Technical Review of Hexabromocyclododecane (HBCD)

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

Office of Environmental Information

Environmental Analysis Division

Analytical Support Branch

January 25, 2016
Purpose 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 technical 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 the 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 with a brief review of the key findings 

presented in the following 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 158
4

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 

predominant 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 has 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 expected 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 degraded 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 subchronic 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 studies evaluating the carcinogenic potential of HBCD in humans or animals 

were not available so no determination of carcinogenicity could be made. HBCD, however, does 

not appear to be genotoxic. 

The developmental and reproductive toxicity of HBCD have been investigated in several 

rodent studies, and there is evidence of developmental and 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 methods, 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 considered 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 isomers (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 (+)-enantiomers, 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 determining 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; Dagani 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 under 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 than 
8

10 million to 50 million pounds during reporting years 1994, 1998, 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-containing 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 Koc value of 7.6×104, determined from a log Kow 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 

4.6×10-5atm-m3/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 model 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 

particulates. 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 also 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 8,974–13,085 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 compartments, as reported in numerous monitoring 

studies. Typically, γ-HBCD is the predominant diastereomer found in sediments and air, while αHBCD is the predominant diastereomer found in biota (Harrad et al., 2009b; de Wit et al., 2010; 

Marvin et al., 2011; 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 dehalogenation, 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 biodegradability (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%, respectively, in aerobic freshwater sediment, and 6 and 10%, respectively, in 
10

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 (ECB, 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 environmental 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. EPA Hexabromocyclododecane Action Plan (U.S. EPA, 2010b), and estimations performed using EPISuiteTM v4.10 

software (U.S. EPA, 2011b). Primary sources were consulted when available; however, data that 

originated with individual manufacturers and industry consortia were typically only available 

from secondary sources.
11

2.2. SYNONYMS/IDENTIFIERS 

(ACS, 2011; ChemIDPlus, 2011)

Cyclododecane, hexabromo1,2,5,6,9,10-Hexabromocyclododecane

Cyclododecane, 1,2,5,6,9,10-hexabromoBromkal 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 reported 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, 1H nuclear 

magnetic resonance (NMR) and 13C NMR data (Arsenault et al., 2007). 
12

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

physical-chemical property and environmental fate endpoints (Attachment A). Experimental 

values for log Kow (5.625), vapor pressure (4.70×10-7 mm Hg), and water solubility (8.6 μg/L), 

were used in the EPISuiteTM 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 EPISuiteTM v.4.10 calculations (U.S. EPA, 2011b). 

Table 2-1. Physical and chemical properties of HBCD

Property Valuea Reference

Molecular formula C12H18Br6

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 (14C-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 Kow 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., 2011 
13

Table 2-1. Physical and chemical properties of HBCD

Property Valuea 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

Koc 7.6×104(estimated) U.S. EPA, 2011b (EPISuiteTM

v4.10)

Henry’s Law 

constant

4.6×10-5atm-m3/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/cm3

2.38 g/cm3

(technical product)

Great Lakes Chemical 

Corporation, 1994 and Albemarle 

Corporation, 1994 as cited in 

ECB, 2008

Flash point Not applicable

aValues 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, EPISuiteTM v.4.10 was used to estimate values for 

physical-chemical property and environmental fate endpoints (Attachment A). Experimental 

values for log Kow (5.625), vapor pressure (4.70×10-7 mm Hg), and water solubility (8.6 μg/L), 
14

were used in the EPISuiteTM 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 EPISuiteTM 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 isomerization 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 αHBCD 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 pentabromocyclododecene (PBCD) occurs by a 

radical mechanism as shown in Figure 2-2 (Kroschwith, 1993 as cited in ECB, 2008; Barontini et 

al., 2001).

Source: Kirk-Othmer, 1993; ECB, 2008

Br Br

Br Br

Br Br

HBCD

Br Br

Br Br

light (hv)

Br

Br Br

Br Br

Br

.

. + Br

pentabromo- cyclododecene

+ HBr

Figure 2-2. Dehalogenation of HBCD by a radical mechanism
15

2.4. PRODUCTION METHODS

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 

perchloroethylene-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 et 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- double 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 stereoisomers. 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). Consistent 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 building 

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 6:13:80, 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 distribution 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 mattress 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 Koc to soil 

adsorption and mobility (Swann et al., 1983), an estimated Koc value of 7.6×104, calculated based 

on a log Kow 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 4.6×10-5atm-m3/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 (Great 
18

Lakes Chemical Corp., 1990 as cited in HSDB, 2011; Lyman et al., 1990; Stenzel and Nixon, 

1997). Using this Henry’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 river1

and 24.7 days for a model lake2. However, adsorption to suspended solids and sediment will 

reduce 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., 2009b). 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, 1988; 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 biota 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 transport 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/second, wind velocity of 5 m/second.

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

respectively, were determined for HBCD in rainbow trout (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., 2002; Covaci et al., 2006; ECB, 2008; U.S. EPA, 2010b; Letcher et al., 2009; 

Shaw et al., 2012). Higher concentrations of HBCD were measured in eggs from wild peregrine 

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; Marvin 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, the 

majority 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). In 
20

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

constituted 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 rates 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) estuarine 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 diastereomer 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 Degradation.

A summary of the environmental fate and persistence characteristics of HBCD is presented in 

Table 2-2.

Table 2-2. Environmental Degradation of HBCD

Property Valuea Reference

Air

PhotodegradationPhoto-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., 2009a
21

Table 2-2. Environmental Degradation of HBCD

Property Valuea 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 14C-labeled HBCD (4.67 mg/kg dry 

weight) after 112 days; method based on OECD 308

Davis et al., 2006 

Half-life: >120 daysb(estimated), based on a 15% decrease in total 

initial radioactivity in abiotic freshwater sediment of 14C-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 14C-labeled HBCD 

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

Davis et al., 2006

Half-life: >120 daysb(estimated), based on a 33% decrease in total 

initial radioactivity in abiotic freshwater sediment of 14C-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 daysb(estimated), based on a 10% decrease in total 

initial radioactivity in viable soil of 14C-labeled HBCD after 113 days; 

method based on OECD 307 using HBCD at 3.04 mg/kg dry weight

Davis et al., 2006
22

Table 2-2. Environmental Degradation of HBCD

Property Valuea Reference

Half-life: >120 daysb(estimated), based on a 6% decrease in total 

initial radioactivity in abiotic soil of 14C-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 daysb (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 

aValues are experimental unless otherwise indicated. 

bA 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 hydroxyl radicals in the atmosphere. An estimated rate constant of 

5.01×10-12 cm3/molecules-second at 25 °C for this reaction corresponds to a half-life of 26 hours, 

assuming an atmospheric hydroxyl radical concentration of 1.5×106 molecules/cm3and 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-HBCD 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 methanol 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 H2O2 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 

solutions. 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 freshwater 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 half-lives >120 days in soil and sediment due to 

minimal degradation being observed. Initial concentrations of 14C radiolabeled HBCD (α-, β-, 

and γ14C-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-cyclododecatriene (see 

Figure 2-4). 

Figure 2-4. Stepwise reductive dehalogenation of HBCD

Further degradation of 1,5,9-cyclododecatriene 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 aerobic 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 interconversion of HBCD 

isomers or α- and -HBCD degradation. The initial concentrations of HBCD were 0.025–0.089 
25

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) study. 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 Davis 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 spectrometry (LC/MS) chromatograms, and poor extraction of HBCD 

leading to HBCD recoveries 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 inoculum (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 during 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 this 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 daysin aerobic soil. An initial total 14C-HBCD concentration of 

3.0–4.7 mg/kg dry weight in the sediment and soil systems was used, allowing for quantification 
26

of individual isomers, metabolite identification, and mass balance evaluation (Davis et al., 2006; 

NICNAS, 2012). Although very high spiking rates can be toxic 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 biologically 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 βstereoisomers). 

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 ml 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 halflives of HBCD with sediment collected from Schuylkill 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, 14C-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, HBCD 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 ml 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 

microcosms were then incubated in the dark at 20 °C for up to 120 days. The disappearance halflives 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 anaerobic 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 than 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 recoveries 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 studies 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 using 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 14C-radiolabelled HBCD, 

quantify isomers other than γ-HBCD, identify degradation products, or establish a mass balance, 

but also procedural problems with contamination of sediment, incomplete HBCD extraction, and 

occurrence of an interfering peak in the LC/MS chromatograms corresponding to γ-HBCD).
29

Furthermore, 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 samples 

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 unknown, 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 typical 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 anaerobic 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 are 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 90% 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 pentabromocyclododecanols (PBCD-OHs) and tetrabromocyclododecadiols (TBCD-OHs) and LinA2 converts 

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. Environmental Distribution Assessment. U.S. EPA used the level III fugacity model 

within the EPISuiteTM 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/hour 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. Observed viable and abiotic 

degradation of HBCD during the most reliable 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 be 
31

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 (1×1020 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 EPISuiteTM model outputs. 

Table 2-3. Environmental distribution of HBCD 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); however, 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 

transformation 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 transformation
32

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 14C 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 500 °C in 100% nitrogen gas 

or air (21 mol % oxygen), hydrogen bromide was the major degradation product. Other highmolecular weight brominated and non-brominated organic compounds, such as PBCDs, TBCD, 

naphthalene, and biphenyl were detected under experimental conditions using Fourier transform 

infrared spectroscopy (FTIR) and gas-chromatographic (GC)/MS analysis (Barontini et al., 

2001). 

Debromination 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, TBCDOH, 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-[

14C]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, 

hexabromododecanedioic 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). These 
33

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 metabolic 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 metabolites 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 HBCD 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 sources of enrichment (Covaci et al., 

2006; Law et al., 2006; ECB, 2008; Marvin et al., 2011; NICNAS, 2012). 

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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 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 v1.00):

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)

For Henry LC Comparison Purposes:

User-Entered Henry LC: not entered

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 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 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 sorbed 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-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 (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 method)

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 hr
43

ATTACHMENT B: ECOSAR Output 
44

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 fatty 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 (LD50) of 

1258 mg/kg in male Charles 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 (LC50) of >200 mg/L (IRDC, 1977; Lewis and Palanker, 

1978; IRDC, 1978b; Gulf South Research Institute, 1988). An LD50 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 ml, 

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 

liver (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 postnatal 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, sperm 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 

effects 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 vehicle 
46

used (corn oil) may have affected observations at higher doses, including: increased mortality 

during 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 generation (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 

preputial 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 spontaneous 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 HBCD 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 and 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). However, 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 accumulation 

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 authors 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 [14C]-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 [14C]-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 [14C]-HBCD 

was detected in body tissues (excluding the gastrointestinal tract) 8 hours after dosing. In a 
48

similar study, Szabo et al. (2010) administered a single dose of 3 mg γ-[

14C]-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 and 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 samples (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). 

Abdallah 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 concentration 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 serum (=12) of pregnant women and their infants in the 35th

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 and 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 high 
50

contribution of γ-HBCD (39%) to the total HBCD serum concentration compared with findings 

from other studies.

In general, measured concentrations of HBCD within 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 1993 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-HBCD, 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 administered 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 (Szabo 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 differences 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 resulted 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 [14C]-HBCD/kg to male and female SpragueDawley rats, most of the administered [14C] 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. [14C]-HBCD concentrations in most tissues decreased over the course of 72 hours, but 

concentrations decreased more slowly in adipose tissue. The concentration of [14C] 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 consecutive daily doses of HBCD (500 mg/kgday) 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 Japanese). 
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 HBCD 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 SpragueDawley rats following a single oral dose (7 – 9 mg/kg) of [14C]-HBCD (Yu and Atallah, 1980). 

Unmetabolized HBCD, however, comprised all of the [14C] 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 isomeric 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 lactation (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 concentration of α- and γ-HBCD in 

the liver (expressed as μg/g lipid) of F1 males and females increased in a dose-related fashion. βHBCD 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 were 
53

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

respectively, 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 concentrations 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 detected 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 stereoisomers 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. αHBCD 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 rapid 

metabolism of γ-HBCD. Szabo et al. (2010) reported that approximately 80% of an administered 

dose of γ[14C]-HBCD was metabolites and approximately 15% was stereoisomerized HBCD 

products in the liver and feces. β- and γ-HBCD were detected in the liver, whereas α-, β-, and γHBCD 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 et 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 multiple isomers of monohydroxy-
54

pentabromocyclododecene, dihydroxy-pentabromocylcododecene, and dihydroxypentabromocyclododecadiene, while in liver and adipose tissue only a single monohydroxyptentabromocyclododecane metabolite was detected. The extent of metabolism 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 terminal 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 [14C]-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 (excluding 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 within 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 male Wistar rats administered 

five consecutive daily oral doses (500 mg/kg-day) of HBCD. The cumulative fecal excretion of 
55

HBCD increased in a linear fashion over 5 days of administration. Both the percentage of 

cumulative excretion measured (32 – 35%) for each 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[14C]-

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 γHBCD. The extent of excretion of β-HBCD was found to be similar to that of γ-HBCD (Sanders 

et al., 2013). Elimination, both 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 female 
56

rat 4-hours post-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 the 

treated animals. The study authors identified anLD50 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 LD50 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 dose 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 LD50 was greater than 

5,000 mg/kg.

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 LD50

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 female 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

Species/strain/test Exposure Result Notes Reference

Rat/Charles 

River/LD50

Single gavage (corn 

oil)

LD50 >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/ LD50 test of 

HBCD residue

Single gavage (corn 

oil)

LD50 = 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/LD50 Single gavage (30% 

aqueous tragacanth)

LD50 >6,400 

mg/kg; data not 

shown

7-day observation period; increasing 

apathy, trembling and late 

mortalities; peritonitis at necropsy.

BASF, 1990

Rat/SpragueDawley/limit test 

of HBCD bottoms

Single gavage (corn 

oil)

LD50 >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 lesions were observed.

Pharmakon 

Research 

International Inc., 

1990

Rat/NR/LD50 test Single gavage (corn 

oil)

LD50 >10,000 

mg/kg

No significant changes observed. Gulf South 

Research 

Institute, 1988
58

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 control 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 discharge. 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 period. 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 LC50 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 LC50 as greater than 200 mg/L. All animals 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 adult 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 LD50 for dermal toxicity of 3,969 mg/kg.

Male and female New Zealand white rabbits (3/sex) were administered 0.5 ml 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/LD50

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/LC50 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/LC50 test 1 hr. whole-body 

inhalation exposure 

to GLS-S6-41A

LC50 >200 mg/L No significant effects 

observed.

Gulf South Research 

Institute, 1988
60

concluded that HBCD is not a primary dermal irritant or a corrosive material. 

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 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 dermally toxic to 

rabbits. 

Male and female New Zealand albino rabbits (3/sex) were administered 0.5 ml 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 Dermal LD50 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 days 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 Photosensitization 

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 challenge 
61

dose greater than 0.05% elicited a positive response. The increase in the concentration of 

induction and challenge 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 ml 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 eye irritant. 

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 ml 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 subacute 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 the 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 Guideline 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 population. 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 continuous 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 of 
63

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, however, 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 weight 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 and 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 metaphysis (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 (CED = 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-day 
64

groups were untreated for a 14-day recovery period prior to necropsy. Animals were observed 

twice daily for mortality and morbidity, 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 (Week 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 in 

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 findings 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). Additionally, 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 cited 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% dose 
65

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 was 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 mg/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 necropsied, 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 points (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 

endpoints (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, Functional 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 after 
66

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 measures 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 transient 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 (Week 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, however, 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, respectively). 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 hypertrophy 
67

observed in 3/5 females in the 1,000 mg/kg-day group (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 female 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 effects 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 adverse 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, 0.16, 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/kgday. 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 relative 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 as “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 

temporary 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 doses 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, however, 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 pentabromocyclododecanes 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 same composite mixture. HBCD was dissolved in acetone, mixed with corn oil 

(allowing the acetone to evaporate), and then mixed with feed particles that were subsequently 
69

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

concentration 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 for 

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 parental (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. Litters 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 were 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 and 
70

microscopic pathology3, clinical chemistry4, bone analysis5, analysis of liver HBCD content, 

evaluation of apolar retinoid content of the liver, and assessment of serum hormone levels6. 

Hematological/immunological evaluation included an evaluation of standard red blood cell 

variables7, 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 males/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 obtained 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) depending 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 pathophysiology, 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 

 

3At necropsy, body weight and the weights of pituitary, thyroid, 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.

4Clinical chemistry included evaluation of serum albumin, alkaline phosphatase, alanine aminotransferase, total 

cholesterol, creatinine, glucose, total protein, and urea.

5Length of tibia, trabecular bone mineral density, cortical bone mineral density, total bone mineral density, total 

mineral content, total area, cortical area, and cortical thickness.

6Serum thyroid hormone concentrations; activity of 17 alpha-hydroxylase-17,20-desmolase (CYP17) in adrenals and 

of aromatase (CYP19) in ovaries.

7Red 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 more 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 unaffected 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 

weight 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, were 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 preputial 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 pathology 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 

of 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 variables 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), tibia cortical area 

(10.5% in males at 100 mg/kg-day), and tibia cortical thickness (5.2% in males at 100 mg/kgday). 

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 significant 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. Female 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 interactions 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 thymus weight in males; and length and trabecular bone density of the tibia 
73

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 of 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 were 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. Therefore, 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 diet. 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/kgday 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. Timeweighted 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 for 
74

an individual dietary concentration (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. Twenty 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 pyramidal 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 for 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 of pregnancy were 

not affected by HBCD exposure in this study. In parental rats, the relative weight of the thyroid 
75

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, the 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 any 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-response 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 (approximately 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 increased 
76

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)

Malesc Femalesc

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

T3 (ng/ml) 1.09 1.13 1.06 0.93** ND ND ND ND

T4 (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

T3 (ng/ml) 0.96 0.93 0.88* 0.89** ND ND ND ND

T4 (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 ND
78

Abbreviations: BW, body weight; PND, postnatal day; ppm, parts per million; PNW, postnatal week; ND, no data; 

NA, not applicable 

aGrade of change: (±) minimal; (+) slight; (++) moderate; (+++) severe

bNumber of animals with each grade

cn=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 through 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, 138, 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 geotaxis 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 PND 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 tested in a 
79

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 number 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 examined 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, liver 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 estrous 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 relative 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 rather 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 adults 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 relative testis weight was 
80

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 weanlings. 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 5αdihydrotestosterone (DHT) was observed in F1 males at 1,500 ppm. A significant decrease in the 

number of primordial follicles in the ovary was observed among 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 high-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 thyroid 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 and female pups in the highest dose 

group on PND 7, 14, and 21. These decreased body weights were associated with decreased food 
81

consumption. Relative liver weight was increased in the highest dose group for both F2 males 

and females. The study authors noted 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 opening 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 findings 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 (approximately 14 mg/kg-day) based on 

increased TSH in F0 females. A reproductive LOAEL of 1,500 ppm (approximately 138 mg/kgday) 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 follicle 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 6/24a 20/23b

F0 female 0/24 0/24 5/25a 11/23b

F1 male 0/24 0/24 2/22 11/24b

F1 female 0/24 1/24 5/24a 13/24b

aSignificantly different from control, p < 0.05.

bSignificantly 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 food 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 weight of live fetuses; or incidence of external or 
83

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 the 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 aforementioned 

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 dose 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 HBCD 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 dose 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 exposure 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, and 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 behavioral 

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 active 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 BioTest 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 tests 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 positive 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

Resultsb

–S9 +S9 Notes Reference

Eukaryotic systems, in vitro

S. typhimurium TA98, 

TA100, TA1537

3,000 μg

in DMSO – –

Doses ≥1,000 μg were partially 

insoluble.

Pharmakologisches 

Institute, 1978

S. typhimurium TA98, 

TA100, TA1535, 

TA1537, TA1538

250 μg 

(Firemaster, FM100, Lot 53, white 

powder)

in DMSO

– –

Doses ≥250 μg were insoluble. Industrial 

BioTest 

Laboratories, 

1990

1,000 μg (FM100, 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 – –

Zeiger 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. Doseresponse 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/ml (-S9) 

and

250 μg/m (+S9) 

in DMSO – (T) – (T)

Doses 750 – 2,500 were partially 

insoluble, and fully insoluble 

>2,500 μg/ml. 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, 1996
87

Table 3-6. Summary of genotoxicity studies of HBCD 

Test/species/strain/ 

route

Test doses

(per plate)a

Resultsb

–S9 +S9 Notes Reference

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 (χ2). 

Highest dose (1,000 μg/well) was 

cytotoxic.

Ethyl 

Corporation, 

1985b

Reversion assay

CHO/V79/Sp5 and 

SPD8

Intragenic 

recombination at hprt

locus in Sp5 (nonhomologous 

recombination) and 

SPD8 (homologous 

recombination) 

duplication cell lines

3–20 μg/ml

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

aLowest 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 attempted 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 correlations 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 with 
88

performance in the neurobehavioral tests. These authors conclude, however, that the potential 

for 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 order 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, brainstem 

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 development 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 evidence 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 HBCD
90

3.10. REFERENCES

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92

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study in albino rats. Test material GLS-S6-41A [unpublished]. Consumer Product Testing, 

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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 letter. Submitted under TSCA Section 8D; EPA Document No. 86-900000265; NTIS No. 

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aberrations in human peripheral blood lymphocytes with cover letter dated 12/12/1996. 

Submitted under TSCA Section 8D; EPA Document No. 86970000358; NTIS No. OTS0573552. 

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photosensitization of skin. [Article in Japanese; English abstract]. Eisei Shikenjo Hokoku 

111:18–24.

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additives in pregnant rats (7). Fetal toxicity of HBCD. Oyo Yakuri (Pharmacometrics) 29:981–

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Testing of Chemicals: Repeated Dose 28-Day Oral Toxicity Study in Rodents, Guideline 407.

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031290. Submitted under TSCA Section 8D; EPA Document No. 86-900000379; NTIS No. 

OTS0522942. 

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94

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flame retardant hexabromocyclododecane alpha: effect of dose, timing, route, repeated Exposure, 

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95

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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. Bioconcentration 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 birds (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 EC50 based on reduced growth in the marine algae, Skeletonema costatum) (Walsh et al., 

1987), and observed chronic aquatic 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 

kestrels (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 shrimp, 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. Studies 
97

conducted at concentrations above the water solubility of HBCD are essentially testing the 

effects at the maximum HBCD concentration 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. 

However, 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., 2012b; Hong et 

al., 2014a; Hu et al., 2009; Palace et al., 2008, 2010; 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 similar 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 bioaccumulation 

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 Database8, and Toxic Substances Control Act (TSCA)-related databases 

(i.e., TSCATS I9, TSCATS II10, and publicly available 8e and FYI notices11). In addition, the 

TOXLINE12 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 Pollution 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, effects on growth and reproduction, organ histopathology, abnormal behavior) were 

used to set toxicity effect levels such as no-observed-effect concentrations (NOECs) and lowestobserved-effect concentrations (LOECs). Although studies in which the nominal concentrations 

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 these studies, as none of these studies 

 

8Available at http://cfpub.epa.gov/ecotox/.

9Available at http://www.syrres.com/what-we-do/databaseforms.aspx?id=384.

10Available at http://yosemite.epa.gov/oppts/epatscat8.nsf/ReportSearch?OpenForm.

11Available at http://www.epa.gov/opptintr/tsca8e/pubs/8eandfyisubmissions.html#search.

12Available 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, and 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 activity (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 αHBCD, 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., 2012b). 

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 differences 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 dimethyl 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) recommends 
100

that a preliminary 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 basis.” 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 been 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, insufficient 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 

biomagnification) 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 biomagnification 

studies (Du et al., 2012a, 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 environment (e.g., through an organism’s 
101

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 burdens 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 toxicity 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 Section 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 

typea,b

Duration and test 

endpoint

Concentration of 

HBCDc

Comments Reference

Acute aquatic toxicity

Freshwater algae

Green algae 

(Selanastrum 

capricornutum)

S, Md,e 96-hour EC50 >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, Uf 96-hour EC50 >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, Md,e 48-hour EC50 >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 test concentration.

Graves and 

Swigert, 1997a 48-hour NOEC 0.0032 mg/L

Water flea (D. magna 

Straus)

S, Ug 48-hour EC50 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/L
103

Table 4-1. Ecological toxicity values for HBCD 

Species Experimental 

typea,b

Duration and test 

endpoint

Concentration of 

HBCDc

Comments Reference

Freshwater fish

Rainbow trout 

(Oncorhynchus 

mykiss)

F, Md,e 96-hour LC50 >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 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.

Graves and 

Swigert, 1997b 96-hour NOEC 0.0025 mg/L

Bluegill sunfish 

(Lepomis 

macrochirus 

Rafinesque)

S, Uh,i,j 96-hour LC50 >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, Uk 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 unknown. 

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 significantly 

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. Associated 

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/L
104

Table 4-1. Ecological toxicity values for HBCD 

Species Experimental 

typea,b

Duration and test 

endpoint

Concentration of 

HBCDc

Comments Reference

Zebrafish (D. rerio) 

embryos

R, Uk 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) concentration 

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, and 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, Uk 72-hour NOEC 

(survival, hatching 

rate, malformations)

0.13 mg/L1 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 

concentrations 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., 2013
105

Table 4-1. Ecological toxicity values for HBCD 

Species Experimental 

typea,b

Duration and test 

endpoint

Concentration of 

HBCDc

Comments Reference

Zebrafish (D. rerio) 

embryos

R, Uk 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 formation.

Du et al., 2012b

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/L
106

Table 4-1. Ecological toxicity values for HBCD 

Species Experimental 

typea,b

Duration and test 

endpoint

Concentration of 

HBCDc

Comments Reference

Estuarine/marine algae

Algae (S. costatum) S, Mm,n 72-hour EC50 >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, Mn,o 72-hour EC50 

(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 EC50 of 52 μg/L. Primary report not 

available. 

Desjardins et al., 

2005 as cited in 

ECB, 2008

Algae (S. costatum) S, Me 72-hour EC50 >0.01 mg/L Test solutions were prepared using dimethylformamide 

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, Mp 72-hour EC50

(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. 

Maximum 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. EC50 ranges 

reflect results in the different media tested.

Walsh et al., 

1987

Algae (Thalassiosira 

pseudonana)

S, Mp 72-hour EC50

(growth)

0.05 – 0.37 mg/L

(0.08 mg/L in SW)

Algae (Chlorella Sp.) S, Mp 96-hour EC50

(growth) 

>1.5 mg/L
107

Table 4-1. Ecological toxicity values for HBCD 

Species Experimental 

typea,b

Duration and test 

endpoint

Concentration of 

HBCDc

Comments Reference

Estuarine/marine invertebrate

Sea urchin 

(Psammechinus 

miliaris) embryos

R, Uk 72-hour LOEC 

(morphological 

abnormalities)

0.064 mg/Ll 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 

was 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. 

Anselmo et al., 

2011

72-hour NOEC 0.032 mg/Ll

Estuarine/marine fish - No data

Chronic aquatic toxicity

Freshwater invertebrate

Water flea (D. 

magna)

F, Md,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 Chemistry 

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, Mq 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 sediment
108

Table 4-1. Ecological toxicity values for HBCD 

Species Experimental 

typea,b

Duration and test 

endpoint

Concentration of 

HBCDc

Comments Reference

Amphipod (H. 

azteca)

F, Mq 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 concentration.

Thomas et al., 

2003b

Amphipod (H. 

azteca)

F, Mm 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 

were 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., 

2003a
109

Table 4-1. Ecological toxicity values for HBCD 

Species Experimental 

typea,b

Duration and test 

endpoint

Concentration of 

HBCDc

Comments Reference

Worm

(Lumbriculus 

variegatus)

S, Mr,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 

conducted 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, Mp,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 rarus)

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 investigations 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., 

2008
110

Table 4-1. Ecological toxicity values for HBCD 

Species Experimental 

typea,b

Duration and test 

endpoint

Concentration of 

HBCDc

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 behavior, 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, NSp 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, Uk 14 – 16-day LOEC 

(morphological 

abnormalities)

0.064 mg/Ll 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, respectively) from day 0 to 16 postfertilization. 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). Hatching 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/Ll
111

Table 4-1. Ecological toxicity values for HBCD 

Species Experimental 

typea,b

Duration and test 

endpoint

Concentration of 

HBCDc

Comments Reference

Estuarine/marine fish

Marine medaka 

(Oryzias melastigma) 

enbryos

R, Uk 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 0.050 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 investigations 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, Uk 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 metamorphosis 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 larvae through 

completion of metamorphosis.

Foekema et al., 

2014
112

Table 4-1. Ecological toxicity values for HBCD 

Species Experimental 

typea,b

Duration and test 

endpoint

Concentration of 

HBCDc

Comments Reference

Terrestrial toxicity and phytotoxicity

Avian species

Japanese quail 

(Coturnix coturnix 

japonica)

Diet exposedv 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 details were provided.

MOEJ, 2009 as 

cited in UNEP, 

2010

Japanese quail (C. 

coturnix japonica)

Diet exposedv 6-week LOAEL 

(reduced chick 

survival)

15 ppm (2.1 mg/kgday)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/kgday) 

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 exposedw,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 average tissue concentrations of 163.5 ng αHBCD/g ww, 13.9 ng β-HBCD/g ww, and 2.6 ng γHBCD/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 control eggs during incubation, but egg 

shell thickness was not affected. No effect on 

reproductive success was identified.

Fernie et al., 

2011
113

Table 4-1. Ecological toxicity values for HBCD 

Species Experimental 

typea,b

Duration and test 

endpoint

Concentration of 

HBCDc

Comments Reference

American kestrel (F. 

sparverius), male

Diet exposedx LOAEL 

(increased testes 

weight in unpaired 

males) 

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. Unpaired males 

(exposed for 3 weeks) were used for evaluation of 

testis 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 testis 

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 counts. 

Marteinson et 

al., 2011

American kestrel (F. 

sparverius)

Diet exposedx 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 

courtship 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 reduced 

incubation nest temperature.

Marteinson et 

al., 2012
114

Table 4-1. Ecological toxicity values for HBCD 

Species Experimental 

typea,b

Duration and test 

endpoint

Concentration of 

HBCDc

Comments Reference

Chicken (Gallus 

gallus domesticus)

In ovo exposedy 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/ml; 

respective nominal doses of HBCD were 0, 50, 100, 

300, 1,000, and 10,000 ng/g egg. Eggs were observed 

for 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)

apH and water hardness were within accepted ranges (6 – 8.5 and 50 – 250 mg/L as CaCO3, respectively) unless noted otherwise.

bExperiment type: S = static, R = renewal, F = flow-through, M = measured, U = unmeasured, NS = not specified.

cmg/L = ppm, units are in terms of HBCD concentration.
115

dTest composition was an isomeric mixture of HBCD (8.5% beta isomer, 6% alpha isomer, and 79.1% gamma isomer).

eSolvent: dimethylformamide (DMF).

fSolvent: Cremophor RH40.

gSolvent: Tween80.

hReplicates per concentration were not used.

iDissolved oxygen concentrations decreased to a minimum level of 1.9 mg O/L.

j

Insufficient acclimatization of test organism.

kSolvent: DMSO.

lConcentrations in mg/L were calculated from molar concentrations using a molecular weight of 641.7 g/mole.

mTest substance was an isomeric mixture of HBCDs (5.15% beta isomer, 7.67% alpha isomer, and 83.04% gamma isomer).

nA single concentration was tested.

oSaturated solution was tested.

pSolvent: acetone.

qTest substance was an isomeric mixture of HBCDs (19.3% beta isomer, 5.8% alpha isomer, and 74.9% gamma isomer).

rSolvent: ethyl acetate.

sSediment of a coarser grain size.

tTest composition was an isomeric mixture of HBCD (6.3% beta isomer, 9.4% alpha isomer, and 84.3% gamma isomer).

uSolvent: peanut oil.

vTest substance was an isomeric mixture of HBCDs (30% beta isomer, 27% alpha isomer, and 43% gamma isomer).

wTest substance was HBCD technical mixture (HBCD-TM), composed of 6% alpha isomer, 8% beta isomer, and 80% gamma isomer.

xVehicle: safflower oil

yTest substance was technical HBCD (HBCD-TM; 11% beta isomer, 12% alpha isomer, and 77% gamma isomer) initially dissolved in acetone to increase its 

solubility 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 = heat 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-effect-concentration; NQ = not quantified; PROD = pentoxyresorufinO-deethylase; ROS = reactive oxygen species; SV-BA = sinus venosus-bulbus arteriosus; T4 = thyroxine; TOC = total organic carbon; ww = wet weight
116

4.2.1. Acute Aquatic Toxicity. Acute aquatic toxicity 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, 2008 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., 2012b; Hu et al., 2009; Wu 

et al., 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 embryos, and 0.064 mg/L 

(nominal) in sea urchin embryos, indicate high acute hazard concern. 

Walsh et al. (1987) reported measured 72-hour EC50 values in S. costatum ranging from 

0.009 to 0.012 mg/L based on reduced growth rate in five different types of saltwater 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. EC50 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 did not result in crystal formation, and multiple concentrations 

were tested. EC50 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 concentrations were 

measured. Desjardins et al. (2004 as cited in ECB, 2008 and IUCLID, 2005) reported 
117

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 EC50 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 indicated 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 postfertilization. Survival of embryos/larvae was significantly reduced at all tested 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, was 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. Increased 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, and 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 et 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 

observed 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) for 

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 cardiomyocytes 
119

(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. (2012b) exposed 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 

lowest 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, measured 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-HBCD. 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 heart 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. (2012b) 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 embryo 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 of 
120

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. These 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-fertilization. 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 concentration 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. 

Experimental 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 and Swigert, 1997; Siebel-Sauer and Bias, 1987 as cited in ECB, 2008). Jatzek (1990) 

reported a LOEC of 10 mg/L and an EC50 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 that 
121

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 saltwater fish species. 

ECOSAR-predicted 96-hour LC50 values for estuarine/marine fish were above the water 

solubility limit13.

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-recommended 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 not 

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 reported 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 

 

13For the ECOSAR predictive tool, values that exceed a Kow 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.
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 edema, 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 life 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). Each 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 thyroid 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 mechanisms 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 antioxidant 

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 concentration 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 intraperitoneal 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: the 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 acute 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 Program (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 LC50

values for estuarine/marine fish and mysid shrimp and the chronic value (ChV) for 

estuarine/marine fish were greater than the water solubility limit14. The predicted ChV for mysid 

shrimp was 0.0007 mg/L, indicating very high expected toxicity. The values of the 

physicochemical parameters (log Kow = 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 

500 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-adverseeffect-level (LOAEL) and no-observed-adverse-effect-level (NOAEL) values of 15 and 5 ppm, 

respectively, based on significant reduction of survival of chicks hatched from eggs of HBCDfed quails (MOEJ, 2009 as cited in UNEP, 2010). 

Several studies have been conducted examining effects of HBCD on American kestrels 

(F. sparverius). Kobiliris (2010) reported a reduced “corticosterone response” (where 

 

14For the ECOSAR predictive tool, values that exceed a Kow 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 weight 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 the 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 contributed 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 

observed 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 survival 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 have

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 

HBCD, 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 αHBCD > 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 Asia 
127

(Marvin 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 subsequent 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 Concentrationa Comments Reference

Aquatic invertebrates

Green mussels (Perna viridis) and blue 

mussels (Mytilus edulis)

NDb – 1,400 ng/g lw ƩHBCDc 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 – 2002d:

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 eeld; 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 – 11.99 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., 

2014
129

Table 4-2. Environmental monitoring data for HBCD 

Species Concentrationa 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 (Oreochromis 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 NDe

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., 2012
130

Table 4-2. Environmental monitoring data for HBCD 

Species Concentrationa 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 lwf Birkenes, Norway Salamova and Hites, 

2013 1.39 ± 0.75 ng/g lwf Bukit Kototabang, Indonesia

1.45 ± 0.78 ng/g lwf Cape Grim, Tasmania

0.66 ± 0.16 ng/g lwf De Aar, South Africa

0.35 ± 0.10 ng/g lwf Dhulikhel, Nepal

21.3 ± 7.7 ng/g lwf Downsview, Toronto, Ontario, Canada

0.84 ± 0.13 ng/g lwf Frasardale, Ontario, Canada

1.65 ± 0.94 ng/g lwf Košetice, Czech Republic

0.62 ± 0.09 ng/g lwf Reykjavik, Iceland

0.76 ± 0.65 ng/g lwf Tula, American Samoa, United States

0.27 ± 0.03 ng/g lwf Whistler, British Columbia, Canada

ND Malin Head, Ireland
131

Table 4-2. Environmental monitoring data for HBCD 

Species Concentrationa 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 collected 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 eggsg 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)

Nesting 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., 2015
132

Table 4-2. Environmental monitoring data for HBCD 

Species Concentrationa Comments Reference

Light-vented bulbul (Pycnonotus 

sinensis), long-tailed shrike (Lanius 

schach), and oriental magpie-robin 

(Copsychus saularis)

e-Waste locationsh:

ƩHBCD = 5.1 – 73 ng/g lw

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 liveri 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)

Muir et al., 2006
133

Table 4-2. Environmental monitoring data for HBCD 

Species Concentrationa Comments Reference

Polar bear (U. maritimus) 0.9 – 41.1 ng/g lwj 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

aww = wet weight; lw = lipid weight; dw = dry weight

bND = not detected

cƩHBCD = summed concentration of α-, β-, and γ- isomers.

dValues are the range of mean values for different locations.

eHBCD values were not provided for each species. The mean and range values are composite values of the eight species.

fValues are mean ± standard error.

gValues are the range of median values for different locations.

hValues are the range of median values for different species.

iHBCD was detected in 2 of 141 animals.

jRange 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; BAFs 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 may 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 greater 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 bioaccumulation 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 BCFa,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 BCFa,c 8,974d,e,f and 13,085d,e,f Drottar and Krueger, 

2000

Fathead minnow 

(Pimephales promelas)

32-day BCF 18,100f,g Veith et al., 1979

Mirror carp (Cyprinus 

carpio morpha noblis)

30-day exposure and 30-day 

depuration BCFh

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 catfish 

(Hypostomus plecostomus)

Log BAFi.j 4.8 – 7.7 for HBCD isomers 

(α-HBCD had higher BAFs than 

β- and γ-HBCD)

He et al., 2013

Freshwater food webk Log BAFl,m α-HBCD: 2.58 – 6.01 

β-HBCD: 3.24 – 5.58

γ-HBCD: 3.44 – 5.98

ƩHBCDs: 2.85 – 5.98i

Wu et al., 2011

Ragworm (Hediste 

diversicolor)

28-day BSAFn,o ≤0.006 Haukås et al., 2010b

28-day assimilation efficiency 67%

Bivalve (Corbicula 

fluminea)

Log BSAFn,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 BSAFn,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 BMFq α-HBCD: 9.2

β-HBCD: 4.3

γ-HBCD: 7.2

Law et al., 2006
136

Table 4-3. Bioaccumulation studies for HBCD

Species Duration and test endpoint Value Reference

Zebrafish (Danio rerio) 42-day exposure and 21-day 

depuration BMFr

α-HBCD: 12.33 – 29.71

β-HBCD: 7.34 – 11.63

γ-HBCD: 7.61 – 7.76

Du et al., 2012a

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 BMFs

±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-HBCDt 1.7 Letcher et al., 2009

Harbor seal (Phoca 

vitulina)

BMF for α-HBCDu 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)

BMFv 1.3 – 2.6 for α-HBCD in e-waste 

area; 4.9 – 30 for α-HBCD in 

urban area; 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)

BMFw 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 Norwayx

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, wwbased); 0.8 – 26 (α-HBCD, lwbased); 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)

TMFy 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 webbb

TMFaa 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 webcc TMFdd 2.22, 2.18, and 2.19 for ̶+α-

, -α-, and ±α-HBCD, 

respectively 

Wu et al., 2010

Marine aquatic food webee TMFdd 1.74 and 1.68 for α-HBCD and 

ƩHBCD, respectively

Zhang et al., 2013

Freshwater food webee 2.58 and 2.44 for α-HBCD and 

ƩHBCD, respectively

aBCF calculated at steady state.

bBCF calculated using tissue concentration based on wet weight.

cBCF calculated using whole-fish tissue concentrations.

dTest substance was an isomeric mixture of HBCDs (4.5% beta isomer, 6.4% alpha isomer, and 79.1% gamma 

isomer).

eBCF 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 steady state was achieved at the higher concentration. See text for additional discussion.

fSolvent used.

gA depuration phase was excluded from the study.

hBCFs were calculated separately for viscera, gill, skin, and muscle (tissues listed in order of decreasing 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.

iBAF calculated using muscle tissue concentration based on lipid weight.

jBAF calculated from concentrations of HBCD in samples collected from a river running through a highly 

industrialized area in South China.

kSpecies included Chinese mysterysnail (Cipangopaludina chinensis), prawn (Macrobrachium nipponense), mud 

carp (C. molitorella), crucian carp (Carassius uratus), northern snakehead (Ophicephalus argus), and water snake 

(Enhydris chinensis).

lBAF calculated using concentration of HBCD in whole organisms (wet weight).

mBAF calculated from concentrations of HBCD in samples collected 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

nBSAF = lipid-normalized tissue concentration ÷ TOC-normalized sediment concentration.

oHBCD-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.

pBSAF calculated from concentrations of HBCD in samples collected from 0 to 44.6 km downstream of a North 

Carolina wastewater treatment plant that treats wastewater from a textile manufacturer.

qThe BMF was calculated from the equation BMF = (AE × F)/kd, where F is the feeding rate on a lipid basis and kd

is the depuration rate constant and AE is the assimilation efficiency.

rFish food was spiked with α-, β-, or γ-HBCD.

sFish food was spiked with (±)α-, (±)β-, or (±)γ-HBCD.

tAdipose 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.

uBMFs were calculated 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).

vThe lipid-normalized 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.

wHBCD 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 HBCD 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.

xThe 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.

yTMF was calculated as the average rate of increase of HBCD from one trophic level to the next.

zβ-HBCD was not detected in samples according to study authors.

aaTMF was calculated as the slope of the curve of the trophic level versus the logarithm of the body residue.

bbSamples were collected from pelagic and benthic food webs, including zooplankton, benthic invertebrates, benthic 

and pelagic fish, jellyfish, and a piscivorous bird species.

ccSpecies include Chinese mysterysnail (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.

ddTMF was calculated using the same method as in Brandsma et al., 2015.

eeMany 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 diastereomer accumulation in various species and the extent to which 
139

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 

physiological 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. (2012a) 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-HBCD (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 literature 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. Concentrations 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 fish 
140

exposed 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 

conducted 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 indicate time when steadystate 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 bioaccumulative 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 

freshwater 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- and 
141

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 and 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 diastereomers distributed to liver, brain, and muscle. Assimilation 

efficiencies of HBCD for individual diastereomers (percent of HBCD isomer in the fish relative 
142

to the amount in the bolus fed to the fish at the start of the experiment) were >100% for βHBCD, 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. 

(2012a) 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 trout 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 remainder of the 

experiment for α-HBCD (overall depuration rate was not determined). Depuration rates of 

0.44×10-2and 0.48×10-2per day were found for β-HBCD and γ-HBCD, respectively. Steadystate was not reached for any of the diastereomers within the 52-day exposure period. 

Additional studies are available that support the conclusion that HBCD has the potential 

to biomagnify. Studies of zebrafish by Du et al. (2012a, 2013) reported diastereo- and 

enantiomer-specific biomagnification. When BMFs were calculated for 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 lower 
143

than those from Du et al. (2012a). 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, reporting 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 between 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 (Delphinapterus 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 determine 
144

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 in 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 aquatic 

and terrestrial organisms and biomagnify in food webs is high. 
145

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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-hexabromoMOL 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-Entered 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 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 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 Hrs
157

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 (KOCWIN 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 biodegradation: 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 hr
158

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-hexabromoCAS 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 Sol: 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

------------------------------------------------

Br

Br CH2 CH2

Br

Br

CH2

CH2

Br

Br

CH2

CH2

Cyclododecane, 1,2,5,6,9,10-hexabromo- 
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 : Earthworm 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.

Neutral 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)

--------------------------------------------------