[Federal Register Volume 81, Number 106 (Thursday, June 2, 2016)]
[Proposed Rules]
[Pages 35275-35290]
From the Federal Register Online via the Government Publishing Office [www.gpo.gov]
[FR Doc No: 2016-12464]


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ENVIRONMENTAL PROTECTION AGENCY

40 CFR Part 372

[EPA-HQ-TRI-2015-0607; FRL-9943-55]
RIN 2025-AA42


Addition of Hexabromocyclododecane (HBCD) Category; Community 
Right-to-Know Toxic Chemical Release Reporting

AGENCY: Environmental Protection Agency (EPA).

ACTION: Proposed rule.

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SUMMARY: EPA is proposing to add a hexabromocyclododecane (HBCD) 
category to the list of toxic chemicals subject to reporting under 
section 313 of the Emergency Planning and Community Right-to-Know Act 
(EPCRA) and section 6607 of the Pollution Prevention Act (PPA). EPA is 
proposing to add this chemical category to the EPCRA section 313 list 
because EPA believes HBCD meets the EPCRA section 313(d)(2)(B) and (C) 
toxicity criteria. Specifically, EPA believes that HBCD can reasonably 
be anticipated to cause developmental and reproductive effects in 
humans and is highly toxic to aquatic and terrestrial organisms. In 
addition, based on the available bioaccumulation and persistence data, 
EPA believes that HBCD should be classified as a persistent, 
bioaccumulative, and toxic (PBT) chemical and assigned a 100-pound 
reporting threshold. Based on a review of the available production and 
use information, members of the HBCD category are expected to be 
manufactured, processed, or otherwise used in quantities that would 
exceed a 100-pound EPCRA section 313 reporting threshold.

DATES: Comments must be received on or before August 1, 2016.

ADDRESSES: Submit your comments, identified by Docket ID No. EPA-HQ-
TRI-2015-0607, by one of the following methods:
     Federal eRulemaking Portal: http://www.regulations.gov. 
Follow the online instructions for submitting comments. Do not submit 
electronically any information you consider to be Confidential Business 
Information (CBI) or other information whose disclosure is restricted 
by statute.
     Mail: Document Control Office (7407M), Office of Pollution 
Prevention and Toxics (OPPT), Environmental Protection Agency, 1200 
Pennsylvania Ave. NW., Washington, DC 20460-0001.
     Hand Delivery: To make special arrangements for hand 
delivery or delivery of boxed information, please follow the 
instructions at http://www.epa.gov/dockets/where-send-comments-epa-dockets#hq.

Additional instructions on commenting or visiting the docket, along 
with more information about dockets generally, is available at http://www.epa.gov/dockets/commenting-epa-dockets.

[[Page 35276]]


FOR FURTHER INFORMATION CONTACT: 
    For technical information contact: Daniel R. Bushman, Toxics 
Release Inventory Program Division (7409M), Office of Pollution 
Prevention and Toxics, Environmental Protection Agency, 1200 
Pennsylvania Ave. NW., Washington, DC 20460-0001; telephone number: 
(202) 566-0743; email: [email protected].
    For general information contact: The Emergency Planning and 
Community Right-to-Know Hotline; telephone numbers: toll free at (800) 
424-9346 (select menu option 3) or (703) 412-9810 in Virginia and 
Alaska; or toll free, TDD (800) 553-7672; or go to http://www.epa.gov/superfund/contacts/infocenter/.

SUPPLEMENTARY INFORMATION:

I. General Information

A. Does this notice apply to me?

    You may be potentially affected by this action if you manufacture, 
process, or otherwise use HBCD. The following list of North American 
Industrial Classification System (NAICS) codes is not intended to be 
exhaustive, but rather provides a guide to help readers determine 
whether this document applies to them. Potentially affected entities 
may include:
     Facilities included in the following NAICS manufacturing 
codes (corresponding to Standard Industrial Classification (SIC) codes 
20 through 39): 311*, 312*, 313*, 314*, 315*, 316, 321, 322, 323*, 324, 
325*, 326*, 327, 331, 332, 333, 334*, 335*, 336, 337*, 339*, 111998*, 
211112*, 212324*, 212325*, 212393*, 212399*, 488390*, 511110, 511120, 
511130, 511140*, 511191, 511199, 512220, 512230*, 519130*, 541712*, or 
811490*.

* Exceptions and/or limitations exist for these NAICS codes.
     Facilities included in the following NAICS codes 
(corresponding to SIC codes other than SIC codes 20 through 39): 
212111, 212112, 212113 (corresponds to SIC code 12, Coal Mining (except 
1241)); or 212221, 212222, 212231, 212234, 212299 (corresponds to SIC 
code 10, Metal Mining (except 1011, 1081, and 1094)); or 221111, 
221112, 221113, 221118, 221121, 221122, 221330 (Limited to facilities 
that combust coal and/or oil for the purpose of generating power for 
distribution in commerce) (corresponds to SIC codes 4911, 4931, and 
4939, Electric Utilities); or 424690, 425110, 425120 (Limited to 
facilities previously classified in SIC code 5169, Chemicals and Allied 
Products, Not Elsewhere Classified); or 424710 (corresponds to SIC code 
5171, Petroleum Bulk Terminals and Plants); or 562112 (Limited to 
facilities primarily engaged in solvent recovery services on a contract 
or fee basis (previously classified under SIC code 7389, Business 
Services, NEC)); or 562211, 562212, 562213, 562219, 562920 (Limited to 
facilities regulated under the Resource Conservation and Recovery Act, 
subtitle C, 42 U.S.C. 6921 et seq.) (corresponds to SIC code 4953, 
Refuse Systems).
     Federal facilities.
    To determine whether your facility would be affected by this 
action, you should carefully examine the applicability criteria in part 
372, subpart B of Title 40 of the Code of Federal Regulations. If you 
have questions regarding the applicability of this action to a 
particular entity, consult the person listed under FOR FURTHER 
INFORMATION CONTACT.

B. What action is the Agency taking?

    EPA is proposing to add a hexabromocyclododecane (HBCD) category to 
the list of toxic chemicals subject to reporting under EPCRA section 
313 and PPA section 6607. As discussed in more detail later in this 
document, EPA is proposing to add this chemical category to the EPCRA 
section 313 list because EPA believes HBCD meets the EPCRA section 
313(d)(2)(B) and (C) toxicity criteria.

C. What is the Agency's authority for taking this action?

    This action is issued under EPCRA sections 313(d) and 328, 42 
U.S.C. 11023 et seq., and PPA section 6607, 42 U.S.C. 13106. EPCRA is 
also referred to as Title III of the Superfund Amendments and 
Reauthorization Act of 1986.
    Section 313 of EPCRA, 42 U.S.C. 11023, requires certain facilities 
that manufacture, process, or otherwise use listed toxic chemicals in 
amounts above reporting threshold levels to report their environmental 
releases and other waste management quantities of such chemicals 
annually. These facilities must also report pollution prevention and 
recycling data for such chemicals, pursuant to section 6607 of the PPA, 
42 U.S.C. 13106. Congress established an initial list of toxic 
chemicals that comprised 308 individually listed chemicals and 20 
chemical categories.
    EPCRA section 313(d) authorizes EPA to add or delete chemicals from 
the list and sets criteria for these actions. EPCRA section 313(d)(2) 
states that EPA may add a chemical to the list if any of the listing 
criteria in EPCRA section 313(d)(2) are met. Therefore, to add a 
chemical, EPA must demonstrate that at least one criterion is met, but 
need not determine whether any other criterion is met. Conversely, to 
remove a chemical from the list, EPCRA section 313(d)(3) dictates that 
EPA must demonstrate that none of the following listing criteria in 
EPCRA section 313(d)(2)(A)-(C) are met:
     The chemical is known to cause or can reasonably be 
anticipated to cause significant adverse acute human health effects at 
concentration levels that are reasonably likely to exist beyond 
facility site boundaries as a result of continuous, or frequently 
recurring, releases.
     The chemical is known to cause or can reasonably be 
anticipated to cause in humans: Cancer or teratogenic effects, or 
serious or irreversible reproductive dysfunctions, neurological 
disorders, heritable genetic mutations, or other chronic health 
effects.
     The chemical is known to cause or can be reasonably 
anticipated to cause, because of its toxicity, its toxicity and 
persistence in the environment, or its toxicity and tendency to 
bioaccumulate in the environment, a significant adverse effect on the 
environment of sufficient seriousness, in the judgment of the 
Administrator, to warrant reporting under this section.
    EPA often refers to the EPCRA section 313(d)(2)(A) criterion as the 
``acute human health effects criterion;'' the EPCRA section 
313(d)(2)(B) criterion as the ``chronic human health effects 
criterion;'' and the EPCRA section 313(d)(2)(C) criterion as the 
``environmental effects criterion.''
    EPA published in the Federal Register of November 30, 1994 (59 FR 
61432) (FRL-4922-2), a statement clarifying its interpretation of the 
EPCRA section 313(d)(2) and (d)(3) criteria for modifying the EPCRA 
section 313 list of toxic chemicals.

II. Background Information

A. What is HBCD?

    HBCD is a cyclic aliphatic hydrocarbon consisting of a 12-membered 
carbon ring with 6 bromine atoms attached (molecular formula 
C12H18Br6). HBCD has 16 possible 
stereoisomers. Technical grades of HBCD consist predominantly of three 
diastereomers, [alpha]-, [szlig]- and [gamma]-HBCD (Ref. 1). HBCD may 
be designated as a non-specific mixture of all isomers 
(hexabromocyclododecane, Chemical Abstracts Service Registry Number 
(CASRN) 25637-99-4) or as a mixture of the three main diastereomers 
(1,2,5,6,9,10-hexabromocyclododecane, CASRN 3194-55-6) (Ref 1). The 
main use of HBCD is as a flame retardant in expanded polystyrene foam 
(EPS) and

[[Page 35277]]

extruded polystyrene foam (XPS) (Ref. 2). EPS and XPS are used 
primarily for thermal insulation boards in the building and 
construction industry. HBCD may also be used as a flame retardant in 
textiles including: upholstered furniture, upholstery seating in 
transportation vehicles, draperies, wall coverings, mattress ticking, 
and interior textiles, such as roller blinds (Ref. 2). In addition, 
HBCD is used as a flame retardant in high-impact polystyrene for 
electrical and electronic appliances such as audio-visual equipment, as 
well as for some wire and cable applications (Ref. 2).
    Concerns for releases and uses of HBCD have been raised because it 
is found world-wide in the environment and wildlife and has also been 
found in human breast milk, adipose tissue and blood (Ref. 1). HBCD is 
known to bioaccumulate and biomagnify in the food chain and has been 
detected over large areas and in remote locations in environmental 
monitoring studies (Ref. 1).

B. How is EPA proposing to list HBCD under EPCRA section 313?

    HBCD is identified through two primary CASRNs 3194-55-6 
(1,2,5,6,9,10-hexabromocyclododecane) and 25637-99-4 
(hexabromocyclododecane) (Ref. 1). EPA is proposing to create an HBCD 
category that would cover these two chemical names and CASRNs. The HBCD 
category would be defined as: Hexabromocyclododecane and would only 
include those chemicals covered by the following CAS numbers:
     3194-55-6; 1,2,5,6,9,10-Hexabromocyclododecane.
     25637-99-4; Hexabromocyclododecane.

As a category, facilities that manufacture, process or otherwise use 
HBCD covered under both of these names and CASRNs would file just one 
report.
    In addition to listing HBCD as a category, EPA is proposing to add 
the HBCD category to the list of chemicals of special concern. There 
are several chemicals and chemical categories on the EPCRA section 313 
chemical list that have been classified as chemicals of special concern 
because they are PBT chemicals (see 40 CFR 372.28(a)(2)). In a final 
rule published in the Federal Register of October 29, 1999 (64 FR 
58666) (FRL-6389-11), EPA established the PBT classification criteria 
for chemicals on the EPCRA section 313 chemical list. For purposes of 
EPCRA section 313 reporting, EPA established persistence half-life 
criteria for PBT chemicals of 2 months in water/sediment and soil and 2 
days in air, and established bioaccumulation criteria for PBT chemicals 
as a bioconcentration factor (BCF) or bioaccumulation factor (BAF) of 
1,000 or higher. Chemicals meeting the PBT criteria were assigned 100-
pound reporting thresholds. With regards to setting the EPCRA section 
313 reporting thresholds, EPA set lower reporting thresholds (10 
pounds) for those PBT chemicals with persistence half-lives of 6 months 
or more in water/sediment or soil and with BCF or BAF values of 5,000 
or higher, these chemicals were considered highly PBT chemicals. The 
data presented in this proposed rule support classifying the HBCD 
category as a PBT chemical category with a 100-pound reporting 
threshold.

III. What is EPA's evaluation of the toxicity, bioaccumulation, and 
environmental persistence of HBCD?

    EPA evaluated the available literature on the human health 
toxicity, ecological toxicity, bioaccumulation potential, and 
environmental persistence of HBCD (Ref. 1). Unit III.A. provides a 
review of the human health toxicity studies and EPA's conclusions 
regarding the human health hazard potential of HBCD. Unit III.B. 
discusses the ecological toxicity of HBCD, Unit III.C. contains 
information on the bioaccumulation potential of HBCD, and Unit III.D. 
provides information on the environmental persistence of HBCD.

A. What is EPA's review of the human health toxicity data for HBCD?

    1. Toxicokinetics. HBCD is absorbed via the gastrointestinal tract 
and metabolized in rodents (Refs. 3, 4, 5, and 6). Once absorbed, HBCD 
is distributed to a number of tissues, including fatty tissue, muscle, 
and the liver (Refs. 7, 8, 9, 10, 11, and 12). Elimination of HBCD is 
predominantly via feces (as the parent compound), but it is also 
eliminated in urine (as secondary metabolites) (Refs. 3, 4, and 5). 
HBCD has been detected in human milk, adipose tissue, and blood (Refs. 
13, 14, 15, 16, 17, 18, 19, 20, 21, 22, 23, and 24). The composition of 
HBCD isomers in most rodent toxicity studies resembles that of 
industrial grade HBCD, which may differ from human exposure to certain 
foods that have been shown to contain elevated fractions of [alpha]-
HBCD (Ref. 25).
    2. Effects of acute exposure. HBCD was not found to be highly toxic 
in acute oral, inhalation, and dermal studies in rodents. One study 
reported an oral median lethal dose (LD50) of >10,000 
milligrams per kilogram (mg/kg) in Charles River rats (Ref. 26). 
Another study by the same researchers, however, reported an 
LD50 of 680 mg/kg for females and 1,258 mg/kg for males in 
Charles River CD rats (Ref. 27). Two other studies reported an oral 
LD50 of >5,000 mg/kg in Sprague-Dawley rats and >10,000 mg/
kg in NR rats (Refs. 28 and 29). An oral study in NR mice reported an 
LD50 of >6,400 mg/kg (Ref. 30). Acute inhalation studies in 
rats have generally concluded that HCBD is not highly toxic, with a 
median lethal concentration (LC50) reported by Gulf South 
Research Institute of >200 milligrams per liter (mg/L) (Refs. 26, 27, 
29, 31). Acute dermal toxicity studies have generally shown HBCD not to 
be highly toxic in rabbits (Refs. 27, 29, 31, and 32). One dermal study 
reported an LD50 of 3,969 mg/kg (Ref. 27). Additionally, 
HBCD is not a dermal irritant in rabbits (Refs. 27, 29, and 31), but it 
is a mild skin allergen in guinea pigs (Ref. 32). Acute eye irritation 
studies have concluded that HBCD is a primary eye irritant (Ref. 27) 
and a mild, transient ocular irritant (Ref. 29).
    3. Effects of short-term and subchronic exposure. In subacute and 
subchronic studies, HBCD demonstrated effects on the thyroid and liver 
(Refs. 8, 33, 34, and 35). In a subacute study, van der Ven et al. 
(Ref. 8) exposed Wistar rats (5/sex/dose) by gavage to a mixture of 
HBCD dissolved in corn oil at concentrations resulting in doses of 0.3, 
1.0, 3.0, 10, 30, 100, and 200 milligrams per kilogram per day (mg/kg/
day) for 28 days. The isomeric composition of the HBCD was 10.3% 
[alpha], 8.7% [beta], and 81.0% [gamma]. The authors reported a 
benchmark dose lower bound confidence limit (BMDL) of 29.9 mg/kg/day 
for an increase in pituitary weight, a BMDL of 1.6 mg/kg/day for an 
increase in thyroid weight, and a BMDL of 22.9 mg/kg/day for an 
increase in liver weight. The increase in thyroid weight was the most 
sensitive end point observed and, according to research by EPA, is 
considered relevant to humans (Ref. 36). Additionally, histopathology 
of the thyroid demonstrated that thyroid follicles were smaller, 
depleted, and had hypertrophied epithelium in female rats.
    In another subacute study, HBCD was administered orally by gavage 
in corn oil to Sprague-Dawley Crl:CD BR rats for 28 days at doses of 0, 
125, 350, or 1,000 mg/kg/day (6 rats/sex/dose in 125 and 350 mg/kg/day 
groups and 12 rats/sex/dose in the control and 1,000 mg/kg/day groups) 
(Ref. 33). At the end of 28 days, 6 rats/sex/dose were necropsied, 
while the remaining rats in the control and 1,000 mg/kg/day groups were 
untreated for a 14-day recovery period prior to necropsy. The authors 
reported

[[Page 35278]]

increased absolute and liver to body weight ratios in females, but the 
authors considered the findings to be adaptive and not adverse. This 
study also identified a no-observed-adverse-effect level (NOAEL) of 
1,000 mg/kg/day.
    In an older subacute study (Ref. 37), an HBCD product was 
administered to Sprague-Dawley rat (10/sex/group) at doses of 0, 1, 
2.5, and 5% of the diet for 28 days. Doses were calculated to be 0, 
940, 2,410, 4,820 mg/kg/day. Mean liver weight (both absolute and 
relative) was increased in all dose groups, but no microscopic 
pathology was detected. Thyroid hyperplasia was observed in some 
animals at all doses in addition to slight numerical development of the 
follicles and ripening follicles in the ovaries at the high dose. The 
authors concluded that these observed effects were not pathologic and 
reported a NOAEL of 940 mg/kg/day (Ref. 37).
    In a subchronic study, Chengelis (Refs. 34 and 35) administered 
HBCD by oral gavage in corn oil daily to Crl:CD(SD)IGS BR rats (15/sex/
dose) at dose levels of 0, 100, 300, or 1,000 mg/kg/day for 90 days. At 
the end of 90 days, 10 rats/sex/dose were necropsied, while the 
remaining rats were untreated for a 28-day recovery period prior to 
necropsy. The authors reported significant treatment-related changes in 
rats, including decreased liver weight and histopathological changes, 
but the authors considered these changes mild, reversible, and 
adaptive. Decreased liver weight accompanied by the observed 
histopathological changes, however, can be considered an adverse 
effect. Therefore, EPA identified a lowest-observed-adverse-effect 
level (LOAEL) of 100 mg/kg/day based on these changes.
    In an older subchronic study (Ref. 38) an HBCD product was 
administered to Sprague-Dawley rats (10/sex/group) at doses of 0, 0.16, 
0.32, 0.64, and 1.28% of the diet for 90 days. Doses were calculated to 
be 0, 120, 240, 470, and 950 mg/kg/day. An increase in relative liver 
weight was observed and was accompanied by fatty accumulation. The 
pathology report concluded that although fat was visible 
microscopically in treated rats, the change was not accompanied by any 
pathology, and therefore could not be defined as ``fatty liver.'' No 
histological changes were found in any other organ. The authors 
concluded that the increased liver weight and the fat deposits, both of 
which were largely reversible when administration of HBCD was stopped, 
were the result of a temporary increase in the activity of the liver. 
They identified a NOAEL of 950 mg/kg/day.
    4. Carcinogenicity. No adequate studies were found evaluating the 
carcinogenicity of HBCD in animals or humans. One non-guideline study 
(Ref. 39) was cited in the U.S. EPA's Flame Retardant Alternatives for 
Hexabromocyclododecane (HBCD): Final Report (Ref. 40), but this study 
was not adequate to draw conclusions regarding carcinogenicity.
    5. Developmental and reproductive toxicity. The developmental and 
reproductive toxicity of HBCD have been investigated in several 
studies. In a 1-generation study that included additional 
immunological, endocrine and neurodevelopmental endpoints, van der Ven 
et al. (Ref. 9) exposed Wistar rats (10/sex/dose) to a composite 
mixture of technical-grade HBCD (10.3% [alpha], 8.7% [beta], and 81.0% 
[gamma]) in the diet at concentrations resulting in doses of 0.1, 0.3, 
1.0, 3.0, 10, 30, or 100 mg/kg/day. In the highest dose group (100 mg/
kg/day) body weight decreases of 7-36% in males and 10-20% in females 
were observed in first generation (F1) pups. The authors observed 
decreases in kidney and thymus weight in both F1 males and females. 
Decreases in testes, adrenal, prostate, heart, and brain weights in F1 
males were also observed. No histopathological changes, however, were 
observed in any of these organs. Other developmental effects were 
observed, including: Immune system effects, indications of liver 
toxicity, and decreases in bone mineral density at very low doses 
(i.e., <1.3 mg/kg/day). The authors noted that the vehicle used (corn 
oil) may have affected some observations at higher doses, including: 
Increased mortality during lactation, decreased liver weight in males, 
decreased adrenal weight in females, decreased plasma cholesterol in 
females, and other immunological markers of toxicity. Increased 
anogenital distance was observed in males at 100 mg/kg on postnatal day 
(PND) 4, but not on PND 7 or 21. There was no effect on preputial 
separation. The time to vaginal opening was delayed in females at the 
100 mg/kg dose. There were no effects of HBCD exposure on thyroid 
hormones triiodothyronine (T3) and thyroxine (T4) in either the 
parental or F1 animals. There were no effects on thyroid weight or 
thyroid pathology in the F1 animals (parents were not examined). The 
most sensitive endpoints with valid benchmark dose (BMD)/BMDL ratios 
for female rats were decreased bone mineral density with a BMDL of 
0.056 mg/kg/day (BMD of 0.18 mg/kg/day) at a benchmark response (BMR) 
of 10% and decreased concentrations of apolar retinoids in the liver 
with a BMDL of 1.3 mg/kg/day (BMD = 5.1 mg/kg/day) at a BMR of 10%. The 
most sensitive endpoint with a valid BMD/BMDL ratio for male rats was 
an increased IgG response to sheep red blood cells with a BMDL of 0.46 
mg/kg/day (BMD = 1.45 mg/kg/day) at a BMR of 20%. There were no 
significant effects of HBCD exposure on any measure of reproduction, 
including: Mating success, time to gestation, duration of gestation, 
number of implantation sites, pup mortality (at birth and throughout 
lactation), or sex ratios within a litter. Therefore, a BMDL for 
reproductive toxicity could not be derived for this study.
    Saegusa et al. (Ref. 41) exposed pregnant Sprague-Dawley rats (10/
sex/dose) to HBCD from gestation day 10 until PND 20 at dietary 
concentrations of 0, 100, 1,000, or 10,000 parts per million (ppm) in a 
soy-free diet. The authors observed increased relative thyroid weight 
and decreased T3 levels in F1 male Sprague-Dawley rats at postnatal 
week (PNW) 11 following dietary exposure to 1,000 ppm (approximately 
146.3 mg/kg/day) HBCD. The authors also reported a significant 
reduction in the number of CNPase-positive oligodendrocytes at 10,000 
ppm (approximately 1,504.8 mg/kg/day). EPA identified a maternal LOAEL 
of 10,000 ppm (about 1,504.8 mg/kg/day) based on increased incidence of 
thyroid follicular cell hypertrophy, and a developmental LOAEL of 1,000 
ppm (about 146.3 mg/kg/day) based on increased relative thyroid weight 
and decreased T3 levels in F1 males at PNW 11. Changes in reproductive 
endpoints (e.g., the number of implantation sites, live offspring, sex 
ratio) were not observed. Therefore, a LOAEL for reproductive toxicity 
could not be determined for this study.
    Ema et al. (Ref. 42) administered HBCD to groups of male and female 
Crl:CD(SD) rats (24/sex/dose, as a mixture of [alpha]-HBCD, [beta] -
HBCD, and [gamma]-HBCD with proportions of 8.5, 7.9, and 83.7%, 
respectively) in the diet at concentrations of 0, 150, 1,500, or 15,000 
ppm from 10 weeks prior to mating through mating, gestation, and 
lactation. The authors reported a decrease in the number of primordial 
follicles in F1 female rats at 1,500 ppm (approximately 138 mg/kg/day) 
and a significant increase in the number of litters lost in the F1 
generation at 15,000 ppm (approximately 1,363 mg/kg/day). These authors 
reported no other significant treatment-related effects in any 
generation for indicators of reproductive health, including: Estrous 
cyclicity, sperm count and morphology, copulation index, fertility 
index,

[[Page 35279]]

gestation index, delivery index, gestation length, number of pups 
delivered, number of litters, or sex ratios. The authors reported a 
reduced viability index on day 4 and day 21 of lactation among second 
generation (F2) offspring at 15,000 ppm (approximately 1,363 mg/kg/
day). They observed additional developmental effects at doses as low as 
1,500 ppm (approximately 115 and 138 mg/kg/day for F1 males and 
females, respectively), including: An increase in dihydrotestosterone 
(DHT) in F1 males and an increased incidence of animals with decreased 
thyroid follicle size in both sexes and generations. These authors 
reported no effects on sexual development indicated by anogenital 
distance, vaginal opening, or preputial separation among F1 or F2 
generations. The percentage of pups with completed eye opening on PND 
14 was significantly decreased compared to controls in F2 females at 
1,500 ppm and in F2 males and females at 15,000 ppm. Fewer F2 females 
exposed to 15,000 ppm HBCD completed the mid-air righting reflex 
(76.9%) than control F2 females (100%). These findings were not 
consistent over generations or sexes and were not considered treatment 
related. No other effects of HBCD exposure on the development of 
reflexes were observed in either F1 or F2 progeny. EPA identified a 
maternal LOAEL of 150 ppm (about 14 mg/kg/day) based on increased 
thyroid-stimulating hormone (TSH). A reproductive LOAEL of 1,500 ppm 
(about 138 mg/kg/day) was identified based on a decreased number of 
primordial follicles in the ovary observed in F1 females. A 
developmental LOAEL of 15,000 ppm (about 1,142 mg/kg/day for males and 
1,363 mg/kg/day for females) was identified based on increased pup 
mortality during lactation in the F2 generation.
    Murai et al. (Ref. 43) fed female Wistar rats HBCD in the diet at 
concentrations of 0, 0.01, 0.1, or 1% throughout gestation (Days 0-20). 
Dams in the high-dose group demonstrated a statistically significant 
decrease (8.4%) in food consumption and increase in liver weight (13%) 
in comparison with controls. There were no treatment-related effects on 
maternal or fetal body weight. There were no effects on the number of 
implants; number of resorbed, dead, or live fetuses; body weight of 
live fetuses; or incidence of external or visceral abnormalities. A few 
skeletal variations were present but were also observed in controls and 
not considered significant. There were no effects on weaning or 
survival. The European Commission (Ref. 44) used the study's data to 
calculate the doses to be 0, 7.5, 75, and 750 mg/kg/day (based on the 
assumption of a mean animal weight of 200 grams (g) and food 
consumption of 15 g/day). They concluded that the offspring NOAEL was 
750 mg/kg/day and the maternal LOAEL was 750 mg/kg/day based on a 13% 
liver weight increase in the high dose group.
    Eriksson et al. (Ref. 45) conducted a study that examined behavior, 
learning, and memory in adult mice following exposure to HBCD on PND 
10. The authors administered a single oral dose of HBCD (mixture of, 
[alpha]-, [beta]-, and [gamma]-diastereoisomers) dissolved in a fat 
emulsion at 0, 0.9, or 13.5 mg/kg/day on PND 10 to male and female NMRI 
mice. The authors concluded that exposure on PND 10 affected 
spontaneous motor behavior, learning, and memory in adult mice in a 
dose-dependent manner. The authors identified the lowest exposure 
level, 0.9 mg/kg, as the LOAEL based on significantly reduced mean 
locomotor activity compared with controls during the first 20-minute 
interval of testing. EPA, however, identified a LOAEL of 13.5 mg/kg/day 
based on decreased habituation, locomotion, and rearing during all 
intervals. This study was not conducted according to current guidelines 
(Ref. 46) and Good Laboratory Practices; therefore, EPA reserves 
judgment on the significance of these findings.
    6. Genotoxicity. A limited number of studies investigated the 
genotoxicity of HBCD. These studies indicate that HBCD is not likely to 
be genotoxic (Refs. 47, 48, 49, 50, 51, 52, 53, and 54).
    7. Conclusions regarding the human hazard potential of HBCD. The 
available evidence indicates that HBCD has the potential to cause 
developmental and reproductive toxicity at moderately low to low doses. 
While there were some indications of liver toxicity in some short-term 
and subchronic studies, the evidence for these effects is not 
sufficient to support listing. The available evidence for developmental 
and reproductive toxicity, however, is sufficient to conclude that HBCD 
can be reasonably anticipated to cause moderately high to high chronic 
toxicity in humans based on the EPCRA section 313 listing criteria 
published in the Federal Register of November 30, 1994 (59 FR 61432) 
(FRL-4922-2).

B. What is EPA's review of the ecological toxicity of HBCD?

    HBCD can cause effects on survival, growth, reproduction, 
development, and behavior in aquatic and terrestrial species. Observed 
acute toxicity values as low as 0.009 mg/L for a 72-hour 
EC50 (i.e., the concentration that is effective in producing 
a sublethal response in 50% of test organisms) based on reduced growth 
in the marine algae Skeletonema costatum (Ref. 55) indicate high acute 
aquatic toxicity. Observed chronic aquatic toxicity values as low as 
0.0042 mg/L (maximum acceptable toxicant concentration (MATC)) for 
reduced size (length) of surviving young in water fleas (Daphnia magna) 
(Ref. 56) indicate high chronic aquatic toxicity. Reduced chick 
survival in Japanese quails (Coturnix coturnix japonica) fed a 15 parts 
per million (ppm) HBCD diet (2.1 mg/kg/day) (Ref. 57 as cited in Ref. 
58) and altered reproductive behavior (reduced courtship and brood-
rearing activity) and reduced egg size in American kestrels (Falco 
sparverius) fed 0.51 mg/kg/day (Refs. 59, 60, 61, and 62) indicate high 
toxicity to terrestrial species as well.
    Assessment of HBCD's aquatic toxicity is complicated by its low 
water solubility and differences in the solubility of the three main 
HBCD isomers, which makes testing difficult and interpretation 
uncertain for studies conducted above the water solubility. Studies 
conducted at concentrations above the water solubility of HBCD are 
essentially testing the effects at the maximum HBCD concentration 
possible. In some acute and chronic aquatic toxicity studies conducted 
using methods, test species, and endpoints recommended by EPA, no 
effects were reported at or near the limit of water solubility. 
However, water solubility is not considered a limiting factor for 
hazard determination for aquatic species since there are studies 
showing adverse effects at or below the water solubility of HBCD. In 
addition, the potential for HBCD to bioaccumulate, biomagnify, and 
persist in the environment, significantly increases concerns for 
effects on aquatic organisms.
    A wide range of effects of HBCD have been reported in fish (e.g., 
developmental toxicity, embryo malformations, reduced hatching success, 
reduced growth, hepatic enzyme and biomarker effects, thyroid effects, 
deoxyribonucleic acid (DNA) damage to erythrocytes, and oxidative 
damage) and in invertebrates (e.g., degenerative changes, morphological 
abnormalities, decreased hatching success, and altered enzyme activity) 
(Refs. 63, 64, 65, 66, 67, 68, 69, 70, 71, 72, 73, and 74). Reduced 
thyroid hormone (triiodothyronine, T3, and thyroxine, T4) levels in 
rainbow trout (Oncorhynchus mykiss) (Refs. 68 and 69), are similar to 
those observed in mammals. Reduced T4 levels were also

[[Page 35280]]

reported in birds exposed to HBCD (Ref. 61).
    1. Acute aquatic toxicity. Adverse effects observed following acute 
exposure were found in studies with marine algae, including EPA-
recommended estuarine/marine algae species Skeletonema costatum (Ref. 
75 as cited in Refs. 44 and 76, Refs. 55 and 77), a series of short-
term (72 to 120-hour) early life stage tests with zebrafish (Danio 
rerio) embryos (Refs. 64, 65, 67, and 72), and short-term (72-hour) 
results from an early life stage test with sea urchin embryos (Ref. 
63). Effects in these studies, reported at concentrations as low as 
0.009 mg/L (measured) in algae, 0.01 mg/L (nominal) in zebrafish 
embryos, and 0.064 mg/L (nominal) in sea urchin embryos, indicate high 
acute toxicity. Walsh et al. (Ref. 55) reported measured 72-hour 
EC50 values in Skeletonema costatum ranging from 0.009 to 
0.012 mg/L based on reduced growth rate in five different types of 
saltwater media (0.010 mg/L in seawater itself). The study tested two 
other marine algal species, Chlorella sp. and Thalassiosira pseudonana, 
that were also found to be inhibited by HBCD, albeit at higher 
concentrations than Skeletonema costatum. EC50 values for 
reduced growth in these species were 0.05-0.37 mg/L (0.08 mg/L in 
seawater) for Thalassiosira pseudonana and >1.5 mg/L for Chlorella sp.
    Subsequent studies by Desjardins et al. (Ref. 75) confirmed the 
high acute toxicity of HBCD to Skeletonema costatum. In these studies, 
single concentrations were tested, but the assays were conducted 
without solvent and the concentrations were measured. Desjardins et al. 
(Ref. 75) reported approximately 10% inhibition of growth in 
Skeletonema costatum exposed to 0.041 mg/L for 72 hours. Desjardins et 
al. (Ref. 77) found that a saturated solution of 0.0545 mg/L resulted 
in 51% growth inhibition after 72 hours of exposure. The latter result 
corresponds to an approximate EC50 of 0.052 mg/L.
    Zebrafish embryo studies reported a variety of effects on embryos 
and larvae at low HBCD concentrations. In the Deng et al. (Ref. 64) 
study, developmental toxicity endpoints were assessed at 96 hours post-
fertilization in embryos/larvae exposed to HBCD starting 4 hours post-
fertilization. Survival of embryos/larvae was significantly reduced at 
all tested concentrations, making the low concentration of 0.05 mg/L 
the lowest-observed-effect-concentration (LOEC) in this study; a no-
observed-effect-concentration (NOEC) was not established. Embryonic 
malformation rate was significantly increased and larval growth 
significantly decreased at >=0.1 mg/L. Malformations included epiboly 
deformities, yolk sac and pericardial edema, tail and heart 
malformations, swim bladder inflation, and spinal curvature. Embryo 
hatching rate was reduced only at the high concentration of 1 mg/L. 
Heart rate, a marker for cardiac developmental toxicity, was 
significantly decreased at all tested concentrations. Associated 
mechanistic studies suggest the mechanism for developmental toxicity 
involves the generation of reactive oxygen species (ROS) and the 
consequent triggering of apoptosis genes. Increased ROS formation 
(indicative of oxidative stress) was observed at a nominal 
concentration of 0.1 mg/L. In the same study, zebrafish embryos exposed 
to HBCD exhibited increased expression of pro-apoptotic genes (Bax, 
P53, Puma, Apaf-1, caspase 3, and caspase-9), decreased expression of 
anti-apoptotic genes (Mdm2 and Bcl-2), and increased activity of 
enzymes involved in apoptosis (caspase-3 and caspase-9) with LOECs of 
0.05-1 mg/L.
    Hu et al. (Ref. 67) found that hatching of zebrafish embryos was 
delayed at 0.002 mg/L, the lowest concentration tested, and other 
concentrations up to and including 0.5 mg/L, but not the two high 
concentrations of 2.5 and 10 mg/L. The same authors observed an 
increase in heat shock protein (Hsp70) at 0.01 mg/L and an increase in 
malondialdehyde activity, used as a measure of lipid peroxidation, at 
0.5 mg/L. The activity of superoxide dismutase was increased at 0.1 mg/
L, but decreased at 2.5 and 10 mg/L. The authors concluded that HBCD 
can cause oxidative stress and over expression of Hsp70 in acute 
exposures of zebrafish embryos.
    Du et al. (Ref. 65) exposed zebrafish embryos 4 hours post-
fertilization to each of three diastereomers of HBCD ([alpha]-, [beta]-
, and [gamma]-HBCD) individually at nominal concentrations of 0.01, 
0.1, and 1.0 mg/L. Hatching success was reduced after 68 hours of 
exposure to [gamma]-HBCD at the lowest concentration (0.01 mg/L), but a 
higher concentration of [alpha]- or [beta]-HBCD (0.1 mg/L) was 
necessary to reduce hatching success. After 92 hours, survival was 
reduced at concentrations of 0.01, 0.1, and 1 mg/L of [gamma]-, [beta]-
, and [alpha]-HBCD, respectively. Growth, measured as body length of 
larvae after 92 hours of exposure, was reduced at 0.1 mg/L of [beta]- 
and [gamma]-HBCD and at 1 mg/L of [alpha]-HBCD. After 116 hours of 
exposure, malformations were observed at all test concentrations of 
[beta]- and [gamma]-HBCD and at 0.1 mg/L and above for [alpha]-HBCD. 
Effects on heart rate varied depending upon the length of exposure; 
reduced heart rate was observed at 0.1 mg/L of [beta]- and [gamma]-HBCD 
or 1 mg/L of [alpha]-HBCD at 44 hours and at 0.1 mg/L of [alpha]- and 
[beta]-HBCD at 92 hours, whereas [gamma]-HBCD resulted in an increase 
in heart rate at 1 mg/L at 92 hours. An increase in generation of ROS 
was observed after 116 hours at 0.1 mg/L of [beta]- and [gamma]-HBCD 
and at 1 mg/L of [alpha]-HBCD. Activities of caspase-3 and caspase-9 
enzymes, indicative of apoptosis, were increased after 116 hours at 0.1 
mg/L of [gamma]-HBCD and at 1 mg/L of [alpha]- and [beta]-HBCD. The 
authors ranked the HBCD diastereomers in the following order for 
developmental toxicity to zebrafish: [gamma]-HBCD > [beta] HBCD > 
[alpha]-HBCD.
    Effects indicative of oxidative stress, as seen in the zebrafish 
embryo studies, were also found in clams. Zhang et al. (Ref. 74) 
measured parameters indicative of antioxidant defenses and oxidative 
stress after 1, 3, 6, 10, and 15 days of exposure to low nominal 
concentrations of HBCD ranging from 0.000086 to 0.0086 mg/L in the clam 
Venerupis philippinarum. Increases in ethyoxyresorufin-o-deethylase 
(EROD) activity, glutathione (GSH) content, and DNA damage were 
observed in clams exposed to 0.00086 mg/L, while increased lipid 
peroxidation (LPO) was observed at 0.0086 mg/L. These same effects were 
observed at lower concentrations as the length of exposure increased.
    Anselmo et al. (Ref. 63) exposed sea urchin (Psammechinus miliaris) 
embryos to HBCD in an early life stage test. Newly-fertilized embryos 
were exposed to HBCD at nominal concentrations of 0, 9, 25, 50, and 100 
nanomolar (nM) (0, 0.0058, 0.016, 0.032, and 0.064 mg/L, respectively) 
in dimethyl sulfoxide solvent and evaluated at 72 hours post-
fertilization. A significant increase in morphological abnormalities 
was found at a nominal concentration of 100 nM HBCD (0.064 mg/L), the 
highest concentration tested. Observed malformations included short or 
deformed larval arms and slight edema around the larval body. The NOEC 
for this effect at 72 hours was 0.032 mg/L.
    2. Chronic aquatic toxicity. A measured MATC of 0.0042 mg/L, based 
on reduced size (length) of surviving young water fleas (Daphnia 
magna), indicates high chronic toxicity (Ref. 56). This study reported 
additional effects, including decreased reproductive rate and decreased 
mean weight of surviving young at 0.011 mg/L. Other effects reported 
following chronic exposure to HBCD included degenerative changes in the 
gills of clams (Macoma balthica), manifested by the increased frequency

[[Page 35281]]

of nuclear and nucleolar abnormalities and the occurrence of dead 
cells, at nominal concentrations of >=0.1 mg/L (50-day LOEC) (Ref. 71), 
a nominal MATC of 0.045 mg/L for increased morphological abnormalities 
in sea urchin (P. miliaris) embryos exposed to HBCD for up to 16 days 
in an early life stage test (Ref. 63), and a nominal MATC of 0.03 mg/L 
for increased malformation rate in marine medaka (Oryzias melastigma) 
embryos exposed to HBCD for 17 days in an early life stage test (Ref. 
66). The developmental abnormalities in medaka included yolk sac edema, 
pericardial edema, and spinal curvature (Ref. 66). Mechanistic findings 
in this study included increases in heart rate and sinus venosus-bulbus 
arteriosus (SV-BA) distance, which are markers for cardiac development, 
induction of oxidative stress and apoptosis, and suppression of 
nucleotide and protein synthesis.
    Thyroid effects were reported in juvenile rainbow trout 
(Oncorhynchus mykiss) following dietary exposure to HBCD (Refs. 68 and 
69). Each of the diastereomers of HBCD (administered separately via 
diet at concentrations of 5 ng/g of [alpha]-, [beta]-, or [gamma]-HBCD 
for up to 56 days) disrupted thyroid homeostasis, as indicated by lower 
free circulating T3 and T4 levels.
    The mechanisms of the effects on fish and invertebrates following 
chronic exposure were similar to those found in acute studies. Effects 
observed in fish include increased formation of ROS resulting in 
oxidative damage to lipids, proteins, and DNA, decreased antioxidant 
capacities in fish tissue (e.g., brains, hepatocytes, or erythrocytes), 
and increasing levels of EROD (detoxification enzyme) and 
PentoxyResorufin-O-Deethylase (PROD, detoxification enzyme) levels in 
hepatocytes of fish exposed to the nominal concentration of >=0.1 mg/L 
(corresponds to ~0.2 mg/g whole fish (wet weight)) for 42 days (Ref. 
73). Ronisz et al. (Ref. 70) found a significant increase in hepatic 
cytosolic catalase activity in rainbow trout (Oncorhynchus mykiss) 5 
days after a single intraperitoneal injection of 50 mg/kg was 
administered. The same authors observed reductions in liver somatic 
index (LSI) and EROD activity in a 28-day study in which rainbow trout 
were injected intraperitoneally with HBCD on days 1 and 14 at a dose 
somewhat less than 500 mg/kg. Zhang et al. (Ref. 74) observed the 
following signs of oxidative stress in clams (V. philippinarum) after 
15 days of exposure to HBCD: The activities of antioxidant enzymes 
(EROD, superoxide dismutase (SOD), and glutathione-S-transferase 
(GST)), as well as GSH content, were increased at 0.000086 mg/L, the 
lowest concentration tested. In addition, LPO was increased at 0.00086 
mg/L and DNA damage was increased at 0.0086 mg/L.
    3. Terrestrial toxicity and phytotoxicity. Japanese quail (Coturnix 
coturnix japonica) exposed for 6 weeks to an isomeric mixture of HBCD 
in the diet experienced a reduction in hatchability at all tested 
concentrations (12-1,000 ppm) (Ref. 57). Additional effects included a 
significant reduction in egg shell thickness starting at 125 ppm, 
decreases in egg weights and egg production rates starting at 500 ppm, 
increases in cracked eggs starting at 500 ppm, and adult mortality at 
1,000 ppm. A subsequent test, conducted at lower dietary 
concentrations, determined LOAEL and NOAEL values of 15 and 5 ppm, 
respectively, based on significant reduction of survival of chicks 
hatched from eggs of quails fed HBCD (Ref. 57).
    Several studies have been conducted examining effects of HBCD on 
American kestrels (Falco sparverius). Kobiliris (Ref. 78) reported a 
reduced ``corticosterone response'' (where ``corticosterone response'' 
was defined as a stimulation of the adrenal cortex to produce and 
release corticosterone into the bloodstream), reduced flying activities 
of juvenile males during hunting behavior trials, and delayed response 
times of juvenile females during predator avoidance behavior trials in 
American kestrels exposed in ovo to 164.13 ng/g wet weight. Kestrels 
exposed via the diet to 0.51 mg/kg/day beginning 3 weeks prior to 
pairing and continuing until the first chick hatched began to lay eggs 
6 days earlier than controls and laid larger clutches of smaller eggs 
(Ref. 59). Although the technical mixture of HBCD stereoisomers 
contained predominantly [gamma]-HBCD (80% of the mixture), the main 
isomer found in eggs was [alpha]-HBCD (>90% of the total HBCD in eggs). 
In a subsequent study, Marteinson et al. (Ref. 61) exposed kestrels to 
dietary HBCD at the same dose (0.51 mg/kg/day) and found increased 
testes weight in unpaired males, a marginally significant effect on 
testis histology in unpaired males (increased number of seminiferous 
tubules containing elongated spermatids; p = 0.052), marginally 
increased testosterone levels in breeding males (increased at the time 
the first egg was laid; p = 0.054), and no significant effect on sperm 
counts. Plasma T4 levels were reduced in breeding males throughout the 
study, which the authors took to suggest that thyroid disruption that 
may have contributed to the observed increase in testes weight. 
Marteinson et al. (Ref. 62) found altered reproductive behavior in both 
sexes of kestrels fed 0.51 mg/kg/day, including reduced activity in 
both sexes during courtship and in males during brood rearing, which 
may have contributed to the observed reduction in incubation nest 
temperature and also to the reduced egg size reported previously by 
Fernie et al. (Ref. 58). In a 22-day study of chickens (Gallus gallus 
domesticus) exposed to HBCD in ovo, reduced pipping success was 
observed at 100 ng/g egg (Ref. 79).
    The accumulation and toxicity of [alpha]-, [beta]-, and [gamma]-
HBCDs in maize have been studied (Ref. 80). The order of accumulation 
in roots was [beta]-HBCD > [alpha]-HBCD > [gamma]-HBCD and in shoots it 
was [beta]-HBCD > [gamma]-HBCD > [alpha]-HBCD. In maize exposed to 2 
[mu]g/L HBCD, the inhibitory effect of the diastereomers on the early 
development of maize as well as the intensities of hydroxyl radical and 
histone H2AX phosphorylation followed the order [alpha]-HBCD > [beta]-
HBCD > [gamma]-HBCD, which indicates diastereomer-specific oxidative 
stress and DNA damage in maize. The study confirmed that for maize 
exposed to HBCDs, the generation of reactive oxygen species was one, 
but not the only, mechanism for DNA damage.
    4. Conclusions regarding the ecological hazard potential of HBCD. 
HBCD has been shown to cause acute toxicity to aquatic organisms at 
concentrations as low as 0.009 mg/L and chronic toxicity at 
concentrations as low as 0.0042 mg/L. Toxicity to terrestrial species 
has been observed at doses as low as 0.51 mg/kg/day. The available 
evidence shows that HBCD is highly toxic to aquatic and terrestrial 
species.

C. What is EPA's review of the bioaccumulation data for HBCD?

    HBCD has been shown in numerous studies to bioaccumulate in aquatic 
species and biomagnify in aquatic and terrestrial food chains (Ref. 1). 
BCFs for HBCD in fish in the peer-reviewed literature range as high as 
18,100 (Refs. 81, 82, and 83). Some of the bioaccumulation values for 
fish species and a freshwater food web are shown in Table 1. The 
complete listing of the available bioaccumulation data and more details 
about the studies can be found in the ecological assessment (Ref. 1).

[[Page 35282]]



                         Table 1--HBCD BCF and BAF Data for Fish and Freshwater Food Web
----------------------------------------------------------------------------------------------------------------
                                      Duration and test
              Species                      endpoint                   Value                    Reference
----------------------------------------------------------------------------------------------------------------
Rainbow trout (Oncorhynchus         35-day BCF...........  8,974 and 13,085..........  Ref. 81.
 mykiss).
Fathead minnow (Pimephales          32-day BCF...........  18,100....................  Ref. 82.
 promelas).
Mirror carp (Cyprinus carpio        30-day exposure and    [alpha]-HBCD: 5,570-11,500  Ref. 83.
 morpha noblis).                     30-day depuration     [beta]-HBCD: 187-642......
                                     BCF.                  [gamma]-HBCD: 221-584.....
Mud carp (Cirrhinus molitorella),   Log BAF..............  4.8-7.7 for HBCD isomers    Ref. 84.
 nile tilapia (Tilapia nilotica),                           ([alpha][dash]HBCD had
 and suckermouth catfish                                    higher BAFs than [beta]-
 (Hypostomus plecostomus).                                  and [gamma][dash]HBCD)
                                                            (BAFs ranged from ~63,000
                                                            to 50,000,000).
Freshwater food web...............  Log BAF..............  [alpha]-HBCD: 2.58-6.01...  Ref. 85.
                                                           [beta]-HBCD: 3.24-5.58....
                                                           [gamma]-HBCD: 3.44-5.98...
                                                           [Sigma]HBCDs: 2.85-5.98...
                                                           (BAFs range from ~700 to
                                                            950,000).
----------------------------------------------------------------------------------------------------------------

    Drottar and Kruger (Ref. 81) provided strong evidence that HBCD 
bioaccumulates in a study conducted according to established guidelines 
(OECD Test Guideline (TG) 305 and Office of Prevention, Pesticides and 
Toxic Substances (OPPTS) 850.1730). In this study, BCFs of 13,085 and 
8,974 were reported in rainbow trout (O. mykiss) exposed to 0.18 and 
1.8 [micro]g/L, respectively. Concentrations of HBCD in tissue reached 
steady-state at day 14 for fish exposed to 1.8 [micro]g/L and, during 
the subsequent depuration stage, a 50% reduction of HBCD from edible 
and non-edible tissue and whole fish was reported on days 19 and 20 
post-exposure. In fish exposed to 0.18 [micro]g/L, an apparent steady-
state was reached on day 21, but on day 35, the tissue concentration of 
HBCD in fish increased noticeably; thus, steady-state was not achieved 
according to study authors, and BCF values (for the exposure 
concentration of 0.18 [micro]g/L) were calculated based on day 35 
tissue concentrations. Clearance of 50% HBCD from tissue of 0.18 
[micro]g/L exposed fish occurred 30-35 days post-exposure.
    Veith et al. (Ref. 82) further supports the conclusion that HBCD 
bioaccumulates in a study conducted prior to the establishment of 
standardized testing guidelines for bioconcentration studies. The study 
reported a BCF of 18,100 following exposure of fathead minnows to 6.2 
[micro]g/L; the BCF was identified as a steady-state BCF, but the 
report does not indicate the time when steady-state was reached. A 
depuration phase was not included in this study. Zhang et al. (Ref. 83) 
calculated BCFs for each HBCD diastereomer in mirror carp and found 
strong evidence that [alpha]-HBCD (BCF of 5,570-11,500) is much more 
bioaccumulative than [beta]- and [gamma]-HBCD (BCF of 187-642); BCF 
values that were normalized to lipid content were much higher (30,700-
45,200 for [alpha]-HBCD, 1,030-1,900 for [beta]-HBCD, and 950-1,730 for 
[gamma]-HBCD) than non-normalized BCFs.
    BAFs, which capture accumulation of HBCD from diet as well as water 
and sediment, were calculated for freshwater food webs in 
industrialized areas of Southern China in two separate field studies. 
He et al. (Ref. 84) calculated log BAFs of 4.8-7.7 (corresponding to 
BAFs of 63,000-50,000,000) for HBCD isomers in carp, tilapia, and 
catfish, and found higher BAFs for [alpha]-HBCD than [beta]- and 
[gamma]-HBCD. In a pond near an e-waste recycling site, Wu et al. (Ref. 
85) calculated log BAFs of 2.85-5.98 for HBCD (corresponding to BAFs of 
700-950,000) in a freshwater food web. Log BAFs for each diastereomer 
in this study were comparable to one another (see Table 1). La Guardia 
et al. (Ref. 86) calculated log BAFs in bivalves and gastropods 
collected downstream of a textile manufacturing outfall; these ranged 
from 4.2 to 5.3 for [alpha]- and [beta]-HBCD (BAFs of 16,000-200,000), 
and from 3.2 to 4.8 for [gamma]-HBCD (BAFs of 1,600-63,000).
    In general, [alpha]-HBCD bioaccumulates in organisms and 
biomagnifies through food webs to a greater extent than the [beta]- and 
[gamma]- diastereomers. Uncertainty remains as to the balance of 
diastereomer accumulation in various species and the extent to which 
bioisomerization and biotransformation rates for each isomer affect 
bioaccumulation potential. Some authors (e.g., Law et al., Ref. 87) 
have proposed that [gamma]-HBCD isomerizes to [alpha]-HBCD under 
physiological conditions, rather than uptake being diastereisomer-
specific. To test this theory, Esslinger et al. (Ref. 88) exposed 
mirror carp (Cyprinus carpio morpha noblis) to only [gamma]-HBCD and 
found no evidence of bioisomerization. In contrast, when Du et al. 
(Ref. 89) exposed zebrafish (Danio rerio) to only [gamma]-HBCD, they 
found detectable levels of [alpha]-HBCD in fish tissue, suggesting that 
bioisomerization occurred. Marvin et al. (Ref. 90) hypothesized that 
differences in accumulation could also be due in part to a combination 
of differences in solubility, bioavailability, and uptake and 
depuration kinetics.
    Zhang et al. (Ref. 91) calculated diastereomer-specific BCFs in 
algae and cyanobacteria ranging from 174 to 469. For the cyanobacteria 
(Spirulina subsalsa), the BCF for [alpha]-HBCD (350) was higher than 
the BCFs for [beta]-HBCD (270) and [gamma]-HBCD (174). However, for the 
tested alga (Scenedesmus obliquus), the BCF for [beta]-HBCD (469) was 
higher than that for the other isomers (390-407).
    In summary, HBCD has been shown in numerous studies to be highly 
bioaccumulative in aquatic species and biomagnify in aquatic and 
terrestrial food chains; however, diastereomer- and enantiomer-specific 
mechanisms of accumulation are still unclear.

D. What is EPA's review of the persistence data for HBCD?

    There are limited data available on the degradation rates of HBCD 
under environmental conditions. A short summary of the environmental 
fate and persistence data for HBCD is presented in Table 2; additional 
details about this data can be found in the HBCD hazard assessment 
(Ref. 1).

[[Page 35283]]



               Table 2--Environmental Degradation of HBCD
------------------------------------------------------------------------
             Property                   Value             Reference
------------------------------------------------------------------------
                                   Air
------------------------------------------------------------------------
Photodegradation.................  Photo-induced    Ref. 9.2.
                                    isomerization
                                    of [gamma]-
                                    HBCD to
                                    [alpha]-HBCD
                                    in indoor dust
                                    with a
                                    measured
                                    decrease in
                                    HBCD
                                    concentration
                                    concurrent
                                    with an
                                    increase of
                                    pentabromocycl
                                    ododecenes
                                    (PBCDs) in
                                    indoor dust.
                                   Indirect         Ref. 93.
                                    photolysis
                                    half-life: 26
                                    hours AOPWIN
                                    v1.92
                                    (estimated).
------------------------------------------------------------------------
                                  Water
------------------------------------------------------------------------
Hydrolysis.......................  Not expected     Ref. 44.
                                    due to lack of
                                    functional
                                    groups that
                                    hydrolyze
                                    under
                                    environmental
                                    conditions and
                                    low water
                                    solubility
                                    (estimated).
------------------------------------------------------------------------
                                Sediment
------------------------------------------------------------------------
Aerobic conditions...............  No               Refs. 76 and 94.
                                    biodegradation
                                    observed in 28-
                                    day closed-
                                    bottle test.
                                   Half-life: 128,  Ref. 95.
                                    92, and 72
                                    days for
                                    [alpha]-,
                                    [gamma]-, and
                                    [beta]-HBCD,
                                    respectively
                                    (estimated),
                                    based on a 44%
                                    decrease in
                                    total initial
                                    radioactivity
                                    in viable
                                    freshwater
                                    sediment.
                                   Half-life: >120
                                    days
                                    (estimated),
                                    based on a 15%
                                    decrease in
                                    total initial
                                    radioactivity
                                    in abiotic
                                    freshwater
                                    sediment.
                                   Half-life: 11    Ref. 96.
                                    and 32 days
                                    (estimated) in
                                    viable
                                    sediment
                                    collected from
                                    Schuylkill
                                    River and
                                    Neshaminy
                                    creek,
                                    respectively.
                                   Half-life: 190
                                    and 30 days
                                    (estimated) in
                                    abiotic
                                    sediment
                                    collected from
                                    Schuylkill
                                    River and
                                    Neshaminy
                                    creek.
Anaerobic conditions.............  Half-life: 92    Ref. 95.
                                    days
                                    (estimated),
                                    based on a 61%
                                    decrease in
                                    total initial
                                    radioactivity
                                    in viable
                                    freshwater
                                    sediment.
                                   Half-life: >120
                                    days
                                    (estimated),
                                    based on a 33%
                                    decrease in
                                    total initial
                                    radioactivity
                                    in abiotic
                                    freshwater
                                    sediment.
                                   Half-life: 1.5   Ref. 96.
                                    and 1.1 days
                                    (estimated) in
                                    viable
                                    sediment
                                    collected from
                                    Schuylkill
                                    River and
                                    Neshaminy
                                    creek.
                                   Half-life: 10
                                    and 9.9 days
                                    (estimated) in
                                    abiotic
                                    sediment
                                    collected from
                                    Schuylkill
                                    River and
                                    Neshaminy
                                    creek.
------------------------------------------------------------------------
                                  Soil
------------------------------------------------------------------------
Aerobic conditions...............  Half-life: >120  Ref. 95.
                                    days
                                    (estimated),
                                    based on a 10%
                                    decrease in
                                    total initial
                                    radioactivity
                                    in viable soil.
                                   Half-life: >120
                                    days
                                    (estimated),
                                    based on a 6%
                                    decrease in
                                    total initial
                                    radioactivity
                                    in abiotic
                                    soil.
                                   Half-life: 63    Ref. 96.
                                    days
                                    (estimated) in
                                    viable soil
                                    amended with
                                    activated
                                    sludge.
                                   Half-life: >120
                                    days
                                    (estimated) in
                                    abiotic soil..
Anaerobic conditions.............  Half-life: 6.9   Ref. 96.
                                    days
                                    (estimated) in
                                    viable soil
                                    amended with
                                    activated
                                    sludge.
                                   Half-life: 82
                                    days
                                    (estimated) in
                                    abiotic soil
                                    using a
                                    nominal HBCD
                                    concentration
                                    of 0.025 mg/kg
                                    dry weight.
------------------------------------------------------------------------

    1. Abiotic degradation. HBCD is not expected to undergo significant 
direct photolysis since it does not absorb radiation in the 
environmentally available region of the electromagnetic spectrum that 
has the potential to cause molecular degradation (Ref. 97). Although 
HBCD is expected to exist primarily in the particulate phase in the 
atmosphere, a small percentage may also exist in the vapor phase based 
on its vapor pressure (Refs. 22, 90, 98, and 99). HBCD in the vapor 
phase will be degraded by reaction with photochemically produced 
hydroxyl radicals in the atmosphere. An estimated rate constant of 5.01 
x 10-12 cubic centimeters (cm\3\)/molecules-second at 25 
[deg]C for this reaction corresponds to a half-life of 26 hours, 
assuming an atmospheric hydroxyl radical concentration of 1.5 x 10\6\ 
molecules/cm\3\ and a 12-hour day (Refs. 93 and 100).
    Photolytic isomerization of HBCD has been described in both indoor 
dust samples and in samples of HBCD standards dissolved in methanol 
using artificial light (Ref. 92). After 1 week in the presence of 
light, indoor dust containing predominantly [gamma]-HBCD was found to 
decrease in [gamma]-HBCD and increase in [alpha]-HBCD concentration. 
There was a measured decrease in HBCD concentration concurrent with an 
increase in PBCDs in the indoor dust exposed to artificial light. The 
three diastereomerically-pure HBCD standards ([alpha]-, [beta]-, and 
[gamma]-HBCD) that were dissolved in methanol also began to 
interconvert within 1 week, resulting in a decrease in [gamma]-HBCD 
concentration and an increase in [alpha]-HBCD concentration.
    HBCD is not expected to undergo hydrolysis in environmental waters 
due to lack of functional groups that hydrolyze under environmental 
conditions and the low water solubility of HBCD (Ref. 44).
    Observed abiotic degradation of HBCD during simulation tests based 
on Organisation for Economic Cooperation and Development (OECD) methods 
307 and 308 was approximately 33% in anaerobic freshwater sediment, 15% 
in aerobic freshwater sediment, and 6% in aerobic soil after 112-113 
days (Refs. 44 and 95). The results from these studies correspond to 
estimated half-lives >120 days in soil and sediment due to minimal 
degradation being observed. Initial concentrations of \14\C 
radiolabeled HBCD ([alpha]-, [beta]-, and [gamma]- \14\C-HBCD in a 
ratio of 7.74:7.84:81.5) were 3.0-4.7 mg/kg dry weight in the sediment 
and soil systems. HBCD degradation observed under abiotic conditions 
was attributed to abiotic reductive dehalogenation (Refs. 44, 76, and 
95). Degradation proceeded through a stepwise process to form

[[Page 35284]]

tetrabromocyclododecene, dibromocyclododecadiene (DBCD), and 1,5,9-
cyclododecatriene (Refs. 44 and 95). Further degradation of 1,5,9-
cyclododecatriene was not observed. In this study, HBCD degradation 
occurred faster in sediment than in soil and faster under anaerobic 
conditions compared to aerobic conditions (Refs. 44 and 95).
    Previous OECD 308 and 307 based simulation tests from the same 
authors (Davis et al. 2005, Ref. 96) presented results suggesting 
faster abiotic degradation, particularly in sediment under anaerobic 
conditions, but were performed at much lower HBCD concentrations and 
measured only [gamma]-HBCD (Refs. 44, 76, 90, 96, and 101). In this 
study, abiotic degradation half-lives in freshwater sediments were 30-
190 days under aerobic conditions and 9.9-10 days under anaerobic 
conditions. Estimated half-lives in abiotic soil were >120 days under 
aerobic conditions and 82 days under anaerobic conditions. This study 
evaluated [gamma]-HBCD only and did not address interconversion of HBCD 
isomers or [alpha]- and [beta]-HBCD degradation. The initial 
concentrations of HBCD were 0.025-0.089 mg/kg dry weight in the 
sediment and soil systems, nearly 100 times less than the HBCD 
concentrations used in the subsequent Davis et al. 2006 study (Ref. 
95). Higher concentrations of HBCD (3.0-4.7 mg/kg dry weight) in the 
Davis et al. 2006 study (Ref. 95) allowed for quantification of 
individual isomers, metabolite identification and mass balance 
evaluation (Refs. 95 and 101). Additionally, the Davis et al. 2005 
study (Ref. 96) was considered to be of uncertain reliability for 
quantifying HBCD persistence because of concerns regarding potential 
contamination of sediment samples, an interfering peak corresponding to 
[gamma]-HBCD in the liquid chromatography/mass spectrometry (LC/MS) 
chromatograms, and poor extraction of HBCD leading to HBCD recoveries 
of 33-125% (Refs. 44 and 101).
    2. Biotic degradation. A few studies on the biodegradation of HBCD 
were located. A closed bottle screening-level test for ready 
biodegradability (OECD Guideline 301D, EPA OTS 796.3200) was performed 
using an initial HBCD concentration of 7.7 mg/L and an activated 
domestic sludge inoculum (Refs. 76 and 94). No biodegradation was 
observed (0% of the theoretical oxygen demand) over the test period of 
28 days under the stringent guideline conditions of this test.
    Degradation of HBCD during simulation tests with viable microbes, 
based on OECD methods 307 and 308, was approximately 61% in anaerobic 
freshwater sediment, 44% in aerobic freshwater sediment, and 10% in 
aerobic soil after 112-113 days (Refs. 44 and 95). The results from 
this study correspond to estimated HBCD half-lives of 92 days in 
anaerobic freshwater sediment, 128, 92, and 72 days for [alpha]-, 
[gamma]-, and [beta]-HBCD, respectively in aerobic freshwater sediment, 
and >120 days in aerobic soil. An initial total \14\C-HBCD 
concentration of 3.0-4.7 mg/kg dry weight in the sediment and soil 
systems was used, allowing for quantification of individual isomers, 
metabolite identification, and mass balance evaluation (Refs. 95 and 
101). Although very high spiking rates can be toxic to microorganisms 
in biodegradation studies and lead to unrealistically long estimated 
half-lives, the results of this study did not suggest toxicity to 
microorganisms. Tests with viable microbes demonstrated increased HBCD 
degradation compared to the biologically-inhibited control studies. In 
combination, these studies suggest that HBCD will degrade slowly in the 
environment, although faster in sediment than in soil, faster under 
anaerobic conditions than aerobic conditions, faster with microbial 
action than without microbial action, and at different rates for 
individual HBCD diastereomers (slower for [alpha]-HBCD than for the 
[gamma]- and [beta]-stereoisomers).
    The same researchers (Ref. 76) previously conducted a water-
sediment simulation test for commercial HBCD based on OECD guideline 
308 using nominal HBCD concentrations of 0.034-0.089 mg/kg dry weight 
(Refs. 44, 76, and 102). Aerobic and anaerobic microcosms were pre-
incubated at 20 [deg]C for 49 days and at 23 [deg]C for 43-44 days, 
respectively. HBCD was then added to 14-37 g dry weight freshwater 
sediment samples in 250 ml serum bottles (water:sediment ratio of 1.6-
2.9) and the microcosms were sealed and incubated in the dark at 20 
[deg]C for up to 119 days. For the aerobic microcosms, the headspace 
oxygen concentration was kept above 10-15%. This study evaluated only 
[gamma]-HBCD and did not address interconversion of HBCD isomers or 
[alpha]- and [beta]-HBCD degradation. Disappearance half-lives of HBCD 
with sediment collected from Schuylkill River and Neshaminy creek were 
11 and 32 days in viable aerobic sediments, respectively (compared to 
190 and 30 days in abiotic aerobic controls, respectively), and 1.5 and 
1.1 days in viable anaerobic sediments, respectively (compared to 10 
and 9.9 days in abiotic anaerobic controls).
    Data from these tests suggest that anaerobic degradation is faster 
than aerobic degradation of HBCD in viable and abiotic sediments and 
that degradation is faster in viable conditions than abiotic 
conditions. While these findings are consistent with Davis et al. 2006 
(Ref. 95), the actual degradation rates in this study are much faster. 
However, results from this study do not provide a reliable indication 
of HBCD persistence. A mass balance could not be established because 
only [gamma]-HBCD was used to quantify HBCD concentrations, \14\C-
radiolabelled HBCD was not used, and degradation products were not 
identified; therefore, apparent disappearance of HBCD in this study may 
not reflect biodegradation. In addition, there were concerns that 
contaminated sediment may have been used, HBCD extraction was 
incomplete (HBCD recovery varied from 33 to 125%), and an interfering 
peak was observed in the LC/MS chromatograms corresponding to [gamma]-
HBCD (Refs. 44 and 101).
    Similarly, a soil simulation test was conducted based on OECD 
guideline 307 for commercial HBCD using 50 g dry weight sandy loam soil 
samples added to 250 ml serum bottles (Refs. 44, 76, 96, and 103). The 
moisture content was 20% by weight. Aerobic and anaerobic microcosms 
were pre-incubated at 20 [deg]C for 35 days and at 23 [deg]C for 43 
days, respectively. Activated sludge was added to the soil at 5 mg/g, 
and HBCD was added to the soil to achieve a nominal concentration of 
0.025 mg/kg dry weight. The microcosms were then incubated in the dark 
at 20 [deg]C for up to 120 days. The disappearance half-lives were 63 
days in viable aerobic soil (compared to >120 days in abiotic aerobic 
controls) and 6.9 days in viable anaerobic soil (compared to 82 days in 
abiotic anaerobic controls). As in the sediment studies, HBCD 
degradation in soil occurred faster under anaerobic conditions compared 
to aerobic conditions, and faster in viable conditions than abiotic 
conditions. The disappearance half-lives in soil were slower than those 
in sediment.
    Biological processes were suggested to be responsible for the 
increased degradation of HBCD in this study using viable conditions, 
relative to abiotic conditions; however, degradation was not adequately 
demonstrated in soil because no degradation products were detected and 
only [gamma]-HBCD was used to quantify HBCD concentrations, making it 
impossible to calculate a mass balance. HBCD recoveries on day 0 of the 
experiment were well below (0.011-0.018 mg/kg dry weight) the nominal 
test concentrations (0.025 mg/kg dry weight), suggesting rapid 
adsorption of HBCD to soil and poor extraction methods (Refs. 44 and 
101).

[[Page 35285]]

    In studies using 0.025-0.089 mg/kg HBCD (Davis et al. 2005, Ref. 
96), the estimated half-life values were shorter than studies using 
3.0-4.7 mg/kg HBCD (Davis et al. 2006, Ref. 95) by approximately one 
order of magnitude for aerobic viable sediment (11-32 days compared 
to72-128 days) and anaerobic viable sediment (1.1-1.5 days compared to 
92 days). The viable aerobic soil half-life using lower concentrations 
of HBCD (Davis et al. 2005, Ref. 96) was less than half of the half-
life based on the higher HBCD concentration (63 days compared to >120 
days) (Davis et al. 2006, Ref. 95). Both Davis et al. studies (Refs. 95 
and 96) suggest that HBCD degrades faster in sediment than in soil, 
faster under anaerobic conditions than aerobic conditions, and faster 
with microbial action than without microbial action. HBCD is poorly 
soluble, and it was suggested that at higher concentrations of HBCD, 
degradation is limited by mass transfer of HBCD into microbes. However, 
results from the Davis et al. 2005 study (Ref. 96) likely overestimate 
the rate of HBCD biodegradation, for the reasons noted previously 
(primarily, failure to use \14\C-radiolabelled HBCD, quantify isomers 
other than [gamma]-HBCD, identify degradation products, or establish a 
mass balance, but also procedural problems with contamination of 
sediment, incomplete HBCD extraction, and occurrence of an interfering 
peak in the LC/MS chromatograms corresponding to [gamma]-HBCD).
    It is important to note that the rapid biodegradation rates from 
Davis et al. 2005 (Ref. 96) are not consistent with environmental 
observations. HBCD has been detected over large areas and in remote 
locations in environmental monitoring studies (Refs 1 and 104). Dated 
sediment core samples indicate slow environmental degradation rates 
(Refs. 44, 90, 96, and 101). For example, HBCD was found at 
concentrations ranging from 112 to 70,085 [mu]g/kg dry weight in 
sediment samples collected at locations near a production site in 
Aycliffe, United Kingdom two years after the facility was closed down 
(Ref. 44). Monitoring data do not provide a complete, quantitative 
determination of persistence because HBCD emission sources, rates, and 
quantities are typically unknown, and all environmental compartments 
are not considered. However, the monitoring data do provide evidence in 
support of environmental persistence. In addition, the widespread 
presence of HBCD in numerous terrestrial and aquatic species indicates 
persistence in the environment sufficient for bioaccumulation to occur 
(Ref. 1).

IV. Rationale for Listing HBCD and Lowering the Reporting Threshold

A. What is EPA's rationale for listing the HBCD category?

    HBCD has been shown to cause developmental effects at doses as low 
as 146.3 mg/kg/day (LOAEL) in male rats. Developmental effects have 
also been observed with a BMDL of 0.056 mg/kg/day (BMD of 0.18 mg/kg/
day) based on effects in female rats and a BMDL of 0.46 mg/kg/day (BMD 
of 1.45 mg/kg/day) based on effects in male rats. HBCD also causes 
reproductive toxicity at doses as low 138 mg/kg/day (LOAEL) in female 
rats. Based on the available developmental and reproductive toxicity, 
EPA believes that HBCD can be reasonably anticipated to cause 
moderately high to high chronic toxicity in humans. Therefore, EPA 
believes that the evidence is sufficient for listing the HBCD category 
on the EPCRA section 313 toxic chemical list pursuant to EPCRA section 
313(d)(2)(B) based on the available developmental and reproductive 
toxicity data.
    HBCD has been shown to be highly toxic to both aquatic and 
terrestrial species with acute aquatic toxicity values as low as 0.009 
mg/L and chronic aquatic toxicity values as low as 0.0042 mg/L. HBCD is 
highly toxic to terrestrial species as well with observed toxic doses 
as low as 0.51 and 2.1 mg/kg/day. In addition to being highly toxic, 
HBCD is also bioaccumulative and persistent in the environment, which 
further supports a high concern for the toxicity to aquatic and 
terrestrial species. EPA believes that HBCD meets the EPCRA section 
313(d)(2)(C) listing criteria on toxicity alone but also based on 
toxicity and bioaccumulation as well as toxicity and persistence in the 
environment. Therefore, EPA believes that the evidence is sufficient 
for listing the HBCD category on the EPCRA section 313 toxic chemical 
list pursuant to EPCRA section 313(d)(2)(C) based on the available 
ecological toxicity data as well as the bioaccumulation and persistence 
data.
    HBCD has the potential to cause developmental and reproductive 
toxicity at moderately low to low doses and is highly toxic to aquatic 
and terrestrial organisms; thus, EPA considers HBCD to have moderately 
high to high chronic human health toxicity and high ecological 
toxicity. EPA does not believe that it is appropriate to consider 
exposure for chemicals that are moderately high to highly toxic based 
on a hazard assessment when determining if a chemical can be added for 
chronic human health effects pursuant to EPCRA section 313(d)(2)(B) 
(see 59 FR 61440-61442). EPA also does not believe that it is 
appropriate to consider exposure for chemicals that are highly toxic 
based on a hazard assessment when determining if a chemical can be 
added for environmental effects pursuant to EPCRA section 313(d)(2)(C) 
(see 59 FR 61440-61442). Therefore, in accordance with EPA's standard 
policy on the use of exposure assessments (See November 30, 1994 (59 FR 
61432, FRL-4922-2), EPA does not believe that an exposure assessment is 
necessary or appropriate for determining whether HBCD meets the 
criteria of EPCRA section 313(d)(2)(B) or (C).

B. What is EPA's rationale for lowering the reporting threshold for 
HBCD?

    EPA believes that the available bioaccumulation and persistence 
data for HBCD support a classification of HBCD as a PBT chemical. HBCD 
has been shown to be highly bioaccumulative in aquatic species and to 
also biomagnify in aquatic and terrestrial food chains. While there is 
limited data on the half-life of HBCD in soil and sediment, the best 
available data supports a determination that the half-life of HBCD in 
soil and sediment is at least 2 months. This determination is further 
supported by the data from environmental monitoring studies, which 
indicate that HBCD has significant persistence in the environment. The 
widespread presence of HBCD in numerous terrestrial and aquatic species 
also supports the conclusion that HBCD has significant persistence in 
the environment. Therefore, consistent with EPA's established policy 
for PBT chemicals (See 64 FR 58666, October 29, 1999) (FRL-6389-11) EPA 
is proposing to establish a 100-pound reporting threshold for the HBCD 
category.

V. References

    The following is a listing of the documents that are specifically 
referenced in this document. The docket includes these documents and 
other information considered by EPA, including documents that are 
referenced within the documents that are included in the docket, even 
if the referenced document is not itself physically located in the 
docket. For assistance in locating these other documents, please 
consult the person listed under FOR FURTHER INFORMATION CONTACT.

1. USEPA, OEI. 2016. Technical Review of Hexabromocyclododecane 
(HBCD) CAS

[[Page 35286]]

Registry Numbers 3194-55-6 and 25637-99-4. January 25, 2016.
2. USEPA, OEI. 2014. Economic Analysis of the Proposed Rule to add 
HBCD to the List of TRI Reportable Chemicals. March 28, 2014.
3. Arita, R., Miyazaki, K., Mure, S. 1983. Metabolic test of HBCD. 
Test on chemical substances used in household items. Studies on 
pharmacodynamics of HBCD (unpublished). In: Toxicology summary: HBCD 
(HBCD), Albemarle, S.A. Department of Pharmacy, Hokkaido University 
Hospital, Japan.
4. Yu, C.C., Atallah, Y.H. 1980. Pharmacokinetics of HBCD in rats 
(unpublished). Vesicol Chemical Corporation, Rosemont, IL.
5. Szabo, D.T., Diliberto, J.J., Hakk, H. et al. 2010. 
Toxicokinetics of the flame retardant HBCD gamma: Effect of dose, 
timing, route, repeated exposure, and metabolism. Toxicol. Sci. 
117(2):282-293.
6. Szabo, D.T., Diliberto, J.J., Hakk, H., Huwe, J.K., Birnbaum, 
L.S. 2011. Toxicokinetics of the flame retardant 
hexabromocyclododecane alpha: Effect of dose, timing, route, 
repeated Exposure, and metabolism. Toxicol. Sci. 121(2):234-244.
7. Reistad, T., Fonnum, F., Mariussen, E. 2006. Neurotoxicity of the 
pentabrominated diphenyl ether mixture, DE-71, and HBCD (HBCD) in 
rat cerebellar granule cells in vitro. Arch. Toxicol. 80(11):785-
796.
8. van der Ven, L.T.M., Verhoef, A., van de Kuil, T., Slob, W., 
Leonards, P.E.G., Visser, T.J., Hamers, T., Herlin, M., Hakansson, 
H., Olausson, H., Piersma, A.H., Vos, J.G. 2006. A 28-day oral dose 
toxicity study enhanced to detect endocrine effects of 
hexabromocyclododecane in Wistar rats. Toxicological Sciences 94(2): 
281-292.
9. van der Ven, L.T.M., van de Kuil, T., Leonards, P.E., et al. 
2009. Endocrine effects of HBCD (HBCD) in a one-generation 
reproduction study in Wistar rats. Toxicol Lett 185:51-62. Including 
supplementary tables.
10. Brandsma, S.H., van der Ven, L.T.M., De Boer, J. and Leonards, 
P.E. 2009. Identification of hydroxylated metabolites of 
hexabromocyclododecane in wildlife and 28-days exposed Wistar rats. 
Environ. Sci. Technol. 43, 6058-6063.
11. Hakk, H., Szabo, D.T., Huwe, J., Diliberto, J. and Birnbaum, 
L.S. 2012. Novel and distinct metabolites identified following a 
single oral dose of [alpha]- or [gamma]-hexabromocyclododecane in 
mice. Environ. Sci. Technol. 46:13494-13503.
12. Sanders, J.M., Knudsen, G.A. and Birnbaum, L.S. 2013. The fate 
of [beta]-hexabromocyclododecane in female C57BL/6 mice. 
Toxicological Sciences 134(2): 251-257.
13. Antignac, J.P., Cariou, R., Maume, D., et al. 2008. Exposure 
assessment of fetus and newborn to brominated flame retardants in 
France: preliminary data. Mol. Nutr. Food Res. 52(2):258-265.
14. Weiss, J., Wallin, E., Axmon, A., et al. 2006. Hydroxy-PCBs, 
PBDEs, and HBCDDs in serum from an elderly population of Swedish 
fishermen's wives and associations with bone density. Environ. Sci. 
Technol. 40(20):6282-6289.
15. Kakimoto, K., Akutsu, K., Konishi, Y., et al. 2008. Time trend 
of HBCD in the breast milk of Japanese women. Chemosphere 
71(6):1110-1114.
16. Rawn, D.F.K., Ryan, J.J., Sadler, A.R. et al. 2014. Brominated 
flame retardant concentrations in sera from the Canadian Health 
Measures Survey (CHMS) from 2007 to 2009. Environment International 
63: 26-34.
17. Abdallah, M. and Harrad, S. 2011. Tetrabromobisphenol-A, 
hexabromocyclododecane and its degradation products in UK human 
milk: Relationship to external exposure. Environment International, 
37: 443-448.
18. Meijer, L., Weiss, J., Van Velzen, M., et al. 2008. Serum 
concentrations of neutral and phenolic organohalogens in pregnant 
women and some of their infants in The Netherlands. Environ. Sci. 
Technol. 42(9):3428-3433.
19. Thomsen, C., Molander, P., Daae, H.L., et al. 2007. Occupational 
exposure to HBCD at an industrial plant. Environ. Sci. Technol. 
41(15):5210-5216.
20. Fangstrom, B., Strid, A., Bergman, A. 2005. Temporal trends of 
brominated flame retardants in milk from Stockholm mothers, 1980-
2004. Department of Environmental Chemistry, Stockholm University, 
Stockholm, Sweden. Available online at: http://www.imm.ki.se/Datavard/PDF/mj%C3%B6lk_poolade_NV%20rapport%202005%20modersmjolk.pdf.
21. Fangstrom, B., Athanassiadis, I., Odsjo, T., et al. 2008. 
Temporal trends of polybrominated diphenyl ethers and HBCD in milk 
from Stockholm mothers, 1980-2004. Mol. Nutr. Food Res. 52(2):187-
193.
22. Covaci, A., Gerecke, A.C., et al. 2006. Hexabromocyclododecanes 
(HBCDs) in the Environment and Humans: A Review. Environ. Sci. 
Technol. 40: 3679-3688.
23. Johnson-Restrepo, B., Adams, D.H., et al. 2008. 
Tetrabromobisphenol A (TBBPA) and Hexabromocyclododecanes (HBCDs) in 
tissues of humans, dolphins, and sharks from the United States. 
Chemosphere 70: 1935-1944.
24. Toms, L-M.L., Guerra, P., Eljarrat, E., Barcel[oacute], D., 
Harden, F.A., Hobson, P., et al. 2012. Brominated flame retardants 
in the Australian population: 1993-2009. Chemosphere 89:398-403.
25. Schecter, A., Szabo, D.T., Miller, J., Gent, T.L., Malik-Bass, 
N., Petersen, M., Paepke, O., Colacino, J.A., Hynan L.S., Harris, 
T.R., Malla, S., Birnbaum, L.S. 2012. Hexabromocyclododecane (HBCD) 
stereoisomers in U.S. food from Dallas, TX. Environmental Health 
Perspectives 120(9): 1260-1264.
26. IRDC (International Research and Development Corporation). 1977. 
Acute toxicity studies in rabbits and rats with HBCD with 
attachments. Submitted under TSCA Section 8E; EPA Document No. 88-
7800065; NTIS No. OTS0200051.
27. IRDC (International Research and Development Corporation). 1978. 
Acute toxicity studies in rabbits and rats with residue of HBCD with 
attachments and cover letter dated 030178. Submitted under TSCA 
Section 8E; EPA Document No. 88-7800088; NTIS No. OTS0200466.
28. Pharmakon Research International Inc. 1990. Acute exposure oral 
toxicity study in rats (83 EPA/OECD) with attachments and cover 
letter dated 030890. Submitted under TSCA Section 8D; EPA Document 
No. 86-900000166; NTIS No. OTS0522237.
29. Gulf South Research Institute. 1988. Initial submission: Letter 
from Ethyl Corp to USEPA regarding technical and toxicity data on 
brominated flame retardants including HBCD. EPA Document No. FYI-
OTS-0794-0947; NTIS No. OTS0000947.
30. BASF. 1990. Report on the study of the acute oral toxicity of 
HBCD in the mouse with cover letter dated 03-12-90. Submitted under 
TSCA Section 8D; EPA Document No. 86-900000383; NTIS No. OTS0522946.
31. Lewis, A.C., Palanker, A.L. 1978. A dermal LD50 study 
in albino rabbits and an inhalation LC50 study in albino 
rats. Test material GLS-S6-41A (unpublished). Consumer Product 
Testing, Fairfield, NJ; Experiment Reference No. 78385-2. Client: 
Saytech Inc.
32. Momma, J., Kaniwa, M., Sekiguchi, H., Ohno, K., Kawasaki, Y., 
Tsuda, M., Nakamura, A., Kurokawa, Y. 1993. Dermatological 
evaluation of a flame retardant, hexabromocyclododecane (HBCD) on 
guinea pig by using the primary irritation, sensitization, 
phototoxicity, and photosensitization of skin. (Article in Japanese; 
English abstract). Eisei Shikenjo Hokoku 111:18-24.
33. Chengelis, C. 1997. A 28-day repeated dose oral toxicity study 
of HBCD in rats. Study No. WIL-186004. WIL Research Laboratories, 
Inc. Ashland, OH.
34. Chengelis, C. 2001. An oral (gavage) 90-day toxicity study of 
HBCD in rats. Study No. WIL-186012. WIL Research Laboratories, Inc. 
Ashland, Ohio.
35. Chengelis, C. 2002. Amendment to the Final Report for: An oral 
(gavage) 90-day toxicity study of HBCD in rats. Study No. WIL-
186012. WIL Research Laboratories, Inc. Ashland, Ohio.
36. Hill, R.N., Crisp, T.M., Hurley, P.M., Rosenthal, S.L., and 
Singh, D.V. 1998. Risk assessment of thyroid follicular cell tumors. 
Environ. Health Perspect. 106, 447-457.
37. Zeller, H. and Kirsch, P. 1969. Hexabromocyclododecane: 28-day 
feeding trials with rats. BASF unpublished laboratory study. As 
cited in USEPA. 2001. High Production Volume (HPV) data summary and 
test plan for hexabromocyclododecane (HBCD) CAS No. 3194-55-6. 
Prepared by the American Chemistry Council's Brominated Flame 
Retardant Industry Panel (BFRIP), Arlington, VA.
38. Zeller, H. and Kirsch, P. 1970. Hexabromocyclododecane: 90-day

[[Page 35287]]

feeding trials with rats. BASF unpublished laboratory study. As 
cited in USEPA. 2001. High Production Volume (HPV) data summary and 
test plan for hexabromocyclododecane (HBCD) CAS No. 3194-55-6. 
Prepared by the American Chemistry Council's Brominated Flame 
Retardant Industry Panel (BFRIP), Arlington, VA.
39. Kurokawa, Y., Inoue, T., Uchida, Y., et al. 1984. Carcinogenesis 
test of flame retarder hexabromocyclododecane in mice. Hardy, M.; 
Albemarle Corporation, personal communication, Department of 
Toxicology, National Public Health Research Institute, Biological 
Safety Test Research Center. Unpublished, translated from Japanese. 
As cited in reference 40.
40. USEPA. 2014. Flame Retardant Alternatives for 
Hexabromocyclododecane (HBCD): Final Report.
41. Saegusa, Y., Fujimoto, H., Woo, G., et al. 2009. Developmental 
toxicity of brominated flame retardants, tetrabromobisphenol A and 
1,2,5,6,9,10-HBCD, in rat offspring after maternal exposure from 
mid-gestation through lactation. Reprod. Toxicol. 28(4):456-67.
42. Ema, M., Fujii, S., Hirata-Koizumi, M., et al. 2008. Two-
generation reproductive toxicity study of the flame retardant HBCD 
in rats. Reprod. Toxicol. 25(3):335-351.
43. Murai, T., Kawasaki, H., Kanoh, S. 1985. Studies on the toxicity 
of insecticides and food additives in pregnant rats (7). Fetal 
toxicity of HBCD. Oyo Yakuri (Pharmacometrics) 29:981-986 (in 
Japanese with English abstract).
44. European Commission. 2008. Risk Assessment: 
Hexabromocyclododecane CAS-No.: 25637-99-4 EINECS No.: 247-148-4, 
Final Report May 2008. Luxembourg: Office for Official Publications 
of the European Communities.
45. Eriksson, P., Fischer, C., Wallin, M., et al. 2006. Impaired 
behaviour, learning and memory, in adult mice neonatally exposed to 
HBCD (HBCDD). Environ. Toxicol. Pharmacol. 21(3):317-322.
46. USEPA. 1998. Guidelines for neurotoxicity risk assessment. Risk 
Assessment Form. Federal Register. 63 FR 26926, May 14, 1998 (FRL-
6011-3).
47. Industrial Bio-Test Labs. 1990. Mutagenicity of two lots of FM-
100 lot 53 and residue of lot 3322 in the absence and presence of 
metabolic activation with test data and cover letter. Submitted 
under TSCA Section 8D; EPA Document No. 86-900000267; NTIS No. 
OTS0523259.
48. Litton Bionetics Inc. 1990. Mutagenicity evaluation of 421-32b 
(Final report) with test data and cover letter. Submitted under TSCA 
Section 8D; EPA Document No. 86-900000265; NTIS No. OTS0523257.
49. SRI Research Institute. 1990. In vitro microbiological 
mutagenicity studies of four CIBA-GEIGY corporation compounds (Final 
report) with test data and cover letter. Submitted under TSCA 
Section 8D; EPA Document No. 86-900000262; NTIS No. OTS0523254.
50. Zeiger, E., Anderson, B., Haworth, S., et al. 1987. Salmonella 
mutagenicity tests: III. Results from the testing of 255 chemicals. 
Environ. Mutagen. 9 (Suppl. 9):1-110.
51. Huntingdon Research Center. 1978. Ames metabolic activation test 
to assess the potential mutagenic effect of compound no. 49 with 
cover letter dated 031290. Submitted under TSCA Section 8D; EPA 
Document No. 86-900000385; NTIS No. OTS0522948.
52. Pharmakologisches Institute. 1978. Ames test with hexabromides 
with cover letter dated 031290. Submitted under TSCA Section 8D; EPA 
Document No. 86-900000379; NTIS No. OTS0522942.
53. Ethyl Corporation. 1985. Genetic toxicology Salmonella/
microsomal assay on HBCD with cover letter dated 030890. Submitted 
under TSCA Section 8D; EPA Document No. 86-900000164; NTIS No. 
OTS0522235.
54. Microbiological Associates Inc. 1996. HBCD (HBCD): chromosome 
aberrations in human peripheral blood lymphocytes with cover letter 
dated 12/12/1996. Submitted under TSCA Section 8D; EPA Document No. 
86970000358; NTIS No. OTS0573552.
55. Walsh, G.E., Yoder, M.J., McLaughlin, L.L., et al. 1987. 
Responses of marine unicellular algae to brominated organic 
compounds in six growth media. Ecotoxicol. Environ. Saf. 14:215-222.
56. Drottar, K.R., Krueger, H.O. 1998. Hexabromocyclododecane 
(HBCD): A flow-through life-cycle toxicity test with the cladoceran 
(Daphnia magna). Report #439A-108. Wildlife International Ltd, 
Easton, MD, pp 78. Submitted under TSCA Section 8D; EPA Document No. 
86980000152; OTS0559490.
57. MOEJ (Ministry of the Environment, Japan). 2009. 6-Week 
administration study of 1,2,5,6,9,10-hexabromocyclododecane for 
avian reproduction toxicity under long-day conditions using Japanese 
quail. Report. Ministry of the Environment, Japan. Research 
Institute for Animal Science in Biochemistry & Toxicology (as cited 
in Ref. 58).
58. UNEP (United Nations Environmental Program). 2010. 
Hexabromocyclododecane draft risk profile. United Nations 
Environmental Program, Stockholm Convention.
59. Fernie, K.J., Marteinson, S.C., Bird, D.M., et al. 2011. 
Reproductive changes in American kestrels (Falco sparverius) in 
relation to exposure to technical hexabromocyclododecane flame 
retardant. Environ. Toxicol. Chem. 30:2570-2575.
60. Marteinson, S.C., Bird, D.M., Shutt, J.L., et al. 2010. Multi-
generational effects of polybrominated diphenylethers exposure: 
Embryonic exposure of male American kestrels (Falco sparverius) to 
DE-71 alters reproductive success and behaviors. Environ. Toxicol. 
Chem. 29: 1740-1747.
61. Marteinson, S.C., Kimmins, S., Letcher, R.J., et al. 2011. Diet 
exposure to technical hexabromocyclododecane (HBCD) affects testes 
and circulating testosterone and thyroxine levels in American 
kestrels (Falco sparverius). Environ. Res. 111:1116-1123.
62. Marteinson, S.C., Bird, D.M., Letcher, R.J., et al. 2012. 
Dietary exposure to technical hexabromocyclododecane (HBCD) alters 
courtship, incubation and parental behaviors in American kestrels 
(Falco sparverius). Chemosphere 89:1077-1083.
63. Anselmo, H.M.R., Koerting, L., Devito, S., et al. 2011. Early 
life developmental effects of marine persistent organic pollutants 
on the sea urchin Psammechinus miliaris. Ecotox. Environ. Safe. 
74:2182-2192.
64. Deng, J., Yu, L., Liu, C., et al. 2009. Hexabromocyclododecane-
induced developmental toxicity and apoptosis in zebrafish embryos. 
Aquat. Toxicol. 93(1):29-36.
65. Du, M., Zhang, D., Yan, C., et al. 2012. Developmental toxicity 
evaluation of three hexabromocyclododecane diastereoisomers on 
zebrafish embryos. Aquat. Toxicol. 112-113:1-10.
66. Hong, H., Li, D., Shen, R., et al. 2014. Mechanisms of 
hexabromocyclododecanes induced developmental toxicity in marine 
medaka (Oryzias melastigma) embryos. Aquat. Toxicol. 152:173-185.
67. Hu, J., Liang, Y., Chen, M., et al. 2009. Assessing the toxicity 
of TBBPA and HBCD by zebrafish embryo toxicity assay and biomarker 
analysis. Environ. Toxicol. 24:334-342.
68. Palace, V.P., Pleskach, K., Halldorson, T., et al. 2008. 
Biotransformation enzymes and thyroid axis disruption in juvenile 
rainbow trout (Oncorhynchus mykiss) exposed to 
hexabromocyclododecane diastereoisomers. Environ. Sci. Technol. 
42(6):1967-1972.
69. Palace, V., Park, B., Pleskach, K., et al. 2010. Altered 
thyroxine metabolism in rainbow trout (Oncorhynchus mykiss) exposed 
to hexabromocyclododecane (HBCD). Chemosphere 80(2):165-169.
70. Ronisz, D., Farmen Finne, E., Karlsson, H., et al. 2004. Effects 
of the brominated flame retardants hexabromocyclododecane (HBCDD), 
and tetrabromobisphenol A (TBBPA), on hepatic enzymes and other 
biomarkers in juvenile rainbow trout and feral eelpout. Aquat. 
Toxicol. 69:229-245.
71. Smolarz, K. and Berger, A. 2009. Long-term toxicity of 
hexabromocyclododecane (HBCDD) to the benthic clam Macoma balthica 
(L.) from the Baltic Sea. Aquat. Toxicol. 95(3):239-247.
72. Wu, M., Zuo, Z., Li, B., et al. 2013. Effects of low-level 
hexabromocyclododecane (HBCD) exposure on cardiac development in 
zebrafish embryos. Ecotoxicology 22:1200-1207.
73. Zhang, X., Yang, F., Zhang, X., et al. 2008. Induction of 
hepatic enzymes and oxidative stress in Chinese rare minnow 
(Gobiocypris rarus) exposed to waterborne hexabromocyclododecane 
(HBCDD). Aquat. Toxicol. 86(1):4-11.
74. Zhang, H., Pan, L., Tao, Y. 2014. Antioxidant responses in clam

[[Page 35288]]

Venerupis philippinarum exposed to environmental pollutant 
hexabromocyclododecane. Environ. Sci. Pollut. Res. 21:8206-8215.
75. Desjardins, D., MacGregor, J.A., Krueger, H.O. 2004. 
Hexabromocyclododecane (HBCD): A 72 hour toxicity test with the 
marine diatom (Skeletonema costatum), Final report. Wildlife 
Internation Ltd, Easton, MD, pp 66. As cited in Refs. 44 and 76.
76. IUCLID. 2005. Hexabromocyclododecane IUCLID dataset. Submitted 
to U.S. EPA's High Production Volume (HPV) Chemical Program.
77. Desjardins, D., MacGregor, J.A., Krueger, H.O. 2005. Final 
report. Chapter 1, Hexabromocyclododecane (HBCD): A 72-hour toxicity 
test with the marine diatom (Skeletonema costatum) using a co-
solvent. Chapter 2, Hexabromocyclododecane (HBCD): A 72-hour 
toxicity test with the marine diatom (Skeletonema costatum) using 
generator column saturated media. Wildlife International Ltd, 
Easton, MD, pp19. As cited in Ref. 44.
78. Kobiliris, D. 2010. Influence of embryonic exposure to 
hexabromocyclododecane (HBCD) on the corticosterone response and 
``fight or flight'' behaviors of captive American kestrels. Thesis 
submitted to McGill University in partial fulfilment of the 
requirements of the degree of Masters of Science. Department of 
Natural Resource Sciences, McGill University, Montreal, Canada.
79. Crump, D., Egloff, C., Chiu, S., et al. 2010. Pipping success, 
isomer-specific accumulation, and hepatic mRNA expression in chicken 
embryos exposed to HBCD. Toxicol. Sci. 115:492-500.
80. Wu, T., Wang, S., Huang, H., et al. 2012. Diastereomer-specific 
uptake, translocation, and toxicity of hexabromocyclododecane 
diastereoisomers to maize. J. Agr. Food Chem. 60:8528-8534.
81. Drottar, K.R. and Krueger, H.O. 2000. Hexabromocyclododecane 
(HBCD): A flow-through bioconcentration test with the rainbow trout 
(Oncorhynchus mykiss). Report# 439A-111. Wildlife International Ltd, 
Easton, MD, pp 1-137. Submitted under TSCA Section FYI; EPA Document 
No. 84010000001; OTS0001392.
82. Veith, G.D., Defoe, D.L., Bergstedt, B.V. 1979. Measuring and 
estimating the bioconcentration factor of chemicals in fish. J. Fish 
Res. Board Can. 36:1040-1048.
83. Zhang, Y., Sun, H., Ruan, Y. 2014. Enantiomer-specific 
accumulation, depuration, metabolization and isomerization of 
hexabromocyclododecane (HBCD) diastereomers in mirror carp from 
water. J. Haz. Mater. 264:8-15.
84. He, M., Luo, X., Yu, L., et al. 2013. Diasteroisomer and 
enantiomer-specific profiles of hexabromocyclododecane and 
tetrabromobisphenol A in an aquatic environment in a highly 
industrialized area, South China: Vertical profile, phase partition, 
and bioaccumulation. Environ. Poll. 179:105-110.
85. Wu, J., Guan, Y., Zhang, Y., et al. 2011. Several current-use, 
non-PBDE brominated flame retardants are highly bioaccumulative: 
Evidence from field determined bioaccumulation factors. Environ. 
Int. 37:210-215.
86. La Guardia, M.J., Hale, R.C., Harvey, E., et al. 2012. In situ 
accumulation of HBCD, PBDEs, and several alternative flame-
retardants in the bivalve (Corbicula fluminea) and gastropod (Elimia 
proxima). Environ. Sci. Technol. 46:5798-5805.
87. Law, K., Palace, V.P., Halldorson, T., et al. 2006. Dietary 
accumulation of hexabromocyclododecane diastereoisomers in juvenile 
rainbow trout (Oncorhynchus mykiss) I: Bioaccumulation parameters 
and evidence of bioisomerization. Environ. Toxicol. Chem. 
25(7):1757-1761.
88. Esslinger, S., Becker, R., M[uuml]ller-Belecke, A., et al. 2010. 
HBCD stereoisomer pattern in mirror carps following dietary exposure 
to pure [gamma]-HBCD enantiomers. J. Agric. Food Chem. 58:9705-9710.
89. Du, M., Lin, L., Yan, C., et al. 2012. Diastereoisomer- and 
enantiomer-specific accumulation, depuration, and bioisomerization 
of hexabromocyclododecanes in zebrafish (Danio rerio). Environ. Sci. 
Technol. 46:11040-11046.
90. Marvin, C.H., Tomy, G.T., Armitage, J.M., et al. 2011. 
Hexabromocyclododecane: Current understanding of chemistry, 
environmental fate and toxicology and implications for global 
management. Environ. Sci. Technol. 45:8613-8623. Including 
supporting information document.
91. Zhang, Y., Sun, H., Zhu, H., et al. 2014. Accumulation of 
hexabromocyclododecane diastereomers and enantiomers in two 
microalgae, Spirulina subsalsa and Scenedesmus obliquus. Ecotox. 
Environ. Safe. 104:136-142.
92. Harrad, S; Abdallah, MA; Covaci, A. (2009a) Causes of 
variability in concentrations and diastereomer patterns of 
Hexabromocyclododecanes in indoor dust. Environment International 
35:573-579.
93. USEPA. 2011. EPI Suite results for CAS 003194-55-6. Download EPI 
SuiteTM v4.0. U.S. Environmental Protection Agency. Available online 
at http://www.epa.gov/opptintr/exposure/pubs/episuitedl.htm (see 
section 2, attachment A in Ref. 1).
94. Schaefer, E.C. and Haberlein, D. 1996. Hexabromocyclododecane 
(HBCD): Closed bottle test. 439E-102, Wildlife International Ltd, 
Easton, MD, USA (as cited in Ref. 44).
95. Davis, J.W., Gonsior, S.J., Markham, D.A., et al. 2006. 
Biodegradation and product identification of 
[\14\C]hexabromocyclododecane in wastewater sludge and freshwater 
aquatic sediment. Environ. Sci. Technol. 40:5395-5401. Including 
supporting information document.
96. Davis, J.W., Gonsior, S.J., Marty, G.T., et al. 2005. The 
transformation of hexabromocyclododecane in aerobic and anaerobic 
soils and aquatic sediments. Water Res. 39:1075-1084.
97. Hazardous Substance Data Bank. 2011. 1,2,5,6,9,10-
Hexabromocyclododecane. Hazardous Substances Data Bank. Part of the 
National Library of Medicine's Toxicology Data Network (TOXNET7). 
Bethesda, MD. Available online at http://toxnet.nlm.nih.gov/cgi-bin/sis/htmlgen?HSDB (accessed May 31, 2011).
98. Bidleman, T.F. 1988. Atmospheric processes. Environ. Sci. 
Technol. 22(4):361-367.
99. Stenzel, J.I., Nixon, W.B. 1997. Hexabromocyclododecane (HBCD): 
Determination of the vapor pressure using a spinning rotor gauge 
with cover letter dated 08/15/1997. Chemical Manufacturers 
Association. Submitted under TSCA Section 8D. OTS0573702.
100. USEPA. 1993. Determination of rates of reaction in the gas-
phase in the troposphere. 5. Rate of indirect photoreaction: 
Evaluation of the atmospheric oxidation computer program of Syracuse 
Research Corporation for estimating the second-order rate constant 
for the reaction of an organic chemical with hydroxyl radicals. 
Washington, DC: U.S. Environmental Protection Agency. EPA744R93001.
101. National Industrial Chemicals Notification and Assessment 
Scheme. 2012. Hexabromocyclododecane. Priority existing chemical 
assessment report. Volume 34. Commonwealth of Australia: Australia. 
National Industrial Chemicals Notification and Assessment Scheme. 
PEC34.
102. Davis, J.W., Gonsior, S.J., Marty, G.T. 2003. Evaluation of 
aerobic and anaerobic transformation of hexabromocyclododecane in 
aquatic sediment systems. Project Study ID 021081, 87 pp. DOW 
Chemical Company: Midland, MI, USA. Submitted under TSCA Section 
FYI; EPA Document No. 84040000010; FYI-1103-01472, pg. 440.
103. Davis, J.W., Gonsior, S.J., Marty, G.T. 2003. Evaluation of 
aerobic and anaerobic transformation of hexabromocyclododecane in 
soil. Project Study ID 021082, 61 pp. DOW Chemical Company: Midland, 
MI, USA. Submitted under TSCA Section FYI; EPA Document No. 
84040000010; FYI-1103-01472, pg. 379.
104. USEPA. 2010. Hexabromocyclododecane (HBCD) action plan. U.S. 
Environmental Protection Agency. August 18, 2010.

VI. What are the Statutory and Executive Orders reviews associated with 
this action?

    Additional information about these statutes and Executive Orders 
can be found at http://www2.epa.gov/laws-regulations/laws-and-executive-orders.

[[Page 35289]]

A. Executive Order 12866: Regulatory Planning and Review and Executive 
Order 13563: Improving Regulation and Regulatory Review

    This action is not a significant regulatory action and was 
therefore not submitted to the Office of Management and Budget (OMB) 
for review under Executive Orders 12866 (58 FR 51735, October 4, 1993) 
and 13563 (76 FR 3821, January 21, 2011).

B. Paperwork Reduction Act (PRA)

    This action does not contain any new information collection 
requirements that require additional approval by OMB under the PRA, 44 
U.S.C. 3501 et seq. OMB has previously approved the information 
collection activities contained in the existing regulations and has 
assigned OMB control numbers 2025-0009 and 2050-0078. Currently, the 
facilities subject to the reporting requirements under EPCRA section 
313 and PPA section 6607 may use either EPA Toxic Chemicals Release 
Inventory Form R (EPA Form 1B9350-1), or EPA Toxic Chemicals Release 
Inventory Form A (EPA Form 1B9350- 2). The Form R must be completed if 
a facility manufactures, processes, or otherwise uses any listed 
chemical above threshold quantities and meets certain other criteria. 
For the Form A, EPA established an alternative threshold for facilities 
with low annual reportable amounts of a listed toxic chemical. A 
facility that meets the appropriate reporting thresholds, but estimates 
that the total annual reportable amount of the chemical does not exceed 
500 pounds per year, can take advantage of an alternative manufacture, 
process, or otherwise use threshold of 1 million pounds per year of the 
chemical, provided that certain conditions are met, and submit the Form 
A instead of the Form R. Since the HBCD category would be classified a 
PBT category, it is designated as a chemical of special concern, for 
which Form A reporting is not allowed. In addition, respondents may 
designate the specific chemical identity of a substance as a trade 
secret pursuant to EPCRA section 322, 42 U.S.C. 11042, 40 CFR part 350.
    OMB has approved the reporting and recordkeeping requirements 
related to Forms A and R, supplier notification, and petitions under 
OMB Control number 2025-0009 (EPA Information Collection Request (ICR) 
No. 1363) and those related to trade secret designations under OMB 
Control 2050-0078 (EPA ICR No. 1428). As provided in 5 CFR 1320.5(b) 
and 1320.6(a), an Agency may not conduct or sponsor, and a person is 
not required to respond to, a collection of information unless it 
displays a currently valid OMB control number. The OMB control numbers 
relevant to EPA's regulations are listed in 40 CFR part 9 or 48 CFR 
chapter 15, and displayed on the information collection instruments 
(e.g., forms, instructions).

C. Regulatory Flexibility Act (RFA)

    I certify that this action will not have a significant economic 
impact on a substantial number of small entities under the RFA, 5 
U.S.C. 601 et seq. The small entities subject to the requirements of 
this action are small manufacturing facilities. The Agency has 
determined that of the 55 entities estimated to be impacted by this 
action, 42 are small businesses; no small governments or small 
organizations are expected to be affected by this action. All 42 small 
businesses affected by this action are estimated to incur annualized 
cost impacts of less than 1%. Thus, this action is not expected to have 
a significant adverse economic impact on a substantial number of small 
entities. A more detailed analysis of the impacts on small entities is 
located in EPA's economic analysis (Ref. 2).

D. Unfunded Mandates Reform Act (UMRA)

    This action does not contain an unfunded mandate of $100 million or 
more as described in UMRA, 2 U.S.C. 1531-1538, and does not 
significantly or uniquely affect small governments. This action is not 
subject to the requirements of UMRA because it contains no regulatory 
requirements that might significantly or uniquely affect small 
governments. Small governments are not subject to the EPCRA section 313 
reporting requirements. EPA's economic analysis indicates that the 
total cost of this action is estimated to be $372,973 in the first year 
of reporting (Ref. 2).

E. Executive Order 13132: Federalism

    This action does not have federalism implications as specified in 
Executive Order 13132 (64 FR 43255, August 10, 1999). It will not have 
substantial direct effects on the States, on the relationship between 
the national government and the States, or on the distribution of power 
and responsibilities among the various levels of government.

F. Executive Order 13175: Consultation and Coordination With Indian 
Tribal Governments

    This action does not have tribal implications as specified in 
Executive Order 13175 (65 FR 67249, November 9, 2000). This action 
relates to toxic chemical reporting under EPCRA section 313, which 
primarily affects private sector facilities. Thus, Executive Order 
13175 does not apply to this action.

G. Executive Order 13045: Protection of Children From Environmental 
Health Risks and Safety Risks

    EPA interprets Executive Order 13045 (62 FR 19885, April 23, 1997) 
as applying only to those regulatory actions that concern environmental 
health or safety risks that EPA has reason to believe may 
disproportionately affect children, per the definition of ``covered 
regulatory action'' in section 2-202 of the Executive Order. This 
action is not subject to Executive Order 13045 because it does not 
concern an environmental health risk or safety risk.

H. Executive Order 13211: Actions Concerning Regulations That 
Significantly Affect Energy Supply, Distribution, or Use

    This action is not subject to Executive Order 13211 (66 FR 28355, 
May 22, 2001), because it is not a significant regulatory action under 
Executive Order 12866.

I. National Technology Transfer and Advancement Act (NTTAA)

    This rulemaking does not involve technical standards and is 
therefore not subject to considerations under section 12(d) of NTTAA, 
15 U.S.C. 272 note.

J. Executive Order 12898: Federal Actions To Address Environmental 
Justice in Minority Populations and Low-Income Populations

    EPA has determined that this action will not have 
disproportionately high and adverse human health or environmental 
effects on minority or low-income populations as specified in Executive 
Order 12898 (59 FR 7629, February 16, 1994). This action does not 
address any human health or environmental risks and does not affect the 
level of protection provided to human health or the environment. This 
action adds an additional chemical to the EPCRA section 313 reporting 
requirements. By adding a chemical to the list of toxic chemicals 
subject to reporting under section 313 of EPCRA, EPA would be providing 
communities across the United States (including minority populations 
and low income populations) with access to data which they may use to 
seek lower exposures and consequently reductions in chemical risks for 
themselves and their children. This information can also be used by 
government agencies and others to identify potential problems, set 
priorities, and take appropriate steps to

[[Page 35290]]

reduce any potential risks to human health and the environment. 
Therefore, the informational benefits of the action will have positive 
human health and environmental impacts on minority populations, low-
income populations, and children.

List of Subjects in 40 CFR Part 372

    Environmental protection, Community right-to-know, Reporting and 
recordkeeping requirements, and Toxic chemicals.

    Dated: May 16, 2016.
Gina McCarthy,
Administrator.
    Therefore, it is proposed that 40 CFR chapter I be amended as 
follows:

PART 372--[AMENDED]

0
1. The authority citation for part 372 continues to read as follows:

    Authority:  42 U.S.C. 11023 and 11048.


0
2. In Sec.  372.28, amend the table in paragraph (a)(2) as follows:
0
a. Revise the heading for the second column, and
0
b. Alphabetically add the category ``Hexabromocyclododecane (This 
category includes only those chemicals covered by the CAS numbers 
listed here)'' and list ``3194-55-6 (1,2,5,6,9,10-
Hexabromocyclododecane)'' and ``25637-99-4 (Hexabromocyclododecane)''
    The additions to read as follows:


Sec.  372.28  Lower thresholds for chemicals of special concern.

    (a) * * *
    (2) * * *

------------------------------------------------------------------------
                                                            Reporting
                                                          threshold (in
                     Category name                        pounds unless
                                                        otherwise noted)
------------------------------------------------------------------------
 
                              * * * * * * *
Hexabromocyclododecane (This category includes only                  100
 those chemicals covered by the CAS numbers listed
 here)................................................
 3194-55-6 1,2,5,6,9,10-Hexabromocyclododecane........  ................
25637-99-4 Hexabromocyclododecane.....................  ................
 
                              * * * * * * *
------------------------------------------------------------------------

* * * * *
0
3. In Sec.  372.65, paragraph (c) is amended by adding alphabetically 
an entry for ``Hexabromocyclododecane (This category includes only 
those chemicals covered by the CAS numbers listed here)'' to the table 
to read as follows:


Sec.  372.65  Chemicals and chemical categories to which this part 
applies.

* * * * *
    (c) * * *

------------------------------------------------------------------------
                     Category name                       Effective date
------------------------------------------------------------------------
 
                              * * * * * * *
Hexabromocyclododecane (This category includes only               1/1/17
 those chemicals covered by the CAS numbers listed
 here)................................................
 3194-55-6 1,2,5,6,9,10-Hexabromocyclododecane........  ................
25637-99-4 Hexabromocyclododecane.....................  ................
 
                              * * * * * * *
------------------------------------------------------------------------

[FR Doc. 2016-12464 Filed 6-1-16; 8:45 am]
 BILLING CODE 6560-50-P