1260 volume 120 | number 9 | September 2012 • Environmental Health Perspectives

Research

Hexabromocyclododecane (HBCD) is a 

brominated aliphatic cyclic hydrocarbon flame 

retardant (BFR) used in polystyrene foam for 

thermal insulation in buildings, upholstery textiles, and electrical equipment (Covaci et al. 

2006). HBCD is a major BFR produced globally, and before the phase-out of certain polybrominated diphenyl ethers (PBDEs), it was the 

third most abundant BFR in North America, 

after tetrabromobisphenol A and PBDEs 

(Johnson-Restrepo et al. 2008). HBCD enters 

the environment during its production and 

by leaching from consumer products (Covaci 

et al. 2006). Human exposure occurs through 

dust inhalation and dust and food ingestion 

(Abdallah and Harrad 2009; Schecter et al. 

2009; Thomsen et al. 2008). HBCD is hydrophobic and lipophilic, and it bioaccumulates 

with half-lives of 2, 60, and 240 days in air, 

water, and sediment, respectively (Abdallah and 

Harrad 2009; Germer et al. 2006; Haukås et al. 

2009; Johnson-Restrepo et al. 2008; Kunisue 

et al. 2007; Thomsen et al. 2008; Törnkvist 

et al. 2011). Terminal elimination half-lives 

in nonoccupationally exposed humans were 

calculated to be on average 64 days (range 

23–219 days) (Geyer et al. 2004), using the 

daily intake for total HBCD in adult humans 

based on a Swedish market basket study (Lind 

et al. 2002). The authors attributed the high 

variability observed in half-lives to differences 

in exposure, total body fat content, and biological differences that affect bioaccumulation, 

uptake, metabolism, and elimination. Abdallah 

and Harrad (2011) calculated a human half-life 

of 165 days for α-HBCD (representing 75% 

of the maximum half-life of 219 days for the 

HBCD mixture), whereas a half-life of 55 days 

was estimated for the β- and γ-isomers (25% of 

219 days).

HBCD is categorized as a persistent organic 

pollutant (POP) because of its persistence, 

toxicity, and abilities to bioaccumulate and to 

travel over long distances. It is currently on 

the European Chemicals Agency candidate list 

of substances of very high concern (European 

Chemicals Agency 2010). The U.S. Environmental Protection Agency (EPA) has developed 

an action plan for HBCD and is considering 

adding it to its list of “chemicals of concern” 

(U.S. EPA 2010).

Increased attention has focused on the 

individual stereoisomers that constitute the 

HBCD commercial mixture. In laboratory 

animal studies, individual stereoisomers that 

compose the commercial mixture have been 

shown to have different biological properties 

(Szabo et al. 2010, 2011a). There are three 

main stereoisomers in the commercial 

mixture: α-HBCD (10–13%), β-HBCD 

(1–12%), and γ-HBCD (75–89%) (Heeb 

et al. 2005). γ-HBCD dominates in the 

commercial mixture and in the environment, whereas biota show a predominance of 

α-HBCD (Arsenault et al. 2007; Law et al. 

2008). The largest tissue depots found after 

oral exposure to γ-HBCD in mice included 

the liver, brain, blood, and fat; however, 

tissue levels were low because of γ-HBCD’s 

rapid metabolism and elimination (Szabo 

et al. 2010). For γ-HBCD, the mouse biological half-life was estimated at 1–4 days with 

Address correspondence to A. Schecter, University 

of Texas School of Public Health, 6011 Harry Hines 

Blvd., V8.122E, Dallas, TX 75390-9128 USA. 

Telephone: (214) 336-8519. Fax: (214) 648-1081. 

E-mail: arnold.schecter@utsouthwestern.edu

We thank J. Huwe and H. Hakk of the U.S. 

Department of Agriculture for their expertise and 

generous time in advising the authors on various 

aspects of this manuscript.

This study was partially funded by the Gustavus 

and Louise Pfeiffer Research Foundation and partially funded by the National Institute of Environmental Health Sciences (NIEHS). Support in part 

was also provided by an appointment to the Research 

Participation Program for the U.S. Environmental 

Protection Agency (EPA), Office of Research and 

Development, administered by the Oak Ridge 

Institute for Science and Education through an interagency agreement between the U.S. Department 

of Energy and the U.S. EPA. M. Petersen and 

O. Paepke are both employed by Eurofins Gfa 

GmbH Laboratory, Hamburg, Germany. Support for 

J. Colacino was provided by training grants from the 

NIEHS (T32 ES007062) and the National Human 

Genome Research Institute (T32 HG00040). 

This manuscript does not reflect policies of the 

U.S. EPA, NIEHS, National Cancer Institute, or 

National Institutes of Health. This manuscript 

reflects only the conclusions of the authors.

The authors declare they have no actual or potential competing financial interests.

Received 20 January 2012; accepted 18 May 2012.

Hexabromocyclododecane (HBCD) Stereoisomers in U.S. Food 

from Dallas, Texas

Arnold Schecter,1 David T. Szabo,2 James Miller,1 Tyra L. Gent,1 Noor Malik-Bass,1 Malte Petersen,3 Olaf Paepke,3

Justin A. Colacino,4 Linda S. Hynan,5 T. Robert Harris,1 Sunitha Malla,1 and Linda S. Birnbaum6,7

1University of Texas School of Public Health, Dallas, Texas, USA; 2National Center for Environmental Assessment, U.S. Environmental 

Protection Agency, Arlington, Virginia, USA; 3Eurofins Gfa GmbH Laboratory, Hamburg, Germany; 4Department of Environmental 

Health Sciences, University of Michigan School of Public Health, Ann Arbor, Michigan, USA; 5Department of Clinical Sciences and 

Department of Psychiatry, University of Texas Southwestern Medical Center, Dallas, Texas, USA; 6National Cancer Institute, National 

Institutes of Health, Department of Health and Human Services, Washington, DC, USA; 7National Institute of Environmental Health 

Sciences, National Institutes of Health, Department of Health and Human Services, Research Triangle Park, North Carolina, USA

Background: Hexabromocyclododecane (HBCD) is a brominated flame retardant used in 

polystyrene foams in thermal insulation and electrical equipment. The HBCD commercial mixture 

consists mainly of α, β, and γ stereoisomers. Health concerns of HBCD exposure include alterations 

in immune and reproductive systems, neurotoxic effects, and endocrine disruption. Stereoisomerspecific levels of HBCD have not been measured previously in U.S. food.

Objectives: We measured HBCD stereoisomer levels in U.S. foods from Dallas, Texas, 

supermarkets.

Methods: Convenience samples of commonly consumed foods were purchased from supermarkets 

in Dallas in 2009–2010. Food samples included a wide variety of lipid-rich foods: fish, peanut butter, poultry, pork, and beef. Thirty-six individual food samples were collected in 2010 and analyzed 

for α-, β-, and γ-HBCD stereoisomers using liquid chromatography–tandem mass spectrometry 

(LC-MS/MS). Ten pooled food samples previously collected in 2009 for a study of total HBCD 

levels using gas chromatography–mass spectrometry (GC-MS), were reanalyzed for α-, β-, and 

γ-HBCD stereoisomers using LC-MS/MS.

Results: Of the 36 measured individual foods, 15 (42%) had detectable levels of HBCD. 

Median (ranges) of α- and γ-HBCD concentrations were 0.003 (< 0.005–1.307) and 0.005 

(< 0.010–0.143) ng/g wet weight (ww), respectively; β-HBCD was present in three samples 

with a median (range) of 0.003 (< 0.005–0.019) ng/g ww. Median levels (range) for α-, β-, and 

γ-HBCD, in pooled samples were 0.077 (0.010–0.310), 0.008 (< 0.002–0.070), and 0.024 

(0.012–0.170) ng/g ww, respectively.

Conclusions: α-HBCD was detected most frequently and at highest concentrations, followed by 

γ-, and then β-HBCD, in food samples from Dallas, Texas. Food may be a substantial contributor 

to the elevated α-HBCD levels observed in humans. These data suggest that larger and more representative sampling should be conducted.

Key words: Dallas, Texas; food; HBCD; hexabromocyclododecane; stereoisomers. Environ Health 

Perspect 120:1260–1264 (2012). http://dx.doi.org/10.1289/ehp.1204993 [Online 31 May 2012]
HBCD stereoisomers in food from Dallas, Texas

Environmental Health Perspectives • volume 120 | number 9 | September 2012 1261

limited capability for bioaccumulation (Szabo 

et al. 2010). This is in contrast to α-HBCD, 

which has a biological half-life of 17 days in 

mice. α-HBCD bioaccumulates and its tissue 

distribution is dictated by lipophilicity, with 

higher levels being detected in mouse adipose, 

liver, muscle, and skin tissues (Szabo et al. 

2011a). α-HBCD predominates in biota 

in part because of the rapid metabolism of 

γ-HBCD to α-HBCD (Szabo et al. 2010). 

Toxicity studies suggest that the commercial 

mixture of HBCD is an endocrine disruptor 

and developmental neurotoxicant; specifically, 

HBCD commercial mixtures have been 

associated with changes in rat thyroid systems 

(Palace et al. 2010; Saegusa et al. 2009), 

altered function of human natural killer 

cells (Hinkson and Whalen 2009, 2010), 

and neurotoxic effects such as decreased fine 

manipulative abilities and lower attention in 

children (Roze et al. 2009). Studies on the 

health effects of individual stereoisomers are 

extremely limited. The difficulty of analytical 

separation, quantities, and purity needed 

for both in vivo and in vitro experiments 

have limited the progress of toxicity studies 

examining individual HBCD stereoisomers.

Because of the different physical, chemical, 

and biological properties of the HBCD stereoisomers, there is a growing demand to design 

studies characterizing individual HBCD stereoisomers. Although studies have investigated 

HBCD in the environment (Haukås et al. 

2009; Law et al. 2008), dust (Harrad et al. 

2009), human serum (Roosens et al. 2009), 

human milk (Ryan et al. 2006), and animal 

tissues (Johnson-Restrepo et al. 2008; Zegers 

et al. 2005), few have examined HBCD contamination of food. Such information is notably sparse for foods found in the United States. 

We previously reported “total HBCD” levels 

in pooled samples of U.S. food but did not differentiate between stereoisomers (Schecter et al. 

2009). Because ingestion is believed to be a 

major exposure route for HBCD (Covaci et al. 

2006), and to examine stereoisomer-specific 

dietary concentrations from food consumed 

by the U.S. population, we report here, for 

the first time, concentrations of α-, β-, and 

γ-HBCD stereoisomers in selected foods from 

supermarkets in Dallas, Texas.

Material and Methods

Sample collection. In this study, food samples 

were collected in Dallas, Texas, in 2009 and 

2010. Further study will be needed to determine representative levels for U.S. food in 

general.

Analysis of individual food samples collected in 2010 for HBCD stereoisomers used 

liquid chromatography/electrospray negative 

ionization with tandem mass spectrometry 

(LC-MS/MS). A convenience collection of 

foods (n = 36) was purchased in 2010 from 

supermarkets in Dallas. The 2010 selection 

of foods was based on findings of quantifiable 

HBCD levels in our previous study of total 

HBCD concentration in U.S. food (Schecter 

et al. 2009). Food sampled in this analysis 

included fish, peanut butter, turkey, pork, and 

beef. After collection, the foods were frozen 

at –80°C and sent on dry ice to Eurofins Gfa 

GmbH Laboratory (Hamburg, Germany) for 

stereoisomer-specific HBCD analysis. Each 

food sample was individually analyzed for 

HBCD stereoisomer levels using LC-MS/MS.

Reanalysis of pooled U.S. food samples collected in 2009 for HBCD stereoisomers used 

LC-MS/MS. In our previous study (Schecter 

et al. 2009), 31 pooled samples from Dallas 

supermarkets were analyzed for “total HBCD” 

(Schecter et al. 2009). The foods selected were 

a convenience sample of commonly consumed foods purchased at local supermarkets, 

including fish, peanut butter, and poultry. 

Samples were frozen at –80oC and shipped 

on dry ice to Hamburg, Germany, and stored 

in a deep freezer in Hamburg until analyzed 

by gas chromatography–mass spectroscopy 

(GC-MS). Thirteen of the 31 (42%) pooled 

samples contained detectable levels of HBCD. 

Of these 13 pooled samples, 10 had adequate 

amounts of sample remaining for additional 

analysis and were chosen for stereoisomerspecific analysis to validate and compare to 

our previous findings of total HBCD levels. Reanalysis of the 10 pooled 2009 food 

samples for HBCD stereoisomers was conducted using LC-MS/MS.

Chemical analyses. The isotope dilution 

analysis is a technique used to provide the 

best precision and accuracy for these chemical analyses. Three native HBCD standards 

(α-, β-, and γ-HBCD stereoisomers; purity 

of ≥ 98%) and one carbon-13 (13C)-labeled 

standard (γ-HBCD) were obtained from 

Wellington Laboratories, (Guelph, Ontario, 

Canada). The internal standard, 13C-labeled 

γ-HBCD, was added to the homogenized fraction. Soxhlet extraction was performed with 

hexane:acetone (4:1). The lipid extract was 

separated, further purified with sulfuric acid, 

then cleaned up with aluminum oxide (2% 

water). The final extract was reduced and dried 

under a light stream of nitrogen and then dissolved in 100 μL methanol for LC-MS/MS

analysis. The measurements were performed 

using LC-MS/MS detection (Varian 1200L 

LC-MS/MS Triple Quadrupole) with solvents water/acetone/methanol (H2O/ACN/

MeOH; 20/30/50% vol/vol) [solution A (A)] 

and ACN/MeOH (30/70% vol/vol) [solution B (B)]. The gradient program consisted 

of 2 min at 100% (A)/0% (B), ramped to 

25% (A)/75% (B) in 13.5 min and held for 

4 min, then ramped to 100% (A)/0% (B) in 

1 min and held for 6 min. A C18 guard cartridge [4 mm × 2.0 mm inner diameter (i.d.); 

Phenomenex, Torrance, CA, USA] and a 

Synergy 4u Fusion RP C-18 column (100 mm 

× 2.0 mm i.d., 80A; Phenomenex) were used 

for liquid chromatographic separation.

Quality assurance/quality control measures 

included a multipoint calibration curve, recalibration of each sequence of analyses (with 

a minimum of one blank in each batch of 

a maximum of 10 samples), and duplicate 

analyses of approximately 20% of the positive 

samples. Acceptable reproducibility/accuracy 

was determined by analyzing a laboratory 

reference sample of pooled fish oil. The ratio 

between the assigned value and the determined 

value (i.e., recovery) ranged from 87–110% 

for α-HBCD (mean 96.5%; relative standard 

deviation of 8.1%; n = 7); and recovery rates 

ranged between 80–113% for γ-HBCD (mean 

93.3%; reproducibility variation coefficient of 

10.7%; n = 7). The concentration of β-HBCD 

in this reference sample was below the limit 

of quantification, and therefore recovery rates 

were unobtainable.

Total HBCD levels were calculated as the 

sum of α-, β-, and γ-HBCD stereoisomer 

levels for each individual food examined, 

and analyzed with the values below the limit 

of detection (LOD) set to zero as well as to 

one-half the LOD (LOD/2). We have elected 

to report values below the LOD as equal to 

LOD/2 instead of zero because the results 

were similar.

Statistical analysis. Continuous measures 

are described using medians, range, and means. 

To estimate the total distribution of HBCD 

stereoisomers in U.S. food, we calculated the 

proportion of the three HBCD stereoisomers in 

each individually analyzed sample as the mean 

stereoisomer divided by the mean of the total 

HBCD with nondetected food values estimated 

as LOD/2, weighting each food sample equally. 

We used IBM SPSS software, version 19 (SPSS 

Inc., Chicago, IL, USA) to analyze these data.

Results

Table 1 shows stereoisomer-specific HBCD 

levels for 36 individually analyzed foods 

purchased in 2010. For all foods except 

peanut butter, the LODs for each of the 

HBCD stereoisomers are 0.005 ng/g for 

α- and β-HBCD and 0.010 ng/g for 

γ-HBCD (Table 1). LODs for peanut butter 

were 0.020 ng/g for α- and β-HBCD and 

0.040 ng/g for γ-HBCD. Fifteen of 36 individual food samples (42%) had at least one 

detectable stereoisomer. α-HBCD was present 

in 13 samples (36%) with a median (range) of 

0.003 (< 0.005–1.307) ng/g wet weight (ww); 

β-HBCD was present in 3 samples (8%) with 

a median of 0.003 (< 0.005–0.019) ng/g ww; 

γ-HBCD was present in 8 samples (22%) with 

a median of 0.005 (< 0.010–0.143) ng/g ww 

with median values calculated after setting 

nondetectable levels to LOD/2. Total HBCD 
Schecter et al.

1262 volume 120 | number 9 | September 2012 • Environmental Health Perspectives

levels for these stereoisomers, calculated as the 

sum of the three stereoisomers, had a median 

of 0.012 (< LOD to 1.366) ng/g ww. The 

mean proportion of each HBCD stereoisomer 

in all 36 samples was 78.0% α-HBCD, 3.8% 

β-HBCD, and 18.1% γ-HBCD; the mean 

proportion of each HBCD stereoisomer for 

the 15 food samples with at least one detectable stereoisomer was 93.6% α-HBCD, 5.3% 

β-HBCD, and 1.8% γ-HBCD. Of all the 

foods studied, canned sardines stand out with 

an α-HBCD level of 1.307 ng/g ww. Smoked 

turkey sausages had the next highest α-HBCD 

level of 0.479 ng/g ww. HBCD levels were 

< LOD for fresh deli meats and fish, chili with 

beans, and bacon.

Table 2 presents stereoisomer-specific analysis of 10 pooled food samples purchased in 

2009, as well as total HBCD levels previously 

measured by GC-MS (Schecter et al. 2009). 

Total HBCD levels as measured by GC-MS 

and total levels calculated by adding individual 

stereoisomer levels by LC-MS/MS were similar, with < 0.051-ng/g ww difference between 

estimates for 9 of the 10 samples. Both α- and 

γ-HBCD were detected in all 10 analyzed 

samples, whereas β-HBCD was detected in 8 of 

10 analyzed samples, with nondetectable levels of 0.004 and 0.002 ng/g for fresh catfish 

and canned chili, respectively. Levels for α-, 

β-, and γ-HBCD [median (range)] were 0.077 

(0.010–0.310), 0.008 (< 0.002–0.070), and 

0.024 (0.012–0.170) ng/g ww, respectively, with 

median values calculated after setting nondetectable levels to LOD/2. Total levels as measured by 

GC-MS were higher than those calculated from 

LC-MS/MS with two exceptions: a) sliced ham, 

where the totals calculated from LC-MS/MS 

were higher, and b) canned chili, where the 

totals were equal.

Table 3 lists available data on stereoisomerspecific HBCD levels (reported on a wetweight basis) in foods worldwide. Although 

comparative data for specific foods in each 

country are not available, it is notable that 

Romania (0.040–0.250 ng/g ww), Sweden 

(0.005–0.630 ng/g ww), United Kingdom 

(0.104–0.730 ng/g ww), and Norway 

(0.015–1-0.769 ng/g ww) have reported some 

of the lowest total HBCD levels (Dirtu and 

Covaci 2010; Driffield et al. 2008; Knutsen 

et al. 2008; Törnkvist et al. 2011), and the 

Netherlands has reported the highest seafood 

total HBCD level (0.200–230.0 ng/g ww) 

(van Leeuwen and de Boer 2008). Higher 

HBCD levels have been measured in food 

samples from the Netherlands, Japan, and 

Scotland (Fernandes et al. 2008; Nakagawa 

et al. 2010; van Leeuwen and de Boer 2008) 

than in samples tested in the present study.

Discussion

Although γ-HBCD is the dominant stereoisomer in commercial HBCD mixtures, 

α-HBCD was the main stereoisomer found in 

23 of 46 analyses (10 composite and 13 individual) as has also been reported in other food 

studies and in biota (Driffield et al. 2008; 

Knutsen et al. 2008; Law et al. 2008; Nakagawa 

et al. 2010). This stereoisomer shift from what 

is in commercial mixtures has been hypothesized to be due to several reasons, including 

increased thermodynamic stability (Heeb et al. 

2010), biological stability, and persistence of 

the α-HBCD stereoisomer (Szabo et al. 2011a) 

as compared to γ-HBCD. Rapid metabolism of 

Table 1. HBCD levels in individual U.S. food samples purchased and analyzed in 2010 using LC-MS/MS 

(ng/g ww) (n = 36).

Stereoisomer levelsa

Food Totalb α β γ

Sardines in water 1.366 1.307 < 0.005 0.056

Smoked turkey sausages 0.518 0.479 < 0.005 0.036

Fresh salmon, store A 0.446 0.327 < 0.005 0.116

Fresh salmon, store B 0.410 0.346 < 0.005 0.061

Sardines in olive oil #1 0.270 0.262 < 0.005 < 0.010

Fresh catfish, store B 0.162 0.115 0.016 0.031

Fresh deli sliced turkey, store B 0.155 0.135 < 0.005 0.017

Fresh tilapia, store A 0.148 < 0.005 < 0.005 0.143

Chili with beans #1 0.105 < 0.005 < 0.005 0.100

Sardines in olive oil #2 0.067 0.059 < 0.005 < 0.010

Fresh deli sliced ham, store C 0.051 0.027 0.019 < 0.010

Smoked turkey sausages #1 0.049 0.029 0.015 < 0.010

Creamy peanut butter #1 0.040 < 0.020 < 0.020 < 0.040

Creamy peanut butter #2 0.040 < 0.020 < 0.020 < 0.040

Creamy peanut butter #3 0.040 < 0.020 < 0.020 < 0.040

Smoked turkey sausages #2 0.032 0.024 < 0.005 < 0.010

Fresh catfish, store A 0.016 0.008 < 0.005 < 0.010

Fresh catfish, store C 0.014 0.006 < 0.005 < 0.010

Chili with beans #2 0.010 < 0.005 < 0.005 < 0.010

Chili with beans #3 0.010 < 0.005 < 0.005 < 0.010

Bacon #1 0.010 < 0.005 < 0.005 < 0.010

Bacon #2 0.010 < 0.005 < 0.005 < 0.010

Bacon #3 0.010 < 0.005 < 0.005 < 0.010

Fresh deli sliced beef, store C 0.010 < 0.005 < 0.005 < 0.010

Fresh deli sliced turkey, store C 0.010 < 0.005 < 0.005 < 0.010

Fresh deli sliced chicken, store C 0.010 < 0.005 < 0.005 < 0.010

Fresh deli sliced beef, store A 0.010 < 0.005 < 0.005 < 0.010

Fresh deli sliced ham, store A 0.010 < 0.005 < 0.005 < 0.010

Fresh deli sliced turkey, store A 0.010 < 0.005 < 0.005 < 0.010

Fresh deli sliced chicken, store B 0.010 < 0.005 < 0.005 < 0.010

Fresh tilapia, store C 0.010 < 0.005 < 0.005 < 0.010

Fish sticks #1 0.010 < 0.005 < 0.005 < 0.010

Fish sticks #2 0.010 < 0.005 < 0.005 < 0.010

Fish sticks #3 0.010 < 0.005 < 0.005 < 0.010

Fresh tilapia, store C 0.010 < 0.005 < 0.005 < 0.010

Fresh deli sliced beef, store C 0.010 < 0.005 < 0.005 < 0.010

Median 0.012 0.003 0.003 0.005

Mean 0.114 0.089 0.004 0.021

aLODs based on a 20-g sample size. bTotal HBCD levels calculated as sum of stereoisomers with values < LOD set to 

LOD/2; mean and median values also calculated after setting values < LOD to LOD/2.

Table 2. HBCD levels in U.S. food samples purchased in 2009 (Schecter et al. 2009) analyzed using GC-MS 

and reanalyzed using LC-MS/MS in 2010 (n = 10) (ng/g ww).

Totala

(GC-MS)b

Total 

(LC-MS/MS)

Stereoisomer levels

Food α β γ

Canned sardines 0.593 0.550 0.310 0.070 0.170

Fresh salmon 0.352 0.123 0.099 0.005 0.019

Peanut butter 0.300 0.282 0.223 0.007 0.052

Fresh tilapia 0.180 0.130 0.090 0.010 0.030

Sausages 0.151 0.119 0.078 0.013 0.028

Fresh catfish 0.133 0.109 0.050 < 0.004c 0.057

Sliced turkey 0.124 0.113 0.076 0.019 0.018

Sliced chicken breast 0.098 0.048 0.025 0.008 0.015

Sliced ham 0.024 0.036 0.012 0.006 0.018

Canned chili 0.023 0.023 0.010 < 0.002c 0.012

Median 0.142 0.116 0.077 0.008 0.024

aTotal HBCD levels measured by laboratory and not calculated as sum of stereoisomers. bData from Schecter et al. 2009. cValues < LOD.
HBCD stereoisomers in food from Dallas, Texas

Environmental Health Perspectives • volume 120 | number 9 | September 2012 1263

γ-HBCD in vivo (Szabo et al. 2010) and in vitro

(Zegers et al. 2005), along with biological 

stereoisomerization of γ-HBCD to α-HBCD 

(Szabo et al. 2010), is also suspected to be a 

contributing factor to the decreased levels of 

γ-HBCD and increased levels of α-HBCD in 

biota, and subsequently animal-based foods 

consumed in the United States and elsewhere.

Stereoisomer-specific concentrations of 

HBCD measured in this U.S. food study were 

lower than HBCD levels measured in food 

samples from some other countries (Fernandes 

et al. 2008; Nakagawa et al. 2010; van Leeuwen 

and de Boer 2008). Some of the foods tested 

for our study may not have been as lipid rich as 

the foods tested for other studies (for example, 

certain types of fish), which could account for 

some disparities (Hayward et al. 2006). Higher 

HBCD food levels reported in other studies 

may also be due to a greater use of HBCD in 

those countries (Morose 2006). In the present 

study, we did not note an association between 

higher lipid levels and higher HBCD levels 

(results not shown).

HBCD stereoisomer levels reported for 

the present study and for comparisons with 

other studies are reported on a wet weight 

basis, which is more representative of food 

as eaten then food levels reported on a lipid 

basis (Goscinny et al. 2011). In contrast, we 

and many authors report measured values of 

lipid-soluble compounds in biologic samples 

from humans on a lipid basis to reflect exposure and body burden. In comparing studies it is important to note whether the data 

is reported on a wet weight or lipid basis 

because levels of the lipophilic HBCDs will be 

higher when reported on a lipid basis. Several 

other studies have also reported evidence suggesting that lipid-rich fish and meats are a 

major source of HBCD exposure from foods 

(Dirtu and Covaci 2010; Driffield et al. 2008; 

Fernandes et al. 2008; Knutsen et al. 2008; 

Nakagawa et al. 2010; Törnkvist et al. 2011; 

van Leeuwen and de Boer 2008). However, it 

is also becoming clear that HBCD contamination is not restricted to fish or meat.

LC-MS/MS measurements confirmed the 

high levels of HBCD previously reported in a 

sample of peanut butter assayed in 2009 using 

GC-MS (300 and 282 ng/g ww, respectively) 

(Schecter et al. 2009). Driffield et al. (2008) 

also detected HBCD in nuts bought in the 

United Kingdom. Individual PBDE and 

polychlorinated biphenyl (PCB) congeners, 

pesticides, and a number of perfluorinated 

compounds (PFCs) were found at much lower 

levels than HBCD in the same pooled peanut 

butter samples (Schecter et al. 2009, 2010a). 

However, HBCD stereoisomer levels were 

nondetectable in the three individual peanut 

butter samples tested for the present study. 

There are many ways that the differing levels 

found in individual and pooled peanut butter 

samples may be explained. HBCD found in 

dust may contaminate foods during preparation in the indoor environment (Abdallah and 

Harrad 2009). Also, HBCD in soil has been 

shown to be transferred into vegetables in some 

instances (Li et al. 2011). Heavy metal accumulation in plants can serve as an indicator of 

soil contamination; chromium, copper, and 

zinc have been transferred to and accumulated 

in the roots and leaves of peanut plants grown 

in soils contaminated with those metals (Ching 

et al. 2008). It may be possible that uptake 

of HBCD from soil by peanuts from select 

regions might have contributed to HBCD levels measured in this study. The distribution of 

HBCD in other plant tissues is also found to 

be stereoisomer specific, with higher levels of 

α-HBCD in stems and leaves and higher levels 

of γ-HBCD in roots (Li et al. 2011).

Food contamination levels seen in this 

study and in our previous study (Schecter 

et al. 2009) differed among samples of the 

same type of food, for example among different samples of turkey sausages. This may 

reflect differences in the source of the foods, 

differences in handling, and dissimilarities in 

ingredients. Although some variation was seen 

between foods, the variation in levels among 

food samples purchased in Dallas may be similar to variation likely to be observed in foods 

from other U.S. locations. The variation might 

be partially due to the different environments 

in which the animals are raised and the differences in feeds used, with cattle being primarily 

raised outdoors and swine and poultry primarily coming from indoor environments. Thus, 

it is fair to say that husbandry practices and 

feed likely influence animal concentrations of 

HBCD, but packaging may also contribute to 

the levels detected.

The data we present here concerning 

HBCD stereoisomers (α-, β-, and γ-HBCD), 

and data on perfluorinated compounds 

(PFCs), PCBs, pesticides, and PBDEs in the 

same pooled food samples collected in 2009 

(Schecter et al. 2009, 2010b), document multiple chemical contamination of U.S. food 

with various POPs and endocrine disrupting 

compounds. Interestingly, concentrations of 

HBCDs are found at higher levels than PCBs, 

pesticides, and PFCs in those pooled foods 

measured, and exceeded PBDEs in samples 

of tilapia, ham, sliced chicken breast, sliced 

turkey, and peanut butter (Schecter et al. 2009, 

2010a). When comparing the total HBCD levels from the 2009 study with total HBCD levels 

calculated in this study, the total HBCD levels 

between the two studies are similar, i.e., consistent with the use of different assay methods. 

It is noteworthy that calculated HBCD intake 

previously reported for Dallas, Texas, foods 

of 15.3 ng/day, 2.1 × 10–7 mg/kg-bw/day for 

a 70-kg individual (Schecter et al. 2009), is 

< 10 mg/kg-bw/day, which is the no observed 

adverse effect level in rats after exposure 

to the HBCD commercial mixture. This 

level is used as the critical effect level by the 

European Union to characterize risk (European 

Commission 2008). To our knowledge, 

human-based benchmarks are not available.

The toxicity of mixtures of both classical 

and emerging POPs reported in this study 

and previously is unknown (Schecter et al. 

2010a, 2010b). An analysis of a larger and 

more representative sampling of Texas and 

other U.S. foods for a wide range of POPs, 

including HBCD stereoisomers, is indicated. 

Exposure and health outcome research should 

continue because of the lack of data regarding 

levels of HBCD in food, human dietary and 

other intake levels, and health effects. Special 

attention to the effects of HBCD stereoisomers 

from food is warranted for the developing 

young because total body burdens in infantile mice are 10- and 2.5-times higher than 

adult levels after exposure to γ-HBCD and 

α-HBCD, respectively (Szabo et al. 2011b). 

These differences would lead to higher concentrations of the HBCD stereoisomers in target 

Table 3. HBCD levels [range (ng/g ww)] in food worldwide.

HBCD levels

Country Food α β γ Total Reference

USA Seafood, meat, peanut butter 0.006–1.307 0.006–0.070 0.012–0.170 0.010–1.366 Present study

Japan Seafood 0.010–18.30 0.010–2.440 0.020–56.60 0.010–77.30 Nakagawa et al. 2010

Norway Seafood 0.005–0.653 0.003–0.063 0.007–0.053 0.015–0.769 Knutsen et al. 2008

United Kingdom Meat, sugars, vegetables, fruits, nuts 0.055–0.290 0.030–0.320 0.019–0.120 0.104–0.730 Driffield et al. 2008

The Netherlands Seafood NA NA NA 0.200–230.0 van Leeuwen and de Boer 2008

Scotland Seafood NA NA NA 0.030–12.10 Fernandes et al. 2008

Romania Meat, dairy, oil NA NA NA 0.040–0.250 Dirtu and Covaci 2010

Sweden Fish, meat, dairy, eggs, fats NA NA NA 0.005–0.630 Törnkvist et al. 2011

NA, not available.
Schecter et al.

1264 volume 120 | number 9 | September 2012 • Environmental Health Perspectives

tissues during critical windows of development. 

Further research on the toxicity of individual 

HBCD stereoisomers and mixtures of these, 

other POPs, and other toxic chemicals detected 

in food, is strongly indicated.

Finding HBCD in food from our previous 

Texas food study and HBCD stereoisomers 

in this study suggests that contamination of 

U.S. food with HBCD is ongoing (Schecter 

et al. 2009). These data also suggest that larger 

and more representative sampling of U.S. food 

should be conducted. Future studies should 

quantify to what extent U.S. foods are contaminated with HBCD stereoisomers, estimate what 

the toxicity of each and all may be, and fully 

elucidate the mechanisms of stereoisomer shifts 

for HBCD in biota and the environment.

References

Abdallah M, Harrad S. 2009. Personal exposure to HBCDs and 

its degradation products via ingestion of indoor dust. 

Environ Int 35:870–876.

Abdallah M, Harrad S. 2011. Tetrabromobisphenol-A, hexabromocyclododecane and its degradation products in UK human 

milk: relationship to external exposure. Environ Int 37:443–448.

Arsenault G, Konstantinov A, Marvin CH, MacInnis G, McAlees A, 

McCrindle R, et al. 2007. Synthesis of the two minor isomers, 

d- and e-1,2,5,6,9,10-hexabromocyclododecane, present 

in commercial hexabromocyclododecane. Chemosphere 

68:887–892.

Ching JA, Binag CA, Alejandro GJD. 2008. Uptake and distribution of some heavy metals in peanut (Arachis hypogaea L.) 

grown in artificially contaminated soils. Phillipine Agric 

Sci 91(2):134–142.

Covaci A, Gerecke AC, Law RJ, Voorspoels S, Kohler M, 

Heeb NV, et al. 2006. Hexabromocyclododecanes (HBCDs) 

in the environment and humans: a review. Environ Sci 

Technol 40:3679–3688.

Dirtu AC, Covaci A. 2010. Estimation of daily intake of organohalogenated contaminants from food consumption and indoor 

dust ingestion in Romania. Environ Sci Technol 44:6297–6304.

Driffield M, Harmer N, Bradley E, Fernandes AR, Rose M, 

Mortimer D, et al. 2008. Determination of brominated flame 

retardants in food by LC-MS/MS: diastereoisomer-specific 

hexabromocyclododecane and tetrabromobisphenol A. 

Food Addit Contam 25:895–903.

European Chemicals Agency. 2010. Candidate List of Substances 

of Very High Concern for Authorisation. Candidate List 

Table. Helsinki. Available: http://echa.europa.eu/web/guest/

candidate-list-table [accessed 9 July 2012].

European Commission. 2008. Risk Assessment: Hexabromocyclododecane. CAS-No. 25637-99-4. EINECS No. 247-148-4. 

Available: http://esis.jrc.ec.europa.eu/doc/risk_assessment/

REPORT/hbcddreport044.pdf [accessed 16 July 2012]. 

Fernandes A, Dicks P, Mortimer D, Gem M, Smith F, Driffield M, 

et al. 2008. Brominated and chlorinated dioxins, PCBs 

and brominated flame retardants in Scottish shellfish: 

methodology, occurrence and human dietary exposure. Mol 

Nutr Food Res 52:238–249.

Germer S, Piersma AH, van der Ven L, Kamyschnikow A, Fery Y, 

Schmitz H-J, et al. 2006. Subacute effects of the brominated flame retardants hexabromocyclododecane and tetrabromobisphenol A on hepatic cytochrome P450 levels in 

rats. Toxicology 218:229–236.

Geyer HJ, Schramm KW, Darnerud PO, Aune M, Feicht EA, 

Fried KW. 2004. Terminal elimination half-lives of the 

brominated flame retardants TBBPA, HBCD, and lower 

brominated PBDEs in humans. Organohalogen Compounds 

66:3867–3872.

Goscinny S, Vandevijvere S, Maleki M, Van Overmeire I, 

Windal I, Hanot V, et al. 2011. Dietary intake of hexabromocyclododecane diastereoisomers (a-, b-, and g-HBCD) in 

the Belgian adult population. Chemosphere. 84:279–288.

Harrad S, Abdallah MA-E, Covaci A. 2009. Causes of variability in 

concentrations and diastereomer patterns of hexabromocyclododecanes in indoor dust. Environ Int 35:573–579.

Haukås M, Hylland K, Berge JA, Nygård T, Mariussen E. 2009. 

Spatial diastereomer patterns of hexabromocyclododecane (HBCD) in a Norwegian fjord. Sci Total Environ 

407:5907–5913.

Hayward SJ, Lei YD, Wania F. 2006. Comparative evaluation of 

three high-performance liquid chromatography–based Kow 

estimation methods for highly hydrophobic organic compounds: polybrominated diphenyl ethers and hexabromocyclododecane. Environ Toxicol Chem 25:2018–2027.

Heeb NV, Graf H, Bernd Schweizer W, Lienemann P. 2010. 

Thermally-induced transformation of hexabromocyclo dodecanes and isobutoxypenta bromocyclododecanes in flameproofed polystyrene materials. Chemosphere 80:701–708.

Heeb NV, Schweizer WB, Kohler M, Gerecke AC. 2005. Structure 

elucidation of hexabromocyclododecanes—a class of 

compounds with a complex stereochemistry. Chemosphere 

61:65–73.

Hinkson NC, Whalen MM. 2009. Hexabromocyclododecane 

decreases the lytic function and ATP levels of human 

natural killer cells. J Appl Toxicol 29:656–661.

Hinkson NC, Whalen MM. 2010. Hexabromocyclododecane 

decreases tumor-cell-binding capacity and cell-surface 

protein expression of human natural killer cells. J Appl 

Toxicol 30:302–309.

Johnson-Restrepo B, Adams DH, Kannan K. 2008. Tetrabromobisphenol A (TBBPA) and hexabromocyclododecanes 

(HBCDs) in tissues of humans, dolphins, and sharks from 

the United States. Chemosphere 70:1935–1944.

Knutsen HK, Kvalem HE, Thomsen C, Frøshaug M, Haugen M, 

Becher G, et al. 2008. Dietary exposure to brominated flame 

retardants correlates with male blood levels in a selected 

group of Norwegians with a wide range of seafood consumption. Mol Nutr Food Res 52:217–227.

Kunisue T, Takayanagi N, Isobe T, Takahashi S, Nakatsu S, 

Tsubota T, et al. 2007. Regional trend and tissue distribution of brominated flame retardants and persistent organochlorines in raccoon dogs (Nyctereutes procyonoides) 

from Japan. Environ Sci Technol 42:685–691.

Law RJ, Herzke D, Harrad S, Morris S, Bersuder P, Allchin CR. 

2008. Levels and trends of HBCD and BDEs in the European 

and Asian environments, with some information for other 

BFRs. Chemosphere 73:223–241.

Li Y, Zhou Q, Wang Y, Xie X. 2011. Fate of tetrabromobisphenol A 

and hexabromocyclododecane brominated flame retardants 

in soil and uptake by plants. Chemosphere. 82:204–209.

Lind Y, Aune M, Atuma S, Becker W, Bjerselius R, Glynn A, 

et al. 2002. Food intake of the brominated flame retardants: 

PBDE’s and HBCD in Sweden. Organohalogen Compounds 

58:181–184.

Morose G. 2006. An Overview of Alternatives to Tetrabromobisphenol A (TBBPA) and Hexabromocyclododecane 

(HBCD). Lowell, MA:Lowell Center for Sustainable 

Production, University of Massachusetts Lowell.

Nakagawa R, Murata S, Ashizuka Y, Shintani Y, Hori T, 

Tsutsumi T. 2010. Hexabromocyclododecane determination 

in seafood samples collected from Japanese coastal areas. 

Chemosphere 81:445–452.

Palace V, Park B, Pleskach K, Gemmill B, Tomy G. 2010. Altered 

thyroxine metabolism in rainbow trout (Oncorhynchus mykiss) 

exposed to hexabromocyclododecane (HBCD). Chemosphere 

80:165–169.

Roosens L, Abdallah MA-E, Harrad S, Neels H, Covaci A. 2009. 

Exposure to hexabromocyclododecanes (HBCDs) via dust 

ingestion, but not diet, correlates with concentrations in 

human serum: preliminary results. Environ Health Perspect 

117:1707–1712.

Roze E, Meijer L, Bakker A, Van Braeckel KNJA, Sauer PJJ, 

Bos AF. 2009. Prenatal exposure to organohalogens, including brominated flame retardants, influences motor, cognitive, and behavioral performance at school age. Environ 

Health Perspect 117:1953–1958.

Ryan JJ, Wainman BC, Schecter A, Moisey J, Kosarac I, Sun WF. 

2006. Trends of the brominated flame retardants, PBDEs and 

HBCD, in human milks from North America. Organohalogen 

Compounds 68:778–781.

Saegusa Y, Fujimoto H, Woo G-H, Inoue K, Takahashi M, 

Mitsumori K, et al. 2009. Developmental toxicity of brominated flame retardants, tetrabromobisphenol A and 

1,2,5,6,9,10-hexabromocyclododecane, in rat offspring after 

maternal exposure from mid-gestation through lactation. 

Reprod Toxicol 28:456–467.

Schecter A, Colacino J, Haffner D, Patel K, Opel M, Päpke O, 

et al. 2010a. Perfluorinated compounds, polychlorinated 

biphenyls, and organochlorine pesticide contamination in 

composite food samples from Dallas, Texas, USA. Environ 

Health Perspect 118:796–802.

Schecter A, Colacino J, Patel K, Kannan K, Yun SH, Haffner D, 

et al. 2010b. Polybrominated diphenyl ether levels in 

foodstuffs collected from three locations from the United 

States. Toxicol Appl Pharmacol 243:217–224.

Schecter A, Haffner D, Colacino J, Patel K, Päpke O, Opel M, 

et al. 2009. Polybrominated diphenyl ethers (PBDEs) and 

hexabromocyclodecane (HBCD) in composite U.S. food 

samples. Environ Health Perspect 118:357–362.

Szabo DT, Diliberto JJ, Hakk H, Huwe JK, Birnbaum LS. 2010. 

Toxicokinetics of the flame retardant hexabromocyclododecane gamma: effect of dose, timing, route, repeated 

exposure, and metabolism. Toxicol Sci 117:282–293.

Szabo DT, Diliberto JJ, Hakk H, Huwe JK, Birnbaum LS. 2011a. 

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

exposure, and metabolism. Toxicol Sci 121:234–244.

Szabo DT, Diliberto JJ, Huwe JK, Birnbaum LS. 2011b. 

Accumulation and distribution of HBCD alpha and gamma 

in developing mice differs from that in adults. Toxicol Sci 

123:256–263.

Thomsen C, Knutsen HK, Liane VH, Frøshaug M, Kvalem HE, et al. 

2008. Consumption of fish from a contaminated lake strongly 

affects the concentrations of polybrominated diphenyl 

ethers and hexabromocyclododecane in serum. Mol Nutr 

Food Res 52:228–237.

Törnkvist A, Glynn A, Aune M, Darnerud PO, Ankarberg EH. 

2011. PCDD/F, PCB, PBDE, HBCD and chlorinated pesticides in a Swedish market basket from 2005—levels and 

dietary intake estimations. Chemosphere 83:193–199.

U.S. EPA (U.S. Environmental Protection Agency). 2010. 

Hexabromocyclododecane (HBCD) Action Plan Summary. 

Available: http://www.epa.gov/oppt/existingchemicals/

pubs/actionplans/RIN2070-AZ10_HBCD%20action%20

plan_Final_2010-08-09.pdf [accessed 9 July 2012].

van Leeuwen SPJ, de Boer J. 2008. Brominated flame retardants 

in fish and shellfish—levels and contribution of fish consumption to dietary exposure of Dutch citizens to HBCD. 

Mol Nutr Food Res 52:194–203.

Zegers BN, Mets A, van Bommel R, Minkenberg C, Hamers T, 

Kamstra JH, et al. 2005. Levels of hexabromocyclododecane in harbor porpoises and common dolphins from 

western European seas, with evidence for stereoisomerspecific biotransformation by cytochrome P450. Environ 

Sci Technol 39:2095–2100.