and memory in adult NMRI mice following exposure to HBCD on postnatal day (PND) 10. At 0.9 mg/kg, the authors reported significantly reduced mean locomotor activity. Studies evaluating the carcinogenic potential of HB CD in humans or animals were not available , so no determination of carcinogenicity could be made. HBCD, however, does not appear to be genotoxic. The available toxicity data for HBCD includes evidence of developmental and reproductive toxicity in rats an d mice. Based on the available data, EPA concludes that there is sufficient evidence to support a concern for moderately high to high toxicity of HBCD.47

3.2. KINETICS AND METABOLISM In humans, HBCD has been detected in breast milk, adipose tissue, blood , and umbilical serum (Antignac et al., 2008; Weiss et al., 2006; Kakimoto et al. 2008; Rawn et al., 2014; Abdallah and Harrad 2011; Meijer et al., 2008; Thomsen et al., 2007; Fangstrom et al., 2005; 2008; Covaci et al., 2006; Johnson

- Restrepo, 2008). Howev er, the majority of data regarding HBCD kinetics and metabolism comes from rodent studies (Arita et al., 1983; translated from Japanese; Yu and Atallah, 1980; Szabo et al., 2010; Szabo et al., 2011a; b; Reistad et al., 2006; van der Ven et al., 2006; 2009; Brandsma et al., 2009; Hakk et al., 2012; Sanders et al., 2013). Results from oral administration studies in rats indicate that ingested HBCD is: (1) moderately absorbed, (2) distributed to a number of tissues with preferential distribution and accumulati on of unchanged HBCD in fatty tissue, and (3) eliminated predominantly as unchanged HBCD and metabolites in feces. These studies provide qualitative evidence that HBCD can be metabolized in rodents.
3.2.1. Absorption. Several animal studies have examined absorption of HBCD. Arita et al. (1983; translated from Japanese) studied the absorption of HBCD in four male Wistar rats following administration of a 500 mg/kg dose administered by gavage (powder suspended in corn oil) for five consecutive days. These au thors measured intestinal absorption of HBCD using isolated sections of small intestines from two rats. They observed that approximately 12% of the administered HBCD was found in the intestinal tissue. Yu and Atallah (1980) administered a single dose of [ 1 4 C]

- HBCD dissolved in olive oil (7

– 9 mg/kg) by gavage to Sprague

- Dawley rats (eight females and two males) and observed that the administered [ 14 C]

- HBCD was partially absorbed in the gastrointestinal tract. HBCD reached a peak concentration in the blood at 4 hours post dosing and then gradually declined. Approximately 43% of the administered [ 14 C]

- HBCD was detected in body tissues (excluding the gastrointestinal tract) 8 hours after dosing. In a48

similar study, Szabo et al. (2010) administered a single dos e of 3 mg γ

- [ 14 C]

- HBCD/kg dissolved in oil by gavage to female C57BL/6 mice and reported that 85% of the administered dose was absorbed after exposure. In a follow

- up study, approximately 90% of α

- HBCD administrated was absorbed after exposure (Szabo et al ., 2011b).

3.2.2. Distribution. Several studies have assessed the distribution of HBCD in human tissues. HBCD has been detected in breast milk, adipose tissue, blood , and umbilical serum (Antignac et al., 2008; Weiss et al., 2006; Kakimoto et al. 2008; Rawn et al., 2014; Abdallah and Harrad 2011; Meijer et al., 2008; Thomsen et al., 2007; Fangstrom et al., 2005; 2008; Covaci et al., 2006; Johnson

- Restrepo, 2008 ; Kim an d Oh, 2014 ). The composition of HBCD isomers in most rodent toxicity studies resembles that of industrial grade HBCD, which may differ from human exposure to certain foods that have been shown to contain elevated fractions of α

- HBCD (Schecter et al., 2012)
. Antignac et al. (2008) analyzed samples from 26 mother

- newborn pairs in France for the presence of brominated flame retardants. α

- HBCD was detected in approximately half of the adipose samples (1

– 3 ng/g lipid weight) and in 7 of the breast milk sampl es (2.5

– 5 ng/g lipid weight). No β

- or γ - isomers were detected in any of the samples. The authors were unable to measure HBCD in serum samples, however, due to analytical limitations. Weiss et al. (2006) reported HBCD (mostly the α

- HBCD diastereomer but also γ

- HBCD) concentration in human serum from 53 wives of Swedish fisherman (identified by the authors as a high exposure group) ranging from <0.24

– 3.4 ng/g fat with a median concentration of 0.46 ng/g fat (n=50). In a study of Japanese women, α

- HBCD (0 .43

– 1.9 ng/g lipid weight) was detected in all 11 pooled breast milk samples collected between 1988 and 2006, γ

- HBCD (0.24

– 2.60 ng/g lipid) was detected in 7 of the 11 samples, and β

- HBCD was not detected (<0.2 ng/g lipid) (Kakimoto et al., 2008). Abda llah and Harrad (2011) measured HBCD in 34 human milk samples from Birmingham, UK.49

HBCD was detected in all samples and the average concentration of all HBCD isomers was 5.96 ng/g lipid weight. α

- HBCD comprised 62

– 95% of HBCD detected in the study, while β

- HBCD and γ

- HBCD comprised 2

– 18% and 3

– 33% of HBCD detected in the study, respectively. In an analysis of HBCD concentrations in sera from the Canadian Health Measures Survey, researchers detected HBCD in all samples analyzed (Rawn et al., 2014). In most samples, HBCD was detected as <1 ng/g lipid, similar to global concentrations. α

- HBCD was the biggest contributor to total HBCD, although β

- HBCD and γ

- HBCD were detected in 23% and 35% of the samples, respectively. No differences in total HBCD concent ration were associated with age
or sex. Researchers have also found that HBCD can be transferred through the placenta during critical stages of infant development (Meijer et al., 2008). These authors detected HBCD in the maternal serum (n=69) and cord ser um (=12) of pregnant women and their infants in the 35 th week of pregnancy. Maternal serum concentrations of HBCD ranged from below the limit of detection to 7.4 ng/g lipid (median value = 0.7 ng/g lipid) and cord serum concentrations ranged from 0.2 to 4. 3 ng/g lipid (median value = 0.2 ng/g lipid). More recently, significant correlations between HBCD concentrations in sera from mothers and infants have been found, suggesting that that maternal transfer is important (Kim and Oh, 2014). In an occupational study, Thomsen et al. (2007) measured HBCD isomer concentrations in airborne dust samples and in serum of ten male workers exposed to HBCD in the workplace. There was no correlation, however, between HBCD concentrations in airborne dust samples collected a nd HBCD concentrations in blood serum. The mean HBCD concentration in serum was 190 ng/g lipids (median = 101 ng/g lipids; range = 6

– 856 ng/g lipids). Additionally, HBCD was not detected (detection limit = 1 ng/g lipids) in any of the serum samples taken from a reference group with no occupational exposure to HBCD. The authors reported a notably high50

contribution of γ

- HBCD (39%) to the total HBCD serum concentration compared with findings from other studies. In general, measured concentrations of HBCD wit hin humans have risen over the past couple of decades. Fangstrom et al. (2005; 2008) investigated the temporal trends of brominated flame retardants in breast milk of women in Stockholm, Sweden from 1980 through 2004. A large increase of HBCD (isomers not determined) was measured in breast milk from these women from 1980 (0.084 ng/g lipid) to 2004 (0.39 ng/g lipid) (Fangstrom et al., 2005). A similar increase in HBCD concentration in breast milk was reported by the same authors in a more recent analysis of this study population (Fangstrom et al., 2008). This trend was also observed in HBCD concentration in the breast milk of Japanese women. HBCD concentration in breast milk samples from women in Osaka, Japan was below the limit of detection from 1973 through 1983, increased to 0.43 ng/g lipid in 1988, and measured between 1 and 4 ng/g lipid from 1993 through 2006 (Kakimoto et al., 2008). No temporal trend, however, was apparent in a study conducted using human breast milk samples collected in Australia from 1 993 to 2009 (Toms et al., 2012). Studies on the distribution of HBCD in rodents have measured HBCD concentrations in organs such as the liver, kidneys, and brain. HBCD was identified in the liver (1,248 ± 505 μg/kg, w/w) and brain tissue (49 ± 13 μg/kg, w/ w) of 3 male Wistar rats 72 hours after a single intraperitoneal (i.p.) injection of 18.7 ± 0.7 mg/kg (Reistad et al., 2006). The authors reported that the relative isomeric concentration in the liver was approximately 3.2, 17.5, and 78.6% for
α - HBCD, β

- HB CD, and γ

- HBCD, respectively. The relative isomeric concentration in the brain was approximately 1.9, 3.3, and 95.9% for α

- HBCD, β

- HBCD, and γ

- HBCD, respectively. The relative total isomeric composition of the test compound, however, was 1.6, 8.1, and 90.3 % for
α - HBCD, β

- HBCD, and γ

- HBCD, respectively. These results suggest that HBCD administered by i.p. injection is distributed to the liver and the brain, and that the isomeric composition of HBCD in the liver and brain differs from the composition of the a dministered material.51

Similar results were obtained in single oral dose toxicity studies. Following a single gavage exposure of γ

- HBCD to female C57BL/6 mice, the largest percentage of the dose was observed in the liver, followed by the skin and muscle (S zabo et al., 2010). In a similar study, an administered dose of α

- HBCD was found primarily in adipose, liver muscle, and skin (Szabo et al., 2011b). Accumulation was dose dependent with nonlinear accumulation at higher doses. In a study investigating the d ifferences in tissue distribution of α

- and γ

- HBCD between adult and developing mice, the authors reported that while distribution patterns were similar between adult and developing mice, concentrations of each diastereomer were higher in the pup’s liver, fat, kidney, brain, blood, muscle, and lungs than the adult’s (Szabo et al., 2011a). This finding suggests that developmental stage exposure to α

- and γ

- HBCD may result in increased body burden compared with adults. A 10

- day gavage exposure to γ

- HBCD resu lted in increased deposition in the adipose tissue and blood compared to a single gavage exposure (Yu and Atallah, 1980). Eight hours after administration of a single oral dose of 7

– 9 mg [ 14 C]

- HBCD/kg to male and female Sprague

Dawley rats, most of the a dministered [ 14 C] was detected in the fat (20%), muscle (14%), and liver (7%), with smaller amounts (<1%) in the blood, heart, lung, gonads, uterus, spleen, kidney, and brain. [ 14 C]

- HBCD concentrations in most tissues decreased over the course of 72 hours, but concentrations decreased more slowly in adipose tissue. The concentration of [ 14 C] in the fat and muscle of females 48 hours after exposure was higher than that of males measured at the same time point. Following oral administration of five consecutiv e daily doses of HBCD (500 mg/kg

day) to Wistar rats, HBCD was detected in fatty tissue collected from the periphery of the kidney (0.33

– 0.67 mg/g tissue), but not in the spleen, kidneys, pancreas, liver, or heart (Arita et al., 1983; translated from Jap anese).52

3.2.3. Metabolism. There is a growing body of evidence demonstrating that HBCD is metabolized in humans and rodents. Abdallah and Harrad (2011) detected two HBCD degradation products for the first time in human tissues. The degradation products pentabromocyclododecenes (average concentration of 0.04 ng/g lipid weight; n=9) and tetrabromocyclododecadienes (average concentration of 0.15 ng/g lipid weight; n=25) were both detected in human milk samples from Birmingham, UK. Toxicokinetic studies of H BCD orally administered to rats provide evidence that HBCD can be metabolized in rodents. In one study, HBCD was reported to be rapidly metabolized to unknown polar metabolites in the blood, muscle, liver, kidney, and nonfat tissues of Sprague

Dawley rats following a single oral dose (7

– 9 mg/kg) of [ 14 C]

- HBCD (Yu and Atallah, 1980). Unmetabolized HBCD, however, comprised all of the [ 14 C] detected in fat. In a more recent study, four different groups of hydroxylated HBCD metabolites were identified in rat adipose, liver, lung, and muscle tissue by liquid and gas chromotography combined with mass spectrometry: monohydroxy metabolites of HBCD, pentabromocyclododecene (PBCDe), tetrabromocyclododecene (TBCDe), and dihydroxy

- HBCD (Brandsma et al., 2009). The is omeric composition of HBCD detected in the livers of Wistar rats following a 28

- day gavage study with exposure beginning at the age of 7 weeks (van der Ven et al., 2006) and a dietary reproduction study with exposure occurring during gestation and lactatio n (van der Ven et al., 2009) was different from the isomeric composition of the administered technical compound. Much lower ratios of γ

- HBCD to α

- HBCD were observed in the liver compared with the administered compound. In van der Ven et al. (2009), the con centration of α

- and γ

- HBCD in the liver (expressed as μg/g lipid) of F1 males and females increased in a dose

- related fashion. β

HBCD was only detected at the highest exposure concentrations. The ratio of γ/α isomers was 8.1 in the administered technical mixture of HBCD, but the ratio in the livers of exposed F1 animals ranged from 0.04 to 2.2 in females and from 0.2 to 1.4 in males. Concentrations were53

higher in females over the entire range of doses (four and six times higher for γ

- and α

- isomers, respe ctively, at the highest doses). In van der Ven et al. (2006), the average ratio between γ

- and
α - HBCD was 8.1 in the administered technical mixture, whereas the average ratio in the liver was 2.9 in females and 1.9 in males. Females had higher HBCD concent rations in the liver, and the ratio of lipid

- normalized γ

- HBCD to α

- HBCD in the liver decreased with increasing exposure concentration (liver lipid concentrations did not differ among dose groups) and the duration of oral exposure. HBCD isomers have been d etected in the blubber of harbor porpoises and common dolphins stranded on the coasts of Ireland, Scotland, The Netherlands, Belgium, France, and Spain (Zegers et al., 2005). Tissues from 10 harbor porpoises and 9 common dolphins were analyzed for HBCD ste reoisomers by liquid chromatography

- mass spectrometry (LC/MS) and were found to contain only α

- HBCD. A follow

- up in vitro study of microsomes (from Wistar rats and a harbor seal) treated with equimolar mixtures of α

- , β

- , and γ

- HBCD demonstrated that the β

- and γ

- HBCD isomers were metabolized by CYP450, but the α

- HBCD isomer was not. α

HBCD was also found to be the predominant isomer identified in serum samples taken from Swedish fishermen’s wives (Weiss et al., 2006). Further evidence has demonstrated rapi d metabolism of γ

- HBCD. Szabo et al. (2010) reported that approximately 80% of an administered dose of γ[ 14 C]

- HBCD was metabolites and approximately 15% was stereoisomerized HBCD products in the liver and feces. β

- and γ

- HBCD were detected in the liver, wh ereas α - , β

- , and γ

HBCD were observed in the fat and feces. The bile, serum, and urine contained 100% metabolites. In a more recent study, α

- HBCD and γ

- HBCD produced distinct metabolites after a single administration of HBCD to C57BL/6 female mice (Hakk e t al., 2012). In mice exposed to
α - HBCD, four hydroxylated metabolites were detected in fecal extracts, and one of these metabolite isomers was found in liver, brain, and adipose tissue. Fecal extracts from mice exposed to γ

- HBCD, however, contained multip le isomers of monohydroxy

-54

pentabromocyclododecene, dihydroxy

- pentabromocylcododecene, and dihydroxy

pentabromocyclododecadiene, while in liver and adipose tissue only a single monohydroxy

ptentabromocyclododecane metabolite was detected. The extent of meta bolism of β

- HBCD in C57BL/6 female mice was found to be similar to that of γ - HBCD (Sanders et al., 2013).

3.2.4. Excretion. In humans, Geyer et al. (2004) estimated the HBCD total body half

- life and the terminal half

- life from adipose tissue as 64 days ( with a range of 23

– 219 days). The total body half

- life was estimated using daily intake and total body burden under steady state conditions in
non - occupationally exposed adult humans according to a one compartment, open pharmacokinetic model. The termina l half

- life was estimated from applying an interspecies adjustment to the terminal half

- life in rats found in the literature. No other human data on elimination of HBCD were found in the literature. In rodents, HBCD is eliminated predominantly in the feces following either single or repeat oral administration. Following a single oral dose of 7

– 9 mg [ 14 C]

- HBCD/kg, Yu and Atallah (1980) reported that 0.12% of the administered HBCD was detected in urine 8 hours after administration. Total HBCD in the body (e xcluding the gastrointestinal tract) diminished to 24% of the administered dose after 24 hours and 18% of the administered dose after 48 hours. Approximately 77% of the administered HBCD was eliminated in the feces and 16% was eliminated in the urine withi n 72 hours. The rate of elimination increased until 12 hours after administration, and it decreased thereafter. Unchanged HBCD was not detected in urine. Of the radioactivity detected in feces, about 72% was unextractable (i.e., unmetabolized HBCD) and
was not further characterized. The elimination curves for both urine and feces were biphasic, with the peak rate of elimination at 12 hours post dosing. Arita et al. (1983; translated from Japanese) detected HBCD in the feces, but not in the urine, of four ma le Wistar rats administered five consecutive daily oral doses (500 mg/kg

- day) of HBCD. The cumulative fecal excretion of55

HBCD increased in a linear fashion over 5 days of administration. Both the percentage of cumulative excretion measured (32

– 35%) for e ach of the 5 days of dosing and the percentage of administered dose excreted per day (29

– 37%) were roughly constant. Szabo et al. (2010) observed that approximately 80% of an administered dose of γ[ 14 C]

- HBCD was eliminated in the urine (30%) and feces (50%) four days following a single gavage dose of 3 mg/kg. The authors observed no dose dependency in fecal or urinary elimination of γ

HBCD. The extent of excretion of β

- HBCD was found to be similar to that of γ

- HBCD (Sanders et al., 2013). Elimination, b oth whole body and from individual tissues, of α

- HBCD was found to be biphasic and was excreted in the feces as parent and metabolites, whereas urine only contained metabolites (Szabo et al., 2011b). The presence of polar metabolites in the urine and blood were an important factor in determining the initial whole

- body half

- life. Initial half

- lives were 1

– 3 days and terminal half

- lives of 17 days were observed. These findings suggest the potential for α

- HBCD bioaccumulation. Abdallah and Harrad (2011) also concluded, using a pharmacokinetic model for humans, that α

- HBCD had a higher bioaccumulation potential and longer half

- life than β

- and γ

- HBCD.
3.3. EFFECTS OF ACUTE EXPOSURE 3.3.1. Acute Oral Toxicity Studies. Several studies assessing acute oral toxicity of HBCD have been conducted (shown in Table 3

- 1). Study summaries are below.
Acute Oral Toxicity Studies in Rats (IRDC, 1977) Male and female Charles River CD rats (5/sex) were administered 10,000 mg/kg HBCD by gavage in corn oil. Mortality was not observed in rats 14 days following gavage. In some males, hypoactivity was observed 2.5 hours post

- gavage, and corneal opacity and ptosis were observed on Days 6

– 14. A single case of diarrhea and of hypoactivity was observed in a female56

rat 4

- hours p ost
- gavage. There was no control group with which to compare the treated animals.
Acute Toxicity Studies in Rats with Residue of HBCD (IRDC, 1978a) Male and female Charles River CD rats (5/sex/dose group) were administered 215, 464, 1,000, 2,150, 4,640, or 10,000 mg/kg HBCD by gavage in corn oil and observed over 14 days. Within the first two days following gavage, an increased incidence of mortality was observed at ≥1,000 mg/kg in both males and females. There was no control group with which to compare t he treated animals. The study authors identified anLD 50 of 1258 mg/kg in males and 680 mg/kg in females.
Report on the Study of the Acute Oral Toxicity of NBCD in the Mouse (BASF, 1990) An LD 50 of > 6,400 mg/kg was identified in an acute study in which HBCD was administered to mice in a single gavage exposure as a 30% tragacanth suspension. Mice were observed for 7 days, and the observed clinical signs of toxicity included increased apathy, trembling, and late mortalities. The necropsies revealed peritonitis.
Acute Exposure Oral Toxicity Study in Rats (Pharmakon Research International Inc., 1990) HBCD was administered by gavage in corn oil to male and female Sprague

- Dawley rats (5/sex) at a single dos e level of 5,000 mg/kg. The single dose was followed by a 14

- day observation period. Observations include poor grooming, ptosis, chromodacryorrhea, lacrimation, and swelling of the face. The study authors concluded that the LD 50 was greater than 5,000 mg/k g.
Letter from Ethyl Corp to USEPA Re: Technical and Toxicity Data on Brominated Flame Retardants Including HBCD (Gulf South Research Institute, 1988)57

Male and female adult Wistar albino rats (5/ sex) received a single gavage dose of 1,000, 3,000, 5,000, or 10,000 mg/kg HBCD in corn oil and were observed for 14 days. Effects following oral exposure to HBCD were not observed. The study authors concluded that the LD 50 was >10,000 mg/kg.

Abbreviations: NR, not reported 3.3.2. Acute Inhalation Toxicity Studies. The toxicity of HBCD by the inhalation route of exposure has been assessed in several acute studies (shown in Table 3

- 2). Study summaries are below.
Acute Inhalation Toxicity Study in Rats (IRDC, 1977) Male and fe male Charles River CD rats (5/sex) were placed in cages in stainless steel chamber with a constant airflow. A total of 339.65 grams of HBCD dust powder was disseminated in a 4 hour exposure period. The rate of powder disseminated was calculated to be Table 3 - 1. Acute oral toxicity of HBCD and HBCD bottoms and residue S pecies/strain/test Exposure Result Notes Reference Rat/Charles River/LD 50 Single gavage (corn

oil) LD 50 >10,000

mg/kg No mortality; transient hypoactivity

and diarrhea; corneal opacity and ptosis in 3/5 males, which did not resolve by end of 14

- day observation. IRDC, 1977

Rat/Charles River CD/ LD 50 test of HBCD residue Single gavage (corn

oil) LD 50 = 1258

mg/kg (male); 680 mg/kg (female) Tested Firemaster 100; increased

activity, eye squint, dyspnea, lacrimation, and nasal discharge. Weight loss in all animals; recovery noted by conclusion of study. IRDC, 1978a

Mouse/NR/LD 50 Single gavage (30%

aqueous tragacanth) LD 50 >6,400

mg/kg; data not shown 7

- day observation period; increasing

apathy, trembling and late mortalities; peritonitis at necropsy. BASF, 1990 Rat/Sprague

Dawley/limit test of HBCD bottoms Single gavage (corn

oil) LD 50 >5,000

mg/kg; clinical signs This study tested HBCD bottoms

described as black solids, lacrimation, and facial swelling resolved by post

- exposure day 4; no gross lesio ns were observed. Pharmakon Research International Inc., 1990 Rat/NR/LD 50 test Single gavage (corn

oil) LD 50 >10,000

mg/kg No significant changes observed. Gulf South Research Institute, 198858

1.415 grams per minute. The chamber dust concentration was calculated to be 202.14 mg/L. Mortality was not observed. All animals demonstrated an increase in body weight and slight dyspnea was observed during the 4

- hour exposure duration, but there was no contro l group with which to compare the treated animals.
Acute Inhalation Toxicity Study in Rats with HBCD (IRDC, 1978b) Male and female Charles River CD rats (5/sex) were exposed via inhalation to 22.9 mg/L HBCD for 4 hours followed by a 14

- day observation period. After 30 minutes of exposure, rats exhibited eye squint, slight dyspnea, salivation, lacrimation, and nasal dis charge. As the exposure duration progressed, more rats exhibited symptoms. All rats demonstrated significant body weight reduction and one female rat died on Day 3. The 9 remaining rats appeared to recover and gain weight by the end of the observation peri od. Pathological examination revealed dried blood around the nares, patchy red areas on the lungs, and a distended stomach filled with yellow liquid. There was no control group with which to compare the treated animals.

An Inhalation LC 50 Study In Albino Rats (Lewis and Palanker, 1978) Male and female albino rats (5/sex) were exposed to 200 mg/L HBCD via inhalation for one hour and observed for two weeks. Mortality was not observed and the study authors identified the LC 50 as greater than 200 mg/L. All an imals demonstrated an increase in body weight, although control data was not provided.
Letter from Ethyl Corp to USEPA Re: Technical and Toxicity Data on Brominated Flame Retardants Including HBCD (Gulf South Research Institute, 1988)59

Male and female adul t Wistar albino rats (5/per sex) were exposed to 200 mg/L HBCD for one hour and observed for 14 days. Effects following inhalation exposure to HBCD were not observed.

Abbreviations: NR, not reported 3.3.3. Acute Dermal Toxicity and Dermal Irritation Studies Acute Toxicity Studies in Rats with Residue of HBCD (IRDC, 1978a) Male and female New Zealand white rabbits (10/sex/dose) were administered 1,250, 2,500, 5,000, 10,000, or 20,000 mg/kg HBCD on shaved skin. The skin of one male and one female of each group was also abraded. The administration site was occluded for 24 hours and then washed. Mortality was observed daily for 14 days after application. The study authors identified a combined male and female LD 50 for dermal toxicity of 3,969 mg/kg. Male and female New Zealand white rabbits (3/sex) were administered 0.5 m l HBCD on shaved skin. The skin of 3 rabbits was also abraded. The administration site was occluded for 4 hours and then washed. Examinations were carried out at 24, 48, and 72 hours. The authors Table 3 - 2. Acute inhalation toxicity of HBCD and a residue of HBCD Species/strain/test Exposure Result Notes Reference Rat/Charles River/LD 50 4

- hr exposure to

Firemaster 100 dust: 202.14 mg/L No mortality;

slight dyspnea at the end of exposure only All rats gained weight over

the observation period, but no control data were reported. IRDC, 1977

Rat/Wistar/limit test 1

- hr exposure to 200

mg/L GLS

- S6

- 41A (highest possible concentration) No mortality; no

clinical signs All rats gained weight over

the observation period, but no control data were reported. Lewis and Palanker, 1978

Rat/Charles River/LC 50 test of Firemaster 100 residue (liquid) 4

- hr exposure to

Firemaster 100 residue: 22.9 mg/L One death (female, Day 3 post

- exposure);

dyspnea, salivation, lacrimation, and nasal discharge during exposure, resolved by Day 3; significant body weight loss through Day 3, but beginning to recover by Day 14 post exposure; signs of nasal and lung irritation at gross necropsy of rat that died. IRDC, 1978b

Rat/NR/LC 50 test 1 hr. whole

- body

inhalation exposure to GLS

- S6

- 41A LC 50 >200 mg/L No significant effects observed. Gulf South Research Institute, 198860

concluded that HBCD is not a primary dermal irritant or a corrosive material.
Letter from Ethyl Corp to USE PA Re: Technical and Toxicity Data on Brominated Flame Retardants Including HBCD (Gulf South Research Institute, 1988) Male and female New Zealand albino rabbits (3/sex/dose) were administered 500, 2,000, 5,000, or 8,000 mg/kg HBCD on shaved and occluded skin. The skin of 3 rabbits was also abraded. After 14 days of observation, the authors concluded that HBCD is not derma lly toxic to rabbits. Male and female New Zealand albino rabbits (3/sex) were administered 0.5 m l HBCD on shaved and occluded skin. The skin of 3 rabbits was also abraded. Observations at 24 and 72 hours demonstrated no evidence of dermal irritation.

A D ermal LD 50 Study in Albino Rabbits (Lewis and Palanker, 1978) Male and female New Zealand albino rabbits (3/sex/dose) were administered 500, 2,000, 5,000, and 8,000 mg/kg HBCD on shaved and occluded skin. The skin of 3 rabbits was also abraded. After 14 da ys of observation, the authors concluded that HBCD is not dermally toxic to rabbits.
Dermatological Evaluation of a Flame Retardant, Hexabromocyclododecane (HBCD), on Guinea Pig by Using the Primary Irritation, Sensitization, Phototoxicity, and Photosens itization of Skin (Momma et al., 1993) The skin sensitization test was carried out according to the maximization test method of Magnusson and Kligman. HBCD was dissolved in olive oil to a concentration of 0.05, 0.5, or 5%. When induction of sensitization occurred, challenge doses of 0.005, 0.05, 0.5, or 5% (dissolved in acetone) were applied. The induction dose of greater than 0.5% and the challenge61

dose greater than 0.05% elicited a positive response. The increase in the concentration of induction and ch allenge doses did not further increase the percentage of positive responses or the intensity of response. The authors concluded that HBCD is a mild skin allergen.
3.3.4. Acute Eye Irritation Studies Acute Toxicity Studies in Rats with Residue of HBCD (IRDC, 1978a) Male and female New Zealand white rabbits (3/sex) were administered 0.1 m l HBCD in the cupped conjunctival sac of the right eye. Examinations were carried out until 14 days after administration. The authors concluded that HBCD is a primary ey e irr itant.
Letter from Ethyl Corp to USEPA Re: Technical and Toxicity Data on Brominated Flame Retardants Including HBCD (Gulf South Research Institute, 1988) Male and female New Zealand rabbits (3/sex) were administered 0.1 m l HBCD with no wash for 24 hours and observed for 7 days. The authors concluded that HBCD was a mild, transient ocular irritant.

3.4 EFFECTS OF REPEATED EXPOSURE
3.4.1. Effects of Short - term and Subchronic Exposure Only a few subacute and subchronic toxicity studied of HBCD were identified in the available literature. HBCD demonstrated effects mainly on the thyroid and liver in both subacute and subchronic studies. In two studies, however, most of these effects were not significant after a recovery period of 14 or 28 days. In a s ubacute study, Van der Ven et al. (2006) exposed Wistar rats (5/sex/dose) by gavage to a mixture of HBCD dissolved in corn oil at concentrations resulting in doses of 0.3,62

1.0, 3.0, 10, 30, 100, and 200 mg/kg

- day for 28 days. The isomeric composition of th e HBCD was 10.3% α, 8.7% β, and 81.0% γ. Internal dosing was verified by analysis of HBCD enantiomers in the liver of a subset of animals per dose group. Exposure started at the age of 11 weeks, and rats were observed for 28 days according to OECD Guidelin e 407 (OECD, 1995). Food consumption, body weight, and clinical observations were assessed during the exposure period. At the end of the exposure period, animals were necropsied. All standard organ and tissue assessments were performed. Dose

- response data were analyzed using the benchmark dose approach in PROAST software. The authors used the likelihood ratio test for model selection based on the modeling methodology outlined by Slob (2002). The critical effect dose (CED, also referred to as the benchmark dose (BMD)) was calculated at a predefined critical effect size (CES) of 10 or 20%, depending on the endpoint. The authors identified a CES of 10% as the default, noting that this effect size would cover the hazard for the most sensitive subjects in a popu lation. A CES of 20% was used for liver weight and immune parameters as lower effects sizes could represent a nontoxic physiologic response. The 5% lower confidence bound of the CED is considered a benchmark dose lower bound confidence limit (BMDL) for con tinuous data. In the case of a more than 10

- fold difference between CED and BMDL, the calculated BMDLs were not considered statistically valid. There was a significant dose

- dependent weight increase of the pituitary, thyroid, and liver in females (BMDL of 29.9, 1.6, and 22.9 mg/kg

- day, respectively, and reported CED of 50.6, 3.4, and 29.9 mg/kg

- day, respectively). Histopathology of the thyroid demonstrated that thyroid follicles were smaller, depleted, and had hypertrophied epithelium in female rats. There was a decrease in thymus weight in males (BMDL of 104.2 mg/kg

- day and reported CED of 176.6 mg/kg

- day). No effects on sperm quantity or quality were noted. There was a dose

- dependent decrease in total thyroxine (TT4) in females (BMDL of 55.5 mg/kg

- day and reported CED of63

76.1 mg/kg

- day) and there were no effects of HBCD on total triiodothyronine (TT3) levels. In the liver, a dose

- dependent increase in T4

- UDP glucuronosyltransferase (T4

- UGT) activity was observed in both males and females. A valid BMDL, how ever, was not determined for this parameter due to high statistical uncertainty. Dose

- dependent decreases in apolar liver retinoid concentrations (e.g., retinyl palmitate, retinol) were also reported for females, but these likely reflect increased liver we ight as there was no effect on total apolar retinoid concentration in the liver. Females demonstrated dose

- dependent decreases in concentrations of alkaline phosphatase (BMDL of 18.9 mg/kg

- day and CED of 33.9 mg/kg

- day) and glucose (BMDL of 70.8 mg/kg

- day and CED of 100.0 mg/kg

- day) along with increases in total protein/albumin (BMDL of 142.7 and 197.5 mg/kg

- day, respectively, and CED of 194.3 and 316.8 mg/kg

- day, respectively). Males demonstrated dose

- dependent decreases in glucose (BMDL of 57 mg/kg

- day an d CED of 98.1 mg/kg

- day). There were some reported effects on mean red blood cell volume and mean erythrocyte hemoglobin concentration in males, but these effects were below the CES. Increased mineral density of trabecular bone at the femur and tibia metap hysis (BMDL of 69.9 and 49.3 mg/kg

- day, respectively, and CED of 139.3 and 91.1 mg/kg

- day, respectively) were observed in female rats. The most sensitive endpoint was the 10% increase in thyroid weight in females with a calculated BMDL of 1.6 mg/kg

- day (CE D = 3.4 mg/kg

- day). According to EPA research, adverse rodent non

- cancer thyroid effects due to chemically

- induced thyroid

- pituitary disruption are considered relevant to humans (Hill et al., 1998). In another subacute study, HBCD was administered orally by gavage in corn oil to Sprague

- Dawley Crl:CD BR rats for 28 days at doses of 0, 125, 350, or 1,000 mg/kg

- day (6 rats/sex/dose in 125 and 350 mg/kg

- day groups and 12 rats/sex/dose in the control and 1,000 mg/kg

- day groups) (Chengelis, 1997 as cited by U.S . EPA, 2001). At the end of 28 days, 6 rats/sex/dose were necropsied, while the remaining rats in the control and 1,000 mg/kg

- day64

groups were untreated for a 14

- day recovery period prior to necropsy. Animals were observed twice daily for mortality and morb idity, clinical signs were recorded daily, and body weight and food consumption were measured weekly. Functional Observational Battery and motor activity evaluations were carried out prior to study initiation, during the last week of HBCD administration (W eek 3), and during the recovery period (Week 5). Samples for hematology and serum chemistry evaluations were collected at the primary necropsy (Day 28) and recovery necropsy (Day 42). All standard organ and tissue assessments were performed. Body weights, weight gain, and food consumption were unaffected by HBCD. No changes in the Functional Observational Battery and motor activity tests were reported. At Week 4, a statistically significant increase in mean activated partial thromboplastin time was noted i n males at 1,000 mg/kg

- day along with a significant decrease in females at 1,000 mg/kg

- day. The authors noted that these findings were not, however, toxicologically significant. Some serum chemistry findings were statistically significant, but these findin gs were not dose related or toxicologically significant. No treatment

- related gross or microscopic lesions were observed. Absolute liver weight was significantly increased at 28 days in males (1,000 mg/kg

- day) and females (350 and 1,000 mg/kg

- day). Additio nally, liver to body weight ratios in the 350 and 1,000 mg/kg

- day male and female dose groups were significantly increased at 28 days. There was, however, an absence of histologic and serum chemistry changes. At the recovery necropsy, only female absolute and liver to body weight ratios were significantly increased. The authors considered these liver findings as adaptive rather than an adverse effect. The authors reported a NOAEL of 1,000 mg/kg

- day. In an older subacute study (Zeller and Kirsch 1969 as cit ed by U.S. EPA, 2001) an HBCD product was administered to Sprague

- Dawley rat (10/sex/group) at doses of 0, 1, 2.5, and 5% of the diet for 28 days. Doses were calculated to be 0, 940, 2410, 4820 mg/kg

- day. No more details on study methodology were available . No clinical signs were observed at the 1% dose65

level and no significant change in mean body weight was observed at the 1 and 2.5% dose levels. Mean liver weight (both absolute and relative) was increased in all dose groups, but no microscopic pathology w as detected. Thyroid hyperplasia was observed in some animals at all doses in addition to slight numerical development of the follicles and ripening follicles in the ovaries at the high dose. No gross or microscopic changes were detected in any other organ nor were clinical chemistry changes detected. The authors attributed the increased liver weight to hyperactivity as a result of increased thyroid activity and concluded that the increased liver weights were not pathologic. They identified a NOAEL or 940 m g/kg

- day. In a subchronic study, Chengelis (2001; 2002) administered HBCD by oral gavage in corn oil daily to Crl:CD(SD)IGS BR rats (15/sex/dose) at dose levels of 0, 100, 300, or 1,000 mg/kg

- day for 90 days. At the end of 90 days, 10 rats/sex/dose were n ecropsied, while the remaining rats were untreated for a 28

- day recovery period prior to necropsy. Animals were observed twice daily for mortality and morbidity, while body weight and food consumption were measured weekly. Blood was collected at various po ints (Week 3, 13, and 17) for hematology, serum chemistry, and hormone (T3, T4, thyroid

- stimulating hormone (TSH)) measurements. Urine was collected prior to necropsy for urinalysis. Ocular examinations were carried out prior to study initiation and during Week 12 and 15. Functional Observational Battery and locomotor activity evaluations were carried out on 5 animals/sex/dose prior to study initiation, during the last week of HBCD administration (Week 13), and during the recovery period. Reproductive endpo ints (such as estrous cyclicity and sperm motility/viability, morphology, and count) were measured along with standard organ and tissue assessments. There were no treatment

- related effects on mortality and no changes in body weight, food consumption, Func tional Observational Battery and locomotor activity, hematology, ocular lesions, or reproductive parameters were observed. Significant effects on clinical chemistry, hormone levels, organ weights, and histology findings were observed at Week 13 but not aft er66

the recovery period (Week 17). At Week 13, liver weight was increased at 100 mg/kg

- day. Clinical chemistry changes at Week 13 included an increase in albumin (all dose levels for males), total protein (all dose levels for females and 1,000 mg/kg

- day for males), globulin (300 and 1,000 mg/kg

- day for females), and chloride (all doses for both sexes). Additionally, gamma glutamyltransferase levels were significantly increased in the 1,000 mg/kg

- day dose groups. All of these increased serum chemistry measure s were above or at the high end of historical control ranges. Significantly decreased alkaline phosphatase was also noted in all female dose groups at Week 13 (but not Week 17). Other significant serum chemistry findings were either not dose dependent or t ransient changes that were not significant at Week 13 or 17. T4 levels were significantly decreased at Week 13 in all male dose groups and the 300 and 1,000 mg/kg

- day female dose groups, but these decreases were not significant after the recovery period (W eek 17). TSH was significantly elevated in all male dose groups and all female dose groups at Week 13, but there were no significant findings after the recovery period (Week 17). There were no significant effects on T3. No gross lesions were reported, how ever, HBCD

- related histopathological changes were identified in the liver and thyroid at Week 13. Minimal hepatocellular vacuolation occurred in 10% of control males and roughly 50% of males in all dose groups. Mild and moderate hepatocellular vacuolation was observed in females in the 300 and 1,000 mg/kg

- day dose groups. Minimal to mild hepatocellular hypertrophy was also observed in high dose females. An increased incidence of minimal thyroid follicular cell hypertrophy was observed in males at 300 and 1, 000 mg/kg

- day (reported incidences were 1/10, 1/10, 5/10, and 7/10 for controls, 100, 300, and 1,000 mg/kg

- day groups, respectively) and females at 300 and 1,000 mg/kg

- day (reported incidences were 4/10 and 3/10 for 300 and 1,000 mg/kg

- day groups, respecti vely). Additionally, mild thyroid follicular hypertrophy was observed in 4/10 females in the 1,000 mg/kg

- day group. The only histopathological finding at Week 17 was minimal thyroid follicular hypertrophy67

observed in 3/5 females in the 1,000 mg/kg

- day grou p (versus 0/0, 0/0, and 0/0 in the control group, 100, and 300 mg/kg

- day groups, respectively). Histopathological changes were accompanied by increased liver weights. Both absolute and relative liver weights were significantly increased in all male and fem ale dose groups at Week 13, but there were no significant, dose

- related increases at Week 17. Increased prostate weight was also reported in the 1,000 mg/kg

- day group, but this increase did not persist into the recovery period and there were no observed ef

fects on sperm. The authors concluded that these treatment

- related changes were mild, reversible, and adaptive. They identified a NOAEL of 1,000 mg/kg

- day. Decreased liver weight accompanied by histopathological changes, however, can be considered an adve rse effect. Therefore, EPA identified a LOAEL of 100 mg/kg

- day based on these changes. In an older subchronic study (Zeller and Kirsch 1970 as cited by U.S. EPA, 2001) an HBCD product was administered to Sprague

- Dawley rat (10/sex/group) at doses of 0, 0.1 6, 0.32, 0.64, and 1.28% of the diet for 90 days. Doses were calculated to be 0, 120, 240, 470, and 950 mg/kg

day. Doses up to 0.64% produced no adverse clinical signs, no change in body weight, and no change in clinical chemistry results. An increase in r elative liver weight was observed and was accompanied by fatty accumulation. The pathology report concluded that although fat was visible microscopically in treated rats, the change was not accompanied by any pathology, and therefore could not be defined a s “fatty liver”. No histological changes were found in any other organ. The authors concluded that the increased liver weight and the fat deposits, both of which were largely reversible when administration of HBCD was stopped, were the result of a temporar y increase in the activity of the liver. They identified a NOAEL of 950 mg/kg

- day.68

3.4.2 Carcinogenicity No studies evaluating the carcinogenic potential of HBCD in humans or animals were found in the available literature. One non

- guideline study (Kurokawa et al., 1984) was cited in the U.S. EPA ’s Flame Retardant Alternatives for Hexabromocyclododecane (HBCD) : Final Report (U.S. EPA, 2014 ), but this study was not adequate to draw conclusions regarding carcinogenicity.

3.5. DEVELOPMENTAL AND REPRODUCTIVE TOXICITY van der Ven et al. (2009) In a one

- generation study that included additional immunological, endocrine and neurodevelopmental endpoints, van der Ven (2009) exposed Wistar rats (10/sex/dose) to a composite mixture of technical

- grade HBCD in the diet at concentrations resulting in dos es of 0.1, 0.3, 1.0, 3.0, 10, 30, or 100 mg/kg

- day. This study followed OECD Guidelines 415 and 407 (OECD, 1983; OECD, 1995). The isomeric composition of the HBCD was 10.3% α, 8.7% β, and 81.0% γ. The liver concentrations of HBCDs in F1 control rats, howev er, indicated exposure of F1 control rats to HBCD. For example, in the male F1 livers the concentration of HBCD in the control group were comparable to those in the three lowest dose groups. The authors stated that there were traces of tetra

- and pentabrom ocyclododecanes in the mixture, and hypothesized that brominated dioxins and furans were not likely to be present due to the aliphatic structure of HBCD and the lack of aryl hydrocarbon receptor (AhR) activation in a previous 28

- day study that used the sam e composite mixture. HBCD was dissolved in acetone, mixed with corn oil (allowing the acetone to evaporate), and then mixed with feed particles that were subsequently69

pelletized. The use of corn oil as a carrier for HBCD resulted in a higher

- than

- usual
con centration of oil (10% ratio of oil volume/feed weight) in the diet
. A carrier control group received a standard low

- fat diet (5% lipids) supplemented with corn oil and a negative control group received the standard low

- fat diet. Males were exposed fo r 70 days prior to mating and females were exposed for 14 days prior to mating. Exposure was continued for both sexes during a 2

- week mating period, and for females throughout gestation, parturition, and lactation. Food intake and body weight were measured throughout the study. Males were euthanized at the end of the mating period. Females were euthanized following weaning of their litters, and the number of uterine implantation sites was recorded. Reproductive variables that were assessed for the parenta l (P) generation included the number of successful matings per dose group, mean time to gestation, and length of gestation. At the time of birth, the number of live and dead pups, sex, anogenital distance, and gross pathology was noted for the pups. Litter s were not standardized, and pups continued to be exposed to HBCD through the dams’ milk and the pelletized diet. F1 body weight was measured on PND 4, 7, 14, and 21. Anogenital distance was also measured on Day 21. Early mortality, time of vaginal opening , and time of preputial separation were monitored throughout early lactation. On PND 21, two male and two female pups from each litter were euthanized, and their body weights and weights of reproductive organs were recorded. Sperm counts and morphology wer e assessed from the cauda epididymides of the males. The remaining F1 pups continued to be exposed to HBCD in the same manner as the P generation. Just prior to the end of the study, at approximately 8 weeks of age, the remaining F1 pups were assigned to separate groups for necropsy (5/sex/dose), immunological testing (4 males/dose), or neurobehavioral testing (6/sex/dose plus additional males and females from selected dose groups). Necropsy was conducted at 11 weeks of age and included gross and70

microsco pic pathology 3 , clinical chemistry 4 , bone analysis 5 , analysis of liver HBCD content, evaluation of apolar retinoid content of the liver, and assessment of serum hormone levels 6

. Hematological/immunological evaluation included an evaluation of standard red blood cell variables 7 , an analysis of subpopulations of immune cells and natural killer cell activity in the spleen, and an assessment of white cell counts and fractions in femoral bone marrow. Antibody titers were also measured in additional groups (4 mal es/dose) that were injected intraperitoneally with sheep red blood cells at Day 0 and Day 15. Titers were evaluated in blood on Day 0 (control), Day 7 (IgM analysis), and Day 21 (IgG analysis) post

- injection. T

- tests were used to compare the results obtain ed for the two control diets, with the acknowledgement that the small sample sizes limit the statistical power to discriminate between the two groups. Dose

- response data were analyzed using the benchmark dose approach with PROAST software. The authors used the likelihood ratio test for model selection and referred to the modeling methodology outlined by Slob (2002). The benchmark dose (BMD, or critical effect dose, CED) was calculated at a predefined benchmark response (BMR or critical effect size, CES) dep ending on the endpoint. CES is the size of the effect that is considered to be close to the border between adverse and nonadverse, which was defined for each individual parameter on the basis of its known physiological homeostatic variation and of its path ophysiology, including irreversibility or adverse follow

- up effects. Most effects were assessed at a CES of 10%, while developmental parameters were assessed at a CES of 5%. Some immune and liver parameters

3 At necropsy, body weight and the weights of pituitary, thy roid, adrenals, popliteal lymph nodes, spleen, thymus, heart, liver, kidneys, brain, and reproductive organs were recorded. Samples of these tissues were also preserved for histological evaluation. 4 Clinical chemistry included evaluation of serum albumin, alkaline phosphatase, alanine aminotransferase, total cholesterol, creatinine, glucose, total protein, and urea. 5 Length of tibia, trabecular bone mineral density, cortical bone mineral density, total bone mineral density, total mineral content, total area, cortical area, and cortical thickness. 6 Serum thyroid hormone concentrations; activity of 17 alpha

- hydroxylase

- 17,20

- desmolase (CYP17) in adrenals and of aromatase (CYP19) in ovaries. 7 Red cell count, hemoglobin concentration and associated variables , hematocrit, mean cell volume, reticulocyte percentage, and platelet count.71

were assessed at a CES of 20%. In the case of a m ore than 10

- fold difference between BMD and BMDL, the calculated BMDLs were considered insufficiently informative. The authors did not observe any parental mortality. Body weight was reduced in parental males during the first 3 weeks of exposure and was u naffected in parental females. Food intake, however, was reduced during week 3 of gestation in parental females. Growth and development were affected by HBCD exposure in the F1 pups. There was a decrease in body weight in F1 males (7

– 36%) and females (10

– 20%) beginning on PND 4 and persisting throughout the duration of the study at the highest dose administered. The authors noted that these observations may be related to decreased food intake, which was significant at some stages. In males, testes weigh t demonstrated a dose

- dependent decrease, kidney and thymus weights were decreased in a dose

- dependent manner for both males and females, and adrenals, prostate, heart, and brain weights were decreased in F1 males. No histopathological changes, however, we re observed in any of these organs. Additionally, there was no effect of HBCD exposure on serum chemistry in the F1 generation. Increased anogenital distance was observed in males at 100 mg/kg on PND 4, but not on Days 7 or 21. There was no effect on prep utial separation. The time to vaginal opening was delayed in females at the 100 mg/kg dose. There were no effects of HBCD exposure on thyroid hormones (T3 and T4) in either the parental or F1 animals. There were no effects on thyroid weight or thyroid path ology in the F1 animals (parents were not examined). There was no significant treatment

- related effect on CYP17 activity in the adrenals or for CYP19 activity in the ovary. There was, however, a strong correlation between CYP19 and internal concentration
o f the γ

- HBCD when analyzed on group averages. The authors observed several immune system effects. An increase in IgG antibody titers in response to immunization with sheep red blood cells was observed in 8

- week

- old F1 males.72

White blood cell count in bone marrow was increased. There was a decrease of lymphocyte fraction and increase of the polymorph neutrophil population. There were decreases in the sum of apolar retinoids in the liver (21.4 and 24.1% in males and females, respectively, at 100 mg/kg

- day) and in the sum of apolar retinoids per mg/g of liver (18.5 and 17.2% in males and females, respectively, at 100 mg/kg

- day). In F1 females, there was a dose

- related decrease in bone mineral density (20.4% at 100 mg/kg

- day). Other small changes in bone varia

bles were also observed in the F1 generation, including: tibia length (4.2 and 1.2% in males and females, respectively, at 100 mg/kg

- day), tibia total mineral content (10.7% in males at 100 mg/kg

- day), tibia total area (8.9% in males at 100 mg/kg

- day), tib ia cortical area (10.5% in males at 100 mg/kg

- day), and tibia cortical thickness (5.2% in males at 100 mg/kg

day). The authors observed that the two control diets demonstrated statistically significant differences for various study endpoints in F1 animals . For examples, mortality of F1 pups during lactation was higher in the carrier control group (i.e., standard diet supplemented with corn oil) than in animals fed the standard low

- fat diet. Also, F1 male rats in the carrier oil group had a statistically si gnificant decrease in liver weight, adrenal weight, white blood cell count, and fraction of splenic natural killer cells. F1 male rats in the carrier oil group had a statistically significant increase in thymus weight and fraction of splenic CD

- 8 cells. Fe male rats in the carrier oil group had decreased plasma cholesterol, length of tibia, and adrenal weight. These females also had increased trabecular bone mineral density. Based on comparison of control groups, the study authors noted potential interaction s between carrier oil and HBCD exposure for food intake and body weight during gestation; body weights of F1 pups during lactation (except male pups on PND 7); liver concentration of apolar retinoids in males; decreased adrenal weight in males; increased t hymus weight in males; and length and trabecular bone density of the tibia73

in females. The authors noted, however, that the endpoints showing effects at low doses did not appear to be affected by the carrier oil in the diet. The authors reported BMDLs for several developmental endpoints in this study. The most sensitive endpoints with valid BMD/BMDL ratios for female rats were decreased bone mineral density with a BMDL of 0.056 mg/kg

- day (BMD of 0.18 mg/kg

- day) at a BMR of 10% and decreased concentrations o f apolar retinoids in the liver with a BMDL of 1.3 mg/kg

- day (BMD = 5.1 mg/kg

- day) at a BMR of 10%. The most sensitive endpoint with a valid BMD/BMDL ratio for male rats was an increased IgG response to sheep red blood cells with a BMDL of 0.46 mg/kg

- day ( BMD = 1.45 mg/kg

- day) at a BMR of 20% . The authors also reported recalculated BMDLs expressed as observed responses against internal liver concentrations per animal for effects detected at final necropsy (refer to original study for more detail). There wer e no significant effects of HBCD exposure on any measure of reproduction, including mating success, time to gestation, duration of gestation, number of implantation sites, pup mortality (at birth and throughout lactation), or sex ratios within a litter. Th erefore, a BMDL for reproductive toxicity could not be derived for this study.
Saegusa et al. (2009) In a study designed to examine the developmental effects of HBCD, Saegusa et al. (2009) exposed pregnant Sprague

- Dawley rats (10/sex/dose) to HBCD from gestation day 10 until PND 20 at dietary concentrations of 0, 100, 1,000, or 10,000 ppm in a soy

- free di et. Measured HBCD intake by dams as reported by the authors were 8.1 ± 1.1, 80.7 ± 15.9, or 803.2 ± 101.9 mg/kg

day for gestation day 10 to 20; 14.3 ± 2.3, 138.7 ± 29.4, or 1404.8 ± 184.5mg/kg

- day for PNDs 1 to 9; and 21.3 ± 5.7, 212.9 ± 77.3, or 2,231.3 ± 445.0 mg/kg

- day for PNDs 9 to 20. Time

weighted average (TWA) doses were calculated by multiplying the HBCD intake (mg/kg

- day) by the number of inclusive days of exposure for each time point. The sum of each time point for74

an individual dietary concentrat ion (100, 1,000, and 10,000 ppm) was divided by the total number of inclusive days (33 days) of exposure. TWA doses were calculated to be 0, 14.8, 146.3, and 1,504.8 mg/kg

- day (U.S. EPA, 2014 ). On PND 1, the number, weights, and anogenital distance of neonates were recorded. On PND 2, litters were culled randomly to eight offspring per dam (comprised of four males and four females). On PND 20, dosing was terminated and all dams were killed. Twent y male and twenty female offspring per group were subjected to prepubertal necropsy and histopathological and thyroid

- related hormone assessment. The remaining F1 males and females were allocated to four rats per cage and given a regular diet and water ad libitum. These animals were maintained for adult examination when they were 11 weeks old. The effects of HBCD on neuronal migration were observed on PNW 11 by measuring the variability in the distribution of neurons located within and lateral to the pyrami dal cell layer of the hippocampal CA1 region. The authors also evaluated the effect of HBCD on oligodendroglial development by measuring areas immunoreactive for CNPases and the number of CNPase

- positive oligodendrocytes surrounding myelinated axons in the cerebral cortical area. Prepubertal necropsies, in which the organs were removed and weighed, were conducted on PND 20 and a histopathological assessment was performed. Dams were subjected to an implantation count. Female offspring were monitored daily fo r vaginal opening from PND 26 and all male pups were examined for preputial separation from PND 34. The age at the onset of puberty was recorded for the offspring allocated for adult examination. Estrous cycles of females were examined by daily microscopic observation of vaginal smears from PNW 8 to PNW 11. At PNW 11, offspring were sacrificed and tissues subjected to histopathological assessment and thyroid

- related hormone measurement. Maternal body weight during gestation and lactation and the duration o f pregnancy were not affected by HBCD exposure in this study. In parental rats, the relative weight of the thyroid7 5

and histopathological changes, including diffuse follicular cell hypertrophy, were significantly increased in females in the 10,000 ppm dose group. Changes were also observed in the other treatment groups, but these changes were not statistically significant. No changes in the number of implantation sites, number of live offspring, or male ratio were observed. Multiple changes were identified in the F1 generation. Selected findings are summarized in Table 3

- 3. No changes in body weight, anogenital distance, or onset of puberty in males or females were observed. At PND 20, statistically significant increases in relative liver weight and diffuse vacuolar degeneration of liver cells were observed in both males and females at 10,000 ppm. T3 levels in males were significantly decreased and TSH levels significantly elevated at the highest dose, while T4 levels were unaltered at any dose. At PNW 11, th e authors observed a significant dose

- dependent increase in relative thyroid weights in male rats at 1,000 and 10,000 ppm. Relative liver weight was elevated in the 100 ppm dose group males, but changes were not observed at higher doses or in females at an y dose level. Relative epididymides weight was also elevated in 100 ppm dose group males, but changes were not observed at higher doses. Additionally, T3 levels were significantly decreased at 1,000 and 10,000 ppm compared to controls, but a dose

- respons e relationship was unclear. Serum T4 levels and TSH concentrations were unchanged at PNW 11. Additionally, a statistically significant, dose

- dependent decrease in the number of CNPase

- positive oligodendrocytes in the cingulate deep cortex was observed in males at the highest dose (10,000 ppm) at PNW 11 . The authors reported no significant difference in the distribution of hippocampal CA1 neurons. The corpus callosum area was also unaffected at all doses. EPA identified a maternal LOAEL of 10,000 ppm (appro ximately 1,504.8 mg/kg

- day) based on increased incidence of thyroid follicular cell hypertrophy. EPA identified a developmental LOAEL of 1,000 ppm (approximately 146.3 mg/kg

- day) based on increased76

relative thyroid weight and decreased T3 levels in F1 male rats at PNW 11. Changes in reproductive endpoints (e.g., the number of implantation sites, live offspring, sex ratio) were not observed. Therefore, a LOAEL for reproductive toxicity could not be determined for this study.77

Table 3 - 3. Summary of effects in parental and F1 rats after dietary, gestational, lactational, and postnatal exposure to HBCD HBCD (ppm in diet) Males c Females c Generation 0 100 1,000 10,000 0 100 1,000 10,000 F0 Thyroid Relative weight (mg/100 g BW) ND ND ND ND 5.73 6.75 6.30 7.47 * Histopathology: diffuse follicular cell hypertrophy (±/+/++/+++) a ND ND ND ND 3/10 (0/3/0/0) b 5/10

(2/3/0/0) 6/10

(1/3/2/0) 9/10 #

(0/3/4/2) §§ F1 Relative organ weights, PND 20 Liver (g/100 g BW) 3.68 3.82 3.98 4.66 * 3.77 3.83 4.01 4.83 * Relative organ weights, PNW 11 Liver (g/100 g BW) 3.45 3.81 ** 3.58 3.53 3.35 3.59 3.44 3.30 Thyroid (mg/100 g BW) 4.85 5.66 5.78 * 6.20 ** 8.20 6.84 7.35 7.72 Epididymides (mg/100 g BW) 0.23 0.21 * 0.22 0.21 NA NA NA NA Thyroid

- related hormones PND 20 T 3 (ng/ml) 1.09 1.13 1.06 0.93 ** ND ND ND ND T 4 (μg/dl) 4.39 4.20 4.78 4.20 ND ND ND ND TSH (ng/ml) 5.40 6.66 6.07 7.00 * ND ND ND ND PNW 11 T 3 (ng/ml) 0.96 0.93 0.88 * 0.89 ** ND ND ND ND T 4 (μg/dl) 4.77 4.84 5.21 5.20 ND ND ND ND TSH (ng/ml) 4.74 5.81 5.36 4.96 ND ND ND ND Histopathology PND 20 Liver: Vacuolar degeneration, liver cells, diffuse (+/++) a 0/10 0/10 0/10 6/10 * (6/0) 0/10 0/10 0/10 6/10 * (0/6) §§ PNW 11 Adrenal: Vacuolar degeneration, diffuse, cortical cells (+/++) a 0/10 0/10 0/10 4/10 * (2/2) § ND ND ND ND78

Abbreviations: BW, body weight; PND, postnatal day; ppm, parts per million; PNW, postnatal week; ND, no data; NA, not applicable a Grade of change: (±) minimal; (+) slight; (++) moderate; (+++) severe b Number of animals with each grade c n=10 rats/sex/group * Significantly different from the controls by Dunnett’s test or Dunnett

- type rank

- sum test (p<0.05) ** Significantly different from the controls by Dunnett’s test or Dunnett

- type rank

- sum test (p<0.01) # Significantly different from the controls by Fisher’s exact probability test (p<0.05) § Significantly different from the controls by Mann

- Whitney’s U

- test (p<0.05) §§ Significantly different from the controls by Mann

- Whitney’s U

- test (p<0.01) Source: Saegusa et al. (2009)
Ema et al. (2008) In a two

- generation reproductive toxicity study, Ema et al. (2008) administered HBCD to Crl:CD(SD) rats conducted according to OECD Guideline 416 (OECD, 2001), U.S. EPA Guidelines (U.S. EPA, 1991; 1996), and Good Laboratory Practices. Groups of male and female rats (24/sex/dose) were fed HBCD (as a mixture of α

- HBCD, β

- HBCD, and γ

- HBCD with proportions of 8.5, 7.9, and 83.7%, respectively) in the diet at concentrations of 0, 150, 1,500, or 15,000 ppm from 10 weeks prior to mating throu gh mating, gestation, and lactation. Mean calculated daily doses throughout the study were reported by the authors as: 10.2, 101, and 1,008 mg/kg for F0 males; 14.0, 141, and 1,363 mg/kg for F0 females; 11.4, 115, and 1,142 mg/kg for F1 males; and 14.3, 13 8, and 1,363 mg/kg for F1 females. Reproductive and developmental milestones were monitored, including: number of implantation sites, fertility index, estrous cycles, pinna unfolding, incisor eruption, eye opening, surface righting reflex, negative geotax is reflex, mid

- air righting reflex, and anogenital distance. Twenty

- four male and 24 female F1 pups in each group were selected as parents on PNDs 21

– 25 and the authors ensured equal body weights among groups. Litter sizes were adjusted to 4/sex/dose on P ND 4. Culled pups received complete external and internal necropsies, however, tissues from these pups were not collected. At 4 weeks of age, 10 F1 rats/sex/dose were tested for spontaneous locomotor activity. At 6 weeks of age, 10 F1 rats/sex/dose were te sted in a79

water
- filled maze. Hematological variables and serum chemistries (10/sex/dose), serum hormone levels (8/sex/dose), and sperm parameters (all males) were evaluated for randomly selected adult males and females of the F0 and F1 generations. The num ber of primordial follicles was counted in 10 F1 females from each group. All parental animals received complete necropsy, including measurement of organ weights and gross evaluation of tissues. For all dose groups, organs showing gross alteration were exa mined microscopically. Comprehensive microscopic evaluations were conducted on control and high

- dose parental animals. All F1 weanlings not selected to parent the next generation and all F2 weanlings were necropsied on PND 26. For F1 and F2 weanlings, live r sections from the control and high

- dose groups were evaluated microscopically and thyroids were evaluated microscopically for one male and one female from each dam. There was no significant treatment

- related mortality and no significant differences in clinical signs between control and HBCD

- treated parental animals of the F0 or F1 generations. There were no significant treatment

- related effects in any generation on es trous cyclicity, sperm counts and morphology, copulation index, fertility index, gestation index, delivery index, gestation length, number of pups delivered, number of litters, or sex ratios. Statistically significant increases in mean absolute and relativ e liver weights were observed in F0 adult males and F1 weanling males and females exposed to ≥1,500 ppm, and in F0 adult females and F1 male and female adults exposed to 15,000 ppm. The authors concluded that these liver weight changes were an adaptive rat her than a toxic response. Mean absolute and relative thyroid weights were significantly increased in F0 and F1 adults exposed to HBCD at 15,000 ppm. A significant increase in relative brain weight and relative pituitary gland weight was observed in F1 adu lts at 150 ppm, but there were no significant effects observed at higher doses. Also, a significant decrease in absolute brain weight was observed in male and female weanlings and adults in the 15,000 ppm dose group. While a significant increase in relativ e testis weight was80

observed in F1 male weanlings at 150, 1,500, and 15,000 ppm, there were no significant effects on either absolute or relative testis weights in F1 adults at any dose level. Additionally, these effects were not observed in F2 male weanli ngs. The authors concluded that these organ weight changes were not toxicologically significant because they were inconsistent across age, sex, and generation. A significant increase in the number of litters lost at the highest dose was observed in the F1 parental generation; incidences (number lost/number pregnant) were 1/23, 1/23, 0/21, and 8/21. A statistically significant decrease in follicle

- stimulating hormone (FSH) was observed in F0 males at 1,500 ppm, whereas, an increase in FSH in F0 females was reported at 15,000 ppm (data on sex hormone levels not provided by study authors). An increase in 5α

dihydrotestosterone (DHT) was observed in F1 males at 1,500 ppm. A significant decrease in the number of primordial follicles in the ovary was observed amo ng F1 females treated with HBCD at 1,500 and 15,000 ppm compared to controls. A decrease in primordial follicle count is normally considered a biomarker of an adverse reproductive effect because no recovery is possible, indicating that HBCD is potentially a reproductive toxicant. Significant decreases in serum T4 were observed in F0 males and females at 15,000 ppm, but changes were not observed in F1 males or females. Significantly increased TSH was observed in all HBCD

- treated F0 females and in mid

- and hi gh

- dose F1 females. Hypertrophy of thyroid follicular cells was not observed upon histological examination of tissues from F0 and F1 rats. Increased TSH levels at 150 ppm in F0 females, however, were not accompanied by histopathological changes in the thyr oid or decreased T4 levels. An increase in the incidence of animals with decreased thyroid follicle size was observed in both sexes and generations among rats exposed to 1,500 and 15,000 ppm (Table 3

- 4). Body weight was significantly decreased in F2 male a nd female pups in the highest dose group on PND 7, 14, and 21. These decreased body weights were associated with decreased food81

consumption. Relative liver weight was increased in the highest dose group for both F2 males and females. The study authors note d that changes in absolute and relative organ weights (brain, pituitary, thymus, kidney, spleen, adrenals, testes, epididymis, seminal vesicle, prostate, ovaries, and uterus) were observed in adults and weanlings of both sexes and generations, but were not likely of toxicological significance because the observed changes likely reflected the decrease in body weight observed primarily at the highest dose. The pup viability index was significantly reduced by on Day 4 and Day 21 of lactation among F2 offspring at the highest dose in comparison with controls (Table 3

- 5). There were no effects on sexual development indicated by anogenital distance, vaginal opening, or preputial separation among F1 or F2 generations. The percentage of pups with completed eye openi ng on PND 14 was significantly decreased compared to controls in F2 females at 1,500 ppm and in F2 males and females at 15,000 ppm. F2 females exposed to 15,000 ppm HBCD completed the mid

- air righting reflex (76.9%) than control F2 females ( 100%). These fi ndings were not consistent over generations or sexes and were not considered treatment related. No other effects of HBCD exposure on the development of reflexes were observed in either F1 or F2 progeny. EPA identified a maternal LOAEL of 150 ppm (approxima tely 14 mg/kg

- day) based on increased TSH in F0 females. A reproductive LOAEL of 1,500 ppm (approximately 138 mg/kg

day) was identified based on the decrease in the number of primordial follicles in the ovary observed among F1 females treated with HBCD. A developmental LOAEL of 15,000 ppm (approximately 1,142 mg/kg

- day for males and approximately 1,363 mg/kg

- day for females) was identified based on increased pup mortality during lactation in the F2 generation.82

Table 3 - 4. Incidence of decreased thyroid foll icle size in rats exposed to HBCD in the diet for two generations

Generation/sex Dietary HBCD concentration (ppm) 0 150 1,500 15,000 F0 male 0/24 0/24 6/24 a 20/23 b F0 female 0/24 0/24 5/25 a 11/23 b F1 male 0/24 0/24 2/22 11/24 b F1 female 0/24 1/24 5/24 a 13/24 b

a Significantly different from control, p < 0.05. b Significantly different from control, p < 0.01. Source: Ema et al. (2008)
Table 3 - 5. Summary of pup viability in F2 rats after dietary, gestational, lactational, and postnatal exposure to HBCD HBCD (ppm in diet) F2 Generation 0 150 1500 15,000 Viability Index During Lactation (%) Day 0 98.6 97.7 96.0 97.8 Day 4 86.9 87.3 92.1 68.4 * Day 21 85.0 89.6 71.3 49.7 ** Abbreviations: ppm, parts per million a n=24/sex/group except for b Values are given as the number of animals that showed abnormal findings * Significantly different from the controls (p<0.05) ** Significantly different from the controls (p<0.01) Source: Ema et al. (2008)
Murai et al. (1985) Female Wistar rats were fed HBCD in the diet at concentrations of 0, 0.01, 0.1, or 1% throughout gestation (Days 0

– 20) (Murai et al., 1985; Japanese with English abstract). Dams in the high

- dose group had a statistically significant decrease (8.4%) in f ood consumption and increase in liver weight (13%) in comparison with controls. There were no treatment

- related effects on maternal or fetal body weight. There were no effects on the number of implants; number of resorbed, dead, or live fetuses; body weigh t of live fetuses; or incidence of external or83

visceral abnormalities. A few skeletal variations were present but were also observed in controls and not considered significant. There were no effects on weaning or survival. The European Commission (European Commission, 2008) used the study’s data to calculate the doses to be 0, 7.5, 75, and 750 mg/kg

- day (based on the assumption of a mean animal weight of 200g and food consumption of 15 g/day). They concluded that the offspring NOAEL was 750 mg/kg

- day and th e maternal LOAEL was 750 mg/kg

- day based on 13% liver weight increase in the high dose group.
Eriksson et al. (2006) The potential effects of HBCD on behavior, learning, and memory were assessed in male mice that were exposed on PND 10 (Eriksson et al., 2006). Neonatal NMRI mice of both sexes were given a single oral dose of HBCD (mixture of, α

- , β

- , and γ

- diastereomers ) dissolved in a fat emulsion at 0, 0.9, or 13.5 mg/kg

- day on PND 10. Mice received no further exposure to HBCD. Only male mice were used for neurotoxicological observation. At 3 months, male mice were tested for effects on spontaneous behavior and for performance in a swim maze. To assess spontaneous motor behavior, locomotion, rearing, and total activity were recorded for each of 10 mice randomly selected from 3 to 4 different litters from each dose group. Each mouse was tested only once for each endpoint, but each endpoint was evaluated over three consecutive 20

- minute intervals. The authors defined “habituation” as a decrease in the afor ementioned variables in response to the diminished novelty of the test chamber with time. For the swim maze, which is a test of spatial learning ability and relearning ability, a total of 12

– 17 mice were randomly selected from 3 to 4 litters for each dos e group. Each mouse was given 5 trials in the swim maze on 5 consecutive days; the objective of the test was to remember and locate a submerged platform after being placed on the platform for 30 seconds and moved to another location.84

There were no clinical signs of dysfunction in the treated mice throughout the study, nor were any significant body weight changes observed. Control mice displayed a pattern of habituation (i.e., greatest activity during the first 20

- minute interval and decreased activity with each successive 20

- minute interval). However, mice exposed to the highest dose of HBCD (13.5 mg/kg) displayed a statistically significant reduction in habituation. For each indicator of activity (locomotion, rearing, and total activity), mice exposed to HB CD at 13.5 mg/kg exhibited statistically significantly less activity than controls during the first 20

- minute interval, equivalent to controls during the second 20

- minute interval, and greater than controls during the last 20

- minute interval. Mice exposed to 0.9 mg/kg HBCD had statistically significantly reduced mean locomotor activity and rearing than controls during the first 20

- minute interval, but did not differ from controls in subsequent intervals. In the swim maze test, mice exposed to the highest do se of HBCD had a significantly longer search time on Day 4 of the test than either controls or mice given 0.9 mg/kg HBCD. Furthermore, on Day 5 of the test, when the location of the platform was changed, mice given 13.5 mg/kg HBCD took significantly longer than controls to find the platform at the new location. Eriksson et al. (2006) concluded that exposure on PND 10 affected spontaneous motor behavior, learning, and memory in adult mice in a dose

- dependent manner. The authors identified the lowest exposur e level, 0.9 mg/kg, as the LOAEL based on significantly reduced mean locomotor activity compared with controls during the first 20

- minute interval of testing. EPA, however, identified a LOAEL of 13.5 mg/kg

- day based on decreased habituation, locomotion, an d rearing in all intervals. This study was not conducted according to current guidelines (U.S. EPA, 1998) and Good Laboratory Practices, therefore EPA reserves judgment on the significance of these findings. The authors used too few dose groups and the beh avioral alterations were induced at doses that did not produce clinical signs or affect weight gain.85

Additionally, effects due to litter size were not considered. However, this study did demonstrate good repeatability for control values and for relevant ac tive substances tested several times.

3.6. GENOTOXICITY A limited number of studies investigated the genotoxicity of HBCD. These studies, however, indicate that HBCD is not likely to be genotoxic. These studies are summarized in Table 3

- 6. The majority of these studies were standard Ames tests for detecting mutagenic potential in Salmonella typhimurium ( S. typhimurium ). Most Ames tests conducted with HBCD yielded negative results (Industrial Bio

- Test Labs, 1990; Litton Bionetics, Inc., 1990; SRI Research Institute, 1990; Zeiger et al., 1987; Huntingdon Research Center, 1978; Pharmakologisches Institute, 1978). Two Ames tests showed positive, dose

- dependent results for strain TA1535; one for TA1535 only using a liquid residue of HBCD in DMSO (Industrial Bi oTest Laboratories, 1990); and one for strains TA1535 and TA100 using an unidentified mixture characterized only as HBCD bottoms in acetone (Ethyl Corporation, 1985a). Both of these strains detect reversions by base pair substitution. However, the Ames tes ts in the strains that were positive in these two studies (TA1535 and TA100) were negative in the other studies cited above. In mammalian systems, a reverse mutation assay with Chinese hamster ovary (CHO) Sp5 and SPD8 cell lines exposed to HBCD yielded pos itive results (Helleday et al., 1999). A test of unscheduled DNA synthesis with rat hepatocytes exposed to HBCD bottoms was also positive with a dose

- response relationship (Ethyl Corporation, 1985b). The reverse mutation assay in CHO cells (Helleday et al. , 1999) and the unscheduled DNA synthesis assay in F344 hepatocytes (Ethyl Corporation, 1985b) were not repeated by any other group. Several assays performed to determine the genotoxicity of HBCD were negative even when testing at cytotoxic concentrations , including: one in yeast (Litton Bionetics Inc., 1990),86

one detecting chromosomal aberrations in human peripheral lymphocytes in vitro (Microbiological Associates Inc., 1996), and one in vivo mouse micronucleus test following intraperitoneal injection of HBCD (BASF, 2000).
Table 3 - 6. Summary of genotoxicity studies of HBCD Test/species/strain/ route Test doses

(per plate) a Results b Notes Reference

– S9 +S9 Eukaryotic systems, in vitro S. typhimurium TA98, TA100, TA1537 3,000 μg

in DMSO

– – Doses ≥1,000 μg were partially insoluble. Pharmako

logisches Institute, 1978 S. typhimurium TA98, TA100, TA1535, TA1537, TA1538 250 μg

(Firemaster, FM

100, Lot 53, white powder) in DMSO – – Doses ≥250 μg were insoluble. Industrial BioTest Laboratories, 1990 1,000 μg (FM

100, Lot 3322, liquid residue) in DMSO – + (TA1535 only) Positive in TA1535 at highest dose only; lower doses showed positive trend with dose. S. typhimurium TA98, TA100, TA1535 10,000 μg

in DMSO

– – Insoluble at 10,000 μg. Huntingdon Research Center, 1978 S. typhimurium TA98, TA100, TA1535, TA1537 10,000 μg

in DMSO

– – Z ei ger et al., 1987 S. typhimurium TA98, TA100, TA1535, TA1537, TA1538 5,000 μg

in DMSO

_ _ SRI Research Institute, 1990 S. typhimurium TA98, TA100, TA1535, TA1537 50 μg

(HBCD bottoms ) in acetone +

(TA1535 and 100 only) + (TA100 only) No cytotoxicity observed. Dose

response only in TA1535 (

– S9) ≥100 μg/plate. TA100 positive at highest dose only (5,000 ug/plate). All doses had a black precipitate thought to be carbon
. Ethyl Corporation, 1985a
S. typhimurium TA98, TA100, TA1535, TA1537, TA1538 50 μg

– – Litton Bionetics, 1990 Prokaryotic non - mammalian systems, in vitro Saccharomyces cerevisiae D4 50 μg

– – Litton Bionetics, 1990 Mammalian systems, in vitro Chromosomal aberration test In human peripheral blood lymphocytes 750 μg/m l (

- S9)

and 250 μg/m (+S9) in DMSO

– (T)

– (T) Doses 750
– 2,500 were partially insoluble, and fully insoluble >2,500 μg/m l . Repeated test for two harvest time points: 20 hr (

- S9) or 4 hr (+S9) incubations, and 20 or 44 hr incubations (

- S9 and +S9). Microbiological Associates, 199687

Table 3 - 6. Summary of genotoxicity studies of HBCD Test/species/strain/ route Test doses

(per plate) a Results b Notes Reference

– S9 +S9 Unscheduled DNA Synthesis rat/F344 male/primary hepatocytes 5
– 1,000 μg/well

in acetone (HBCD bottoms) + NA Five highest doses (from 5 μg/well) showed an increased response with dose over solvent control, but only four highest were statistically significant (χ 2 ). Highest dose (1,000 μg/well) was cytotoxic. Ethyl Corporation, 1985b

Reversion assay CHO/V79/Sp5 and SPD8 Intragenic recombination at hprt loc us in Sp5 (non

homologous recombination) and SPD8 (homologous recombination) duplication cell lines 3

– 20 μg/m l

in DMSO

+ NA A statistically significant increase in reversion frequency was observed in both assays in the highest dose group as determined by linear regression analysis. Helleday et al., 1999

Mammalian systems, in vivo Micronucleus test mouse/NMRI/intraperi toneal injection 2,000 mg/kg

in DMSO

– (T) NA BASF, 2000

a Lowest effective dose for positive results; highest dose tested for negative results. b + = positive, ± = equivocal or weakly positive,

– = negative, T = cytotoxicity, NA = not applicable, ND = no data.

3.7. OTHER CHRONIC EFFECTS Weiss et al. (2006) measured the concentration of polychlorinated biphenyls (PCBs), polybrominated diphenyl ethers (PBDEs), and HBCDs in the blood serum of Swedish fishermen’s wives (53 women with a median age of 62 years) and a ttempted to correlate these measurements with indicators of potential adverse effects on bone (i.e., bone mineral density and markers of bone metabolism). HBCD isomers, PBDEs, and PCBs were detected in the serum of these women; however, there were no corre lations between any of the bone health indicators and the identified chemicals in the serum. Kiciński et al (2012) assessed the relationship between PBDE exposure and neurobehavioral function using biomonitoring data from Flanders, Belgium. They concluded that PBDE exposure was associated with changes in the motor function and serum levels of thyroid hormones, but HBCD did not show significant associations with88

performance in the neurobehavioral tests. These authors conclude, however, that the potential fo r exposure misclassification in this study may have been likely due to a large number of observations for which PBDE levels were below the level of quantification. As a result, the authors had to create categories that were not biologically relevant in ord er to analyze the data.
3.8. POTENTIAL MODES OF ACTION Lilienthal et al. (2009) reported the results of the neurobehavioral assessment from the
one - generation study conducted by van der Ven et al. (2009). The general study design is described in the summary of van der Ven et al. (2009) in Section 3.5. Lilienthal et al. (2009) examined 110 day

- old F1 rats for haloperidol

- induced catalepsy to determine possible effects of HBCD on the dopaminergic system . One month after the catalepsy measurements, brain stem auditory evoked potentials (BAEPs) to broadband click and frequency

- specific tone stimuli were recorded in males and females to examine potential effects on auditory functions. The study authors concluded that the observed dose

- dependent decreases in latencies may be due to HBCD

- related effects on dopaminergic activity or to HBCD

- related induction of metabolizing liver enzymes resulting in enhanced metabolism of haloperidol. HBCD

- related effects on BAEP were observed in the low frequency range and only in male offspring. The study authors hypothesized that HBCD exerts a cochlear effect on males based on the results of the BAEPs. The authors noted that an alternative explanation could be HBCD

- induced changes in retinoids, which may be involved in develop ment of the inner ear. Too little information is available in this study to determine the significance of its findings
. 3.9. CONCLUSION Based on the available data (summary of key data is shown in Figure 3

- 1), EPA concludes that there is sufficient eviden ce to support a concern for moderately high to high toxicity from exposure to HBCD based on its potential developmental and reproductive effects
.89 Figure 3 - 1. Exposure - response array for developmental and reproductive toxicity studies of HBCD90

3.10. REFERENCES Abdallah, M; and Harrad , S
. (2011) Tetrabromobisphenol

- A, hexabromocyclododecane and its degradation products in UK human milk: relationship to external exposure. Environment International, 37: 443

– 448. Antignac, JP; Cariou, R; Maume, D; et al. (2008) Exposure assessment of fetus and new born to brominated flame retardants in France: preliminary data. Mol Nutr Food Res 52(2):258

– 265. Arita, R; Miyazaki, K; Mure, S. (1983) Metabolic test of HBCD. Test on chemical substances used in household items. Studies on pharmacodynamics of HBCD [unpu blished]. In: Toxicology summary: HBCD (HBCD), Albemarle, S.A. Department of Pharmacy, Hokkaido University Hospital, Japan. BASF. (1990) Report on the study of the acute oral toxicity of HBCD in the mouse with cover letter dated 031290. Submitted under T SCA Section 8D; EPA Document No. 86

- 900000383; NTIS No. OTS0522946. BASF. (2000) Cytogenetic study in vivo with of HBCD in the mouse micronucleus test after two intraperitoneal administrations. Experimental Toxicology and Ecology, BASF Aktiengesellschaft , 67056 Ludwigshafen, Germany; Project No. 26MO100/004018. As cited in European Commission (2008). Brandsma, S. H.; Van der Ven, L. T.; De Boer, J.; and Leonards, P. E. (2009). Identification of hydroxylated metabolites of hexabromocyclododecane in wildlife and 28

- days exposed Wistar rats. Env iron. Sci. Technol. 43, 6058

– 6063
. Chengelis, C. ( 1997 ) . A 28

- day repeated dose oral toxicity study of HBCD in rats. Study No. WIL

- 186004. WIL Research Laboratories, Inc. Ashland, OH. Available online at http://yosemite .epa.gov/oppts/epatscat8.nsf/ALLIDS/E90A4C7EC4F02C0785256930004C0892/ $FILE/84970000012.pdf?OpenElement Chengelis, C. (2001). An oral (gavage) 90

- day toxicity study of HBCD in rats. Study No. WIL

186012. WIL Research Laboratories, Inc. Ashland, Ohio. Avail able online at http://www.epa.gov/oppt/tsca8e/pubs/8ehq/2002/jan02/fyi_0102_01424a.pdf
. Chengelis, C. (2002). Amendment to the Final Report for: An oral (gavage) 90

- da y toxicity study of HBCD in rats. Study No. WIL

- 186012. WIL Research Laboratories, Inc. Ashland, Ohio. Available online at http://www.epa.gov/oppt/tsca8e/pubs/8ehq/200 2/sep02/fyi_0902_01424b2.pdf
. Covaci, A., A. C. Gerecke, et al. (2006). Hexabromocyclododecanes (HBCDs) in the Environment and Humans: A Review. Environ Sci Technol 40 : 3679

- 3688.91

Ema, M; Fujii, S; Hirata

- Koizumi, M; et al. (2008) Two

- generation reproductive toxicity study of the flame retardant HBCD in rats. Reprod Toxicol 25(3):335

– 351. Eriksson, P; Fischer, C; Wallin, M; et al. (2006) Impaired behaviour, learning and memory, in adult mice neonatally exposed to HBCD (HBCDD). Environ Toxicol Pha rmacol 21(3):317

– 322. Ethyl Corporation. (1985a) Genetic toxicology Salmonella/microsomal assay on HBCD with cover letter dated 030890. Submitted under TSCA Section 8D; EPA Document No. 86

- 900000164; NTIS No. OTS0522235. Ethyl Corporation. (1985b) Genetic toxicology rat hepatocyte primary culture/DNA repair test on HBCD with cover letter dated 030890. Submitted under TSCA Section 8D; EPA Document No. 86

- 900000163; NTIS No. OTS0522234. European Commission (2008). Risk Ass essment: HBCD CAS

- No.: 25637

- 99

- 4 EINECS No.: 247

- 148

- 4, Final Report May 2008. Luxembourg. Available online at http://ec.europa.eu/health/ph_risk/committees/04_sch er/docs/scher_o_093.pdf
. Fangstrom, B; Strid, A; Bergman, A. (2005) Temporal trends of brominated flame retardants in milk from Stockholm mothers, 1980

- 2004. Department of Environmental Chemistry, Stockholm University, Stockholm, Sweden. Available online at http://www.imm.ki.se/Datavard/PDF/mj%C3%B6lk_poolade_NV%20rapport%202005%20mod ersmjolk.pdf
. Fangstrom, B; Athanassiadis, I; Odsjo, T; et al. (2008) Temporal trends of polybrominated diphenyl ethers and HBCD in milk from Stockholm mothers, 1980

- 2004. Mol Nutr Food Res 52(2):187

– 193. Geyer, HJ; Schramm, K

- W; Darnerud, PO; Aune, M; Feicht, EA; Fried, KW; Henkelmann, B; Lenoir, D; Schmid, P; McDon ald, TA. (2004) Terminal elimination half

- lives of the brominated flame retardants TBBPA, HBCD, and lower brominated PBDEs in humans. Organohalogen Compounds 66: 3820

– 3825. Gulf South Research Institute. (1988) Initial submission: Letter from Ethyl Corp to USEPA re technical and toxicity data on brominated flame retardants including HBCD. EPA Document No. FYI

- OTS

- 0794

- 0947; NTIS No. OTS0000947. Hakk, H.; Szabo, D. T.; Huwe, J.; Diliberto, J.; and Birnbaum, L. S. (2012). Novel and distinct metabolites id entified following a single oral dose of α - or γ - hexabromocyclododecane in mice. Environ. Sci. Technol
. 46: 13494

– 13503. Helleday, T; Tuominen, KL; Bergman, A; et al. (1999) Brominated flame retardants induce intragenic recombination in mammalian cells. Mutat Res 439(2):137

– 147. Hill, R. N., Crisp, T. M., Hurley, P. M., Rosenthal, S. L., and Singh, D. V. (1998). Risk assessment of thyroid follicular cell tumors. Environ. Health Perspect. 106, 447

– 457.92

Huntingdon Research Center. (1978) Ames metabolic ac tivation test to assess the potential mutagenic effect of compound no. 49 with cover letter dated 031290. Submitted under TSCA Section 8D; EPA Document No. 86

- 900000385; NTIS No. OTS0522948. Industrial Bio

- Test Labs. (1990) Mutagenicity of two lots of FM

- 100 lot 53 and residue of lot 3322 in the absence and presence of metabolic activation with test data and cover letter. Submitted under TSCA Section 8D; EPA Document No. 86

- 900000267; NTIS No. OTS0523259. IRDC (International Research and Development Cor poration). (1977) Acute toxicity studies in rabbits and rats with HBCD with attachments. Submitted under TSCA Section 8E; EPA Document No. 88

- 7800065; NTIS No. OTS0200051. IRDC (International Research and Development Corporation). (1978a) Acute toxicity studies in rabbits and rats with residue of HBCD with attachments and cover letter dated 030178. Submitted under TSCA Section 8E; EPA Document No. 88

- 7800088; NTIS No. OTS020 0466. IRDC (International Research and Development Corporation). (1978b) Acute inhalation toxicity study in rats with HBCD with attachments and cover letter dated 042478. Submitted under TSCA Section 8E; EPA Document No. 88

- 7800137; NTIS No. OTS0200488. Johnson

- Restrepo, B. ; D. H. Adams ; et al. (2008). Tetrabromobisphenol A (TBBPA) and Hexabromocyclododecanes (HBCDs) in tissues of humans, dolphins, and sharks from the United States. Chemosphere 70 : 1935

- 1944. Kakimoto, K; Akutsu, K; Konishi, Y; et al. (2008) Time trend of HBCD in the breast milk of Japanese women. Chemosphere 71(6):1110

– 1114. Kiciński, M; Viaene, MK; Den Hond, E; Schoeters, G; Covaci, A; Dirtu, AC; Nelen, V; Bruckers, L; Croes, K; Sioen, I; Baeyens, W; Van Larebeke, N; Nawrot, TS. (2012) Neurobehavioral function and low

- level exposure to brominated flame retardants in adolescents: a cross - sectional study. Environ Health 11:86. Kim, U

- J; Oh, J

- E. (2014) Tetrabromobisphenol A and hexabromocyclododecane flame retardants in infant moth er paired serum samples, and their relationships with thyroid hormones and environmental factors. Environmental Pollution 184: 193

– 200. Lewis, AC; Palanker, AL. (1978) A dermal LD50 study in albino rabbits and an inhalation LC50 study in albino rats. Test material GLS

- S6

- 41A [unpublished]. Consumer Product Testing, Fairfield, NJ; Experiment Reference No. 78385

- 2. Client: Saytech Inc. Lilienthal, H; van der Ven, LT; Piersma, AH. (2009) Effects of the brominated flame retardant HBCD (HBCD) on dopamine

- dependent behavior and brainstem auditory evoked potentials in a one - generation reproduction study in Wistar rats. Toxicol Lett 185:63

– 72.93

Litton Bionetics Inc. (1990) Mutagenicity evaluation of 421

- 32b (Final report) With test data and cover let ter. Submitted under TSCA Section 8D; EPA Document No. 86

- 900000265; NTIS No. OTS0523257. Meijer, L; Weiss, J; Van Velzen, M; et al. (2008) Serum concentrations of neutral and phenolic organohalogens in pregnant women and some of their infants in The Net herlands. Environ Sci Technol 42(9):3428

– 3433. Microbiological Associates Inc. (1996) Hexabromocyclododecane (HBCD): chromosome aberrations in human peripheral blood lymphocytes with cover letter dated 12/12/1996. Submitted under TSCA Section 8D; EPA Docu ment No. 86970000358; NTIS No. OTS0573552. Momma, J; Kaniwa, M; Sekiguchi, H; Ohno, K; Kawasaki, Y; Tsuda, M; Nakamura, A; Kurokawa, Y. (1993) Dermatological evaluation of a flame retardant, hexabromocyclododecane (HBCD) on guinea pig by using the primar y irritation, sensitization, phototoxicity, and photosensitization of skin. [Article in Japanese; English abstract]. Eisei Shikenjo Hokoku 111:18

– 24. Murai, T; Kawasaki, H; Kanoh, S. (1985) Studies on the toxicity of insecticides and food additives in pregnant rats (7). Fetal toxicity of HBCD. Oyo Yakuri (Pharmacometrics) 29:981

– 986 (in Japanese with English abstract)
. OECD (Organisation for Economic Co

- operation and Development). (1983). Guideline for the Testing of Chemicals: One

- Generation Reproduction Toxicity Study, Guideline 415. Available online at http://www.oecd.org/chemicalsafety/risk

- assessment/1948458.pdf
. OECD (Organisation for Economic Co

- operation and Development). (1995). Guideline for the Testing of Chemicals: Repeated Dose 28

- Day Oral Toxicity Study in Rodents, Guideline 407. Available online at http://www.oecd.org/env/ehs/testing/37477972.pdf
. OECD (Organisation for Economic Co

- operation and Development). (2001). Guideline for the Testing of Chemicals: Two

- Gene ration Reproduction Toxicity Study, Guideline 416. Available online at http://www.oecd.org/chemicalsafety/risk

- assessment/1948466.pdf
. Pharmakologisches Institute. (1978) Ames test with hexabromides with cover letter dated 031290. Submitted under TSCA Section 8D; EPA Document No. 86

- 900000379; NTIS No. OTS0522942. Pharmakon Research International Inc. (1990) Acute exposure oral toxicity study in rats (83 EPA/OECD) with attach ments and cover letter dated 030890. Submitted under TSCA Section 8D; EPA Document No. 86

- 900000166; NTIS No. OTS0522237. Rawn, D.F.K.; Ryan, J.J.; Sadler, A.R. et al. (2014). Brominated flame retardant concentrations in sera from the Canadian Health Mea sures Survey (CHMS) from 2007 to 2009. Environment International 63: 26

– 34.94

Reistad, T; Fonnum, F; Mariussen, E. (2006) Neurotoxicity of the pentabrominated diphenyl ether mixture, DE

- 71, and HBCD (HBCD) in rat cerebellar granule cells in vitro. Arch Toxicol 80(11):785

– 796. Saegusa, Y; Fujimoto, H; Woo, G et al. (2009) Developmental toxicity of brominated flame retardants, tetrabromobisphenol A and 1,2,5,6,9,10

- HBCD, in rat offspring after maternal exposure from mid

- gestation through lactation. Repro d Toxicol 28(4):456

- 67. Sanders, J.M.; Knudsen, G.A.; and Birnbaum, L.S. (2013). The fate of β

hexabromocyclododecane in female C57BL/6 mice. Toxicological Sciences 134(2): 251

– 257. Schecter, A; Szabo, DT; Miller, J; Gent, TL; Malik

- Bass, N; Petersen, M; Paepke, O; Colacino JA; Hynan LS; Harris, TR; Malla, S; Birnbaum, LS. (2012) Hexabromocyclododecane (HBCD) stereoisomers in U.S. food from Dallas, TX. Environmental Health Perspectives 120(9): 1260

– 1264. Slob, W. (2002) Dose

–

response modeling of continuous endpoints. Toxicol. Sci. 66, 298

– 312. SRI Research Institute. (1990) In vitro microbiological mutagenicity studies of four CIBA

GEIGY corporation compounds (Final report). With test data and cover letter. Submitted under TSCA Section 8D; EPA Document No. 86

- 900000262; NTIS No. OTS0523254. Sz abo, DT; Diliberto, JJ; Hakk, H; et al. (2010) Toxicokinetics of the flame retardant HBCD gamma: effect of dose, timing, route, repeated exposure, and metabolism. Toxicol Sci 117(2):282

– 293
. Szabo, DT; Diliberto, JJ; Huwe, JK; Birnbaum, LS. (2011a) Differ ences in tissue distribution of HBCD alpha and gamma between adult and developing mice. Toxicol Sci 123(1):256

– 263. Szabo, DT; Diliberto, JJ; Hakk, H; Huwe, JK; Birnbaum, LS. (2011b) Toxicokinetics of the flame retardant hexabromocyclododecane alpha: effe ct of dose, timing, route, repeated Exposure, and metabolism. Toxicol Sci 121(2):234

– 244. Thomsen, C; Molander, P; Daae, HL; et al. (2007) Occupational exposure to HBCD at an industrial plant. Environ Sci Technol 41(15):5210

– 5216. Toms, L

- ML; Guerra, P; Eljarrat, E; Barceló, D; Harden, FA; Hobson, P, et al. (2012) Brominated flame r e tardants in the Australian population: 1993

– 2009. Chemosphere 89:398

– 40 3
. U.S. EPA (Environmental Protection Agency). (1991) Guidelines for developmental toxicity risk assess ment. Federal Register 56(234):63798

– 63826. Available online at http://www.epa.gov/iris/backgr

- d.htm
.95

U.S. EPA (Environmental Protection Agency). (1996) Guidelines for reproductive toxicity risk assessm ent. Federal Register 61(212):56274

– 56322. Available online at http://www.epa.gov/iris/backgr

- d.htm
. U.S. EPA (Environmental Protection Agency). (1998) Guidelines for neurotoxicity risk assessment. Federal Register 63(93):26926

– 26954. Available online at http://www.epa.gov/iris/backgr

- d.htm
. U.S. EPA ( U.S. Environmental Protection Agency). (2001) High Production Volume (HPV) data summary and test plan for hexabromocyclododecane (HBCD) CAS No. 3194

- 55

- 6. Prepared by the American Chemistry Council’s Brominated Flame Retardant Industry Panel (BFRIP), Arlin gton, VA. Available online at: http://www.epa.gov/oppt/chemrtk/pubs/summaries/cyclodod/c13459.pdf
. U.S. EPA (Environmental Protection Agency). ( 2014 ) Flame Retardant Alternatives for Hexabromocyclododecane (HBCD): Final Report . Available online at http://www2.epa.gov/sites/production/files/2014

- 06/documents/hb cd_report.pdf
. van der Ven, LT; Verhoef, A; van de Kuil, T; Slob, W; Leonards, PEG; Visser, TJ; Hamers, T; Herlin, M; Hakansson, H; Olausson, H; Piersma, AH; Vos, JG. (2006) A 28

- day oral dose toxicity study enhanced to detect endocrine effects of hexabro mocyclododecane in Wistar rats. Toxicological Sciences 94(2): 281

– 292. van der Ven, LT; van de Kuil, T; Leonards, PE; et al. (2009) Endocrine effects of HBCD (HBCD) in a one

- generation reproduction study in Wistar rats. Toxicol Lett 185:51

– 62. Including supplementary tables. Weiss, J; Wallin, E; Axmon, A; et al. (2006) Hydroxy

- PCBs, PBDEs, and HBCDDs in serum from an elderly population of Swedish fishermen's wives and associations with bone density. Environ Sci Technol 40(20):6282

– 6289. Yu, CC; Atallah, YH. (1980) Pharmacokinetics of HBCD in rats [unpublished]. Vesicol Chemical Corporation, Rosemont, IL. As cited in European Commission (2008)
. Zegers, BN; Mets, A; Van Bommel, R; et al. (2005) Levels of HBCD in harbor porpoises and common dolphins from western European seas, w ith evidence for stereoisomer

- specific biotransformation by cytochrome P450. Environ Sci Technol 39(7):2095

– 2100. Zeiger, E; Anderson, B; Haworth, S; et al. (1987) Salmonella mutagenicity tests: III. Results from the testing of 255 chemicals. Environ Muta gen 9(Suppl. 9):1

– 110. Zeller, H. and Kirsch, P. 1969. Hexabromocyclododecane: 28

- day feeding trials with rats. BASF unpublished laboratory study. As cited in U.S. EPA. 2001. Zeller, H. and Kirsch, P. 1970. Hexabromocyclododecane: 90

- day feeding t rials with rats. BASF unpublished laboratory study. As cited in U.S. EPA. 2001.96 4. ECOLOGICAL HAZARD ASSESSMENT
4.1. SUMMARY HBCD presents a significant concern for adverse effects on the environment. This conclusion is based on the observed potential for bioaccumulation, biomagnification, altered reproductive behavior, as well as high acute and chronic toxicities. Bioconcentrat ion factors (BCFs) and biomagnification factors (BMFs) as high as 18,100 and 29.7 respectively have been observed in fish (Du et al., 2012a; Law et al., 2006; Zhang et al., 2014b). BMF values of 26 (lipid

- weight) and 1.6

- 3 have also been observed in bird s (Haukås et al., 2010a) and mammals (Shaw et al., 2012) respectively. Observed acute toxicity values as low as 0.009 mg/L for a 72 hour EC 50 based on reduced growth in the marine algae, Skeletonema costatum) (Walsh et al., 1987), and observed chronic aqua tic toxicity values as low as 0.0042 mg / L (maximum acceptable toxicant concentration (MATC) ) for reduced size (length) of surviving young in Daphnia magna (Drottar and Krueger, 1998) , indicate high acute and chronic aquatic toxicity. Reduced chick survival in Japanese quails (Coturnix coturnix japonica) fed a 15 ppm HBCD diet (2.1 mg/kg body weight

- day) (MOEJ, 2009 as cited in UNEP, 2010) and altered reproductive behavior (reduced courtship and brood

- rearing activity) and reduced egg size in American kestre ls (Falco sparverius) fed 0.51 mg/kg body weight

- day (Fernie et al., 2011; Marteinson et al., 2010, 2011, 2012) indicate high terrestrial toxicity as well. The model

- predicted (ECOSAR) chronic toxicity value for the estuarine/marine invertebrate, mysid shr imp , was 0.0007 mg/L which is also consistent with an expectation of high chronic aquatic toxicity for HBCD. Assessment of HBCD aquatic toxicity is complicated by the low water solubility of the chemical and differences in the solubility of the three main HBCD isomers , which makes testing difficult and interpretation uncertain for studies conducted above the water solubility. Studies97

conducted at concentrations above the water solubility of HBCD are essentially testing the effects at the maximum HBCD conc entration possible. In contrast with the studies cited above, other acute and chronic aquatic toxicity studies conducted using methods, test species, and endpoints recommended by the EPA reported no effects at or near the limit of water solubility. Howev er, water solubility is not considered a limiting factor for hazard determination for aquatic species since there are studies showing adverse effects at or below the water solubility of HBCD (see Section 4.2 for a more detailed discussion of this issue)
. In addition, the potential for HBCD to bioaccumulate, biomagnify, and persist in the environment, significantly increases concerns for effects on aquatic organisms. A wide range of effects of HBCD have been reported in fish (e.g., developmental toxicity, embryo malformations, reduced hatching success, reduced growth, hepatic enzyme and biomarker effects, thyroid effects, DNA damage to erythrocytes, and oxidative damage) and invertebrates (e.g., degenerative changes, morphological abnormalities, decreased hatching success, and altered enzyme activity) in supporting studies that assessed endpoints beyond those evaluated in guideline studies (Anselmo et al., 2011; Deng et al., 2009; Du et al., 2012 b ; Hong et al., 2014a; Hu et al., 2009; Palace et al., 2008, 2 010; Ronisz et al., 2004; Smolarz and Berger, 2009; Wu et al., 2013; Zhang et al., 2008, 2014a). Effects on the thyroid in fish (reduced thyroid hormone ( triiodothyronine, T3, and thyroxine, T4 ) levels in rainbow trout ( Palace et al., 2008, 2010) ) , are sim ilar to those observed in mammals (see Chapter 3, Human Health Hazard Assessment). Reduced T4 levels were also reported in birds exposed to HBCD (Marteinson et al., 2011). In the following sections, available data on the ecological toxicity and bioaccumu lation of HBCD and its diastereomers are presented and described in further detail.98

4.2. ECOLOGICAL TOXICITY Databases searched in 2011 and 2015 to identify studies on ecological toxicity were ECOTOX (which contains three databases: AQUIRE, PHYTOTOX, and TERRETOX), the Office of Pesticide Programs (OPP) Environmental Fate and Effects Division (EFED) Pesticide Ecotoxicity Database 8 , and Toxic Substances Control Act (TSCA)

- related databases (i.e., TSCATS I 9 , TSCATS II 10 , and publicly available 8e and FYI noti ces 11 ). In addition, the TOXLINE 12 database was searched to identify population effects studies and additional aquatic toxicity studies. This assessment includes not only studies that followed standard test guidelines (e.g., Office of Chemical Safety and Po llution Prevention [OCSPP], Organisation for Economic Co

- operation and Development [OECD]), but also non

- standard toxicity tests that followed procedures that were scientifically sound. Only clearly adverse signs of toxicity (e.g., lethality, immobility, e ffects on growth and reproduction, organ histopathology, abnormal behavior) were used to set toxicity effect levels such as no

- observed

- effect concentrations (NOECs) and lowest

observed

- effect concentrations (LOECs). Although studies in which the nominal c oncentrations greatly exceeded water solubility (e.g., Siebel

- Sauer and Bias, 1987 as cited in ECB, 2008; Jatzek, 1990; Calmbacher, 1978) are described in this assessment, the conclusion of concern for adverse effects on the environment is not based on the se studies, as none of these studies

8 Available at http://cfpub.epa.gov/ecotox/
. 9 Available at h ttp://www.syrres.com/what

- we - do/databaseforms.aspx?id=384
. 10 Available at http://yosemite.epa.gov/oppts/epatscat8.nsf/ReportSearch?OpenForm
. 11 Available at http://www.epa.gov/opptintr/tsca8e/pubs/8eandfyisubmissions.html#search
. 12 Available at http://toxnet.nlm.nih.gov/ cgi

- bin/sis/htmlgen?TOXLINE
.99

reported toxic effects. Studies were considered to be invalid if the control response was abnormal. Commercial mixtures of HBCD, or technical HBCD (HBCD

- TM), consist primarily of three diastereomers : α - HBCD, β

- HBCD, an d γ

- HBCD. Each of these diastereomers has two enantiomers (+ and ̶ ); the ratio between these two enantiomers is often reported as the enantiomeric fraction (EF). Enantiomers of chiral compounds such as HBCD often exhibit differences in biological activit y (Zhang et al., 2009). As summarized in de Wit et al. (2010), γ - HBCD tends to predominate in air, α

- HBCD is the most abundant diastereomer in biota, and α

HBCD, β

- HBCD, and γ

- HBCD have been found in relatively similar proportions in marine sediments. More recent environmental monitoring studies are reporting concentrations of individual diastereomers (e.g., Chen et al., 2011) and ecotoxicological studies are reporting acute and chronic effects of individual diastereomers of HBCD (e.g., Du et al., 2012 b ). HBCD is poorly water soluble and tends to partition to organic phases in the aquatic environment (e.g., lipids, suspended solids; see Marvin et al., 2011). Assessment of aquatic toxicity is complicated by the low water solubility of the chemical and differ ences in the solubility of the three main HBCD isomers , which makes testing difficult and interpretation uncertain for studies conducted above the water solubility . Aquatic studies available to assess ecotoxicity concerns generally used solvents such as di methyl sulfoxide (DMSO) to solubilize HBCD into solution and generally did not characterize the impact that solvents had on test solutions. According to OECD guidance for testing difficult substances (OECD, 2000), “because of the potential for interaction with the test substance resulting in an altered response in the test, their use should be restricted to situations where no other acceptable method of media preparation is available.” If use of a solvent is deemed necessary, OECD (2000) recommends100

that a p reliminary experiment be conducted that determines solubility in test solutions under testing conditions. OECD (2000) includes a caveat that “a high level of uncertainty is often associated with solubility estimates for poorly water

- soluble substances. It is therefore often not possible to specify an upper exposure limit with a high degree of confidence. Exposure concentrations which exceed a “reported” water solubility should therefore not necessarily be considered invalid but reviewed on a case

- by

- case ba sis.” Studies in which an organic solvent was used were included in this assessment; the use of a solvent is noted in the study summary. For studies in which nominal concentrations greatly exceeded the water solubility of HBCD, a comment stating such has b een included in the study description. It should be noted that water solubility studies available for HBCD indicate differences among the water solubilities of the three diastereomers , α - , β

- , and γ

- HBCD (0.0488, 0.0147, and 0.0021

– 0.0034 mg/L for α

- , β

- , and γ

- HBCD, respectively), which suggests that overall water solubility of technical

- grade HBCD may vary depending on its isomeric composition. Also, a solvent may differentially alter the solubility of each of the HBCD diastereomers. Presently, insuffic ient information is available to address whether use of solvents affects the outcomes of available studies or to establish a more definitive limit of water solubility. The nature of HBCD (e.g., it is persistent, bioaccumulates, and has potential for biomag nification) suggests that effects observed above the limit of solubility in standard guideline studies may remain relevant for hazard characterization. Bioaccumulation studies indicate that HBCD will accumulate in tissue (see Section 4.3). Results of bioma gnification studies (Du et al., 2012 a , 2013; Haukås et al., 2010a; Law et al., 2006; Letcher et al., 2009; Shaw et al., 2012; Sun et al., 2012) also suggest that exposure of aquatic and terrestrial organisms to HBCD may be from both diet and the ambient en vironment (e.g., through an organism’s101

respiratory and dermal surfaces). Environmental monitoring data show frequent occurrence of HBCD in biota over broad geographical areas and in remote locations where no demonstrable sources have been shown to account for the occurrence of HBCD (e.g., Allchin and Morris, 2003; De Boer et al., 2004 as cited in UNEP, 2010; Dodder et al., 2014; Ilyas et al., 2013; Klosterhaus et al., 2012; McKinney et al., 2011; Miljeteig et al., 2009; Muir et al., 2006; Salamova and Hites , 2013; Tomy et al., 2008; Zegers et al., 2005). Together, these factors support a conclusion that HBCD is persistent in the environment (see Chapter 2, Chemistry and Environmental Fate, for a more detailed discussion of environmental persistence of HBCD) and within the tissues of environmental species and, thus, exposure and HBCD body burden are not limited by HBCD water solubility. In summary, since it is not currently possible to specify an upper water solubility limit for HBCD with confidence, body burd ens of HBCD in aquatic organisms may originate from both the diet and ambient water, and HBCD can bioaccumulate, hazard determination in this report included the consideration of data from all available aquatic studies. Table 4

- 1 presents the measured toxi city values (and predicted value for mysid shrimp) indicating acute and chronic toxicity of HBCD to aquatic and terrestrial organisms. Section 4.2.1 discusses the acute aquatic toxicity data for HBCD; chronic aquatic toxicity studies are summarized in Sect ion 4.2.2; Section 4.2.3 contains the results of predicted aquatic toxicity values using U.S. EPA models for types of species that were not found in the available toxicity studies; Section 4.2.4 details the available information on terrestrial toxicity and phytotoxicity; and Section 4.2.5 discusses population effects.102

Table 4 - 1. Ecological toxicity values for HBCD Species Experimental

type a,b Duration and test

endpoint Concentration of

HBCD c Comments Reference Acute aquatic toxicity Freshwater algae Green algae ( Selanastrum capricornutum ) S, M d,e 96

- hour EC 50 >0.0037 mg/L Nominal test concentrations ranged from 1.5 to 6.8 μg/L. The mean measured concentration of HBCDs (3.7 μg/L) at the maximum nominal test concentration of 6.8 μg/L (with solvent) was similar to the independently measured water solubility limit (8.6 μg/L, see Chemistry Chapter). No effects were observed at any test concentration. Roberts and Swigert, 1997 96

- hour NOEC 0.0037 mg/L

Green algae ( Scenedesmus subspicatus ) S, U f 96

- hour EC 50 >500 mg/L Nominal test concentrations ranged from 7.81 to 500 mg/L. Actual concentrations of HBCD in test solution are unknown. No effects were observed at the highest tested concentration. Primary report not available. Siebel

- Sauer and Bias, 1987 as cited in ECB, 2008 96

- hour NOEC 500 mg/L
Freshwater invertebrate Water flea ( D. magna ) F, M d,e 48

- hour EC 50 >0.0032 mg/L Nominal test concentrations were 1.5, 2.2, 3.2, 4.6, and 6.8 μg/L. Corresponding mean measured concentration of HBCDs were 2.4, 1.8, 2.1, 2.3, and 3.2 μg/L (with solvent). The maximum nominal concentration was similar to the independently measured water solubility of 8.6 μg/L (see Chemistry Chapter). No effects were observed at any tes t concentration. Graves and Swigert, 1997a 48

- hour NOEC 0.0032 mg/L

Water flea ( D. magna Straus) S, U g 48

- hour EC 50 146.34 mg/L Nominal test concentrations ranged from 0.01 to 1,000 mg/L and greatly exceeded the independently measured water solubility of 0.0086 mg/L (see Chemistry Chapter). Actual concentrations of HBCD in test solution are unknown. Appearance of test solution was not reported and, thus, the occurrence of undissolved test substance in solution is unknown and may have affected Daphnia mobility at higher test concentrations. Jatzek, 1990 48

- hour LOEC 10 mg/L 48

- hour NOEC 1 mg/L103

Table 4 - 1. Ecological toxicity values for HBCD Species Experimental

type a,b Duration and test

endpoint Concentration of

HBCD c Comments Reference Freshwater fish Rainbow trout ( Oncorhynchus mykiss ) F, M d,e 96

- hour LC 50 >0.0025 mg/L Nominal test concentrations were 1.5, 2.2, 3.2, 4.6, and 6.8 μg/L. Corresponding mean measured concentration of HBCDs were 0.75, 1.5, 2.3, 2.3, and 2.5 μg/L (with solvent). The maximum nominal c oncentration was similar to the independently measured water solubility of 8.6 μg/L (see Chemistry Chapter). No effects were observed at any test concentration. Graves and Swigert, 1997b 96

- hour NOEC 0.0025 mg/L

Bluegill sunfish ( Lepomis macrochirus Rafinesque) S, U h,i,j 96

- hour LC 50 >100 mg/L Nominal test concentrations ranged from 10 to 100 mg/L and greatly exceeded the independently measured water solubility of 0.0086 mg/L (see Chemistry Chapter). Actual concentrations of HBCD in test solutions are unknown. A white flocculate was formed on the surface of the water in all test solutions. No effects were observed at any test concentration. Calmbacher, 1978 96

- hour NOEC 100 mg/L

Zebrafish ( Danio rerio ) embryos S, U k 92

- hour LOEC (survival) 0.05 mg/L Early life stage test. Four

- hour post

- fertilization zebrafish embryos were exposed to nominal concentrations of 0, 0.05, 0.1, 0.5, or 1 mg/L until 96 hours post

- fertilization (92

- hour exposure). Actual concentrations of HBCD in test solutions are unkn own. Embryo survival was significantly reduced at all tested concentrations; thus, a NOEC was not established. Heart rate was also significantly decreased at all tested concentrations. Embryo growth was significantly decreased and malformation rate was sig nificantly increased at ≥0.1 mg/L. Malformations included epiboly deformities, yolk sac and pericardial edema, tail and heart malformations, swim bladder inflation, and spinal curvature. Hatching rate was reduced only at the high concentration of 1 mg/L. A ssociated mechanistic studies suggest that the mechanism for developmental toxicity involves the generation of ROS and the consequent triggering of apoptosis genes. Deng et al., 2009 92

- hour NOEC <0.05 mg/L104

Table 4 - 1. Ecological toxicity values for HBCD Species Experimental

type a,b Duration and test

endpoint Concentration of

HBCD c Comments Reference Zebrafish ( D. rerio ) embryos R, U k 96

- hour NOEC (survival, hatching rate, malformations) 10 mg/L Early life stage test. Zebrafish embryos, within 2 hours post
- fertilization, were exposed to nominal concentrations of 0, 0.002, 0.01, 0.1, 0.5, 2.5, or 10 mg/L for 96 hours. Solvent (DMSO) conc entration was no more than 0.05%. Actual concentrations of HBCD in test solutions are unknown. There were no effects on embryo survival or malformations. Hatching of embryos was significantly delayed at 0.002

– 0.5 mg/L, but not at higher concentrations, a nd hatching success was not affected. The relevance of delayed hatching at low concentrations without any effect on hatching rate is uncertain. Associated mechanistic investigations showed that HBCD can cause oxidative stress and Hsp70 (heat shock protein) overexpression, potentially inducing acute toxicity to zebrafish embryo in a short

- term exposure. Hu et al., 2009

Zebrafish ( D. rerio ) embryos R, U k 72

- hour NOEC (survival, hatching rate, malformations) 0.13 mg/L 1 Early life stage test. Zebrafish embryos, within 1 hour post
- fertilization, were exposed to nominal HBCD concentrations of 0, 2, 20, or 200 nM (0, 0.0013, 0.013, or 0.13 mg/L, respectively) with a DMSO concentration of 50 μL/L for 72 hours. Actual concentr ations of HBCD in test solution are unknown. There were no significant adverse effects on survival, hatching rate, or malformation rate (heart or whole embryo) at any concentration. Significant increases in cardiac arrhythmia and (in contrast to Deng et al ., 2009) heart rate were observed at ≥0.013 and ≥0.0013 mg/L, respectively. Associated mechanistic studies showed no increase in apoptosis around the heart, but upregulation of a subset of genes related to cardiac development. Wu et al., 2013105

Table 4 - 1. Ecological toxicity values for HBCD Species Experimental

type a,b Duration and test

endpoint Concentration of

HBCD c Comments Reference Zebrafish ( D. rerio ) embryos R, U k LOEC: γ

- HBCD (survival, hatching, malformations) 0.01 mg/L Early life stage test. Four

- hour post

- fertilization zebrafish embryos were exposed to three diastereomers of HBCD (α

- , β

- , and γ

- HBCD) individually at nominal concentrations of 0, 0.01, 0.1, and 1.0 mg/L in 0.005% DMSO until 120 hours post

- fertilization (hpf), corresponding to a 116

- hour exposure. Actual concentrations of HBCD in test solution are unknown. All diastereomers affected embryo development in the order: γ

- HBCD > β

- HBCD > α

- HBCD. For γ

- HBCD; hatching success (72 hpf) and survival (96 hpf) were decreased and malformation rate (120 hpf) increased, at ≥0.01 mg/L. Growth (96 hpf) was decreased at ≥0.1 mg/L. For β

- HBCD, malformation rate was increased at ≥0.01 mg/L, but decreases in hatching success, survival, and growth were only seen at ≥0.1 mg/L. For α - HBCD, the most sensitive endpoints, hatching success and malformation rate, were affected at ≥0.1 mg/L, while survival and growth were decreased only at 1.0 mg/L. Heart rate was affected by all isomers at ≥0.1 mg/L, decreased at both 48 and 96 hpf by α - HBCD and β

- HBCD, but decreased at 48 hpf and increased at 96 hpf by γ

- HBCD. Associated mechanistic studies found that all three isomers induced apoptosis through ROS format ion. Du et al., 2012 b NOEC: γ

- HBCD <0.01 mg/L LOEC: β

- HBCD (malformations) 0.01 mg/L NOEC: β

- HBCD <0.01 mg/L LOEC: α

- HBCD (hatching, malformations) 0.1 mg/L
NOEC: α

- HBCD 0.01 mg/L106

Table 4 - 1. Ecological toxicity values for HBCD Species Experimental

type a,b Duration and test

endpoint Concentration of

HBCD c Comments Reference Estuarine/marine algae Algae ( S. costatum ) S, M m,n 72

- hour EC 50 >0.041 mg/L The test was performed to study effects on algal growth near the limit of water solubility (no solvent). A single concentration was tested, which produced approximately 10% inhibition of growth rate. Primary report not available. Desjardins et al., 2004 as cited in ECB, 2008 and IUCLID, 2005 72

- hour EC ~10 (growth) 0.041 mg/L
Algae ( S. costatum ) S, M n,o 72

- hour EC 50 (growth) 0.052 mg/L A single concentration was tested that represented the maximum achievable amount of test substance solubilized into solution without the use of a solvent, measured at 54.5 μg/L. At the 72

- hour observation, 51% growth inhibition was observed. This result corresponds to an EC 50 of 52 μg/L. Primary report not available. Desjardins et al. , 2005 as cited in ECB, 2008

Algae ( S. costatum ) S, M e 72

- hour EC 50 >0.01 mg/L Test solutions were prepared using dimethyl formamide solvent (0.1 ml /L). Nominal concentrations of 0.64, 1.6, 4.0, and 10 μg/L were tested. No significant effect on growth or biomass was found at any concentration. Concentrations were measured at the start of the experiment, but results were reported as nominal. Primary report not available. Desjardins et al., 2005 as cited in ECB, 2008 72

- hour NOEC 0.01 mg/L

Algae ( S. costatum ) S, M p 72

- hour EC 50 (growth) 0.009

– 0.012 mg/L

(0.010 mg/L in SW) The tests were conducted with seawater (SW), Rila

marine mix (RIL), Instant Ocean synthetic sea salts (IO), Forty Fathoms marine mix (40

- F), Utikem seawater compound (UTI), or HW Marinemix (HAW). Growth rates of each species were similar in all media. Maxi mum test concentrations were selected by determining the maximum achievable test concentration (in acetone) that did not result in crystal formation; the range of concentrations tested was not provided. HBCD affected growth of all three species tested, but Chlorella less than the others. EC 50 ranges reflect results in the different media tested. Walsh et al., 1987 Algae ( Thalassiosira pseudonana ) S, M p 72

- hour EC 50 (growth) 0.05

– 0.37 mg/L

(0.08 mg/L in SW) Algae ( Chlorella Sp.) S, M p 96

- hour EC 50 (growth) >1.5 mg/L107

Table 4 - 1. Ecological toxicity values for HBCD Species Experimental

type a,b Duration and test

endpoint Concentration of

HBCD c Comments Reference Estuarine/marine invertebrate Sea urchin ( Psammechinus miliaris ) embryos R, U k 72

- hour LOEC (morphological abnormalities) 0.064 mg/L l Early life stage test. Embryos were exposed to HBCD at nominal concentrations of 0, 9, 25, 50, and 100 nM (0, 0.0058, 0.016, 0.032, and 0.064 mg/L, respectively). After 72 hours of exposure, incidence of morphological abnormalities in sea urchin embryos wa s significantly increased at 0.064 mg/L (short or deformed larval arms and slight edema around the larval body). Hatching success was not affected at any concentration. Additional results from this study are provided in the chronic toxicity section. Ansel mo et al., 2011 72

- hour NOEC 0.032 mg/L l

Estuarine/marine fish - No data Chronic aquatic toxicity Freshwater invertebrate Water flea ( D. magna ) F, M d,e 21

- day LOEC (reduced length of surviving young) 0.0056 mg/L Nominal test concentrations were 0.85, 1.7, 3.4, 6.8, and 13.6 μg/L. Corresponding mean measured concentration of HBCDs were 0.87, 1.6, 3.1, 5.6, and 11 μg/L (with solvent). The maximum nominal concentration was similar to the independently measured water solubility of 8.6 μg/L (see Chemi stry Chapter). In addition to the reduction in length of surviving young seen at 0.0056 mg/L, reduced weight of surviving young and impaired reproduction were observed at 0.011 mg/L. Drottar and Krueger, 1998 21

- day MATC 0.0042 mg/L 21

- day NOEC 0.0031 mg/L

Amphipod ( Hyalella azteca ) F, M q 28

- day LOEC (reduced survivability) 500 mg/kg dry

weight sediment Non

- GLP range

- finding study conducted in the

presence of 2 and 5% TOC. Further study details were not provided. Thomas et al., 2003a, b 28

- day NOEC 100 mg/kg dry

weight sediment108

Table 4 - 1. Ecological toxicity values for HBCD Species Experimental

type a,b Duration and test

endpoint Concentration of

HBCD c Comments Reference Amphipod ( H. azteca ) F, M q 28

- day NOEC 1,000 mg/kg dry weight sediment Tests were conducted using spiked sediment with 5%

TOC. Nominal concentrations ranged from 31 to 1,000 mg/kg dry weight sediment. Measurement of test concentrations indicated that HBCD tended to remain in sediment with measured concentrations in sediment ranging from 78.2 to 122% of nominal test concentrations for the highest test concentration; measured concentrations for lower test concentrations were more variable. Concentrations of HBCD in overlying water were consistently below the limit of quantitation (LOQ; 0.1 mg/L). Effect levels were based on nominal concentrations of HBCD in sediment. No effects were observed at any test co ncentration. Thomas et al., 2003b

Amphipod ( H. azteca ) F, M m 28

- day NOEC 1,000 mg/kg dry weight sediment Tests were conducted using spiked sediment with 2%

TOC. Nominal concentrations ranged from 31 to 1,000 mg/kg dry weight sediment. Measurement of test concentrations indicated that HBCD tended to remain in sediment with measured concentrations in sediment ranging from 82.8 to 115% of nominal test concentrations for the highest test concentration; measured concentrations for lower test concentrations wer e more variable. Concentrations of HBCD in overlying water were consistently below the LOQ (0.1 mg/L). Effect levels were based on nominal concentrations of HBCD in sediment. No effects were observed at any test concentration. Thomas et al., 2003a109

Table 4 - 1. Ecological toxicity values for HBCD Species Experimental

type a,b Duration and test

endpoint Concentration of

HBCD c Comments Reference Worm (L umbriculus variegatus) S, M r,s 28

- day LOEC (reduction in worm number) 28.7 mg/kg dry

weight sediment The nominal test concentrations were 0.05, 0.5, 5, 50,

and 500 mg/kg dry weight sediment. Corresponding measured concentrations (geometric mean of day 0 and day 28) were ND, 0.2, 3.1, 28.7, and 303.2 mg/kg dry weight. No HBCD was detected in the overlaying water or in the pore water. Study details were excerpted from a secondary source (ECB, 2008). Total number of worms was significantly reduced at ≥28.7 mg/kg and biomass was significantly reduced at ≥303.2 mg/kg. A study with Chironomus riparius was conduc ted concurrently, but was considered to be invalid because the average emergence rate in the solvent control was unusually low and variation among solvent control replicates was high. Oetken et al., 2001 as cited in ECB, 2008 28

- day NOEC 3.1 mg/kg dry weight

sediment

Freshwater fish Rainbow trout ( O. mykiss ) F, M p,t 88

- day NOEC 0.0037 mg/L Nominal test concentrations were 0.43, 0.85, 1.7, 3.4, and 6.8 μg/L. Corresponding mean measured concentrations of HBCDs were 0.25, 0.47, 0.83, 1.8, and 3.7 μg/L (with solvent). The maximum nominal concentration was similar to the independently measured water solubility of 8.6 μg/L (see Chemistry Chapter). No effects were observed at any test concentration. Drottar et al., 2001

Chinese rare minnow ( Gobiocypris ra rus ) R, U 42

- day NOEC 0.5 mg/L Nominal HBCD concentrations ranged from 1 to 500 μg/L. After 42 days of exposure, corresponding internal HBCD levels ranged from 34 to 654 μg/g (wet weight) in whole fish. No mortality was observed. Associate mechanistic inv estigations found DNA damage in erythrocytes, induction of EROD and PROD in hepatocytes, and ROS formation in brain tissue at ≥0.1 mg/L. Zhang et al., 2008110

Table 4 - 1. Ecological toxicity values for HBCD Species Experimental

type a,b Duration and test

endpoint Concentration of

HBCD c Comments Reference European flounder ( Platichthys flesus ) Sediment exposed 78

- day NOEC 8,000 μg/g TOC in sediment Flounder were exposed to HBCD for 78 days via

spiked sediment and food. A maximum level of 446 mg/g lipid was reached in muscle from an animal exposed to the combination 3,000 μg/g lipid in feed and 800 μg/g TOC in sediment. No effects were seen on behavio r, survival, growth, organ weight or histopathology, enzyme activity, or plasma thyroid hormone levels in any group. Kuiper et al., 2007 Sediment and diet exposed 78

- day NOEC 800 μg/g TOC in sediment and 3,000 μg/g lipid in food
Estuarine/marine invertebrate Clam ( Macoma balthica ) S, NS p 50

- day LOEC (degenerative changes in the gills) 0.1 mg/L Nominal test concentrations were 0, 0.1, and 0.25 mg/L. Degenerative changes in the gills manifested by the increased frequency of nuclear and nucleolar abnormalities and the occurrence of dead cells were observed at all test concentrations and, thus, a NOEC was not determined. Smolarz and Berger, 2009 50

- day NOEC <0.1 mg/L Mysid shrimp predicted ChV 0.0007 mg/L ECOSAR, 2012 Sea urchin ( P. miliaris ) embryos R, U k 14

– 16

- day LOEC (morphological abnormalities) 0.064 mg/L l Early life stage test for sea urchin embryos. Embryos were exposed to HBCD at nominal concentrations of 0, 9, 25, 50, and 100 nM (0, 0.0058, 0.016, 0.032, and 0.064 mg/L, re spectively) from day 0 to 16 post

fertilization. After 16 days of exposure, incidence of morphological abnormalities in sea urchin embryos was significantly increased at 0.064 mg/L (short or deformed larval arms and slight edema around the larval body). Ha tching success was not affected at any concentration. Additional results from this study are provided in the acute toxicity section. MATC calculated for this review. Anselmo et al., 2011 14

– 16

- day MATC 0.045 mg/L 14

– 16

- day NOEC 0.032 mg/L l111

Table 4 - 1. Ecological toxicity values for HBCD Species Experimental

type a,b Duration and test

endpoint Concentration of

HBCD c Comments Reference Estuarine/marine fish Marine medaka ( Oryzias melastigma ) enbryos R, U k 17

- day LOEC (malformations) 0.05 mg/L Early life stage test. Embryos were exposed to technical HBCD at nominal concentrations of 0, 0.005, 0.020, and 0.050 mg/L until the first fry stage (17 days post
- fertilization) to assess effects on developmental toxicity. The final solvent (DMSO) concentration was 0.01%. There were no effects on hatch success, survival rate, or hatch

- out time. Malformation rate was significantly increased at 0.0 50 mg/L. The developmental abnormalities included yolk sac edema, pericardial edema and spinal curvature. Heart rate and SV

- BA distance, which are markers for cardiac development, were both increased in a dose

- related manner. Associated mechanistic investi gations found that HBCD induced oxidative stress and apoptosis, and suppressed nucleotide and protein synthesis. MATC calculated for this review. Hong et al., 2014a 17

- day MATC 0.03 mg/L 17

- day NOEC 0.02 mg/L

Sole ( Solea solea ) embryos R, U k NOEC 0.25 mg/L Early life stage test in which fertilized eggs were exposed to nominal concentrations of 25, 80, or 250 μg/L starting at 12 hours post

- fertilization until hatching at 5

– 6 days and then maintained unexposed until completion of meta morphosi s at around 40 days post
- fertilization. The final solvent (DMSO) concentration was 0.1%. Tissue concentrations in larvae after 6 days of exposure ranged from 2,280 to 12,400 μg/g ww. HBCD had no effect on hatching success or survival or development of larv ae through completion of metamorphosis. Foekema et al., 2014112

Table 4 - 1. Ecological toxicity values for HBCD Species Experimental

type a,b Duration and test

endpoint Concentration of

HBCD c Comments Reference Terrestrial toxicity and phytotoxicity Avian species Japanese quail ( Coturnix coturnix japonica ) Diet exposed v 6

- week LOAEL (reduction in hatchability) 125 ppm Reductions in egg shell thickness were observed at >125 ppm, and at 500 ppm, decreased egg weights and egg production rates and increased cracked eggs were reported. Adult mortality was reported at 1,000 ppm. Study details were excerpted from a secondary source and few study deta ils were provided. MOEJ, 2009 as cited in UNEP, 2010

Japanese quail ( C. coturnix japonica ) Diet exposed v 6

- week LOAEL (reduced chick survival) 15 ppm (2.1 mg/kg

day) A second study that tested lower HBCD concentrations

was performed in order to establish a NOAEL and LOAEL. Study details were excerpted from a secondary source. 6

- week NOAEL 5 ppm (0.7 mg/kg

day) American kestrel ( F. sparverius ) In ovo exposed LOAEL (behavioral effects) 164.3 ng/g wet

weight in egg Reduced corticosterone response in male nestling

kestrels, reduced flying activities in juvenile males, and delayed response time to predator avoidance in juvenile females. Effect level determined as concentrations in eggs. Kobiliris, 2010

American kestrel ( F. sparverius ), female Diet exposed w,x LOAEL (reduced egg size) 0.51 mg/kg

- day HBCD dissolved in safflower oil was injected into the brains of dead cockerels daily; kestrels were fed a ration of cockerels equivalent to approximately 0.51 mg/kg

- day. Dietary exposure began 3 weeks prior to pairing and continued through courtship, egg laying, and incubation, until the first chick hatched (approximately 75 days). Exposed kestrels laid eggs with ave rage tissue concentrations of 163.5 ng α HBCD/g ww, 13.9 ng β

- HBCD/g ww, and 2.6 ng γ

HBCD/g ww. Exposed birds displayed reduced time from pairing to egg laying and laid larger clutches of smaller eggs (volume, mass). Exposed eggs lost more weight than con trol eggs during incubation, but egg shell thickness was not affected. No effect on reproductive success was identified. Fernie et al., 2011113

Table 4 - 1. Ecological toxicity values for HBCD Species Experimental

type a,b Duration and test

endpoint Concentration of

HBCD c Comments Reference American kestrel ( F. sparverius ), male Diet exposed x LOAEL (increased testes weight in unpaired males) 0.51 mg/k g - day HBCD dissolved in safflower oil was injected into the brains of dead cockerels daily; kestrels fed from the cockerels ad libitum and received a dose of approximately 0.51 mg/kg

- day. Unpaired males (exposed for 3 weeks) were used for evaluation of te stis weight and histology; breeding males (exposed for 4 weeks pre

- pairing through one reproductive cycle until 2 days before chicks hatched, for a total of ~75 days of exposure) were used for sperm counts and levels of testosterone and thyroid hormone in blood. Testes weight was significantly increased versus controls in the unpaired males, possibly related to thyroid disruption, as total and free T4 levels were reduced in breeding males throughout the 75

- day exposure. Marginal effects were observed on tes tis histology in unpaired males (increased number of seminiferous tubules containing elongated spermatids; p = 0.052) and testosterone level in breeding males (increased at the time the first egg was laid; p = 0.054). No effects were observed on sperm coun ts. Marteinson et al., 2011

American kestrel ( F. sparverius ) Diet exposed x LOAEL (altered reproductive behavior in both sexes) 0.51 mg/kg

- day HBCD dissolved in safflower oil was injected into the brains of dead cockerels daily; kestrels fed from the cockerels ad libitum and received a dose of approximately 0.51 mg/kg

- day. Exposure began 4 weeks prior to pairing and continued throughout the court ship period, through egg incubation until 2 days before chicks hatched, for a total of ~75 days of exposure. Reduced activity was observed in both sexes during courtship and in males during brood rearing, possibly contributing to reduced egg size and reduc ed incubation nest temperature. Marteinson et al., 2012114

Table 4 - 1. Ecological toxicity values for HBCD Species Experimental

type a,b Duration and test

endpoint Concentration of

HBCD c Comments Reference Chicken ( Gallus gallus domesticus ) In ovo exposed y 22

- day LOEC (pipping success) 100 ng/g egg HBCD was injected into the air cell of chicken eggs prior to incubation. Nominal stock concentrations of HBCD in safflower oil were 0, 0.25, 0.5, 1.5, 5, and 50 mg/ ml ; corresponding measured stock concentrations were 0, 0.22, 0.43, NQ, 4.98, and NQ mg/m l ; respective nominal doses of HBCD were 0, 50, 100, 300, 1,000, and 10,000 ng/g egg. Eggs were observed fo r pipping success for 22 days. Reduced pipping success was observed at 100 (three independent experiments) and 10,000 ng/g, but not at 300 or 1,000 ng/g egg. Crump et al., 2010

Terrestrial invertebrate Earthworm ( Eisenia fetida ) Soil exposed 56

- day LOEC (reproduction) 235 mg/kg dry soil Nominal concentrations of 78.5, 157, 313, 625, 1,250, 2,500, and 5,000 mg/kg dry soil were tested. Worms were exposed for 56 days. Measured concentrations at 56 days were <1.35 (control), 51.5, 128, 235, 543, 1,070, 2,020 and 3,990 mg/kg dry soil. There was no effect on survival (day 28). Reproduction (day 56) was significantly reduced at ≥235 mg/kg soil. Study details were excerpted from a secondary source. Aufderheide et al., 2003 as cited in ECB, 2008 56

- day NOEC 128 mg/kg dry soil

Plants Corn ( Zea mays ), cucumber ( Cucumis sativa ), onion (

Allium cepa ), ryegrass ( Lolium perenne ), soybean ( Glycine max ), tomato ( Lycopersicon esculentum ) Soil exposed 21

- day NOEC >6,200 mg/kg soil Nominal test concentrations were 40, 105, 276, 725, 1,904, and 5,000 mg/kg. Mean measured levels were 0 (control), 31.3, 97.8, 297, 764, 2,230, and 6,200 mg/kg dry soil. No effects on seedling emergence, survival, shoot dry weight, or height were observed at any concentration in any species. Study details were excerpted from a secondary source. Porch et al., 2002 as cited in ECB, 2008)

a pH and water hardness were within accepted ranges (6

– 8.5 and 50

– 250 mg/L as CaCO 3 , respectively) unless noted otherwise. b Experiment type: S = static, R = renewal, F = flow

- through, M = measured, U = unmeasured, NS = not specified. c mg/L = ppm , units are in terms of HBCD concentration
.115 d Test composition was an isomeric mixture of HBCD (8.5% beta isomer, 6% alpha isomer, and 79.1% gamma isomer). e Solvent: d imethylformamide (DMF). f Solvent: Cremophor RH40. g Solvent: Tween80. h Replicates per concentration were not used. i Dissolved oxygen concentrations decreased to a minimum level of 1.9 mg O/L. j Insufficient acclimatization of test organism. k Solvent: DMSO. l Concentrations in mg/L were calculated from molar concentrations using a molecular weight of 641.7 g/mole. m Test substance was an isomeric mixture of HBCDs (5.15% beta isomer, 7.67% alpha isomer, and 83.04% gamma isomer). n A single concentration was tested
. o Saturated solution was tested. p Solvent: acetone. q Test substance was an isomeric mixture of HBCDs (19.3% beta isomer, 5.8% alpha isomer, and 74.9% gamma isomer). r Solvent: ethyl acetate. s Sediment of a coarser grain size. t Test composition was an isome ric mixture of HBCD (6.3% beta isomer, 9.4% alpha isomer, and 84.3% gamma isomer). u Solvent: peanut oil. v Test substance was an isomeric mixture of HBCDs (30% beta isomer, 27% alpha isomer, and 43% gamma isomer). w Test substance was HBCD technical mixture (HBCD

- TM), composed of 6% alpha isomer, 8% beta isomer, and 80% gamma isomer. x Vehicle: safflower oil y Test substance was technical HBCD (HBCD

- TM; 11% beta isomer, 12% alpha isomer, and 77% gamma isomer) initially dissolved in acetone to increase its solub ility in organic safflower oil; acetone was then evaporated under nitrogen gas to yield a safflower oil/HBCD

- TM solution. Abbreviations: DMSO = dimethyl sulfoxide; EROD = ethyoxyresorufin

- O

- deethylase; GLP = good laboratory practice; Hsp70 = h eat shock protein; LOAEL = lowest

- observed

- adverse

- effect

- level; LOEC = lowest

- observed

- effect

- concentration; LOQ = limit of quantification; MATC = maximum acceptable toxicant concentration; NOAEL = no - observed

- adverse

- effect

- level; NOEC = no

- observed

- effe ct
- concentration; NQ = not quantified; PROD = pentoxyresorufin

O

- deethylase; ROS = reactive oxygen species; SV

- BA = sinus venosus

- bulbus arteriosus; T4 = thyroxine; TOC = total organic carbon; ww = wet weight116

4.2.1. Acute Aquatic Toxicity. Acute aquatic to xicity data for HBCD are presented in Table 4

- 1. Adverse effects observed following acute exposure were found in studies with marine algae, including the EPA

- recommended estuarine/marine algae species S. costatum (Desjardins et al., 2004 as cited in ECB, 2 008 and IUCLID, 2005; Desjardins et al., 2005 as cited in ECB, 2008; Walsh et al., 1987), a series of short

- term (72

– 120

- hour) early life stage tests with zebrafish embryos performed in China (Deng et al., 2009; Du et al., 2012 b ; Hu et al., 2009; Wu et a l., 2013), and short

- term (72

- hour) results from an early life stage test with sea urchin embryos (Anselmo et al., 2011). Effects in these studies, reported at concentrations as low as 0.009 mg/L (measured) in algae, 0.01 mg/L (nominal) in zebrafish embryo s, and 0.064 mg/L (nominal) in sea urchin embryos, indicate high acute hazard concern. Walsh et al. (1987) reported measured 72

- hour EC 50 values in S. costatum ranging from 0.009 to 0.012 mg/L based on reduced growth rate in five different types of saltw ater media (0.010 mg/L in seawater itself). The study tested two other marine algal species, Chlorella sp. and Thalassiosira pseudonana , that were also found to be inhibited by HBCD, albeit at higher concentrations than S. costatum . EC 50 values for reduced growth in these species were 0.05

– 0.37 mg/L (0.08 mg/L in seawater) for T. pseudonana and >1.5 mg/L for Chlorella sp . In these assays, maximum test concentrations were selected by determining the maximum achievable test concentration (in acetone) that d id not result in crystal formation, and multiple concentrations were tested. EC 50 values were based on measured concentrations. Subsequent studies by Desjardins et al. (2004 as cited in ECB, 2008 and IUCLID, 2005) confirmed the high acute toxicity of HBCD to S. costatum. In these studies, single concentrations were tested, but the assays were conducted without solvent and the conce ntrations were measured. Desjardins et al. (2004 as cited in ECB, 2008 and IUCLID, 20 05 ) reported117

approximately10% inhibition of growth in S. costatum exposed to 0.041 mg/L for 72 hours. Desjardins et al. (2005 as cited in ECB, 2008) found that a saturated solution of 0.0545 mg/L resulted in 51% growth inhibition after 72 hours of exposure. The latter result corresponds to an approximate EC 50 of 0.052 mg/L. The zebrafish embryo studies reported a variety of effects on embryos and larvae at low HBCD concentrations. These studies were performed using DMSO solvent, and results are based on nominal concentrations, as actual concentrations were not measured. The actual effective concentrations are likely to be lower than the nominal concentrations indicat ed below. In the Deng et al. (2009) study, developmental toxicity endpoints were assessed at 96 hours post

- fertilization in embryos/larvae exposed to HBCD starting 4 hours post

fertilization. Survival of embryos/larvae was significantly reduced at all tes ted concentrations, making the low concentration of 0.05 mg/L the LOEC in this study; a NOEC was not established. Embryonic malformation rate was significantly increased and larval growth significantly decreased at ≥0.1 mg/L. Malformations included epiboly deformities, yolk sac and pericardial edema, tail and heart malformations, swim bladder inflation, and spinal curvature. Embryo hatching rate was reduced only at the high concentration of 1 mg/L. Heart rate, a marker for cardiac developmental toxicity, wa s significantly decreased at all tested concentrations. Associated mechanistic studies suggest that the mechanism for developmental toxicity involves the generation of reactive oxygen species (ROS) and the consequent triggering of apoptosis genes. Increase d ROS formation (indicative of oxidative stress) was observed at a nominal concentration of 0.1 mg/L. In the same study, zebrafish embryos exposed to HBCD exhibited increased expression of pro

- apoptotic genes (Bax, P53, Puma, Apaf

- 1, caspase

- 3, and caspase

- 9),118

decreased expression of anti

- apoptotic genes (Mdm2 and Bcl

- 2), and increased activity of enzymes involved in apoptosis (caspase

- 3 and caspase

- 9) with LOECs of 0.05

– 1 mg/L. In contrast to the results of Deng et al. (2009), Hu et al. (2009) found no effects on survival or malformation rates in zebrafish embryos exposed to concentrations up to 10 mg/L (highest concentration tested) for 96 hours, starting 2 hours post

- fertilization. Hatching was delayed at 0.002 mg/L, the lowest concentration tested, an d other concentrations up to and including 0.5 mg/L, but not the two high concentrations of 2.5 and 10 mg/L. Hatching success was not affected, so the relevance of the observed hatching delay at lower concentrations is uncertain. The same authors observed an increase in heat shock protein (Hsp70) at 0.01 mg/L and an increase in malondialdehyde activity, used as a measure of lipid peroxidation, at 0.5 mg/L. The activity of superoxide dismutase was increased at 0.1 mg/L, but decreased at 2.5 and 10 mg/L. Hu e t al. (2009) concluded that HBCD can cause oxidative stress and overexpression of Hsp70 in acute exposures of zebrafish embryos. Similar to the results seen by Hu et al. (2009), Wu et al. (2013) observed no effects on survival or embryo malformation rates, but also no effect on hatching, in a 72

- hour toxicity test of zebrafish embryos exposed within 1 hour after fertilization to nominal concentrations up to 200 nM HBCD (0.13 mg/L; highest concentration tested). In contrast to the decrease in heart rate obse rved by Deng et al. (2009), heart rate was increased after 48 hours of exposure to 200 nM HBCD (0.13 mg/L) and after 60 hours of exposure to 2 nM HBCD (0.0013 mg/L). The same authors found cardiac arrhythmia in embryos exposed to 20 nM HBCD (0.013 mg/L) fo r 72 hours. Corresponding to the observed cardiac arrhythmia and increased heart rate, Wu et al. (2013) found increased expression of genes involved in conduction system development and cardiac morphogenesis (Tbx5 and Nkx2.5) and calcium ion transport in c ardiomyocytes119

(Atp2a2b and RyR2) at concentrations of 0.0013 and 0.13 mg/L, respectively. However, again in contrast to the results of Deng et al. (2009), there was no increase in apoptotic cells around the heart or caspase

- 3 activity. Du et al. ( 2012 b ) ex posed zebrafish embryos 4 hours post

- fertilization to each of three diastereomers of HBCD (α

- , β

- , and γ

- HBCD) individually at nominal concentrations of 0.01, 0.1, and 1.0 mg/L. Hatching success was reduced after 68 hours of exposure to γ

- HBCD at the lowes t concentration (0.01 mg/L), but a higher concentration of α

- or β

- HBCD (0.1 mg/L) was necessary to reduce hatching success. After 92 hours, survival was reduced at concentrations of 0.01, 0.1, and 1 mg/L of γ

- , β

- , and α

- HBCD, respectively. Growth, measur ed as body length of larvae after 92 hours of exposure, was reduced at 0.1 mg/L of β

- and γ

- HBCD and at 1 mg/L of
α - HBCD. After 116 hours of exposure, malformations were observed at all test concentrations of β

- and γ

- HBCD and at 0.1 mg/L and above for α

- H BCD. Effects on heart rate varied depending upon the length of exposure; reduced heart rate was observed at 0.1 mg/L of β

- and γ

- HBCD or 1 mg/L of α

- HBCD at 44 hours and at 0.1 mg/L of α

- and β

- HBCD at 92 hours, whereas γ - HBCD resulted in an increase in he art rate at 1 mg/L at 92 hours. An increase in generation of ROS was observed after 116 hours at 0.1 mg/L of β

- and γ

- HBCD and at 1 mg/L of α

- HBCD. Activities of caspase

- 3 and caspase

- 9 enzymes, indicative of apoptosis, were increased after 116 hours at 0. 1 mg/L of γ

- HBCD and at 1 mg/L of α

- and β

- HBCD. Du et al. ( 2012 b ) ranked the HBCD diastereomers in the following order for developmental toxicity to zebrafish: γ

- HBCD > β

- HBCD > α

- HBCD. Effects indicative of oxidative stress, as seen in the zebrafish emb ryo studies, were also found in clams. Zhang et al. (2014a) measured parameters indicative of antioxidant defenses and oxidative stress after 1, 3, 6, 10, and 15 days of exposure to low nominal concentrations of120

HBCD ranging from 0.000086 to 0.0086 mg/L in the clam Venerupis philippinarum . Increases in ethyoxyresorufin

- O

- deethylase (EROD) activity, glutathione (GSH) content, and DNA damage were observed in clams exposed to 0.00086 mg/L, while increased lipid peroxidation (LPO) was observed at 0.0086 mg/L. T hese same effects were observed at lower concentrations as the length of exposure increased. Results observed at the end of the 15

- day exposure are described in the chronic toxicity section. Anselmo et al. (2011) exposed sea urchin ( Psammechinus miliaris ) embryos to HBCD in an early life stage test. Newly

- fertilized embryos were exposed to HBCD at nominal concentrations of 0, 9, 25, 50, and 100 nM (0, 0.0058, 0.016, 0.032, and 0.064 mg/L, respectively) in DMSO solvent and evaluated at 72 hours post

- fertili zation. Concentrations were not measured. There was no effect on hatching success of embryos at any concentration. A significant increase in morphological abnormalities was found at a nominal concentration of 100 nM HBCD (0.064 mg/L), the highest concentra tion tested. Observed malformations included short or deformed larval arms and slight edema around the larval body. The NOEC for this effect at 72 hours was 0.032 mg/L. Results at later time points are described in the chronic toxicity section. Experimen tal studies conducted following EPA

- recommended study guidelines in freshwater fish, Daphnia , and green algae mostly reported no lethal or sublethal effects at concentrations approaching saturation (Calmbacher, 1978; Graves and Swigert, 1997a, b; Roberts a nd Swigert, 1997; Siebel

- Sauer and Bias, 1987 as cited in ECB, 2008). Jatzek (1990) reported a LOEC of 10 mg/L and an EC 50 of 146.34 mg/L for Daphnia immobility based on nominal test concentrations that greatly exceeded measured water solubility values for HBCD; however, because the appearance of the test solutions was not reported, it is possible that121

reduction in daphnid swimming was related to the presence of insoluble material at higher concentrations. No data were found for acute toxicity of HBCD to sa ltwater fish species. ECOSAR

- predicted 96

- hour LC 50 values for estuarine/marine fish were above the water solubility limit 13

. 4.2.2. Chronic Aquatic Toxicity. Chronic aquatic toxicity data are presented in Table 4

- 1. A high concern for hazard was determined based on reduced size (length) of surviving young daphnids, resulting in a measured MATC of 0.0042 mg/L (Drottar and Krueger, 1998). The study followed EPA

- r ecommended guidelines and used a flow

- through test system. This study reported additional effects, including decreased reproductive rate and decreased mean weight of surviving young at 0.011 mg/L. Mortality of adult daphnids in HBCD treatment groups was no t significantly different from control mortality. A chronic study in rainbow trout by the same researchers, conducted following EPA recommended guidelines, found no effects at concentrations approaching saturation (Drottar et al., 2001). Other effects re ported following chronic exposure to HBCD included degenerative changes in the gills of clams ( Macoma balthica ), manifested by the increased frequency of nuclear and nucleolar abnormalities and the occurrence of dead cells, at nominal concentrations of ≥0. 1 mg/L (50

- day LOEC) (Smolarz and Berger, 2009), a nominal MATC of 0.045 mg/L for increased morphological abnormalities in sea urchin ( P. miliaris ) embryos exposed to HBCD for up to 16 days in an early life stage test (Anselmo et al., 2011), and a nominal MATC of 0.03 mg/L for increased malformation rate in marine medaka ( Oryzias melastigma ) embryos exposed

13 For the ECOSAR predictive tool, values that exceed a K ow cut
- off or water solubility limit are outside the domain of the ECOSAR class training data sets. Thus, even though experimental values that exceed the limit of solubility are considered in this hazard assessment, predicted values above the limit of solubi lity would be considered irrelevant due to the limitations of the predictive tool.122

to HBCD for 17 days in an early life stage test (Hong et al., 2014a). The developmental abnormalities in medaka included yolk sac edema, pericardial ede ma, and spinal curvature (Hong et al., 2014a). Mechanistic findings in this study included increases in heart rate and sinus
venosus - bulbus arteriosus (SV

- BA) distance, which are markers for cardiac development, induction of oxidative stress and apoptosis, and suppression of nucleotide and protein synthesis. In contrast, Foekema et al. (2014) observed no effects on mortality or development through metamorphosis (approximately 40 days post

- fertilization) in sole ( Solea solea ) embryos exposed in an early li fe stage test to concentrations of HBCD up to 0.25 mg/L for 6 days, starting at 12 hours post

- fertilization. Thyroid effects were reported in juvenile rainbow trout ( Oncorhynchus mykiss ) following dietary exposure to HBCD (Palace et al., 2008, 2010). Eac h of the diastereomers of HBCD (administered separately via diet at concentrations of 5 ng/g of α

- , β

- , or γ - HBCD for up to 56 days) disrupted thyroid homeostasis, as indicated by lower free circulating T3 and T4 levels (Palace et al., 2008, 2010). No thy roid or other effects were observed in European flounder ( Platichthys flesus ) following 78 days of diet or sediment exposure to maximum concentrations of 3,000 μg/g lipid in food and 8,000 μg/g total organic carbon (TOC), respectively (Kuiper et al., 2007) . In two sediment toxicity tests (Thomas et al., 2003a, b) with Hyalella azteca exposed to spiked sediment in the presence of 2 or 5% TOC, no effects were seen at the highest dose tested (1,000 mg/kg dry weight sediment) in the definitive studies (reduced survival was reported at 500 mg/kg sediment in the range

- finding study). Number of worms, Lumbriculus variegatus , was significantly reduced at ≥28.7 mg/kg dry weight sediment (measured) in a 28

- day study (Oetken et al., 2001 as cited in ECB, 2008).123

The me chanisms of the effects on fish and invertebrates following chronic exposure were similar to those found in acute studies. Effects observed in fish include increased formation of ROS resulting in oxidative damage to lipids, proteins, and DNA, decreased ant ioxidant capacities in fish tissue (e.g., brains, hepatocytes, or erythrocytes), and increasing levels of EROD (detoxification enzyme) and pentoxyresorufin

- O

- deethylase (PROD, detoxification enzyme) levels in hepatocytes of fish exposed to the nominal conc entration of ≥0.1 mg/L (corresponds to ~0.2 mg/g whole fish [wet weight]) for 42 days (Zhang et al., 2008). Ronisz et al. (2004) found a significant increase in hepatic cytosolic catalase activity in rainbow trout ( O. mykiss ) 5 days after a single intraper itoneal injection of 50 mg/kg was administered. The same authors observed reductions in liver somatic index (LSI) and EROD activity in a 28

- day study in which O. mykiss were injected intraperitoneally with HBCD on days 1 and 14 at a dose somewhat less than 500 mg/kg. Zhang et al. (2014a) observed the following signs of oxidative stress in clams ( V. philippinarum ) after 15 days of exposure to HBCD: t he activities of antioxidant enzymes (EROD, superoxide dismutase [SOD], and glutathione

- S

- transferase [GST]), as well as GSH content, were increased at 0.000086 mg/L, the lowest concentration tested. In addition, LPO was increased at 0.00086 mg/L and DNA damage was increased at 0.0086 mg/L.
4.2.3. Aquatic Toxicity Predictions. No laboratory data were found on ac ute toxicity of HBCD to estuarine/marine fish, and limited data were found on acute toxicity to estuarine/marine invertebrates and chronic toxicity to estuarine/marine fish and invertebrates. Therefore, EPA’s Ecological Structural Activity Relationship Pro gram (ECOSAR, 2012) was used to determine if adequate aquatic toxicity values could be predicted for estuarine/marine invertebrates and fish.124

HBCD was classified into the Neutral Organics ECOSAR class. The predicted 96

- hour LC 50 values for estuarine/marine fish and mysid shrimp and the chronic value (ChV) for estuarine/marine fish were greater than the water solubility limit 14 . The predicted ChV for mysid shrimp was 0.0007 mg/L, indicating very high expected toxicity. The values o f the physicochemical parameters (log K ow = 5.625; water solubility = 0.0086 mg/L) used for predicting the aquatic toxicity of HBCD are presented in Chapter 2 (Chemistry and Environmental Fate) and the predicted aquatic toxicity values are depicted in the ECOSAR output in Attachment B.
4.2.4. Terrestrial Toxicity and Phytotoxicity. Terrestrial toxicity and phytotoxicity data are presented in Table 4

- 1. Japanese quail exposed for 6 weeks to an isomeric mixture of HBCD in the diet experienced a reduction in hatchability at all tested concentrations (12

– 1,000 ppm) (MOEJ, 2009 as cited in UNEP, 2010). Additional effects included a significant reduction in egg shell thickness starting at 125 ppm, decreases in egg weights and egg production rates starting at 50 0 ppm, increases in cracked eggs starting at 500 ppm, and adult mortality at 1,000 ppm. A subsequent test, conducted at lower dietary concentrations, determined lowest

- observed

- adverse

effect

- level (LOAEL) and no

- observed

- adverse

- effect

- level (NOAEL) value s of 15 and 5 ppm, respectively, based on significant reduction of survival of chicks hatched from eggs of HBCD

fed quails (MOEJ, 2009 as cited in UNEP, 2010). Several studies have been conducted examining effects of HBCD on American kestrels ( F. sparveri us ). Kobiliris (2010) reported a reduced “corticosterone response” (where

14 For the ECOSAR predictive tool, values that exceed a K ow cut
- off or water solubility limit are outside the domain of the ECOSAR class training data sets. Thus, even though experimental values that exceed the limit of solubility are considered in this hazard assessment, predicted values above the limit of solubility would be considered irrelevant due to the limitations of the predictive tool.125

“corticosterone response” was defined as a stimulation of the adrenal cortex to produce and release corticosterone into the bloodstream), reduced flying activities of juvenile males during hunting behavior trials, and delayed response times of juvenile females during predator avoidance behavior trials in American kestrels exposed in ovo to 164.13 ng/g wet weight. Kestrels exposed via the diet to 0.51 mg/kg body weight

- day beginning 3 weeks prior to pairing and continuing until the first chick hatched began to lay eggs 6 days earlier than controls and laid larger clutches of smaller eggs (Fernie et al., 2011). Although the technical mixture of HBCD stereoisomers contained predominantly γ - HBCD (80% of the mixture), the main isomer found in eggs was α - HBCD (>90% of the total HBCD in eggs). In a subsequent study, Marteinson et al. (2011) exposed kestrels to dietary HBCD at the same dose (0.51 mg/kg kestrel

- day) and found increased testes we ight in unpaired males, a marginally significant effect on testis histology in unpaired males (increased number of seminiferous tubules containing elongated spermatids; p = 0.052), marginally increased testosterone levels in breeding males (increased at th e time the first egg was laid; p = 0.054), and no significant effect on sperm counts. Plasma T4 levels were reduced in breeding males throughout the study, which the authors took to suggest that thyroid disruption that may have contributed to the observed increase in testes weight. Marteinson et al. (2012) found altered reproductive behavior in both sexes of kestrels fed 0.51 mg/kg body weight

- day, including reduced activity in both sexes during courtship and in males during brood rearing, which may have co ntributed to the observed reduction in incubation nest temperature and also to the reduced egg size reported previously by Fernie et al. (2011). In a 22

- day study of chickens ( Gallus gallus domesticus ) exposed to HBCD in ovo , reduced pipping success was ob served at 100 ng/g egg (Crump et al., 2010).126

Aufderheide et al. (2003) reported LOEC and NOEC values of 235 and 128 mg/kg dry soil, respectively, based on reproductive effects in earthworms following 56 days of exposure. There was no effect on worm surviv al at soil concentrations up to 3,990 mg/kg dry soil (measured), the highest level tested. Studies using monocot and dicot plant species exhibited no toxicity up to the maximum test concentration of 6,200 mg/kg soil (measured) (Porch et al., 2002 as cited in ECB, 2008). The accumulation and toxicity of α

- , β

- , and γ

- HBCDs in maize h ave been studied (Wu et al., 2012b). The order of accumulation in roots was β

- HBCD > α

- HBCD > γ - HBCD and in shoots it was β

- HBCD > γ

- HBCD > α

- HBCD. In maize exposed to 2 μg/L HBC D, the inhibitory effect of the diastereomers on the early development of maize as well as the intensities of hydroxyl radical and histone H2AX phosphorylation followed the order α

HBCD > β

- HBCD > γ - HBCD, which indicates diastereomer

- specific oxidative stress and DNA damage in maize. The study confirmed that for maize exposed to HBCDs, the generation of reactive oxygen species was one, but not the only, mechanism for DNA damage.
4.2.5. Population Effects. HBCD was first detected in fish and sediments in Swedish rivers in 1995 (Sellström et al., 1998). Prior to that, HBCD had only been found in environmental samples in Japan (Watanabe and Tatsukawa, 1989). Although much monitoring data are available indicating the occurrence of HBCD in biota over large areas and in remote locations, studies are not available that directly characterize effects observed in exposed populations in relation to environmentally

- monitored HBCD levels , however, interspecific plant interactions may enhance the bioavailability of HBCDs (Li et al., 2011)
. HBCD has been widely detected in biota and sediment in the northern hemisphere; most studies of HBCD occurrence in the environment have been conducted in Europe and Asia127

(Marv in et al., 2011; Klosterhaus et al., 2012 ; Law et al., 2014 ). Wu et al. (2012a) found higher levels of HBCD (by at least one order of magnitude) in wildlife near brominated flame retardant production facilities and electronic waste recycling sites in China than in other parts of the country. Temporal trends in HBCD concentrations are not as clear. Wu et al. (2012a) suggested that as polybrominated diphenyl ethers (PBDEs) and other flame retardants have been phased out of production, the use of HBCD (and sub sequent environmental concentrations of HBCD) has increased. Some concentrations of HBCD in wildlife populations are presented in Table 4

- 2. In general, concentrations reported in wildlife represent levels at which HBCD has accumulated in tissue, whereas laboratory environmental toxicity studies reported effects related to exposure concentrations.128

Table 4 - 2. Environmental monitoring data for HBCD

Species Concentration a Comments Reference Aquatic invertebrates Green mussels ( Perna viridis ) and blue mussels ( Mytilus edulis ) ND b

– 1,400 ng/g lw ƩHBCD c Coastal waters of Asia (Cambodia, China, Hong Kong, India, Indonesia, Japan, Malaysia, the Philippines, and Vietnam) Isobe et al., 2012
Mussels ( Mytilus spp. ) 0.27 ng α - HBCD/g dw (range: 0

– 1.5 ng α - HBCD/g dw); 0.068 ng β

- HBCD /g dw (range: 0

– 0.4 ng β

- HBCD /g dw ); 0.69 ng γ

- HBCD/g dw (range: 0

– 2.5 ng γ

- HBCD/g dw) California coast Dodder et al., 2014

Fish Common carp ( Cyprinus carpio ), flathead catfish ( Pylodictus olivaris ), channel catfish ( Ictalurus punctatus ), redhorse sucker ( Moxostoma sp. ), and gizzard shad ( Dorosoma cepedianum ) In 1999

– 2002 d : 3.2

– 13 ng/g lw (carp); 7.2 ng/g lw (flathead catfish); 4.2

– 20 ng/g lw (channel catfish); ND

– 20 ng/g lw (sucker); 4.6

– 22 ng/g lw (shad) In 2006

– 2007 : 54

– 4,640 ng/g lw (carp); 698 ng/g lw (flathead catfish); 60

– 3,680 ng/g lw (channel catfish); 16

– 1,400 ng/g lw (sucker); 13

– 290 ng/g lw (shad) Hyco, Dan, and Roanoke Rivers in Virginia and North Carolina (region of textile production) Chen et al., 2011

Eel ( Anguilla anguilla ) and brown trout ( Salmo trutta ) 399.6

– 3,216 ng ƩHBCD/g ww in eel d ; 20.3

– 2341.2 ng ƩHBCD/g ww in trout Downstream of an HBCD manufacturing plant on the River Skerne (Durham, United Kingdom) Allchin and Morris, 2003 Lake trout ( Salmo trutta forma lacustris ) 168 ng/g lw (range: 49

– 324 ng/g lw) Lake Geneva, Switzerland Cheaib et al., 2009 Pelagic saithe ( Pollachius virens ) and demersal cod ( Gadus morhua ) 7.13

– 11.9 9 ng/g lw (saithe); 3.86

– 13.9 ng/g lw (cod) 4 Norwegian coast Bustnes et al., 2012 European eel (

A. anguilla ) 510 ng/g lw (range: 7

– 9,500 ng/g lw) Flanders, Belgium; α

- HBCD was the predominant isomer (average 75%), followed by γ

- HBCD (20%) and β

- HBCD (4%) in almost all eel samples Malarvannan et al., 2014129

Table 4 - 2. Environmental monitoring data for HBCD

Species Concentration a Comments Reference Crucian carp ( Carassius auratus ) Mean 16 ng ƩHBCD/g lw in females (mean 12 ng α - HBCD/g lw, mean 1.4 ng β

- HBCD/g lw, and mean 2.4 ng γ

- HBCD/g lw); Mean 9.7 ng ƩHBCD/g lw in males (mean 7.7 ng α - HBCD/g lw, mean 0.70 ng β

- HBCD/g lw, and mean 1.3 ng γ

- HBCD/g lw); Mean 5.7 ng ƩHBCD/g lw in eggs (mean 4.2 ng α - HBCD/g lw, mean 1.1 ng β

- HBCD/g lw, and mean 0.49 ng γ

- HBCD/g lw) Seomjin, Nam, and Nakdong Rivers, Korea Jeong et al., 2014

Nile tilapia ( Ore ochromis niloticus ) 1.6 and 3.3 ng α

- HBCD/g dw; concentrations of β

- and γ

- HBCD were ND Municipal dumpsite in Surabaya, Indonesia Ilyas et al., 2013 White croaker ( Genyonemus lineatus ) and shiner surfperch ( Cymatogaster aggregata ) Croaker: ND

– 4.5 ng ƩHBCD/g lw, ND

– 4.5 ng α - HBCD/g lw, ND ng β

- HBCD/g lw, ND

– 1.4 ng γ - HBCD/g lw; Surfperch: 2.5

– 24.7 ng ƩHBCD/g lw, 2.5

– 19.6 ng α - HBCD/g lw, ND

– 5.1 ng β

- HBCD/g lw, ND

– 2.5 ng γ

- HBCD/g lw San Francisco Bay Klosterhaus et al., 2012

Eight species of wild fish ( Oxygymnocypris stewartii, Schizopygopsis younghusbandi , Schizothorax macropogon , Schizothorax o’connori , Schizothorax waltoni , Gymoncypris waddellii , Gymoncypris przewalskii , and Racoma tibetanus ) 2.12 ng ƩHBCD/g lw (range: ND

– 13.7); 1.53 ng α - HBCD/g lw; 0.59 ng γ

- HBCD/g lw; β

- HBCD was ND e High mountain lakes and rivers on the Tibetan Plateau, China Zhu et al., 2013

Aquatic mammals Beluga whales ( Delphinapterus leucas ) 10

– 20 ng ƩHBCD/g lw Canadian Arctic Tomy et al., 2008 Harbor porpoise ( Phocoena phocoena ) and common dolphin ( Delphinus delphis ) 100

– 5,100 ng ƩHBCD /g lw (medians; porpoise) and 200

– 900 ng ƩHBCD /g lw blubber (medians; dolphin) Coasts of Scotland, Ireland, The Netherlands, Belgium, France, and Spain Zegers et al., 2005

Indo

- Pacific humpback dolphin ( Sousa chinensis ) and finless porpoise ( Neophocaena phocaenoides ) 32

– 519 ng/g lw (dolphin) and 4.1

– 501 ng/g lw (porpoise) in blubber Hong Kong Jin et al., 2012130

Table 4 - 2. Environmental monitoring data for HBCD

Species Concentration a Comments Reference Striped dolphin ( Stenella coeruleoalba ) 290

– 940 ng/g lw dolphins found stranded in 2003; 10

– 310 ng/g lw from archived dolphin blubber samples, 1978

– 1992 Japan Isobe et al., 2009
Harbor seal ( Phoca vitulina ) 4.4

– 19.3 ng ƩHBCD/g lw in adults, 2.0

– 11.7 ng ƩHBCD/g lw in pups; 2.8

– 19.3 ng α - HBCD /g lw in adults, 1.3

– 10.2 ng α - HBCD /g lw in pups; ND

– 0.4 ng β

- HBCD /g lw in adults, ND

– 0.4 ng β

- HBCD /g lw in pups; ND

– 2.0 ng γ - HBCD /g lw in adults, ND

– 1.6 ng γ - HBCD /g lw in pups San Francisco Bay Klosterhaus et al., 2012

Terrestrial plants Pine ( Pinus spp. ), ginkgo ( Ginkgo biloba ), locust tree ( Robinia pseudoacacia ), weeping willow ( Salix babylonica ), poplar ( Populus spp. ), and cypress ( Cupressus spp. ) 26

– 3,400 ng/g lw ƩHBCD Beijing, China Hu et al., 2011

Full

- grown hardwood and coniferous trees (e.g., pine [ Pinus spp. ], fir [ Abies spp. ], and spruce [ Picea spp. ]) 1.30 ± 0.50 ng/g lw f Birkenes, Norway Salamova and Hites, 2013 1.39 ± 0.75 ng/g lw f Bukit Kototabang, Indonesia 1.45 ± 0.78 ng/g lw f Cape Grim, Tasmania 0.66 ± 0.16 ng/g lw f De Aar, South Africa 0.35 ± 0.10 ng/g lw f Dhulikhel, Nepal 21.3 ± 7.7 ng/g lw f Downsview, Toronto, Ontario, Canada 0.84 ± 0.13 ng/g lw f Frasardale, Ontario, Canada 1.65 ± 0.94 ng/g lw f Košetice, Czech Republic 0.62 ± 0.09 ng/g lw f Reykjavik, Iceland 0.76 ± 0.65 ng/g lw f Tula, American Samoa, United States 0.27 ± 0.03 ng/g lw f Whistler, British Columbia, Canada ND Malin Head, Ireland131

Table 4 - 2. Environmental monitoring data for HBCD

Species Concentration a Comments Reference Birds Peregrine falcon ( Falco peregrinus ) and sparrow hawk (

Accipiter nisus ) 71

– 1,200 ng/g lw (peregrine falcon eggs); 84

– 19,000 ng/g lw (sparrow hawk muscle) De Boer et al., 2004 as cited in UNEP, 2010 Peregrine falcon ( F. peregrinus ) 0.02 ng/g ww geometric mean in nestlings (range: 0

– 0.03 ng/g ww) Canadian Great Lakes Basin Fernie and Letcher, 2010 Peregrine falcon ( F. peregrinus ) Canada: 230

– 15,000 ng ƩHBCD/g lw; 170

– 15,000 ng α

- HBCD/g lw; 1.9

– 39 ng γ

- HBCD/g lw; Spain: 0.9

– 1,600 ng ƩHBCD/g lw; 47

– 1,600 ng α - HBCD/g lw; 0.9

– 11 ng γ

- HBCD/g lw Fail

- to
- hatch eggs were coll ected from active nests in the Canadian Great Lakes

- St. Lawrence River and New Brunswick, Canada; In and from active nests in Guadalajara and Bilbao, Spain Guerra et al., 2012

Ivory gull ( Pagophila eburnea ) 14

– 272 ng/g lw in eggs g The Norwegian and Russian Arctic Miljeteig et al., 2009 Bald eagle ( Haliaeetus leucocephalus ) 0.13 ± 0.07 ng/g ww in nestlings (average ± standard error) Great Lakes watershed Venier et al., 2010 Herring gull ( Larus argentatus ), glaucous

- winged gull ( Larus glaucescens ), California gull ( Larus californicus ), and ring

- billed gull ( Larus delawarensis ) 2.2

– 9.0 ng/g ww arithmetic mean (marine ecosystem); 6.7

– 16.6 ng/g ww arithmetic mean (non

- Great Lakes freshwater ecosystem); 2

– 12 ng/g ww median (Great Lakes ecosystem) N esting colonies spanning Atlantic to Pacific Canada Chen et al., 2012

Barn owl ( Tyto alba ) 8.61

– 28.8 ng α

- HBCD/g lw in muscle, liver, preen gland, and adipose tissue; ND β

- HBCD; ND

– 1.97 ng γ - HBCD/g lw Flanders, Belgium and Normandy, France Eulaers et al., 2014

Black

- tailed gull ( Larus crassirostris ), domestic pigeon ( Columba livia var. domestica ), Pacific loon ( Gavia pacifica ), grey heron (

Ardea cinerea ), little egret ( Egretta garzetta ), and great egret ( Ardea alba ) 408

– 2,300 ng/g lw (black

- tailed gull); ND

– 112 ng/g lw (pigeon); 333

– 3,130 ng/g lw (loon); 48

– 11,800 ng/g lw (herons and egrets) Urbanized coastal region of South Korea Hong et al., 2014b

Double

- crested cormorant ( Phalacrocorax auritus ) 21.6

– 39.0 ng ƩHBCD/g lw in eggs (21.6

– 39.0 ng α - HBCD/g lw, ND

– 1.1 ng β

- HBCD/g lw, ND

– 2.7 ng γ - HBCD/g lw) San Francisco Bay Klosterhaus et al., 2012
Great blue heron (

Ardea herodias ) and double

- crested cormorant ( P. auritus ) 1.7

– 15.5 ng/g ww in eggs (herons), 0.57

– 0.77 ng/g ww in eggs (cormorants) Salish Sea region of the west coast of North America Miller et al., 2015132

Table 4 - 2. Environmental monitoring data for HBCD

Species Concentration a Comments Reference Light

- vented bulbul ( Pycnonotus sinensis ), long

- tailed shrike ( Lanius schach ), and oriental magpie

- robin ( Copsychus saularis ) e - Waste locations h : ƩHBCD = 5.1

– 73 ng/g l w α - HBCD = 5.1

– 48 ng/g lw β

- HBCD = ND γ - HBCD = ND

– 1.5 ng/g lw Urban locations: ƩHBCD = 15

– 380 ng/g lw α - HBCD = 13

– 290 ng/g lw β

- HBCD = ND γ - HBCD = ND Rural locations: ƩHBCD = 2.8

– 16 ng/g lw α - HBCD = 2.7

– 15 ng/g lw β

- HBCD = ND γ - HBCD = ND

– 2.5 ng/g lw E

- waste, urban, and rural locations in the Pearl River Delta, South China Sun et al., 2012

Mammals Arctic fox ( Vulpes lagopus ) ND

– 36 ng/g lw in liver i Svalbard, Norway Andersen et al., 2015 Polar bear ( Ursus maritimus ) ND

– 109 ng/g lw Arctic region: Bering

- Chukchi Sea (Alaska), Amundsen Gulf, Western Hudson Bay, Foxe Basin/Gulf of Boothia, Lancaster Sound/Jones Sound, Northeastern Baffin Island, Southeastern Baffin Island (Canada), Scoresbysund (East Greenland), and Svalbard (Norway) Mu ir et al., 2006133

Table 4 - 2. Environmental monitoring data for HBCD

Species Concentration a Comments Reference Polar bear ( U. maritimus ) 0.9

– 41.1 ng/g lw j Alaska, Canada, East Greenland, and Svalbard McKinney et al., 2011 Ringed seal ( Pusa hispida ) ND

– 18.1 ng/g lw α

- HBCD East Greenland Vorkamp et al., 2011

a ww = wet weight; lw = lipid weight; dw = dry weight b ND = not detected c ƩHBCD = summed concentration of α

- , β

- , and γ

- isomers. d Values are the range of mean values for different locations. e HBCD values were not provided for each species. The mean and range values are composite value s of the eight species. f Values are mean ± standard error. g Values are the range of median values for different locations. h Values are the range of median values for different species. i HBCD was detected in 2 of 141 animals. j Range is based on geometric mean values of 11 subpopulations.134

4.3. BIOACCUMULATION POTENTIAL Bioaccumulation in aquatic organisms can be estimated using parameters such as BCFs, bioaccumulation factors (BAFs), biota

- sediment accumulation factors (BSAFs), BMFs, trophic magnification factors (TMFs), and assimilation efficiencies. BCFs, BAFs, BSAFs, and assimilation efficiencies describe the degree to which a chemical can be concentrated in the tissues of an organism. BCFs account for exposure from the ambient environment only (typically water) through an organism’s respiratory and dermal surfaces; BA Fs account for absorption from all routes of exposure, including diet, in the natural environment (Arnot and Gobas, 2006); BSAFs are BAFs for organisms living in sediment; and assimilation efficiency represents an estimate of the fraction of the BAF that m ay be accumulated from the diet. A BCF, BAF, or BSAF greater than 1 implies bioaccumulation from the environment. A BMF expresses the accumulation of a contaminant in a given species from the species in the trophic level below it and is calculated as the concentration of a contaminant in the organism divided by the concentration of the contaminant in its prey; this is similar to a TMF, which is the average rate of increase of a contaminant from one trophic level to the next in a food web. A BMF or TMF grea ter than 1 indicates potential for biomagnification through the food chain, such that a predator species would have higher body burdens of the chemical than its prey species. Table 4

- 3 shows data from studies that quantitatively evaluated the bioaccumulat ion or biomagnification potential of HBCD.135

Table 4 - 3. Bioaccumulation studies for HBCD Species Duration and test endpoint Value Reference Cyanobacteria ( Spirulina subsalsa ) 7

- day BCF a,b α - HBCD: 350 β

- HBCD: 270 γ - HBCD: 174 Zhang et al., 2014c

Green algae ( Scenedesmus obliquus ) α - HBCD: 407 β

- HBCD: 469 γ - HBCD: 390 Rainbow trout ( Oncorhynchus mykiss ) 35

- day BCF a,c 8,974 d,e,f and 13,085 d,e,f Drottar and Krueger, 2000 Fathead minnow ( Pimephales promelas ) 32

- day BCF 18,100 f,g Veith et al., 1979 Mirror carp ( Cyprinus carpio morpha noblis ) 30

- day exposure and 30

- day depuration BCF h α - HBCD: 5,570

– 11,500 β

- HBCD: 187

– 642 γ - HBCD: 221

– 584 Zhang et al., 2014b Mud carp ( Cirrhinus molitorella ), nile tilapia ( Tilapia nilotica ), and suckermouth catfis h ( Hypostomus plecostomus ) Log BAF i.j 4.8

– 7.7 for HBCD isomers (α

- HBCD had higher BAFs than β

- and γ

- HBCD) He et al., 2013

Freshwater food web k Log BAF l,m α - HBCD: 2.58

– 6.01 β

- HBCD: 3.24

– 5.58 γ - HBCD: 3.44

– 5.98 ƩHBCDs: 2.85

– 5.98 i Wu et al., 2011

Ragworm ( Hediste diversicolor ) 28

- day BSAF n,o ≤0.006 Haukås et al., 2010b 28

- day assimilation efficiency 67% Bivalve ( Corbicula fluminea ) Log BSAF n,p 0.4

– 0.9 (α

- and β

- HBCD),

- 0.5

– 0 (γ

- HBCD) La Guardia et al., 2012 Log BAF 5

– 5.3 (α

- and β

- HBCD), 4.6

– 4.8 (γ

- HBCD) Gastropod ( Elimia proxima ) Log BSAF n,p 0

– 0.6 (α

- and β

- HBCD),

- 1.5

–

- 0.3 (γ

- HBCD) Log BAF 4.2

– 5 (α

- and β

- HBCD), 3.2

– 4.5 (γ

- HBCD) Benthivorous barbel ( Barbus graellsii ) BSAF 0.1

– 1.44 van Beusekom et al., 2006 Pelagic bleak (

Alburnus alburnus ) 0.14

– 1.23 Rainbow trout ( O. mykiss ) 4

- day assimilation efficiencies β

- HBCD: >100% α - HBCD: 78% γ - HBCD: 46% Haukås et al., 2009
Rainbow trout ( O. mykiss ) 52

- day BMF q α - HBCD: 9.2 β

- HBCD: 4.3 γ - HBCD: 7.2 Law et al., 2006136

Table 4 - 3. Bioaccumulation studies for HBCD Species Duration and test endpoint Value Reference Zebrafish ( Danio rerio ) 42

- day exposure and 21

- day depuration BMF r α - HBCD: 12.33

– 29.71 β

- HBCD: 7.34

– 11.63 γ - HBCD: 7.61

– 7.76 Du et al., 2012 a Assimilation efficiency α - HBCD: 40.84

– 54.67% β

- HBCD: 33.59

– 47.71% γ - HBCD: 29.31

– 40.80% Zebrafish ( D. rerio ) 42

- day exposure and 21

- day depuration BMF s ±α

- HBCD: 0.062

– 0.289 ±β

- HBCD: 0.043

– 0.080 ±γ

- HBCD: 0.034

– 0.067 Du et al., 2013 Assimilation efficiency ±α

- HBCD: 28.99

– 73.71% ±β

- HBCD: 30.82

– 52.54% ±γ

- HBCD: 16.28

– 53.53% Polar bear ( Ursus maritimus ) BMF for α

- HBCD t 1.7 Letcher et al., 2009 Harbor seal ( Phoca vitulina ) BMF for α

- HBCD u 1.6

– 3.0 (alewife [ Alosa pseudoharengus ] to harbor seal); 0.54

– 1.0 (Atlantic herring [ Clupea harengus ] to harbor seal); 0.89

– 1.6 (Atlantic mackerel [ Scomber scombrus ] to harbor seal) Shaw et al., 2012

Light

- vented bulbul ( Pycnonotus sinensis ), long

- tailed shrike ( Lanius schach ), and oriental magpie

- robin ( Copsychus saularis ) BMF v 1.3

– 2.6 for α

- HBCD in e

- waste area; 4.9

– 30 for α

- HBCD in urban are a; 1.1

– 1.4 for α - HBCD in rural area; 0.5 for γ - HBCD in e

- waste area; 2.3 for γ - HBCD in urban area Sun et al., 2012

Common kestrel ( Falco tinnunculus ), eagle owl ( Bubo bubo ), and little owl ( Athene noctua ) BMF w 0.07 (α

- HBCD, sparrow to kestrel); 0.21 (α

- HBCD, rat to kestrel); 5.5 (α

- HBCD, sparrow to owl); 16 (α

- HBCD, rat to owl); 0.28 (γ

- HBCD, sparrow to kestrel) Yu et al., 2013

Coastal marine food web in Norway x BMF for Great black

- backed gull ( Larus marinus ) 3.1

– 487 (α

- HBCD, ww

- based); 2.8

– 23 (α

- HBCD, lw

- based); 0.2

– 2.4 (γ

- HBCD, ww

- based); 0.03

– 8.4 (γ

- HBCD, lw

- based); 2.6

– 132 (ƩHBCD, ww

- based); 1.5

– 5.8 (ƩHBCD, lw

- based) Haukås et al., 2010a

BMF for Common eider ( Somateria mollissima ) 14

– 1,285 (α

- HBCD, ww

based); 0.8

– 26 (α

- HBCD, lw

based); 4.5

– 116 (γ

- HBCD, ww - based); 0.2

– 2.0 (γ

- HBCD, lw

- based); 10

– 425 (ƩHBCD, ww - based); 0.5

– 7.9 (ƩHBCD, lw

- based)137

Table 4 - 3. Bioaccumulation studies for HBCD Species Duration and test endpoint Value Reference

BMF for Shore crab ( Carcinus maenas ) 0.1 (α

- HBCD, ww

- based); 0.2 (α

- HBCD, lw

- based); 0.3 (γ

- HBCD, ww

- based); 0.9 (γ

- HBCD, lw

- based); 0.2 (ƩHBCD, ww

- based); 0.5 (ƩHBCD, lw

- based) TMF y 9.1 (α

- HBCD, ww

- based); 2.6 (α

- HBCD, lw

- based); 0.9 (γ

- HBCD, ww

- based); 0.3 (γ

- HBCD, lw

- based) Arctic marine food web TMF 2.1 (α

- HBCD) and 0.5 (γ

- HBCD) z Tomy et al., 2008 Western Scheldt estuarine food web bb TMF aa 2.2 (α

- HBCD, total food web); 0.3 (γ

- HBCD, total food web); 3.7 (α

- HBCD, benthic food web); 0.2 (γ

- HBCD, benthic food web); 2.0 (α

- HBCD, pelagic food web); 0.3 (γ

- HBCD, pelagic food web) Brandsma et al., 2015

Freshwater food web cc TMF dd 2.22 , 2.18 , and 2.19 for ̶+ α -
, - α - , and ±α

- HBCD, respectively Wu et al., 2010 Marine aquatic food web ee TMF dd 1.74 and 1.68 for α

- HBCD and ƩHBCD, respectively Zhang et al., 2013 Freshwater food web ee 2.58 and 2.44 for α

- HBCD and ƩHBCD, respectively

a BCF calculated at steady state. b BCF calculated using tissue concentration based on wet weight. c BCF calculated using whole

- fish tissue concentrations. d Test substance was an isomeric mixture of HBCDs (4.5% beta isomer, 6.4% alpha isomer, and 79.1% gamma isomer). e BCF based on a measured exposure concentration of 1.8 μg/L (8974) was considered to be more reliable than the BCF based on the 0.18 μg/L concentration (13,085) because water concentrations were maintained more consistently and ste ady state was achieved at the higher concentration. See text for additional discussion. f Solvent used. g A depuration phase was excluded from the study. h BCFs were calculated separately for viscera, gill, skin, and muscle (tissues listed in order of decreas ing BCF). The range of BCFs for these four tissues is provided for each enantiomer. BCFs were calculated by dividing the uptake rate constant by the depuration rate constant. i BAF calculated using muscle tissue concentration based on lipid weight. j BAF cal culated from concentrations of HBCD in samples collected from a river running through a highly industrialized area in South China. k Species included Chinese mysterysnail ( Cipangopaludina chinensis ), prawn ( Macrobrachium nipponense ), mud carp ( C. molitorell a ), crucian carp ( Carassius uratus ), northern snakehead ( Ophicephalus argus ), and water snake ( Enhydris chinensis ). l BAF calculated using concentration of HBCD in whole organisms (wet weight). m BAF calculated from concentrations of HBCD in samples collecte d from a natural pond in an electronic waste recycling site in South China.138

Table 4 - 3. Bioaccumulation studies for HBCD Species Duration and test endpoint Value Reference
n BSAF = lipid

- normalized tissue concentration ÷ TOC

- normalized sediment concentration. o HBCD

- contaminated sediment and food (blue mussels; Mytilus edulis ) samples were collected from a highly contaminated fjord on the west coast of Norway. Ragworms were collected from a control site and exposed for 28 days to contaminated sediment or food. p BSAF calculated from concentrations of HBCD in samples collected from 0 to 44.6 km downstre am of a North Carolina wastewater treatment plant that treats wastewater from a textile manufacturer. q The BMF was calculated from the equation BMF = (AE × F)/k d , where F is the feeding rate on a lipid basis and k d is the depuration rate constant and AE is the assimilation efficiency. r Fish food was spiked with α

- , β

- , or γ - HBCD. s Fish food was spiked with (±)α

- , (±)β

- , or (±)γ

- HBCD. t Adipose tissue, brain, and liver samples were collected from polar bears and blubber tissue was collected from ringed seals ( Pusa hispida ) in East Greenland. Ringed seal is the predominant prey item for polar bears in this region. The BMF of 1.7 was calculated using the adipose tissue concentration in polar bears; HBCD was not detected in brain and liver samples. u BMFs were cal culated using HBCD concentrations in liver and blubber samples collected from harbor seals that had stranded along the northwest Atlantic coast and HBCD concentrations in three prey species (alewife, Atlantic herring, and Atlantic mackerel). v The lipid

- nor malized BMF was calculated as the HBCD concentration in muscle of terrestrial passerine birds divided by the HBCD concentration in their stomach contents. Birds were sampled from e

- waste, urban, and rural locations of South China. w HBCD concentrations were measured in muscle tissue of predators (common kestrel, eagle owl, and little owl) and prey (Eurasian tree sparrow [ Passer montanu] ) and brown rat [ Rattus norvegicus ]) in Beijing. Since no significant differences were observed in the concentrations of HBC D between the two owl species, the data for the two owl species were combined and treated as one population. BMFs were calculated using lipid

- normalized concentrations of predators divided by that of prey. x The following species were sampled: blue mussel ( M. edulis ; soft tissue), lugworm ( Arenicola marina ; whole organism), shore crab ( C. maenas ; muscle), common eider ( S. mollissima ; egg contents), and great black

- backed gull ( L. marinus ; egg contents) from a location that was 2.8

– 6.4 km away from a point source at the head of the fjord. y TMF was calculated as the average rate of increase of HBCD from one trophic level to the next. z β

- HBCD was not detected in samples according to study authors. aa TMF was calculated as the slope of the curve of the trophic l evel versus the logarithm of the body residue. bb Samples were collected from pelagic and benthic food webs, including zooplankton, benthic invertebrates, benthic and pelagic fish, jellyfish, and a piscivorous bird species. cc Species include Chinese mystery snail ( C. chinensis ), prawn ( M. nipponense ), mud carp ( C. molitorella ), crucian carp ( C. uratus ), northern snakehead ( O. argus ), and water snake ( E. chinensis ). Samples were collected from a natural pond in an e

- waste recycling site in South China. dd TMF w as calculated using the same method as in Brandsma et al., 2015. ee Many species were sampled from Tianjin, China. Abbreviations: dw = dry weight; lw = lipid weight; ww = wet weight HBCD has been shown in numerous studies to bioaccumulate and biomagnify in aquatic and terrestrial food chains. In general, α

- HBCD bioaccumulates in organisms and biomagnifies through food webs to a greater extent than the β

- and γ

- diastereomers . Uncertainty remains as to the balance of diastere o mer accumulation in various species and the extent to which139

bioisomerization and biotransformation rates for each isomer affect bioaccumulation potential. Some authors (e.g., Law et al., 2006) have proposed that γ

- HBCD isomerizes to α

- HBCD under physiologi cal conditions, rather than uptake being diastereisomer

- specific. To test this theory, Esslinger et al. (2010) exposed mirror carp ( Cyprinus carpio morpha noblis ) to only γ

- HBCD and found no evidence of bioisomerization. In contrast, when Du et al. ( 2012 a ) exposed zebrafish ( Danio rerio ) to only γ

- HBCD, they found detectable levels of α

- HBCD in fish tissue, suggesting that bioisomerization occurred. Marvin et al. (2011) hypothesized that differences in accumulation could also be due in part to a combination of differences in solubility, bioavailability, and uptake and depuration kinetics. Zhang et al. (2014c) calculated diastereomer

- specific BCFs in algae and cyanobacteria ranging from 174 to 469. For the cyanobacteria ( Spirulina subsalsa ), the BCF for α

- H BCD (350) was higher than the BCFs for β

- HBCD (270) and γ

- HBCD (174). However, for the tested alga ( Scenedesmus obliquus ), the BCF for β

- HBCD (469) was higher than that for the other isomers (390

– 407). BCFs for HBCD in fish in the peer

- reviewed literat ure range as high as 18,100, as shown in Table 4

- 3 (Drottar and Krueger, 2000; Veith et al., 1979; Zhang et al., 2014b). Drottar and Krueger (2000) provided strong evidence that HBCD bioaccumulates in a bioconcentration test that was conducted according to guidelines OECD Test Guideline (TG) 305 and Office of Prevention, Pesticides and Toxic Substances (OPPTS) 850.1730. In this study, BCFs of 13,085 and 8,974 were reported in rainbow trout ( O. mykiss ) exposed to 0.18 and 1.8 μg/L, respectively. Concentratio ns of HBCD in tissue reached steady

- state at day 14 for fish exposed to 1.8 μg/L and, during the subsequent depuration stage, a 50% reduction of HBCD from edible and non

- edible tissue and whole fish was reported on days 19 and 20 post

- exposure. In fish140

exp osed to 0.18 μg/L, an apparent steady

- state was reached on day 21, but on day 35, the tissue concentration of HBCD in fish increased noticeably; thus, steady

- state was not achieved according to study authors, and BCF values (for the exposure concentration of 0.18 μg/L) were calculated based on day 35 tissue concentrations. Clearance of 50% HBCD from tissue of 0.18 μg/L exposed fish occurred 30

– 35 days post

- exposure. Veith et al. (1979) further supports a conclusion that HBCD bioaccumulates in a study
co nducted prior to the establishment of standardized testing guidelines for bioconcentration studies. The study reported a BCF of 18,100 following exposure of fathead minnow to 6.2 μg/L; the BCF was identified as a steady

- state BCF, but the report does not i ndicate time when steady

state was reached. A depuration phase was not included in this study. Zhang et al. (2014b) calculated BCFs for each diastereomer in mirror carp and found strong evidence that α

- HBCD (BCF of 5,570

– 11,500) is much more bioaccumulat ive than β

- and γ

- HBCD (BCF of 187

– 642); BCF values that were normalized to lipid content were much higher (30,700

– 45,200 for
α - HBCD, 1,030

– 1,900 for β

- HBCD, and 950

– 1,730 for γ

- HBCD) than non

- normalized BCFs. BAFs, which capture accumulation of HBCD from diet as well as water and sediment, were calculated for freshwater food webs in industrialized areas of Southern China in two separate field studies. He et al. (2013) calculated log BAFs of 4.8

– 7.7 (corresponding to BAFs of 63,000

– 50,000,000) for HBCD isomers in carp, tilapia, and catfish, and found higher BAFs for α

- HBCD than β

- and γ

- HBCD. In a pond near an e

- waste recycling site, Wu et al. (2011) calculated log BAFs of 2.85

– 5.98 for ƩHBCD (corresponding to BAFs of 700

– 950,000) in a fres hwater food web. Log BAFs for each diastereomer in this study were comparable to one another (see Table 4

- 3). La Guardia et al. (2012) calculated log BAFs in bivalves and gastropods collected downstream of a textile manufacturing outfall; these ranged from 4.2 to 5.3 for α

- and141

β

- HBCD (BAFs of 16,000

– 200,000), and from 3.2 to 4.8 for γ

- HBCD (BAFs of 1,600

– 63,000). BSAFs calculated in studies of invertebrates and fish are generally lower than reported BCFs and BAFs. Haukås et al. (2010b) reported BSAFs ≤0.006 calculated from lipid

- normalized concentrations of HBCD in ragworms and HBCD concentrations normalized to total organic content in sediment, indicating very low bioavailability of HBCD from sediments. Ragworm tissue concentrations were all less than the limit of detection. The pattern of diastereomers in sediments was found to generally resemble the composition of technical HBCD (i.e., predominantly γ

- HBCD). This study also found that in ragworms exposed to HBCD through a diet of contaminated mussels (containing diastereomer contributions of 48% α - HBCD, 7% β

- HBCD, and 45% g

- HBCD) , the tissue concentration of α

- HBCD was greater than that of β

- HBCD or γ

- HBCD, suggesting selective bioaccumulation of the α

- diastereomer
. Log BSAFs calculated in bivalves a nd gastropods collected downstream of a textile manufacturing outfall ranged from 0 to 0.9 (for α

- and β

- HBCD) and from

- 1.5 to 0 (for γ - HBCD) (La Guardia et al., 2012). These correspond to BSAFs of 1

– 8 for α

- and β

- HBCD and 0.03

– 1 for γ

- HBCD. BSAFs in benthivorous barbell ( Barbus graellsii ) and pelagic bleak ( Alburnus alburnus ) were calculated based on measured concentrations of HBCD reported in Eljarrat et al. (2004, 2005 as cited in van Beusekom et al., 2006) and ranged from 0.1 to 1.44 and from 0.14 to 1.23, respectively (van Beusekom et al., 2006). In Haukås et al. (2009), juvenile rainbow trout exposed via a single gavage dose of 0.61 mg/g wet weight (measured) reached a maximum accumulation of HBCD by day 8, with considerable amounts of all diast ereomers distributed to liver, brain, and muscle. Assimilation efficiencies of HBCD for individual diastereomers (percent of HBCD isomer in the fish relative142

to the amount in the bolus fed to the fish at the start of the experiment) were >100% for β

HBCD, 78% for α

- HBCD, and 46% for γ

- HBCD; values were calculated on day 4. A 71% reduction in HBCD was observed by day 21, primarily from the brain and liver. No clear elimination was observed from muscles. In zebrafish exposed to HBCD for 42 days, Du et al. ( 20 12 a ) measured diastereomer

- specific assimilation efficiencies of 40.84

– 54.67% for
α - HBCD, 33.59

– 47.71% for β

- HBCD, and 29.31

– 40.80% for γ

- HBCD. Du et al. (2013) reported similar assimilation efficiencies in zebrafish (see Table 4

- 3). Haukås et al. (2010b) reported a 28

- day assimilation efficiency of 67% in ragworms. Biomagnification of HBCD was demonstrated by Law et al. (2006), who reported BMFs of 9.2 (α

- HBCD), 4.3 (β

- HBCD), and 7.2 (γ

- HBCD). Uptake of HBCD into muscle from the diet of rainbow tro ut was exponential for α

- HBCD with a doubling time of 8.2 days, exponential for β

- HBCD with a doubling time of 17.1 days, and linear for γ

- HBCD with a rate constant of 0.006 per day. Depuration was rapid during the first 14 days and slower for the remainde r of the experiment for α

- HBCD (overall depuration rate was not determined). Depuration rates of 0.44×10 - 2 and 0.48×10 - 2 per day were found for β

- HBCD and γ

- HBCD, respectively. Steady

state was not reached for any of the diastereomers within the 52

- day exp osure period. Additional studies are available that support the conclusion that HBCD has the potential to biomagnify. Studies of zebrafish by Du et al. ( 2012 a , 2013) reported diastereo

- and enantiomer

- specific biomagnification. When BMFs were calculated f or diastereomers without accounting for specific enantiomers, after 42 days of exposure and a 21

- day depuration period,
α - HBCD was shown to biomagnify to a greater extent than β

- and γ

- HBCD (maximum BMFs of 29.71, 11.63, and 7.76, respectively). Enantiomer

- specific BMFs calculated in zebrafish by Du et al. (2013) followed a similar diastereomer pattern, although the BMF values were much lower143

than those from Du et al. ( 2012 a ). Additionally, the results of Du et al. (2013) suggest that the (+) enantiomers of β

- and γ

- HBCD are selectively magnified compared to their (−) enantiomers. This pattern did not hold true for α

- HBCD. Letcher et al. (2009) found evidence of biomagnification of HBCD from the ringed seal to the polar bear in an East Greenland food web, r eporting a BMF of 1.7. BMFs for α

- HBCD in a harbor seal food web varied according to prey fish species, but ranged from 0.54 to 3.0 (Shaw et al., 2012). Shaw et al. (2012) calculated higher BMFs from prey fish to the livers of adult male harbor seals than to the blubber of those seals. BMFs for α

- HBCD in gulls and common eiders in a coastal marine food web in Norway provide evidence of biomagnification, ranging from 3.1 to 1,285 when calculated on a wet weight basis and from 2.8 to 26 when calculated on a lipid

- weight basis (Haukås et al., 2010a). In terrestrial food webs in China, both Sun et al. (2012) and Yu et al. (2013) found evidence of biomagnification (see Table 4

- 3), with BMFs up to 30 in passerine birds and up to 16 in owls. Yu et al. (2013) found more ( ̶ ) α - HBCD in predator species than (+) α

- HBCD, but other studies do not agree, suggesting that enantiomer biomagnification may be species

- specific. Tomy et al. (2008) describes the extent of trophic transfer (transfer and accumulation of HBCD betw een trophic levels) by calculating TMFs of 2.1 and 0.5 for α

- and γ

- HBCD, respectively, based on the Arctic marine food web. Samples of blubber were taken and analyzed from the beluga whale ( D elphinapterus leucas ), narwhal ( Monodon monoceros ), and walrus ( Odobenus rosmarus ), while whole organisms were analyzed for arctic cod ( Boreogadus saida ), shrimp ( Pandalus borealis and Hymenodora glacialis ), clams ( Mya truncate and Serripes groenlandica ) , deepwater redfish ( Sebastes mentella ), and mixed zooplankton to determine144

HBCD concentrations in the tissue of animals of different trophic levels in order to establish whether HBCD biomagnifies between trophic levels. Brandsma et al. (2015) studied trophic magnification of HBCD through benthic and pelagic food webs i n the Western Scheldt estuary, The Netherlands, and found similar results:
α - HBCD concentrations increased and γ

- HBCD concentrations decreased with an increase in trophic level (TMFs of 2.2 and 0.3, respectively). In a freshwater food web studied near an
e - waste recycling site in South China, Wu et al. (2010) calculated enantiomer

- specific TMFs for
α - HBCD of 2.18

– 2.2, and found evidence that as HBCD migrates up through the food web,
α - HBCD increases and γ

- HBCD decreases, while β

- HBCD comprises a very low proportion of ∑HBCD. This pattern, also demonstrated by data in Haukås et al. (2010a), becomes more prominent at upper trophic levels. In marine and freshwater food webs, Zhang et al. (2013) calculated TMFs greater than 1 for α

- HBCD and ƩHBCD. In summary, while HBCD has been shown in numerous studies to bioaccumulate and biomagnify in aquatic and terrestrial food chains, diastereomer

- and enantiomer

- specific mechanisms of accumulation are still unclear.
4.4 CONCLUSIONS There is evidence for high concern for both acute and chronic toxicity to aquatic and terrestrial organisms exposed to HBCD. HBCD can cause effects on survival, growth, reproduction, development, and behavior. The potential of HBCD to bioaccumulate in aqua tic and terrestrial organisms and biomagnify in food webs is high.145

4.5. REFERENCES Allchin, CR; Morris, S. (2003) Hexabromocyclododecane (HBCD) diastereoisomers and brominated diphenyl ether congener (BDE) residues in edible fish from the rivers Skerne a nd Tees, U.K. Organohalogen Compd 61:41

– 44. Andersen, MS; Fuglei, E; König, M; et al. (2015) Levels and temporal trends of persistent organic pollutants (POPs) in arctic foxes ( Vulpes lagopus ) from Svalbard in relation to dietary habits and food availabil ity. Sci Total Environ 511:112

– 122. Anselmo, HMR; Koerting, L; Devito, S; et al. (2011) Early life developmental effects of marine persistent organic pollutants on the sea urchin Psammechinus miliaris . Ecotox Environ Safe 74:2182

– 2192. Arnot, JA; Gobas, FAPC. (2006) A review of bioconcentration factor (BCF) and bioaccumulation factor (BAF) assessments for organic chemicals in aquatic organisms. Environ Rev 14:257

– 297. Aufderheide, J; Jones, A; MacGregor, JA; et al. (2003) Effect of hexabromocyclododecane on the survival and reproduction of the earthworm, Eisenia fetida . ABC study No. 47222, ABC Laboratories, Inc. and Wildlife International Ltd: Columbia, MO and Easton, MD, pp 94 [as cited in ECB, 2008]
. Brandsma, SH; Leonards, PEG; Leslie, HA; et al. (20 15) Tracing organophosphorus and brominated flame retardants and plasticizers in an estuarine food web. Sci Total Environ 505:22

– 31. Bustnes, JO; Borgå, K; Dempster, T; et al. (2012) Latitudinal distribution of persistent organic pollutants in pelagic and demersal marine fish on the Norwegian coast. Environ Sci Technol 46:7836

– 7843. Calmbacher, CW. (1978) The acute toxicity of HBCD lot 990

- 17 to the bluegill sunfish Lepomis macrochirus rafinisque with test data and cover letter. Union Carbide Corporation, Tarrytown, NY. UCES project No. 11506

- 03

- , 9 pp. Submitted under TSCA Section 8D; EPA Document No. 900000268; OTS0523260. Cheaib, Z; Grandjean, D; Kupper, T; et al. (2009) Brominated flame retardants in fish of Lake Geneva (Switzerland). Bull Environ Contam Toxicol 82:522

– 527. Chen, D; La Guardia, MJ; Luellen, DR; et al. (2011) Do temporal and geographical patterns of HBCD and PBDE flame retardants in U.S. fish reflect evolving industrial u sage? Environ Sci Technol 45:8254

– 8261. Chen, D; Letcher, RJ; Burgess, NM; et al. (2012) Flame retardants in eggs of four gull species ( Laridae ) from breeding sites spanning Atlantic to Pacific Canada. Environ Pollut 168:1

– 9.146

Crump, D; Egloff, C; Chiu, S ; et al. (2010) Pipping success, isomer

- specific accumulation, and hepatic mRNA expression in chicken embryos exposed to HBCD. Toxicol Sci 115:492

– 500. De Boer, J; Leslie, HA; Leonards, PEG; et al. (2004) Screening and time trend study of decabromodipheny lether and hexabromocyclododecane in birds. Abstract. The Third International Workshop on Brominated Flame Retardants, June 6

– 9, Toronto, Canada. pp 125

– 128 [as cited in UNEP, 2010]. de Wit, CA; Herzke, D; Vorkamp, K. (2010) Brominated flame retardants in the Arctic environment

– trends and new candidates. Sci Total Environ 408:2885

– 2918. Deng, J; Yu, L; Liu, C; et al. (2009) Hexabromocyclododecane

- induced developmental toxicity and apoptosis in zebrafish embryos. Aquat Toxicol 93(1):29

– 36. Desjardins, D; MacGregor, JA; Krueger, HO. (2004) Hexabromocyclododecane (HBCD): A 72

- hour toxicity test with the marine diatom ( Skeletonema costatum ), Final report. Wildlife Internation Ltd, Easton, MD, pp 66 [as cited in ECB, 2008 and IUCLID, 2005]. Desjardins, D; MacGregor, JA; Krueger, HO. (2005) Final report. Chapter 1, Hexabromocyclododecane (HBCD): A 72

- hour toxicity test with the marine diatom ( Skeletonema costatum ) using a co

- solvent. Chapter 2, Hexabromocyclododecane (HBCD): A 72

- hour toxicity test with the marine diatom ( Skeletonema costatum ) using generator column saturated media. Wildlife International Ltd, Easton, MD, pp19 [as cited in ECB, 2008]
. Dodder, NG; Maruya, KA; Ferguson, PL; et al. (2014) Occurrence of contaminants of emerging concern in mussel s ( Mytilus spp. ) along the California coast and the influence of land use, storm water discharge, and treated wastewater effluent. Marine Poll Bull 81:340

– 346. Drottar, KR; Krueger, HO. (1998) Hexabromocyclododecane (HBCD): A flow

- through life

cycle toxic ity test with the cladoceran ( Daphnia magna ). Report #439A

- 108. Wildlife International Ltd, Easton, MD, pp 78. Submitted under TSCA Section 8D; EPA Document No. 86980000152; OTS0559490. Drottar, KR; Krueger, HO. (2000) Hexabromocyclododecane (HBCD): A flo w - through bioconcentration test with the rainbow trout ( Oncorhynchus mykiss ). Report# 439A

- 111. Wildlife International Ltd, Easton, MD, pp 1

– 137. Submitted under TSCA Section FYI; EPA Document No. 84010000001; OTS0001392. Drottar, KR; MacGregor, JA; Krueg er, HO. (2001) Hexabromocyclododecane (HBCD): An early life

- stage toxicity test with the rainbow trout ( Oncorhynchus mykiss ). Wildlife International Ltd, Easton, MD, pp 1

– 102. Submitted under TSCA Section 8E; EPA Document No. 88020000023; OTS0574262.147

Du, M; Lin, L; Yan, C; et al. ( 2012 a ) Diastereoisomer

- and enantiomer

- specific accumulation, depuration, and bioisomerization of hexabromocyclododecanes in zebrafish ( Danio rerio ). Environ Sci Technol 46:11040

– 11046. Du, M; Zhang, D; Yan, C; et al. (2012 b ) D evelopmental toxicity evaluation of three hexabromocyclododecane diastereoisomers on zebrafish embryos. Aquat Toxicol 112

– 113:1

– 10. Du, M; Lin, L; Yan, C; et al. (2013) Enantiomer

- specific bioaccumulation and depuration of hexabromocyclododecanes in zebra fish ( Dania rerio ). J Haz Mater 248

– 249:167

– 171. ECB (European Chemical Bureau). (2008) Risk assessment hexabromocyclododecane, CAS No: 25637

- 99

- 4, EINECS No: 247

- 148

- 4, Final Draft. European Chemical Bureau. Swedish Chemicals Agency: Sundbyberg, Sweden; pp 492. ECOSAR. (2012) ECOSAR: v. 1.11. U.S. Environmental Protection Agency, OPPT

- Risk Assessment Division. Available online at http://www.epa.gov/opptintr/newchems/tools/21ecosar.ht m (accessed June 17, 2015). Eljarrat, E; de la Cal, A; Raldua, D; et al. (2004) Occurence and bioavailability of polybrominated diphenyl ethers and hexabromocyclododecane in sediment and fish from Cinca river, a tributary of the Ebro river (Spain). Environ Sci Technol 38:2603

– 2608. Eljarrat, E; de la Cal, A; Duran, DC; et al. (2005) Brominated flame retardents in Alburnus alburnus from Cinca river basin. Environ Pollut 133:501

– 508 [as cited in van Beusekom et al., 2006]. Esslinger, S; Becker, R; Mü ller

- Belecke, A; et al. (2010) HBCD sereoisomer pattern in mirror carps following dietary exposure to pure γ

- HBCD enantiomers. J Agric Food Chem 58:9705

– 9710. Eulaers, I; Jaspers, VLB; Pinxten, R; et al. (2014) Legacy and current

- use brominated flame reta rdants in the barn owl. Sci Total Environ 472:454

– 462. Fernie, KJ; Letcher, RJ. (2010) Historical contaminants, flame retardants, and halogenated phenolic compounds in peregrine falcon ( Falco peregrinus ) nestlings in the Canadian Great Lakes basin. Enviro n Sci Technol 44:3520

– 3526. Fernie, KJ; Marteinson, SC; Bird, DM; et al. (2011) Reproductive changes in American kestrels ( Falco sparverius ) in relation to exposure to technical hexabromocyclododecane flame retardant. Environ Toxicol Chem 30:2570

– 2575. F oekema, EM; Parron, ML; Mergia, MT; et al. (2014) Internal effect concentrations of organic substances for early life development of egg

- exposed fish. Ecotox Environ Safe 101:14

– 22.148

Graves, WC; Swigert, JP. (1997a) Hexabromocyclododecane (HBCD): A 48

- hour flow

- through acute toxicity test with the Cladoceran ( Daphnia magna ). Report# 439A

- 102. Wildlife International Ltd, Easton, MD. Submitted under TSCA Section 8D; EPA Document No. 86970000792; OTS0559002. Graves, WC; Swigert, JP. (1997b) Hexabromocyclododecane (HBCD): A 96

- hour flow

- through acute toxicity test with the rainbow trout ( Oncorhynchus mykiss ). Report# 439A

- 101. Wildlife Internationl Ltd, Easton, MD, pp 70. Submitted under TSCA Section 8D; EPA Document No. 86980000086; OTS055942 4. Guerra, P; Alaee, M; Jiménez, B; et al. (2012) Emerging and historical brominated flame retardants in peregrine falcon ( Falco peregrinus ) eggs from Canada and Spain. Environ Int 40:179

– 186. Haukås, M; Mariussen, E; Ruus, A; et al. (2009) Accumulation and disposition of hexabromocyclododecane (HBCD) in juvenile rainbow trout ( Oncorhynchus mykiss ). Aquat Toxicol 95:144

– 151. Haukås, M; Hylland, K; Nygård, T; et al. (2010a) Diastereomer

- specific bioaccumulation of hexabromocyclododecane (HBCD) in a coasta l food web, Western Norway. Sci Total Environ 408:5910

– 5916. Haukås, M; Ruus, A; Hylland, K; et al. (2010b) Bioavailability of hexabromocyclododecane to the polychaete Hediste diversicolor : exposure through sediment and food from a contaminated fjord. Env iron Toxicol Chem 29:1709

– 1715. He, M; Luo, X; Yu, L; et al. (2013) Diasteroisomer and enantiomer

- specific profiles of hexabromocyclododecane and tetrabromobisphenol A in an aquatic environment in a highly industrialized area, South China: vertical profil e, phase partition, and bioaccumulation. Environ Poll 179:105

– 110. Hong, H; Li, D; Shen, R; et al. (2014a) Mechanisms of hexabromocyclododecanes induced developmental toxicity in marine medaka ( Oryzias melastigma ) embryos. Aquat Toxicol 152:173

– 185. Hong , SH; Shim, WJ; Han, GM; et al. (2014b) Levels and profiles of persistent organic pollutants in resident and migratory birds from an urbanized coastal region of South Korea. Sci Total Environ 470

– 471:1463

– 1470. Hu, J; Liang, Y; Chen, M; et al. (2009) Asse ssing the toxicity of TBBPA and HBCD by zebrafish embryo toxicity assay and biomarker analysis. Environ Toxicol 24:334

– 342. Hu, J; Jim, J; Wang, Y; et al. (2011) Levels of polybrominated diphenyl ethers and hexabromocyclododecane in the atmosphere and tre e back from Beijing, China. Chemosphere 84:355

– 360.149

Ilyas, M; Sudaryanto, A; Setiawan, IE; et al. (2013) Characterization of polychlorinated biphenyls and brominated flame retardants in sludge, sediment and fish from municipal dumpsite at Surabaya, Indone sia. Chemosphere 93:1500

– 1510. Isobe, T; Ochi, Y; Ramu, K; et al. (2009) Organohalogen contaminants in striped dolphins ( Stenella coeruleoalba ) from Japan: present contamination status, body distribution and temporal trends (1978

– 2003). Marine Poll Bull 5 8:396

– 401. Isobe, T; Ogawa, SP; Ramu, K; et al. (2012) Geographical distribution of non

- PBDE

- brominated flame retardants in mussels from Asian coastal waters. Environ Sci Pollut Res 19:3107

– 3117. IUCLID. (2005) International Uniform ChemicaL Information Database (IUCLID) Hexabromocyclododecane dataset. Submitted to U.S. EPA’s High Production Volume (HPV) Chemical Program by Albemarle Corporation . Available online at http://www.epa.gov/chemrtk/pubs/summaries/cyclodod/c13459rr.pdf (accessed June 17, 2015). Jatzek HJ. (1990) Determination of the acute toxicity of hexabromid S to the waterflea Daphnia magna Straus. 1184/88, BASF Indolfab. Submitted under TSCA Section 8D; EPA Document No. 86900000392; OTS0522955
. Jeong, GH; Hwang, NR; Hwang, E; et al. (2014) Hexabromocyclododecanes in crucian carp and sediment from the major rivers in Korea. Sci Total Environ 470

– 471:1471

– 1478. Jin, L; Lam, JCW; Murphy, MB; et al. (2012) Temporal trends of polybrominated diphenyl ethers and hexabromocyclododecanes in marine mammals with special reference to Hong Kong, South China. In: Loganathan, BG; Lam, PKS, eds. Global Contamination Trends of Persistent Organic Chemicals. Bo ca Raton, FL: CRC Press, pp 497

– 513. Klosterhaus, SL; Stapleton, HM; La Guardia, MJ; et al. (2012) Brominated and chlorinated flame retardants in San Francisco Bay sediments and wildlife. Environ Int 47:56

– 65. Kobiliris, D. (2010) Influence of embryonic exposure to hexabromocyclododecane (HBCD) on the corticosterone response and “fight or flight” behaviors of captive American kestrels. Thesis submitted to McGill University in partial fulfilment of the requirements of the degree of Masters of Science. Depa rtment of Natural Resource Sciences, McGill University, Montreal, Canada. Kuiper, RV; Cantoacuten, RF; Leonards, PE; et al. (2007) Long

- term exposure of European flounder ( Platichthys flesus ) to the flame

- retardants tetrabromobisphenol A (TBBPA) and hexa bromocyclododecane (HBCD). Ecotoxicol Environ Saf 67(3):349

– 360. La Guardia, MJ; Hale, RC; Harvey, E; et al. (2012) In situ accumulation of HBCD, PBDEs, and several alternative flame

- retardants in the bivalve ( Corbicula fluminea ) and gastropod ( Elimia pr oxima ). Environ Sci Technol 46:5798

– 5805.150

Law, K; Palace, VP; Halldorson, T; et al. (2006) Dietary accumulation of hexabromocyclododecane diastereoisomers in juvenile rainbow trout ( Oncorhynchus mykiss ) I: bioaccumulation parameters and evidence of bioiso merization. Environ Toxicol Chem 25(7):1757

– 1761. Law, RJ; Covaci, A; Harrad, S; et al. (2014) Levels and trends of PBDEs and HBCDs in the global environment: status at the end of 2012. Environ Int 65:147

– 158. Letcher, RJ; Gebbink, WA; Sonne, C; et al. (2009) Bioaccumulation and biotransformation of brominated and chlorinated contaminants and their metabolites in ringed seals ( Pusa hispida ) and polar bears ( Ursus maritimus ) from East Greenland. Environ Int 35:1118

– 1124. Li, Y; Zhou, Q; Wang, Y; et al. (2011) Fate of tetrabromobisphenol A and hexabromocyclododecane brominated flame retardants in soil and uptake by plants. Chemosphere 82:204

– 209. Malarvannan, G; Belpaire, C; Geeraerts, C; et al. (2014) Assessment of persistent brominated and chlorinated organic contaminants in the European eel ( Anguilla anguilla ) in Flanders, Belgium: Levels, profiles, and health risk. Sci Total Environ 482

– 483:222

– 233. Marteinson, SC; Bird, DM, Shutt, JL; et al. (2010) Multi

- generational effects of polybrominated diphe nylethers exposure: embryonic exposure of male American kestrels (Falco sparverius) to DE

- 71 alters reproductive success and behaviors. Environ Toxicol Chem 29: 1740

- 1747. Marteinson, SC; Kimmins, S; Letcher, RJ; et al. (2011) Diet exposure to technical hexabromocyclododecane (HBCD) affects testes and circulating testosterone and thyroxine levels in American kestrels ( Falco sparverius ). Environ Res 111:1116

– 1123. Marteinson, SC; Bird, DM; Letcher, RJ; et al. (2012) Dietary exposure to technical hexabromocyclododecane (HBCD) alters courtship, incubation and parental behaviors in American kestrels ( Falco sparverius ). Chemosphere 89:1077

– 1083. Marvin, CH; Tomy, GT; Armitage, JM; et al. (2011) Hexabromocyclododecane: current understanding of chemist ry, environmental fate and toxicology and implications for global management. Environ Sci Technol 45:8613

– 8623. McKinney, MA; Letcher, RJ; Aars, J; et al. (2011) Flame retardants and legacy contaminants in polar bears from Alaska, Canada, East Greenland a nd Svalbard, 2005

– 2008. Environ Int 37:365

– 374. Miljeteig, C; Strøm, H; Gavrilo, MV; et al. (2009) High levels of contaminants in ivory gull Pagophila eburnea eggs from the Russian and Norwegian Arctic. Environ Sci Technol 43(14):5521

– 5528.151

Miller, A; El liott, JE; Elliott, KH; et al. (2015) Brominated flame retardant trends in aquatic birds from the Salish Sea region of the west coast of North America, including a mini

- review of recent trends in marine and estuarine birds. Sci Total Environ 502:60

– 69. MO EJ (Ministry of the Environment, Japan). (2009) 6

- Week administration study of 1,2,5,6,9,10

- hexabromocyclododecane for avian reproduction toxicity under long

- day conditions using Japanese quail. Report. Ministry of the Environment, Japan. Research Institut e for Animal Science in Biochemistry & Toxicology [as cited in UNEP, 2010]. Muir, DCG; Backus, S; Derocher, AE; et al. (2006) Brominated flame retardants in polar bears ( Ursus maritimus ) from Alaska, the Canadian Arctic, East Greenland, and Svalbard. Envi ron Sci Technol 40(2):449

– 455. OECD (Organisation for Economic Co

- operation and Development). (2000) OECD Series on assessment and testing number 23: Guidance document on aquatic toxicity testing of difficult substances and mixtures. Organization for Econ omic Co

- operation and Development, document ENV/JM/MONO(2000)6. Oetken, M; Ludwichowski, K; Nagel, R. (2001) Validation of the preliminary EU

- concept of assessing the impact of chemicals to organisms in sediment by using selected substances. Dresden Unive rsity of Technology, Dresden, Germany, Institute of Hydrobiology. UBA

- FB 299 67 411, pp 97 [as cited in ECB, 2008]. Palace, VP; Pleskach, K; Halldorson, T; et al. (2008) Biotransformation enzymes and thyroid axis disruption in juvenile rainbow trout ( Onco rhynchus mykiss ) exposed to hexabromocyclododecane diastereoisomers. Environ Sci Technol 42(6):1967

– 1972. Palace, V; Park, B; Pleskach, K; et al. (2010) Altered thyroxine metabolism in rainbow trout ( Oncorhynchus mykiss ) exposed to hexabromocyclododecane (HBCD). Chemosphere 80(2):165

– 169. Porch, JR; Kendall, TZ; Krueger, HO. (2002) Hexabromocyclododecane (HBCD): A toxicity test to determine the effects of the test substance on seedling emergence of six species of plants. Wildlife International Ltd, East on, MD, pp 126 [as cited in ECB, 2008]. Roberts, CA; Swigert, JP. (1997) Hexabromocyclododecane (HBCD): A 96

- hour toxicity test with the freshwater alga ( Selenastrum capricornutum ). Report# 439A

- 103. Wildlife International Ltd, Easton, MD. Submitted under TSCA Section 8D; EPA Document No. 86970000799; OTS0573662
. Ronisz, D; Farmen Finne, E; Karlsson, H; et al. (2004) Effects of the brominated flame retardants hexabromocyclododecane (HBCDD), and tetrabromobisphenol A (TBBPA), on hepatic enzymes and other biomarkers in juvenile rainbow trout and feral eelpout. Aquat Toxicol 69:229

– 245.152

Salamova, A; Hites, RA. (2013) Brominated and chlorinated flame retardants in tree bark from around the globe. Environ Sci Technol 47:349

– 354. Sellström, U; Kierkegaard, A; de Wit, C; et al. (1998) Polybrominated diphenyl ethers and hexabromocyclododecane in sediment and fish from a Swedish river. Environ Toxicol Chem 17:1065

– 1072. Shaw, SD; Berger, ML; Weijs, L; et al. (2012) Tissue

- specific accumulation of polybrominated diphenyl ethers (PBDEs) including deca

- BDE and hexabromocyclododecanes (HBCDs) in harbor seals from the northwest Atlantic. Environ Int 44:1

– 6. Siebel

- Sauer; Bias. (1987) Algal growth inhibition test. Hexabromid S. Test report no 1184/87. BASF Labor fur Ö kologie, pp 12 [as cited in ECB, 2008]. Smolarz, K; Berger, A. (2009) Long

- term toxicity of hexabromocyclododecane (HBCDD) to the benthic clam Macoma balthica (L.) from the Baltic Sea. Aquat Toxicol 95(3):239

– 247. Sun, Y; Luo, X; Mo, L; et al. (2012) H exabromocyclododecane in terrestrial passerine birds from e - waste, urban and rural locations in the Pearl River Delta, South China: levels, biomagnification, diastereoisomer

- and enantiomer

- specific accumulation. Environ Pollut 171:191

– 198. Thomas, S; Kre uger, HO; Kendall, TZ. (2003a) Hexabromocyclododecane (HBCD): a prolonged sediment toxicity test with Hyalella azteca using spiked sediment with 2% total organic carbon. Report# 439A

- 119B. Wildlife International Ltd, Easton, MD, pp 103. Submitted under TSC A Section FYI; EPA Document No. 84040000010; FYI

- 1103

- 01472, pg 150. Available online at http://java.epa.gov/oppt_chemical_search/proxy?filena me=2003

- 11

- FYI

- 03

- 01472A_fyi_1103_01472a.pdf (accessed June 18, 2015). Thomas, S; Kreuger, HO; Kendall, TZ. (2003b) Hexabromocyclododecane (HBCD): a prolonged sediment toxicity test with Hyalella azteca using spiked sediment with 5% total organic carbon. Report# 439A

- 120, Final. Wildlife International Ltd, Easton, MD, pp 103. Submitted under TSCA Section FYI; EPA Document No. 84040000010; FYI

- 1103

- 01472, pg 253. Available online at http://java.epa.gov/oppt_chemical_search/proxy?filename=2003

- 11

- FYI

- 03

- 01472A_fyi_1103_01472a.pdf (accessed June 18, 2015). Tomy, GT; Pleskach, K; Oswald, T; et al. (2008) Enantioselective bioaccumulation of hexabro mocyclododecane and congener

- specific accumulation of brominated diphenyl ethers in an Eastern Canadian Arctic marine food web. Environ Sci Technol 42:3634

– 3639. UNEP (United Nations Environmental Program). (2010) Hexabromocyclododecane draft risk profile . United Nations Environmental Program, Stockholm Convention. Available online at http://chm.pops.int/Convention/POPsReviewCommittee/hrPOPRCMeetings/POPRC5/POPRC5F ollowupcommunications/HBCDInvitationforcommentsondraftRP/tabid/742/language/en

US/Default.aspx (accessed June 18, 2015).153

van Beusekom, OC; Eljarrat, E; Barceloacute, D; et al. (2006) Dynamic modeling of food

- chain accumulation of brominated flame retardants in fish from the Ebro River Basin, Spain. Environ Toxicol Chem 25(10):2553

– 2560. Veith, GD; Defoe, DL; Bergstedt, BV. (1979) Measuring and estimating the bioconcentration factor of chemicals in fish. J Fish Res Board Can 36:1040

– 1048. Venier, M; Wierda, M; Bowerman, WW; et al. (2010) Flame retardants and organochlorine pollutants in bald eagle plasma from the Great Lakes region. Chemosphere 80:1234

– 1240. Vorkamp, K; Rig, FF; Bossi, R; et al. (2011) Temporal trends of hexabromocyclododecane, polybrominated diphenyl ethers and polychlorinated biphenyls in ringed seals from East Greenland. Environ Sci Technol 45:1243

– 1249. Walsh, GE; Yoder, MJ ; McLaughlin, LL; et al. (1987) Responses of marine unicellular algae to brominated organic compounds in six growth media. Ecotoxicol Environ Saf 14:215

– 222. Watanabe, I; Tatsukawa, R. (1989) Anthropogenic brominated aromatics in the Japanese environment . Proceedings of the Workshop on Brominated Aromatic Flame Retardants, Swedish National Chemicals Inspectorate, Solna, Sweden, October 24

– 26, pp 63

– 71 [as cited in Sellström et al. 1998]. Wu, J; Guan, Y; Zhang, Y; et al. (2010) Trophodynamics of hexabromo cyclododecanes and several other non

- PBDE brominated flame retardants in a freshwater food web. Environ Sci Technol 44:5490

– 5495. Wu, J; Guan, Y; Zhang, Y; et al. (2011) Several current

- use, non - PBDE brominated flame retardants are highly bioaccumulative: evidence from field determined bioaccumulation factors. Environ Int 37:210

– 215. Wu, J; Zhang, Y; Luo, X; et al. (2012a) A review of polybrominated diphenyl ethers and alternative brominated flame retardants in wildlife from China: levels, trends, and bio accumulation characteristics. J Environ Sci 24:183

– 194. Wu, T; Wang, S; Huang, H; et al. (2012b) Diastereomer

- specific uptake, translocation, and toxicity of hexabromocyclododecane diastereoisomers to maize. J Agr Food Chem 60:8528

– 8534. Wu, M; Zuo, Z; L i, B; et al. (2013) Effects of low

- level hexabromocyclododecane (HBCD) exposure on cardiac development in zebrafish embryos. Ecotoxicology 22:1200

– 1207. Yu, L; Luo, X; Zheng, X; et al. (2013) Occurrence and biomagnification of organohalogen pollutants in two terrestrial predatory food chains. Chemosphere 93:506

– 511.154

Zegers, BN; Mets, A; Van Bommel, R; et al. (2005) Levels of hexabromocyclododecane in harbor porpoises and common dolphins from western European seas, with evidence for stereoisomer

- specific b iotransformation by cytochrome p450. Environ Sci Technol 39(7):2095

– 2100. Zhang, X; Yang, F; Zhang, X; et al. (2008) Induction of hepatic enzymes and oxidative stress in Chinese rare minnow ( Gobiocypris rarus ) exposed to waterborne hexabromocyclododecane (HBCDD). Aquat Toxicol 86(1):4

– 11. Zhang, Y; Yang, F; Luo, C; et al. (2009) Bioaccumulative characteristics of hexabromocyclododecanes in freshwater species from an electronic waste recycling area in China. Chemosphere 76:1572

– 1578. Zhang, Y; Sun, H; L iu, F; et al. (2013) Hexabromocyclododocanes in limnic and marine organisms and terrestrial plants from Tiangin, China: diastereomer

- and enantiomer

- specific profiles, biomagnification, and human exposure. Chemosphere 93:1561

– 1568. Zhang, H; Pan, L; Tao, Y. (2014a) Antioxidant responses in clam Venerupis philippinarum exposed to environmental pollutant hexabromocyclododecane. Environ Sci Pollut Res 21:8206

– 8215. Zhang, Y; Sun, H; Ruan, Y. (2014b) Enantiomer

- specific accumulation, de puration, metabolization and isomerization of hexabromocyclododecane (HBCD) diastereomers in mirror carp from water. J Haz Mater 264:8

– 15. Zhang, Y; Sun, H; Zhu, H; et al. (2014c) Accumulation of hexabromocyclododecane diastereomers and enantiomers in two microalgae, Spirulina subsalsa and Scenedesmus obliquus
. Ecotox Environ Safe 104:136

– 142. Zhu, N; Fu, J; Gao, Y; et al. (2013) Hexabromocyclododecane in alpine fish from the Tibetan Plateau, China. Environ Poll 181:7

– 13.155

Attachment A: EPI Suite Output for HBCD , CAS Registry Number 3194 - 55 - 6 SMILES : BrC(C(Br)CCC(Br)C(Br)CCC(Br)C(Br)C1)C1 CHEM : Cyclododecane, 1,2,5,6,9,10

- hexabromo

MOL FOR: C12 H18 Br6 MOL WT : 641.70

------------------------------ EPI SUMMARY (v4.11)

-------------------------- Physical Property Inputs: Log Kow (octanol

- water): 5.63 Boiling Point (deg C) :

------ Melting Point (deg C) : 190.00 Vapor Pressure (mm Hg) : 4.7E

- 007 Water Solubility (mg/L): 0.0086 Henry LC (atm

- m3/mole) :

---- -- Log Octanol

- Water Partition Coef (SRC): Log Kow (KOWWIN v1.68 estimate) = 7.74 Boiling Pt, Melting Pt, Vapor Pressure Estimations (MPBPVP v1.43): Boiling Pt (deg C): 462.03 (Adapted Stein & Brown method) Melting Pt (deg C): 180.03 (Mean or Weighted MP) VP(mm Hg,25 deg C): 1.29E

- 008 (Modified Grain method) VP (Pa, 25 deg C) : 1.72E

- 006 (Modified Grain method) VP (exp database): 4.70E

- 07 mm Hg (6.27E

- 005 Pa) at 21 deg C Subcooled liquid VP: 2.01E

- 005 mm Hg (21 deg C, user

- entered VP ) : 0.00268 Pa (21 deg C, user

- entered VP ) Water Solubility Estimate from Log Kow (WSKOW v1.42): Water Solubility at 25 deg C (mg/L): 0.003653 log Kow used: 5.63 (user entered) melt pt used: 190.00 deg C Water Sol (Exper. database match) = 0.0086 mg/L (25 deg C) Exper. Ref : TSCATS; OTS0523262 Water Sol Estimate from Fragments: Wat Sol (v1.01 est) = 0.0030722 mg/L ECOSAR Class Program (ECOSAR v1.11): Class(es) found: Neutral Organics Henrys Law Constant (25 deg C) [HENRYWIN v3.20]: Bond Method : 1.72E

- 006 atm

- m3/mole (1.74E

- 001 Pa

- m3/mole) Group Method: 6.43E

- 011 atm

- m3/mole (6.52E

- 006 Pa

- m3/mole) Exper Database: 4.60E

- 05 atm

- m3/mole (4.66E+000 Pa

- m3/mole)156

For Henry LC Comparison Purposes: User

- Entered Henry LC: not entered Henrys LC [via VP/WSol estimate using User

- En tered or Estimated values]: HLC: 4.614E

- 005 atm

- m3/mole (4.676E+000 Pa

- m3/mole) VP: 4.7E

- 007 mm Hg (source: User

- Entered) WS: 0.0086 mg/L (source: User

- Entered) Log Octanol

- Air Partition Coefficient (25 deg C) [KOAWIN v1.10]: Log Kow used: 5.63 (user ente red) Log Kaw used:

- 2.726 (exp database) Log Koa (KOAWIN v1.10 estimate): 8.356 Log Koa (experimental database): None Probability of Rapid Biodegradation (BIOWIN v4.10): Biowin1 (Linear Model) : 0.1650 Biowin2 (Non

- Linear Model) : 0.0000 Expert Survey Bio degradation Results: Biowin3 (Ultimate Survey Model): 1.9548 (months ) Biowin4 (Primary Survey Model) : 3.1342 (weeks ) MITI Biodegradation Probability: Biowin5 (MITI Linear Model) :

- 0.4234 Biowin6 (MITI Non

- Linear Model): 0.0000 Anaerobic Biodegradation Probability: Biowin7 (Anaerobic Linear Model): 2.5072 Ready Biodegradability Prediction: NO Hydrocarbon Biodegradation (BioHCwin v1.01): Structure incompatible with current estimation method! Sorption to aerosols (25 Dec C)[AEROWIN v1.00]: Vapor pressure (liquid/subcooled): 0.00268 Pa (2.01E

- 005 mm Hg) Log Koa (Koawin est ): 8.356 Kp (particle/gas partition coef. (m3/ug)): Mackay model : 0.00112 Octanol/air (Koa) model: 5.57E

- 005 Fraction sorbed to airborne particulates (phi): Junge

- Pankow model : 0.0389 Mackay model : 0.0822 Octanol/air (Koa) model: 0.00444 Atmospheric Oxidation (25 deg C) [AopWin v1.92]: Hydroxyl Radicals Reaction: OVERALL OH Rate Constant = 5.0148 E

- 12 cm3/molecule

- sec Half

- Life = 2.133 Days (12

- hr day; 1.5E6 OH/cm3) Half

- Life = 25.595 Hrs157

Ozone Reaction: No Ozone Reaction Estimation Fraction sorbed to airborne particulates (phi): 0.0605 (Junge

- Pankow, Mackay avg) 0.00444 (Koa method) Note: the sorbed fraction may be resistant to atmospheric oxidation Soil Adsorption Coefficient (KOCWI N v2.00): Koc : 9.678E+004 L/kg (MCI method) Log Koc: 4.986 (MCI method) Koc : 7.689E+004 L/kg (Kow method) Log Koc: 4.886 (Kow method) Aqueous Base/Acid

- Catalyzed Hydrolysis (25 deg C) [HYDROWIN v2.00]: Total Kb for pH > 8 at 25 deg C : 1.174E

- 007 L/mol

- sec Kb Half

- Life at pH 8: 1.871E+005 years Kb Half

- Life at pH 7: 1.871E+006 years (Total Kb applies only to esters, carbmates, alkyl halides) Bioaccumulation Estimates (BCFBAF v3.01): Log BCF from regression

- based method = 3.382 (BCF = 2408 L/kg wet

- wt) Log Biotransformation Half

- life (HL) = 0.8413 days (HL = 6.939 days) Log BCF Arnot

- Gobas method (upper trophic) = 3.372 (BCF = 2354) Log BAF Arnot

- Gobas method (upper trophic) = 3.643 (BAF = 4391) log Kow used: 5.63 (user entered) Volatilization from Water: Henry LC: 4.6E

- 005 atm

- m3/mole (Henry experimental database) Half

- Life from Model River: 34.83 hours (1.451 days) Half

- Life from Model Lake : 592.3 hours (24.68 days) Removal In Wastewater Treatment: Total removal: 89.74 percent Total biodegradati on: 0.75 percent Total sludge adsorption: 88.87 percent Total to Air: 0.12 percent (using 10000 hr Bio P,A,S) Level III Fugacity Model: Mass Amount Half

- Life Emissions (percent) (hr) (kg/hr) Air 0.582 51.2 1000 Water 5.68 1.44e+003 1000 Soil 56.1 2.88e+003 1000 Sediment 37.7 1.3e+004 0 ; Persistence Time: 2.61e+003 hr158

Attachment B: ECOSAR Output

SMILES : BrC(C(Br)CCC(Br)C(Br)CCC(Br)C(Br)C1)C1 CHEM : Cyclododecane, 1,2,5,6,9,10

- hexabromo

CAS Num: 003194

- 55

- 6 ChemID1: MOL FOR: C12 H18 Br6 MOL WT : 641.70 Log Kow: 7.742 (EPISuite Kowwin v1.68 Estimate) Log Kow: 5.625 (User Entered) Log Kow: (PhysProp DB exp value

- for comparison only) Melt Pt: (User Entered for Wat Sol estimate) Melt Pt: (deg C, PhysProp DB exp value for Wat Sol estimate) Wat S ol: 0.001343 (mg/L, EPISuite WSKowwin v1.43 Estimate) Wat Sol: 0.0086 (mg/L, User Entered) Wat Sol: 0.0086 (mg/L, PhysProp DB exp value)

-------------------------------------- Values used to Generate ECOSAR Profile

-------------------------------------- Log Kow: 5.625 (User Entered) Wat Sol: 0.0086 (mg/L, User Entered)

------------------------------------------------ Available Measured Data from ECOSAR Training Set

------------------------------------------------159

No Data Available

-------------------------------------- ECOSAR v1.1 Class

- specific Estimations

-------------------------------------- Neutral Organics Predicted ECOSAR Class Organism Duration End Pt mg/L (ppm) ================= ============ == ====== ====== ========== Neutral Organics : Fish 96

- hr LC50 0.293 * Neutral Organics : Daphnid 48

- hr LC50 0.232 * Neutral Organics : Green Algae 96

- hr EC50 0.690 * Neutral Organics : Fish ChV 0.042 * Neutral Organics : Daphnid ChV 0.057 * Neutral Organics : Green Algae ChV 0.381 * Neutral Organics : Fish (SW) 96

- hr LC50 0.376 * Neutral Organics : Mysid 96

- hr LC50 0.024 * Neutral Organics : Fish (SW) ChV 0.374 * Neutral Organics : Mysid (SW) ChV 0.000709 Neutral Organics : Ea rthworm 14

- day LC50 469.484 * Note: * = asterisk designates: Chemical may not be soluble enough to measure this predicted effect. If the effect level exceeds the water solubility by 10X, typically no effects at saturation (NES) are reported.

-- ---------------------------- Class Specific LogKow Cut

- Offs

------------------------------ If the log Kow of the chemical is greater than the endpoint specific cut

- offs presented below, then no effects at saturation are expected for those endpoints. Neut ral Organics:

---------------- Maximum LogKow: 5.0 (Fish 96

- hr LC50; Daphnid LC50, Mysid LC50) Maximum LogKow: 6.0 (Earthworm LC50) Maximum LogKow: 6.4 (Green Algae EC50) Maximum LogKow: 8.0 (ChV)

--------------------------------------------------