Supporting Information 

Hexabromocyclododecane: Current Understanding of Chemistry, 

Environmental Fate and Toxicology and Implications for Global 

Management 

Christopher H. Marvin1* and Gregg T. Tomy2,3*; James M. Armitage4,5; Jon A. Arnot4; 

Lynn McCarty6; Adrian Covaci7 and Vince Palace2,3 

1

 Water Science and Technology Directorate, Environment Canada, Burlington, Ontario 

L7R 4A6 Canada 

2

 Fisheries & Oceans Canada, Winnipeg, Manitoba R3T 2N6 Canada 

3

 University of Manitoba, Faculty of Environment and Geography, Winnipeg, Manitoba, 

R3T 2N2 Canada 

4

 University of Toronto Scarborough, Ontario, M1C 1A4, Canada 

5

 Aarhus University Hospital, Department of Occupational Medicine, Aarhus, Denmark 

6

 LS McCarty Scientific Research and Consulting, Ontario, Canada 

7

 Toxicological Centre, University of Antwerp, 2610 Wilrijk, Belgium 

*To whom correspondence may be addressed. 

Email: chris.marvin@ec.gc.ca and/or gregg.tomy@dfo-mpo.gc.ca

Phone : (CHM: 905-319-6919; GTT: 204-983-5167) 

72 pages 

1 figure 

22 tables 
S-2

Section Contents Page 

S.1 

S.2 

S.3 

S.4 

S.5 

S.6 

S.7 

S.8 

S.9 

S.10

Physical-chemical properties 

Additional details on data characterizing the degradation of 

HBCD 

Additional details on data and approaches used to 

characterize Overall Persistence (POV) 

Additional details on data characterizing the 

bioaccumulation potential of HBCD 

Modeled bioavailability, uptake efficiencies and inherent 

bioaccumulation factors (BAFs) 

Biotransformation and bioisomerization half-life estimation 

Additional details on data and approaches to characterize 

Long-range Transport (LRT) Potential 

Summary of monitoring data from near-point source, 

industrialized regions and remote sites 

Analytical Chemistry 

Additional Toxic Endpoints 

Literature Cited 

S-3 

S-15 

S-20 

S-24 

S-25 

S-30 

S-38 

S-41 

S-43 

S-47 

S-63 
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Section S.1: Physical-Chemical Properties

Measured and modeled physical-chemical properties for (technical) t-HBCD and the 

three main diastereoisomers (α-, β- and γ-HBCD) were compiled and reviewed as 

detailed below. The physical-chemical properties considered of most relevance for 

environmental chemistry include: molar mass (M; g·mol-1), melting point (TM; 

oC or K), 

water solubility (S; mol·L

-1), vapor pressure (P; Pa), Henry’s law constant (H; Pa·m3

·mol1

), and the dimensionless equilibrium partition coefficients between air and water (KAW), 

octanol and water (KOW), and octanol and air (KOA). Based on this review of the data, 

values were selected to obtain thermodynamically consistent values, or Final Adjusted 

Values (FAVs), using recommended methods (SI-1). The FAVs were then used for 

evaluating toxicity testing data and for mass balance modeling. Much of the present 

review is based on a previous unpublished assessment of HBCD (SI-2). 

Molar mass 

The molar mass (molecular weight) of HBCD is 641.7 g·mol-1 (SI-3). 

Melting point 

The melting points for the HBCD diastereoisomers have been reported as 179 to 181 °C, 

170 to 172 °C and 207 to 209 °C for α-, β-, and γ-HBCD, respectively (SI-3). The melting 

intervals for two commercial HBCD products are 175 to 183 °C and 187 to 195 °C (SI-3). 

A melting point of 190 °C (463 K) was selected as an average value as input data for 

EUSES in the European Union Risk Assessment Report (EU RAR) (SI-4) and is selected 

as representative for t-HBCD in this critical review; however, different isomeric mixtures 

will result in different melting points for t-HBCD. 

Fugacity ratio 

The fugacity ratio F (unitless) is required to correct for the influence of intermolecular 

interactions of the solid phase and is necessary for estimating the properties of the 
S-4

subcooled liquid state of the chemicals which are used for the thermodynamic 

consistency calculations. The fugacity ratio can be estimated as 

F = exp(6.79(1 – TM/T)) (S.1) 

where T is the system temperature (K). F is 0.023 for t-HBCD (TM = 463 K) assuming 

that the entropy of fusion is 56.5 J·K

-1

·mol-1 (SI-5). For example, for vapor pressure P, the 

solid state property data (subscript S) can be corrected to the subcooled liquid state values 

(subscript L) as: 

PL = PS/F (S.2) 

The fugacity ratio is expected to be different for the diastereoisomers due to different 

melting points and different entropies of fusion; however, isomer specific entropies of 

fusion are not available. The diastereoisomer specific Fs are 0.029, 0.036, and 0.015 for 

α-, β-, and γ-HBCD, respectively, using the assumed entropy of fusion above. 

Considering there is uncertainty for the entropy of fusion at the melting point for all 

diastereoisomer specific calculations, a value of 0.023 was selected for F for 

simplification of the present calculations. Isomer-specific measurements for the entropy 

of fusion are required to obtain more accurate estimates of the subcooled liquid state 

property values. 

Water solubility 

Table S.1 summarizes two studies that have measured the solid state water solubilities SS

for the predominant HBCD isomers using the generator column method (SI-6,7). These 

tests are in general accordance with OECD guidelines and are compliant with Good 

Laboratory Practise (GLP). The test substance in the first study was a composite sample 

consisting of 8% α-, 5.37% β- and 86.63% γ-HBCD (SI-6). The test substance in the 

second study was a composite sample consisting of 6.0% of α-, 8.5% of β-, and 79.1% γHBCD; however, no diastereoisomer-specific analyses were conducted (SI-7). The first 

study (SI-6) had fewer impurities and unknowns then the second study (SI-7) and all 
S-5

isomers could be quantified; therefore, the data from (SI-6) are considered to be the most 

reliable measurements available. The SS (μg·L

-1) are 48.8±1.9, 14.7±0.5, 2.1±0.2 for α-, 

β-, and γ-HBCD, respectively at 20 °C (SI-6). The water solubility limit for t-HBCD was 

reported as 65.6 μg·L

-1, i.e., the sum of the individual diastereoisomers (SI-6). 

When t-HBCD is added to water all three diastereoisomers are assumed to dissolve as 

discussed and illustrated in the EU RAR (SI-8): 

“γ-HBCD will reach its maximum solubility at 2.1 μg/l (with a total 

HBCD dissolved concentration of 2.4 μg/l), β-HBCD at 14.7 μg/l (with a 

total HBCD dissolved concentration of 39 μg/l) and α-HBCD at 48.8 μg/l 

(with a total dissolved HBCD concentration of 65.6 μg/l). At this level, 

65.6 μg/l dissolved HBCD, the total content of HBCD in the water is 610 

μg/l, with 544.4 μg/l representing non-dissolved γ- and β-HBCD.” 

In other words, testing for possible effects of t-HBCD at concentrations greater than 

about 2 or 3 μg·L

-1 will result in exposures to test organisms of non-dissolved 

(precipitate) γ-HBCD, thus compromising the reliability of the water based toxicity test 

results (SI-9). This also has implications for interpreting in vitro test data. 

Table S.1. Solid state water solubility data for HBCD at 20 °C as summarized in (SI-8). 

Measured water solubilities SS for α-, β- and γ-HBCD correspond to approximately 0.08, 

0.02 and 0.003 μM, respectively. 

Test substance Water solubility (μg·L

-1)* Reference 

α-HBCD 48.8±1.9 (CV = 2.8%) 

β-HBCD 14.7±0.5 (CV = 3.4%) 

γ-HBCD 2.1±0.2 (CV = 10.5%) 

(SI-6) 

Σ HBCD (technical product) 65.6 (SI-6) 

HBCD (technical product) 3.4±2.3* (SI-7) 

*25 °C; CV – Coefficient of Variation 
S-6

Model estimates for the water solubility were also obtained and reviewed. These 

estimates do not differentiate between diastereoisomers. When a log KOW value of 5.63 is 

used as input (see below), EPI Suite Ver. 4.0 estimates SS values of 1.3 μg·L

-1 and 3.0 

μg·L

-1 from the WSKOWWIN Ver. 1.41 and WATERNT Ver. 1.01 submodels, 

respectively (SI-10). These SS estimates are generally consistent with the measured 

values for γ-HBCD. The SPARC model Ver. 4.2 (http://ibmlc2.chem.uga.edu/sparc/) 

predicts a SS of 0.31 μg·L

-1 for 1,2,5,6,9,10-HBCD, which is consistent with the 

diastereoisomer configurations (vicinal bromine substitution). The SPARC SS estimate is 

about 1 order of magnitude lower than the measured values for γ-HBCD. 

The measured SS for the individual HBCD diastereoisomers from Study 1 (SI-6) are 

considered to be the most reliable measurements available. The SS (mol·m

-3) estimates 

are 7.60×10-5, 2.29×10-5, 3.27×10-6 for α-, β-, γ-HBCD diastereoisomers, respectively. 

These estimates correspond to subcooled liquid water solubility values SL (mol.m-3) of 

3.27×10-3, 9.86×10-4, 1.41×10-4 for α-, β-, γ-HBCD diastereoisomers, respectively. A 

value of 3.0 μg·L

-1 (SS = 4.67×10-6 mol.m-3, SL = 2.01×10-4 mol.m-3) is selected as a 

“representative” value for t-HBCD for the thermodynamic consistency calculations. 

Vapor pressure 

Measured values of the solid state vapor pressure PS for HBCD include 1.6×10-5 Pa at 20 

oC using the effusion method and 6.27×10-5 Pa at 21 oC using the spinning rotor method. 

The effusion method used in the aforementioned test and the spinning rotor method are 

not recommended for substances with vapor pressures <10-4 Pa (SI-11). The composition 

of HBCD in the effusion method is undetermined and the following percentages of 

diastereoisomers were reported for the spinning rotor method: α-HBCD 6.0; β-HBCD 

8.5; and γ-HBCD 79.1. According to OECD guidelines (SI-11), both of these methods are 

not recommended for substances with vapor pressures as low as the values measured for 

HBCD; however, both methods produced similar values. Wania (SI-12) reported a liquid 

state vapor pressure PL (subcooled) of 2.41×10-5 Pa at 25oC for t-HBCD using the gas 

chromatographic-retention time method, which is suitable for substances with vapor 

pressures as low as 10-5 Pa. This corresponds to a PS of 7.23×10-7 Pa (F = 0.023). 
S-7

EPI Suite Ver. 4.0 estimates a PS of 2.25×10-6 Pa (SI-3) and SPARC Ver. 4.2 estimates a 

PS of 2.63×10-9 Pa (http://ibmlc2.chem.uga.edu/sparc/). The EPI Suite estimate is about 3 

times higher than the measured estimate by Wania whereas the SPARC estimate is 275 

times lower. 

At present there are no isomer-specific estimates or measurements for vapor pressure. 

Goss and colleagues (SI-13) used the COSMOtherm model to calculate KOW and KAW

partitioning of α-, β-, γ-HBCD diastereoisomers at 25oC (see details below). From the 

isomer-specific quantum chemical software estimates of KAW and from isomer-specific 

estimates of water solubility SS it is possible to estimate PS for the three isomers as 

PS = KAW.RT.SS (S.3) 

where R is the gas constant (8.314 Pa·m

3

(mol.K)-1) and T is temperature (298 K). This 

approach assumes that temperature differences (20 to 25 oC) are minimal. The resulting 

isomer-specific estimates for PS are 2.72×10-10 Pa, 3.58×10-11 Pa, and 1.86×10-11 Pa for α-

, β-, γ-HBCD, respectively. These model estimates are about 104 times lower than the 

values measured by Wania. Goss et al. (SI-13) note that “the absolute values of the 

calculated air-water partition coefficients may be considerably off the real values”. 

COSMOtherm also underestimates KAW for hexachlorocyclohexane (HCH) (SI-13). Since 

the KOW estimates for HCH and HBCD are similar to measured estimates, this suggests 

that errors in COSMOtherm KAW estimates relate to errors in P. Thus, the COSMOtherm 

estimates for KAW and PS are not considered reliable; however, in relative terms the PS

results are similar to the measured values for SS such that values for γ-HBCD are about 1 

order of magnitude less than values for α-HBCD. 

There are limited data and conflicting results for vapor pressure. Based a review of the 

current available data, the measured PS of 7.23×10-7 Pa (PL = 3.11×10-5 Pa based on F of 

0.023) using the gas saturation method is considered the most reliable estimate for the 

commercial HBCD mixture (t-HBCD). This value was selected as an input for 
S-8

thermodynamic consistency analyses (see below). Estimates for the diastereoisomers 

were derived by assuming the PL for γ-HBCD is approximately 2.0×10-5 Pa and that the 

relative differences in values estimated by COSMOtherm are applicable such that PL for 

α- and β-HBCD are approximately 2.93×10-4 Pa and 3.86×10-5 Pa, respectively. Due to 

the limited data and uncertainty in estimating vapor pressure, greater uncertainty is 

considered for these values compared to the water solubility estimates. 

Solubility in Octanol 

There are no known reported measurements for the solubility of t-HBCD or the 

diastereoisomers in octanol. Estimates are provided below in the section on the octanolwater partition coefficient. 

Henry’s law constant (Air-water partition coefficient) 

There are no measured estimates for Henry’s law constant H (Pa.m3.mol-1) currently 

available. The dimensionless air-water partition coefficient KAW is H/RT. EPI Suite Ver. 

4.0 estimates (HENRYWIN Ver. 3.2) for Henry’s law constants are 0.174 and 6.52×10-6

Pa.m3.mol-1 using the bond and group methods, respectively (SI-10). The group method 

estimate is considered erroneous. The log KAW estimate corresponding to the bond 

method is -4.15. H can also be estimated from vapor pressure and water solubility as 

PS/SS or PL/SL. In this manner the selected values of 2.41×10-5 Pa and 2.01×10-4 mol.m-3

correspond to an estimate for H of 0.120 Pa.m3.mol-1 (log KAW = -4.32). SPARC Ver. 4.2 

estimates a value for H of 0.0055 Pa.m3.mol-1 (log KAW = -5.66) using the same 

approach, i.e., PL/SL (http://ibmlc2.chem.uga.edu/sparc/). The estimates from EPI Suite 

bond method and the relative solubilities of the selected values for vapor pressure and 

water solubility are consistent, independent estimates. The EPI Suite bond method 

estimate is selected for thermodynamic consistency calculations for t-HBCD. 

Diastereoisomer estimates for KAW have been reported by Goss et al. using the 

COSMOtherm model (SI-13). The log KAW values are -8.84, -9.20, and -8.64 for α-, β-, γHBCD, respectively. These values correspond to H values of 3.6×10-6, 1.6×10-6, 5.7×10-6

Pa.m3.mol-1 for α-, β-, γ-HBCD, respectively. As discussed above and by the authors the 
S-9

absolute estimates are considered to be erroneous, but the relative values are considered 

reasonable. Diastereoisomer estimates of log KAW from selected values for PL/SL are -

4.27, -4.63, and -4.07 for α-, β-, γ-HBCD, respectively. These estimates are dependent on 

the values for PL/SL, which have already been selected for thermodynamic consistency

calculations. Independent estimates for the diastereoisomers were derived by assuming 

log KAW for γ-HBCD is equal to the EPI Suite estimate (-4.15) and that the relative 

differences in values estimated by COSMOtherm are applicable such that log KAW for αand β-HBCD are -4.35 and -4.71, respectively. 

A Henry’s law constant of 0.75 Pa.m3.mol-1 for t-HBCD was reported as the ratio of 

selected values for PS and SS (SI-3). The different values for PS and SS selected in this 

critical review result in an estimate of 0.120 Pa.m3.mol-1. The EPI Suite bond method 

estimate (SI-10) of 0.174 Pa.m3.mol-1 provides an independent and consistent estimate for 

inclusion in the FAV calculations. 

Octanol-water partition coefficient 

A measured log KOW of HBCD of 5.625 at 25±0.05°C using the generator column 

method and following GLP practices has been reported (SI-4). The test sample was a 

composite containing 8.5% β-, 6.0% α- and 79.1% γ-HBCD (total HBCD 93.6%). The 

study was generally carried out according to the guidelines, although the concentration of 

the stock solution was approximately 0.2% instead of 1%. The generator column method 

for mixtures may result in excess, non-dissolved quantities of diastereoisomers in the 

water column (SI-4). 

Diastereoisomer-specific log KOW values were estimated for α- β- and γ-HBCD using 

reversed-phase high-performance liquid chromatography (RP-HPLC) methods (SI-14). 

The estimated log KOW values for α-, β- and γ-HBCD were 5.07±0.09, 5.12±0.09 and 

5.47±0.10, respectively. The calibration compounds were chlorobenzenes and 

polychlorinated biphenyls, which as discussed by the authors, are not ideal because of the 

structural dissimilarities between the organochlorines and HBCD (SI-14). Assuming the 

solubility in octanol is similar for the diastereoisomers, the KOW estimates are relatively 
S-10

consistent with measurements of water solubility such that KOW increases with decreasing 

water solubility. The RP-HPLC derived KOW values for the diastereoisomers are 

comparable but lower than the measured KOW estimate for commercial HBCD using the 

generator column method. 

Table S.2 lists 11 model estimates including EPI Suite Ver. 4.0, SPARC Ver 4.2 and 

COSMOfrag model estimates. The mean value of the model estimates for log KOW is 6.73 

with a standard deviation of 0.76. These model estimates do not differentiate between 

diastereoisomers. Goss et al. (SI-13) reported diastereoisomer log KOW estimates of 5.59, 

5.44 and 5.53, for α-, β- and γ-HBCD, respectively, using the COSMOtherm model. The 

COSMOtherm model estimates are of the same general magnitude as the generator 

column KOW measurement for t-HBCD and the diastereoisomers determined from 

calibration using RP-HPLC. A notable difference is the relative ranking of 

hydrophobicity between the RP-HPLC and COSMOtherm estimates. 

There are uncertainties in the measured and estimated values for KOW for the commercial 

mixture and the diastereoisomers. The solubilities of the diastereoisomers in octanol are 

expected to be similar and differences in KOW are expected to be the result of differences 

in water solubility. SPARC Ver. 4.2 calculates an activity coefficient in octanol γO of 

12.88 (unitless) and a solid state solubility in octanol SOS of 16.21 mol.m-3 (liquid state 

solubility in octanol SOL of 698 mol.m-3) for HBCD. The SOL value was selected as input 

for the thermodynamic consistency calculations for the commercial HBCD mixture and 

the diastereoisomers. 

The measured log KOW value of 5.63 was selected as input for the consistency 

calculations for t-HBCD. For the diastereoisomers the estimates by Goss et al. were 

selected as input for the consistency calculations. Although these estimates were derived 

in silico they are close to measured estimates and considering the potential 

underestimation of HBCD using the generator column method and the lack of ideal 

calibration substances for the RP-HPLC method, the selected values are considered as 

reasonable starting point for the thermodynamic consistency calculations.
S-11

Table S.2. Model estimates for the octanol-water partition coefficient (KOW). 

log KOW Model Reference 

7.74 EPI Suite Ver. 4.0 (SI-10) 

6.45 ALPOGPs (SI-15) 

5.23 AC logP (SI-15) 

6.75 AB/logP (SI-15) 

6.14 miLogP (SI-15) 

6.66 CosmoFrag (SI-15) 

6.69 SPARC Ver. 4.2 http://ibmlc2.chem.uga.edu/sparc/

6.96 ALOGP (SI-15) 

6.30 MLOGP (SI-15) 

8.01 XLOGP2 (SI-15) 

7.10 XLOGP3 (SI-15) 

6.73 Average log KOW 

0.76 Standard Deviation 

Thermodynamic Consistency Calculations for Chemical Properties 

The methods of Schenker et al. (SI-1) were used to obtain thermodynamically consistent 

estimates for the solubility and partitioning properties of t-HBCD and the three 

diastereoisomers. The data selected as input for the calculation of thermodynamically 

consistent FAVs are listed in Table S.3. An effort was made to maximize input 

parameters, but to ensure that data sources were reliable and independent. In this context 

“independent” means that physical-chemical properties used in the harmonization 

calculations were obtained from distinct measurements or estimates. For example, KAW

estimates derived from ratios of PS and SS were not included in the calculations because 

the selected estimates of PS and SS were already used as input. Reliability was considered 

high if multiple data sources provided similar values. Model estimates were considered as 

well as measured data. The harmonization method also requires and provides relative 

variance values for chemical properties. These values provide guidance as to the relative 

uncertainty of the different properties. The relative variance value inputs were selected 

based on professional judgement of the general reliability of the various parameters. 
S-12

The issue of the mutual solubility of water and octanol can be important when seeking to 

obtain thermodynamic consistency in solubility and partitioning parameters (SI-16,17). 

KOW measurement techniques, such as the shake-flask and slow stir methods, result in the 

mixing of water and octanol phases, which results in the presence of some octanol in the 

water phase and of some water in the octanol phase. Measurements of the octanol-air 

partition coefficient (KOA) and of the solubility in octanol, on the other hand, are obtained 

using “dry” octanol, i.e., octanol that has not been saturated with water. Similarly, 

measurements of water solubility and air-water partition coefficients are typically 

performed with pure water. To account for the possible influence of the mutual solubility 

of water and octanol on the phase partitioning of highly hydrophobic substances, 

relationships have been proposed for converting measured KOW values to ratios of 

concentrations in pure octanol and pure water CO/CW. For the thermodynamic 

consistency analysis, the equation recommended by Beyer et al. (SI-17) was used to 

convert KOW to CO/CW prior to adjustment and to convert the adjusted CO/CW back to 

KOW after adjustment. 

Table S.3. Selected values for physical-chemical properties and relative variance values 

used as inputs for thermodynamic consistency calculations. 

Substance PL, Pa SL, mol·m-3 SOL, mol·m-3 Log KAW Log KOW

 a Log KOA 

t-HBCD 2.41×10-5 2.01×10-4 698 -4.15 5.63 10.71 b 

α-HBCD 2.93×10-4 3.27×10-3 698 -4.35 5.59 9.62 b 

β-HBCD 3.86×10-5 9.86×10-4 698 -4.71 5.44 10.50 b 

γ-HBCD 2.00×10-5 1.41×10-4 698 -4.15 5.53 10.79 b 

Relative variance 

t-HBCD 3 2 4 4 3 

α-HBCD 5 1 4 5 4 

β-HBCD 5 1 4 5 4 

γ-HBCD 5 1 4 5 4 

a

 log KOW values corresponding with log (Co/Cw) values of 6.02, 5.97, 5.76, and 5.89 for t-, α-, β-, and γHBCD, respectively. 

b

 values not used as inputs for calculation of FAVs because they are not “independent” values, i.e., they 

were derived from vapor pressure and octanol solubility estimates (see text). The values are included for 

comparisons with FAVs. 
S-13

Table S.4 lists the thermodynamically consistent FAVs for the relevant solubility and 

partitioning properties. The percentages of adjustments were modest for changes from the 

selected input values to the calculated FAVs for most properties suggesting that the data 

are generally consistent and reasonable. Notably, all of the KOW values were slightly 

increased and the solubility in octanol decreased in all cases except for α-HCBD for 

which it increased (largest percentage adjustment). Notably, the adjustment suggests that 

α-HCBD is about 4 times more soluble in octanol than γ-HCBD. This may be partially 

explained by the high symmetry of the α-isomer (SI-3) and may partially contribute to 

increased bioaccumulation potential in the environment for α-HCBD. The relative 

ranking of the FAVs for KOW are consistent with the rankings of the measured water 

solubilities and the independent KOW estimates by Hayward et al. (SI-14) that were not 

used in this procedure. The FAVs for KOA are also comparable to the estimates that were 

not included in the consistency calculations. 

The harmonization method was applied to t-HBCD in this critical review; however, a 

thermodynamic consistency calculation should generally not be used for mixtures, but 

rather for discrete substances with distinct properties. Considering the uncertainty in the 

FAVs for t-HBCD in the context of the greater uncertainties associated with actual 

emissions, mode-of-entry, isomerization and degradation in the environment, the errors in 

calculating “representative” FAVs for t-HBCD are considered to be comparatively low. 
S-14

Table S.4. Final adjusted values (FAVs at 25 oC), relative variance values, percent 

adjustments, and assigned confidence factors used in the present study for mass balance 

modeling and data analyses obtained from the thermodynamic consistency calculations 

(SI-1). 

Substance PL, Pa SL, mol·m-3 SOL, mol·m-3 Log KAW Log KOW

 a Log KOA 

t-HBCD 3.03×10-5 2.33×10-4 380 -4.28 5.77 10.46 

α-HBCD 3.00×10-4 2.77×10-3 1340 -4.36 5.38 9.96 

β-HBCD 4.29×10-5 9.89×10-4 630 -4.76 5.47 10.47 

γ-HBCD 2.42×10-5 1.68×10-4 300 -4.23 5.80 10.40 

Relative variance 

t-HBCD 1.95 1.27 2.13 2.13 1.95 3.45 

Α-HBCD 2.70 0.82 2.20 2.70 2.20 4.50 

Β-HBCD 2.70 0.82 2.20 2.70 2.20 4.50 

Γ-HBCD 2.70 0.82 2.20 2.70 2.20 4.50 

Percent adjusted 

t-HBCD 26% 16% -45% -26% 39% 26% 

α-HBCD 2% -15% 92% -2% -38% 2% 

β-HBCD 11% 0% -9% -10% 7% 11% 

γ-HBCD 21% 19% -57% -18% 87% 21% 

Confidence factor (Cf) 

t-HBCD 2.5 2.3 3 3 3 3 

α-HBCD 4 2 4 4 4 4 

β-HBCD 4 2 4 4 4 4 

γ-HBCD 4 2 4 4 4 4 

a

 log KOW values corresponding with log (Co/Cw) values of 6.22, 5.68, 5.81, 6.25 for t-, α-, β-, and γHBCD, respectively 

Estimating uncertainty 

A confidence factor Cf (or more appropriately an uncertainty factor) can be used to 

express the degree to which parameter X may deviate from the median M in a lognormal 

distribution (SI-18,19). For example, a Cf of 2 suggests that 95% of all of the values in 

the distribution are within 2 and 0.5 times M. A 95% probability can be expressed as: 

 
S-15

probability < X <Cf M = 0.95

Cf

M













⋅ (S.4) 

Thus, the variance in the lognormal distribution increases with an increase in Cf. Values 

of confidence can be derived empirically or subjectively or in combination. 

Section S.2: Additional details on data characterizing the degradation 

of HBCD 

Air 

The atmospheric degradation half-life (HLAIR) can be calculated using the following 

equation: 

[ ]

ln 2

k OH

HL

OH

AIR = (S.5) 

where kOH is the 2nd-order rate constant for reactions for OH radicals (cm3/molecule/s) 

and [OH] is the concentration of OH radicals (molecule/cm3). The kOH of HBCD 

estimated by AOPWIN v1.92 is 6.12x10-12 cm3

/molecule/s. According to the model 

documentation, 90% of the AOPWIN estimations are within a factor of 2 of empirical 

data and 95% are within a factor of 3. Based on this model performance, it is assumed 

that the uncertainty in the kOH estimation for HBCD is a factor of 3 (in either direction). 

Estimated [OH] concentrations are available in the literature (SI-10,20-22), typically 

expressed as either 12 h (daylight) averages or 24 h averages. Different [OH] are 

presented in Table S.5 along with the corresponding HLAIR. 

Table S.5. Range in estimated HLAIR for HBCD based on estimated [OH]. 

[OH] 

mol./cm3 

Region Time Period 

(h) 

HLAIR 

(d) 

Min 

(d) 

Max 

(d) 

1.5x106 Global 12 1.7 0.6 5.2 

9.2x105 Global 24 1.4 0.5 4.3 

9.8x105 N. 

America 

24 1.3 0.4 4.0 
S-16

8.5x105 S. America 24 1.5 0.5 4.6 

1.1x106 Global 24 1.2 0.4 3.6 

9.7x105

 Global 24 1.4 0.5 4.1 

As shown in Table S.5, the assumed uncertainty in kOH results in ranges in HLAIR that 

span the threshold criterion for persistence in air (2 d) rather than the variability in [OH]. 

Water 

As discussed in the main text, the available empirical data on biodegradation in the water 

column is considered unreliable. Biodegradation potential can be assessed using 

BIOWINTM (SI-10), which provides quantitative output which is then used to bin 

compounds into different qualitative categories (e.g., does NOT biodegrade fast, Time 

frame: Months, Time frame: Weeks). Model output for HBCD is presented in Table S.6. 

Note that no distinction can be made between isomers or enantiomers. 

Table S.6. BIOWINTM v4.10 model output for HBCD (not isomer specific) as presented 

in Arnot et al. (SI-2). 

Model Output Interpretation 

BIOWIN1 (linear) 0.1650 Does NOT biodegrade fast 

BIOWIN2 (non-linear ) 0.0000 Does NOT biodegrade fast 

BIOWIN3 (Ultimate deg) 1.9548 Time frame : Months 

BIOWIN4 (Primary deg) 3.1342 Time frame: Weeks 

BIOWIN5 (MITI linear) -0.4234 NOT Readily Degradable 

BIOWIN6 (MITI non-linear) 0.0000 NOT Readily Degradable 

BIOWIN7 (anaerobic linear) 2.5072 Biodegrades FAST 

The model output (BIOWIN7 vs. BIOWIN1) is in qualitative agreement with 

observations that biodegradation proceeds quicker in anaerobic environments compared 

to aerobic. However, the model output cannot be used directly to compare with the 

persistence criterion for water (60 d). Aronson et al. (SI-23) related model output for 

BIOWIN3 to experimental half-lives in an attempt to provide some guidance on 
S-17

extrapolating from qualitative BIOWIN output to quantitative half-live estimates. 

BIOWIN3 output for HBCD suggests that ultimate degradation occurs over a time frame 

of months. The default water half-life associated with this category, based on 

experimental primary degradation data for 44 compounds, was suggested to be 120 d. 

However, the empirical half-life data for the compounds in this category ranged from 1–

1420 d (median = 85 d, 10% trimmed mean = 126 d). 

Arnot et al. (SI-24) used BIOWINTM model output and empirical half-life data to derive 

regression equations which relate numerical model output to half-life estimated directly. 

These regressions are presented in Table S.7 along with the corresponding estimated halflife values. 

Table S.7. General equations (SI-24) used to estimate aerobic environmental 

biodegradation half-lives from BIOWINTM model output and corresponding values 

calculated for HBCD. 

Model Regression Equation r2 Half-life (d) 

BIOWIN1 (linear) log10(t1/2) = -1.32x + 2.24 0.72 105.2 

BIOWIN3 (Ultimate deg) log10(t1/2) = -1.07x + 4.20 0.77 128.3 

BIOWIN4 (Primary deg) log10(t1/2) = -1.46x + 6.51 0.78 85.9 

where x is the numerical output generated for each respective BIOWINTM model (see previous Table above). 

Hydrolysis (abiotic) 

The aqueous hydrolysis rate program [HYDROWIN, (v1.67), EPI Suite (v4.0)] was 

unable to estimate a hydrolysis rate constant for HBCD likely due to the lack of 

hydrolysable functional groups. The reaction of HBCD with water, therefore, can be 

assumed to be negligible. 

Sediments and Soils 

Davis et al. reported on the biodegradation of t-HBCD in freshwater sediments and soils 

(SI-25). Using OECD test guidelines 307 and 308, the authors demonstrated that the rate 
S-18

of loss of HBCD at 20oC was appreciably faster under anoxic conditions in both media. 

Relative to biologically sterile controls, biotransformation of HBCD was faster in the 

presence of microorganisms and t1/2’s values ranged from 11 to 32 days (aerobic) and 1.1 

to 1.5 days (anaerobic) in sediment. In soil, t1/2’s under aerobic and anaerobic conditions 

were 63 and 6.9 days, respectively. Initial t-HBCD concentrations in these experiments 

were in the range of 10–100 ug/kg dry weight. Because of the poor analytical sensitivity, 

the authors were unable to characterize degradation products. 

In a subsequent study, Davis et al. showed that degradation of HBCD proceeded by 

sequential loss of Br2 (SI-26). Estimated t1/2’s for the α-, γ- and β-isomers were 128, 92 

and 72 daysrespectively in freshwater sediments under aerobic conditions (SI-26). Initial 

concentrations of α-, γ- and β-HBCD in these experiments were reported to be in the 

range of 4–5 mg/kg dry weight. While there were no statistically significantly 

differences between the measured values, these measurements were greater than what 

was previously reported on by the same authors (SI-25). The authors note that the latter 

study was conducted at HBCD concentrations (mg/kg) much greater than the earlier 

study (μg/kg) and at high test concentrations, degradation kinetics may be limited by the 

mass transfer of chemical into the microbes (SI-26). 

Additional data on HBCD degradation kinetics in aerobic soil was cited in the European 

Commission Risk Assessment Report (SI-4), taken from an unpublished study conducted 

by the Dow Chemical Company (see Davis et al. 2004 in ref 4). The initial concentration 

of t-HBCD (14C radiolabelled) was reported to be approximately 3 mg/kg dry weight. 

Results were presented as % of initial nominal radioactivity; t-HBCD fell from 

97.6±0.2% at Day 0 to 88.2±1.4% at Day 112 for t-HBCD and α-HBCD fell from 

8.6±0.5% to 8.2±0.2%. Metabolites were not quantified. These data were considered to 

be indicative of negligible degradation and no degradation half-lives were estimated. 

There have been two studies on the degradation of HBCD in wastewater sludge (SI26,27). Both studies suggest that the α-isomer is the most recalcitrant. The rank order of 

the individual biotransformation rates as reported by Gerecke et al. (SI-27) were β ≈ γ > 
S-19

α-isomer and γ <β < α-isomer as reported by Davis et al. (SI-25). Because of the 

pronounced difference in the test concentration of t-HBCD used (ca. 90-fold) and based 

on the earlier argument, it is not surprising that there were differences in the estimated 

t1/2’s of t-HBCD between the studies: of 0.66 (SI-27) and 15 days (SI-26). 

Photodegradation. 

Harrad et al. (SI-28) studied the photo-induced degradation of HBCD under two different 

exposure conditions: (i) individual standard solutions (in methanol) and (ii) isomer 

fortified indoor dust samples (SI-28). Using artificial lighting, photo-induced 

isomerization of γ- to the α-isomer was demonstrated in both scenarios. Under the first 

scenario, the authors also showed that each isomer can isomerize to produce different 

proportions of the other 2 diastereoisomers and that conversion of γ- to the α-isomer was 

the most predominant (SI-28). There was also a measured decrease in concentration of 

ΣHBCD concomitant with an increase in concentration of the degradation product, 

pentabromo-cyclododecane (PBCDs), in the simulation study with indoor dust. No mass 

balance or measurement of any other degradation product was reported. As one would 

expect, the photolytic degradation of the isomers on indoor dust was not enantiomerselective, i.e., EF = 0.5 throughout the duration of the study. 

Mammalian Biotransformation Pathways 

Brandsma et al. (SI-29) reported on the in vivo metabolism of t-HBCD in Wistar rats and 

on the detection of HBCD degradates in biota. The authors identified debromination and 

hydroxylation as the major degradation pathways and were able to identify, PBCDs, 

tetrabromocyclododecane (TBCD) and the corresponding mono- and di-OH forms of 

HBCD, PBCD and TBCD in select tissues of exposed rats. The mono-OH-HBCD was 

also detected in tern eggs and mono-OH-HBCD and –PBCD was detectable in blubber of 

harbour seal (SI-29). 

Potential degradation products and environmental relevance
S-20

The temporal evolution of concentrations of α-, β-, and γ-hexabromocyclododecanes 

(HBCDs), and pentabromocyclododecenes (PBCDs — degradation products of HBCDs) 

was studied in separate aliquots of a well-homogenized indoor dust (SI-28). These were: 

(a) exposed to natural light, and (b) kept in the dark. Results revealed a slow degradative 

loss of HBCDs via elimination of HBr. Tetrabromocyclododecadienes (TBCDs) were 

detected in all sediments from nine English lakes, while pentabromocyclododecenes 

(PBCDs) present in some (SI-30). This suggests HBCD can degrade via sequential loss of 

HBr. 

Section S.3: Additional details on data and approaches used to 

characterize Overall Persistence (POV) 

Table S.8. Degradation half-lives selected for HBCD benchmarking (SI-2). 

Media Default HL 

 (d) 

Cf Lower Bound 

(d) 

Upper Bound 

(d) 

Air 1.3 3 0.4 4.0 

Water 

(fresh, marine) 

85 10 8.5 850 

Sediments 

(fresh, marine) 

35 6 6.0 210 

Soil 85 10 8.5 850 

Physical-Chemical Properties for Benchmark Chemicals 

The chemicals selected for the POV benchmarking exercise representing POPs were as 

follows: HCB, PCB28, PCB101, PCB180, Aldrin and Heptachlor. Non-POPs include 

biphenyl, p-cresol, atrazine and tetrachloromethane and candidate POPs (Candi-POPs) 

include α-, β-, γ-HCH, and PBDE-99. These substances represent a range of physicalchemical properties. Most of these substances have been used in previous POPs 

benchmarking assessments (SI-31). Aldrin and heptachlor were selected because these 

POPs have shorter half-lives in air than most POPs and these half-life estimates are 
S-21

comparable to those for HBCD. Table S.9 lists the partitioning properties and Table S.10 

lists the degradation half-lives (HLs). 
Table S.9. Selected physical-chemical properties for the model benchmarking simulations. Data are from (31) 

heptachlor which are from (SI-1,32). 

Status Name Abbrev MW (g.mol-1) log KOW log POP hexachlorobenzene (HCB) HCB 284.8 5.61 -1POP 2,4,4'-trichlorobiphenyl (PCB 28) PCB28 257.6 5.66 -1POP 2,2',4,5,5'-pentachlorobiphenyl (PCB 101) PCB101 326.4 6.38 -2POP 2,2',3,4,4',5,5'-hexachlorobiphenyl (PCB 180) PCB180 395.3 7.15 -2POP Aldrin Ald 364.9 6.25 -2POP Heptachlor Hept 373.4 5.96 -1POP alpha-hexachlorocyclohexane α-HCH 290.8 3.88 -3POP beta-hexachlorocyclohexane β-HCH 290.8 3.91 -4POP gamma-hexachlorocyclohexane γ-HCH 290.8 3.76 -3POP 2,2',4,4',5-pentabromodiphenyl ether PBDE99 564.7 6.76 -3Non-POP Biphenyl Biphenyl 154.2 4.06 -1Non-POP p-cresol p-cresol 108.1 1.97 -4

Non-POP Atrazine Atrazine 215.7 2.73 -6Non-POP tetrachloromethane CCl4 153.8 2.83 0.
Table S.10. Selected primary transformation half-lives (HL; days) for the model benchmarking simulations. except for aldrin and heptachlor in which environmental HLs were selected based on (10,24,32) and fish HLAvian and mammalian HLs (Av/Mam) for aldrin and heptachlor were assumed to be 1/3 estimates for fish. Tra(HLN) in biota are listed normalized to 1 kg body size and scaled to mass (Mi; kg) specific values (HLi) for d

models as HLi = HLN x (Mi/1)0.25 (SI-33). 

Status Abbrev HL_Air HL_Water HL_Soil HL_Sediment HL_Fish (1 kPOP HCB 650 1000 1500 9000 860 POP PCB28 6 300 450 2700 190 POP PCB101 20 1200 1800 10800 370 POP PCB180 80 2200 3300 19800 1100 POP aldrin 0.2 1000 1000 2000 75 POP heptachlor 1.5 25 50 150 90 POP α-HCH 50 440 660 3960 55 

POP β-HCH 50 900 1350 8100 90 

POP γ-HCH 50 400 600 3600 120 

POP PBDE99 20 850 1275 7650 285 Non-POP biphenyl 3 15 22.5 135 5.2 Non-POP p-cresol 0.3 5 7.5 45 0.35 Non-POP atrazine 0.4 50 75 450 0.3 Non-POP CCl4 900 150 225 1350 0.2 
S-24

Model Selection 

The POV benchmarking exercise was reported in Arnot et al. (SI-2) using the Risk 

Assessment Identification And Ranking (RAIDAR) model v2.0. Figure S.1 illustrates 

POV estimates for HBCD against POP and non-POP chemicals (Tables S.9 and S.10) 

using The OECD Tool model (SI-34). POV estimates for HBCD range from 12 d 

(assuming lower bound environmental half-lives in Table S.8) to 1220 d (assuming upper 

bound environmental half-lives in Table S.8). Assuming the median environmental halflives in Table S.8, the POV for HBCD 123 d. In general, POV predictions for HBCD are 

more similar to POV values for the non-POP benchmark chemicals than to POV values for 

the POP benchmark chemicals. 

Figure S.1. Overall persistence (POV; d) and characteristic travel distance (CTD; km) 

calculated by The OECD Tool for HBCD and benchmark POP and non-POP chemicals 

(Tables S.9 and S.10). The results are based on “worst case” Tool simulations. 

10

100

1000

10000

100000

1000000

1 10 100 1000 10000

The Tool P ov (d)

CTD, km

POPs

non-POPs

t-HBCD (median HLs)

t-HBCD (lower bound HLs)

t-HBCD (upper bound HLs)
S-25

Section S.4: Additional details on data characterizing the 

bioaccumulation potential of HBCD

Data used to derive empirical BAFs based on (SI-35) and (SI-30) are presented in Table 

S.12 and S.13 below. 

Table S.12. Empirical BAFs (L/kg) for α-HBCD calculated using data from (SI-35). 

CDissolved 11 pg/L 

Species Lipid Content 

(-) 

CLIPID 

(ng/g lipid) 

CWET WEIGHT 

(pg/kg) 

log BAF 

(L/kg) 

Walleye 0.178 6.53 1.16x106 5.02 

Whitefish 0.120 1.22 1.46x105 4.12 

Mussels 0.095 8.25 7.84x105 4.85 

Zooplankton 0.097 8.64 8.38x105 4.88 

Emerald shiner 0.160 8.70 1.39x106 5.10 

Goldeye 0.160 8.58 1.37x106 5.10 

White sucker 0.152 3.66 5.65x105 4.70 

Table S.13. Empirical BAFs (L/kg) for α-, β- and γ-HBCD calculated using data from 

(SI-30). 

BAF L/g lipid* log BAF L/kg** 

Isomer Mean Min Max Mean Min Max 

Α-HBCD 5900 1200 23000 5.47 4.78 6.06 

Β-HBCD 1300 250 3500 4.81 4.10 5.24 

Γ-HBCD 810 110 3200 4.61 3.74 5.20 

* in same units as originally published; muscle tissue 

** recalculated following Harrad et al. (SI-36); assumes fish is 5% lipid overall and 

equilibrium between all lipids in body 
S-26

Section S.5: Modeled bioavailability, uptake efficiencies and inherent 

bioaccumulation factors (BAFs) 

Isomer-specific physical-chemical property estimates are currently available for α-, βand γ-HBCD, allowing the model calculations shown below to be undertaken; no 

distinctions are made between enantiomers of each isomer i.e., they are assumed to have 

identical properties. 

 

Bioavailability 

Reported water solubilities for α-, β- and γ-HBCD are approximately 50, 15 and 2 μg/L, 

respectively. The higher water solubility of the α isomer in comparison to the other two 

isomers translates into higher bioavailability in aquatic ecosystems only when the water 

solubility limits are exceeded (i.e., > 15 μg/L). At environmentally-relevant 

concentrations (i.e., pg/L to ng/L range), the bioavailable fraction in the water column is 

a function of partitioning between the dissolved and solid phases in the water column, not 

the water solubility limit. For neutral organic compounds, partitioning is largely 

controlled by the hydrophobicity of the compound, typically represented by the octanolwater partition coefficient (KOW). This parameter can readily be used to estimate 

partitioning to environmental solids assuming that absorption to organic carbon (OC) is 

the dominant mode of interaction. Once the extent of sorption to organic carbon is 

estimated (i.e. organic carbon partitioning coefficient, KOC), the freely dissolved i.e. 

bioavailable fraction of a compound in the water column (FDissolved) can be estimated 

using the equations shown below: 

 

Dissolved POC DOC

Dissolved

Dissolved M M M

M

F

+ +

= (S.6) 

where MDissolved, MPOC and MDOC are the mass of the compound in the dissolved phase, 

associated with particulate organic carbon (POC) and dissolved organic carbon (DOC) 

respectively. Dividing all terms by MDissolved (or CDissolved VDissolved) yields the following 

equation relating FDisssolved to partition coefficients (KPOC, KDOC) and volumes of each 

phase in the water column (VDissolved, VPOC, VDOC): 
S-27

Dissolved

DOC

DOC

Dissolved

POC

POC

Dissolved

V

V

K

V

V

K

F

+ +

=

1

1

 (7) 

FDissolved is calculated in the following illustrative example based on the parameterization 

for Lake Ontario compiled for modeling purposes by Arnot & Gobas (SI-37). 

Concentrations of particulate and dissolved organic carbon in water column: 

CPOC = 2.5x10-7 kg/L 

CDOC = 2.0x10-6 kg/L 

Assuming the density of POC and DOC = 1 kg/L, the volumes of POC, DOC and water 

(dissolved) per liter of bulk water are: 

VPOC = 2.5x10-7 kg/L 

VDOC = 2.0x10-6 kg/L 

VDIS = 1 - VPOC - VDOC 

While more sophisticated approaches for estimating organic carbon-water partition 

coefficients are available (e.g. polyparameter linear free energy relationships), the 

approach adopted here is to use KOW, which is reasonable as a first-approximation for a 

neutral apolar compound such as HBCD. KPOC and KDOC are calculated as in Arnot & 

Gobas (SI-37) as: 

KPOC = ρPOC KOW (S.8) 

KDOC = ρDOC KOW (S.9) 

where ρPOC and ρDOC are proportionality constants (L/kg OC) which relate to sorptive 

capacity of organic carbon to that of octanol. The assumed values for these constants are 

shown below. 

The selected proportionality constants are ρPOC = 0.35 L/kg and ρDOC = 0.08 L/kg. 

Calculated FDissolved for each isomer are presented in Table S.14 using the FAVs (see 

Section S.1) and COSMOTherm values (SI-13). The reason for including both sets of 
S-28

estimates is that they differ not only in magnitude but also in relative values among the 

isomers. 

Table S.14. Estimated dissolved fraction of HBCD isomers in the water column using 

parameters representing Lake Ontario 

Source Log KOW FDissolved 

Arnot et al. FAVs (SI-33) 

α-HBCD 

β-HBCD 

γ-HBCD 

5.38 

5.47 

5.80 

94.4% 

93.2% 

86.5% 

COSMOTherm 

α-HBCD 

β-HBCD 

γ-HBCD 

5.59 

5.44 

5.53 

91.2% 

93.6% 

92.3% 

Based on the FAVs, the freely dissolved fraction of α-HBCD is only marginally higher 

than the freely dissolved fraction of γ-HBCD despite having a 25-fold higher water 

solubility. Using the COSMOTherm property estimates, the freely dissolved fraction of 

all isomers is essentially identical. Based on these calculations (and those for sediment 

porewater, not shown here), differences in the bioavailable fraction in the environment 

between the isomers are not likely to explain the extent of enrichment of α-HBCD in 

biota. 

Uptake efficiencies across the gills and gastrointestinal tract 

Empirical relationships relating the uptake efficiency for transport across the gills and 

gastrointestinal tract (GIT) to KOW are available in the literature for aquatic organisms 

(SI-37). The approach is based on two-film resistance theory and similar relationships 

have also been derived for humans and other air-breathing organisms (SI-38-40). The 

relationships (SI-37) are shown below for uptake efficiency across the gill (EW) and GIT 

(ED): 

1

85.1( (155 / ))− EW = + K OW (12) 

7 1

0.3( 10 )0.2

− − ED = × ⋅ K OW + (13) 
S-29

Values calculated for HBCD isomers are presented in Table S.15. 

Table S.15. Estimated uptake efficiencies across the gills and GIT in aquatic organisms 

for HBCD isomers based on estimated log KOW.

Source Log KOW EW ED 

Arnot et al. FAVs (33) 

α-HBCD 

β-HBCD 

γ-HBCD 

5.38 

5.47 

5.80 

54.0% 

54.0% 

54.0% 

48.3% 

47.9% 

45.7% 

COSMOTherm 

α-HBCD 

β-HBCD 

γ-HBCD 

5.59 

5.44 

5.53 

54.0% 

54.0% 

54.0% 

47.2% 

48.0% 

47.6% 

As shown in Table S.15, uptake efficiencies across the gills and GIT are very similar for 

all HBCD isomers and therefore also cannot explain the extent to which α-HBCD is 

preferentially accumulated in biota. 

Modeled Bioaccumulation factors (BAF) assuming perfect persistence

By assuming no biotransformation (i.e. whole body metabolic rate constant = 0), the 

aquatic food chain bioaccumulation model described in Arnot & Gobas (SI-37) can be 

used to evaluate the inherent bioaccumulation potential of each isomer i.e. based on 

physical-chemical properties alone. The bioaccumulation model calculations are useful 

because they integrate the considerations described above and also trophodynamics. To 

be consistent with calculations made above, the Lake Ontario parameter set was used and 

bioaccumulation factors (BAFs) were calculated assuming the FAVs described above and 

the COSMOTherm values (SI-13). In order to obtain an environmentally-relevant isomer 

profile in both media as well as ratio between concentrations in overlying water versus 

sediments, water column and sediment concentrations used as inputs to the model were 

based on measurements in English lakes (SI-30). As noted above, other system 

parameters (e.g. CPOC, CDOC) are based on Lake Ontario as are the organisms included 

and feeding relationships. Hence, direct comparisons to empirical concentrations and 

BAFs in biota reported in (Harrad et al. (SI-41)) are not appropriate. The selected input 

concentrations are shown below in Table S.16. 
S-30

Table S.16: Assumed input water (total) and sediment concentrations based on (SI-30). 

Isomer CWATER

(pg/L) 

% CSEDIMENT

(ng/g dw) 

% 

α-HBCD 8.1 17 0.16 13 

β-HBCD 6.2 13 0.13 11 

γ-HBCD 32 69 0.93 76 

Modeled log BAFs using the FAV property values are shown in Table S.17a whereas 

model output using the COSMOTherm property values are shown in Table S.17b. 

Table S.17a. Modeled log BAFs using the Arnot et al. FAV physical-chemical property 

estimates and water (total) and sediment concentrations from (SI-30). 

Organism α-HBCD β-HBCD γ-HBCD Ratio α : γ

Phytoplankton 3.79 3.87 4.15 0.44 

Zooplankton 3.84 3.92 4.24 0.40 

Diporeia 3.93 4.00 4.29 0.44 

Mysid 4.29 4.36 4.64 0.44 

Sculpin 4.08 4.19 4.61 0.29 

Alewife 4.26 4.37 4.79 0.29 

Smelt 4.04 4.16 4.61 0.27 

Lake trout 5.09 5.24 5.77 0.21 

Table S.17b. Modeled log BAFs using the COSMOTherm physical-chemical property 

estimates and water (total) and sediment concentrations from (SI-30) 

Organism α-HBCD β-HBCD γ-HBCD Ratio α : γ

Phytoplankton 3.97 3.84 3.92 1.13 

Zooplankton 4.04 3.89 3.98 1.14 

Diporeia 4.08 3.98 4.10 0.94 

Mysid 4.44 4.34 4.46 0.95 

Sculpin 4.33 4.15 4.29 1.10 

Alewife 4.52 4.33 4.46 1.15 

Smelt 4.31 4.12 4.27 1.09 

Lake trout 5.43 5.19 5.35 1.19 

As shown in Table S.17a, bioaccumulation of the γ-HBCD is favoured by a factor of 

approximately 2–5 in comparison to α-HBCD based on the FAVs. The largest difference 
S-31

is found for the highest trophic level consumer (Lake trout), reflecting the fact that 

preferential accumulation (based solely on log KOW) is accentuated along each step of the 

food chain. However, since bioaccumulation of γ-HBCD is favoured over α-HBCD, 

these additional calculations clearly demonstrate that differences in physical-chemical 

properties/bioavailability do not explain the extent to which α-HBCD is preferentially 

accumulated in biota. The model results obtained using the COSMOTherm property 

values show the same directionality of preferential uptake (i.e. α-HBCD over γ-HBCD) 

but the magnitude of the increased accumulation (maximum ~ 20%) is insufficient to 

explain field observations. Similar conclusions are obtained using terrestrial 

bioaccumulation models (SI-38). The main finding based on these theoretical calculations 

is that biotransformation and/or bioisomerization favouring the accumulation of α-HBCD 

are the most plausible driving factors behind what is observed in the field as opposed to 

the physical-chemical properties. 

Section S.6: Biotransformation and bioisomerization half-life estimation 

Fish Biotransformation 

Law et al. (SI-42) conducted a dietary uptake (feeding) and depuration study. Fish (~0.25 

kg) were fed diastereomer specific diets and first-order “depuration” rate constants of 

0.0044 and 0.0048 d-1 with respective half-lives of 157 d (±71 SE) and 144 d (±60 SE) 

were reported for β- and γ-HBCD. Estimates for α-HBCD could not be determined 

because “elimination” did not follow first-order kinetics. The study shows the potential 

for bioisomerization of the different isomers whereby the isomers are apparently 

converted from one isomeric form to another. Some of the HBCD is not being eliminated 

or depurated from the organism but merely transformed into another diastereomer; 

therefore these “elimination” rates do not necessarily account for the loss of HBCD from 

the organism via transformation to a metabolite. 

Table S.18 summarizes the molar elimination and formation of the diastereoisomers 

during the “elimination” phase based on the data reported in Table 2 of Law et al. (see 

also Figure 1 from that study) (SI-42). The data suggest that γ-HBCD is predominantly 
S-32

converted to α-HBCD (70%) and only about 30% of the total elimination half-life is 

actually attributable to “elimination of HBCD” from the organism through the formation 

of a metabolite. Data also suggest that β-HBCD is predominantly converted to α-HBCD; 

however, the molar change during the depuration period actually suggests a net “gain” of 

total HBCD. It is not possible for the fish to create HBCD; therefore, based on the 

general uncertainty of the reported measurements it can be assumed that there is no 

apparent “loss of HBCD” from the organism for β-HBCD. The data also suggest that αHBCD is predominantly eliminated with minimal conversion to the other 

diastereoisomers. Based on the data for γ-HBCD (high conversion to α-HBCD), some of 

the γ-HBCD isomer being formed from α-HBCD in the α-HBCD exposure study is 

expected to be converted back to α-HBCD during the experiment. It would seem that γHBCD is converted to α-HBCD faster than α-HBCD is converted to γ-HBCD, which 

would support the overall observation for the attenuation of γ-HBCD isomers in food 

webs. Furthermore, if relative bioisomerization rates approximate the relative rates of 

thermal isomerization, γ-HBCD would be converted to α-HBCD at rates approximately 

10 times faster than α-HBCD is converted to γ-HBCD (SI-43). 

Table S.18. Molar mass balance for diastereoisomer bioisomerization and actual total 

elimination (fecal egestion, gill respiration and biotransformation). 

Elimination 

phase 

Other isomers formed 

HBCD (+) or lost (-) 

isomer 

fed to 

fish 

Day 

56 

Day 

168 

Apparent 

loss of 

“parent” 

isomer γ β α

Net loss of 

“parent” 

isomer 

from 

organism 

% actual 

elimination 

of “parent” 

isomer 

α-HBCD 1.4 0.5 0.9 -0.23 -0.1 N/A 0.9 ~100 

β-HBCD 0.4 0.3 0.1 0 N/A +0.2 -0.1 ~0 

γ-HBCD 0.4 0.17 0.23 N/A -0.01 +0.16 0.07 ~30.4 

Table S.18 suggests that α-HBCD is biotransformed primarily to a metabolite, whereas γHBCD is predominantly biotransformed to α-HBCD. β-HBCD does not appear to be 

biotransformed to a metabolite. Our interpretation of the data is somewhat different from 

the interpretation of Law et al. (SI-42) who suggested that α-HBCD is recalcitrant to 

bioisomerization. Non-first order elimination kinetics were reported for α-HBCD. Indeed 
S-33

non-first order kinetics may be the result of the relatively faster rates of conversion of γHBCD back to α-HBCD. A “pseudo-elimination rate constant” based on the molar 

quantity at the start of the elimination (Day 56) and the molar quantity at the end of the 

elimination phase (Day 168) is estimated as 0.009 d-1 for a 0.25 kg fish. From the net loss 

it appears that most of α-HBCD is actually biotransformed to a metabolite and eliminated 

from the fish, corresponding to an elimination half-life of about 75 d for α-HBCD. 

Table S.19 summarizes the parameters required to estimate biotransformation and 

bioisomerization rates using a mass balance estimation method for fish when applied to 

the Law et al. dataset for rainbow trout (SI-44). There are three possible 

biotransformation rate constant (kM) estimates that could be considered. The first is “loss 

of parent” diastereomer. This includes bioisomerization from one diastereomer to a 

different diastereomer (and subsequent bioisomerization reactions) and the formation of 

metabolites (elimination). This would be calculated based on the “total depuration” rate 

constant (kT) reported by Law et al. (SI-42) The second is biotransformation to a 

metabolite that is no longer HBCD. This value is based on the reported total depuration 

rate constant (or estimated depuration rate constant for α-HBCD) and a correction for the 

actual loss of parent diastereomer (Table S.18), i.e., 30% of the reported total depuration 

rate constant for γ-HBCD. The third transformation rate constant is the difference 

between the two estimated rate constants which is assumed to be the result of 

bioisomerization to another isomeric form of HBCD but without elimination from the 

fish, i.e., 70% of the reported total depuration rate constant for γ-HBCD. It must be 

recognized that the measured data are uncertain and that possible interconversion back 

and forth between the different isomeric forms at competing, and presumably different 

rates, further confounds the interpretation of the data and the ability to accurately 

estimate bioisomerization and biotransformation (elimination) rate constants. Since the 

bioaccumulation models used in the present study do not currently account for 

bioisomerization only the second rate constant (loss of HBCD from the organism) is 

considered relevant for exposure modelling at this time. 
S-34

The estimated total elimination rate is so slow for β-HBCD, that no estimate for 

biotransformation is possible using the mass balance method. The estimated 

biotransformation rate constants for rainbow trout (~0.25 kg) were 1.7×10-3 and 3.0×10-4

d

-1 for α-, and γ-HBCD, respectively. The estimation method provides screening level 

information regarding confidence in the estimated kM values (SI-44,45). Although an 

estimate for γ-HBCD is calculated, the percentage of positive values is less than 25% 

corresponding to a low confidence category assignment for the estimated kM value (SI44,45). Although an estimate for α-HBCD is calculated, the output confidence factor is 

greater than 10.3 corresponding to a “low” confidence category assignment for the 

estimated kM value (SI-45). Using a rearrangement of Equation 2.5, the corresponding 

primary biotransformation half-lives normalized HLN to a 1 kg fish are 3330 and 720 d 

for α-, and γ-HBCD, respectively. Thus, the analysis from the measured data from Law et 

al. (SI-42) suggest “very slow” (SI-2) and low confidence primary biotransformation halflives for α-, and γ-HBCD. 
S-35

Table S.19. Mass balance model summary for estimating biotransformation and 

bioisomerization rate constants from a dietary feeding study in rainbow trout (SI-42). 

Diastereomer 

Parameter α-HBCD β-HBCD γ-HBCD 

Value Cƒ Value Cƒ Value Cƒ

W (kg) 0.25 1.5 0.25 1.5 0.25 1.5 

LF (kg/kg) 0.045 1.5 0.045 1.5 0.045 1.5 

T (°C) 11.5 1.05 11.5 1.05 11.5 1.05 

GD (kg/d) 0.005W 2 0.005W 2 0.005W 2 

COX (mg/L) 9.9 1.25 9.9 1.25 9.9 1.25 

ED N/R a N/A 0.41 2 0.41 2 

KOW 2.40x105 3 2.95x105 3 6.31x105 3 

kT (d-1) b 9.0x10-3 4.4 c 4.41x10-4 4.6 d 1.44x10-3 4.2 d

kM (d-1) 1.7x10-3 13 No estimate possible 3.0x10-4 7 

Cf: confidence factor; kT: total elimination rate constant; W: whole body weight of the fish; LF: whole body 

lipid fraction of the fish; T: water temperature; GD: feeding rate; COX: dissolved oxygen concentration; ED: 

chemical transfer efficiency between the gastrointestinal tract and the organism; KOW: octanol-water 

partition coefficient; kM: primary biotransformation half-life (loss of NA: not applicable; NR: not reported. 

a

 Model default calculation 

b

 Reported or calculated elimination rate constants for “net loss of parent isomer”. For α-HBCD, the rate 

constant described in the text is assumed first-order, for β-HBCD and γ-HBCD, the reported rate constants 

are corrected for the % of actual elimination of the parent isomer as reported in Table 2.6. Since the value 

for β -HBCD is apparently “no net loss” a value of 10% of the reported value was assumed 

c

 Assumed 

d

 Calculated from the coefficient of variation from the empirical elimination half-life estimates following (

SI-44,45) 

There are complicating factors with these data other than bioisomerization. The Law et al. 

(SI-42) data are sampled from muscle tissue and not whole body elimination rate constant 

estimates. The fish were being dosed via the food and HBCD would be subject to firstpass at the liver, which could be the predominant site for bioisomerization as well as 

biotransformation of the HBCD isomers to metabolites. Whole body distribution 

(toxicokinetics) is expected to be slow and changes in the liver are expected to be faster 

than changes in the muscle tissue. The preferential solubility in storage lipids (solubility 

in octanol) for α-HBCD may be another complicating factor in the analysis. Furthermore, 

the study recognized that steady state was not reached. 
S-36

Drottar and Krueger (SI-46) conducted a flow through bioconcentration test for t-HBCD 

at two different exposure concentrations for rainbow trout (0.035 kg). The measured 

average concentrations were 0.18 and 1.8 μg·L-1, for the low and high levels, 

respectively. The whole body bioconcentration factors (BCFs) were 13,000 (not at steady 

state) and 9,000 L·kg-1, respectively. This study reports total elimination rate constants of 

0.0228 and 0.0359 d-1 for t-HBCD for the low and high levels, respectively. It is noted 

that the sponsor of the study did not want the first two sample times during the depuration 

phase (days 1 and 3) analyzed and thus these time points were not included in the 

elimination rate constant estimates. The implications are unclear. The total uptake and 

elimination rate constants were calculated by BIOFAC and were used as inputs into the 

mass balance calculations for estimating kM. The mass balance calculations for primary 

biotransformation of t-HBCD to metabolites provide kM estimates of 0.0027 d-1 (Cf = 7.4) 

and 0.012 d-1 (Cf = 6.7) for the low and high levels, respectively. According to the 

screening level information regarding confidence in the estimated kM values, both 

estimates are considered “high” (SI-45). The study documents some experimental 

difficulties maintaining a constant water concentration for the low level exposure test. 

These difficulties may result in errors in the BIOFAC rate constant estimates which are 

derived by fitting a model to the chemical concentration time trends. The 

biotransformation rate constants from both exposures correspond to primary 

biotransformation half-lives of 260 and 56 d for the mass of the fish in the study (0.035 

kg). Using a rearrangement of Equation 2.5, the corresponding primary biotransformation 

half-lives normalized HLN to a 1 kg fish are 1080 and 235 d for t-HBCD for the low and 

high levels, respectively. These biotransformation rates are considered “very slow” and 

“slow”, respectively (SI-2). 

Nyholm et al. (SI-47) evaluated the dietary exposure and biotransformation of various 

brominated flame retardants, including HBCD, in zebrafish. There was low recovery for 

HBCD in this study and only relative concentrations could be determined. Thermal 

degradation during the analysis is considered to be a plausible explanation for the low 

recovery. A “high bioaccumulation potential” was suggested by the authors, although 

bioaccumulation parameters such as elimination rate constants were not reported. 
S-37

A QSAR for predicting screening level primary biotransformation rates in fish estimates 

a kM of 0.0071 d-1 for a 1 kg fish (SI-2). This rate constant corresponds to a HLN of about 

100 d. 

In consideration of all of the factors for the biotransformation data available for fish a 

HLN of 250 d is selected for mass balance modelling with a Cf of 7. Thus, 95% range of 

the generated estimates for HLN will be between 35 and 1750 d. 

Mammalian Biotransformation 

Geyer et al. reported total elimination half-life estimates for rats and humans for t-HBCD 

(SI-48). The data for humans were derived from body burdens and daily intake rates. It is 

unclear how the data for rats were determined. The reported total elimination half-lives 

are 64 d (reported range of 23 to 219) and 8 d for humans and rats, respectively. The total 

elimination half-life includes elimination processes such as respiration, fecal egestion, 

urinary excretion and biotransformation. The mass balance method for fish was refined 

for application to mammalian half-life data assuming 70 kg for the human and 0.25 kg for 

the rat. The estimated HLN (normalized to 1 kg) are 25 d (Cf = 3.4) and 13 d (Cf = 3.2) 

from the reported human and rat total elimination half-life data. A HLN value of 20 d (Cf

= 5) was selected for the mass balance exposure modelling for t-HBCD. The 95% range 

of the generated estimates for HLN will be between 4 and 100 d. This HLN value is about 

1 order of magnitude faster than the HLN value selected for fish. 

In vitro biotransformation studies have been conducted for laboratory rats and a harbour 

seal (sampled shortly after death) using microsomal liver preparations (SI-49). Significant 

reductions in LC-MS chromatogram peaks in the in vitro assays for rats (incubations of 

1:1:1 isomeric mixture of α-, β- and γ-HBCD) showed that β- and γ-HBCD 

diastereromers were biotransformed; however, the peak of the α-HBCD did not decrease 

“significantly”. The incubations of a 1:1:1 mixture of the HBCD isomers with the 

harbour seal microsomes showed average decreases of the parent isomers by 69, 60, and 

17% for the β-, γ-, and α-HBCD isomers, respectively. β- and γ-HBCD were also tested 

separately and new peaks of brominated compounds (indicative of metabolites) were only 
S-38

formed when NADPH was added confirming a cytochrome P-450 mediated 

biotransformation (SI-49). Unfortunately, it does not appear that α-HBCD was analyzed 

for the possibility of forming peaks in the same isolated manner. Therefore the possible 

biotransformation of α-HBCD by cytochrome P-450 cannot be confirmed or refuted. 

Although α-HBCD was reduced to a lesser degree some formation of α-HBCD from γHBCD (bioisomerization) may have been possible during the simultaneous incubations. 

Invertebrate Biotransformation 

For earthworms, estimated biota-soil accumulation factors (wet weight-soil/wet weightworm) ranged between 0.03 and 0.08 based on the total concentration of HBCD in worm 

tissue and estimates for the wet weight concentration in soil after 28 day of exposures to a 

range of soil concentrations (SI-4). The diastereomer specific biota-soil accumulation 

factor for α-HBCD is more than one order of magnitude higher than the value for γHBCD (SI-4). The study suggests a degree of bioisomerization of γ-HBCD to α-HBCD or 

preferential biotransformation of γ-HBCD. It is recognized that this is not a valid BCF or 

BSAF study. 

Biotransformation Rate Summary 

The selected half-life values are based on a critical review of the limited available data 

and thus require a degree of expert judgement. It is noteworthy that the primary 

transformation half-lives in water, soil, and sediment are shorter than the half-lives in 

fish. While this may be possible, it is counterintuitive, especially with regards to the 

degree in which the values differ (e.g. aerobic sediment – 35 d; 1 kg fish – 250 d). The 

data for fish are not definitive, but are supported by bioaccumulation measurements in the 

environment showing biomagnification in food webs (low biotransformation rates). This 

suggests that the environmental half-lives may be underestimates. Therefore, some of the 

mass balance model simulations will explore the possibility of half-lives in water, soil 

and sediment that are greater than median values. 

In summary, there is conflicting evidence for diastereomer-specific biotransformation and 

bioisomerization rates and pathways based on in vitro, laboratory and monitoring data. 
S-39

These data are for different species under different conditions and natural variability is 

expected among and between individuals and species. It is recognized that α-HBCD is 

usually the predominant diastereoisomer in monitoring data particularly for upper trophic 

level species (e.g., marine mammals, birds), but the precise reason for this cannot be 

elucidated from the data. It may be a result of preferential biotransformation to 

metabolites of the β- and γ-HBCD isomers and a limited capacity for biotransformation 

for α-HBCD in the food webs. It could also be that bioisomerization rates for γ-HBCD to 

α-HBCD are relatively faster than biotransformation rates for α-HBCD resulting in the 

preferential accumulation of α-HBCD. Different relative solubilities for the diastereomers 

in storage lipids may also be a factor. The uncertainty associated with bioisomerization 

and biotransformation limits the ability to accurately model isomer-specific HBCD 

bioaccumulation and subsequent exposures to organisms in the environment. 

Section S.7: Additional details on data and approaches to characterize 

Long-range Transport (LRT) Potential

A LRT potential benchmarking exercise used the same compounds and property values 

as were used to conduct the POV benchmarking exercise (see Section S.3). Arnot et al. 

(2009) (SI-2) present model output based on Characteristic Travel Distance (CTD), 

calculated using The OECD Tool (SI-34) and Arctic Contamination Potential (eACP10), 

calculated using GloboPOP (SI-12). CTD is the distance at which the modeled air 

concentration reaches 1/e (i.e. 37%) of the original source level. CTD results are 

summarized for HBCD and POP and non-POP chemicals in Figure S.1. eACP10 is the 

fraction of total emissions released that is present in the Arctic region after 10 years of 

constant release. Results for eACP10 are summarized in Table S.20. In both exercises, 

there is overlap between POPs and non-POPs, confounding the classification of HBCD. 
S-40

Table S.20. eACP10 model output for benchmarking exercise for HBCD (SI-2). 

Status Chemical Scenario eACP10 

POP HCB 

PCB28 

PCB101 

PCB180 

Aldrin 

Heptachlor 

Default HLs 

Default HLs 

Default HLs 

Default HLs 

Default HLs 

Default HLs 

1.61% 

0.17% 

0.53% 

0.65% 

0.04% 

0.004% 

Non-POP Biphenyl 

p-cresol 

Atrazine 

CCl4 

Default HLs 

Default HLs 

Default HLs 

Default HLs 

0.004% 

0.004% 

0.016% 

0.18% 

Newly Listed α-HCH 

β-HCH 

γ-HCH 

PBDE99 

Default HLs 

Default HLs 

Default HLs 

Default HLs 

0.55% 

0.53% 

0.50% 

0.37% 

Under Review t-HBCD 

t-HBCD 

t-HBCD 

Default HLs 

Lower Bound HLs 

Upper Bound HLs 

0.05% 

0.005% 

0.40% 

Influence of CTD and emission strength on presence in remote regions

The importance of considering emission strength in addition to LRT P when considering 

monitoring data from remote sites is illustrated below. In this example, five hypothetical 

chemicals with CTD ranging from 100 – 2000 km are included. The modeled air 

concentrations at the remote site (CR,O) assuming an initial air concentration (CO) of 1 

pg/m3 and a distance of 2000 km from source region to remote site for each of the five 

chemicals are shown in Table S.21. The relative emission strength at the source region 

required to obtain an identical air concentration in the remote region (CR,E = 0.37 pg/m3) 

is also shown. For chemicals with low CTD (100 km), the relative emission strength is 

required to be many orders of magnitude higher than the chemical with high CTD (2000 

km). However, the relative emission strength for chemicals with CTDs between 500–

2000 km is only required to range over approximately 1 order of magnitude for identical 

air concentrations to be obtained at the remote site. 
S-41

Table S.21. CTD versus emission strength as determinant of air concentration in remote 

regions 

Distance to Remote Region 2000 km CO = 1 pg/m3 

CTD 

(km) 

CR,O in remote 

region 

Relative Emission 

Strength at Source 

CR,E 

(pg/m3) 

2000 0.37 1.0 0.37 

1500 0.26 1.4 0.37 

1000 0.14 2.7 0.37 

500 0.018 20.1 0.37 

100 2.1x10-9 1.8x108 0.37 

The purpose of this simplified analysis is to show that detection in remote regions does 

not necessarily mean that the compound of interest has a particularly high LRT P. 

Emission strength, mode of entry and spatial distribution of emissions are also important 

factors to consider, especially when comparing monitoring data for a suite of chemicals. 
S-42

Section S.8: Summary of monitoring data from near-point source, 

industrialized regions and remote sites 

Table S.22. Summary of monitoring data (ΣHBCD) from locations near known point 

sources, industrialized regions (no known direct sources) and remote sites (i.e. Arctic) as 

reported in Arnot et al. (SI-2). 

Compartment/Biota Units Near Point 

Source 

(range) 

Industrialized 

Regions 

(range) 

Remote Sites 

(range) 

Air pg m-3 10 – >1000 0.1 – 600 1 – 6.1 

Water μg l-1 0.006 – 140 

Soil μg kg-1 dw 265 – 815 

Sediment μg kg-1 dw 1 – 1000 <0.1 – 30 <DL – 4.3 

Aquatic invertebrates ng g-1 lw <DL – 1.9 

Benthic invertebrates ng g-1 lw 3100 – 17335 0.5 – 100* 1.4 

Fish 

(lower trophic level) 

ng g-1 lw 5 – 25* 

Fish 

(upper trophic level) 

ng g-1 lw 1000 – 50000 10 – 300 0.4 – 2 

Birds ng g-1 lw 25 – 150 20 – 175 

Marine mammals ng g-1 lw <5 – 19210** 30 – 750 <0.1 – 45 

Terrestrial mammals ng g-1 lw 

Humans ng g-1 lw <DL – 190 

* includes food basket survey data; ** largely reflects reported range in harbour porpoises inhabiting

coastal waters of the UK (Law et al. 2006, 2008) – median ΣHBCD values ranged from 30 – 1860 ng g-1 lw 

over the period 1994 – 2006; dw = dry weight, lw = lipid-normalized, DL = reported detection limit 

Monitoring data published more recently are broadly consistent with the approximate 

ranges presented in Table S.22. For example, Vorkamp et al. (SI-50) reported α-HBCD 

levels between <1.35 and 18.1 ng g-1 lw in ringed seals from east Greenland (Remote 

Site) sampled over the period 1986 – 2008. Reported concentrations of ΣHBCDs in 
S-43

water, sediment and fish from lakes in England (Industrialized Region) ranged between 

0.00008 – 0.00027 μg l-1, 0.9 – 4.8 ug kg-1 dw, and 14 – 290 ng g-1 lw, respectively (SI30). The main gap with respect to abiotic monitoring data is the paucity of measurements 

in soils (and sediments) from background and remote regions. Data on HBCD levels in 

soils may be particularly informative because this compartment is often the largest 

reservoir for hydrophobic contaminants; the long degradation half-lives reported for 

HBCD in (aerobic) soils also implies that this compartment may be important in the longterm as a storage medium and potential secondary source. Additional data for terrestrial 

mammals (wildlife) would also be valuable. 

With respect to monitoring data in the Arctic, there appears to be a distinct relationship 

between proximity to main production/use regions (e.g. Europe = high production/use, N. 

America = low production/use) and levels observed (SI-2,50). For example, reported 

total HBCD concentrations in walrus, beluga and narwhal blubber inhabiting Arctic 

regions closer to North America (e.g. Davis Strait) sampled in 1998 were 0.6, 1.4 and 3.4 

ng·g-1 lw respectively (SI-51). These values are lower than those reported in mammals 

from Svalbard where the mean HBCD (total) concentrations in blubber of ringed seal 

sample in 2003 was 20 ng g-1 lw and the mean total HBCD concentrations in polar bears 

sampled in 2002–2003 ranged from approximately 12 to 45 ng g-1 lw. Mean total HBCD 

concentrations in polar bears inhabiting Greenland were similar to Svalbard (SI-52) but 

reported concentrations in polar bears from the Bering Strait-Chukchi Sea area were 

substantially lower with a mean reported value of 0.4 ng g-1 lw (range <0.1 to 35.1 ng g-1

lw). If these concentrations are representative, it implies that i) long-range transport of 

chemical emitted from Europe has a more limited impact on the western Arctic and ii) 

historical HBCD inputs to the Western Arctic, likely emanating from North American 

and Asia-Pacific sources, are substantially lower than inputs impacting Svalbard. 

Hypothesis ii) is consistent with available information on the distribution of global 

HBCD production and use. Further detail is provided in Arnot et al. (SI-2) and Vorkamp 

et al. (SI-50). 

 
S-44

Section S.9: Analytical Chemistry 

Sample preparation. 

A general trend to apply specific sample preparation methods, including extraction, 

clean-up and fractionation has been observed in the last years (SI-53). Abdallah et al. (SI54) has included an in-cell purification using Florisil and Hydromatrix during the PLE. 

Even using this approach, the extracts had to be further purified with concentrated H2SO4

followed by solid phase extraction (SPE) on Florisil. PLE-based methods have been used 

also by Eljarrat et al. (SI-55) for the extraction of HBCD isomers from lyophilized human 

milk and by Fredriksen et al. (SI-56) for marine biological samples. Recently, ultrasonicassisted extraction (UAE) has also been used for the extraction of HBCD isomers from 

sewage sludge using a mixture of DCM-acetonitrile (1:1) (SI-57). Microwave assisted 

extraction (MAE) with acetone/n-hexane (1/3, v/v) has been employed for the extraction 

of HBCD isomers from marine sediments (SI-58). 

HBCD has been extracted also from various polymers or from consumer products based 

on these polymers. The extraction efficiency of PLE, MAE and UAE were compared for 

the recovery of HBCD from styrenic polymeric plastics (SI-59). PLE and MAE resulted 

in complete extraction of HBCD (recovery >95%). UAE, finally, offered relatively low 

extraction recoveries (10–50%). In another study, HBCD was extracted together with 

other BFRs in a short step by ultrasonic-supported dissolution and precipitation (USDP) 

from styrenic polymers. To analyse HBCD in flame-retarded textiles (e.g. curtains), 

three different extraction methods (Soxhlet, UAE, and soaking extractions with toluene 

and DCM) were compared (SI-60). During Soxhlet extraction using toluene, the percent 

contribution of α-HBCD to total HBCDs increased slightly and that of γ-HBCD 

decreased, indicating that γ-HBCD was isomerised to some extent at the boiling point of 

toluene. For UAE, the temperature of the water bath can easily increase over time during 

the extraction, which might also lead to isomerisation. Soaking extraction with DCM was 

chosen as the most facile procedure to extract HBCD diastereomers from textiles (SI-60). 

A cleanup based on dispersive SPE with primary–secondary amine was used recently for 

sewage sludge extracts (SI-57). Yet, it has been shown that such cleanup is not sufficient 
S-45

for complete matrix removal as matrix effects showed high ion suppression as high as 

50% for all 3 HBCD diastereoisomers. The retention behaviour of individual HBCD 

isomers on silica gel and Florisil was investigated using diverse mobile phase solvents 

and accounting for matrix effects. The β-HBCD diastereomer is substantially retained on 

both Florisil and silica regardless of the solvent used and therefore it undergoes selective 

loss during clean-up (SI-61). This sequence is counterintuitive to sequences based on 

reverse-phase chromatography with a C18-column, in which the α- (and not the β-) isomer 

elutes first when using a polar solvent. These results indicate that care should be taken 

when isolating HBCDs and other molecular diastereomers from environmental and 

biological samples, and that reported concentrations of β-HBCD in the literature may be 

negatively biased. 

Analysis 

Traditionally, HBCD has been determined via GC−ECNI/MS. However, GC techniques 

have a number of serious limitations, such as interconversion of the HBCD isomers 

above 160oC, degradation in dirty injection systems or above 240°C, and weaker 

structural identification in ECNI/MS. In contrast, α-, β- and γ-HBCD isomers can be 

separated easily using reversed-phase liquid chromatography and determined by mass 

spectrometry (LC/MS-MS). Two additional isomers (δ-HBCD and ε-HBCD) have been 

isolated from technical products with δ-HBCD identified recently in fish. Furthermore, 

HBCD stereoisomers can be resolved on an enantioselective LC column (SI-62).

However, LC-MS analysis is prone to ion suppression resulting in decreased sensitivity. 

This problem is avoided by thorough sample clean-up to remove interfering components 

and by using 13C-labelled and 2H-labelled HBCDs as internal standards to compensate for 

potential variations in sensitivity during and between sample runs. 

A few recent studies have investigated the comparability of results issued using GC or 

LC-separation methods. However, the results were not conclusive due to a large variation 

in the methodologies employed in each studies and the type of samples, fish (SI-63) and 

dust (SI-54) for which the comparison was achieved. Although it was concluded that LCESI-MS-MS results in reliable data and most important individual isomeric 
S-46

concentrations, it was acknowledge that GC-ECNI-MS can give a good estimation of the 

total HBCDs, if appropriate internal standards and calculation methods are used. 

An important gap is the lack of harmonized methods for the determination of HBCDs in 

environmental matrices. Regular intercalibration studies and participation in laboratory 

proficiency studies (when available) are important so as to maintain high quality of 

analytical data. Up to now, only a limited number of intercalibration studies have been 

performed with a rather limited number of participating laboratories (SI-64). The 

laboratories were able to determine total HBCD concentrations in the marine samples 

with satisfying quality (RSD <35%). However, the analysis of samples with (very) low 

HBCD contamination (< 2 ng/g lipid weight) should be improved. There are currently no 

certified reference materials certified for HBCDs that can be used for method validation. 

Up to now, indicative values for HBCDs have been issued for several reference materials, 

but the certification of HBCDs in a wider range of environmentally-relevant materials is 

needed.

As discussed previously, LC tandem MS methods are now generally the preferred choice 

for analysis of HBCD. These methods now find routine application in many laboratories; 

some have been further developed to include more comprehensive suites of analytes, 

including other BFRs, such as tetrabromobisphenol A (TBBPA) (SI-56,65). 

Some studies have described the use of ion trap MS for determination of HBCD (SI66,67); however, Gómara et al. (SI-67) state that the typically monitored MRM transition

for HBCD (m/z 640.6 → m/z 78.9 and m/z 80.9) corresponding to [M-H]+

 to Br- is below 

the conventional ion trap low mass cut-off. Interestingly, the authors overcame this 

difficulty through forcing formation of a chlorine adduct of HBCD, and subsequent 

detection of a stable quasi-molecular ion as a result of fragmentation. The formation of 

chlorine adducts of HBCD, corresponding to ions at m/z 676.6, have been reported (SI66-69). Adduct formation should be considered as a major factor in quantitation of 

HBCD; some of the previous studies employed an ammonium acetate mobile phase 

modifier to reduce chlorine adduct formation, which can result in formation of an acetate 
S-47

adduct (SI-69). Galindo-Iranzo et al. (SI-69) employed two different ammonium salt 

mobile phase modifiers to encourage (ammonium chloride) or inhibit (ammonium 

acetate) chlorine adduct formation; they found chloride adduct formation to result in 

method LODs more than an order of magnitude lower than inhibition. 

Biological processes, such as metabolism, could result in a shift in the enantiomeric 

composition of HBCD (SI-70). The three HBCD diastereomers (α-, β- and γ-HBCD) 

most prevalent in the technical mixture (~95%) are each represented by a corresponding 

pair of enantiomers. These enantiomers have been isolated using enantioselective HPLC 

and their structures confirmed by X-ray diffraction (SI-71). The elution order of the 

HBCD enantiomers as determined by Heeb et al. (2007) is (-) α-HBCD, (-) β-HBCD, (+) 

α-HBCD, (+) β-HBCD, (+) γ-HBCD and (-) γ-HBCD. 

It is commonly known that enantiomers, although simply mirror images of each other, 

can have vastly different biological properties/activities and degradation rates. Therefore, 

to fully understand the environmental relevance of HBCD in the environment and its 

potential health effects, enantioselective analyses are required. Over the past several 

years, a number of studies have described enantioselective methods for analysis of 

HBCD, and subsequently applied these methods to environmental samples. These 

methods have largely been based on LC/MS/MS techniques using the mass spectrometer 

in electrospray ionization (ESI) negative ion mode (SI-51,70). Janak et al. (SI-70) 

produced one of the first studies involving enantioner-specific determination of HBCD; 

the authors reported enantioselective accumulation of both α- and γ-HBCD in a variety of 

marine species from the Western Scheldt Estuary (the Netherlands), including an 

apparent enrichment of the (+) α-HBCD enantiomer in whiting and bib liver samples. 

Extraction and clean up methods for determination of HBCD enatiomers do not deviate 

from those for the parent diastereomers; however, a chiral stationary phase is required for 

separation prior to mass spectral determination. Most methods use a Nucleodex β-PM 

(Machery-Nagel, Düren, Germany) chiral stationary phase for enantiospecific separation 

of the HBCD diastereomers. This stationary phase is a permethylated β-cyclodextrin. 
S-48

Guerra et al. (SI-72) also investigated the applicability of the Chiradex and Chirasep 

stationary phases for separation of HBCD enantiomers, but only the Nucleodex β-PM 

phase was found to be satisfactory. Application of state-of-the-art LC/MS/MS technology 

to the determination of HBCD enantiomers results in very low instrument detection limits 

(IDLs) and overall method detection limits (MDLs). For example, the method of Guerra 

et al. (SI-72) demonstrates IDLs and MDLs in the low- to sub-pg and low- to sub-ng/g, 

respectively. 

Section S10: Additional Toxic Endpoints 

Carcinogenicity 

Despite an early report to the contrary (SI-73), most studies report negative results for 

HBCD’s ability to induce mutagenesis or genetic recombination leading to 

carcinogenicity. This has been the case for results obtained in yeast (SI-74), bacteria (SI75), human peripheral blood lymphocytes (SI-75) and in rats treated with high doses of 

HBCD. Carcinogenesic compounds may also induce their effects via epigenetic means 

(not due to genetic mutations or recombinations). For example, altered DNA methylation 

has been observed in many types of cancers and may be among the earliest cellular 

changes observed during oncogenesis. However, no impacts of HBCD on global or gene 

specific DNA methylation have been observed in fish (SI-75) or mammalian models (SI76). 

Additional Thyroid-Disruption 

Schriks et al. (SI-77) also examined the ability of HBCD to interact with α and β thyroid 

hormone receptors using specific reporter gene assays. Specifically, the ability of HBCD 

to affect T3 mediated activation of the two receptors in an in vitro test system was 

examined. Green monkey kidney fibroblasts (CV-1) transfected with the two thyroid 

receptors linked to luciferase reporter genes were utilized for these studies. While they 

had seen thyroid potentiation activity for HBCD with a T-screen assay previously (SI-78), 

they reported no effect of HBCD on T3 receptor activation in this study. The different 

results from this study may be explained by the fact that CV-1 cells do not have 
S-49

endogenous thyroid receptors, and therefore may also lack other sites that were targeted 

by HBCD compared with the cells that were used for their previous studies. 

Conversion of the less active form of thyroid hormone, thyroxine (T4), to the active form, 

triiodothyronine (T3) is accomplished by deiodinase enzymes. It has been suggested that 

this too may be impacted by HBCD. Lower activity of T4 deiodination was evident in 

livers of fish exposed to HBCD in rainbow trout. It may be that elevated levels of T3 that 

were concurrently observed in these fish was responsible for down regulating the 

deiodinase activity (SI-79). However, subsequent work by the same authors showed that 

rainbow trout exposed to HBCD had elevated deiodinase enzyme activity (SI-80). 

Finally, Crump et al. (SI-81) found no altered expression in any of the three deiodinase 

mRNAs that they examined. Future research is needed to determine the potential for 

HBCD to affect specific deiodinase activities that may influence thyroid hormone 

turnovers rates. 

Immunological Effects of HBCD 

Van der Ven et al. (SI-82) also reported that some immunological parameters were 

affected by HBCD. For example, in males white blood cell count in the bone marrow was 

increased, there was a decline in lymphocyte fraction and an increase of the polymorph 

neutrophil population in circulation, lower thymus weight, increased popliteal lymph 

nodes, and enlargement of the marginal zone in the spleen. In females thymus weight 

declined. To test function of the immune system selected F1 rats were challenged with 

the NK activity assay. This challenge involves inoculation with a thymus dependent 

antigen (sheep red blood cells) and identification of specific antibodies. The response was 

shown to be potentiated by HBCD exposure with a BMDL of 0.46 MKD. Van der Ven et 

al. (SI-82) speculate that these effects may also be linked to thyroidal disruption. Recent 

experimental evidence has documented lower circulating thyroid hormones and 

reproductive disturbance in an HBCD exposed avian model. Captive adult American 

kestrels (Falco sparverius) were exposed to environmentally relevant concentrations of 

HBCD delivered via the diet throughout pair bnding, mating and eggs laying, by injecting 

the compound into the brain of day old cockerels (SI-83). Egg laying was accelerated, 
S-50

and clutch size increased among HBCD fed birds, but these eggs were significantly 

smaller than eggs derived from kestrels fed a reference diet. These effects were 

associated with in ovo HBCD concentrations of 1000-9600 ng/g. Elevated testosterone 

and testis size and reduced reproductive behaviour were identified in parental males. 

As with whole animal reproductive studies, in vitro examinations of HBCDs effect on 

gonadal function have returned conflicting results. Expression of estrogen responsive 

genes (ERα, ERβ and THRα) were downregulated in HepG2 human hepatocarcinoma

cells incubated with HBCD at 0.3 ng/ml for 72 hours (SI-76). Hamers et al. (SI-84) 

investigated the ability of the technical mixture of HBCD, as well as the individual 

diastereoisomers, to affect several receptors. In stably transfected CALUX® cell lines, 

antiandrogenic and anitprogestagenic activity was observed in the low μM range with α

and γ having greater effects than β and TM-HBCD. Estrogenic antagonism was also 

shown for β and γ diastereoisomers but not for α or the TM-HBCD (SI-84). 

Gregoraszczuk et al. (SI-85) demonstrated altered capacity of porcine ovarian follicular 

cells to produce reproductive steroids (progesterone, testosterone and estradiol) when 

exposed to a fish liver extract containing high concentrations of HBCD. However, the 

fish oil extracts also contained relatively high concentrations of PCBs, DDT and 

metabolites, hexachlorocyclobenzene (HCH) and toxaphenes, making specific cause and 

effect relationships for HBCD difficult to draw. Cantón and Van den Berg (86) reported 

that HBCD diastereoisomers affected activity of the enzyme aromatase which catalyzes 

the conversion of testosterone to estradiol, in an in vitro test system. However, Kuiper et 

al. (87) found that aromatase was not affected in either testes or ovaries of European 

flounder exposed to dietary and sediment borne HBCD for 78 days. Taken together, these 

results indicate that while there is some evidence for HBCD’s potential to impact 

reproduction and its molecular control, more mechanistic work is required. 

There is strong recent evidence that HBCD affects lipid metabolism. Canton et al. (SI-88) 

examined gene expression in rats from the 30 and 100 mg/kg dose groups from Germer et 

al. (SI-89). HBCD most strongly affected the peroxisome proliferator-activated receptor 

(PPAR)-mediated regulation of lipid metabolism, triacylglycerol metabolism, cholesterol 
S-51

biosynthesis, androstenedione, testosterone, estrogen and glutathione metabolism. 

Selected genes from these affected pathways were further examined showing that 

cholesterol biosynthesis and lipid metabolism genes were down-regulated, especially in 

female rats. Canton et al. (SI-88) provide a hypothesis for the mechanisms by which 

HBCD affects cholesterol and lipid synthesis in hypothyroid female rats identified in the 

van der Ven et al. (SI-90) study. They suggest that thyroxine depletion down-regulates 

cholesterol synthesis by decreasing synthesis of the HMG-CoA enzyme which catalyzes 

the conversion of HMG-CoA to mevalonate, a precursor of cholesterol. Furthermore, 

they note that a simultaneous increase in serum cholesterol may be related to lower 

catabolism of circulating low-density lipoproteins when thyroxine concentrations are low. 

Crump et al. (SI-81) also provided evidence that HBCD may affect lipid metabolism in an 

avian animal model. Expression ofTHRSP14-a decreased in a dose dependent manner 

with exposure to HBCD. While the function of THSRP14-a is not fully known, it is 

thought to play a role in T4 mediated lipogenesis. Crump et al. (SI-81) also found that 

liver fatty acid binding protein (L-FABP), which in involved in metabolism and 

intracellular transport of lipids, was downregulated by HBCD. This gene may be 

responsive to activation by T3, and therefore, reduced expression may be secondary to 

thyroid hormone imbalance following HBCD exposure. 

HBCD may also affect the function of natural killer (NK) cells, specialized lymphocytes 

which are capable of removing tumour and antibody coated cells and virally infected 

cells. Using short term exposures (1hr – 6d) of NK cells derived from human blood, 

Hinkson and Whalen (2009) showed that HBCD could affect ATP energy reserves of the 

NK cells as well as their lytic ability toward tumour cells (SI-91). Effects were evident 

even with short term exposure (1 hr), but the lowest observed effect level was 1 μM for a 

decrease in tumour-lysing function after 6 d of exposure to HBCD. 

Reproductive and Hormone Mediated Effects 

Effects secondarily related to HBCD’s thyroidal impacts are increasingly being 

examined. Reproduction, lipid metabolism, nervous system function and development, 
S-52

and even the process of adapting to a new osmotic environment in salmon smolts (SI-92), 

have been studied. Several detailed studies have examined reproduction in rats treated 

with HBCD. Early studies found few discernable effects of HBCD on reproduction in 

orally exposed rats (SI-93,94). More recent reproductive trials have extended on this 

work by including multigenerational effects and by extending the oral dosing of HBCD 

beyond the period of gestation (SI-82,95,96), but again, few effects consistent with 

HBCD toxicity have been reported. Ema et al. (SI-96) fed rats HBCD enriched diets for 

10 weeks prior to mating, continuing through mating, gestation and lactation periods for 2 

successive generations. No effects of HBCD on reproductive parameters were reported. 

Primordial follicle numbers were reduced in F1 females at 1500 ppm and higher. 

Primordial follicles represent the total female reproductive potential, and although they 

did not see any other reproductive effects in these females, the study was only based on 

one litter from each dam so that the lifelong capacity of the affected females was not 

assessed. No tissue concentrations of HBCD were reported, making relationships 

between reproductive toxicity and tissue burdens impossible. 

Developmental toxicity was assessed in rats exposed to HBCD via the diet after maternal 

exposure from mid-gestation through lactation to day 20 after delivery (SI-95). A 

subgroup of the offspring were necropsied after the HBCD treatment but others were 

maintained until 11 wks when they had reached puberty to examine reproductive 

parameters. HBCD exposure did not affect maternal weight, food consumption, 

implantations, or duration of the pregnancy. The number of live offspring, their sex ratio 

and body weights and anogenital distance were also not affected by HBCD. However 

maternal hypothyroidism was associated with impaired oligodendroglial cell 

development in offspring of dams, but only at higher exposures. 

Van der Ven et al. (SI-82) also exposed rats to dietary HBCD exposure for 10 weeks prior 

to mating in males and for 2 weeks prior to mating in females. Exposure was continued 

through gestation and lactation, and in their F1 offspring from weaning until final 

necropsy. HBCD accumulated to a greater degree in females relative to males across all 

dose groups (an average of 4-6 fold higher in the higher dose groups) in the F1 generation 
S-53

but there were no effects on mating success, gestation or mortality of pups. HBCD also 

decreased bodyweight of F1 in a dose dependent manner starting at post natal day 4 and 

continuing until the end of the study (11 wks of age). The authors note that this may be 

related to decreased food intake of the HBCD exposed pups. Other effects affected at low 

exposures included a decrease in the weight of the testis in the F1 generation, decreased 

trabecular bone mineral density and a decline in circulating apolar retinoids. The authors 

note that testis is responsive to androgen antagonists and cite a recent study (SI-84) 

showing that HBCD can interact with the progesterone and androgen receptors. Finally, 

Van der Ven et al. (SI-82) also reported that thyroidal disruption may be linked to 

immune system activation (see SI). 

 

Nervous System Effects 

There is growing evidence that HBCD may affect the nervous system, with secondary 

effects on learning and motor activity. Reistad et al. (SI-97) demonstrated the ability of 

HBCD to accumulate in brain tissue. Using an in vitro test system these same authors 

showed that cerebellar granule cell (CGC) death increased dose dependently. Cell death 

was inhibited by the n-methyl D-aspartate (NMDA) receptor antagonist MK801. 

Activation of the NMDA receptor may trigger several downstream pathways leading to 

calcium influx and apoptosis, ROS formation, and cellular damage. Inhibited toxicity by 

MK801 suggests that HBCD induced toxicity through one or more of these effects. The 

involvement of ROS was further implicated because 50μM tocopherol in the incubation 

medium inhibited HBCD-induced cell death but calcium free medium and caspase 

inhibitor had no effect. Nuclear observations suggested that HBCD induced apoptosis in 

24% of cells but HBCD did not induce DNA laddering or caspase induction. 

It seems likely that the toxicity of HBCD to nervous cells demonstrated by Reistad et al. 

(97) may be manifested in compromised membrane function and neurotransmitter 

release. Dingemans et al. (SI-98) found that HBCD affected basal and depolarizationmediated cytosolic calcium levels in an in vitro assay using neurendocrine PC 12 (Rat 

pheochromocytoma) cells. HBCD applied immediately before depolarization had no 

effect on intracellular Ca2+ making direct interaction with voltage gated calcium 
S-54

channels (VGCC) unlikely. Involvement of specific VGCC were investigated by 

employing blockers (nifedipine, GVIA, or MVIIC x-conotoxins) of the different types of 

VGCC (L-, N-, or P/Q-type, respectively). If the effect of HBCD were specific to a given 

type of VGCC, then application of HBCD with the inhibitor of that type of channel 

should result in no greater effect that the inhibitor itself. However, HBCD elicited a 

greater Ca2+ influx even in the presence of each of the inhibitors, indicating that the 

effect of HBCD was not VGCC specific. HBCD had no effect on vesicular catecholamine 

release, investigated amperometrically using carbon fiber electrodes, in the same cells. 

The potential effects of technical-HBCD on rat brain synaptosome uptake of the 

neurotransmitters dopamine, glutamate and γ-amino n-butyric acid (GABA) was 

investigated in vitro (SI-99). Pre-exposure of isolated synaptosomes to HBCD for noncompetitively inhibited the subsequent uptake of dopamine. Glutamate uptake was 

inhibited at just 1 μM, but inhibition never rose above 50%, even at 50 μM. No effect on 

GABA uptake was elicited with HBCD exposure. To examine whether the mechanism of 

inhibited neurotransmitter uptake was related to an effect on membrane potential, uptake 

of the marker Tetra[3H]phenylphosphonium bromide (TPP+) was performed after preexposure to HBCD. HBCD only slightly reduced the uptake of TPP, which correlated 

well with its effect on glutamate but not dopamine, suggesting that HBCD elicited its 

effect only partially by altering membrane potential. 

Neurotoxic effects of HBCD have also been determined at the tissue and organism level. 

In their study of the reproductive effects of HBCD, Saegusa et al. (SI-95) also examined 

neuronal migration in the brain as a measure of altered brain development. HBCD had no 

effect on the distribution of neurons located within, or lateral to, the pyramidal cell layer 

of the hippocampal CA1 region. However, there was a significant reduction in the 

number of CNPase-positive oligodendrocytes in the cingulate deep cortex among rats. 

This is a measure of mature oligodendrocytes, and the authors noted that this result is 

likely a result of the hypothyroidism reported in animals from this group. 
S-55

Lilienthal et al. (SI-100) examined potential disturbances by technical-HBCD on the 

dopaminergic system and attempted to link disturbances with altered auditory function. 

Rats used in this study were adults of the F1 generation from the study previously 

described by van der Ven et al. (SI-82). The effect of HBCD exposure on haloperidol 

induced catalepsy was evaluated first. Haloperidol inhibits uptake of dopamine by 

synapses and results in prolonged latencies of movement. To examine HBCD effect on 

this process, rats were injected with haloperidol and placed in 3 postures; front paws on a 

bar, clinging to an incline or on a box with their legs in holes. Latencies to first 

movements from these unnatural positions were measured. HBCD decreased the latencies 

of all three cataleptic movements, and the effect was more pronounced in females than in 

males. In males, only foreleg retraction from the box resulted in a significant effect. 

Lilienthal et al. (SI-100) suggested that effects were a result of HBCD inhibiting 

vescicular and membrane transport of dopamine, allowing dopamine to remain active 

longer. Alternatively, HBCD could increase clearance rates of haloperdidol thereby 

affecting catalepsis endpoints. Haloperidol is metabolized by CYP1A1, CYP3A4, and 

CYP2D6 and the authors cite Germer et al. (SI-89) as having reported induction of 

several CYP enzymes in rats exposed to HBCD. It should be noted that the effects of 

HBCD in the study by Lilienthal et al. (SI-100) were more significant in females and 

HBCD can induce metabolic enzymes activity to a greater extent in females (SI-89). 

One month after the catalepsy experiments, brainstem auditory evoked potentials 

(BAEPs) were used to evaluate hearing (SI-100). For the BAEP measurements, rats were 

sedated and electodes were implanted to the auditory brainstem and the response was 

measured as auditory clicks were issued of different strengths to establish a threshold. 

HBCD caused increased thresholds for BAEP in the 0.5 to 4 kHz range by 5 to 9 dB, but 

only in males. In addition to the increased threshold, HBCD appeared to induce and 

increased latency period in the lower frequency range. The authors note that the isolated 

effect on the lower frequency range suggests that HBCD affects the apical cochlea which 

is known to be dependent on thyroxine for development. 
S-56

In a preliminary report, Eriksson et al. (SI-101) reported that neonatal mice exposed to 

0.9 to 13.5 mg TM-HBCD/kg bodyweight on day 10 exhibited developmental neurotoxic 

effects, including changes in spontaneous behaviour, learning and memory when assessed 

at 3 moths of age. Timing for the HBCD exposure was targeted at the critical phase of 

neonatal development in mice (first 3-4 weeks of life) when the maturation processes of 

the CNS is particularly vulnerable. In the follow up report neonatal mice exposed to a 

single oral gavage dose of 20% fat emulsion vehicle containing TM-HBCD at 0.9 or 13.5 

mg/kg were assessed at 3 months for spontaneous movement over a sixty minute period. 

HBCD exposure reduced activity during the initial 20 minute period, and mice from the 

high dose group showed greater activity during the final 20 minutes of the 60 minute 

assessment, relative to the reference group receiving only the vehicle by gavage. The 

ability to learn to locate a platform in a swimming test was also examined. Mice from the 

13.5 mg/kg group learned to locate the platform significantly more slowly than mice from 

the refernce or low dose groups. When the platform was repositioned to examine relearning behaviour, this group again underperformed relative to the other groups. Similar 

behavioural aberrations observed in animals exposed to polychlorinated biphenyls 

(PCBs) and polybrominated diphenyl ethers (PBDEs) are cited by Eriksson et al. (2006). 

Finally, linkages between HBCD exposure, thyroid disruption and brain development 

were reported using a fish animal model. The effects of HBCD exposure on 

smoltification in Atlantic salmon were examined as there is growing evidence that 

exposure to chemicals during freshwater phase of salmon life can affect smoltification 

(SI-92). Interactions of HBCD with this process are even more relevant considering that 

the profound osmoregulatory adjustments required during the transition from freshwater 

to saltwater are thought to be regulated by thyroid hormones. Olfactory imprinting 

responsible for salmon returning to their natal streams is related to a T4 surge that may 

mediate the development of olfactory neurons during critical developmental windows. 

Salmon parr were exposed to HBCD and then transferred to seawater. The peak of 

thyroxine appeared a week earlier in HBCD exposed fish than in controls. HBCD also 

attenuated the olfactory responses of HBCD exposed fish. The authors report that they 

had shown in an unpublished study that adults olfactory ability was unaffected by HBCD, 
S-57

suggesting that the disturbance in these smolts was not at the level of olfactory mechanics 

but instead at the level of olfaction development. Collectively the studies above suggest 

that there is increasing evidence that HBCD may elicit effects on the nervous system at 

the subcellular, cellular, tissue and whole organism levels. However, additional research 

is required to examine the potential for long term effects in organisms exposed to 

environmentally relevant concentrations of HBCD. 

Oxidative Stress 

There is increasing evidence that HBCD is able to induce oxidative stress and apoptosis, 

and much of this evidence has been derived using fish models (SI-102-104). Altered 

activities of enzymatic antioxidants and indices of oxidative stress have been documented 

in rainbow trout (Oncorhynchus mykiss) and in the marine species eelpout (Zoarces 

viviparous) exposed to dietary HBCD (SI-103) and in zebrafish exposed to waterborne 

HBCD (SI-104). Deng et al. (2009) examined oxidative stress and the apoptosis pathway 

in zebrafish embryos exposed to waterborne HBCD. Based on their results, the authors 

suggested that the apoptotic cell death program may actually be initiated by the 

generation of reactive oxygen species. Zhang et al. (SI-105) examined the ability of 

HBCD to induce oxidative stress in brain of Chinese rare minnows (Gobiocyprinus 

rarus) exposed to waterborne HBCD. Brain was the focus of this study because of its 

high oxygen metabolism and high content of polyunsaturated fatty acid (PUFA) lipids, 

making it susceptible to reactive oxygen mediated damage. Reactive oxygen species and 

protein carbonyl content were induced in brain relative to untreated animals. Zhang et al. 

(SI-105) concluded that HBCD exposure can induce the generation of ROS leading to 

compromised antioxidant defence status and resulting oxidation of proteins, lipids and 

DNA within nervous tissue. 

Zhang et al. (SI-105) also evaluated the ability of HBCD to induce oxidative stress in 

mammalian cells. The capacity of individual HBCD diasteroisomers and their + or – 

enantiomers to induce oxidative stress in human hepatoma (Hep2G) cells exposed to 

environmentally relevant concentrations (0.5, 1.0, 5.0 or 10.0 ug/ml) for 24 hours in vitro 

was examined. Toxicity was evaluated using three different assays (MTT, resazurin, 
S-58

LDH release assay) and by detecting ROS using a fluorescent probe. The assays 

consistently found greater toxicity and ROS generation from exposure to the + 

entantiomers relative to the – forms, with ROS generation being significant at 0.5 μM and 

above for β and γ diastrereoisomers, but only at 5 μM and above with α-HBCD. 
Table S.23. Brief descriptions of studies examining effects of HBCD on thyroid endpoints, reproduction, nerve stress in biological systems where in vivo or in vitro HBCD exposure has been quantified, or where dietary reported. Study numbers are plotted against lowest observed effect levels in Figure 2. * denotes studies for which exposure was quantified as ng ml-1. 

Study Summary 

 Thyroid disruption 

SI-90 

Van der Ven et al. defined a bench mark dietary dose of 1.6 mg HBCD/kg/day that equated to 43 ug/g 

thyroid effects. This included lower thyroid weight (most sensitive parameter), lower circulating TT4activity in female rats (Effect level=43000 ng/g liver w/w). 

SI-89 

Germer et al. studied the same rats as van der Ven et al. and showed induction of CYP2B1 mRNA an

PROD activity, suggesting phenobarbitol type inducing activity of HBCD. They also showed CYP3A

activity induced more in females than males (significant at 3 mg/kg/day which is approximately equal 

(lipid corrected) according to van der Ven et al. derived concentrations. (=920,000 ng/g liver lipid) 

SI-106 

Palace et al. observed higher T3 in γ-HBCD group at day 56 and day 14 depuration, lower T4 in α-, β

and day 7 depuration in γ-HBCD. Lower Deiodinase activity in all groups at day 56. Also gland hypefrom days 56 to day 14 of depuration. From Law et al. approximate concentrations of HBCD in musc265.4 ± 97.9 (ng/g lipid weighted), β = 74.1 ± 29.0 ng/g, γ = 48.6 ± 33.4 ng/g. LOEL was for thyroid 

at 49 ng/g lipid weighted. 

SI-83 Fernie et al. exposed male kestrels to 54 μg/g in food for 3 weeks prior to mating time had lower TT4 

concentrations in plasma was 3396 ng/g (lipid weighted). 

SI-84* Hamers et al.: 1 uM α-HBCD increased cell growth in GH3 pituitary cell line (= 641ng/ml not lipid co

HBCD displaced T4 from TTR at IC50 of 12 and 25 μM respectively. 

SI107* Yamade-Okabe et al. demonstrated HBCD enhanced thyroid receptor mediated gene expression at 3.1

corrected) in presence of the thyroid hormone, triidotyronine (T3). 

SI108* Sakai et al.: HBCD shown to have an affinity for receptors at 5 μM (= 3205 ng/ml) 
SI-78* Schriks et al. reported γ- or TM-HBCD potentiated growth of GH3 rat pituitary cells mediated by T3 

ng/ml not lipid weighted) in presence of T3. 

SI-77* Schriks et al.: HBCD potentiated Xenopus laevis tail tip regression ex vivo in the presence of T3 at 1μ

weighted) 

SI-81* Crump et al.: CYP2H1, CYP3A3, UGT1A9 mRNA were upregulated, and TTR and L-FABP mRNA 

downregulated by α or TM-HBCD at 1μM (= 641 ng/ml not lipid weighted). 

SI109* Fery et al.: TM-HBCD increased CYP3A mRNA expression in primary cultured rat hepatocytes and i

line at >1μM (=641 ng/ml not lipid weighted) 

SI-80 

Palace et al. Juvenile rainbow trout pre-exposed to α-, β- or γ-HBCD diasteroisomers for 32 days had 

T4 delivered via gavage (less in thyroid and more in blood). β- and γ-HBCD exposed groups had low

enzyme actvity. *Assumed similar tissue concentrations to those reported by Palace et al. 2008 (49 ngsimilar methodology. 

SI-96 

Ema et al.: rats were fed HBCD (0, 150, 1500 or 15000 ppm) for 10 wks premating, through mating, gIncreased thyroid weight in males and females at dietary exposures of ≥ 15000 ppm, increased plasma dietary exposures of ≥ 1500ppm, and decreased plasma T4 in F0 males and females at dietary exposu

reported was for elevated plasma TSH at 1500 ppm dietary exposure which equated to 100,000 ng/g/d

SI-92* 

Lower and Moore: Atlantic salmon pre-smolts exposed to 11 ng/L (=0.011 ng/ml) waterborne TM-HB

seawater transfer, then seawater without HBCD for 20 days. Peak of T4 prior to smoltification was eaolfactory responses to conspecific urine was lower among HBCD treated fish. No tissue residues of Hexposure concentration was employed and no tissue residues were reported and there was no lipid we

SI-95 

Saegusa et al.: pregnant female rats were exposed to dietary TM-HBCD from gestational day 10 to 20 

effects on offspring. Thryoid weight increased at 80-212 mg/kg/d and thryoid follicular cell hypertropmg/kg/d in adult females. Among the offspring, Lower circulating T3 was identified at 11 weeks old aincreased TSH and lower T3 was reported at 800-2200 mg/kg/d on post natal day 20. No tissue residu

SI-76* Aniagu et al. exposed human hepatocarcinoma cells (HepG2) to 0.03 or 0.3 ng/ml TM-HBCD for 72 

PCR showed down regulation of thyroid hormone receptor α (THRα) at 0.3 ng/ml (not lipid weighted
SI-95 

Saegusa et al. (2009). Pregnant females rats exposed to dietary TM-HBCD (0,100, 1000, 10,000 ppm 

until 20 days post weaning. Thyroid weight increased with glandular hypertrophy and increased thyro

well as decreased circulating T3 among 10,000 ppm group (=800 mg/kg/d) 

Nervous system effects 

SI-97* Reistad et al.: TM-HBCD induced death of rat cerebral granule cells (CGC) in culture with an LC50 o

ng/ml). 

SI-99* Mariussen and Fonnum showed inhibited uptake of neurotransmitters dopamine (IC50 of 3 μM for HB

brain synaptosomes with a lowest observable effect of 1μM for glutamate uptake (=641 ng/ml). Not li

SI-98* 

Dingemans et al.: HBCD inhibited depolarization evoked increase in intracellular Ca2+ influx via volt

and depolarization-evoked vesciular catecholamine release in rat pheochromocytoma (PC12) cells wi2 μM (=1282 ng/g). Not lipid weighted. 

SI-100 

Lilienthal et al.: elevation of auditory thresholds in male rats exposed to HBCD corresponding to liver weighted (=Bench Mark Dose Level of 5% for this parameter, = 7000 ng/g liver lipid). In females, hawas reduced by HBCD at 16 μg/g in liver with a BMDL of 16.2- μg/g liver lipid. This BMDL was se

clinical intervention at this level of dysfunction in humans = 16,000ng/g liver lipid. 

SI-110 

Eriksson et al. exposed mice to single oral gavage of TM-HBCD (0.9 mg/kg or 13.1 mg/kg) on post nspontaneous behaviour and water maze learning activity in males at 3 months old. Spontaneous behav20 minute interval at 0.9 mg/kg (=dietary exposure. Water maze learning was impaired in 13.1 mg/kg also impaired. 

SI-95 Saegusa et al.: pregnant female rats exposed to dietary TM-HBCD (0, 100, 1000, 10000 ppm in diet) 

days post natal to examine effects on offspring. Male offpspring brains were examined immunohistoc11 had reduced CNPase positive oligodendrocytes in the cingulate deep cortext at 10,000 ppm ( LOE 

Oxidative Stress 

SI104* Hu et al.: zebrafish (2 hrs postfertilization) exposed to aqueous HBCD (0, 0.002, 0.01, 0.5, 2.5 and 10 

analyzed for antioxidant enzyme superoxide dismuatse and the oxidative stress measures, malonaldeh

mg/L but SOD results were mixed. LOEL = 500 ng/ml 
SI-105 

Zhang et al. exposed Chinese rareminnows to waterborne 1-500μg/L HBCD for 14, 28 or 42 days and 

oxygen species and evidence of membrane breakdown (TBARS), protein carbonyl, and inhibited antibrain after 42 days at a LOEL of 10μg/L. Whole body HBCD was measured after 42 days and the 10 

approximately 100 μg/g = 100,000 ng/g) 

SI105* 

Zhang et al. exposed HepG2 cells to individual HBCD diastereoisomers and enantiomers (0.5, 1, 5 or 

measured cell viability by MIT, resazurin assays and LDH leakage as well as reactive oxygen via DC

evident at 

≥0.5 μg/ml (MIT, resazurin) with γ≥β

>

α. (+) enantiomers were more toxic than (-) or racem

sensitive than the other 2 assays. ROS increased at 0.5 μg/ml and above with same order of toxicity o

(LOEL =500 ng/ml) 

Reproduction and Development 

SI-83 

Fernie et al.: kestrels exposed to TM-HBCD (54 μg/d/ individual) exhibited earlier egg laying, increa

slower embryonic growth. Male birds had increased testosterone, were heavier, and exhibited less cou

= 3396 ng/g lipid weighted. 

SI-96 

Ema et al.: rats exposed to HBCD in diet (0, 150, 1500 or 15000 ppm) for 10 wks premating, through lactation had fewer primary folicles in F1 females 

≥1500 ppm, decreased bodyweight associated with 

F1 males and females at 

≥ 15000ppm and more litters lost among F1 parents at

≥ 15000ppm. LOEL w

ng/g/d 

SI-90 

van der Ven et al. Parental generation rats exposed via food to HBCD premating to weaning. F1 gene

bone mineral density (Bench Mark Dose Level (BMDL) = 0.056 mkd which was equivalent to 1.3 μg 

weighted), decreased plasma apolar retinoids (BMDL = 1.3 mkd, =79μg/g liver) and potentiation of im

(BMDL=0.46mkd, = 7 μg/g in liver). LOEL for trabecular bone effects at 1.3 μg/g = 1300 ng/g lipid w

SI-76* Aniagu et al.: exposed human hepatocarcinoma cells (HepG2) to 0.03 or 0.3 ng/ml TM-HBCD for 72 

PCR showed down regulation of estrogen receptor 

α and 

β at 0.3 ng/ml. 

SI-84* Hamers et al.: using CALUX assays showed antiestrogenic activity of 

β- and 

γ-HBCD, IC50 at 5-10 μ

antagonized progesterone receptor at IC50 of 1.5 μM (LOEL = 1025 ng/ml) 

SI104* Hu et al.: zebrafish (2 hrs postfertilization) exposed to aqueous HBCD (0, 0.002, 0.01, 0.5, 2.5 and 10 

hatching identified at 

≥2 ng/ml. 
SI-111 

Japanese MOE Study #08-016 (2009). Listed in Annex E for Japan in the call for more information by 

Sixty pairs of Japanese quail exposed to 5, 15, 45 or 125 ppm HBCD for 6 wks. Hatched eggs reared 

Death increased and overall survival was lower at 

≥15ppm (=2.1 mkd). Reproductive ability index wa

wks exposures. LOEL = 15ppm = 2100ng/g/d. 

SI102* Deng et al. (2009). Four hr post fertilized zebrafish embryos exposed to 0, 0.05, 0.1 or 1 mg/L for 96 

decreased, and apoptosis elevated at 0.5 mg/L. LOEL = 50,000ng/ml. 
S-64

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