## Supporting Information

Hexabromocyclododecane: Current Understanding of Ch emistry, Environmental Fate and Toxicology and Implications for Global Management

Christopher H. Marvin 1* and Gregg T. Tomy 2,3* ; James M. Armitage 4,5 ; Jon A. Arnot 4 ; Lynn McCarty 6 ; Adrian Covaci 7 and Vince Palace 2,3

1 Water Science and Technology Directorate, Environment Canada, Burli ngton, 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, Denmar k

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 tablesS-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 (P OV )
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-63S-3



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

Measured and modeled physical-chemical properties for (technic al) 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 (T M ; o C or K), water solubility (S; mol
· L -1 ), vapor pressure (P; Pa), Henry’s law constant (H; Pa
· m 3

· mol 1 ), and the dimensionless equilibrium partition coefficients between a ir and water (K AW ), octanol and water (K OW ), and octanol and air (K OA ). 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 reporte d 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 val ue 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 t he properties of theS-4

subcooled liquid state of the chemicals which are used for the therm odynamic consistency calculations. The fugacity ratio can be estimated as

F = exp (6.79(1 – T M /T)) (S.1)

where T is the system temperature (K). F is 0.023 for t-HBCD ( T M = 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 s tate values (subscript L) as:

P L = P S /F (S.2)

The fugacity ratio is expected to be different for the diaste reoisomers due to different melting points and different entropies of fusion; however, isomer speci fic 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 m elting point for all diastereoisomer specific calculations, a value of 0.023 was selecte d for F for simplification of the present calculations. Isomer-specific meas urements for the entropy of fusion are required to obtain more accurate estimates of the subc ooled liquid state property values.

Water solubility
Table S.1 summarizes two studies that have measured the solid stat e water solubilities S S 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% γ

HBCD; however, no diastereoisomer-specific analyses were conduct ed ( SI

- 7 ). The first study ( SI

- 6 ) had fewer impurities and unknowns then the second study ( SI

- 7 ) and allS-5

isomers could be quantified; therefore, the data from ( SI

- 6 ) are considered to be the most
reliable measurements available. The S S (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 μ g/l (with a total
HBCD dissolved concentration of 2.4 μ g/l), β -HBCD at 14.7 μ g/l (with a
total HBCD dissolved concentration of 39 μ g/l) and α -HBCD at 48.8 μ g/l (with a total dissolved HBCD concentration of 65.6 μ g/l). At this level,
65.6 μ g/l dissolved HBCD, the total content of HBCD in the water is 610

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

In other words, testing for possible effects of t-HBCD at concent rations greater than
about 2 or 3 μ g · L -1 will result in exposures to test organisms of non-dissolved (precipitate) γ -HBCD, thus compromising the reliability of the water based tox icity 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 summar ized in ( SI

- 8 ). Measured water solubilities S S for α -, β - and γ -HBCD correspond to approximately 0.08, 0.02 and 0.003 μ M, 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 VariationS-6

Model estimates for the water solubility were also obtained and r eviewed. These
estimates do not differentiate between diastereoisomers. When a log K OW value of 5.63 is used as input (see below), EPI Suite Ver. 4.0 estimates S S values of 1.3 μ g · L -1 and 3.0

μ g · L -1 from the WSKOWWIN Ver. 1.41 and WATERNT Ver. 1.01 submodels, respectively ( SI

- 10 ). These S S 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 S S of 0.31 μ g · L -1 for 1,2,5,6,9,10-HBCD, which is consistent with the diastereoisomer configurations (vicinal bromine substitution). The SPAR C S S estimate is about 1 order of magnitude lower than the measured values for γ -HBCD.

The measured S S for the individual HBCD diastereoisomers from Study 1 ( SI

- 6 ) are
considered to be the most reliable measurements available. The S S (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 value s S L (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 μ g · L -1 (S S = 4.67×10 -6 mol.m -3 , S L = 2.01×10 -4 mol.m -3 ) is selected as a “representative” value for t-HBCD for the thermodynamic consistency ca lculations.

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

o C using the effusion method and 6.27×10 -5 Pa at 21 o C using the spinning rotor method. The effusion method used in the aforementioned test and the spinning rotor m ethod 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 percent ages 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 P L

(subcooled) of 2.41×10 -5 Pa at 25 o C for t-HBCD using the gas chromatographic-retention time method, which is suitable for substanc es with vapor pressures as low as 10 -5 Pa. This corresponds to a P S of 7.23×10 -7 Pa (F = 0.023).S-7

EPI Suite Ver. 4.0 estimates a P S of 2.25×10 -6 Pa ( SI

- 3 ) and SPARC Ver. 4.2 estimates a
P S of 2.63×10 -9 Pa (http://ibmlc2.chem.uga.edu/sparc/). The EPI Suite estimate i s 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 measure ments for vapor pressure. Goss and colleagues ( SI

- 13 ) used the COSMOtherm model to calculate K OW and K AW partitioning of α -, β -, γ -HBCD diastereoisomers at 25 o C (see details below). From the isomer-specific quantum chemical software estimates of K AW and from isomer-specific estimates of water solubility S S it is possible to estimate P S for the three isomers as

P S = K AW .RT.S S (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 o C) are minimal. The resulting isomer-specific estimates for P S 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 tim es 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 considerabl y off the real values”.
COSMOtherm also underestimates K AW for hexachlorocyclohexane (HCH) ( SI

- 13 ). Since
the K OW estimates for HCH and HBCD are similar to measured esti mates, this suggests
that errors in COSMOtherm K AW estimates relate to errors in P. Thus, the COSMOtherm
estimates for K AW

and P S are not considered reliable; however, in relative terms the P S results are similar to the measured values for S S 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 P S of 7.23×10 -7 Pa (P L = 3.11×10 -5 Pa based on F of 0.023) using the gas saturation method is considered the most reliable es timate for the commercial HBCD mixture (t-HBCD). This value was selected as an input forS-8

thermodynamic consistency analyses (see below). Estimates f or the diastereoisomers were derived by assuming the P L for γ -HBCD is approximately 2.0×10 -5 Pa and that the relative differences in values estimated by COSMOtherm ar e applicable such that P L 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, gre ater 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-HBC D or the diastereoisomers in octanol. Estimates are provided below in the se ction on the octanolwater partition coefficient.

Henry’s law constant (Air-water partition coefficient) There are no measured estimates for Henry’s law constant H (P a.m 3 .mol -1 ) currently available. The dimensionless air-water partition coefficient K AW 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.m 3 .mol -1 using the bond and group methods, respectively ( SI

- 10 ). The group method estimate is considered erroneous. The log K AW estimate corresponding to the bond method is -4.15. H can also be estimated from vapor pressure and water solubility as
P S /S S or P L

/S L . 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.m 3 .mol -1 (log K AW = -4.32). SPARC Ver. 4.2
estimates a value for H of 0.0055 Pa.m 3 .mol -1 (log K AW = -5.66) using the same approach, i.e., P L

/S L (http://ibmlc2.chem.uga.edu/sparc/). The estimates from EPI Suite bond method and the relative solubilities of the selected values for vapo r pressure and water solubility are consistent, independent estimates. The EPI S uite bond method estimate is selected for thermodynamic consistency calculations for t-HBCD.

Diastereoisomer estimates for K AW have been reported by Goss et al. using the COSMOtherm model ( SI

- 13 ). The log K AW

values are -8.84, -9.20, and -8.64 for α -, β -, γ

HBCD, respectively. These values correspond to H values of 3.6×10 -6 , 1.6×10 -6 , 5.7×10 -6

Pa.m 3 .mol -1 for α -, β -, γ -HBCD, respectively. As discussed above and by the authors theS-9

absolute estimates are considered to be erroneous, but the relative va lues are considered reasonable. Diastereoisomer estimates of log K AW from selected values for P L /S L are - 4.27, -4.63, and -4.07 for α -, β -, γ -HBCD, respectively. These estimates are dependent on
the values for P L

/S L , which have already been selected for thermodynamic consistency calculations. Independent estimates for the diastereoisomers wer e derived by assuming log K AW for γ -HBCD is equal to the EPI Suite estimate (-4.15) and that the re lative differences in values estimated by COSMOtherm are applicable such that log K AW for α and β -HBCD are -4.35 and -4.71, respectively.

A Henry’s law constant of 0.75 Pa.m 3 .mol -1 for t-HBCD was reported as the ratio of
selected values for P S and S S ( SI

- 3 ). The different values for P S and S S selected in this
critical review result in an estimate of 0.120 Pa.m 3 .mol -1 . The EPI Suite  bond method estimate ( SI

- 10 ) of 0.174 Pa.m 3 .mol -1 provides an independent and consistent estimate for
inclusion in the FAV calculations.

Octanol-water partition coefficient A measured log K OW

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, alt hough 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 diast ereoisomers in the water column ( SI

- 4 ).

Diastereoisomer-specific log K OW values were estimated for α - β - and γ -HBCD using reversed-phase high-performance liquid chromatography (RP-HPLC) me thods ( SI

- 14 ). The estimated log K OW 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 idea l because of the structural dissimilarities between the organochlorines and HBCD ( SI

- 14 ). Assuming the solubility in octanol is similar for the diastereoisomers, the K OW estimates are relativelyS-10

consistent with measurements of water solubility such that K OW increases with decreasing water solubility. The RP-HPLC derived K OW values for the diastereoisomers are comparable but lower than the measured K OW 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 K OW is 6.73
with a standard deviation of 0.76. These model estimates do not differenti ate between diastereoisomers. Goss et al. ( SI

- 13 ) reported diastereoisomer log K OW 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 a s the generator
column K OW

measurement for t-HBCD and the diastereoisomers determined fr om calibration using RP-HPLC. A notable difference is the relat ive ranking of hydrophobicity between the RP-HPLC and COSMOtherm estimates.

There are uncertainties in the measured and estimated values for K OW for the commercial
mixture and the diastereoisomers. The solubilities of the diastereoi somers in octanol are expected to be similar and differences in K OW are expected to be the result of differences in water solubility. SPARC Ver. 4.2 calculates an activity coeff icient in octanol γ O of 12.88 (unitless) and a solid state solubility in octanol S OS of 16.21 mol.m -3 (liquid state solubility in octanol S OL

of 698 mol.m -3 ) for HBCD. The S OL value was selected as input for the thermodynamic consistency calculations for the commercial HBCD mixture and
the diastereoisomers.

The measured log K OW value of 5.63 was selected as input for the consistency
calculations for t-HBCD. For the diastereoisomers the estimat es 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 potentia l underestimation of HBCD using the generator column method and the lac k of ideal calibration substances for the RP-HPLC method, the selected val ues are considered as reasonable starting point for the thermodynamic consistency calculations.S-11

Table S.2. Model estimates for the octanol-water partition coefficient (K OW ). log K OW 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 K OW 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-HB CD and the three diastereoisomers. The data selected as input for the calculati on 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 ind ependent. In this context “independent” means that physical-chemical properties used in the har monization
calculations were obtained from distinct measurements or estima tes. For example, K AW estimates derived from ratios of P S and S S were not included in the calculations because
the selected estimates of P S and S S were already used as input. Reliability was considered high if multiple data sources provided similar values. Model estimat es were considered as well as measured data. The harmonization method also requires and provi des relative variance values for chemical properties. These values provide guidance as to the relative uncertainty of the different properties. The relative variance va lue inputs were selected based on professional judgement of the general reliability of the various parame ters.S-12

The issue of the mutual solubility of water and octanol can be importa nt when seeking to obtain thermodynamic consistency in solubility and partitioning parame ters ( SI

- 16,17 ).
K OW measurement techniques, such as the shake-flask and slow stir met hods, 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 (K OA ) and of the solubility in octanol, on the other hand, are obtained using “dry” octanol, i.e., octanol that has not been saturated with wate r. Similarly, measurements of water solubility and air-water partition coeffi cients 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 substanc es, relationships have been proposed for converting measured K OW values to ratios of concentrations in pure octanol and pure water C O /C W . For the thermodynamic consistency analysis, the equation recommended by Beyer et al. ( SI

- 17 ) was used to
convert K OW

to C O /C W

prior to adjustment and to convert the adjusted C O /C W back to
K OW after adjustment.

Table S.3. Selected values for physical-chemical properties and relative va riance values used as inputs for thermodynamic consistency calculations.
Substance P L , Pa S L , mol·m -3 S OL , mol·m -3 Log K AW Log K OW a Log K OA 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 K OW values corresponding with log (Co/Cw) values of 6. 02, 5.97, 5.76, and 5.89 for t-, α -, β -, and γ

HBCD, respectively. b values not used as inputs for calculation of FAVs because they are not “independent” values, i.e., th ey were derived from vapor pressure and octanol solubi lity estimates (see text). The values are included for comparisons with FAVs.S-13

Table S.4 lists the thermodynamically consistent FAVs for the r elevant solubility and partitioning properties. The percentages of adjustments were modest for changes from the selected input values to the calculated FAVs for most properties s uggesting that the data are generally consistent and reasonable. Notably, all of the K OW values were slightly increased and the solubility in octanol decreased in all cases exc ept for α -HCBD for which it increased (largest percentage adjustment). Notably, the a djustment 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 K OW are consistent with the rankings of the measured water solubilities and the independent K OW estimates by Hayward et al. ( SI

- 14 ) that were not used in this procedure. The FAVs for K OA 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 re view; however, a thermodynamic consistency calculation should generally not be used for mixtures, but rather for discrete substances with distinct properties. Consideri ng the uncertainty in the FAVs for t-HBCD in the context of the greater uncertainties a ssociated 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 compar atively low.S-14

Table S.4. Final adjusted values (FAVs at 25 o C), relative variance values, percent adjustments, and assigned confidence factors used in the present st udy for mass balance modeling and data analyses obtained from the thermodynamic consis tency calculations ( SI

- 1 ).
Substance P L , Pa S L , mol·m -3 S OL , mol·m -3 Log K AW Log K OW a Log K OA 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 K OW values corresponding with log (Co/Cw) values of 6. 22, 5.68, 5.81, 6.25 for t-, α -, β -, and γ

HBCD, 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 m edian 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 0.95 y probabilit = M Cf < X < Cf M 

    

⋅ (S.4)

Thus, the variance in the lognormal distribution in creases with an increase in Cf . Values of confidence can be derived empirically or subject ively or in combination.



### Section S.2: Additional details on data characteri zing the degradation
of HBCD

Air The atmospheric degradation half-life (HL AIR ) can be calculated using the following equation:

]

[ 2 ln OH k HL

OH AIR = (S.5)

where k OH is the 2 nd -order rate constant for reactions for OH radicals (cm 3 /molecule/s) and [OH] is the concentration of OH radicals (molec ule/cm 3 ). The k OH of HBCD estimated by AOPWIN v1.92 is 6.12x10 -12 cm 3 /molecule/s. According to the model documentation, 90% of the AOPWIN estimations are wi thin a factor of 2 of empirical
data and 95% are within a factor of 3. Based on th is model performance, it is assumed that the uncertainty in the k OH estimation for HBCD is a factor of 3 (in either di rection).

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 HL AIR .

Table S.5. Range in estimated HL AIR for HBCD based on estimated [OH].
[OH] mol./cm 3 Region Time Period (h) HL AIR

(d) Min (d) Max (d) 1.5x10 6 Global 12 1.7 0.6 5.2 9.2x10 5 Global 24 1.4 0.5 4.3 9.8x10 5 N.

America 24 1.3 0.4 4.0S-16 8.5x10 5 S. America 24 1.5 0.5 4.6 1.1x10 6 Global 24 1.2 0.4 3.6 9.7x10 5 Global 24 1.4 0.5 4.1
As shown in Table S.5, the assumed uncertainty in k OH results in ranges in HL AIR 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 empiri cal data on biodegradation in the water column is considered unreliable. Biodegradation po tential can be assessed using
BIOWIN TM ( 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 fo r HBCD is presented in Table S.6.
Note that no distinction can be made between isomer s or enantiomers.

Table S.6. BIOWIN TM 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 qualit ative agreement with observations that biodegradation proceeds quicker i n anaerobic environments compared to aerobic. However, the model output cannot be us ed 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 onS-17

extrapolating from qualitative BIOWIN output to qua ntitative half-live estimates. BIOWIN3 output for HBCD suggests that ultimate degr adation occurs over a time frame
of months. The default water half-life associated with this category, based on experimental primary degradation data for 44 compou nds, was suggested to be 120 d. However, the empirical half-life data for the compo unds in this category ranged from 1– 1420 d (median = 85 d, 10% trimmed mean = 126 d).

Arnot et al. ( SI

- 24 ) used BIOWIN TM model output and empirical half-life data to deriv e regression equations which relate numerical model o utput 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 BIOWIN TM model output and corresponding values calculated for HBCD.

Model Regression Equation r 2 Half-life (d)
BIOWIN1 (linear) log 10 (t 1/2 ) = -1.32 x + 2.24 0.72 105.2
BIOWIN3 (Ultimate deg) log 10 (t 1/2 ) = -1.07 x + 4.20 0.77 128.3
BIOWIN4 (Primary deg) log 10 (t 1/2 ) = -1.46 x + 6.51 0.78 85.9

where x is the numerical output generated for each respect ive BIOWIN TM 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 H BCD likely due to the lack of hydrolysable functional groups. The reaction of HB CD 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 aut hors demonstrated that the rateS-18

of loss of HBCD at 20 o C was appreciably faster under anoxic conditions in both media. Relative to biologically sterile controls, biotrans formation of HBCD was faster in the presence of microorganisms and t 1/2 ’s values ranged from 11 to 32 days (aerobic) and 1 .1 to 1.5 days (anaerobic) in sediment. In soil, t 1/2 ’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. Beca use 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 Br 2 ( SI

- 26 ). Estimated t 1/2 ’s for the α -, γ - and β -isomers were 128, 92 and 72 daysrespectively in freshwater sediments und er aerobic conditions ( SI

- 26 ). Initial
concentrations of α -, γ - and β -HBCD in these experiments were reported to be in t he range of 4–5 mg/kg dry weight. While there were no statistically significantly differences between the measured values, these meas urements 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 ( μ g/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 aer obic 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 ( 14 C radiolabelled) was reported to be approximately 3 mg/kg dry weight. Results were presented as % of initial nominal radi oactivity; 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 quantif ied. These data were considered to be indicative of negligible degradation and no degr adation half-lives were estimated.

There have been two studies on the degradation of H BCD in wastewater sludge ( SI

26,27 ). Both studies suggest that the α -isomer is the most recalcitrant. The rank order o f 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

t 1/2 ’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 und er 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 tha t conversion of γ - to the α -isomer was the most predominant ( SI

- 28 ). There was also a measured decrease in concentrat ion of
Σ HBCD 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 pro duct was reported. As one would expect, the photolytic degradation of the isomers o n 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 correspondin g mono - and di -OH forms of HBCD, PBCD and TBCD in select tissues of exposed ra ts. 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 relevanceS-20

The temporal evolution of concentrations of α -, β -, and γ -hexabromocyclododecanes (HBCDs), and pentabromocyclododecenes (PBCDs — degr adation products of HBCDs) was studied in separate aliquots of a well-homogeni zed indoor dust ( SI

- 28 ). These were: (a) exposed to natural light, and (b) kept in the d ark. Results revealed a slow degradative loss of HBCDs via elimination of HBr. Tetrabromocyc lododecadienes (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 lo ss of
HBr.



### Section S.3: Additional details on data and approac hes used to characterize Overall Persistence (P OV 

### )

Table S.8. Degradation half-lives selected for HBCD benchmar king ( 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 P OV benchmarking exercise representing POPs were as follows: HCB, PCB28, PCB101, PCB180, Aldrin and He ptachlor. 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 ran ge of physicalchemical properties. Most of these substances have been used in previous POPs benchmarking assessments ( SI

- 31 ). Aldrin and heptachlor were selected because thes e POPs have shorter half-lives in air than most POPs and these half-life estimates areS-21

comparable to those for HBCD. Table S.9 lists the p artitioning properties and Table S.10 lists the degradation half-lives ( HL s).22Table S.9.Selected physical-chemical properties for the model benchmarking simulations. Data are from (31) except for aldrin andheptachlor which are from (SI-1,32).Status NameAbbrev MW (g.mol-1) log KOWlog KAWlog KOAPOP hexachlorobenzene (HCB) HCB 284.8 5.61 -1.51 7.12POP 2,4,4'-trichlorobiphenyl (PCB 28) PCB28 257.6 5.66 -1.93 7.86POP 2,2',4,5,5'-pentachlorobiphenyl (PCB 101) PCB101 326.4 6.38 -2.08 8.83POP 2,2',3,4,4',5,5'-hexachlorobiphenyl (PCB 180) PCB180 395.3 7.15 -2.51 10.17POP AldrinAld 364.9 6.25 -2.01 8.26POP Heptachlor Hept 373.4 5.96 -1.78 7.74POP alpha-hexachlorocyclohexaneα-HCH 290.8 3.88 -3.59 7.48POP beta-hexachlorocyclohexaneβ-HCH 290.8 3.91 -4.83 8.74POP gamma-hexachlorocyclohexaneγ-HCH 290.8 3.76 -3.96 7.72POP 2,2',4,4',5-pentabromodiphenyl ether PBDE99 564.7 6.76 -3.67 11.26Non-POP BiphenylBiphenyl 154.2 4.06 -1.96 6.02Non-POPp-cresolp-cresol 108.1 1.97 -4.26 6.23Non-POP AtrazineAtrazine 215.7 2.73 -6.84 9.57Non-POP tetrachloromethane CCl4153.8 2.83 0.19 2.6423Table S.10.Selected primary transformation half-lives (HL; days) for the model benchmarking simulations. Data are from (SI-31)except for aldrin and heptachlor in which environmentalHLs were selected based on (10,24,32) and fishHLs are from (SI-10,33).Avian and mammalianHLs (Av/Mam) for aldrin and heptachlor were assumed to be 1/3 estimates for fish. Transformation half-lives(HLN) in biota are listed normalized to 1 kg body sizeand scaled to mass (Mi; kg) specific values (HLi) for different species in themodels asHLi=HLNx (Mi/1)0.25(SI-33).Status Abbrev HL_Air HL_Water HL_Soil HL_Sediment HL_Fish (1 kg)HL_Av/Mam (1 kg)POP HCB 650 1000 1500 9000 860 290POP PCB28 6 300 450 2700 190 8.3POP PCB101 20 1200 1800 10800 370 44POP PCB180 80 2200 3300 19800 1100 29000POP aldrin 0.2 1000 1000 2000 75 25POP heptachlor 1.5 25 50 150 90 30POPα-HCH 50 440 660 3960 55 20POPβ-HCH 50 900 1350 8100 90 30POPγ-HCH 50 400 600 3600 120 40POP PBDE99 20 850 1275 7650 285 95Non-POP biphenyl 3 15 22.5 135 5.2 0.2Non-POPp-cresol 0.3 5 7.5 45 0.35 0.2Non-POP atrazine 0.4 50 75 450 0.3 0.2Non-POP CCl4900 150 225 1350 0.2 0.2S-24

Model Selection

The P OV benchmarking exercise was reported in Arnot et al. ( SI

- 2 ) using the Risk Assessment Identification And Ranking (RAIDAR) mode l v2.0. Figure S.1 illustrates
P OV estimates for HBCD against POP and non-POP chemica ls (Tables S.9 and S.10) using The OECD Tool model ( SI

- 34 ). P OV estimates for HBCD range from 12 d (assuming lower bound environmental half-lives in T able S.8) to 1220 d (assuming upper bound environmental half-lives in Table S.8). Assum ing the median environmental halflives in Table S.8, the P OV for HBCD 123 d. In general, P OV predictions for HBCD are
more similar to P OV values for the non-POP benchmark chemicals than to P OV values for
the POP benchmark chemicals.

Figure S.1. Overall persistence (P OV ; 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 “w orst 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 characteriz ing 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 ).

C Dissolved 11 pg/L
Species Lipid Content (-) C LIPID

(ng/g lipid) C WET WEIGHT

(pg/kg) log BAF

(L/kg) Walleye 0.178 6.53 1.16x10 6 5.02 Whitefish 0.120 1.22 1.46x10 5 4.12 Mussels 0.095 8.25 7.84x10 5 4.85 Zooplankton 0.097 8.64 8.38x10 5 4.88 Emerald shiner 0.160 8.70 1.39x10 6 5.10 Goldeye 0.160 8.58 1.37x10 6 5.10 White sucker 0.152 3.66 5.65x10 5 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 tis sue ** recalculated following Harrad et al. ( SI

- 36 ); assumes fish is 5% lipid overall and equilibrium between all lipids in bodyS-26



### Section S.5: Modeled bioavailability, uptake effici encies and inherent bioaccumulation factors (BAFs)
Isomer-specific physical-chemical property estimate s are currently available for α -, β and γ -HBCD, allowing the model calculations shown below to be undertaken; no
distinctions are made between enantiomers of each i somer 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 μ g/L, respectively. The higher water solubility of the α isomer in comparison to the other two isomers translates into higher bioavailability in a quatic ecosystems only when the water solubility limits are exceeded (i.e., > 15 μ g/L). At environmentally-relevant concentrations (i.e., pg/L to ng/L range), the bioa vailable fraction in the water column is a function of partitioning between the dissolved an d solid phases in the water column, not the water solubility limit. For neutral organic com pounds, partitioning is largely controlled by the hydrophobicity of the compound, t ypically represented by the octanolwater partition coefficient (K OW ). 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 coeffic ient, K OC ), the freely dissolved i.e. bioavailable fraction of a compound in the water co lumn (F Dissolved ) can be estimated using the equations shown below:

DOC POC Dissolved Dissolved Dissolved M

M M M F +

+ = (S.6)

where M Dissolved , M POC and M DOC are the mass of the compound in the dissolved phas e, associated with particulate organic carbon (POC) an d dissolved organic carbon (DOC) respectively. Dividing all terms by M Dissolved (or C Dissolved V Dissolved ) yields the following equation relating F Disssolved to partition coefficients (K POC , K DOC ) and volumes of each phase in the water column (V Dissolved , V POC , V DOC ):S-27 Dissolved DOC DOC Dissolved POC POC Dissolved
V V K V V K F

+

+ = 1 1 ( 7 )

F Dissolved is calculated in the following illustrative exampl e 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:

C POC = 2.5x10 -7 kg/L C DOC = 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:

V POC = 2.5x10 -7 kg/L V DOC = 2.0x10 -6 kg/L V DIS = 1 - V POC - V DOC While more sophisticated approaches for estimating organic carbon-water partition coefficients are available (e.g. polyparameter line ar free energy relationships), the approach adopted here is to use K OW, which is reasonable as a first-approximation for a neutral apolar compound such as HBCD. K POC and K DOC are calculated as in Arnot & Gobas ( SI

- 37 ) as:

K POC = ρ POC K OW (S.8)

K DOC = ρ DOC K OW (S.9)

where ρ POC and ρ DOC are proportionality constants (L/kg OC) which relat e to sorptive capacity of organic carbon to that of octanol. The assumed values for these constants are
shown below.

The selected proportionality constants are ρ POC = 0.35 L/kg and ρ DOC = 0.08 L/kg.

Calculated F Dissolved 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 ofS-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 t he water column using parameters representing Lake Ontario

Source Log K OW F Dissolved 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 estimate s, the freely dissolved fraction of all isomers is essentially identical. Based on the se calculations (and those for sediment porewater, not shown here), differences in the bioa vailable fraction in the environment between the isomers are not likely to explain the e xtent of enrichment of α -HBCD in
biota.

Uptake efficiencies across the gills and gastrointestinal tract Empirical relationships relating the uptake efficie ncy for transport across the gills and gastrointestinal tract (GIT) to K OW are available in the literature for aquatic organi sms ( SI

- 37 ). The approach is based on two-film resistance th eory and similar relationships
have also been derived for humans and other air-bre athing organisms ( SI

- 38-40 ). The relationships ( SI

- 37 ) are shown below for uptake efficiency across the gill (E W ) and GIT (E D ):

1

)) / 155 ( 85.1(
− + = OW W K E ( 12 )

1 7 )0.2 10 0.3( −

− + ⋅ × = OW D K E ( 13 )S-29

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

Table S.15. Estimated uptake efficiencies across the gills an d GIT in aquatic organisms for HBCD isomers based on estimated log K OW.

Source Log K OW E W E D 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 m etabolic rate constant = 0), the aquatic food chain bioaccumulation model described in Arnot & Gobas ( SI

- 37 ) can be used to evaluate the inherent bioaccumulation poten tial of each isomer i.e. based on physical-chemical properties alone. The bioaccumul ation model calculations are useful because they integrate the considerations described above and also trophodynamics. To be consistent with calculations made above, the Lak e Ontario parameter set was used and bioaccumulation factors (BAFs) were calculated assu ming 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 conc entrations 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. C POC , C DOC ) are based on Lake Ontario as are the organisms in cluded and feeding relationships. Hence, direct compariso ns 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 concentr ations based on ( SI

- 30 ).

Isomer C WATER

(pg/L) % C SEDIMENT (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 a re shown in Table S.17b.

Table S.17a. Modeled log BAFs using the Arnot et al. FAV physi cal-chemical property estimates and water (total) and sediment concentrat ions 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-ch emical property estimates and water (total) and sediment concentrat ions 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 differenceS-31

is found for the highest trophic level consumer (La ke trout), reflecting the fact that preferential accumulation (based solely on log K OW ) 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 t hat differences in physical-chemical properties/bioavailability do not explain the exten t to which α -HBCD is preferentially
accumulated in biota. The model results obtained u sing 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 (ma ximum ~ 20%) is insufficient to explain field observations. Similar conclusions are obtained using terrestrial bioaccumulation models ( SI

- 38 ). The main finding based on these theoretical calc ulations
is that biotransformation and/or bioisomerization f avouring 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 depurati on 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 ki netics. The study shows the potential
for bioisomerization of the different isomers where by the isomers are apparently
converted from one isomeric form to another. Some o f the HBCD is not being eliminated or depurated from the organism but merely transform ed into another diastereomer;
therefore these “elimination” rates do not necessar ily account for the loss of HBCD from the organism via transformation to a metabolite.

Table S.18 summarizes the molar elimination and for mation of the diastereoisomers during the “elimination” phase based on the data re ported in Table 2 of Law et al. (see also Figure 1 from that study) ( SI

- 42 ). The data suggest that γ -HBCD is predominantlyS-32

converted to α -HBCD (70%) and only about 30% of the total elimina tion 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 per iod actually suggests a net “gain” of total HBCD. It is not possible for the fish to crea te 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 α

HBCD is predominantly eliminated with minimal conve rsion 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 γ

HBCD is converted to α -HBCD faster than α -HBCD is converted to γ -HBCD, which would support the overall observation for the atten uation of γ -HBCD isomers in food webs. Furthermore, if relative bioisomerization rat es 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 bioisomeriz ation and actual total elimination (fecal egestion, gill respiration and b iotransformation).

Elimination phase Other isomers formed (+) or lost (-) HBCD 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 γ

HBCD 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 kinet ics were reported for α -HBCD. IndeedS-33

non-first order kinetics may be the result of the r elatively faster rates of conversion of γ

HBCD back to α -HBCD. A “pseudo-elimination rate constant” based o n the molar quantity at the start of the elimination (Day 56) a nd 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 an d 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 es timate biotransformation and bioisomerization rates using a mass balance estimat ion method for fish when applied to the Law et al. dataset for rainbow trout ( SI

- 44 ). There are three possible biotransformation rate constant ( k M ) estimates that could be considered. The first is “loss of parent” diastereomer. This includes bioisomeriza tion from one diastereomer to a different diastereomer (and subsequent bioisomeriza tion reactions) and the formation of metabolites (elimination). This would be calculated based on the “total depuration” rate constant ( k

T ) reported by Law et al. ( SI

- 42 ) The second is biotransformation to a metabolite that is no longer HBCD. This value is ba sed on the reported total depuration rate constant (or estimated depuration rate constan t 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 th e difference
between the two estimated rate constants which is a ssumed to be the result of
bioisomerization to another isomeric form of HBCD b ut without elimination from the fish, i.e., 70% of the reported total depuration ra te 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 c ompeting, and presumably different rates, further confounds the interpretation of the data and the ability to accurately estimate bioisomerization and biotransformation (el imination) rate constants. Since the bioaccumulation models used in the present study do not currently account for bioisomerization only the second rate constant (los s 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 balanc e 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 k M values ( SI

- 44,45 ). Although an
estimate for γ -HBCD is calculated, the percentage of positive val ues is less than 25% corresponding to a low confidence category assignme nt for the estimated k M value ( SI

44,45 ). Although an estimate for α -HBCD is calculated, the output confidence factor i s greater than 10.3 corresponding to a “low” confiden ce category assignment for the
estimated k

M value ( SI

- 45 ). Using a rearrangement of Equation 2.5, the corre sponding primary biotransformation half-lives normalized HL N to a 1 kg fish are 3330 and 720 d
for α -, and γ -HBCD, respectively. Thus, the analysis from the me asured data from Law et al. ( SI

- 42 ) suggest “very slow” ( SI

- 2 ) and low confidence primary biotransformation half

lives for α -, and γ -HBCD.S-35

Table S.19. Mass balance model summary for estimating biotrans formation and bioisomerization rate constants from a dietary feed ing study in rainbow trout ( SI

- 42 ). Diastereomer
Parameter α -HBCD β -HBCD γ -HBCD

Value C

ƒ Value C ƒ Value C ƒ
W (kg) 0.25 1.5 0.25 1.5 0.25 1.5
L F (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
G D (kg/d) 0.005 W

2 0.005 W

2 0.005 W 2
C OX (mg/L) 9.9 1.25 9.9 1.25 9.9 1.25
E D N/R a N/A 0.41 2 0.41 2
K OW 2.40x10 5 3 2.95x10 5 3 6.31x10 5 3
k T (d -1 ) b 9.0x10 -3 4.4 c 4.41x10 -4 4.6 d 1.44x10 -3 4.2 d
k M (d -1 ) 1.7x10 -3 13 No estimate possible 3.0x10 -4 7 Cf : confidence factor; k T : total elimination rate constant; W : whole body weight of the fish; L F : whole body lipid fraction of the fish; T : water temperature; G D : feeding rate; C OX : dissolved oxygen concentration; E D : chemical transfer efficiency between the gastrointe stinal tract and the organism; K OW : octanol-water partition coefficient; k M : 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-ord er, for β -HBCD and γ -HBCD, the reported rate constants are corrected for the % of actual elimination of th e parent isomer as reported in Table 2.6. Since the value for β -HBCD is apparently “no net loss” a value of 10% o f the reported value was assumed

c Assumed d Calculated from the coefficient of variation from the empirical elimination half-life estimates follo wing ( SI

- 4 4,45 )
There are complicating factors with these data othe r 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 a nd HBCD would be subject to firstpass at the liver, which could be the predominant s ite for bioisomerization as well as
biotransformation of the HBCD isomers to metabolite s. 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 ana lysis. Furthermore, the study recognized that steady state was not reac hed.S-36

Drottar and Krueger ( SI

- 46 ) conducted a flow through bioconcentration test fo r t-HBCD at two different exposure concentrations for rainbo w 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 facto rs (BCFs) were 13,000 (not at steady state) and 9,000 L·kg -1 , respectively. This study reports total eliminatio n rate constants of
0.0228 and 0.0359 d -1 for t-HBCD for the low and high levels, respective ly. It is noted that the sponsor of the study did not want the firs t two sample times during the depuration phase (days 1 and 3) analyzed and thus these time p oints were not included in the elimination rate constant estimates. The implicatio ns are unclear. The total uptake and elimination rate constants were calculated by BIOFA C and were used as inputs into the mass balance calculations for estimating k M . The mass balance calculations for primary biotransformation of t-HBCD to metabolites provide k M 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 k M values, both estimates are considered “high” ( SI

- 45 ). The study documents some experimental difficulties maintaining a constant water concentra tion for the low level exposure test. These difficulties may result in errors in the BIOF AC rate constant estimates which are derived by fitting a model to the chemical concentr ation time trends. The biotransformation rate constants from both exposure s correspond to primary
biotransformation half-lives of 260 and 56 d for th e mass of the fish in the study (0.035 kg). Using a rearrangement of Equation 2.5, the cor responding primary biotransformation
half-lives normalized HL N to a 1 kg fish are 1080 and 235 d for t-HBCD for t he 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 biotransformat ion of various brominated flame retardants, including HBCD, in zeb rafish. 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 su ggested by the authors, although bioaccumulation parameters such as elimination rate constants were not reported.S-37

A QSAR for predicting screening level primary biotr ansformation rates in fish estimates

a k M of 0.0071 d -1 for a 1 kg fish ( SI

- 2 ). This rate constant corresponds to a HL N of about
100 d.

In consideration of all of the factors for the biot ransformation data available for fish a
HL N of 250 d is selected for mass balance modelling wi th a Cf of 7. Thus, 95% range of the generated estimates for HL N will be between 35 and 1750 d.

Mammalian Biotransformation Geyer et al. reported total elimination half-life e stimates for rats and humans for t-HBCD ( SI

- 48 ). The data for humans were derived from body burde ns 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 processe s such as respiration, fecal egestion, urinary excretion and biotransformation. The mass b alance method for fish was refined for application to mammalian half-life data assumin g 70 kg for the human and 0.25 kg for
the rat. The estimated HL N (normalized to 1 kg) are 25 d ( Cf = 3.4) and 13 d ( Cf

= 3.2) from the reported human and rat total elimination h alf-life data. A HL N value of 20 d ( Cf = 5) was selected for the mass balance exposure mod elling for t-HBCD. The 95% range of the generated estimates for HL N will be between 4 and 100 d. This HL N value is about 1 order of magnitude faster than the HL N value selected for fish.

In vitro biotransformation studies have been conduc ted 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 vi tro 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 pe ak 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 (i ndicative of metabolites) were onlyS-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 is olated manner. Therefore the possible
biotransformation of α -HBCD by cytochrome P-450 cannot be confirmed or re futed. Although α -HBCD was reduced to a lesser degree some formation of α -HBCD from γ

HBCD (bioisomerization) may have been possible duri ng 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 tot al concentration of HBCD in worm tissue and estimates for the wet weight concentrati on in soil after 28 day of exposures to a range of soil concentrations ( SI

- 4 ). The diastereomer specific biota-soil accumulatio n factor for α -HBCD is more than one order of magnitude higher th an the value for γ

HBCD ( 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 BC F or BSAF study.

Biotransformation Rate Summary
The selected half-life values are based on a critic al review of the limited available data and thus require a degree of expert judgement. It i s noteworthy that the primary transformation half-lives in water, soil, and sedim ent are shorter than the half-lives in fish. While this may be possible, it is counterintu itive, especially with regards to the degree in which the values differ (e.g. aerobic sed iment – 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 u nderestimates. Therefore, some of the mass balance model simulations will explore the pos sibility of half-lives in water, soil and sediment that are greater than median values.

In summary, there is conflicting evidence for diast ereomer-specific biotransformation and bioisomerization rates and pathways based on in vit ro, laboratory and monitoring data.S-39

These data are for different species under differen t conditions and natural variability is expected among and between individuals and species. It is recognized that α -HBCD is usually the predominant diastereoisomer in monitori ng data particularly for upper trophic level species (e.g., marine mammals, birds), but th e precise reason for this cannot be elucidated from the data. It may be a result of pre ferential biotransformation to
metabolites of the β - and γ -HBCD isomers and a limited capacity for biotransfo rmation
for α -HBCD in the food webs. It could also be that biois omerization 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 dias tereomers in storage lipids may also be a factor. The uncerta inty associated with bioisomerization and biotransformation limits the ability to accurat ely model isomer-specific HBCD bioaccumulation and subsequent exposures to organis ms in the environment.



### Section S.7: Additional details on data and approac hes 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 P OV benchmarking exercise (see Section S.3). Arnot et al. (2009) ( SI

- 2 ) present model output based on Characteristic Trav el Distance (CTD), calculated using The OECD Tool ( SI

- 34 ) and Arctic Contamination Potential (eACP 10 ), calculated using GloboPOP ( SI

- 12 ). CTD is the distance at which the modeled air concentration reaches 1/e (i.e. 37%) of the origina l source level. CTD results are
summarized for HBCD and POP and non-POP chemicals i n Figure S.1. eACP 10 is the fraction of total emissions released that is presen t in the Arctic region after 10 years of
constant release. Results for eACP 10 are summarized in Table S.20. In both exercises, there is overlap between POPs and non-POPs, confoun ding the classification of HBCD.S-40

Table S.20. eACP 10 model output for benchmarking exercise for HBCD ( SI

- 2 ).

Status Chemical Scenario eACP 10

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 CCl 4 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 be low. In this example, five hypothetical chemicals with CTD ranging from 100 – 2000 km are i ncluded. The modeled air concentrations at the remote site (C R,O ) assuming an initial air concentration (C O ) of 1 pg/m 3 and a distance of 2000 km from source region to re mote site for each of the five
chemicals are shown in Table S.21. The relative emi ssion strength at the source region required to obtain an identical air concentration i n the remote region (C R,E = 0.37 pg/m 3 ) 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 c hemicals with CTDs between 500– 2000 km is only required to range over approximatel y 1 order of magnitude for identical
air concentrations to be obtained at the remote sit e.S-41

Table S.21. CTD versus emission strength as determinant of ai r concentration in remote regions
Distance to Remote Region 2000 km C O = 1 pg/m 3

CTD (km) C R,O in remote region Relative Emission

Strength at Source C R,E

(pg/m 3 ) 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 1.8x10 8 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 distri bution of emissions are also important factors to consider, especially when comparing moni toring data for a suite of chemicals.S-42



### Section S.8: Summary of monitoring data from near-p oint source, industrialized regions and remote sites

Table S.22. Summary of monitoring data ( Σ HBCD) from locations near known point sources, industrialized regions (no known direct so urces) 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 μ g l -1 0.006 – 140

Soil μ g kg -1 dw 265 – 815

Sediment μ g 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 refl ects reported range in harbour porpoises inhabiting coastal waters of the UK (Law et al. 2006, 2008) – median Σ HBCD 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, Vorka mp 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. Reporte d concentrations of Σ HBCDs inS-43

water, sediment and fish from lakes in England (Ind ustrialized Region) ranged between
0.00008 – 0.00027 μ g l -1 , 0.9 – 4.8 ug kg -1 dw, and 14 – 290 ng g -1 lw, respectively ( SI

30 ). 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 de gradation half-lives reported for HBCD in (aerobic) soils also implies that this comp artment may be important in the longterm as a storage medium and potential secondary so urce. Additional data for terrestrial mammals (wildlife) would also be valuable.

With respect to monitoring data in the Arctic, ther e 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 nar whal 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 ma mmals from Svalbard where the mean HBCD (total) concentra tions 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 t o 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 Ber ing 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 imp act on the western Arctic and ii) historical HBCD inputs to the Western Arctic, likel y emanating from North American and Asia-Pacific sources, are substantially lower t han inputs impacting Svalbard. Hypothesis ii) is consistent with available informa tion on the distribution of global HBCD production and use. Further detail is provide d 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 preparatio n methods, including extraction, clean-up and fractionation has been observed in the last years ( SI

- 53 ). Abdallah et al. ( SI

54 ) has included an in-cell purification using Florisil and Hydromatrix during the PLE. Even using this approach, the extracts had to be fu rther purified with concentrated H 2 SO 4 followed by solid phase extraction (SPE) on Florisi l. PLE-based methods have been used also by Eljarrat et al. ( SI

- 55 ) for the extraction of HBCD isomers from lyophiliz ed human milk and by Fredriksen et al. ( SI

- 56 ) for marine biological samples. Recently, ultrason icassisted extraction (UAE) has also been used for th e 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) h as 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 plasti cs ( 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, H BCD 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-r etarded textiles (e.g. curtains), three different extraction methods (Soxhlet, UAE, a nd soaking extractions with toluene and DCM) were compared ( SI

- 60 ). During Soxhlet extraction using toluene, the per cent 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 isomerisat ion. 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–seco ndary amine was used recently for sewage sludge extracts ( SI

- 57 ). Yet, it has been shown that such cleanup is not sufficientS-45

for complete matrix removal as matrix effects showe d 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 b ased on reverse-phase chromatography with a C 18 -column, in which the α - (and not the β -) isomer elutes first when using a polar solvent. These resu lts indicate that care should be taken when isolating HBCDs and other molecular diastereom ers from environmental and biological samples, and that reported concentration s 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 inter conversion 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 chroma tography 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 enantiosel ective 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 13 C-labelled and 2 H-labelled HBCDs as internal standards to compensat e for potential variations in sensitivity during and betw een sample runs.

A few recent studies have investigated the comparab ility of results issued using GC or LC-separation methods. However, the results were no t conclusive due to a large variation in the methodologies employed in each studies and t he type of samples, fish ( SI

- 63 ) and dust ( SI

- 54 ) for which the comparison was achieved. Although i t was concluded that LCESI-MS-MS results in reliable data and most importa nt individual isomericS-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 st udies 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 o f intercalibration studies have been performed with a rather limited number of participa ting laboratories ( SI

- 64 ). The
laboratories were able to determine total HBCD conc entrations in the marine samples with satisfying quality (RSD <35%). However, the an alysis of samples with (very) low HBCD contamination (< 2 ng/g lipid weight) should b e improved. There are currently no
certified reference materials certified for HBCDs t hat can be used for method validation. Up to now, indicative values for HBCDs have been is sued 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 n ow generally the preferred choice for analysis of HBCD. These methods now find routin e application in many laboratories; some have been further developed to include more co mprehensive 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 ( SI

66,67 ); 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. Interes tingly, 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 resu lt of fragmentation. The formation of chlorine adducts of HBCD, corresponding to ions at m/z 676.6, have been reported ( SI

66-69 ). Adduct formation should be considered as a major factor in quantitation of HBCD; some of the previous studies employed an ammo nium acetate mobile phase modifier to reduce chlorine adduct formation, which can result in formation of an acetateS-47

adduct ( SI

- 69 ). Galindo-Iranzo et al . ( SI

- 69 ) employed two different ammonium salt mobile phase modifiers to encourage (ammonium chlor ide) or inhibit (ammonium acetate) chlorine adduct formation; they found chlo ride adduct formation to result in method LODs more than an order of magnitude lower t han inhibition.

Biological processes, such as metabolism, could res ult 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 is olated 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 sim ply mirror images of each other, can have vastly different biological properties/act ivities 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 enantiose lective methods for analysis of HBCD, and subsequently applied these methods to env ironmental samples. These methods have largely been based on LC/MS/MS techniq ues 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 enantio ner-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 (th e Netherlands), including an apparent enrichment of the (+) α -HBCD enantiomer in whiting and bib liver samples.

Extraction and clean up methods for determination o f 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. Mo st methods use a Nucleodex β -PM (Machery-Nagel, Düren, Germany) chiral stationary p hase 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 Chirad ex and Chirasep stationary phases for separation of HBCD enantiomer s, 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 rec ombination leading to carcinogenicity. This has been the case for results obtained in yeast ( SI

- 74 ), bacteria ( SI

75 ), human peripheral blood lymphocytes ( SI

- 75 ) and in rats treated with high doses of HBCD. Carcinogenesic compounds may also induce thei r 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 im pacts of HBCD on global or gene specific DNA methylation have been observed in fish ( SI

- 75 ) or mammalian models ( SI

76 ).

Additional Thyroid-Disruption Schriks et al. ( SI

- 77 ) also examined the ability of HBCD to interact wit h α and β thyroid hormone receptors using specific reporter gene assa ys. Specifically, the ability of HBCD to affect T3 mediated activation of the two recepto rs in an in vitro test system was examined. Green monkey kidney fibroblasts (CV-1) tr ansfected with the two thyroid receptors linked to luciferase reporter genes were utilized for these studies. While they had seen thyroid potentiation activity for HBCD wit h a T-screen assay previously ( SI

- 78 ), they reported no effect of HBCD on T3 receptor acti vation in this study. The different results from this study may be explained by the fac t that CV-1 cells do not haveS-49

endogenous thyroid receptors, and therefore may als o 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 hormo ne, 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 respon sible for down regulating the deiodinase activity ( SI

- 79 ). However, subsequent work by the same authors sho wed that rainbow trout exposed to HBCD had elevated deiodina se enzyme activity ( SI

- 80 ). Finally, Crump et al. ( SI

- 81 ) found no altered expression in any of the three d eiodinase 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 fracti on 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 se lected F1 rats were challenged with the NK activity assay. This challenge involves inoc ulation 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 BMD L of 0.46 MKD. Van der Ven et al. ( SI

- 82 ) speculate that these effects may also be linked t o thyroidal disruption. Recent experimental evidence has documented lower circulat ing thyroid hormones and reproductive disturbance in an HBCD exposed avian m odel. Captive adult American kestrels ( Falco sparverius ) were exposed to environmentally relevant concentr ations 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 refer ence 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 (ER α , ER β and THR α ) were downregulated in HepG2 human hepatocarcinoma cells incubated with HBCD at 0.3 ng/ml for 72 hour s ( SI

- 76 ). Hamers et al. ( SI

- 84 ) investigated the ability of the technical mixture o f HBCD, as well as the individual diastereoisomers, to affect several receptors. In s tably transfected CALUX® cell lines, antiandrogenic and anitprogestagenic activity was o bserved 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 (progesteron e, testosterone and estradiol) when exposed to a fish liver extract containing high con centrations of HBCD. However, the fish oil extracts also contained relatively high co ncentrations of PCBs, DDT and metabolites, hexachlorocyclobenzene (HCH) and toxap henes, making specific cause and effect relationships for HBCD difficult to draw. Ca ntón 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 t estes 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 mechan istic work is required.

There is strong recent evidence that HBCD affects l ipid metabolism. Canton et al. ( SI

- 88 ) examined gene expression in rats from the 30 and 10 0 mg/kg dose groups from Germer et al. ( SI

- 89 ). HBCD most strongly affected the peroxisome proli ferator-activated receptor (PPAR)-mediated regulation of lipid metabolism, tri acylglycerol metabolism, cholesterolS-51

biosynthesis, androstenedione, testosterone, estrog en and glutathione metabolism. Selected genes from these affected pathways were fu rther 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 hyp othyroid 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 th e HMG-CoA enzyme which catalyzes the conversion of HMG-CoA to mevalonate, a precurso r of cholesterol. Furthermore, they note that a simultaneous increase in serum cho lesterol 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 decrease d in a dose dependent manner with exposure to HBCD. While the function of THSRP1 4-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 downregulate d by HBCD. This gene may be responsive to activation by T3, and therefore, redu ced 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 c oated 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 af fect ATP energy reserves of the NK cells as well as their lytic ability toward tumo ur cells ( SI

- 91 ). Effects were evident even with short term exposure (1 hr), but the lowes t observed effect level was 1 μ M for a decrease in tumour-lysing function after 6 d of exp osure to HBCD.

Reproductive and Hormone Mediated Effects Effects secondarily related to HBCD’s thyroidal imp acts are increasingly being examined. Reproduction, lipid metabolism, nervous s ystem function and development,S-52

and even the process of adapting to a new osmotic e nvironment in salmon smolts ( SI

- 92 ),
have been studied. Several detailed studies have ex amined reproduction in rats treated with HBCD. Early studies found few discernable eff ects 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 repro ductive parameters were reported.
Primordial follicle numbers were reduced in F1 fema les at 1500 ppm and higher. Primordial follicles represent the total female rep roductive 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 capac ity of the affected females was not assessed. No tissue concentrations of HBCD were rep orted, making relationships between reproductive toxicity and tissue burdens im possible.

Developmental toxicity was assessed in rats exposed to HBCD via the diet after maternal exposure from mid-gestation through lactation to da y 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 puber ty to examine reproductive parameters. HBCD exposure did not affect maternal w eight, food consumption, implantations, or duration of the pregnancy. The nu mber of live offspring, their sex ratio and body weights and anogenital distance were also not affected by HBCD. However maternal hypothyroidism was associated with impaire d oligodendroglial cell development in offspring of dams, but only at highe r 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 of fspring from weaning until final necropsy. HBCD accumulated to a greater degree in f emales relative to males across all dose groups (an average of 4-6 fold higher in the h igher dose groups) in the F1 generationS-53

but there were no effects on mating success, gestat ion or mortality of pups. HBCD also decreased bodyweight of F1 in a dose dependent mann er starting at post natal day 4 and continuing until the end of the study (11 wks of ag e). The authors note that this may be related to decreased food intake of the HBCD expose d 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 ci rculating apolar retinoids. The authors note that testis is responsive to androgen antagoni sts and cite a recent study ( SI

- 84 ) showing that HBCD can interact with the progesteron e and androgen receptors. Finally, Van der Ven et al. ( SI

- 82 ) also reported that thyroidal disruption may be li nked 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 vit ro test system these same authors showed that cerebellar granule cell (CGC) death inc reased dose dependently. Cell death was inhibited by the n-methyl D-aspartate (NMDA) re ceptor antagonist MK801. Activation of the NMDA receptor may trigger several downstream pathways leading to calcium influx and apoptosis, ROS formation, and ce llular damage. Inhibited toxicity by MK801 suggests that HBCD induced toxicity through o ne or more of these effects. The involvement of ROS was further implicated because 5 0μM tocopherol in the incubation
medium inhibited HBCD-induced cell death but calciu m free medium and caspase inhibitor had no effect. Nuclear observations sugge sted 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 nervou s cells demonstrated by Reistad et al. ( 97 ) may be manifested in compromised membrane functio n and neurotransmitter release. Dingemans et al. ( SI

- 98 ) found that HBCD affected basal and depolarization

mediated cytosolic calcium levels in an in vitro as say using neurendocrine PC 12 (Rat pheochromocytoma) cells. HBCD applied immediately b efore depolarization had no effect on intracellular Ca2+ making direct interact ion with voltage gated calciumS-54

channels (VGCC) unlikely. Involvement of specific V GCC were investigated by employing blockers (nifedipine, GVIA, or MVIIC x-co notoxins) of the different types of VGCC (L-, N-, or P/Q-type, respectively). If the ef fect of HBCD were specific to a given type of VGCC, then application of HBCD with the inh ibitor of that type of channel should result in no greater effect that the inhibit or 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 e ffect on vesicular catecholamine release, investigated amperometrically using carbon fiber electrodes, in the same cells.

The potential effects of technical-HBCD on rat brai n 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 fo r noncompetitively inhibited the subsequent uptake of do pamine. Glutamate uptake was inhibited at just 1 μM, but inhibition never rose a bove 50%, even at 50 μM. No effect on GABA uptake was elicited with HBCD exposure. To exa mine whether the mechanism of inhibited neurotransmitter uptake was related to an effect on membrane potential, uptake of the marker Tetra[ 3 H]phenylphosphonium bromide (TPP+) was performed after preexposure to HBCD. HBCD only slightly reduced the up take of TPP, which correlated well with its effect on glutamate but not dopamine, suggesting that HBCD elicited its effect only partially by altering membrane potentia l.

Neurotoxic effects of HBCD have also been determine d 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 alt ered brain development. HBCD had no
effect on the distribution of neurons located withi n, 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 c ingulate deep cortex among rats. This is a measure of mature oligodendrocytes, and t he 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 disturban ces with altered auditory function. Rats used in this study were adults of the F1 gener ation 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 move ment. To examine HBCD effect on this process, rats were injected with haloperidol a nd 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 measu red. HBCD decreased the latencies of all three cataleptic movements, and the effect w as more pronounced in females than in males. In males, only foreleg retraction from the b ox resulted in a significant effect. Lilienthal et al. ( SI

- 100 ) suggested that effects were a result of HBCD inhi biting vescicular and membrane transport of dopamine, allo wing dopamine to remain active longer. Alternatively, HBCD could increase clearanc e rates of haloperdidol thereby affecting catalepsis endpoints. Haloperidol is meta bolized 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 sho uld 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 gre ater extent in females ( SI

- 89 ).

One month after the catalepsy experiments, brainste m auditory evoked potentials (BAEPs) were used to evaluate hearing ( SI

- 100 ). For the BAEP measurements, rats were sedated and electodes were implanted to the auditor y brainstem and the response was measured as auditory clicks were issued of differen t 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 thresho ld, HBCD appeared to induce and increased latency period in the lower frequency ran ge. The authors note that the isolated effect on the lower frequency range suggests that H BCD affects the apical cochlea which is known to be dependent on thyroxine for developme nt.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 exhi bited 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 li fe) when the maturation processes of the CNS is particularly vulnerable. In the follow u p 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 mov ement over a sixty minute period. HBCD exposure reduced activity during the initial 2 0 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 receivi ng 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 sig nificantly 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) a re cited by Eriksson et al. (2006).

Finally, linkages between HBCD exposure, thyroid di sruption and brain development were reported using a fish animal model. The effect s of HBCD exposure on
smoltification in Atlantic salmon were examined as there is growing evidence that exposure to chemicals during freshwater phase of sa lmon life can affect smoltification ( SI

- 92 ). Interactions of HBCD with this process are even more relevant considering that the profound osmoregulatory adjustments required du ring the transition from freshwater to saltwater are thought to be regulated by thyroid hormones. Olfactory imprinting responsible for salmon returning to their natal str eams 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 transferr ed to seawater. The peak of thyroxine appeared a week earlier in HBCD exposed f ish 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 olfac tory 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 eli cit effects on the nervous system at the subcellular, cellular, tissue and whole organis m 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 i nduce oxidative stress and apoptosis, and much of this evidence has been derived using fi sh 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 (Z oarces 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. Ba sed 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 polyunsa turated fatty acid (PUFA) lipids, making it susceptible to reactive oxygen mediated d amage. Reactive oxygen species and protein carbonyl content were induced in brain rela tive to untreated animals. Zhang et al. ( SI

- 105 ) concluded that HBCD exposure can induce the gener ation of ROS leading to compromised antioxidant defence status and resultin g oxidation of proteins, lipids and
DNA within nervous tissue.

Zhang et al. ( SI

- 105 ) also evaluated the ability of HBCD to induce oxida tive stress in mammalian cells. The capacity of individual HBCD di asteroisomers and their + or – enantiomers to induce oxidative stress in human hep atoma (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 di fferent assays (MTT, resazurin,S-58

LDH release assay) and by detecting ROS using a flu orescent probe. The assays consistently found greater toxicity and ROS generat ion from exposure to the + entantiomers relative to the – forms, with ROS gene ration being significant at 0.5 μ M and
above for β and γ diastrereoisomers, but only at 5 μ M and above with α -HBCD.S-59Table S.23.Brief descriptions of studies examining effects ofHBCD on thyroid endpoints, reproduction, nerve function and oxidativestress in biological systems where in vivo or in vitro HBCD exposure has been quantified, or where dietary HBCD exposure wasreported. Study numbers are plotted against lowestobserved effect levels in Figure 2.* denotes studies for which exposure was quantifiedas ng ml-1.StudySummaryThyroid disruptionSI-90Van der Ven et al. defined a bench mark dietary dose of 1.6 mg HBCD/kg/day that equated to 43 ug/g liver (w/w) for disruptedthyroid effects. This included lower thyroid weight(most sensitive parameter), lower circulating TT4,and increased T4-UGTactivity in female rats (Effect level=43000 ng/g liver w/w).SI-89Germer et al. studied the same rats as van der Venet al. and showed induction of CYP2B1 mRNA and CYP2B1/2 protein andPROD activity, suggesting phenobarbitol type inducing activity of HBCD. They also showed CYP3A1/3A3 mRNA, protein andactivity induced more in females than males (significant at 3 mg/kg/day which is approximately equal to 920 μg HBCD/g liver(lipid corrected) according to van der Ven et al. derived concentrations. (=920,000 ng/g liver lipid)SI-106Palace et al. observed higher T3 inγ-HBCD group at day 56 and day 14 depuration, lowerT4 inα-,β-,γ-HBCD at day 56 uptakeand day 7 depuration inγ-HBCD. Lower Deiodinase activity in all groups at day 56. Also gland hypertrophy inα-,β-,γ-HBCDfrom days 56 to day 14 of depuration. From Law et al. approximate concentrations of HBCD in muscle (lipid corrected) wereα265.4 ± 97.9 (ng/g lipid weighted),β= 74.1 ± 29.0 ng/g,γ= 48.6 ± 33.4 ng/g. LOEL was for thyroid gland hypertrophy inγgroupat 49 ng/g lipid weighted.SI-83Fernie et al. exposed male kestrels to 54 μg/g in food for 3 weeks prior to mating time had lower TT4and FT4. The sum of HBCDconcentrations in plasma was 3396 ng/g (lipid weighted).SI-84*Hamers et al.: 1 uMα-HBCD increased cell growth in GH3 pituitary cell line (= 641ng/ml not lipid corrected).α-HBCD andβ-HBCD displaced T4 from TTR at IC50 of 12 and 25 μMrespectively.SI107*Yamade-Okabe et al. demonstrated HBCD enhanced thyroid receptor mediated gene expression at 3.12 μM (=1999 ng/ml not lipidcorrected) 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)S-60SI-78*Schriks et al. reportedγ- or TM-HBCD potentiated growth of GH3 rat pituitary cells mediated by T3 receptors at 100 nM (= 64.1ng/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 1μM (=641 ng/ml not lipidweighted)SI-81*Crump et al.: CYP2H1, CYP3A3, UGT1A9 mRNA were upregulated, and TTR and L-FABP mRNA expression weredownregulated byαor 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 in rat H4IIE rat hepatocyte cellline at >1μM (=641 ng/ml not lipid weighted)SI-80Palace et al. Juvenile rainbow trout pre-exposed toα-,β- orγ-HBCD diasteroisomers for 32 days had altered disposition of labelledT4 delivered via gavage (less in thyroid and more in blood).β- andγ-HBCD exposed groups had lower T4 Outer Ring Deiodinaseenzyme actvity. *Assumed similar tissue concentrations to those reported by Palace et al. 2008 (49 ng/g lipid weighted) based onsimilar methodology.SI-96Ema et al.: rats were fed HBCD (0, 150, 1500 or 15000 ppm) for 10 wks premating, through mating, gestation and lactation.Increased thyroid weight in males and females at dietary exposures of≥15000 ppm, increased plasma TSH in F0 and F1 females atdietary exposures of≥1500ppm, and decreased plasma T4 in F0 males and females at dietary exposures of 15000 ppm. LOELreported 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 exposedto 11 ng/L (=0.011 ng/ml) waterborne TM-HBCD for 30 days prior toseawater transfer, then seawater without HBCD for 20 days. Peak of T4 prior to smoltification was earlier, T3 was lower andolfactory responses to conspecific urine was loweramong HBCD treated fish. No tissue residues of HBCDwere reported. Only oneexposure concentration was employed and no tissue residues were reported and there was no lipid weighting.SI-95Saegusa et al.: pregnant female rats were exposed to dietary TM-HBCD from gestational day 10 to 20 days post natal to examineeffects on offspring. Thryoid weight increased at 80-212 mg/kg/d and thryoid follicular cell hypertrophy was identified at 800-2200mg/kg/d in adult females. Among the offspring, Lower circulating T3 was identified at 11 weeks old at80-212 mg/kg/d, andincreased TSH and lower T3 was reported at 800-2200mg/kg/d on post natal day 20. No tissue residues were determined.SI-76*Aniagu et al. exposed human hepatocarcinoma cells (HepG2) to 0.03 or 0.3 ng/ml TM-HBCD for 72 hours. Quantitative real timePCR showed down regulation of thyroid hormone receptorα(THRα) at 0.3 ng/ml (not lipid weighted).S-61SI-95Saegusa et al. (2009). Pregnant females rats exposed to dietary TM-HBCD (0,100, 1000, 10,000 ppm in diet) from gestation day 10until 20 days post weaning. Thyroid weight increased with glandular hypertrophy and increased thyroidstimulating hormone aswell as decreased circulating T3 among 10,000 ppm group (=800 mg/kg/d)Nervous system effectsSI-97*Reistad et al.: TM-HBCD induced death of rat cerebral granule cells (CGC) in culture with an LC50 of 3μM after 24 hours (=1923ng/ml).SI-99*Mariussen and Fonnum showed inhibited uptake of neurotransmitters dopamine (IC50 of 3 μM for HBCD) andglutamate in ratbrain synaptosomes with a lowest observable effectof 1μM for glutamate uptake (=641 ng/ml). Not lipidweighted.SI-98*Dingemans et al.: HBCD inhibited depolarization evoked increase in intracellular Ca2+influx via voltage gated calcium channelsand depolarization-evoked vesciular catecholamine release in rat pheochromocytoma (PC12) cells with alowest observable effect at2 μM (=1282 ng/g). Not lipid weighted.SI-100Lilienthal et al.: elevation of auditory thresholdsin male rats exposed to HBCD corresponding to liver concentrations of 7 μg/g lipidweighted (=Bench Mark Dose Level of 5% for this parameter, = 7000 ng/g liver lipid). In females, haloperidol induced catelepsywas reduced by HBCD at 16 μg/g in liver with a BMDLof 16.2- μg/g liver lipid. This BMDL was set at 20% based on need forclinical intervention at this level of dysfunctionin humans = 16,000ng/g liver lipid.SI-110Eriksson et al. exposed mice to single oral gavageof TM-HBCD (0.9 mg/kg or 13.1 mg/kg) on post natalday 10 and then measuredspontaneous behaviour and water maze learning activity in males at 3 months old. Spontaneous behaviourwas reduced in the first20 minute interval at 0.9 mg/kg (=dietary exposure.Water maze learning was impaired in 13.1 mg/kg group and re-learning wasalso impaired.SI-95Saegusa et al.: pregnant female rats exposed to dietary TM-HBCD (0, 100, 1000, 10000 ppm in diet) fromgestational day 10 to 20days post natal to examine effects on offspring. Male offpspring brains were examined immunohistochemically at post natal week11 had reduced CNPase positive oligodendrocytes inthe cingulate deep cortext at 10,000 ppm ( LOE dose= 800 mg/kg/d).Oxidative StressSI104*Hu et al.: zebrafish (2 hrs postfertilization) exposed to aqueous HBCD (0, 0.002, 0.01, 0.5, 2.5 and 10 mg/L) for 96hrs and thenanalyzed for antioxidant enzyme superoxide dismuatse and the oxidative stress measures, malonaldehyde.MDA increased at≥0.5mg/L but SOD results were mixed. LOEL = 500 ng/mlS-62SI-105Zhang et al. exposed Chinese rareminnows to waterborne 1-500μg/L HBCD for 14, 28 or 42 days and foundincreased reactiveoxygen species and evidence of membrane breakdown (TBARS), protein carbonyl, and inhibited antioxidants SOD and GSH inbrain after 42 days at a LOEL of 10μg/L. Whole bodyHBCD was measured after 42 days and the 10 μg/L exposure equated toapproximately 100 μg/g = 100,000 ng/g)SI105*Zhang et al. exposed HepG2 cells to individual HBCDdiastereoisomers and enantiomers (0.5, 1, 5 or 10μg/ml) for 24 hours andmeasured cell viability by MIT, resazurin assays and LDH leakage as well as reactive oxygen via DCFH-DA assay. Toxicity wasevident at≥0.5 μg/ml (MIT, resazurin) withγ≥β>α. (+) enantiomers were more toxic than (-) or racemic mixtures. LDH was lesssensitive than the other 2 assays. ROS increased at0.5 μg/ml and above with same order of toxicity ofisomers and enatiomers.(LOEL =500 ng/ml)Reproduction and DevelopmentSI-83Fernie et al.: kestrels exposed to TM-HBCD (54 μg/d/ individual) exhibited earlier egg laying, increased numbers of smaller eggs,slower embryonic growth. Male birds had increased testosterone, were heavier, and exhibited less courtship calling (plasma HBCD= 3396 ng/g lipid weighted.SI-96Ema et al.: rats exposed to HBCD in diet (0, 150, 1500 or 15000 ppm) for 10 wks premating, through mating, gestation andlactation had fewer primary folicles in F1 females≥1500 ppm, decreased bodyweight associated with reduced food consumption inF1 males and females at≥15000ppm and more litters lost among F1 parents at≥15000ppm. LOEL was at 1500ppm = 100000ng/g/dSI-90van der Ven et al. Parental generation rats exposedvia food to HBCD premating to weaning. F1 generation had lower trabecularbone mineral density (Bench Mark Dose Level (BMDL)= 0.056 mkd which was equivalent to 1.3 μg HBCD/gof liver lipidweighted), decreased plasma apolar retinoids (BMDL= 1.3 mkd, =79μg/g liver) and potentiation of immune response by NK cells(BMDL=0.46mkd, = 7 μg/g in liver). LOEL for trabecular bone effects at 1.3 μg/g = 1300 ng/g lipid weighted.SI-76*Aniagu et al.: exposed human hepatocarcinoma cells(HepG2) to 0.03 or 0.3 ng/ml TM-HBCD for 72 hours.Quantitative real timePCR showed down regulation of estrogen receptorαandβat 0.3 ng/ml.SI-84*Hamers et al.: using CALUX assays showed antiestrogenic activity ofβ- andγ-HBCD, IC50 at 5-10 μM. TM- andγ-HBCDantagonized 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 and10 mg/L) for 96hrs. Delayedhatching identified at≥2 ng/ml.S-63SI-111Japanese MOE Study #08-016 (2009). Listed in AnnexE for Japan in the call for more information by Stockholm Convention.Sixty pairs of Japanese quail exposed to 5, 15, 45or 125 ppm HBCD for 6 wks. Hatched eggs reared to 14 days and evaluated.Death increased and overall survival was lower at≥15ppm (=2.1 mkd). Reproductive ability index was lower at same level after 6wks 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 96hrs. Survival and heart ratedecreased, and apoptosis elevated at 0.5 mg/L. LOEL= 50,000ng/ml.S-64 Literature Cited
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