Ecological State of the Science Report on Decabromodiphenyl Ether 

(decaBDE) 

Bioaccumulation and Transformation 

 

Prepared by: 

Environment Canada 

August 2010 
ii

Summary 

In July 2006, the final decision on the screening assessment of substances – 

polybrominated diphenyl ethers (PBDEs) – was published by the Minister of the 

Environment and the Minister of Health in the Canada Gazette, Part I. It was concluded 

that PBDEs—i.e., tetrabromodiphenyl ether (tetraBDE), pentabromodiphenyl ether 

(pentaBDE), hexabromodiphenyl ether (hexaBDE), heptabromodiphenyl ether (heptaBDE), 

octabromodiphenyl ether (octaBDE), nonabromodiphenyl ether (nonaBDE) and 

decabromodiphenyl ether (decaBDE)—which are found in commercial PentaBDE, 

OctaBDE and DecaBDE technical formulations, are entering the environment in a quantity 

or concentration or under conditions that have or may have an immediate or long-term 

harmful effect on the environment or its biological diversity, and thus meet the criteria set 

out in paragraph 64(a) of the Canadian Environmental Protection Act, 1999 (CEPA 1999). 

In addition, it was concluded that all seven PBDEs homologues met criteria for persistence, 

but only tetra- to hexaBDE met the criteria for bioaccumulation as defined in the

Persistence and Bioaccumulation Regulations. The analysis also noted that the higher 

brominated diphenyl ethers, and decaBDE in particular, could accumulate to some degree 

in biota and debrominate to bioaccumulative and persistent transformation products. 

Since the completion of the Ecological Screening Assessment, a large amount of new 

information has been published regarding the accumulation of decaBDE in biota and the 

potential transformation of decaBDE to persistent and bioaccumulative products. The 

purpose of the present report is to provide an updated analysis of bioaccumulation and 

transformation of decaBDE, by summarizing evidence considered in the original Screening 

Assessment, and then examining the related new science published up to August 25, 2009. 

Overall, available data do not show that the decaBDE itself meets the numeric criteria for 

bioaccumulation as defined in the Persistence and Bioaccumulation Regulations. With 

regard to bioaccumulation, potential factors such as low assimilation efficiency and/or 

metabolic transformation appear to be important determinants of accumulation in 

organisms. Nevertheless, some recent studies have shown concentrations of decaBDE to be 

increasing steadily in some wildlife species and there are a few equivocal reports of 

biomagnification factors (BMFs) exceeding 1. In some cases, such as in the tissues of 

kestrel, sparrowhawk, peregrine falcon, glaucous gull, red fox, shark, harbour porpoise and 

whitebeaked dolphin, measured concentrations of decaBDE in tissues are interpreted as 

high. While trophic magnification or bioaccumulation is a potential explanation for these 

high concentrations, it is also very possible that some biota are exposed to very high 

exposure concentrations of decaBDE by consuming contaminated refuse and/or inhabiting 

decaBDE hotspots close to industrialized areas. 

This report also considers it reasonable to conclude that decaBDE may contribute to the 

formation of lesser-brominated PBDEs and other metabolic products in organisms—

potentially those that are bioaccumulative. Although there is some uncertainty, the 

evaluation found evidence that fish and mammals appear to have some capacity to 

metabolically break down decaBDE. In fish, decaBDE appears to form hepta- to 

nonaBDEs, and potentially penta- and hexaBDEs. In mammals, debromination of decaBDE 

down to heptaBDEs has been observed. In both fish and mammals, formation of lower 

brominated PBDEs appears to be very low and only a fraction (typically on the order of a 

few percent) of the total amount of decaBDE administered to the organism. However, some 

rodent studies have made inferences, based on mass balance evaluations, that rates of 

transformation may be higher, with one study suggesting that approximately 45% of the 
iii

total dose of decaBDE was unaccounted for and may have been metabolized to other 

compounds (such as hydroxylated and hydroxymethoxylated PBDEs) and/or bound as 

inextricable residues. 

The evaluation of transformation in the environment identified numerous laboratory studies 

that provide evidence that decaBDE may break down in the environment, particularly as a 

result of photodegradation and biodegradation. Studies of photodegradation of decaBDE 

sorbed to solids in aqueous and dry systems have demonstrated transformation of decaBDE 

to tri- to nonaBDEs, tri- to octabrominated dibenzofurans (octaBDFs) and unidentified 

products. While relevant to the environment, the actual fraction of decaBDE exposed to 

sunlight adsorbed to atmospheric and aquatic particulates, or solids (anthropogenic or 

natural), would be a small fraction of the total amount of decaBDE in the environment. 

Biodegradation studies have also shown potential breakdown of decaBDE mainly to nona-, 

octa- and heptaBDEs, while transformation to triBDEs has also been shown under 

enhanced laboratory conditions. Overall, biodegradation appears to occur at a much slower 

rate than that of phototransformation, with half-lives in the range of several years to several 

decades. The photolytic half-life of decaBDE adsorbed to house dust and exposed to 

sunlight has been reported to range from approximately 1–2 months (assuming 8 hours of 

sunlight per day). 

Modelling of bioaccumulation factors (BAFs) and BMFs was conducted to estimate 

whether transformation products of decaBDE resulting from processes in organisms and in 

the general environment could be bioaccumulative. The evaluation found that many of the 

identified transformation products could be bioaccumulative (i.e., have BAFs in excess of 

5000) and some could biomagnify in food chains. The analysis also indicated potential 

transformation of decaBDE to products (i.e., tetra- to hexaBDEs) that have been 

established as bioaccumulative based on empirical evidence. 

While laboratory studies on the transformation of decaBDE support a conclusion that 

transformation to lower BDEs and BDFs should be occurring in the environment, the 

phenomenon has not been conclusively shown through monitoring studies to occur in the 

environment. This suggests that the process of environmental transformation may be very 

slow and subtle, and possibly may have relevance to a small fraction of the total decaBDE 

reservoir in the environment. Evidence of transformation may be shielded by existing 

patterns of PBDEs in the environment which are dominated by congeners found in the 

commercial products. The fact that fewer studies have historically measured octa- and 

nonaBDE congeners in environmental samples makes the elucidation of debromination 

patterns of decaBDE in the environment challenging. 

 

While this report has focused on decaBDE, its analyses and conclusions provide useful 

inferences on alternative flame retardants with similar chemical structures and use patterns, 

such as decabromodiphenyl ethane (decaBD ethane). DecaBDE and decaBD ethane have 

only minor structural differences relating to the bond between their aromatic rings and, 

thus, these substances may have similarities in physical and chemical properties, 

persistence, transformation patterns, and accumulation in organisms. Based on the 

similarity in properties between decaBDE and decaBD ethane, the presence of decaBD 

ethane in Canadian wildlife, and the potential for decaBD ethane to be used as a large-scale 

replacement for decaBDE, there is also a need to further understand the potential risks from 

decaBD ethane in the environment and its capacity to accumulate in wildlife and transform 

to potentially bioaccumulative products. Understanding the risk from alternatives generally 

will help to ensure that substitutions of flame retardants are made on an informed basis. 
iv

Table of Contents 

1 Introduction................................................................................................................ 1

1.1 Background.................................................................................................................1 

1.1.1 Composition of Commercial DecaBDE................................................................2 

2 Evidence for Bioaccumulation................................................................................... 8 

2.1 Biota Concentration Data............................................................................................9 

2.1.1 Information Evaluated in the Screening Assessment ........................................9 

2.1.2 New Biota Concentration Data ........................................................................10 

2.1.3 Synthesis of Biota Concentration Data............................................................16 

2.2 Bioconcentration, Bioaccumulation and Biomagnification Data .............................17 

2.2.1 Information Evaluated in the Screening Assessment.......................................17 

2.2.2 New Bioaccumulation Data .............................................................................18 

2.2.3 Model Predictions ............................................................................................39 

2.3 Weight-of-Evidence Analysis...................................................................................43 

2.3.1 Summary of Evidence......................................................................................43 

2.3.2 Conclusion Respecting Bioaccumulation ........................................................46 

3 Evidence of Transformation .................................................................................... 46 

3.1 In Vivo Transformation ............................................................................................47 

3.1.1 Information Evaluated in the Screening Assessment for PBDEs ....................47 

3.1.2 New Evidence of Debromination in Vivo ........................................................50 

3.1.3 Conceptual Models of Metabolism in Vivo .....................................................57 

3.1.4 Implications for Bioaccumulation Assessment................................................58 

3.2 Debromination in the Environment ..........................................................................63 

3.2.1 New Evidence of Debromination in the Environment.....................................66 

3.2.2 Conceptual Model of Transformation in the Environment..............................81 

3.2.3 Implications for Bioaccumulation Assessment................................................93 

4 Summary and Conclusions ...................................................................................... 94 

4.1 Consideration Regarding Related Products ..............................................................96 

5 Key Uncertainties and Data Gaps............................................................................ 97 

6 References................................................................................................................ 99 

Appendix A – Biota Concentration Data 

Appendix B – Bioaccumulation, Biomagnification and Trophic Magnification Data. 

Appendix C – Selection of Log Kow for QSAR Modelling of DecaBDE. 

Appendix D – Model Input Parameters for BAF-QSAR and Terrestrial Biomagnification 

Model Predictions for DecaBDE Transformation Products. 

Appendix E – Summary of Environmental Degradation and Debromination Studies for 

DecaBDE. 

Appendix F – Stuctural Diagrams for DecaBDE and DecaBDethane 
v

List of Tables 

Table 1-1: Concentrations (%, w/w) of PBDEs in selected commercial Octa- and DecaBDE 

products (La Guardia et al. 2006) 3 

Table 2-1: BMFs estimated from the decaBDE rainbow trout feeding study by Stapleton et 

al. (2006; and personal communication from HM Stapleton to Environment Canada, 

January 2008; unreferenced)................................................................................................ 19 

Table 2-2: Summary of observed concentrations of decaBDE in freshwater and marine 

biota from the Lumparn Estuary, Baltic Sea and North Atlantic Ocean (Burreau et al. 2004, 

2006) .................................................................................................................................... 21 

Table 2-3: Estimated BMFs for decaBDE in sampled biota from Lumparn Estuary, Baltic 

Sea and Atlantic Ocean pelagic food webs reported by Burreau et al. (2004, 2006) .......... 22 

Table 2-4: Analytical results from Law et al. (2006) for decaBDE, lipid content, organic 

carbon content and δ15N in water, sediments and biota from Lake Winnipeg .................... 25 

Table 2-5: Summary of accumulation factor calculations by United Kingdom (2007a) using 

the findings of Kierkegaard et al. (2007)............................................................................. 36 

Table 2-6: Estimated BMFs from the Thomas et al. (2005) feeding study with captive 

juvenile grey seals (Halichoerus grypus) ............................................................................ 38 

Table 2-7: BAF and BCF predictions for decaBDE using the Arnot-Gobas kinetic model 

(v1.11).................................................................................................................................. 41 

Table 2-8: Wolf BMF predictions for decaBDE made using the terrestrial biomagnification 

model of Gobas et al. (2003)................................................................................................ 42 

Table 3-1: Predictions of the CHEMCAN Model for decaBDE in the “Ontario Mixed 

Wooded Plain” region.......................................................................................................... 83 

Table 3-2: Summary of decaBDE transformation products observed in laboratory studies 

with some relevance to environmental settings ................................................................... 90 

List of Figures 

Figure 1-1: The relationship between BAF, BCF and BMF for upper and lower trophic 

levels predicted by the BAF–QSAR model for hydrophobic, non-metabolized substances . 7 

Figure 3-1: Combined frequency distributions of predicted BAFs for metabolites of 

decaBDE (predictions for middle trophic level).................................................................. 61

Figure 1-2: Conceptual model of possible environmental transformation pathways for 

decaBDE................................................................................................86 
vi

List of Acronyms 

Acronym Description

B bioaccumulative 

BAF–QSAR 

bioaccumulation factor – quantitative 

structure-activity relationship 

BAF bioaccumulation factor 

BCF bioconcentration factor 

BDE brominated dipheynyl ether 

BDE209 congener form of decaBDE 

BDF brominated dibenzofuran 

BDL below detection limit 

BFR brominated flame retardant 

BMF biomagnification factor 

bw body weight 

bw/d body weight per day 

CEPA 1999 

Canadian Environmental Protection Act, 

1999 

CF confidence factor 

Ci/mol curies/mol 

cm centimetre 

CYP1A cytochrome P4501A 

d day or days 

DE diphenyl ether 

decaBDE decabromodiphenyl ether 

DL detection limit 

dw dry weight 

ECNI electron capture negative ionization 

ED dietary assimilation efficiency 

EI electron ionization 

EVA Experimental Value Adjustment 

FID flame ionization detector 

g grams 

GC gas chromatography 

GC/MS gas chromatography / mass spectrometry 

GC–ECD 

gas chromatography – electron capture 

detection 

GI gastrointestinal 

h hour or hours 

heptaBDE heptabromodiphenyl ether 

hexaBDE hexabromodiphenyl ether 

HIPS high-impact polystyrene 

HPLC high-performance liquid chromatography 

HRGC high-resolution mass spectrometry 

HRMS high-resolution mass spectrometry 

k2 elimination rate constant 

kE fecal egestion rate constant 

kG growth rate constant 
vii

kg/d kilograms per day 

kM metabolic rate constant 

Koa octanol-air partition coefficient 

KOAWIN 

model for prediction of octanol-air 

partition coefficient 

Kow octanol-water partition coefficient 

kT total elimination rate 

L litre 

Lipid lipid weight 

LRMS low-resolution mass spectrometry 

M 

molarity, as the number of moles of a 

solute in 1 litre of solution 

MDL method detection limit 

mg/d milligrams per day 

mg/kg milligrams per kilogram 

MITI 

Ministry of International Trade and 

Industry (Japan) 

mL millilitre 

mmol/kg millimole per kilogram 

mPa millipascal 

MMP Molecular Modelling Pro 

n number of samples 

ng/d nanogram per day 

ng/g nanogram per gram 

nm nanometre 

NMR nuclear magnetic resonance 

nonaBDE nonabromodiphenyl ether 

OC organic carbon 

octaBDE octabromodiphenyl ether 

P persistent 

PBDD polybrominated dibenzo-p-dioxin 

PBDE polybrominated diphenyl ether 

PBDF polybrominated dibenzofuran 

PCB polychlorinated biphenyl 

pentaBDE pentabromodiphenyl ether 

pg/g picogram per gram 

QSAR quantitative structure-activity relationship 

QSPR 

quantitative structure-property 

relationships 

sediment BSAFs 

sediment biota-sediment accumulation 

factors 

SIM selected ion monitoring 

soil BSAFs soil biota-soil accumulation factors 

SRM standard reference material 

TCE trichlorethane 

tetraBDE tetrabromodiphenyl ether 
viii

THF tetrahydrofuran 

TMF trophic magnification factor 

TTR transthyretin 

μg micrograms 

μg/kg micrograms per kilogram 

μg/L micrograms per litre 

μL microlitres 

μm micrometer 

μmol/kg micromole per kilogram 

UK United Kingdom 

US United States 

UV ultraviolet 

v/v volume per volume 

W Watts 

w/w weight per weight 

ww wet weight 

WWTP wastewater treatment plant 
1

1 Introduction 

The purpose of this report is to provide an updated analysis of the bioaccumulation and 

environmental transformation of decabromodiphenyl ether (decaBDE), to be considered 

in the context of the information and analyses already published in the final screening 

assessment on polybrominated diphenyl ethers (PBDEs) (Canada 2006). This evaluation 

is considered a state of the science review. While this report does not critique individual 

studies, it considers the reliability of individual studies when forming a weight of 

evidence for persistence, bioaccumulation or inherent toxicity to non-human biota. This 

report considers materials published up to August 25, 2009. 

1.1 Background 

In July 2006, the ministers of Environment and Health published their final screening 

assessment on PBDEs (Canada 2006). The environmental screening assessment 

examined various supporting information and developed conclusions based on a 

weight-of-evidence approach as required under subsection 76.1 of the Canadian 

Environmental Protection Act, 1999 (CEPA 1999) (Canada 1999). As the term 

“screening assessment” implies, this was not an exhaustive review of all available data, 

but rather, it presented the most critical studies and lines of evidence supporting the 

conclusions (e.g., relating to persistence, bioaccumulation, inherent toxicity, risk 

quotients and long-range transport to remote regions like the Arctic). The PBDE 

screening assessment concluded that PBDEs—tetrabromodiphenyl ether (tetraBDE), 

pentabromodiphenyl ether (pentaBDE), hexabromodiphenyl ether (hexaBDE), 

heptabromodiphenyl ether (heptaBDE), octabromodiphenyl ether (octaBDE), 

nonabromodiphenyl ether (nonaBDE) and decaBDE—which are found in commercial 

PentaBDE, OctaBDE and DecaBDE technical formulations, are entering the environment 

in a quantity or concentration or under conditions that have or may have an immediate or 

long-term harmful effect on the environment or its biological diversity and thus meet the 

criterion under paragraph 64(a) of CEPA 1999. In addition, it was concluded that all of 

the PBDEs assessed met criteria for persistence, but only tetra- to hexaBDEs met the 

criterion for bioaccumulation as identified in the Persistence and Bioaccumulation 

Regulations (Canada 2000) under CEPA 1999. It also noted that higher brominated 

PBDEs, and decaBDE in particular, could accumulate to some degree in biota and 

debrominate to bioaccumulative and persistent transformation products. 

Information obtained as of October 2004 was considered for inclusion into the 

environmental screening assessment (Environment Canada 2006a; Environment Canada 

2006b), while information received between November 2004 and October 2005 was 

reviewed but not generally added to the assessment, as these studies supported the 

conclusions of the draft assessment published for public comment in 2004. Since 2004, a 

large amount of information has been and continues to be published on PBDEs and on 

the issues of decaBDE bioaccumulation and transformation. This report examines the 

new data concerning decaBDE and examines the development of lines of evidence 

respecting this substance’s bioaccumulation and transformation. In addition, a Notice of 

Objection, dated February 14, 2007, was submitted to the Minister of the Environment by 
2

the Sierra Legal Defence Fund (now known as Ecojustice Canada) on behalf of the David 

Suzuki Foundation, Environmental Defence, and the Canadian Environmental Law 

Association, requesting the establishment of a Board of Review to clarify the Canadian 

regulatory approach on PBDEs and to recommend changes to the proposed 

Polybrominated Diphenyl Ethers Regulations. In order to consider information and issues 

raised in the Notice of Objection and to inform a decision respecting the formation of a 

Board of Review, this detailed evaluation was prepared to consider the current science 

respecting the bioaccumulation and transformation of decaBDE to bioaccumulative 

products. This report considers published information up to August 25, 2009, identified 

in original literature and review documents. In addition to retrieving the references from a 

literature database search, direct contacts were made with researchers, academics, 

industry and other government agencies to obtain relevant information on decaBDE. 

The analysis presented herein summarizes relevant information on bioaccumulation and 

transformation examined in the PBDE screening assessment (Canada 2006; Environment 

Canada 2006b), and examines how the new science builds upon existing lines of evidence 

respecting bioaccumulation and transformation. Ecological considerations rather than 

those relating to human health are presented. While the screening assessment of PBDEs 

(Health Canada 2004) concluded that worst-case estimates of Canadians’ exposure to 

PBDEs were much lower than the levels of exposure which caused health effects in 

laboratory animals, in light of uncertainties regarding the available database, Health 

Canada supports Environment Canada’s actions to limit the use of PBDEs so that levels 

do not reach a point where they could potentially harm the health of Canadians. The 

evaluation does not consider decaBDE accumulation in humans; however, consideration 

is given to laboratory rodent studies that are used in human health risk assessment since 

such studies also provide insight into impacts relating to mammalian wildlife. 

This state of science report has undergone substantial internal and external written peer 

review/consultation. While external comments were taken into consideration, the final 

contents remain the responsibility of Environment Canada. Additionally, the draft of this 

evaluation was subject to a 60-day public comment period. The critical information and 

considerations upon which the evaluation is based are summarized below. 

1.1.1 Composition of Commercial DecaBDE 

In this evaluation, information on both the congener form of decaBDE (BDE209) and the 

commercial product also known as Decabromodiphenyl Ether (DecaBDE) are 

considered. Since nona- and octaBDEs are also found in DecaBDE formulations, these 

congeners are considered if appropriate. 

According to the World Health Organization (WHO 1994), current manufactured 

formulations of DecaBDE typically contain 

• decaBDE, 97–98% w/w, and 

• other PBDEs (mainly nonaBDE), 0.3–3.0 w/w. 

Older commercial DecaBDE products contained a higher proportion of lower brominated 

PBDEs (mainly nonaBDE and octaBDE isomers) than more recent formulations. For 
3

instance, FR-300-BA, produced in the 1970s and no longer commercially available, 

contained 77.4% decaBDE, 21.8% nonaBDE and 0.8% octaBDE (Norris et al. 1975). 

La Guardia et al. (2006) analyzed compositions of the currently manufactured DecaBDE 

product, Saytex 102E, and compared it with that of Bromkal 82-ODE, which has not 

been manufactured for more than a decade (Table 1-1). They found that Saytex 102E and 

Bromkal 82-ODE contained 96.8% and 91.6% BDE209, respectively. Both formulations 

contained nonaBDEs, with BDE206 in the highest quantity, followed by BDE207 then 

-208. In addition, Bromkal 82-ODE contained 0.56% octaBDEs, with BDE196, -203 and 

-197 identified. OctaBDEs were not identified in the Saytex 102E product. 

Commercial OctaBDE products (no longer in production) also contained decaBDE (La 

Guardia et al. 2006). For instance, Bromkal 79-8DE contained 49.6% decaBDE as well 

as significant nonaBDE (Table 1-1). In comparison, DE-79 contained lesser amounts of 

deca- and nonaBDE (i.e., 1.31 and 13.07%, respectively). 

Table 1-1: Concentrations (%, w/w) of PBDEs in selected commercial Octa- and DecaBDE 

products (La Guardia et al. 2006) 

PBDE Congener Commercial OctaBDE Products Commercial DecaBDE Products 

DE-79 Bromkal 79-8DE Saytex 102E Bromkal 82-ODE 

BDE154 1.07 0.04 nd2 nd 

BDE144 0.1 0.12 nd nd 

hexaBDE1 < 0.02 nd nd nd 

BDE153 8.66 0.15 nd nd 

BDE139 nd nd nd nd 

BDE140 < 0.02 nd nd nd 

BDE138 0.62 nd nd nd 

BDE184 < 0.02 < 0.02 nd nd 

heptaBDE1 < 0.02 nd nd nd 

BDE175/183 42 12.6 nd nd 

BDE191 < 0.02 nd nd nd 

BDE180 1.7 nd nd nd 

BDE171 1.81 0.17 nd nd 

BDE201 0.78 < 0.02 nd nd 

BDE197 22.2 10.5 nd 0.03 

BDE203 4.4 8.14 nd 0.07 

BDE196 10.5 3.12 nd 0.46 

BDE194 < 0.02 nd nd nd 

octaBDE1 < 0.02 nd nd nd 

BDE208 0.19 < 0.02 nd 0.07 

BDE207 11.5 11.2 0.24 4.1 

BDE206 1.38 7.66 2.19 5.13 

BDE 209 1.31 49.6 96.8 91.6 

1

 Categorized only based on degree of bromination 

2

 nd = not detected 
4

1.2 Definitions and Rationale 

For the purposes of this document, it is important to define and distinguish between 

bioaccumulation, bioconcentration and biomagnification. The term “bioaccumulation” 

has been recently summarized by Arnot and Gobas (2006): 

Bioaccumulation is a process in which a chemical substance is absorbed in an organism via all 

routes of exposure as typically occurs in the natural environment, i.e., dietary and ambient 

environment sources. Bioaccumulation is the net result of competing processes of chemical uptake 

and elimination including the diet, respiratory exchange, fecal egestion, metabolic transformation 

of the “parent” compound and growth dilution... Growth dilution is considered a “pseudoelimination” route since the chemical is not actually eliminated by the organism but the 

concentration can be diluted by an increase in the volume of tissue. 

Thus, it is reasonable to expect that if a chemical is detectable in organism tissues, it is 

there as the result of the process of bioaccumulation. Bioaccumulation can be of concern 

because when organisms take up chemicals, the likelihood that those organisms will be 

harmed generally increases with the amount accumulated. In addition, as 

bioaccumulation increases, there is increased likelihood that contaminated organisms will 

cause indirect harm to predators that consume them. Chemicals that are highly 

bioaccumulative are of particular concern because they tend to biomagnify, i.e., increase 

in concentration from prey to predator across several trophic levels. Thus, relatively low 

ambient concentrations of highly bioaccumulative substances have the potential to cause 

both direct toxicity and indirect toxicity due to biomagnification. 

It is important to distinguish between the process of “bioaccumulation” and the term 

“bioaccumulative” as used in the Persistence and Bioaccumulation Regulations

(Canada 2000), a regulation made under CEPA 1999 and as described in Canada’s Toxic 

Substances Management Policy (Canada 1995a, 1995b). According to the Persistence 

and Bioaccumulation Regulations: 

A substance is bioaccumulative 

(a) when its bioaccumulation factor is equal to or greater than 5000; 

(b) if its bioaccumulation factor cannot be determined in accordance with a method referred to in 

section 5, when its bioconcentration factor is equal to or greater than 5000; and 

(c) if neither its bioaccumulation factor nor its bioconcentration factor can be determined in 

accordance with a method referred to in section 5, when the logarithm of its octanol-water 

partition coefficient is equal to or greater than 5.

The Regulations go on to indicate that “[t]he determination of persistence and 

bioaccumulation...must be made...taking into account the intrinsic properties of the 

substance, the ecosystem under consideration and the conditions in the environment.” 

Thus, determining whether these criteria are met involves professional judgement which 

considers the intrinsic properties of the substance and ecosystem under consideration. 

These criteria were first proposed as part of the federal Toxic Substances Management 

Policy (Canada 1995a, 1995b) and were intended to address lipophilic substances with 

the potential to bioaccumulate and biomagnify in aquatic organisms to levels causing 

effects at the top of the food web. Application of a criterion of 5000 for bioaccumulation 
5

factors (BAFs) or bioconcentration factors (BCFs), or 5 for log Kow (octanol-water 

partition coefficient), are recommended in the Policy. Substances that undergo significant 

bioaccumulation and/or biomagnification are considered “bioaccumulative.” As noted 

above, professional judgement is required in the application of these criteria. Notably, the 

Policy indicates that log Kow should be used with caution in predicting the 

bioconcentration and bioaccumulation potential of organic substances (Canada 1995b). 

For instance, many substances can be metabolized and this can function to decrease 

bioaccumulation potential of the parent compound, and thus many substances with a very 

high log Kow (e.g., >5) can have very low bioaccumulation/bioconcentration potential. 

Based on the definition in CEPA 1999 and its regulations which have their basis in the 

Toxic Substances Management Policy, the term “bioaccumulative” is not synonymous 

with the term “bioaccumulation.” Rather, “bioaccumulation” is the process which may 

lead to a substance being considered “bioaccumulative.” Evidence that a substance is 

both persistent and bioaccumulative when combined with evidence of toxicity and release 

or expected release into the environment provides a significant indication of its potential 

to cause ecological harm (Environment Canada 2007). 

Gobas and Morrison (2000) provide a concise definition of the process of 

“biomagnification” as it is currently understood: 

This is the process in which the chemical concentration in an organism achieves a level that 

exceeds that in the organism’s diet, due to dietary absorption. The extent of chemical 

biomagnification in an organism is best determined under laboratory conditions, where organisms 

are administered diets containing a known concentration of chemical, and there is no change in 

chemical uptake through other exposure routes (e.g., respiratory surface, dermis). 

Biomagnification also can be determined under field conditions, based on chemical 

concentrations in the organism and its diet. Biomagnification factors derived under controlled 

laboratory conditions, which exclude uptake through routes other than the diet, are different from 

those determined under field conditions, because field-based biomagnification factors are 

inevitably the result of chemical uptake by all routes of chemical uptake, rather than dietary 

absorption alone. 

The terms “trophic magnification,” “food web magnification” and “food web 

biomagnification” are often used interchangeably to refer to the same phenomenon. This 

is the phenomenon whereby chemical concentrations in organisms increase with trophic 

level, resulting in higher concentrations in predators than in prey (i.e., biomagnification 

through successive trophic levels). Gobas and Morrison (2000) describe “food-chain” 

bioaccumulation of neutral organic substances as 

...the process in which chemical concentrations in organisms increase with each step in the foodchain, resulting in chemical concentrations in predators that are greater than those in their prey. 

Because concentrations of many hydrophobic chemicals in organisms increase as the lipid content 

of the organism increases, the occurrence of food-chain bioaccumulation is detected best by 

comparing chemical concentrations in predators and prey on a lipid weight basis. An increase in 

lipid-based concentrations in organisms with increasing trophic level indicates food-chain 

bioaccumulation. 
6

Potential indicators of biomagnification or trophic magnification include field-measured 

biomagnification factors (BMFs—the ratio in concentration between predator and prey) 

for known predator-prey relationships, or trophic magnification factors (TMFs—

sometimes called food web magnification factors) which examine the incremental 

magnification with each trophic level for the food web as a whole. TMFs also represent 

the average ratio of concentrations between predator and prey through an entire food 

web. A BMF or TMF exceeding 1 indicates that biomagnification or trophic 

magnification is occurring. The use of BMFs and TMFs are generally thought to be more 

appropriate for terrestrial and marine mammals which breathe air rather than water. Note 

that in laboratory feeding studies, dietary BAFs are often estimated by representing the 

relationship between the test organisms and their diet (i.e., synonymous with the 

definition of a BMF). For the purposes of this report, dietary BAFs are referred to as 

BMFs. 

The BAF criterion of 5000 as identified in the Persistence and Bioaccumulation 

Regulations denotes chemicals which are highly bioaccumulative at the organism level 

and may also suggest the potential for biomagnification or trophic magnification. It 

should be noted that BAFs measured or estimated from total water concentrations are 

dependent on the quantity and quality of organic carbon and particulate matter in the 

water column. The available fraction of a chemical in water determines the uptake 

potential of a substance and subsequently the bioaccumulation potential. The BAF–

QSAR model of Arnot and Gobas (2003) provides bioaccumulation predictions for a 

generic aquatic food web (with lower, middle and upper trophic levels) based on the 

current understanding of bioaccumulation processes in aquatic organisms, accounting for 

the available fraction in water. Figure 1-1 illustrates the model-predicted relationship for 

hydrophobic, non-metabolized substances between log Kow, BAF, BCF and BMF for 

aquatic lower and upper trophic levels in the model. The model predictions suggest that 

as log Kow increases toward and above 4–5, the BAF increases toward and above 5000. 

Also, in this range of log Kow, the BMF begins to significantly exceed 1, indicating that 

biomagnification may occur. 

The underlying mechanism for this phenomenon is described by Kelly et al. (2004). For 

non-metabolized chemicals in aquatic organisms, the BMF results from competing 

processes of chemical uptake (respiratory and dietary) and chemical elimination 

(respiratory and gastrointestinal), which depend on organism physiology and chemical 

hydrophobicity. For aquatic organisms, respiratory elimination is significant for less 

hydrophobic chemicals, and elimination processes override gastrointestinal absorption 

processes. However, as hydrophobicity increases (i.e., log Kow > 5), respiratory 

elimination slows and gastrointestinal absorption becomes significant, resulting in 

biomagnification (i.e., a BMF exceeding 1). At very high log Kow, dietary chemical 

absorption efficiency tends to decline because chemicals are too strongly sorbed to the 

food matrix to be absorbed into the organism, resulting in a decline in both the BAF and 

the BMF (Arnot and Gobas 2003, 2004). Regardless of log Kow, when sufficient, 

chemical metabolism often sustains the total elimination rate at a high enough level to 

prevent chemical biomagnification, even for higher log Kow substances (Arnot and Gobas 

2003; Kelly et al. 2004). 
7

This relationship between log Kow, BAF and BMF for non-metabolized substances, 

which is based on current knowledge of bioaccumulation and biomagnification in aquatic 

organisms, supports the use of a BAF criterion of 5000 to indicate when bioaccumulation 

may occur. Note that the BCF is a less conservative endpoint since it does not increase as 

quickly with log Kow as the BAF. However, a lower trophic level BCF of 5000 generally 

corresponds to a log Kow of 5, which is consistent with the log Kow criterion of the 

Persistence and Bioaccumulation Regulations. 

Figure 1-1: The relationship between BAF, BCF and BMF for upper and lower trophic levels 

predicted by the BAF–QSAR model for hydrophobic, non-metabolized substances 

0

1

2

3

4

5

6

7

8

0 2 4 6 8 10 12 14

log Kow

log BAF or BCF

0

1

2

3

4

5

6

7

8

BMF

Upper Trophic Level BAF

Lower Trophic Level BCF

BCF or BAF = 5000

Upper Trophic Level BMF

BMF=1

Two additional measures which can also provide evidence of biomagnification include 

biota-sediment accumulation factors (sediment BSAFs) and biota-soil accumulation 

factors (soil BSAFs). The BSAF represents the steady-state concentration ratio between 

an organism and soil/sediment, on a lipid-normalized and organic-carbon-normalized 

basis. Assuming equal chemical sorptive capacities between lipid and organic carbon, 

sediment BSAFs of 1 are expected when organisms are in equilibrium with sediments 

(i.e., in the absence of food web biomagnification or trophic dilution). However, given 

the observed difference in sorptive capacity between lipid and organic carbon (see Seth et 

al. 1999), BSAFs up to approximately 3 are still consistent with equilibrium conditions. 

Alternatively, the American Society for Testing Materials (ASTM 1997) recommends a 

“cut-off” value of 1.7 to represent equilibrium conditions. Thus, sediment BSAFs that 

exceed approximately 1.7 to 3 suggest that food web biomagnification or an alternative 

magnification process is increasing organism chemical concentrations above equilibrium. 

Although there are currently no guidelines for evaluating soil BSAFs, a similar range 

would likely be appropriate to assess whether a soil BSAF value provides evidence that 

food web biomagnification or an alternative magnification process is increasing organism 

chemical concentrations above equilibrium. 
8

The evaluation of bioaccumulation data in this report mainly relies on a combination of 

measured and model-predicted BAFs, BCFs, BMFs and TMFs, and in a few cases, 

measured sediment and soil BSAFs. Concentrations of decaBDE in organisms are also 

considered subjectively; however, without quantification of exposure concentrations, 

definite conclusions respecting the capacity for decaBDE to bioaccumulate cannot be 

made from biota concentrations alone. Consideration is also given to chemical uptake and 

elimination and metabolism. An organism’s metabolic capacity can reduce the 

bioaccumulation potential of a substance; however, this metabolic capacity can also result 

in the formation of bioaccumulative transformation products. Full evaluation of BAFs, 

BCFs, BMFs and TMFs would consider the bioaccumulation potential of both the parent 

substance and its metabolic product(s). Further discussion on metabolic transformation is 

found later in this report. 

2 Evidence for Bioaccumulation 

The evaluation of new lines of evidence conducted in this section relies on the definitions 

of bioaccumulative and biomagnification summarized in the preceding section. Most 

importantly, the evaluation of whether these new lines of evidence indicate that decaBDE 

is bioaccumulative relies primarily on ratio-based methods for bioaccumulation 

assessment. From the definitions in Section 1.2, it can be argued that the degree of 

chemical accumulation in an organism is best characterized by a ratio comparing the 

concentration of the subject substance in an organism to the appropriate exposure 

concentration in the organism’s environment (i.e., to enable quantification of BCFs, 

BAFs, BMFs, TMFs or BSAFs). 

Other evidence, such as the presence of high concentrations in top predators, is 

sometimes useful as a supporting line of evidence for ratio-based measures of 

bioaccumulation. However, it is recognized that the characterization of biota chemical 

concentrations as “high” is subjective and is based on professional judgement, and thus 

prone to alternative interpretation. 

The evidence provided by studies quantifying chemical residues in tissues is often 

confounded by various factors. For instance, in many monitoring and field studies, 

particularly those showing unusually high concentrations in predator organisms, there is a 

lack of knowledge regarding levels of exposure. In such instances, quantification of the 

BAF or BMF is impossible. Although a high concentration of a chemical in a top 

predator could be due to trophic magnification, it could also be due to high 

geographically localized exposure to a particular chemical, as would be the case if the 

organism was inhabiting or scavenging from waste dumps. It is also possible that uptake 

routes other than the food chain can be occurring, such as inhalation, or direct ingestion 

of plastic. Further, the analysis of decaBDE can be challenging, affected by a number of 

critical factors that may contribute to the overall accuracy of reported levels in the 

environmental samples. For instance, de Boer and Wells (2006) note that particular 

attention is required during sampling and analysis to avoid issues relevant to 

− degradation due to sunlight, or UV light; 
9

− poor solubility; 

− high background concentrations/potential for contamination and frequent 

low concentrations in biota; and 

− thermal degradation. 

In addition, Covaci et al. (2007) describe analytical issues relevant to gas 

chromatography (GC) set-up affecting the precision of congener determination. They 

discuss a number of chromatographic interferences that hamper good quality data, 

particularly in relation to di-, tetra- and hexaBDEs. Apart from using adequate standards, 

Björkland et al. (2003) also highlighted the need to optimize several other parameters 

relevant to GC / mass spectrometry (MS), such as ionization energy, moderating gas 

pressure, ion source temperature and analyzer temperature. 

The analysis of decaBDE in environmental samples has apparently improved over the 

past decade, as demonstrated by a recent inter-laboratory study undertaken to validate 

and harmonize the analytical methodology (Leonards and Duffek 2008). In total, ten 

routine laboratories participated in this evaluation of decaBDE in dust and sediment, 

including nine laboratories from Europe and one from Canada. Overall, the evaluation 

found acceptable accuracy in results from all laboratories when special attention was 

focused on quality assurance / quality control (QA/QC) procedures. 

The next section summarizes the data respecting decaBDE biota concentrations 

considered in the PBDE screening assessment report (Canada 2006; Environment Canada 

2006b). This summary is followed by an analysis of new data available after 2004 and an 

interpretation of their significance in relation to bioaccumulation and biomagnification of 

decaBDE. 

2.1 Biota Concentration Data 

2.1.1 Information Evaluated in the Screening Assessment 

• Norstrom et al. (2002) did not detect decaBDE in herring gull (Larus argentatus) eggs 

from the Great Lakes. 

• Lichota et al. (2004) determined a total PBDE concentration (67% of which was 

decaBDE) of 0.777 μg/kg lipid measured in fat from Vancouver Island marmot 

(Marmota vancouverensis). 

• European Communities (2002) and Law et al. (2003) summarized the results of many 

analyses for decaBDE in fish and marine mammals from Europe and elsewhere. It was 

noted that decaBDE was detected only occasionally, at a concentration close to the 

detection limit of the method used. 

• DecaBDE (i.e., BDE209) was detected once at a concentration of 1.4 μg/kg wet 

weight (ww) in mussels from Japan (sampling conducted from 1981 to 1985) 

(Watanabe et al. 1987). 

• Dodder et al. (2002) analyzed decaBDE concentrations in freshwater fish from the 

northeastern United States and found that all concentrations of decaBDE were below 

the analytical detection limit (ranged from 1.4 to 1.6 μg/kg ww). 
10

• Ikonomou et al. (2002a) conducted sampling from ringed seal (Phoca hispida) blubber 

between 1981 and 2000, but none of the samples contained decaBDE concentrations in 

excess of the procedural blanks. 

• Ikonomou et al. (2000, 2002a, 2002b) analyzed decaBDE in marine biota collected 

from the British Columbia coast and Holman Island, Northwest Territories, between 

1981 and 2000. The levels of decaBDE were equivalent to those of the procedural 

blanks. 

• DecaBDE (i.e., BDE209) was detected in 18 of 21 analyzed eggs of peregrine falcons 

(Falco peregrinus) from Sweden, at concentrations from 28 to 430 μg/kg lipid 

(Lindberg et al. 2004). 

• A study by de Boer et al. (2001) reported on sampling conducted in 1999 in the North 

Sea and the Tees estuary, UK, of marine mammals—harbour seal (Phoca vitulina), 

harbour porpoise (Phocoena phocoena), white-beaked dolphin (Lagenorhynchus 

albirostris) and bottlenose dolphin (Tursiops truncatus)—for levels of BDE209 in 

liver, muscle and blubber tissues. Most samples showed concentrations of BDE209 

below detection limits (ranged from approximately 0.8 to 9.0 μg/kg lipid); however, a 

few samples showed elevated levels in specific tissues (e.g., up to 26 μg/kg lipid in the 

blubber of harbour seal, up to 160 μg/kg lipid in the liver of harbour porpoise, and up 

to 318 μg/kg lipid in white-beaked dolphin). 

• Sampling was conducted (de Boer et al. 2004) to determine the occurrence of 

decaBDE in liver, muscle tissue and eggs of high trophic level bird species from the 

United Kingdom and the Netherlands. Levels of decaBDE were detected in 10 of 28 

liver samples (range < 1.5 to 181 μg/kg lipid), 14 of 28 muscle samples (range < 4.2 to 

563 μg/kg lipid) and 25 of 68 eggs (range < 1.8 to 412 μg/kg lipid). Concentrations in 

Swedish peregrine falcon (Falco peregrinus) eggs, which were re-analyzed in the 

study, were all within 30% of those originally determined by Lindberg et al. (2004). 

Highest concentrations of decaBDE were measured in muscle tissue samples collected 

from United Kingdom heron (Ardea cinerea) and peregrine falcon (Falco peregrinus), 

and in eggs from Swedish peregrine falcon. 

Further details on these studies are provided in Appendix A and in the Supporting 

Working Document for the Ecological Screening Assessment of Polybrominated Diphenyl 

Ethers (Environment Canada 2006b). 

2.1.2 New Biota Concentration Data 

Recently, a large number of studies showing tissue concentrations of decaBDE in various 

global biota have been published. These studies are discussed below and are interpreted 

with reference to bioaccumulation and biomagnification as defined in earlier sections of 

this report. 

Voorspoels et al. (2006a) reported decaBDE concentrations in red fox (Vulpes vulpes) 

from Belgium. Samples of liver (n=30), muscle (n=33) and adipose tissue (n=27) were 

collected from foxes which were being processed for diagnostic screening of rabies 

infection. Concentrations of decaBDE, detected in less than half of the samples of each 

tissue type, ranged from < 9.1 to 760 ng/g lipid for liver, from < 3.9 to 290 ng/g lipid in 

muscle and from < 3.7 to 200 ng/g lipid for adipose tissue. The authors concluded that 
11

these data confirmed “unambiguously” that decaBDE does “bioaccumulate” in red fox. 

However, in a follow-up paper by Voorspoels et al. (2007), decaBDE was not detected in 

the prey species of fox—wood mice (Apodemus sylvaticus) and bank voles 

(Clethrionomys glareolus). Thus, for this food chain, biomagnification was not identified. 

However, the authors noted that the amount of tissue that was available for analysis was 

low due to the small size of the rodents and this caused some elevated limits of 

quantification, ranging from 7.3 to 17 ng/g lipid for BDE209. 

Voorspoels et al. (2006b) reported decaBDE concentrations in birds of prey from 

Belgium which were found dead or severely injured and had to be euthanized. Bird 

species included the common buzzard (Buteo buteo), sparrowhawk (Accipiter nisus), 

long-eared owl (Asio otus), barn owl (Tyto alba) and tawny owl (Stryx aluco). Tissues 

analyzed from each species included brain, fat, liver, muscle and adipose tissue. Levels of 

decaBDE were detected in nearly all serum samples (2–58 ng/g lipid for all species 

combined) and in some liver samples (below detection limit (BDL)—190 ng/g lipid for 

all species combined), but not in any other tissues. Based on these findings, the authors 

concluded that either the exposure to decaBDE was low, or that this congener was poorly 

accumulated. In the follow-up paper by Voorspoels et al. (2007), decaBDE was not 

detected in prey species of buzzard (wood mice, Apodemus sylvaticus; and bank voles, 

Clethrionomys glareolus) or sparrowhawk (great tit, Parus major), meaning that 

biomagnification could not be assessed. 

Jaspers et al. (2006) analyzed liver and muscle samples from seven species of aquatic 

birds and terrestrial predatory birds from Belgium—grey heron (Ardea cinerea) and great 

crested grebe (Podiceps cristatus), both aquatic, and common buzzard (Buteo buteo), 

kestrel (Falco tinnunculus), sparrowhawk (Accipiter nisus), long-eared owl (Asio otus) 

and barn howl (Tyto alba), all terrestrial. DecaBDE (i.e., BDE209) was detected in one of 

two muscle samples from barn owl at 68 ng/g lipid and not detected in muscle tissue of 

any other species (detection limit (DL) = 1–2.25 ng/g lipid). DecaBDE (BDE209) was 

not detected in liver of heron, grebe or buzzard. The range of decaBDE concentrations in 

liver of the remaining species was 52–85 ng/g lipid. However, decaBDE was only 

detected in a few samples from the terrestrial birds, but not in the aquatic birds. The 

authors suggested that terrestrial bird species might be more exposed to decaBDE than 

aquatic species. 

Christensen et al. (2005) reported decaBDE residues in muscle and fat from coastal (n=6) 

and interior (n=6) grizzly bears (Ursus arctos horribilis) in British Columbia, Canada. 

DecaBDE concentrations (i.e., of BDE209) ranged from approximately 1 to 2.77 μg/kg 

lipid in coastal bears and from approximately 0.5 to 41.71 μg/kg lipid in interior bears 

(concentrations read from graph), suggesting a higher exposure to decaBDE for bears 

feeding in interior environments where diets are predominantly terrestrial in nature. The 

authors also used a “bioaccumulation slope” method to characterize the observed 

bioaccumulation of decaBDE in coastal and interior bears. This analysis is evaluated 

further in Section 2.2. 
12

Verreault et al. (2004, 2005) analyzed decaBDE in two Norwegian arctic top predators, 

glaucous gull (Larus hyperboreus) and polar bear (Ursus maritimus), inhabiting Svalbard 

and Bjoroya, Norway. In gull plasma from Svalbard (n=27), decaBDE ranged from 

< 0.05 to 0.33 ng/g ww while egg concentrations (n=20) ranged from below detection to 

170 ng/g lipid. In Bjoroya, gull plasma (n=89) and egg (n=4) concentrations of 410 ng/g 

lipid and 23–53 ng/g lipid, respectively, were determined. In polar bear plasma from 

Svalbard (n=15), decaBDE was detected in only one sample, at a concentration of 

0.1 ng/g ww. The authors claim that decaBDE was “bioaccumulative” to a limited 

degree. However, this claim was based solely on the occurrence of decaBDE in tissues 

without an evaluation of whether the concentrations are high relative to the environment 

or prey. 

Vorkamp et al. (2005) reported concentrations of decaBDE in eggs of peregrine falcon 

(Falco peregrinus) from southwestern Greenland collected from 1986 to 2003. Although 

decaBDE was detected in all samples (n=37) at concentrations ranging from 3.8 to 

250 ng/g lipid, high concentrations were only found in two eggs from 1995 and 2002 and 

the median concentration was 11 ng/g lipid. The authors noted a significantly increasing 

temporal trend in decaBDE concentration in the falcon tissues. They also concluded that 

their results provided evidence of some uptake and bioaccumulation of decaBDE (ruling 

out the idea of a size threshold for membrane permeation of decaBDE). The authors 

characterized decaBDE concentrations as low in falcon eggs. 

SØrmo et al. (2006) reported analyses of decaBDE in biota from Svalbard, Norway. 

DecaBDE (BDE209) was not detected in two calanoid copepod species (method 

detection limit (MDL) = 0.012–1.3 ng/g lipid) but was detected at a concentration of 

7.22 ng/g lipid in ice amphipod (Gammarus wilkitzkii). The decaBDE concentration in 

polar cod (Boreogadus saida; n=7) ranged from 0.05 to 0.42 ng/g lipid with a mean of 

0.2 ng/g lipid. DecaBDE (i.e., BDE209) was detected in only 1 of 6 samples of ringed 

seal (Pusa hispida) blubber, at a concentration of 0.02 ng/g lipid (MDL = 0.014–

0.75 ng/g lipid). Polar bear (Ursus maritimus) adipose tissue (n=4) contained 0.03–

0.16 ng/g lipid of decaBDE with an average concentration of 0.09 ng/g lipid. DeWit et al. 

(2006) summarized additional data from Muir et al. (2006) and Skaare (2004) for polar 

bears in Svalbard. The BDE209 congener (decaBDE) was not detected in the Muir et al. 

(2006) study (MDL = 1 ng/g lipid) and was 1 ng/g lipid in the Skaare (2004) study. 

Chen et al. (2007) reported measurements of decaBDE concentrations in birds of prey 

from northern China including samples of muscle, liver and kidney from common kestrel 

(Falco tinnunculus; n=6), sparrowhawk (Accipiter nisus; n=11), Japanese sparrowhawk 

(Accipiter gularis; n=6), little owl (Athene noctua; n=6), scops owl (Otus sunia; n=6), 

long-eared owl (Asio otus; n=6), upland buzzard (Buteo hemilasius; n=3) and common 

buzzard (Buteo buteo; n=3). The common kestrel had the highest mean concentrations of 

decaBDE in muscle (2150 ng/g lipid), liver (2870 ng/g lipid) and kidney (483 ng/g lipid), 

followed by scops owl (mean range of 59–537 ng/g lipid in muscle, liver and kidney), 

and sparrowhawk (mean range of 83–249 ng/g lipid in muscle, liver and kidney). The 

concentration of decaBDE in tissues from the other species did not exceed 150 ng/g lipid. 

The authors concluded that decaBDE was especially elevated compared to other 

published reports and that this may be related to significant production, usage or disposal 
13

of decaBDE-containing products in China. They also suggested that exposure to 

decaBDE could be higher in terrestrial food chains than in aquatic food chains. 

Lam et al. (2007) assessed PBDEs levels in eggs of birds of southern China. The 

researchers examined eggs (n=5, for each species) of little egret (Egretta garzetta), blackcrowned night heron (Nycticorax nycticorax), Chinese pond heron (Ardeola bacchus) and 

cattle egret (Bubulcus ibi) from Hong Kong, Ziamen and Quanzhou. Levels of decaBDE 

ranged from < 0.5 (in black-crowned night heron from Quanzhou) to 99 ± 130 ng/g lipid 

(in Chinese pond heron from Xiamen). The higher abundance of BDE209 in birds from 

Xiamen appeared to correspond with high production of electronics in Xiamen. Exposure 

concentrations were not determined in this study. 

Johnson-Restropo et al. (2005) reported the findings of a monitoring study of marine 

fish—teleosts, Atlantic stingray (Dasyatis sabina) and sharks (Rhizoprionodon 

terraenovae and Carcharhinus leucas)—and dolphins (Tursiops truncatus) of coastal 

Florida. BDE209 (decaBDE) was not detected (MDL=0.022 ng/g lipid) in muscle 

samples from silver perch (Bairdiella chrysoura), striped mullet (Mugil cephalus), 

spotted seatrout (Cynoscion nebulosus) and red drum (Sciaenops ocellatus), and also not 

detected (MDL=0.022 ng/g lipid) in blubber samples from bottlenose dolphin. The 

measured concentrations of decaBDE in hardhead catfish (Arius felis) and Atlantic 

stingray were reported to be 4.5 ng/g lipid and 0.1 ng/g lipid, respectively, while for 

sharks, the decaBDE concentration ranged from 16.9 ng/g lipid in spiny dogfish (Squalus 

acanthias) to 778 ng/g lipid in bull shark (Carcharhinus leucas). DecaBDE (i.e., 

BDE209) was the most abundant congener in sharks, suggesting either that there was 

preferential exposure to decaBDE or that preferential accumulation of decaBDE was 

occurring in shark species. 

Gauthier et al. (2008) conducted monitoring of bird eggs and reported on temporal trends 

(1982–2006) for PBDEs, and most notably, for BDE209 in pooled samples of herring 

gull (Larus argentatus) eggs from seven colonies on lakes Superior, Michigan, Huron 

and Ontario, and on the Detroit and Niagara rivers. BDE209 concentrations in 2006 egg 

pools ranged from 4.5 to 20 ng/g ww and composed 0.6–4.5% of the total PBDE 

concentrations measured. The authors noted mean doubling times for BDE209 

concentrations in eggs of 2.1–3.0 years. For octa- and nonaBDEs, the mean doubling 

time was determined to be 3–11 years and 2.4–5.3 years, respectively. Based on this rapid 

increase in concentrations over time, they indicate that the potential for decaBDE to 

bioaccumulate appears to have been previously underestimated, but they did not quantify 

exposure concentrations. The researchers also suggested that congener patterns in egg 

samples suggested meta-position metabolic debromination of BDE209 and -207 to 

BDE197. 

Shaw et al. (2007) reported decaBDE concentrations in blubber of harbour seals (Phoca 

vitulina concolor) of the northwest Atlantic coast from 1991 to 2005. The BDE209 

congener (decaBDE) was detected in 2 of 4 samples at concentrations up to 7.4 ng/g 

lipid. 
14

DeWit et al. (2006) summarized decaBDE levels in Arctic biota from many additional 

studies: 

o concentrations of 0.025 to 0.12 ng/g ww in moss from Norway (Schlabach 

et al. 2002); 

o 0.5, 0.1 and 0.8 ng/g lipid in grouse (Lagopus sp.), lynx (Lynx lynx) and 

moose (Alces alces) liver, respectively, from Norway (Mariussen et al. 

2004); 

o detectable (but not quantified) amounts in eggs of birds of prey from 

northern Norway (Herzke et al. 2005); and 

o 3.6–33 ng/g lipid in the muscle tissue of blue mussels (Mytilus edullis) and 

0.98–0.99 ng/g lipid in the liver of Atlantic cod (Gadus morhua) from 

Norway (Fjeld et al. 2004). 

Verreault et al. (2007) monitored decaBDE in egg yolk and plasma of male and female 

glaucous gulls (Larus hyperboreus) from the Norwegian Arctic. DecaBDE (i.e., 

BDE209) was “virtually non-detectable” in egg yolk and plasma, but the authors 

speculated that the presence of detectable concentrations of octa- and nonaBDEs in the 

samples suggested accumulation and subsequent in vivo debromination of decaBDE. 

Although the detection limit for decaBDE was not reported, detection limits for hepta- to 

nonaBDEs were reported to be 0.16 ng/g ww and are presumably the same for decaBDE. 

Ismail et al. (2009) evaluated concentrations of BDE209 along with other flame 

retardants in archived Lake Ontario, Canada, lake trout (Salvelinus namaycush) samples 

collected between 1979 and 2004. Concentrations of BDE209 were consistently detected 

in lake trout samples throughout the study period and these concentrations ranged from 

2.3 to 12 ng/g lipid (0.27 to 1.3 ng/g ww). In contrast to the other PBDE congeners, a 

large increase (approximately fourfold) in lake trout BDE209 occurred between 1998 and 

2004. The study also found a significant increase in BDE209 concentrations between 

1979 and 2004, with an overall doubling time of 19 years (p < 0.05). 

 

Guo et al. (2008) conducted an analysis of PBDE concentrations and compositional 

profiles in tissues of freshwater and marine wild fish, and farmed fish in the Pearl River 

Delta (PRD) of China. BDE209 was detected in 70 of the total 187 samples, with a range 

of 0.39 to 59.9 ng/g dry weight (dw), in skin, gills, gastrointestinal tract, liver and muscle 

samples. However, the authors suggested that this detection frequency was likely 

underestimated due to the much higher analytical detection limit for BDE209 (10 ng/g) 

than that for other lower brominated PBDE congeners (0.2 ng/g). The highest and lowest 

BDE209 ratios (to sum of total analyzed PBDEs) were found in skin and liver, 

respectively (mean/maximum of 48.0/99.2% in skin and 8.2/83.3% in liver). The highest 

and lowest BDE209 concentrations occurred in skin and liver, with the median levels of 

95.5 and 2.54 ng/g lipid (based on pooled tissue samples of skin, gills, gastrointestinal 

tract, liver and muscles for all fish), respectively. The authors proposed that this may be 

attributed to the fact that the liver is a primary tissue for biotransformation of organic 

compounds and that BDE209 has a low half-life time in fish. Guo et al. (2008) suggested 

that BDE209 could be accumulated in gills and gastrointestinal tract since the 

lipid-normalized BDE209 concentrations in gills and gastrointestinal tract from their 
15

study were actually slightly higher than those in liver and muscle. The study concluded 

that BDE209 is accumulative in fish tissues under natural environments, and speculated 

that BDE209 may bioaccumulate significantly, particularly in biota from areas heavily 

polluted with decaBDE, such as the Pearl River Delta. 

Potter et al. (2009) measured PBDE levels in 23 peregrine falcon (Falco peregrinus) 

eggs, obtained between 1993 and 2002 from 13 nests covering 11 locations in the 

Chesapeake Bay region of the United States. The mean BDE209 contribution to total 

PBDEs measured in egg samples was 5.9% and ranged from 0% for eggs in or near 

undeveloped land to 18.6% for eggs from a densely populated area. The highest BDE209 

burden found in study eggs was 48.2 ng/g ww from an urban highway bridge site. The 

order of PBDE congener dominance (from most to least dominant) in the peregrine 

falcon eggs analyzed typically was BDE 153 > BDE 99 > BDE 100 > BDE 154 > 

BDE 209 > BDE 183 > BDE 197 > BDE 47 > BDE 207. The authors suggested that the 

source of the higher brominated congener signature in peregrine falcon eggs remains 

unresolved, and that the answer may relate to the bird’s biotransformation capabilities. 

BDE209 was detected in all 114 eggs in a peregrine falcon (Falco peregrinus) egg study 

in the northeastern United States (Chen et al. 2008). Eggs were collected from 1996 to 

2006 (excluding 1997 and 1998). Concentrations ranged from 1.4 to 420 ng/g ww for 

BDE209. The authors reported that the median concentration of 26 ng/g ww (or 480 ng/g 

lipid) was much higher than observed in concentrations of European birds (e.g., 

Greenland peregrine eggs, Belgian buzzards and sparrowhawks), but comparable to those 

in Chinese kestrels. BDE209 concentrations and their contribution to the total PBDE 

concentrations were significantly higher in urban than in rural eggs (p < 0.005). In 

contrast to the other PBDE congeners, BDE209 concentrations were shown to increase 

significantly (r = 0.348, p < 0.005), with a doubling time of 5 years. In addition to 

BDE209, eight nona- and octaBDE congeners were frequently detected in this study. 

Together with BDE209, they constituted 16–57% of the total PBDE concentration 

determined in urban eggs and 4.9–53% in rural eggs. 

Kunisue et al. (2008) analyzed the spatial trend of PBDE levels and congener patterns in 

avian species by using stored historical tissue samples taken from open sea, Japanese 

coastal and inland birds. PBDEs were detected in all the avian pectoral muscle samples 

analyzed in this study; however, BDE209 was not detected in two of the three open-water 

bird samples, one of two coastal bird samples, and one of four inland bird samples. 

Japanese coastal and inland birds accumulated higher PBDE levels than open sea birds. 

In addition, accumulation of higher brominated congeners such as BDE209 was 

predominant in the omnivorous species that lived closest to areas inhabited by humans. 

For example, the mean concentration of BDE209 in jungle crow (Corvus macrorhynchos) 

samples from 1998 was 440 ng/g lipid. In the jungle crow, BDE153, BDE183 and 

BDE209 were the predominant congeners, and relatively higher proportions of octa- and 

nona-BDE congeners were also found compared with other birds. 
16

Additional studies are also described in Section 2.2. These studies contained sufficient 

information for conducting a ratio-based evaluation of bioaccumulation or 

biomagnification. 

2.1.3 Synthesis of Biota Concentration Data 

The detected concentrations of decaBDE in a wide range of biota provide confirmation 

that this substance can accumulate, at least to some degree, in organisms. However, these 

data provide little evidence that decaBDE may be bioaccumulative. The main 

shortcoming is that the data fail to compare the measured concentrations with either those 

determined in the environment in which the biota reside (e.g., as a BAF or BCF) or the 

prey that the biota consumes (e.g., as a BMF or TMF). 

In the absence of a ratio-based assessment of whether decaBDE is bioaccumulative or 

biomagnifying, it is still possible to subjectively judge whether the observed 

concentrations in top predators appear high, indicating potential trophic magnification. 

Concentrations of decaBDE have been observed to be relatively high (e.g., exceeding 

100 ng/g lipid) in some top predator species, including 

• birds of prey, especially kestrel, sparrowhawk and owl from China; peregrine 

falcon, sparrowhawk and kestrel in the UK and Sweden; peregrine falcon in 

Greenland; and buzzard in Belgium; 

• red fox in Belgium; 

• sharks of coastal Florida; 

• marine mammals such as harbour porpoise and white-beaked dolphin; and 

• some marine bird species, including heron from Sweden and glaucous gulls from 

Byoroya and Svalbard, Norway. 

Of these, the Greenland peregrine falcon results and the coastal Florida shark results may 

provide the strongest evidence that trophic magnification might be occurring, given the 

potential remoteness of the sites of sampling. However, Vorkamp et al. (2005) also note 

that the peregrine falcon (Falco perigrinus) subspecies of Greenland from which samples 

were taken migrates to Central and South America in the winter, following a route along 

the Atlantic coast of North America. As a result, there is a potential that the organisms 

were exposed to locationally specific contaminated environments that may explain their 

relatively high loading of decaBDE. Although high decaBDE concentrations are 

observed in Florida sharks, similarly high concentrations are not observed in teleost fish 

species from the same ecosystem, which would likely make up the prey species for shark. 

While trophic magnification is a potential explanation for this difference, it is also 

possible that the sharks are consuming contaminated refuse. 

For the remaining results, several factors tend to confound the findings with respect to 

whether decaBDE is bioaccumulative. These include 

• the potential that the sampled species inhabit decaBDE hotspots close to 

industrialized areas; and 
17

• the large number of non-detects within the same studies, and species in which 

relatively high concentrations were observed. 

While there are some high BDE209 concentrations in some biota, overall the data from 

surveillance studies directly contradict a generalized conclusion that the concentration of 

decaBDE is relatively high in top predators. The surveillance data show 

• relatively low and often non-detected concentrations of decaBDE in marine fish 

species, even though some of these species such as salmon or cod could be 

considered top predators within their respective food webs (see Appendix A for 

summary of these data); 

• relatively low and often non-detected concentrations of decaBDE in mammalian 

predators such as grizzly bear, polar bear and lynx which feed near the top of their 

respective food webs; 

• relatively low and often non-detected concentrations of decaBDE in marine 

mammals which feed near the top of their respective food webs; 

• relatively low and often non-detected concentrations of decaBDE in 

marine/aquatic birds which feed at the top of their respective food webs; and 

• relatively low and often non-detected concentrations of decaBDE in birds of prey, 

often in the same studies as those where relatively high concentrations were also 

observed. 

Generally speaking, it appears that the higher concentrations of decaBDE in biota are 

anomalies when the broader dataset of decaBDE concentrations is considered. Thus, the 

evidence based on concentrations in top predators fails to provide a strong indication that 

decaBDE is bioaccumulative or biomagnifying in food webs. 

2.2 Bioconcentration, Bioaccumulation and Biomagnification 

Data 

This section first summarizes the data respecting decaBDE bioaccumulation considered 

in the PBDE screening assessment report (Environment Canada 2006a; Environment 

Canada 2006b), then examines new data available after 2004 and its significance. 

2.2.1 Information Evaluated in the Screening Assessment 

• The Japanese Ministry of International Trade and Industry (MITI 1992) determined that 

the BCF for carp ranged from < 5 to < 50 in their 6-week bioconcentration study 

(recalculated to < 3000 by European Communities (2002)). Given the exceptionally 

low water solubility limit of decaBDE it is not expected that this substance will be 

appreciably taken up from the water phase by aquatic organisms. 

• Stapleton et al. (2004) exposed juvenile common carp (Cyprinus carpio) to decaBDE 

(> 98% purity as reported by Cambridge Isotope Laboratories) amended in food on a 

daily basis for 60 days (d) (940 ng/d/fish) followed by a 40-d period in which fish were 

fed clean food. DecaBDE (i.e., BDE209) was not detected in whole fish tissues during 

the exposure or depuration periods (detection limit of approximately 1 μg/kg ww); 
18

however, several peaks were observed in the chromatograms of the exposed fish that 

were not observed in the control fish, suggesting transformation of decaBDE. Their 

results suggested that at least 0.44% of decaBDE was bioavailable in the form of its 

metabolites. This value could be higher if other metabolites were present which were 

not determined in these studies. 

• Kierkegaard et al. (1999) dosed juvenile rainbow trout (Oncorhynchus mykiss) with 

food contaminated with Dow FR-300 (composition not determined in this study, but 

reported as 77.4% decaBDE, 21.8% nonaBDE and 0.8% octaBDE by Norris et al. 

1973, 1974). A small proportion of the test material was taken up during the 120-d 

exposure phase of the experiment, amounting to about 0.02–0.13% based on the muscle 

concentrations of the total hexa- to decaBDEs present, or approximately 0.005% based 

only on the decaBDE concentrations in muscle and the mean dietary dose of Dow 

FR-300. The authors did not report the decaBDE concentration in food, meaning that a 

dietary BAF cannot be estimated from the study data. 

• Metabolism studies using rats (Norris et al. 1973, 1974; El Dareer et al. 1987) 

suggested that decaBDE has a very low bioaccumulation potential in mammalian 

species. For instance, Norris et al. (1973, 1974) dosed male and female rats with 1.0 mg 

of 14C-labelled DecaBDE as a suspension in corn oil and found that around 90.6% of 

the administered 14C-labelled DecaBDE was excreted in the feces within 24 hours (h), 

and by 48 h, all of the administered chemical had been excreted. Tissue accumulation 

studies in which rats were fed diets of decaBDE at a rate of 0.1 mg/kg body weight per 

day showed that bromine contents in various tissues were not significantly greater than 

those of the controls. The bromine content of the adipose tissue of decaBDE-dosed rats 

was found to be significantly increased at the p < 0.03 level but not at the p < 0.01 level 

when compared with the controls (Norris et al. 1973, 1974). 

The bioaccumulation data from these studies are also summarized in Appendix B. 

Additional studies were also reviewed in the screening assessment but their findings did 

not allow estimates of bioaccumulation parameters. Rather, they are more relevant to the 

evaluation of metabolite formation, and are discussed in Section 3. 

2.2.2 New Bioaccumulation Data 

The bioaccumulation data from studies published after 2004 are discussed in this section 

and summarized in Appendix B. 

Studies on Aquatic Species 

Stapleton et al. (2006) exposed 45 juvenile rainbow trout (Oncorhynchus mykiss) to 

spiked food containing decaBDE (purity reported as 98.7%, no further characterization of 

test material for impurities was undertaken) for a period of 5 months. The 45 fish were 

randomly distributed to three experimental tanks while an additional 15 fish that were fed 

non-spiked food were kept in a separate tank. The concentration of decaBDE in the 

spiked food was 940 ng/g and the fish (average weight 91.2 g) were fed at a rate of 1% of 

their body weight per day. One fish from each tank was sacrificed for analysis at 9 time 

points throughout the 5-month exposure period. Blood was sampled from individual fish 

at the initiation of exposure and at three time points during the final 3 months of the 
19

study. Samples of blood serum, intestine, liver and carcass at each sampling time were 

each analyzed for decaBDE and lower brominated BDE congeners. 

The net uptake of decaBDE during the experiment was estimated at 3.2% based on the 

total burden of hepta- through decaBDE congeners present in the carcass, or 3.7% if the 

liver was included in the calculation. The decaBDE concentration on the final day of 

exposure was highest in the liver (342 ng/g ww) followed by the intestine (~60 ng/g ww, 

read from graph), serum (26–40 ng/g ww) and carcass (5.3 ng/g ww). The detection limit 

in this study was 1 ng/g ww. Several hepta-, octa- and nonaBDE congeners were also 

accumulated, potentially, as a result of decaBDE debromination (fish were dosed only 

with decaBDE, and background decaBDE concentrations in the control fish ranged from 

< 0.5 ng/g to 0.5 ng/g). These debromination results are discussed further in Section 3.1. 

Using concentration data from day 112, the reported lipid content of the test fish, and the 

known decaBDE concentration/lipid content in food reported by Stapleton et al. (2006), it 

is possible to estimate BMFs for decaBDE on its own and for the combined burden of 

decaBDE plus debrominated congeners. Table 2-1 summarizes the calculations. 

Table 2-1: BMFs estimated from the decaBDE rainbow trout feeding study by Stapleton et 

al. (2006; and personal communication from HM Stapleton to Environment Canada, 

January 2008; unreferenced) 

 Food1 Carcass1 Serum1 Liver1

DecaBDE concentration (ng/g 

lipid) 9307 204 33002

 11 958 

Fraction of decaBDE in total 

BDE burden n/a 0.25 0.68 0.92 

Total BDE concentration (ng/g 

lipid)3 n/a 831.3 4853 16 163 

DecaBDE BMF n/a 0.02 0.35 1.28 

Total BDE BMF n/a 0.09 0.52 1.74 

1

Lipid contents: food – 10.1%; carcass – 4.5%; serum – 1% (assumed); liver – 2.3%. 2

Median of reported values (26 and 40 ng/g ww) divided by estimated lipid content. 3

Inferred by dividing decaBDE by the fraction of decaBDE in total BDE burden. 

It is expected that the higher lipid content in food (10.1% lipid) compared with carcass 

and liver (4.5% and 2.3%, respectively) may cause the ww BMFs to underestimate the 

bioaccumulation potential of decaBDE, and as a result, lipid-weight BMFs were 

calculated. For decaBDE on its own, the BMFs ranged from 0.02 (carcass) to 1.28 (liver) 

whereas for the total BDE burden, the BMFs ranged from 0.09 to 1.74. 

Based on the findings of other studies (refer to Section 3), it is possible that decaBDE 

could be transformed to other transformation products not analyzed in this study. The 

Stapleton et al. (2006) study did not include these potential products as analytes. If they 

are formed in rainbow trout, then the reported net uptake of neutral BDEs only would 

underestimate the actual total uptake of decaBDE. If metabolites other than hepta-, octa-, 

and nonaBDEs were being formed and persisting in the fish, then the BMFs calculated 

above would underestimate the total accumulation potential of decaBDE-related 

compounds. 
20

Tomy et al. (2004) studied the uptake by juvenile lake trout (Salvelinus namaycush) of 

twelve tetra- to heptaBDEs (Wellington Laboratories, all purities > 96%) plus DecaBDE 

(technical grade, Great Lakes Chemical Corp., purity not provided) from spiked 

commercial fish food. Test fish were exposed to spiked food for 56 d followed by a 112-d 

elimination period. Seventy fish each were exposed to low and high concentrations 

(measured in food) of technical-grade DecaBDE, and a non-exposed control group was 

also monitored (concentrations measured in food). Significant uptake of decaBDE was 

observed for both the low- and high-exposure treatments. For the low-exposure 

treatment, depuration of chemical during the elimination phase was non-detectable (slope 

not significant) and the absorption efficiency, half-life and BMF were not estimated. For 

the high-exposure treatment, the absorption efficiency was estimated at 5.2% (Tomy et 

al. 2004; and personal communication from G Tomy to Environment Canada, July 2009, 

unreferenced) with a half-life of 26 ± 5 d and a BMF1

 of 0.3. Although lower brominated 

PBDE congeners appeared to be bioformed in the fish, it was not possible to include the 

debrominated congeners in the BMF estimates because similar congeners were also 

present in the spiked food or potentially present as impurities in the technical-grade 

DecaBDE used in the study. In addition, the study contained other uncertainties such as 

the use of fiberglass aquaria which may have resulted in some adsorption of test material. 

Ciparis and Hale (2005) examined the bioavailability and accumulation of multiple 

PBDEs, including decaBDE, from sediments and biosolids to the aquatic oligochaete, 

Lumbriculus variegatus. Oligochaetes were exposed to either composted biosolids 

containing 1600 ng/g total PBDEs or artificial sediments spiked with technical Penta- and 

DecaBDE formulations (1300 ng/g total PBDEs). The experimental protocol included a 

28-d uptake phase followed by a 21-d elimination phase. Following solvent extraction, 

clean-up in a size exclusion column, and further purification using solid-phase extraction 

columns, decaBDE was quantified from substrates and tissues on a gas chromatography 

(GC) device equipped with a halogen-selective electrolytic conductivity detector with 

MDLs of 190 ng/g and 20 ng/g for tissues and substrates, respectively. Although 

significant accumulation of lower brominated PBDEs (especially BDE47 and BDE99) 

was observed with both biosolids and spiked sediments, uptake of decaBDE was minimal 

and it was not possible to estimate steady-state sediment BSAFs or kinetic parameters for 

decaBDE accumulation. The authors speculated that the bioavailability of decaBDE was 

limited by its high log Kow (suggesting that desorption from sediment particles is 

minimal) and large molecular size, which may impede its transport across cell 

membranes. 

Burreau et al. (2004, 2006) reported the results for three separate food web monitoring 

programs for PCBs and PBDEs. Burreau et al. (2004) sampled perch (Perca fluviatilis; 

n=120, 33 individuals), roach (Rutilus rutilus; n=23, 8 individuals) and pike (Esox lucius; 

n=51, 25 individuals) in the Lumparn estuary in the Åland archipelago in the Baltic Sea 

 

1

 BMF = αF/kd where α is the absorption efficiency, F is the feeding rate on a lipid basis and kd is the total 

elimination rate constant 
21

and analyzed muscle tissue composites from each species.2

 Burreau et al. (2006) 

described monitoring studies for the Baltic Sea and the Atlantic Ocean (south of Iceland). 

The Baltic Sea study was conducted in 1998 and sampled zooplankton (Calanoid 

crustacea; n=3 net tows), sprat (Sprattus sprattus; n=6), herring (Clupea harengus; n=5) 

and Atlantic salmon (Salmo salar; n=10), while the Atlantic Ocean study was conducted 

in 1999 and sampled zooplankton (Calanoid sp.; n=10), small herring (Clupea harengus; 

n=6), large herring (Clupea harengus; n=10) and Atlantic salmon (n=10). The detection 

limits ranged from 140 to 148 pg/g. Analytical detection limits ranged from 14 to 

14.8 pg/g ww. Determinations of decaBDE were blank-corrected. Table 2-2 provides a 

summary of median observed concentrations in the Baltic Sea and North Atlantic Ocean 

biota samples. 

Table 2-2: Summary of observed concentrations of decaBDE in freshwater and marine 

biota from the Lumparn Estuary, Baltic Sea and North Atlantic Ocean (Burreau et al. 2004, 

2006) 

Location Species 

Median 

concentration 

(ng/g lipid) 

Number of 

samples with 

decaBDE detected 

/ Number of 

samples analyzed 

Lumparn Estuary 

Roach 

Rutilus rutilus 48 

3/8 

Lumparn Estuary 

Perch 

Perca fluviatilis 1.3 

12/33 

Lumparn Estuary 

Pike 

Esox lucius 1.7 

4/25 

Baltic Sea 

Zooplankton 

Calanoid sp. 2.1 

3/3 

Baltic Sea 

Sprat 

Sprattus sprattus 0.082 

3/6 

Baltic Sea 

Herring 

Clupea harengus 0.24 

4/5 

Baltic Sea 

Atlantic salmon 

Salmo salar 0.41 

3/10 

North Atlantic Ocean 

small herring 

Clupea harengus 0.31 

6/6 

North Atlantic Ocean 

large herring 

Clupea harengus 0.039 

3/6 

North Atlantic Ocean 

Atlantic salmon 

Salmo salar not detected 

0/10 

To examine the potential for food web biomagnification of decaBDE in the biota data for 

each Baltic Sea food web (roach-perch-pike; zooplankton-sprat-herring-salmon), the 

authors conducted an analysis of trophic magnification. This involved a regression of 

lipid-normalized concentration vs. δ15N according to the following model: 

 

2

 The disparity between the n of samples and the n of analysis is explained by the fact that only the samples 

in which the congener was detected are reported. 
22

C = A•e

(B•δ15N) 

Where C is the biota concentration (lipid-normalized), A is a constant representing δ15N 

at the base of the food chain and B represents the “biomagnification power” of the 

substance. A positive B-value indicates biomagnification while a negative B-value 

indicates trophic dilution of chemical concentrations. B is similar to a TMF except that a 

TMF is expressed on an arithmetic, rather than logarithmic basis, and the TMF is based 

on a regression with trophic level (estimated from δ15N) rather than the δ15N content 

itself. The B-values for both food webs were not significantly different from zero, 

indicating that biomagnification of decaBDE did not appear to be occurring in these food 

webs. Failure to detect decaBDE in salmon from the Atlantic Ocean precluded a similar 

analysis for this food web. 

Using the reported concentration data, it is also possible to estimate lipid-normalized 

BMFs for specific predator-prey combinations; these are summarized in Table 2-3. BMFs 

range from 0.03 to 5, depending on the predator-prey combination, suggesting that 

biomagnification was taking place in some predator-prey combinations. However, it is 

important to consider that the exact feeding relationships for these food webs are 

unknown, resulting in considerable uncertainty in these BMF estimates. 

Table 2-3: Estimated BMFs for decaBDE in sampled biota from Lumparn Estuary, Baltic 

Sea and Atlantic Ocean pelagic food webs reported by Burreau et al. (2004, 2006) 

Location Predator/Prey 

BMF (lipidnormalized) 

Lumparn Estuary 

perch/roach 

Perca fluviatilis / Rutilus rutilus 0.03 

pike/roach 

Esox lucius /Rutilus rutilus 0.04 

pike/perch 

Esox lucius / Perca fluviatilis 1.31 

Baltic Sea 

sprat/zooplankton 

Sprattus sprattus / Calanoid crustacea 0.04 

herring/sprat 

Clupea harengus / Sprattus sprattus 2.93 

herring/zooplankton 

Clupea harengus / Calanoid sp. 0.11 

salmon/sprat 

Salmo salar / Sprattus sprattus 5.00 

salmon/herring 

Salmo salar /Clupea harengus 1.71 

Atlantic Ocean 

large herring/small herring 

Clupea harengus / Clupea harengus 0.13 

In evaluating the Burreau et al. (2004, 2006) studies, the United Kingdom (2007a) 

cautions that the relatively high levels of decaBDE in procedural blanks and low 

concentrations of decaBDE in biota samples create uncertainty in the overall 
23

biomagnification analysis. Currently, this appears to be a common issue with field studies 

of decaBDE in biota. 

Shaw et al. (2009) studied the bioaccumulation of PBDEs in northwest Atlantic marine 

food webs. To evaluate the transfer of PBDEs from prey to predator, the study compared 

PBDEs measured previously in harbour seal blubber with whole fish samples of seven 

species of fishes comprising the major prey of harbour seals (Phoca vitulina concolor). 

Eighty-seven individual fish (> 35 cm) were collected off the coast of Maine during the 

May–June 2006 Gulf of Maine Trawl Survey of commercial groundfish stocks. Species 

included silver hake (Merluccius bilinearis, n=10), white hake (Urophycis tenuis, n=17), 

Atlantic herring (Clupea harengus, n=20), American plaice (Hippoglossides platessoides, 

n=10), alewife (Alosa pseudoharengus, n=10), and winter flounder (Pseudopleuronectes 

americanus, n=10). Atlantic mackerel (Scomber scombrus, n=10) were caught by hook 

and line from the same area during June 2006. Whole fish were transported to the 

laboratory on ice where standard length and weight were recorded, then frozen and stored 

at −40°C prior to shipment to the analytical laboratory. Fish whole-body samples were 

pooled and homogenized into 17 composites prior to analysis. For extraction, the fish 

sample (~1.5 g of lipids, between 5 and 100 g tissue) was homogenized and mixed with 

sodium sulphate. After addition of the internal PBDE standards, a mixture of cyclohexane 

and dichloromethane was applied to the column for extraction of PBDEs along with other 

lipophilic compounds and fat. The extract was washed, dried, and after solvent 

evaporation, gravimetric lipid determination was performed. The final extract was 

evaporated by a stream of nitrogen to a final volume of 50 μL containing C13 labelled 

BDE139 as recovery standard. The measurements were performed by high-resolution gas 

chromatography / high resolution mass spectrometry (HRGC/HRMS). 

Total PBDE concentrations in fish ranged from 18.3 to 81.5 ng/g, lipid (overall mean 

62 ± 34 ng/g, lipid), compared with total PBDE concentrations in harbour seal samples of 

80 to 25 720 ng/g lipid (overall mean 2403 ± 5406 ng/g, lipid—analyzed for earlier 

study). BDE209 was detected in 35% of the fish samples and in 25% of the harbour seal 

blubber samples. BDE209 concentrations ranged from non-detect (0.2 ng/g lipid) to 

4 ng/g lipid in fish, and from 1.1 to 7.6 ng/g lipid in seal blubber (mean value=1.2 ng/g, 

lipid). The similarity between fish and seal BDE209 concentrations was in contrast to the 

total PBDE concentrations, which were two orders of magnitude higher in the harbour 

seals than in fish. For BDE209, BMFs from fish to seals ranged from 0.67 (American 

plaice) to 0.75 (Atlantic mackerel) to 1.3 (white hake), which the authors suggested 

represented low biomagnifications potential. This contrasted with BMFs for the other 

PBDEs, which averaged from 17 to 76.5, indicating high biomagnification in this marine 

food web. The authors suggested that the presence of higher brominated congeners, 

including BDE209, at measurable levels in fish and seal tissue, along with high 

biomagnification of BDE153, -155, and -154, suggests recent exposure to the octa- and 

decaBDE formulations in this U.S. coastal marine food web. 

Law et al. (2006) conducted a field study of the trophic magnification of decaBDE in a 

pelagic food web of Lake Winnipeg. Samples of fish, plankton, mussels, sediment and 

water were collected from the south basin of the lake near Gimli, Manitoba. Muscle 
24

tissue from multiple fish species were collected between 2000 and 2002, including 

walleye (Stizostedion vitreum; n=5), whitefish (Coregonus clupeaformis; n=5), emerald 

shiner (Notropis atherinoides; n=5), burbot (Lota lota; n=5), white sucker Catostomus 

commersoni; n=5) and goldeye (Hiodon alosoides; n=3). Samples of net plankton (n=5; 

zooplankton and phytoplankton combined) were collected using horizontal tows with 

160-μm nets (precise date not indicated in article). Mussels (Lampsilis radiata; n=5, 

muscle tissue retained for analysis) were collected by divers in 2002. Sediment grab 

samples were collected at 4 locations with only the surficial 2 cm of sediment retained. 

Water was sampled in 2004 using a Teflon column packed with XAD-2 absorbent. Each 

XAD-2 column was used to sample six 54-L samples from 324 L of water collected. 

Samples were pulled through an inline glass-fibre filter (1-μm pore size) and then onto a 

XAD-2 column. 

All samples were analyzed for decaBDE (and several other chemicals) using GC/MS, 

with additional analyses of organic carbon (OC) for sediments and lipid and δ15N for 

biota. The δ15N measurements were used to estimate trophic position. The detection limit 

of the analytical method used was 0.1 μg/kg for biota and sediment samples, and 15 pg/L 

for water. The decaBDE concentration, lipid and OC contents, and estimated trophic 

level of biota are summarized in Table 2-4. 

Using the trophic levels estimated from δ15N data, the rank order of the trophic levels in 

the pelagic food web was estimated to be mussel ⇒ zooplankton, whitefish ⇒ goldeye, 

white sucker ⇒ burbot, walleye (top predators). A regression of the lipid-normalized 

concentration of decaBDE vs. trophic level was used to estimate a TMF of 3.63 (r2=0.46 

p=0.0001) for decaBDE in the pelagic food web. Predator-prey BMFs (on a lipidnormalized basis) were also calculated using the biota dataset. The estimated BMFs for 

decaBDE ranged from 0.1 to 34, depending on the predator-prey combination. 

While estimated TMFs and BMFs are intended to provide a real-world indication of 

trophic magnification and biomagnification of decaBDE in an aquatic food web, it is 

important to consider some of the uncertainties associated with this study. 

Many of the concentrations of decaBDE in the Law et al. (2007) study were near the 

detection limit, increasing uncertainty in these determinations and raising the possibility 

of false positives. In addition, biomagnification was identified using lipid-normalized 

data; however, certain tissues were characterized by very low lipid concentrations (e.g., 

walleye, burbot and mussel muscle). Such low lipid contents result in extremely uncertain 

concentrations expressed on a lipid weight basis. When biomagnification is evaluated in 

this study on the basis of ww concentrations, biomagnification is not shown to occur. 

There is further uncertainty respecting the appropriateness of lipid-normalization for 

decaBDE. This substance has been suggested to bind protein in some situations, although 

based on chemical structure, protein binding is not expected. It is possible that decaBDE 

is subject to non-specific binding in blood plasma (i.e., lipids) (e.g., Han et al. 2007). As 

 

3

 Note that this is the corrected value published in Law et al. (2007). 
25

a result of this binding, preferential accumulation in liver could occur, but this has not 

been established definitively. 

Table 2-4: Analytical results from Law et al. (2006) for decaBDE, lipid content, organic 

carbon content and δ

15N in water, sediments and biota from Lake Winnipeg 

Sample δ

15N Trophic 

position 

Mean lipid1

 or organic 

carbon2

 content 

Average decaBDE 

concentration 

(ng/g lipid1

; ng/g dry 

weight (dw)2 or pg/L3) 

Water 

(dissolved 

phase) 

n/a n/a n/a <15 pg/L 

Sediment n/a n/a 2% 0.63 

Walleye 

(Stizostedion 

vitreum) 

muscle 

17.8 2.4 1.15% 24.7 

Whitefish 

(Coregonus 

clupeaformis)

muscle 

12.0 0.8 8.78% 3.6 

Mussel 

(Lampsilis 

radiata) 

muscle 

9.5 - 0.32% 50.8 

Zooplankton 

(Calanoid 

sp.) 

9.7 1.00 13.67% 1.2 

Emerald 

shiner 

(Notropis 

atherinoides)

muscle 

16.0 1.9 3.18% 40.3 

Goldeye 

(Hiodon 

alosoides) 

muscle 

16.1 1.95 2.34% 41.6 

White sucker 

(Catostomus 

commersoni)

muscle 

15.2 1.7 2.27% 12.0 

Burbot (Lota 

lota) muscle 

16.6 2.2 0.33% 98.7 

n/a – not applicable 

1

For biota samples 

2

For sediment samples 

3

For water samples 
26

The United Kingdom (2007a) also highlighted the following issues regarding the Law et 

al. (2006) study: 

• There is some uncertainty as to how well the δ15N analysis characterizes trophic levels. 

The resulting trophic structure does not necessarily match that which would be 

expected based on species size and feeding characteristics (e.g., emerald shiner has a 

higher estimated trophic level than whitefish but based on size/life history, it would be 

expected to feed lower in the food web). 

• There is uncertainty as to how well the samples represent the food web, since sample 

sizes were small in some cases, and they were collected at varying times between 2000 

and 2002. 

• Information regarding the actual feeding relationships of the analyzed species is 

lacking. It is unclear whether the species examined feed primarily on one another or 

whether other species that were not sampled could also be an important component of 

the food web. 

In a study of eastern Canadian Arctic marine food webs, Tomy et al. (2008) examined the 

extent of trophic transfer of seven PBDE congeners, including BDE209. PBDEs were 

analyzed in the blubber of the beluga whale (Delphinapterus leucas, n=5), narwhal 

(Monodon monoceros, n=5) and walrus (Odobenus rosmarus, n=5). Whole-organism 

homogenates of Arctic cod (Boreogadus saida, n=8), shrimp (Pandalus borealis and 

Hymenodora glacialis, n=5), clams (Mya truncate and Serripes groenlandica, n=5), 

deepwater redfish (Sebastes mentella, n=5) and mixed zooplankton (n=5) were analyzed. 

Samples were collected in various parts of the Canadian eastern Arctic between 1996 and 

2002, and archived. Marine mammal blubber was extracted using a ball mill shaker with 

anhydrous sodium sulphate and hexane:dichloromethane (DCM). Each cell was spiked 

with recovery internal standards, then shaken and left to stand 2–4 h before centrifuging, 

then decanted. The extraction was repeated twice, combining the decanted extracts. Fish, 

shrimp and clam tissues were homogenized with dry ice in a laboratory blender, then 

stored overnight in a –20°C freezer to allow for sublimation of the CO2. Thawed tissue 

was weighed and mixed with pelleted diatomaceous earth (Hydromatrix) then added to a 

cell along with the recovery internal standards RIS and extracted using an accelerated 

solvent extractor (ASE 300). Zooplankton samples were weighed frozen and 

homogenized by directly mixing with Hydromatrix prior to accelerated solvent 

extraction. Void space was filled with sand. After extraction, anhydrous sodium sulphate 

was added to the collection bottles to remove water. Extracts were reduced in volume and 

filtered. Lipid content was determined gravimetrically in an aliquot of extract, while lipid 

was removed from the remainder of the extract by gel permeation chromatography. After 

volume reduction, samples were further cleaned using Florisil according to Law et al. 

(2006). The BDE fraction was reduced in volume to 200 μL and instrument performance 

internal standards (10 μL of 2 ng/μL aldrin) were added. PBDEs in tissue samples were 

analyzed by gas chromatography electron capture negative ion mass spectrometry (GCECNIMS). PBDEs were detected in selected ion monitoring (SIM) mode using the [Br]- 

ions (m/z 79, 81) and an external standard solution containing the BDE-mix (36 BDE 

congeners) and BDE-209 for quantification. Procedural and instrument blanks were used, 

and all PBDE samples were blank-corrected. The relative trophic level of the organisms 
27

was determined by stable isotopes (δ15N), analyzed at the University of Georgia. BMF 

was determined as the trophic level adjusted ratio of the concentration in the predator 

tissue to that of the concentration in the prey tissue. 

The total concentration of the seven congeners analyzed ranged from 0.4 ng/g lipid in 

walrus to 72.9 ng/g lipid in zooplankton. BDE209 mean concentrations ranged from nondetect (walrus, beluga and narwhal) to 18.7 lipid ng/g (zooplankton). BDE209 was found 

to contribute significantly to the body burden of total PBDEs in the lower trophic level 

organisms: 60% in redfish and 75% in arctic cod, for example. Conversely, in the upper 

trophic level organisms like beluga and narwhal, BDE209 accounted for less than 2% of 

the total BDE burden, whereas BDE47 accounts for over 40% in these samples. The 

authors suggested that elevated BDE209 levels in lower trophic level organisms may 

reflect greater exposure to this compound through zooplankton, and limited metabolic 

capabilities with respect to BDE209. BDE209 lipid-adjusted BMFs for individual 

predator–prey relationships were as follows: beluga (blubber):cod (whole body) < 1; 

beluga (blubber):redfish (whole body) < 1; and cod (whole body):zooplankton (whole 

body) < 1. BMFs were not calculated for narwhal:redfish and narwhal:cod. A statistically 

significant TMF of 0.3 was estimated for BDE209 (p=0.002; correlation coefficient R2

=0.25). The authors concluded that BDE209 concentrations decreased with trophic level, 

suggesting metabolic depletion of this congener or reduced assimilation up the food web. 

The authors suggested that these results are consistent with other studies that have shown 

that BDE209 is not an abundant congener in higher trophic level organisms. The authors 

speculated that this could be due to an enhanced BDE209 metabolic capability among 

higher trophic level organisms. 

In a follow-up to their eastern Canadian Arctic food web study, Tomy et al. (2009) 

conducted a study of western Canadian Arctic marine food webs to examine the

trophodynamics of PBDEs. PBDEs were measured in beluga whale (Delphinapterus 

leucas), ringed seal (Phoca hispida), Arctic cod (Boreogadus saida), Pacific herring 

(Clupea pallasi), Arctic cisco (Coregonus autumnalis), pelagic amphipod (Themisto 

libellula) and Arctic copepod (Calanus hyperboreus). The animals selected were from the 

sample archived repository at Fisheries and Oceans Canada. The brominated compounds 

were measured in the blubber of ringed seal and beluga, in the whole organism minus 

liver for the pelagic fish, and pooled composites for the invertebrates. Extraction methods 

were as per Tomy et al. (2008) (see above study). PBDEs were analyzed by 

high-resolution GC-ECNIMS. BDE209 was quantified by isotope dilution using 13C12-

BDE-209 and the m/z values of 486.6/488.6 and 494.6/496.6 for quantification and 

confirmation for the native and isotope internal standard, respectively. Stable isotopes of 

nitrogen, expressed as δ15N, were analyzed at the stable isotope laboratory at the 

University of Winnipeg (Manitoba, Canada). BMFs were determined as per Tomy et al. 

(2008) (see above study). 

The relative trophic level status of the studied individual organisms was established using 

stable isotopes of δ15N. This analysis indicated the following food web: beluga whale > 

ringed seal > Arctic cod > Pacific herring and Arctic cisco > pelagic amphipod > Arctic 

copepod. The total concentration of the seven congeners analyzed (BDE47, -85, -99, 
28

-100, -153, -154 and -209) ranged from 2.6 ng/g lipid in ringed seals to 205.4 ng/g lipid 

in Arctic cod. The median concentrations of BDE209 ranged from 0.04 ng/g lipid in the 

calanus (arctic copepod) to 7.23 ng/g lipid in Arctic cisco. The rank order of median 

BDE209 concentrations was: cisco > Arctic cod > herring > beluga and ringed seal > 

themisto > calanus. The researchers calculated lipid-normalized BMFs for BDE209 as 

follows: ringed seal (blubber):cod (liver) = 0.3; beluga (blubber):cod (liver) = 0.3; beluga 

(blubber):herring (liver) = 0.9; beluga (blubber):cisco (liver) = 0.03; cod (liver):calanus 

(whole body) = 12.7; cod (liver):themisto (whole body) = 4.8. The results suggested that 

depletion of BDE209 is taking place in the higher trophic level animals but that 

biomagnification may be occurring in the lower trophic level species. Although 

biomagnification was apparent in the lower food chain, there is some uncertainty in the 

results. For instance, it appears that the organisms included in the study were collected at 

different times and locations. None of the PBDEs showed a statistically significant 

positive relationship with trophic level, and no statistically significant TMF was found 

for BDE209. 

Wu et al. (2009) evaluated the biomagnification of PBDEs, including BDE209, in a 

highly contaminated freshwater food web from southern China. Wild aquatic species 

representing different trophic levels were sampled in 2006 from a reservoir surrounded 

by several e-waste recycling workshops. Two top predator species, water snake (Enhydris 

chinensis) and northern snakehead (Channa argus), and their prey, mud carp (Cirrhinus 

molitorella), common carp (Cyprins carpio), crucian carp (Carassius auratus), and prawn 

(Macrobrachium nipponense), were sampled. Chinese mysterysnail (Cipangopaludina 

chinensi) was also collected by hand from the shallow water around the reservoir. 

Samples of small organisms (e.g., Chinese mysterysnail, prawn and mud carp) were 

pooled. The body weight and length of samples were measured, then stored at -20°C until 

further treatment. For chemical analysis, samples were thawed and whole-body 

homogenized, after which two sub-samples were taken from each specimen: one for 

PBDEs determination, and one for nitrogen stable isotope analysis. Samples were ground 

with ashed anhydrous sodium sulphate, spiked with surrogate standards, and extracted 

with hexane/acetone (1/1, v/v) for 48 hours. Lipids (gravimetrical method) were 

determined on an aliquot of the extract. PBDEs were analyzed by GC/MS ECNI mode 

and operated in SIM mode. Recoveries of the spiking blanks ranged from 76.9 to 105.2% 

for PBDEs. Relative standard deviations (RSD) of sample triplicates were less than 24% 

for BDE209. Stable isotopes of nitrogen, expressed as δ15N, were analyzed by a flash EA 

112 series elemental analyzer interfaced with a Finigan MAF ConFlo 111 isotope ratio 

mass spectrometer. 

TMFs were determined as per Tomy et al. (2004) using lipid-normalized concentrations 

and the trophic levels of the food web components. The TMF values ranged from 0.26 to 

4.47 for PBDEs. For BDE209, the TMF was 0.26, suggesting trophic dilution, although 

the result was only marginally statistically significant at p=0.053. The authors suggest 

that different PBDE levels in the organisms, different environmental conditions (e.g., 

warmer water temperature), and the different food web composition could be a factor 

when comparing the lower TMF results of this study to northern food web studies. No 

significant correlation between TMFs and log KOW was found for PBDEs, and the authors 
29

suggested that other factors (e.g., metabolism) might play a more important role in 

PBDEs transfer in the food web. 

Yu et al. (2009) analyzed concentrations of 10 PBDE congeners in a Pearl River Estuary 

food web of the Pearl River Delta region of southern China, to understand the 

accumulation behaviour of these substances. Two hundred and fifty-four biota samples 

(four species of invertebrates and ten species of fish) were collected from the Pearl River 

Estuary between 2005 and 2007. These species included sand swimming crab (Ovalipes 

punctatus, Samoan crab (Scylla serrata), ark shell (Tegillarca granosa), oncomelania 

(Oncomelania hupensischiui), common mullet (Mugil cephalus), red eelgoby 

(Odontamblyopus rubicundus), robust tonguefish (Cynoglossus robustus), slimy 

spinefoot (Siganus canaliculatus), silver sillago (Sillago sihama), pompano (Psenopsis 

anomala), Japanese eel (Anguilla japonica), flatheadfish (Platycephalus indicus), large 

yellow croaker (Pseudosiaena crocea), and Bombay duck (Harpodon nehereus). One 

hundred and twenty-four individual or composite samples were analyzed for PBDEs. 

Samples were homogenized, extracted and analyzed by GC/MS ECNI and operated in 

SIM mode. The concentrations of PBDEs in organisms varied from 6.2 to 208 ng/g lipid 

weight. However, BDE209 was detected in only 18% of samples, ranging from nondetect to 1.6 ng/g lipid. Because of its low presence, BDE209 was excluded from further 

analysis of the report and no TMF was calculated for this congener. 

In a study of in vivo and environmental debromination of decaBDE, La Guardia et al. 

(2007) monitored decaBDE concentrations in sediments and aquatic organisms in the 

receiving environment of a wastewater treatment plant (WWTP) located in Roxboro, 

North Carolina. All samples were extracted and purified using size-exclusion 

chromatography and then analyzed for PBDEs using GC/MS in electron capture negative 

ionization (ECNI) mode and electron ionization (EI) mode. Further study details are 

provided in Sections 3.1.2 and 3.2.1. In samples collected in 2002, decaBDE was 

detected in both sediments and tissues of sunfish (Lepomis gibbosus) and crayfish 

(Cambarus puncticambarus sp. c) collected immediately downstream of the WWTP 

outfall. The reported 2002 concentrations from this location in sediments, sunfish and 

crayfish were 1 630 000 μg/kg organic carbon (OC), 2880 μg/kg lipid and 21 600 μg/kg 

lipid, respectively. The much higher concentration in crayfish was attributed to the 

sediment-association of this species and the authors speculated that crayfish could form a 

link for the transfer of decaBDE from sediments to pelagic organisms. 

Based on the La Guardia et al. (2007) results, it is possible to estimate sediment BSAFs 

of 0.0018 for sunfish and 0.013 for crayfish. These are well below values suggested to 

potentially indicate biomagnification (i.e., ~1.7 to 3; refer to Section 1.2). It is possible 

that a combination of low sediment bioavailability and/or metabolic transformation could 

be limiting the bioaccumulation and biomagnification of decaBDE in this system. 

Wang et al. (2007) examined water, sediment and aquatic species collected from a small 

lake in Beijing, China, which receives effluent discharged from a large WWTP. Samples 

were homogenized, extracted and analyzed using HRGC/HRMS using EI ion source. The 

researchers found that average accumulations of 12 PBDEs (total, tri- to heptaBDEs) and 
30

BDE209 were 6.33 and 237.01 μg/kg dw in sediments. BDE209 concentrations in lake 

water and effluent were below the analytical detection limit (not given for water; 1 μg/kg 

(ww or dw unknown) for sediment and biota). High concentrations of BDE209 were 

determined for lichen (1572 μg/kg dw), march brown (Limnodrilus hoffmeisteri; 

11.37 μg/kg dw), coccid (114 μg/kg dw) and the zooplankton Monia rectirostris, Monia 

micrur and Monia macrocopa (151.9 μg/kg dw). Average concentrations in common carp 

(Cyprinus carpio), Java tilapia (Tilapia nilotica), leather catfish (Silurus meridionalis), 

crusian carp (Carassius auratus) and Chinese softshell turtle (Chinemys reevesii) were 

much lower, ranging from below detection to 19.32 μg/kg dw. 

Bioconcentration/bioaccumulation for BDE209 was not identified. In addition, the 

authors found no obvious biomagnification of PBDEs when they analyzed the 

relationship between PBDE concentrations and organism trophic level. 

Xiang et al. (2007) sampled biota and sediment samples for PBDEs, including BDE209, 

from the Pearl River Estuary of China. In sediments they found that BDE209 was the 

dominant congener, ranging from 792 to 4137 ng/g OC in sediment samples (median 

1372 ng/g OC). With respect to biota, they found non-detectable and measurable 

BDE209 concentrations in all biota species. Concentrations ranged up to 532.3 ng/g lipid 

in large yellow croaker (Pseudosiaena crocea; n=13, median=117.4 ng/g lipid), 

623.5 ng/g lipid in silvery pomfret (Platycephalus argenteus; n=10, median=24.4 ng/g 

lipid), 38.4 ng/g lipid in flathead fish (Platycephalus indicus; n=17, median=0.0 ng/g 

lipid), 373.4 ng/g lipid in robust tongue fish (Cynoglossus robustus; n=8, 

median=0.0 ng/g lipid), 150.4 ng/g lipid in Bombay duck (Harpodon nehereus; n=9, 

median=0.0 ng/g lipid), 555.5 ng/g lipid in jinga shrimp (Metapenaeus affinis; n=10, 

median=0.0 ng/g lipid), 405.3 ng/g lipid in greasy-back shrimp (Metapenaeus crocea; 

n=10, median=30.3 ng/g lipid), and 88.5 ng/g lipid in mantis shrimp (Oratosquilla 

oratoria; n=9, median=42.47 ng/g lipid). The study notes that the high BDE209 

concentrations in biota apparently resulted from elevated concentrations of BDE209 in 

local sediments. However, sediment BSAFs were calculated to range from 0 to 0.04 for 

BDE209 and trophic magnification was not deemed to be occurring based on the data 

shown in this study. 

Eljarrat et al. (2007) reported the results of fish (n=29), sediment (n=6) and effluent (n=3) 

sampling conducted in November 2005 from the River Vero in Spain. They found high 

BDE209 concentrations in sediments (up to 12 459 ng/g dw) and fish—barbel (Barbus 

graellsii) and carp (Cyprinus carpio), from non-detectable to 707 ng/g lipid—

downstream of an industrial park containing industries producing textiles and epoxy 

resins and involved with polymide polymerization. Using concentration measured in 

sediments and fish, the authors calculated sediment BSAFs for BDE209 of 0.0011 to 

0.0013, thus suggesting that bioaccumulation was not occurring based on these data. 

DeBruyn et al. (2009) studied marine horse mussels (Modiolus modiolus) and sediment 

collected off the coast of Vancouver Island, British Columbia, Canada near the city of 

Victoria to evaluate and compare patterns of PBDE and PCB bioaccumulation. Samples 

were collected from 14 stations within 800 metres of a municipal outfall, and from three 

reference locations. At each station, three surface sediment samples (0–2 cm) and 15 
31

randomly selected mussels (> 50 mm length) were collected. Mussels were measured for 

shell length and width, total weight, tissue weight, age and sex, prior to compositing and 

homogenizing tissue samples for chemical analysis by station. PBDEs were analyzed 

using HRGC-HRMS. Methods for PBDE determinations and QA/QC procedures were as 

per the U.S. Environmental Protection Agency (U.S. EPA) Methods 1 668A (21) and 

1 614 (22), with some modifications. Sediment BSAFs were calculated (lipid and organic 

carbon normalized) for BDE209. 

 

Sediment BDE209 concentrations were measured at all stations, ranging from 232 pg/g 

dw to 2550 pg/g dw. BDE209 was the predominant congener in sediment near the 

municipal wastewater. Although sediment BDE209 concentrations were highest near the 

wastewater outfall (2550 pg/g dw), the congener exhibited a gradient of decreasing 

predominance with proximity to the outfall, from approximately 80% of total PBDEs at 

reference locations to approximately 40% at the outfall station and in wastewater. 

BDE209 concentrations in mussel tissue ranged from “not detected” to 5305 pg/g dw 

(wastewater outfall station), and were above the detection limit at 7 of 17 stations. BSAFs 

for BDE209 were calculated using sediment and mussel tissue concentrations from the 

sampling stations. At 10 sites, BSAFs could not be calculated, as mussel tissue BDE209 

concentrations were below the limits of quantification. For the remaining sites, sediment 

BSAFs were calculated as: 1.48 (outfall), 0.52 (200 m), 1.59 (400 m), 0.97 (400 m), 3.53 

(reference), 0.94 (reference), and 1.18 (reference). The results generally suggest that 

BDE209 was not accumulating appreciably in the mussels at any site except for the 

reference location. 

Riva et al. (2007) studied the effect of BDE209 on freshwater bivalve zebra mussels 

(Dreissena polymorpha) under laboratory conditions. The primary objective of the study 

was to investigate the potential genotoxicity of BDE209. For this study, several hundred 

mussels were sampled at a depth of 4–5 m from the Italian subalpine great lakes. Mussels 

(still attached to rocks) were transferred to the laboratory, where they were maintained in 

glass tanks at a 12-hour light/dark photoperiod, with constant temperature (20°C) and 

oxygen (> 90% saturation), and fed a suspension of Pseudokirchneriella subcapitata. 

Approximately 150 acclimated mussels per BDE209 concentration were selected for 

exposure tests. DecaBDE (98% purity) was dissolved in isooctane/toluene (9:1 v:v) 

mixture, dissolved in dimethylsulfoxide (DMSO), and then added to water for final 

nominal concentrations of 0.1, 2 and 10 ug/L. Exposure assay aquaria were screened 

against direct sunlight to avoid photodegradation. Water was changed daily and mussels 

were fed 2 hours prior to water renewal. Control groups of mussels receiving either 

freshwater or a solvent blank (DMSO) were also monitored. 

The mussels reached a relatively constant concentration of decaBDE in their tissues after 

48 hours of exposure, for each treatment concentration. The United Kingdom (2008) 

interpreted the graphical representations of tissue concentrations and suggested a BCF for 

decaBDE in mussels on the order of 1000 l/kg or above. Riva et al. (2007) also suggested 

that there was also evidence for the presence of lower brominated congeners in the 

mussels after 168 hours exposure. These were not determined quantitatively but they 
32

were interpreted to have consistent GC analysis peaks with three heptaBDEs, three 

octaBDEs and three nonaBDEs. 

Nyholm et al. (2008) used zebrafish (Danio rerio) to study the extent of transfer of 11 

structurally diverse brominated flame retardants (BFRs) from females to their eggs. The 

tested BFRS included the PBDEs BDE28, BDE183 and BDE209. The tested BDE209 

compound was synthesized at the Department of Environmental Chemistry at Stockholm 

University, and labelled standards (e.g., 13C) BDE209 were purchased (Cambridge 

Isotope laboratories—purity of PBDEs not provided). The adult fish were exposed to 

BFRs via their feed. The mixture of 11 BFRs in ethanol was added to freeze-dried 

chironomids, giving nominal concentrations of 1 and 100 nmol/g of each molecular 

species on a dry weight basis; ethanol was allowed to evaporate. Twenty-three males and 

23 females were used for each dose level. Zebrafish were fed daily at ~2% of their body 

weight and sampled after 0, 3, 7, 14, 28, 35 and 42 days, 24 hours post-feeding. Two fish 

from each sampling period were pooled for analysis. Eggs were collected directly after 

spawning on days 0, 2–3, 6–7, 13–14, 27–28, 34–36, and 41–42. Samples were extracted 

and analyzed by GC/MS. 

Average lipid contents of the fish and eggs were 3.36% and 0.47%, respectively. Female 

zebrafish exposed to high-dose feed (100 nmol/g) had approximately an order of 

magnitude greater BDE209 concentrations (3.8 to 9.6 nmol/g lipid) than those exposed to 

low-dose feed (0.17 to 0.97 nmol/g lipid). Maximum BDE209 concentrations in 

high-dose female zebrafish were measured after 28 d of exposure (9.6 nmol/g lipid), 

while maximum BDE209 in the low-dose fish was measured after day 42 (0.97 nmol/g 

lipid). 

Egg concentrations from fish exposed to high-dose feed (100 nmol/g) were also 

approximately an order of magnitude greater in BDE209 concentrations (3.4 to 11 nmol/g 

lipid) in comparison to those of fish exposed to low-dose feed (0.46 to 2.2 nmol/g lipid). 

Maximum BDE209 concentrations in high-dose eggs were measured after 28 d of 

exposure (11 nmol/g lipid), while maximum BDE209 in the low-dose fish was measured 

after day 14 (2.2 nmol/g lipid). Eggs to fish ratios were calculated by dividing the 

BDE209 concentration at each point in time by the concentration measured in the fish. 

Egg/fish ratios were significantly (p < 0.05) higher than 1 for BDE209 in both exposure 

groups, with a higher ratio (i.e., higher BDE209 transfer) for the low-dose group. The 

authors suggested that the high egg/fish ratio may be influenced by the BFRs binding to 

the lipoproteins needed for egg production, or that the metabolic systems in eggs have a 

lower capacity for transformation than those in fish. 

Marine Mammals and Terrestrial Species 

Huwe and Smith (2007a, 2007b) examined the dietary accumulation, debromination and 

elimination of decaBDE in rats. Sprague-Dawley rats (n=26) were dosed with a 

commercial DecaBDE formulation (DE-83R 98.5% purity; 0.3 μg/g of diet) for a 21-d 

exposure period, which was followed by a 21-d elimination period. Following the 21-d 

exposure period, rats were sacrificed in groups of three for tissue analysis (liver, 

gastrointestinal tract, plasma, and remaining carcass) on days 0, 3, 7, 10, 14 and 21 of the 
33

elimination phase. Control rats were also sacrificed on these days to determine 

background PDBE concentrations (n=3 on day 0 and n=1 on all other days), and 

background values were subtracted from the PBDE determinations in the exposure group. 

Feces were collected from dosed rats daily during the exposure phase and pooled for 

analysis. Rat feed, feces and tissues were analyzed for a suite of PBDEs, including heptato decaBDE congeners. 

Based on the analytical results, it was estimated that only 5% (or 3.6 μg) of the total 

decaBDE dose was retained in rat tissues following the 21-d dosing period, while 

approximately 50% was excreted to feces during this time. In addition to decaBDE, the 

authors concluded that one nonaBDE congener (BDE207) and two octaBDE congeners 

(BDE201 and BDE197) were derived from the uptake of decaBDE. However, the total 

burden of BDE207, -201 and -197 accounted for only 3% of the total decaBDE dose, and 

45% of the dosed decaBDE was unaccounted for in rat tissues and feces. The authors 

speculated that the formation of bound and/or hydroxylated metabolites which were not 

included in their analysis was a likely explanation for the incomplete mass balance of 

decaBDE. 

Based on the observed carcass concentration of decaBDE a “BCF” (analogous to a 

BMF4

) of 0.05 (on a ww basis, justified by the fact that the percent of lipid in food and 

carcass were similar) was calculated. It is not known whether decaBDE concentrations 

achieved steady state in rat tissues during the 21-d exposure period and, as a result, it is 

uncertain how well the reported BMF represents the potential steady-state value. It is also 

likely that the BMF based on a full accounting of decaBDE plus neutral, bound and or 

hydroxylated metabolites could be higher, but the concentrations of bound and 

hydroxylated metabolites were not reported. The study reported the half-life of decaBDE 

in rat tissue based on the observed elimination of decaBDE during the 21-d elimination 

period. First-order half-lives for decaBDE ranged from 3.9 d in plasma to 8.6 d in 

carcass. For the liver and plasma, second-order decay equations were found to represent 

the data well, with relatively rapid distribution phase half-lives of 0.7 and 1.2 d, 

respectively, but longer elimination phase half-lives of 20.2 and 75.9 d, suggesting 

potential persistence of decaBDE in rat tissues following chronic dosing. However, the 

authors cautioned that there was a high level of uncertainty in the second-order estimates. 

Huwe et al. (2008) conducted a study of 29 male Sprague-Dawley rats to determine and 

compare the adsorption, distribution and excretion of PBDEs administered for 21 d as 

either a dust reference material or as a corn oil solution. The dust reference material 

(NIST Standard Reference Material 2585), containing a characterized and homogeneous 

composition of PBDEs, was mixed with rat chow. The corn oil solution contained 

commercial PBDE products DE-71, DE-79 and DE-83 (purity not given). These products 

were first dissolved in toluene and then mixed into a corn oil mixture. Daily doses were 

administered to the rats of either 1 or 6 μg/kg body weight (bw) of the dust/food or corn 

oil mixtures. The rats were randomly divided into five groups: controls (4 rats), low oil (4 

rats), high oil (13 rats), low dust (4 rats) and high dust (4 rats). Most rats were dosed over 

 

4

 The authors used the term “BCF”; however, given that this value was based on a comparison between rat 

tissue and food concentrations, it is actually analogous to a BMF. 
34

a period of 21 d, and then were killed 24 hours after their last feeding. To assess whether 

rats were approaching a steady-state body burden, three groups of rats (n=3) from the 

high-oil-dose group were killed on days 3, 7 and 14 after dosing began. Sampling was 

conducted of the feed and oil mixtures, feces and tissues (epididymal fat, liver, kidney, 

brain, gastrointestinal tract and remaining carcass) for 15 PBDEs (BDE28/38, -47, -85, 

-99, -100, -138, -153, -154, -183, -186, -197, -203, -206, -207 and -209). 

To determine whether steady-state body burdens were reached, only epididymal fat was 

sampled and analyzed. This analysis showed that all major tri- to octaBDEs had reached 

or were approaching steady state after 14 d, with no statistically significant differences 

between PBDE concentrations on days 14 and 21. 

Retention of PBDEs in the body of rats was congener dependent and ranged from 4.0 to 

4.8% of the dose for BDE209, and 10.1 to 22.6% for nonaBDEs, to approximately 69 to 

78% for BDE47, -100, and -153, but did not generally differ between the dust and oil 

treatment groups. The study did not consistently detect nona- and decaBDEs in the 

adipose tissues above that of the controls. Urine contained less than 0.3% of any 

congener. Fecal excretion was the major route of elimination and was described in the 

study as that component of the dose not adsorbed. Fecal excretion was found to reach a 

steady state by day 2, with no statistically significant differences between mean 

concentrations in feces on days 2, 11 or 20. Excretion of BDE209 was approximately 

68%, and ranged from 55.5 to 91.7% for the nonaBDEs. The amount of the BDE209 dose 

not adsorbed or excreted as parent compound ranged from 28% to 31.9%. Metabolic 

transformation products could have accounted for some portion of these percentages. 

Derived BCFs (analogous to BMFs as defined in this study) for adipose tissues were 

inversely related to the degree of bromination, and ranged from 7 to 24 for tri- to 

hexaBDE, 1 to 6 for hepta- to nonaBDEs and < 1 for decaBDE. For the liver tissues, 

BCFs for all PBDEs were below 1 except for BDE206 (BCF=2.4), and -207 (BCF=1.09). 

Hepatic Cyp2B1 and 2B2 mRNA expression increased in rats receiving the higher PBDE 

doses, suggesting potential effects of metabolic activity. The use of PBDE mixtures in 

this study made it impossible to conclusively determine whether metabolic debromination 

had occurred, although some of the findings are suggestive that metabolic transformation 

may have taken place (e.g., higher BMFs for some congeners like BDE206, and the 

detection of hydroxylated tetra- to hexaBDEs in feces). 

Kierkegaard et al. (2007) reported the findings of a 3-month feeding study with dairy 

cows. The study was originally undertaken to measure the long-term mass balance of 

PCBs; however, archived samples from two cows and feed were subsequently analyzed 

for a range of PBDEs. Over the 13-week study period, the cows were kept indoors and 

milk and feces were sampled once per week. The feed consisted of silage, concentrate 

and a mineral supplement which was not deliberately spiked with PBDEs. As a result, the 

PBDE concentration in food generally represented “background” contamination levels. 

The milk and feces samples were pooled according to the following scheme to obtain a 

series of 5 composite samples representing discrete portions of the 13-week study period: 

three 3-week composites and two 2-week composites. In addition, one of the cows was 
35

slaughtered at the end of the 13-week study, and samples of adipose tissue from 6 fat 

compartments as well as tissues from liver, kidney, heart and leg muscle were collected 

for analysis. Silage samples were retained for analysis at three intervals during the 

13-week study, and concentrate and mineral samples were analyzed once during the 

study. All samples/composites of feed, milk, feces and tissue were analyzed for hepta- to 

nonaBDEs using HRMS and were analyzed for decaBDE using LRMS in negative 

chemical ionization mode (limit of quantification = 0.4 to 150 pg/g lipid or dw). 

DecaBDE (i.e., BDE209) was the dominant congener in all matrices except milk, 

suggesting that milk levels were influenced more by the existing burden of PBDEs in 

storage tissue rather than uptake from food. In addition, PBDE levels were higher in the 

adipose storage tissues than in organ tissues. Based on these observations, the authors 

proposed that the cows were in a state of PBDE elimination rather than accumulation and 

that the observed PBDE concentrations may have been influenced by exposure to PBDEs 

prior to initiation of the 13-week experiment. Although a mass-balance analysis of input 

and output fluxes of PBDEs was attempted, it was largely unsuccessful due to a large 

increase in octa-, nona- and decaBDE congeners in the second silage sample relative to 

the first and third samples. It was unclear whether this second sample accurately 

represented the cows’ exposure since the increase of PBDEs in feces did not appear to 

coincide with the greater increase in decaBDE concentration in feed. PBDE 

concentrations in concentrate and mineral supplements were much lower than in the 

silage and did not likely affect the overall mass balance conducted on the PBDEs in this 

study. 

The Government of the United Kingdom (United Kingdom 2007a) conducted a critical 

analysis of the Kierkegaard et al. (2007) study and calculated dietary accumulation 

factors for a cow from silage based on either adipose tissue or whole body and the 

average silage concentration or end silage concentration. They concluded that the high 

variation in silage PBDE concentration led to considerable uncertainty in the study results 

with respect to bioaccumulation. To estimate accumulation factors and BMFs on a 

lipid-normalized basis, they assumed a 4% lipid content in silage. Table 2-5 summarizes 

the calculated accumulation factors and BMFs for nona- and decaBDE congeners from 

the Kierkegaard et al. (2007) study. 

The calculated BMFs and accumulation factors were well below 1 for decaBDE and only 

exceeded 1 for BDE207. For BDE207, it is uncertain whether the estimated values 

represent direct accumulation of BDE207 from food, or accumulation combined with 

bioformation as a result of the debromination of decaBDE. The United Kingdom (2007a) 

study proposed that, for chemicals such as decaBDE which undergo transformation once 

accumulated in organisms, BMF estimates should be based on the total burden of parent 

chemical and the metabolites resulting from accumulation and transformation of the 

parent chemical. In the case of the Kierkegaard et al. (2007) study, it is difficult to do so 

since the lower brominated PBDEs were also present in the feed and their presence in 

tissue could thus have resulted from both accumulation and bioformation. Furthermore, 

based on studies with rats, it is likely that alternative transformation pathways are also 

present in mammals which result in the presence of polar metabolites, bound residues and 
36

water-soluble residues. For a full accounting of the total chemical burden related to the 

accumulation of decaBDE, these would also have to be quantified. 

Table 2-5: Summary of accumulation factor calculations by United Kingdom (2007a) using 

the findings of Kierkegaard et al. (2007) for one cow 

Parameter Congener 

BDE206 BDE207 BDE208 BDE209 

Concentration Data 

Mean concentration in silage 

(ng/kg lipid)1 

4150 2583 1626 98 750 

Silage concentration over 

last feeding period (ng/kg 

lipid)1 

625 450 220 12 000 

Mean adipose tissue 

concentration (ng/kg lipid) 

552 1867 155 3700 

Mean concentration in 

organs/muscle (ng/kg lipid) 

239 740 49 2378 

Estimated mean whole body 

concentration in cow (ng/kg 

lipid)

286 909 65 2576 

Derived Accumulation Factors for Adipose Tissue 

Ratio of mean adipose 

concentration (ng/kg lipid) to 

mean silage concentration 

(ng/kg lipid) 

0.13 0.72 0.095 0.037 

Ratio of mean adipose 

concentration (ng/kg lipid) to 

silage concentration over 

last feeding period (ng/kg 

lipid) 

0.88 4.1 0.70 0.31 

Derived Accumulation Factors for Whole Body 

BMF based on estimated 

whole body concentration 

(ng/kg lipid) to mean silage 

concentration (ng/kg lipid) 

0.069 0.35 0.040 0.026 

BMF based on estimated 

whole body concentration 

(ng/kg lipid) to silage 

concentration over last time 

period (ng/kg lipid) 

0.46 2.0 0.30 0.21 

1

 Assumed lipid content of 4%. 

Thomas et al. (2005) examined the absorption of decaBDE from diet by three captive 

juvenile grey seals (Halichoerus grypus). The captive seals were fed a constant diet of 

herring for 3 months prior to the initiation of the 3-month study (6 months total time). 

During the 3-month study, feeding with herring continued (1–2.5 kg/d), with all fish 

obtained from a single batch caught in the North Sea. The second month of the study 
37

involved a decaBDE exposure phase with the diet supplemented with 12 μg decaBDE per 

day, dissolved in a cod liver oil capsule. For the final month of the study, the diet was 

supplemented with the cod liver oil capsule only to measure elimination of accumulated 

decaBDE. Fish, blood and feces samples were collected and analyzed for PBDEs on a 

weekly basis throughout the 3-month period, while blubber biopsies were taken and 

analyzed for PBDEs 3 times during the study (beginning, 3 d after initiation of exposure 

phase, and end, after 29 d on a decaBDE-free diet). 

The blood decaBDE concentration increased from non-detectable at the start of the 

exposure phase (day 28) to a maximum of approximately 1000 ng/g lipid (value read 

from graph) between 5 and 11 d after the end of the exposure phase. Concentrations of 

decaBDE in blubber ranged from non-detectable to 3.9 ng/g lipid on day 30 (3 d into the 

exposure phase) and from 3.4 to 7.4 ng/g lipid on day 83 (after 29 d on a decaBDE-free 

diet). The percentage of total ingested decaBDE estimated to be in blubber on day 30 

ranged from 36 to 68% (for one of the three seals, decaBDE was not detected in blubber) 

and on day 83, from 11 to 15%. 

During the elimination phase, the decaBDE half-lives in blood were estimated to be 

between 8.5 and 13 d, most likely due to a combination of metabolic 

transformation/elimination and transfer to blubber. The authors suggested that once 

stored in blubber, decaBDE was unlikely to be metabolized. 

The calculation results indicate a relatively high (i.e., exceeding 1) BMF for blood and 

relatively low BMF for blubber. The relatively high blood-based BMF suggests that 

significant magnification could be occurring from the diet to blood and related tissues, 

indicating relatively high bioaccumulation potential. The blood BMFs reported by 

Thomas et al. (2005) may be somewhat uncertain because of other non-lipid constituents 

(proteins) that can be attributed to the overall sorptive capacity of blood, thus making 

blood lipid normalization less accurate. The lower BMFs for blubber could be explained 

by the large storage capacity of fat tissues for lipophilic chemicals—it is possible that 

only a small fraction of the potential steady-state concentration in blubber was reached 

during the 26-d exposure phase. 

Based on the mass balance of measured input flux (in consumed diet) and output flux (in 

feces) of decaBDE, the authors determined an average absorption efficiency of 89% for 

decaBDE. The authors suggested that the high absorption efficiency called into question 

the theories regarding molecular size thresholds for chemical absorption. The relatively 

high apparent absorption efficiency was attributed to the following factors: 

• Large fat reserves (i.e., blubber) which provide a reservoir for lipophilic chemicals. 

• High food absorption efficiency in homeothermic carnivores such as seals, which 

creates strong a fugacity gradient in the gut, resulting in greater chemical uptake. 

• Possible incomplete collection of feces; however, it was estimated that over a 24-hour 

period, not more than 10% of feces was lost. 

• Degradation of decaBDE in the gut. The lower brominated PBDE congeners 

remained constant in feces despite the supplement with decaBDE during the second 

month, suggesting that degradation was not significant. 
38

Table 2-6: Estimated BMFs from the Thomas et al. (2005) feeding study with captive 

juvenile grey seals (Halichoerus grypus) 

Parameter Value Units 

Daily dose of decaBDE 12 μg/day 

Fish feeding rate 1–2.5 kg/day 

Exposure concentration 4.8–12 ng/g ww 

 48–120 ng/g lipid 

Maximum blood concentration (approximate) 1000 ng/g lipid 

Blood BMF 8.3–20.8 g/g lipid 

Median blubber concentration – Day 30 3 ng/g lipid 

Median blubber concentration – Day 83 5.3 ng/g lipid 

Blubber BMF – Day 30 0.025–0.063 g/g lipid 

Blubber BMF – Day 83 0.044–0.11 g/g lipid 

An additional explanation for the high apparent absorption efficiency which was not 

discussed by the authors is the potential formation of bound metabolites and/or the 

formation and subsequent excretion of phenolic and/or methoxylated metabolites. Recent 

studies with Sprague-Dawley rats (Mörck et al. 2003; Huwe and Smith 2007a, 2007b; 

Riu et al. 2008) have indicated that the formation of phenolic and methoxylated 

metabolites may be an important transformation pathway for accumulated decaBDE. 

Thus, if phenolic or methoxylated metabolites were present but not analyzed in the feces, 

it is possible that the output flux was underestimated, resulting in an overestimate of the 

absorption efficiency of decaBDE. Another potential uncertainty relates to an 

inconsistency in identified dosage rate. While the high proportion of total ingested 

decaBDE estimated to be present in blubber on day 30 and 83 (i.e., up to 68%) indicates 

significant decaBDE adsorption, there is some uncertainty in the overall findings of the 

study. 

By using the reported decaBDE concentrations in blood and blubber in conjunction with 

an estimated decaBDE concentration in food, it is possible to estimate blood-based and 

blubber-based BMFs for the Thomas et al. (2005) exposure study. Unfortunately, the 

authors did not report the lipid content of the herring consumed by the seals; however, 

based on values reported in the literature (e.g., Iverson et al. 2002; Jensen et al. 2007) a 

value of 10% lipid was used as a reasonable approximation for the BMF estimates 

summarized in Table 2-6. Note that the blood concentrations do not appear to have 

reached steady state during the exposure phase and it is likely that this also applies to the 

blubber concentrations. Furthermore, the blubber concentrations at day 30 and day 83 are 

unlikely to represent maximum accumulation of decaBDE (maximum accumulation 

would be expected at or after the end of the exposure phase (at day 54). Thus, the 

blood-based and blubber-based BMFs underestimate the steady-state BMF to an 

unknown degree. 
39

Sellström et al. (2005) analyzed decaBDE in soil and earthworm (species not identified), 

samples collected in 2000 from three research stations (with reference plots and sewage 

sludge amended plots) and two farms (reference and amended/flooded soils) in Sweden. 

Soil decaBDE concentrations at the various sites ranged from 0.015 to 22 000 ng/g dw, 

and organic carbon content (based on loss-on-ignition data) ranged from 2.12 to 7.22%. 

Concentrations of decaBDE in worms ranged from 0.99 to 52 000 ng/g lipid. Using these 

data, Sellström et al. (2005) calculated site-specific soil BSAFs for co-occurring worm 

and soil samples. The estimated BSAFs ranged from 0.04 to 0.7 and averaged 0.3. Based 

on these results, the authors concluded that decaBDE was bioavailable in soils. However, 

in this study, decaBDE was not shown to be bioaccumulating in earthworms. The authors 

did not observe any evidence of photolytic debromination in soils. 

The soil BSAFs determined by Sellström et al. (2005) were all below the range that might 

provide evidence of decaBDE biomagnification (i.e., 1.7 to 3; refer to Section 1.2). 

As part of their study of grizzly bears from British Columbia, Christensen et al. (2005) 

conducted a bioaccumulation analysis for decaBDE. Their method involved the 

estimation of a “bioaccumulation slope” based on a comparison of decaBDE 

concentration with the proportion of meat in the diet. The rationale was that if trophic 

magnification was causing an increase in decaBDE concentrations in prey above that in 

consumed vegetation, then the concentration of decaBDE in bears would increase as the 

proportion of prey-derived meat in their diet increased. A positive bioaccumulation slope 

indicates trophic magnification while a negative bioaccumulation slope indicates that 

trophic dilution was taking place. The authors indicated that the bioaccumulation slope 

for decaBDE was negative (albeit not significantly different from 0) but did not provide 

the value. The lack of a significant positive bioaccumulation slope for decaBDE suggests 

that decaBDE was not undergoing trophic magnification in the studied grizzly bear food 

webs based on a diet of meat. The study did not examine the potential significance of 

decaBDE exposure via consumed vegetation or inhaled air. 

In their study of decaBDE in the marine food web of Svalbard, Norway, Sørmo et al. 

(2006) also attempted to estimate BMFs for predator and prey species. Unfortunately, the 

high number of non-detected concentrations precluded the estimation of BMFs for polar 

bear / ringed seal or ringed seal / polar cod combinations. The estimated BMFs for polar 

cod / ice amphipod based on the mean concentrations in each species (except for ice 

amphipod for which there was only one sample) were 0.1 (wet weight basis) and 0.03 

(lipid weight basis), indicating no evidence of biomagnification for this predator-prey 

combination. 

2.2.3 Model Predictions 

Published models exist for predicting bioaccumulation in aquatic food webs and 

biomagnification in terrestrial mammals. The BAF–QSAR model described by Arnot and 

Gobas (2003) has generic applicability to the Canadian environment, and a modified 

version of this model was applied during the Government of Canada’s categorization of 

its Domestic Substances List. This model predicts both BAFs and BCFs for three 

representative fish trophic levels (low, middle and upper) in a generic aquatic food web 
40

based on a standard set of conditions found in the Canadian environment. For the 

terrestrial environment, Gobas et al. (2003) describe a terrestrial biomagnification model 

for adult male wolves (Canis lupus) based on the work of Kelly and Gobas (2003). It 

predicts BMFs for wolves based on chemical log Koa (logarithm of octanol-air 

coefficient), chemical log Kow, and a set of parameters describing the lichen-caribou-wolf 

food web of Bathurst Inlet in the Canadian Arctic. Both of these models incorporate a 

metabolic rate constant as part of chemical elimination, allowing for metabolism 

correction of BAF, BCF and BMF predictions based on field or laboratory observations. 

For both models, however, the default assumption is zero metabolism. 

BAF and BCF predictions for fish were made for decaBDE using the BAF–QSAR model. 

Environment Canada’s review of log Kow values for decaBDE revealed that a log Kow of 

8.7 reported by Wania and Dugani (2003) is considered to represent the most reliable 

value available. For further rationale regarding the selection of log Kow values, refer to 

Appendix C. Two prediction scenarios were conducted: the first with no correction made 

for metabolic transformation, and the second corrected for metabolism based on the 

laboratory observations of Tomy et al. (2004). Tomy et al. determined a half-life of 26 d 

in juvenile lake trout fed a diet containing decaBDE, which is consistent with a total 

elimination rate (kT) of approximately 0.027/d. The Tomy et al. (2004) value was used to 

derive an in vivo-based metabolic rate constant (kM) according to the method of Arnot et 

al. (2008b). In this method, when kT is available, kM is derived according to the following 

equation: 

kM = kT - (k2 + kE + kG) 

where: 

kM = the metabolic rate constant (1/d); 

k2 = the elimination rate constant (parameterized using data from Arnot et al. 2008a); 

kE = fecal egestion rate constant (parameterized using data from Arnot et al. 2008a); and 

kG = growth rate constant (parameterized using data from Arnot et al. 2008a). 

The method of Arnot et al. (2008b) provides for the estimation of confidence factors 

(CFs) for the kM to account for error associated with the in vivo data (i.e., measurement 

variability, parameter estimation uncertainty and model error). A CF of ±3.0 was 

calculated for the available BMF data. 

Because metabolic potential can be related to body weight and temperature (e.g., Hu and 

Layton 2001; Nichols et al. 2006), the kM was further normalized to 15oC and then 

corrected for the body weight of the middle trophic level fish in the Arnot-Gobas model 

(184 g) (Arnot et al. 2008a). The middle trophic level fish was used to represent overall 

model output and is most representative of fish weight likely to be consumed by an avian 

or terrestrial piscivore. After normalization routines, the kM ranges from 0.02 to 0.17 with 

a median value of 0.06. 
41

The BAF and BCF predictions for the middle trophic level fish for decaBDE are 

summarized in Table 2-7. All predicted BCF values are below 5000, which is expected 

given that uptake and elimination via the gills (which BCF accounts for) is limited and 

only important for substances with a log Kow of approximately less than 4.5. 

Table 2-7: BAF and BCF predictions for decaBDE using the Arnot-Gobas kinetic model 

(v1.11) 

kM (metabolismcorrected; days) Log Kow used Arnot-Gobas 

BCF 

Arnot-Gobas 

BAF 

Half-life (days) 

1.93E-02 (2.5%) 8.7 251 161 618 99 

0.058 (average) 8.7 90 29 386 35 

0.17 (97.5%) 8.7 31 4056 12 

0 (no metabolism) 8.7 2570 2 630 268 795 

Note: Bolded values exceed the BAF/BCF criterion of 5000. 

The predicted BAF, when corrected for metabolic transformation, ranges from 4056 to 

161 618 depending on the rate of metabolism. The predicted BAF for an average kM, 

which can be said to represent the typical fish metabolic potential in the Canadian 

environment, was calculated to be 29 386. When a default of no metabolism is used in the 

model, the BAF is several orders of magnitude higher than the BAF calculated with 

average metabolic rate potential. These results demonstrate the influence of both 

chemical partitioning behaviour (i.e., log Kow) and metabolic transformation on the 

bioaccumulation potential of decaBDE. 

The metabolism-corrected BAFs probably provide the best estimate of the 

bioaccumulation potential of decaBDE since metabolic transformation of decaBDE has 

been demonstrated or inferred in most laboratory studies. It is important to note that these 

corrected BAFs could underestimate the total chemical bioaccumulation related to 

decaBDE since they are based on parent chemical only and do not account for the 

additional presence of metabolites in tissues. 

BMF predictions for wolves were made using a spreadsheet version of the Gobas et al. 

(2003) model. The model can be re-parameterized for dietary assimilation efficiency (ED) 

and kM, the default settings of which are 90% and 0%, respectively. In addition to the 

range of potential log Kow values described for the BAF–QSAR model in Appendix C, 

two log Koa estimates were available: 15.27 (Tittlemier et al. 2002) and 18.423 (predicted 

by the QSAR model, KOAWIN). However, the BMF predictions do not vary 

significantly for log Koa or log Kow values in these ranges, and systemic variation of log 

Koa and log Kow made little difference in the predicted output. Four prediction scenarios 

were conducted: the first with no correction for metabolic transformation, and the 

remaining three corrected for metabolism based on the laboratory observations of Huwe 

and Smith (2007a, 2007b). They determined a range of potential half-lives for decaBDE 

in rats based on a combination of first-order and second-order approximations to 

elimination data for carcass, liver and blood plasma. For carcass, a first-order half-life of 

8.6 d was found to best represent the elimination data. For blood plasma and liver, 

second-order models (i.e., with distribution and elimination phases) provided the better 
42

fit, with elimination phase half-lives (representing slower elimination from residual body 

stores) of 75.9 and 20.2 d, respectively. These half-lives were used to infer kM values 

using a similar method as with the BAF–QSAR model, resulting in kM ranging from 

0.0086 to 0.08/d. Because these rates are based on rodent exposures, they also must be 

scaled to the body weight of the wolf (e.g., Hu and Layton 2001). This results in kM

values of 0.004 to 0.03 assuming a rat weight of 0.25 kg and the wolf model body weight 

of 80 kg (Hu and Layton 2001; Arnot et al. 2008b). 

The terrestrial BMF predictions for decaBDE are summarized in Table 2-8. An ED of 

~56% was calculated as outlined in Kelly et al. (2004) using the ED model for humans as 

the best estimator for carnivores. In the absence of a correction for metabolism, the 

predicted BMF was 89 and, when corrected for metabolism BMF predictions, ranged 

from 0.5 to 3.5 depending on the kM used in the model (Table 2-8). These metabolismcorrected and assimilation-efficiency-corrected predictions are within the range of 

experimental BMF values reported for decaBDE for terrestrial and aquatic receptors. The 

metabolism-corrected BMFs probably provide the best estimate of the biomagnification 

potential of decaBDE since metabolic transformation of decaBDE has been demonstrated 

or inferred in most laboratory and captive feeding studies reviewed from the open 

literature. Thus, corrected BMFs suggest a lack of or low level of biomagnification 

largely as a result of metabolism of decaBDE. 

It is important to note that the corrected BMFs could underestimate the total chemical 

biomagnification related to decaBDE since they are based on parent chemical only and 

do not account for the additional presence of metabolites in tissues. Thus, if all 

metabolites were included in the BMF calculations, it is possible that all predicted BMFs 

might be higher. 

Table 2-8: Wolf BMF predictions for decaBDE made using the terrestrial biomagnification 

model of Gobas et al. (2003) 

kM (wolf BW 

normalized) (days) 

Log Kow

used 

Log Koa used Gobas BMF Half-life 

(days) 

0.004 

(based on half-life of 

79.5 d decaBDE in 

plasma)1

8.7 15.2 3.5 173 

0.03 

(based on half-life of 

8.6 d decaBDE in 

carcass)1

8.7 15.2 0.5 23 

0 (no metabolism) 8.7 15.2 89 4119 

Note 

1

Half-lives observed by Huwe and Smith (2007a, 2007b). 
43

2.3 Weight-of-Evidence Analysis 

2.3.1 Summary of Evidence 

This section provides a summary of evidence which is currently available regarding the 

bioaccumulation and biomagnification of decaBDE. It is intended to synthesize the 

existing state of the science regarding whether decaBDE is “bioaccumulative” or may 

biomagnify in food chains. Unequivocal evidence supporting the conclusion that 

decaBDE is bioaccumulative and meets the criteria for bioaccumulation under the 

Persistence and Bioaccumulation Regulations, or is biomagnifying in food chains, was 

not available. Additional considerations and interpretations on the capacity for decaBDE 

to bioaccumulate and/or biomagnify are also summarized in a third section. 

1. Equivocal5 evidence with respect to whether decaBDE has significant potential to 

bioaccumulate or biomagnify in the environment: 

• The findings of Tomy et al (2009) which found BMFs of 12.7 and 4.8 for the 

cod:calanus and cod:themisto feeding relationships, suggesting that biomagnification 

may be occurring at these lower trophic levels in a western Arctic marine food web. 

However, BMFs for other feeding relationships within the same food chain were 

below 1 for higher trophic-level feeding relationships and the calculated TMF was 

also below 1. The study also has some uncertainty since organisms were sampled at 

different times and in different locations. 

• DeBruyn et al. (2009) calculated decaBDE sediment BSAFs for seventeen sites off 

the coast of British Columbia, Canada. At 10 sites, BSAFs could not be calculated as 

mussel tissue BDE209 concentrations were below the limits of quantification. For the 

remaining sites, all calculated BSAFs were low (i.e., ≤ 1.48), except at one reference 

site where the sediment BSAFs was calculated as 3.53, which suggests that, at this 

location, decaBDE may have biomagnified in the mussels. 

• The toxicity study by Riva et al. (2007), which is based on the interpretation of the 

United Kingdom (2008), suggested accumulation of decaBDE in zebra mussel 

potentially up to or exceeding 1000 l/kg (lipid). However, the study was conducted as 

a genotoxicity study and many study attributes relevant for an evaluation of 

bioconcentration were lacking (e.g., measured concentrations in water, precise 

concentrations in organism). Further, exposure concentrations may have exceeded the 

water solubility limit of decaBDE. 

• A feeding study with cows by Kierkegaard et al. (2007) found that all cow-silage 

BMFs were less than 1. However, a full accounting of parent decaBDE plus 

metabolites was not conducted. There was significant uncertainty in the results of this 

study due to the high variability of decaBDE concentrations in the silage feed. 

• The findings of a Lake Winnipeg pelagic food web study by Law et al. (2006) led to 

an estimated TMF of 3.6 for a Lake Winnipeg food web, with many BMFs 

exceeding 1. These observations were based on parent decaBDE only. If both 

decaBDE plus metabolites were considered, it is possible that the TMF and BMFs 

 

5

 Studies cited here are considered by Environment Canada to provide uncertain evidence for decaBDE to 

bioaccumulate or biomagnify in the environment. 
44

might be higher. However, this study had a number of limitations that make the 

results less certain. 

• A feeding study by Stapleton et al. (2006) using juvenile rainbow trout found 

tissue-specific BMFs which ranged up to 1.74 for decaBDE and neutral metabolites. 

Other metabolites which were not measured may have also been present. However, 

BMFs based on serum and carcass data did not exceed 1. 

• Two studies by Burreau et al. (2004, 2006) on food webs of the Baltic Sea evaluated 

biomagnification in two Baltic Sea food webs and found that biomagnification of 

decaBDE did not appear to be occurring. Failure to detect decaBDE in salmon from 

the Atlantic Ocean precluded a similar analysis for this food web. In addition, the 

relatively high levels of decaBDE in procedural blanks and low concentrations of 

decaBDE in biota samples created uncertainty in the overall biomagnification 

analysis. 

• In a feeding study using captive seals by Thomas et al. (2005), blood-serum-based 

BMFs exceeded 1 but blubber BMFs were well below 1. Blubber concentrations are 

unlikely to have reached steady state. The steady-state BMFs for blubber could be 

higher. However, this study reported an unusually high calculated dietary assimilation 

efficiency of 89%, which would not be expected for decaBDE. 

• Although high concentrations of decaBDE have occasionally been observed in birds 

of prey (especially in China and Europe), red fox, shark, marine mammals and a few 

marine bird species, these high observed concentrations are confounded by the 

potential that the sampled species inhabit decaBDE hotspots close to industrialized 

areas of Europe and China. Thus, if environmental exposures were very high, 

organisms could still achieve very high concentrations in their tissues even though 

BAFs or BMFs were very low. 

• Relatively high concentrations of decaBDE were observed in peregrine falcons in 

Greenland. 

• Using the BAF–QSAR model, metabolism-, weight- and temperature-normalized 

aquatic BAF predictions range from approximately 4000 to 162 000 and reflect the 

uncertainty associated with estimating metabolic rates in fish. The average BAF 

corrected for metabolism was calculated to be approximately 30 000. If the BAF–

QSAR model also accounted for the body burden of metabolites, then the BAF 

estimates could potentially be higher than the range predicted for the parent 

compound.

• Using the Gobas et al. (2003) wolf model, metabolism-corrected terrestrial BMF 

predictions can range from slightly above to slightly below 1 depending on the rate of 

metabolism assumed in the wolf model. The moderately high dietary efficiency of 

56% contributes to a BMF greater than 1 when the lower metabolism rate constant is 

used. It should be noted that it is difficult to compare modelled BMF with any of the 

available measured BMFs because none of them are reported for the wolf or other 

terrestrial carnivore food web. Any comparisons involve uncertainties due to different 

species with different metabolic rates, assimilation efficiencies and body weights. 
45

2. Evidence that does not support6 the conclusion that decaBDE has significant potential 

to bioaccumulate or biomagnify in the environment: 

• A feeding study by Huwe and Smith (2007a, 2007b) using rats determined a BMF of 

0.05, while a study by Huwe et al. (2008) obtained a derived BMF of less than 1. A 

full accounting of parent decaBDE and metabolites could have increased the BMF 

estimates. 

• Calculated sediment BSAFs determined by La Guardia et al. (2007), Xiang et al. 

(2007) and Eljarrat et al. (2007) were well below thresholds indicative of 

biomagnification above equilibrium conditions. 

• A study by Sørmo et al. (2006) for a marine food web in Svalbard, Norway, 

identified only very low concentrations and a very low BMF (0.03) for the polar cod / 

ice amphipod predator-prey relationship. Other BMFs could not be calculated 

because of the high frequency of samples in which decaBDE was not detected, 

suggesting a low potential for biomagnification in general in this food web. 

• A study by Sellström et al. (2005) of the accumulation of decaBDE in earthworms 

from soil found that soil BSAFs were well below the threshold range (>1.7 to 3) that 

would indicate biomagnification above equilibrium conditions. 

• Relatively low and often non-detected concentrations of decaBDE were found for a 

wide range of middle to top predators including some marine fish, mammalian 

predators (grizzly bear, polar bear and lynx), marine mammals, marine/aquatic birds 

and birds of prey (multiple studies, often the same ones as those where relatively high 

concentrations were also observed). 

• There was minimal observed uptake of decaBDE by Lumbriculus variegatus exposed 

to decaBDE in biosolids and artificial spiked sediments (Ciparis and Hale 2005). 

• A study by Christensen et al. (2005) on grizzly bears suggested that biomagnification 

is not occurring. 

• A feeding study by Tomy et al. (2004) using juvenile lake trout determined a BMF of 

0.3, using the kinetic method based on parent decaBDE only. If metabolites were 

included in the BMF estimate, it may have been higher. 

3. Additional considerations and evidence: 

• A BCF study by MITI (1992) resulted in an estimated BCF < 3000. Note that this BCF 

value was based on the estimated water solubility of decaBDE and the detection limit 

in tissues, and therefore is highly uncertain. 

• There was low chemical uptake efficiency (i.e., at least 0.44%) and evidence of 

metabolism in carp (Stapleton et al. 2004). 

• There was low chemical uptake efficiency (i.e., from at least 0.02 to 3.2%) and 

evidence of metabolism in rainbow trout (Stapleton et al. 2006; Kierkegaard et al. 

1999). 

 

6

 Studies cited here indicate that decaBDE does not have the potential to bioaccumulate or biomagnify in 

the environment; these studies include both reliable and less certain evidence. 
46

• Metabolism studies using rats (Norris et al. 1973, 1974; El Dareer et al. 1987) 

suggested that decaBDE has low retention (generally less than 5%), low 

bioaccumulation potential in mammalian species, and rapid excretion. 

• Very low chemical assimilation efficiencies in fish suggest that assimilation 

efficiency in carnivorous mammals may be lower than predicted by the ED model for 

humans. However, the predicted ED value is based on a relationship using multiple 

halogenated organics and so is appropriate for modelling decaBDE. Fish do show a 

significantly lower ED than mammals at a log Kow of 8.7 (conservatively estimated at 

56% vs. ~0%) using the relationships summarized in Kelly et al. 2004). Thus, high 

assimilation of halogenated aromatics by mammals would contribute to a potential for 

higher BMF values in mammals. However, there is little evidence from terrestrial 

food web studies to substantiate this. BMF values could be offset by a high rate of gut 

metabolism to lower brominated and hydroxylated bio-transformation products, 

which would limit the ED vs. model predictions of ED. 

2.3.2 Conclusion Respecting Bioaccumulation 

The existing evidence for the bioaccumulation of decaBDE does not support a conclusion 

of “bioaccumulative” as defined in the current Persistence and Bioaccumulation 

Regulations under CEPA 1999. While most available data show that decaBDE has 

limited potential to bioaccumulate or biomagnify in the environment, some evidence 

suggests a higher BAF than previously considered for decaBDE, and some new data 

suggest possible biomagnification. The modelling undertaken to support this evaluation, 

however, shows uncertainty associated with metabolism in fish, as model-predicted 

aquatic BAFs range from below the 5000 criterion to well above 5000. Predicted 

terrestrial carnivore BMF values also range from below 1 to greater than 1 depending on 

the rate of metabolism assumed. Although less relevant than BAF or BMF, experimental 

BCF measures are below the 5000 criterion. The substance is shown to be increasing in 

concentrations in some wildlife species, and some data suggest that decaBDE has reached 

concentrations in some organisms interpreted to be high. DecaBDE (i.e., BDE209) is also 

considered to be contributing to the bioaccumulation potential of total PBDEs as a result 

of metabolism to lower brominated forms (to be discussed later in this report). 

3 Evidence of Transformation 

This section examines the transformation of decaBDE to determine whether there is 

concern respecting its transformation to bioaccumulative products in organisms (i.e., in 

vivo by metabolism) and in the environment (i.e., by abiotic and biodegradation 

processes). The evidence provided by studies quantifying transformation may be 

confounded by the following limitations: 

(a) Unknown or uncharacterized purity of the parent compound, and the presence 

(often unquantified) of lower brominated PBDEs in the materials used (e.g., 

studies typically used decaBDE with a purity of approximately ≥98%). There is a 

potential that lower brominated PBDE congeners may be present as impurities in 

the food used in in vivo transformation studies. Some of the lower congeners may 

have higher accumulation factors than decaBDE. Thus, even if these impurities 
47

were present in the food (or the decaBDE test substance itself) at concentrations 

below the detection limit, they could still accumulate to measurable 

concentrations in the organism during the course of the experiment. 

(b) Analytical challenges, which could influence the accuracy of the results (see 

Section 2). 

(c) Use of laboratory studies to make inferences respecting processes and rates of 

transformation in the environment. 

These issues are discussed in more detail later in this section. 

3.1 In Vivo Transformation 

3.1.1 Information Evaluated in the Screening Assessment for PBDEs 

The following summarizes the evidence regarding debromination of decaBDE which was 

available during and prior to 2004 and considered in the screening assessment for PBDEs 

(Environment Canada 2006b). 

In a decaBDE feeding study with juvenile carp, Stapleton and Baker (2003) and Stapleton 

et al. (2004) found that seven transformed products of decaBDE accumulated in the 

whole fish and liver tissues over the exposure period. These were identified as penta- to 

octaBDEs with only BDE154 and -155 (i.e., hexaBDEs) being positively identified 

(personal communications from HM Stapleton to Environment Canada, July 2009; 

unreferenced). The unidentified hepta- and octaBDE congeners have recently been 

determined to be BDE188, -202 and -197 (Stapleton et al. 2006). It also appeared that 

higher brominated PBDEs, such as octaBDE, may be further debrominating to lower 

brominated PBDE homologues. By comparing the structure of decaBDE with those of its 

identified degradation products, Stapleton and Baker (2003) suggested that this 

debromination may be due to the action of deiodinase enzymes which normally act on 

thyroid hormone. The bioavailability of decaBDE (parent chemical plus metabolites) was 

estimated at 0.44%. This value could be higher if other metabolites were present (not 

determined in these studies). It should be noted that the alternative explanation provided 

by Stapleton and Baker (2003) for the transformation of decaBDE, i.e., that it could have 

also been due to the action of intestinal microflora, has recently been refuted by Benedict 

et al. (2007), whose study confirmed that carp intestinal flora cannot debrominate 

PBDEs. 

In their feeding study with juvenile rainbow trout, Kierkegaard et al. (1999) detected 

several hexa-, hepta-, octa- and nonaBDEs in liver and muscle tissues of the treated fish 

and some of these congeners were not measured in the food. It was not possible to 

determine in this study whether their presence was a result of decaBDE metabolism or 

the efficient adsorption of congeners initially present in trace amounts in the food. 

Metabolism studies using rats (Norris et al. 1973, 1974; El Dareer et al. 1987) suggest 

that decaBDE has low bioaccumulation potential in mammalian species. For instance, 

Norris et al. (1973, 1974) dosed male and female rats with 1.0 mg of 14C-labelled 

decaBDE as a suspension in corn oil and found that approximately 90.6% of the 

administered 14C-labelled decaBDE was excreted in the feces within 24 h, and by 48 h, 
48

all of the administered chemical had been excreted. Tissue accumulation studies in which 

rats were fed diets of decaBDE at a rate of 0.1 mg/kg body weight (bw)/d showed that 

bromine contents in various tissues were not significantly greater than those of the 

controls. The bromine content of the adipose tissue of decaBDE-dosed rats was found to 

be significantly increased at the p<0.03 level but not at the p<0.01 level when compared 

with the controls (Norris et al. 1973, 1974). Further discussion (beyond that presented in 

Canada 2006 or Environment Canada 2006b) respecting the El Dareer et al. (1987) study 

is provided below (also cited as NTP 1986). 

The National Toxicology Program (NTP 1986) and El Dareer et al. (1987) evaluated the 

adsorption of 14C-labelled decaBDE (97.9 to 99.2% pure) using F344/N rats. DecaBDE 

(BDE209) was administered through diet (total food consumption was 50 000 mg/d, on 

days 1–7 and 9–11 amended with unradiolabelled decaBDE and on day 8 with 

radiolabelled decaBDE). The total dose of decaBDE ranged from 3718 to 3826 mg/kg/d. 

The results showed that after 72 hours of exposure, from 91.3 ± 4% to 101 ± 4% of the 

radioactivity was recovered in the excreted feces. In a further study, the researchers fed 

both unlabelled and radiolabelled decaBDE (~22–25 mg/kg bw) in food (total food 

consumption 48 000 mg/d). The rats were fed with unradiolabelled decaBDE in food on 

days 1–7 and 9, or 9–10, or 9–11, and radiolabelled decaBDE in food on day 8. They 

determined that 82.5 to 86.4% of the radiolabelled decaBDE was recovered from the 

feces, and total recovery ranged from 83.2 to 89.3%. Excretion in urine accounted for 

about 0.01% of the administered dose of decaBDE. They detected trace levels of 

radioactivity in most tissues, with the highest levels found in liver, kidney, lung, skin and 

adipose tissue. The authors continued with an intravenous study in which rats were 

injected with 1.07 mg/kg 14C-decaBDE and were sacrificed 72 hours after dosing. The 

results revealed that 74% of the dose was found in the feces and gut contents, suggesting 

that biliary excretion was important. The authors determined that the excreted material 

was mainly unchanged decaBDE (up to 40%) and three metabolites (unidentified) (up to 

49.8%). In a final study, biliary excretion over 4 hours was evaluated after intravenous 

administration of 0.9 mg/kg of 14C-decaBBDE to rats. The authors found that 7.17% of 

the dose appeared in the bile after 4 hours. The researchers indicated that one unidentified 

metabolite was observed in the bile. 

Mörck et al. (2003) conducted a feeding study to investigate possible distribution 

pathways and metabolites of decaBDE in rats. This study involved orally dosing rats with 

radiolabelled decaBDE. The decaBDE substance used in the study was synthesized by a 

multi-step process involving the bromination of 14C-labelled phenol ultimately giving 

14C-decaBDE (purity determined at > 98%). Before administration, the radiolabelled 

decaBDE was diluted in unlabelled decaBDE synthesized using the same process (also 

with a purity of > 98%). This study used the radioactivity associated with non-extractable 

residues, lipid-bound residues, water-soluble residues, phenolic metabolites and neutral 

metabolites/parent compound to make inferences about the uptake and metabolic 

transformation of decaBDE. A total of 8 “conventional” rats and 2 bile-duct-cannulated 

rats were dosed orally by gavage with 3 mmol/kg, 15 Ci/mol radiolabelled decaBDE in a 

dosing vehicle (lutrol/soya phospholipone) designed to maximize decaBDE solubility and 

absorption. Feces were collected for analysis from the conventional and cannulated rats at 
49

24-hour intervals, while bile was collected and analyzed from the cannulated rats at 4, 12, 

24, 48 and 72 hours post-dosing. The conventional rats were sacrificed at day 3 post-dose 

(n=4) and day 7 post-dose (n=3) for analysis of tissues. 

Radioactivity related to decaBDE was present in the feces, bile and tissues of exposed 

rats as a combination of non-extractable residues, lipid-bound residues, water-soluble 

residues, phenolic metabolites and neutral metabolites/parent compound, indicating the 

potential presence of parent decaBDE and debrominated metabolites, as well as 

conjugated metabolites and metabolites bound to lipid or other tissues. In conventional 

rats, phenolic metabolites increased from ~8% to ~14% in the three days post-dose while 

neutral metabolites/parent compound decreased from ~29% to ~17% in the three days 

post-dose. Bile contained predominantly lipid-bound residues (decreasing from ~60% to 

~11% in the three days post-dose) and water soluble residues (increasing from ~9% to 

~62% in the three days post-dose). 

The study found that approximately 90% of the single oral dose was excreted via feces. 

By totalling the amount of 14C collected in urine and feces and subtracting that from 

100%, the authors estimated that 9% remained in the body (i.e., as parent compound and 

metabolites). The highest concentrations of radioactivity (on a fresh weight basis) were 

found in adrenals, kidney, heart and liver after both 3 and 7 d. On a lipid weight basis, 

plasma and liver had the highest levels; levels in other tissues were low. The high 

absorption (relative to previous studies) of decaBDE was likely due to the use of the 

lutrol/soya phospholipone carrier. Ten percent of the dose excreted in feces came from 

bile. Of the total radioactivity excreted in feces, it was inferred that approximately 65% 

of the dose was eliminated as metabolites. In this study, metabolites were those 

characterized as non-extractable, water-soluble, lipid-bound, phenolic metabolites and 

parent compound/neutral metabolites. Thus, there is some uncertainty that the 65% 

included some proportion of bound parent compound or metabolites. 

To characterize decaBDE metabolites, the authors further manipulated and analyzed the 

phenolic and neutral metabolites / parent compound fractions. Detection was conducted 

using GC/MS (ECNI). Based on their findings, the authors inferred, although they did not 

positively determine, the presence of debrominated, phenolic and methoxylated 

metabolites. Non-transformed decaBDE dominated the neutral metabolites / parent 

compound fraction of both feces and tissues, but small amounts (< 0.5%) of three nonabrominated metabolites were also present. Analysis of the phenolic fraction of feces and 

tissues also inferred the presence of hydroxylated and hydroxyl-methoxylated (i.e., 

guiacol-type) substances with five to seven bromine atoms which the researchers inferred 

were penta- to heptaBDEs. 

From the metabolite characterization, the authors speculated that the potential sequence 

of events in metabolic transformation of decaBDE would involve debromination 

followed by hydroxylation of the exposed benzene ring to form an ortho catechol either 

directly via an arene oxide or by monohydroxylation followed by secondary oxidation of 

the resulting phenol. The resulting ortho catechol would then be methylated, potentially 

by the action of catechol-O-methyltransferase to form the observed guaiacol 
50

(hydroxymethoxylated) metabolites. The initial debromination step would have to 

remove least 2 bromines from the hydroxylated/methoxylated ring but appeared to also 

remove bromines from the other ring. 

The authors also speculated that cytochrome P450 enzymes might be responsible for the 

large proportion bound to residues present in the small intestine. The action of 

cytochrome P450 following debromination is consistent with the formation of phenols 

and catechols, as well as the formation of reactive intermediates that bind covalently with 

macromolecules, which would explain the high level of bound residues in the small 

intestine. 

3.1.2 New Evidence of Debromination in Vivo

Fish 

Stapleton et al. (2006) exposed 45 juvenile rainbow trout to spiked food containing 

decaBDE (Cambridge Isotopes Laboratories, > 98% pure) for a period of 5 months. The 

debrominated products accounted for approximately 73% of the total PBDE burden in the 

carcasses. These products were identified as primarily BDE208 (nonaBDE) and BDE202 

and -201 (octaBDEs) with a small fraction of BDE188 (heptaBDE). NonaBDEs 

(primarily BDE207 and -208) accounted for 26% of the burden in serum with only minor 

amounts of octaBDEs present, and untransformed decaBDE accounting for the remainder 

(approximately 68%). In the liver, the burden was primarily decaBDE with only a small 

fraction of lower brominated PBDEs (primarily nonaBDEs). The predominance of 

BDE202 as a product of decaBDE debromination was similar between rainbow trout 

(observed here) and carp from a previous study, and has been viewed as a potential 

marker for transformation as this congener has apparently not been measured in the 

commercial DecaBDE or OctaBDE products. The researchers also found that 

concentrations of the hepta- and octaBDE congeners increased in concentration over the 

duration of the exposure. Based on the total body burden of hepta- to decaBDE, uptake of 

decaBDE was estimated at 3.2% of total dose. Since no measurement was made of other 

metabolites, the researchers suggest that the estimate of uptake could be higher. 

Stapleton et al. (2006) also conducted an in vitro study using rainbow trout liver 

microsome fractions to confirm the metabolic capacity of rainbow trout tissues and 

potential metabolites. This was partnered with an in vitro study using carp liver 

microsome fractions that was also undertaken to confirm the metabolic capacity of carp 

demonstrated by Stapleton et al. (2004). In this study, liver microsomes were prepared 

from three fish and incubated for 1 and 24 h at 25°C. Analysis was conducted for PBDE 

congeners; however, recovery was low for the 1 h incubation (< 60%) in comparison to 

the 24 h incubation (> 80%). The study also had a control consisting of a boiled 

microsome fraction. Both carp and rainbow trout liver microsomes were observed to 

transform decaBDE even after the 1-h incubation. The carp liver microsomes 

debrominated decaBDE to hexa-, hepta-, octa- and nonaBDEs, although very little 

production of the nonaBDE congeners was observed (mirroring the in vivo experiment) in 

both 1- and 24-h incubations. After 24 h, the carp microsomes transformed 65% of the 

decaBDE mass (30% was transformed to hexaBDE congeners). The liver microsomes of 
51

rainbow trout transformed as much as 22% of decaBDE to octa- and nonaBDEs in the 

24-h incubation. The controls showed no transformation of decaBDE. 

The authors concluded that their results supported the hypothesis that deiodinase 

enzymes were catalyzing debromination of decaBDE; however, they also cautioned that 

it was not possible to rule out the concurrent or alternative action of P450 enzymes. 

Stapleton’s work shows that removal of bromine atoms occurs preferentially from the 

meta- or para-substituted positions (personal communication from HM Stapleton to 

Environment Canada, January 2008; unreferenced). 

Tomy et al. (2004) studied the uptake, by juvenile lake trout, of 12 tetra- to heptaBDEs 

plus decaBDE from spiked commercial fish food. Three lower brominated PBDE 

congeners (unknown penta- and hexaBDE, and BDE140) appeared to be bioformed in the 

exposed fish. The authors hypothesized that debromination of decaBDE was a potential 

explanation since they were not specifically dosed to the organisms. However, it should 

be noted that each of the PBDE congeners dosed in this study had purities of > 96%, and 

thus there is uncertainty regarding whether these congeners were in the dosed material as 

impurities. They suggested that the structural similarity of BDEs to thyroxine (T4) could 

mean that deiodinase enzymes were debrominating higher brominated PBDEs to lower 

brominated PBDEs. The authors concluded that the degree biotransformation, especially 

for decaBDE, was likely to vary considerably between species, leading to high potential 

interspecies variability in bioaccumulation. 

La Guardia et al. (2007) examined the potential for in vivo debromination of decaBDE in 

aquatic organisms inhabiting the receiving environment of a WWTP located in Roxboro, 

North Carolina, which, based on releases reported by industry to the U.S. EPA’s Toxics 

Release Inventory, was determined to receive wastewater from a large plastics 

manufacturing facility. The PBDE congener profile was tracked from the WWTP effluent 

to the receiving environment sediments and to biota in order to evaluate whether 

significant debromination was occurring. In 2002, samples of wastewater sludge, 

sediments and biota (sunfish, creek chub and crayfish) were collected. Then in fall 2005, 

samples of wastewater sludge, sediments and biota (sunfish only) were collected. Aquatic 

biota sampling involved the use of minnow traps at a location 15 m downstream of the 

WWTP outfall. All samples were extracted and purified using size-exclusion 

chromatography analyzed for PBDEs GC/MS in ECNI mode and EI mode. 

A total of 23 PBDE congeners were detected in the biota samples. Of these, decaBDE 

was only detected in 2002 samples of sunfish (2880 μg/kg lipid) and crayfish 

(21 600 μg/kg lipid). The much higher concentration in crayfish was attributed to the 

sediment-association of this species and the authors speculated that crayfish could form a 

link from sediments to pelagic organisms. The authors also speculated that the lack of 

detected decaBDE concentrations in chub could be due to an enhanced ability of this 

species to metabolize decaBDE. Chub are closely related to carp, which have previously 

been demonstrated to have an enhanced capability to debrominate decaBDE (Stapleton et 

al. 2004). Two octa- and three heptaBDE congeners which were not detected in sludge 

and sediments were found in chub tissues, suggesting bioformation of these homologues. 
52

Based on these findings, the authors concluded that decaBDE is bioavailable in natural 

environments and could undergo metabolic debromination in the field, resulting in 

bioformation of lower brominated PBDEs. 

Lebeuf et al. (2006) examined the effects of decaBDE and PCB126 on hepatic 

concentrations of PBDEs and methoxy-PBDEs in Atlantic tomcod. The decaBDE used in 

their experiment (DE-83R, Great Lakes Chemical Corp.) consisted of greater than 96% 

BDE209, with BDE153, -183 and -203 detected in amounts of between 0.00024 and 

0.034% on a mass basis. Further characterization revealed that as many as seven 

nonspecified PBDEs were qualitatively detected (including four heptaBDEs and three 

octaBDEs), in addition to the three nonaBDEs. The fish used in the study were captured 

in the St. Lawrence estuary in 2001 and were acclimated to laboratory conditions for 

about 6 months. The fish were fed twice a week with capelin until the beginning of April, 

and then with rainbow smelts until the end of the study. After the acclimation period, 200 

fish were randomly distributed in groups of 25 fish and placed into eight 500-L fibreglass 

tanks. On day 0, fish from half the tanks were anaesthetized in water. Half of these fish 

(i.e., composing 4 tanks) were injected interaperitoneally with a dose of PCB126, and the 

other half were injected with corn oil alone. The PCB126 was injected to evaluate the 

impact of cytochrome P4501A (CYP1A) induction on the biotransformation of injected 

PBDEs contained in the decaBDE commercial product administered to the fish. The fish 

from two of the four tanks that had received PCB126 and the fish from two of the four 

tanks that had received corn oil alone were injected with decaBDE (dose 400 ng/g fish; 

fish from the remaining tanks received a dose of corn oil alone) after 21 d. 

Fish were sampled and analyzed for decaBDE and potential transformation products after 

seven weeks following decaBDE administration. The study found BDE209, -208, -207, 

-206, -203 and three unidentified octaBDEs in the liver of the fish. All these congeners 

were essentially absent in the control fish. These were also measured in the decaBDE 

product administered in the experiment, and thus, their presence cannot be attributed 

exclusively to biotransformation. Also, the presence of these substances could have been 

due to thermal degradation of the decaBDE product during analysis, or due to 

transformation in the tank due to fecal egestion of decaBDE. Despite an increase in 

EROD activity in the liver of tomcod dosed with PCB126 and decaBDE compared to 

decaBDE alone, no further increases of PBDE hepatic concentrations were observed. 

However, depleted concentrations of BDE17 and 6-methoxy-BDE47 (both were 

measured in the control fish) were found in the fish injected with decaBDE compared to 

control fish. The researchers attributed this to activated hepatic metabolic enzymes other 

than CYP1A. Fish with the injected PCB126 showed an even more significant depletion 

of BDE17 than those with decaBDE treatment and significantly lower concentrations of 

BDE203. While the fish in this study only exhibited limited capacity to metabolize 

decaBDE, the study demonstrates the importance of methodologies to evaluate the 

potential sources of substances detected in organisms following administration of 

decaBDE products. 

The finding of Lebeuf et al. (2006) that CYP1A appears to play a limited role in the 

transformation of PBDEs is consistent with that of Valters et al. (2005), who found 
53

limited evidence of cytochrome P450 metabolism in common carp plasma, as 

hydroxylated PBDEs represented only 0.8% of the fraction of the total PBDE compound. 

Benedict at al. (2007) also concluded (at least based on their study using BDE99) that it 

is doubtful that CYPs are responsible for debromination of BDE99. In their evaluation of 

BDE99 debromination in carp microflora and microsomes, they concluded that an 

endogenous process that occurs with approximately equal activities in the intestine and 

liver microsomes appears responsible for debromination of BDE99 to BDE47. They also 

showed that this process can be inhibited by the presence of reverse thyronine. The exact 

metabolic mechanism could not be fully elucidated from their study, but they noted that 

their work suggests that debromination may involve thyroid hormone deiodinases. 

Nyholm et al. (2009) used zebrafish (Danio rerio) as a vertebrate model organism to 

study the uptake, metabolite formation and elimination of 11 structurally diverse BFRs, 

including BDE28, BDE183 and BDE209 (purity not reported). Solutions containing a 

mixture of BFRs were made from stock solutions of individual BFRs in iso-octane or 

tetrahydrofuran, and evaporated almost to dryness. The residues were taken up in ethanol. 

Resulting mixed BFR solutions were added to a chironomid feed, giving nominal 

concentrations of 1 and 100 nmol/g of each molecular species on a dry weight basis. A 

control feed was made with ethanol only. Zebrafish were fed with BFR-treated or control 

chironomid feed for a 42-d exposure period and then with untreated feed for a 14-d 

elimination period. Samples were extracted and analyzed by GC/MS. 

Zebrafish exposed to high-dose feed (100 nmol/g) for 42 d had detectable BDE209 

concentrations after the 2-week elimination period, but concentrations were tenfold lower 

than those of BDE 28 (which was determined to have the highest concentration among 

BFRs analyzed). Uptake efficiency was lowest for BDE209 (< 1%) compared to other 

BFRs studied, and elimination half life was estimated at 6.5 d. No relationship was found 

between the log Kow values of the BFRs and the levels determined in fish. The study 

detected several compounds in fish tissues that may have been metabolites of the dosed 

BFRs. Notably, several hexaBDEs were tentatively identified by full-scan GC/MS and 

these compounds may have been debromination products of BDE183 or BDE209; 

however, one cannot discount the possibility that they were impurities in the material 

dosed to the organisms. No further discussion of potential debromination of BDE209 was 

mentioned in this study. 

Terrestrial Organisms 

Sandholm et al. (2003) conducted a study of the bioavailability, absorption and metabolic 

transformation of decaBDE in Sprague-Dawley rats. A total of 36 male rats weighing 

200–220 g each were dosed with decaBDE at a dosage rate of 2 μmol/kg and the 

bioavailability, elimination and metabolite formation in plasma were monitored over a 

6-d period. One subset of 18 rats was dosed orally by gavage while a second subset of 18 

rats was dosed intravenously. Blood plasma was monitored at regular intervals (at 1, 3, 6, 

24, 48, 72, 96, 120 and 144 h) over the 6-d monitoring period. In addition, plasma 

samples collected from a separate 7-d single-dose study with radiolabelled decaBDE 

(Mörck et al. 2003) were analyzed to determine the total radioactivity associated with 

neutral and phenolic fractions in plasma of rats exposed to decaBDE. 
54

The oral bioavailability was estimated to be approximately 26%, although the sampling 

methodology may have resulted in an artificially high determination of bioavailability. 

Oral bioavailability was defined as the fraction of administered parent compound 

reaching systemic circulation. The authors indicated that the 26% level was much higher 

than that reported in early studies with decaBDE (i.e., by Norris et al. 1975; NTP 1986; 

El Dareer et al. 1987). Based on the analysis of plasma samples from the 7-d study with 

radiolabelled decaBDE, the phenolic radioactivity in plasma at days 3 and 7 was 4 times 

that of the neutral fraction, suggesting significant oxidative metabolism of neutral 

decaBDE, debrominated metabolites, and higher exposure to phenolic metabolites than to 

neutral parent/metabolite compounds. The presence of phenolic metabolites compared 

with neutral compounds indicated that the adsorption of BDE209 would have been 

greater than their calculated level of 26%. The elimination of decaBDE from plasma was 

multi-phased with terminal (i.e., longest) half-lives of 51 h (oral exposure) and 58 h 

(intravenous exposure). 

The neutral compounds in plasma were identified as decaBDE (> 99.5%) as well as three 

nonaBDEs (< 0.5%). In addition to the neutral compounds, 13 phenolic metabolites were 

determined in plasma with the major metabolites inferred to consist of a hydroxyloctaBDE, a hydroxyl-nonaBDE and a guaiacol-type hydroxymethoxy-hexaBDE. Based 

on the observed parent and metabolite composition, the following general sequence for 

metabolic transformation of decaBDE was proposed: 

1. Reductive debromination of decaBDE 

2. Subsequent oxidation to phenolic metabolites 

3. Formation of guaiacol-type metabolites via an arene oxide and dihydrodiol, or via 

two sequential oxidation steps followed by methylation 

The authors also described how hydroxyl-octaBDE and hydroxyl-nonaBDE could have 

affinity for transthyretin (TTR), which normally functions as a transport protein for 

thyroxine in plasma. The binding to TTR could explain the high fraction of phenolic 

metabolites observed in plasma and may cause the phenolic metabolites to persist as 

bound residues in plasma. The binding of phenolic metabolites could disrupt thyroxine 

transport in blood, resulting in adverse hormonal effects. This behaviour was reported by 

the authors to be similar to that reported for PCB metabolites. 

Hakk and Letcher (2003) summarized the findings of Orn and Klasson-Wehler (1998), 

Hakk et al. (2002) and Mörck and Klasson-Wehler (2001) with respect to the fecal and 

biliary metabolites formed by rats dosed with decaBDE. They concluded that decaBDE 

underwent oxidative debromination to guiacols (hydroxyl-methoxylated BDEs), 

hydroxylated BDEs and nonaBDEs. Additional discussions and conclusions by Hakk and 

Letcher (2003) included the following points: 

• A likely metabolic pathway between decaBDE and hydroxylated or 

hydroxymethoxylated BDEs is debromination as a first step, followed by the 

formation of an arene oxide, and subsequent hydroxylation and/or methoxylation. 
55

• PBDE mixtures are inducers of Phase I and Phase II metabolism, suggesting that these 

metabolic pathways are also important for PBDEs. 

• Hydroxylated BDEs are likely to compete with thyroxine for binding to TTR. 

• Guiacol metabolites could further oxidize to quinones, which are highly reactive and 

would bind to cellular macromolecules. Thus, bound residues may make up a 

significant proportion of metabolite body burdens. 

Huwe and Smith (2007a, 2007b) examined the dietary accumulation, debromination and 

elimination of decaBDE in rats. The burden in body tissues and feces of one nonaBDE 

congener (BDE207) and two octaBDE congeners (BDE201 and BDE197) were higher 

than could be explained by the total dose of each congener associated with their 

background concentrations in rat feed. The authors concluded that the elevated burdens 

were the result of reductive debromination of decaBDE. The formation of BDE197 and 

-207 from BDE209 results from meta debromination(s), while para and meta 

debrominations are responsible for BDE201 formation. The authors proposed that the 

action of deiodinase enzymes may be involved in the meta debromination of decaBDE. 

These deiodinase enzymes are normally involved with the meta dehalogenation of the 

thyroid hormone thyroxine and are present in many tissues of the body. 

However, the total burden of BDE207, -201 and -197 accounted for only 1% of the total 

decaBDE dose, suggesting that either debromination was not the primary metabolic 

pathway or that the debrominated products rapidly underwent further metabolism. 

Overall, 45% of the dosed decaBDE was unaccounted for in rat tissues and feces and the 

authors speculated that the formation of bound and/or hydroxylated metabolites that were 

not included in their analysis was a likely explanation for the incomplete mass balance of 

decaBDE. 

Kierkegaard et al. (2007) reported the findings of a 3-month feeding study with two dairy 

cows. The congener profile in adipose and organ tissue differed from the profile in silage 

and it appeared that BDE207, -197, -196 and -182 accounted for a much higher 

proportion of the total burden in tissue than in silage. Tracing of potential degradation of 

decaBDE during extraction and cleanup with radiolabelled standards ruled out photolytic 

debromination as a source of these hepta- to nonaBDEs during extraction/cleanup. The 

authors also discounted higher dietary absorption of these congeners as an explanation for 

the observed differences between feed and tissues since enhanced absorption was not 

apparent for similar congeners (i.e., BDE183 compared to BDE182). Therefore, it was 

concluded that the observed increase of BDE207, -197, -196 and -182 was likely due to 

reductive debromination of decaBDE. The authors proposed the following two potential 

transformation pathways for decaBDE: 

1. Ortho debromination to BDE206, followed by meta debromination to BDE196, 

followed by meta debromination to BDE182 

2. Meta debromination to BDE207, followed by meta debromination to BDE197 

In a study by Riu et al. (2008), Wistar rats were dosed orally with 99.8% pure [14C]-

radiolabelled decaBDE dispersed in peanut oil, gestational days 16 to 19. Urine and feces 
56

were collected daily over this 4-d period. The animals were sacrificed on day 20 of 

gestation (24 h after the last dose of 14C-decaBDE) and the amounts of radioactivity, 

corresponding decaBDE levels, and metabolic profiles were determined in various organs 

and tissues. Extraction and analytical methods were developed to achieve the radiochromatographic separation of decaBDE metabolites formed in vivo. 

More than 19% of the 4-d administered dose was recovered in the body (tissues + 

carcass) of the rats, demonstrating relatively high absorption of decaBDE as well as 

efficient distribution to tissues. The fecal route was the main excretion pathway, 

accounting for approximately 66.3% of the administered dose, whereas urinary excretion 

was very low (accounting for approximately 0.11% of the total dose). In feces extracts, 

approximately 97% of the radiolabel was found to be unchanged decaBDE. 

Characterization of the metabolites in feces and tissues, conducted using liquid 

chromatography/mass spectrometry (LC/MS) nuclear magnetic resonance (NMR), 

identified all 3 nonaBDEs, an unidentified octaBDE and a hydroxylated octaBDE 

derivative. Together (excreta + tissues + carcass), metabolites accounted for 

approximately 7% of the total radioactivity administered to rats. In addition, a very 

limited amount of an additional metabolite was also detected (not enough for structural 

identification); however, it could not be ruled out that the compound corresponded to a 

hydroxylated heptaBDE. Additional unidentified polar metabolites were detected in 

urine, with similar compounds found in the contents of the intestine, but not the feces. 

Riu et al. (2008) found decaBDE residues to be highest in the adrenals, ovaries, liver, 

kidney and heart, and their calculated values were significantly higher than those of 

plasma. Liver was reported as the target tissue for decaBDE (11 mg/kg, corresponding to 

6.5% of the administered dose). However, the highest residual levels were found in 

endocrine glands: the adrenals (33 mg/kg) and ovaries (16 mg/kg). Both decaBDE and its 

metabolites were reported to cross the placental barrier in rats following oral 

administration (approximately 0.5% of the total dose was found in the fetuses). 

Van den Steen et al. (2007) investigated the accumulation in tissues and debromination of 

BDE209 in European starlings (Sturnus vulgaris), using silastic tube implants as modes 

of exposure. The implants were inserted under the skin by a small incision along the ribs. 

The implants contained BDE209 (purity not given, but sourced from Wellington 

Laboratories, Guelph, Canada) dissolved in iso-octane and mixed in peanut oil. The isooctane was evaporated off by heating the oil mixture to a constant weight. After 

preparation, there were no detectable levels of other PBDE congeners in the oil solution. 

In total, seven adult male starlings were used in the study, with four receiving an 

implanted dose of 46.8 ± 2.2 μg BDE209, and with three (the control group) receiving an 

implant with only peanut oil. During the 76-d exposure period, blood samples were taken 

every 3–7 d. The birds were then were euthanized and pectoral muscle and liver were 

excised. Blanks consisting of water instead of blood were included in each sample batch 

and their PBDE concentrations were subtracted from values determined for biotic 

samples. 
57

Before implantation, BDE209 concentrations in blood were below the limit of 

quantitation (i.e., 0.8 and 0.5 μg/L for BDE209 and other BDEs, respectively). 

Accumulated BDE209 peaked on day 10 (16 ± 4.1 μg/L) in the blood of the exposed 

starlings. After this peak, there was a decline to 3.3 ± 0.4 μg/L in blood at the end of the 

76-d exposure period, which suggested elimination. Tissue concentrations were below the 

limit of quantitation in the control group (5.6 and 2.9 μg/kg lipid in muscle and liver, 

respectively). In the exposed group, the muscle concentrations were about two times 

those in liver (i.e., 461 and 340 μg/kg lipid in muscle, compared with 269 and 337 μg/kg 

lipid in liver). Their study found that various PBDEs were present in both the control and 

exposed group muscle and liver tissues, with the differences being most pronounced for 

the nona- (BDE206, -207, and -208) and octaBDEs (BDE196 and -197). The octaBDEs 

BDE203 and -205 did not differ much between the groups. The authors concluded that 

this study provided evidence for the transformation of BDE209 to lower brominated 

PBDEs in songbirds. The significance of these findings is difficult to interpret since the 

composition of the test material was not provided, and thus the accumulation of nonaand octaBDEs from the test material cannot be ruled out. 

Chen et al. (2008) measured BDE209 in all 114 eggs in a peregrine falcon (Falco 

peregrinus) egg study in the northeastern United States. Eggs were collected between 

1996 and 2006 (excluding 1997 and 1998), and concentrations ranged from 1.4 to 

420 ng/g ww. The authors noted that in addition to BDE209, eight nona- and octaBDE 

congeners frequently detected in their peregrine falcon study have rarely been reported in 

wildlife. Together with BDE209, these congeners composed 16–57% of total PBDEs in 

urban eggs and 4.9–53% in rural eggs. The researchers noted that the concentration of 

BDE201 was not in proportion with this congener’s presence in the OctaBDE 

commercial mixture and this was suggestive that it was formed in vivo from higher 

brominated PBDEs. In addition, they detected BDE202 in 107 of 114 eggs. Since 

BDE202 is not known to be found in OctaBDE or DecaBDE mixtures, they also 

speculated that its presence could be attributed to metabolic transformation of higher 

brominated PBDEs. 

3.1.3 Conceptual Models of Metabolism in Vivo

This section summarizes pathways of metabolism based on those proposed in reviewed 

studies. This analysis respects that some studies contain uncertainties (e.g., potential 

impurities in the test substance administered to organisms, lack of chemical specificity in 

analyses of metabolites). Based on the reviewed literature, the amount of decaBDE which 

apparently may have been transformed to lesser-brominated PBDEs like nona-, octa- and 

heptaBDEs in vivo appears to be generally low, on the order of a small percent of the 

total decaBDE dosed. However, some studies have made inferences respecting higher 

levels of metabolism (potentially up to 45 to 65% of the total dose of decaBDE), which 

could be a result of metabolic transformation products including hydroxylated and 

hydroxymethoxylated PBDEs and/or the formation of bound residues. 
58

Fish 

Based on currently available studies and information reviewed in this report by Stapleton 

et al. (2004, 2006), personal communications from HM Stapleton to Environment Canada 

(January 2008 and July 2009) (unreferenced), Kierkegaard et al. (1999) and Tomy et al. 

(2004) (respecting the uncertainties in these latter two studies), the following pathways 

for metabolic transformation of decaBDE are deduced: 

1. DecaBDE (i.e., BDE209) is debrominated in fish as a first step in metabolism, 

producing at least debrominated hepta- to nonaBDEs, but also potentially pentaand hexaBDEs. 

2. Thyroid hormone deiodinase enzymes which normally remove iodine from 

thyroxine appear to be involved in this debromination pathway. 

3. Bromine has been observed to be preferentially removed from the meta and para 

positions. 

Mammals 

According to the discussion and findings of Sandholm et al. (2003), Mörck et al. (2003), 

Hakk and Letcher (2003), Huwe and Smith (2007a, 2007b) and Kierkegaard et al. (2007) 

the following pathways of transformation are deduced: 

1. Reductive debromination to nona-, octa- and heptaBDEs is the likely first step in 

the metabolism of decaBDE. 

2. Similar to fish, debromination may be the result of action by deiodinase enzymes. 

3. The debrominated neutral metabolites then appear to undergo hydroxylation to 

potentially form phenols or catechols, potentially via an arene oxide. This could 

involve the action of cytochrome P450 enzymes. 

4. The potentially formed hydroxylated BDEs may compete with thyroxine for 

binding to TTR, a thyroxine transport protein present in blood serum. 

5. The catechols may then be methylated, potentially by the action of catechol-Omethyltransferase, to form the observed guaicols. 

6. The guiacol metabolites could further oxidize to quinones, which are highly 

reactive and would bind to cellular macromolecules. 

7. The reactive intermediates would also be subject to rapid conjugation via Phase II 

metabolic processes, leading to water-soluble metabolites which would be 

excreted via bile and feces, as was observed in conventional and cannulated rats. 

3.1.4 Implications for Bioaccumulation Assessment 

The existing evidence for the debromination and metabolism of decaBDE suggests 

metabolic pathways which lead to the formation of the following: 

1. Lower brominated PBDEs (down to heptaBDE congeners in mammals and 

potentially down to pentaBDE congeners in fish) 

2. Possibly hydroxylated BDEs (only shown in mammals) 

3. Possibly hydroxymethoxylated BDEs (only shown in mammals) 

4. Unknown products (potential formation in both mammals and fish) 
59

In order to evaluate whether these metabolites have the potential to bioaccumulate, or 

biomagnify in food webs, model predictions were made using the BAF–QSAR model 

and the terrestrial biomagnifications model of Gobas et al. (2003). It is acknowledged 

that in some cases some of the lower brominated PBDEs may have field or laboratory 

measurements of BAFs or BMFs. However, field and laboratory bioaccumulation data 

for the bioformed hydroxylated and hydroxymethoxylated BDE metabolites are generally 

lacking, and application of the models for all types of metabolites was seen as a 

consistent way to inform our knowledge regarding the bioaccumulation and 

biomagnification potential of these metabolites. The PBDE screening assessment 

supporting working document (Environment Canada 2006b) provides a summary of the 

measured aquatic BCFs and BAFs for the PentaBDE and OctaBDE commercial products. 

These studies showed that BCFs and BAFs exceed 5000 for tetra-, penta- and hexaBDEs 

(this is consistent with the findings of the modelling exercise conducted for this review). 

For making BAF predictions with the BAF–QSAR model, the log Kow of each metabolite 

was based on reported measurements wherever possible. In the absence of reported 

measurements, log Kow was estimated using the Experimental Value Adjustment (EVA) 

method in the KOWWIN estimation program. The EVA uses a reference logKow from 

another structurally close substance (in this case decaBDE) and corrects predictions for 

metabolites based on structural differences with decaBDE. For each metabolite, two 

prediction scenarios were conducted: the first with no correction for metabolic 

transformation, and the second corrected for metabolism based on the laboratory 

observations of Stapleton et al. (2004) and Tomy et al. (2004). Where it was not possible 

to estimate kM for lower brominated PBDEs from laboratory data, reasonable 

approximations were made that took into consideration rates of metabolism for similar 

congeners. For hydroxylated and hydroxymethoxylated BDEs, there was no information 

that would allow for estimation of kM, and a value of 0.026/day (i.e., equal to that 

estimated for decaBDE) normalized to the body weight of the model’s middle trophic 

level fish at 15oC (~0.02/day), was chosen for illustrative purposes. These metabolites are 

expected to undergo further metabolism; however, the actual rates of metabolism are 

unknown. The chosen value and resulting BAFs represent reasonable hypothetical 

predictions. Further details of log Kow and kM estimates are provided in Appendix D. 

Figure 3-1 provides frequency histograms of the pooled BAF predictions for all potential 

metabolites and for the middle trophic level. In the absence of metabolism, the BAFs of 

all metabolites are predicted to exceed 5000. However, with consideration given to 

metabolic transformation (a more realistic scenario), the predicted BAFs range from 295 

to approximately 6 000 000. Based on the fit of a normal distribution to the corrected 

BAF histogram, it is estimated that approximately 74% of the metabolite BAF 

predictions exceed 5000. The metabolite groups illustrated to exceed the BAF of 5000 

include hydroxymethoxy-penta- to nonaBDEs, hydroxy-hexaBDEs, and penta- to 

octaBDEs. The remaining 26% of BAF predictions that are below 5000 consist of 

hydroxyl-octa- to nonaBDEs and nonaBDEs. The large proportion of high BAFs, even 

when metabolic transformation is considered, suggests that many decaBDE metabolites 

could potentially be bioaccumulative. 
60

For the hydroxylated and hydroxymethoxy metabolites the actual rate of metabolism is 

unknown and therefore the BAF predictions are uncertain. If the secondary 

transformation of these metabolites occurred much faster than approximately 0.02/d then 

it is possible that the BAFs would be much lower. At the same time, it is possible that in 

certain species the enzymes responsible for Phase II metabolism might be less developed 

or lacking, resulting in persistence of these metabolites in tissues and much higher BAFs 

that approach those of the non-corrected values. 

For the lower brominated PBDE metabolites, the metabolism-corrected BAF predictions 

are considered to have higher certainty than the predictions for the substituted forms 

since the kM values chosen for the model were based on laboratory observations in most 

cases (the one exception was for nonaBDEs, which were assumed to be metabolized at a 

rate based on that determined for decaBDE). It should be noted that the half-lives chosen 

for model inputs were often relatively short (metabolism is relatively fast) compared to 

the maximum half-lives observed in lab studies (refer to Appendix D). Therefore the 

predicted BAFs are less conservative (i.e., lower) than if the longest observed half-lives 

were chosen. Despite the use of less conservative half-lives, the predicted BAFs for 

penta- to octaBDEs exceeded 5000. 

It is acknowledged that decaBDE metabolites are formed in tissues and accumulation in 

tissue from water-phase exposure (as quantified by the BAF) may be of limited relevance 

for decaBDE. However, these BAF predictions are still useful for providing a perspective 

on the bioaccumulation potential of the metabolite chemicals. The model results raise 

concerns that many of the metabolites formed in fish as a result of the accumulation and 

metabolism of decaBDE are potentially bioaccumulative and may be prone to 

biomagnification in food webs, potentially resulting in increased exposure and risk to 

upper trophic level organisms. 
61

Figure 3-1: Combined frequency distributions of predicted BAFs for metabolites of 

decaBDE (predictions for middle trophic level) 

(a) No Correction for Metabolism 

(b) Corrected for Metabolism 

0

1

2

3

4

5

6

7

1.5 2 2.5 3 3.5 4 4.5 5 5.5 6 6.5 7

log BAF

Frequency

Observed

Frequency

Normal

Distribution

0

2

4

6

8

10

12

3.5 4 4.5 5 5.5 6 6.5 7 7.5 8

log BAF

Frequency

Observed

Frequency

Normal

Distribution

BAF = 5000

0

2

4

6

8

10

12

3.5 4 4.5 5 5.5 6 6.5 7 7.5 8

log BAF

Frequency

Observed

Frequency

Normal

Distribution

0

2

4

6

8

10

12

3.5 4 4.5 5 5.5 6 6.5 7 7.5 8

log BAF

Frequency

Observed

Frequency

Normal

Distribution

BAF = 5000
62

BMF predictions for decaBDE metabolites in wolves were made using a spreadsheet 

version of the Gobas et al. (2003) model. Log Kow values were chosen based on the 

rationale described in Appendix D, whereas log Koa for each metabolite was estimated 

using KOAWIN and by inputting corrected logKow values first. Two prediction scenarios 

were modelled: the first did not consider metabolic transformation, while the second 

considered metabolism based on the laboratory observations of Huwe and Smith (2007a, 

2007b). Where it was not possible to estimate kM for lower brominated PBDEs from 

laboratory data, reasonable approximations were made by considering rates of 

metabolism for similar congeners. For hydroxylated and hydroxymethoxylated BDEs, 

there was no information which would allow for the estimation of kM and thus a value of 

0.004/d was selected for illustrative purposes. This value corresponds to the longest 

observed half-life for decaBDE, nonaBDEs and octaBDEs derived by Huwe and Smith 

(2007a, 2007b), normalized to the body weight of wolf used in the Gobas model. The 

hydroxylated and hydroxymethoxylated BDE metabolites are expected to undergo further 

metabolism; however, the actual rates are unknown. 

Calculation of BMFs for the BDEs were corrected for dietary assimilation efficiency 

according to the ED vs. log Kow relationship described in Kelly et al. (2004) for humans 

as a representative homeotherm. In the absence of metabolism, the BMFs of all 

metabolites are predicted to be very high, with BMFs ranging from ~4 to ~89 for 

chemicals within the ranges of log Kow and log Koa estimated for decaBDE metabolites. 

However, with consideration given to metabolic transformation (a more realistic 

scenario), the predicted BMFs are lower, ranging from 5 to 6. All metabolites’ BMF 

values were > 1, primarily due to a relatively high dietary assimilation efficiency 

assumed for these substances. 

In mammals, the actual rate of metabolism is unknown for the hydroxylated and 

hydroxymethoxy metabolites and therefore the BMF predictions are uncertain and may 

be high. If the secondary transformation of these metabolites occurs much faster than 

0.012/d then it is possible that the BMF would be lower and possibly not exceed 1. At the 

same time, it is possible that in certain species, the enzymes responsible for Phase II 

metabolism might be less developed or lacking, resulting in persistence of these 

metabolites and much higher BMFs that approach the BMF values estimated without 

consideration given to metabolic transformation. 

Because these decaBDE metabolites are formed in tissues, the BMF, which quantifies the 

potential for chemical transfer from prey tissues to predator tissues, is seen as an accurate 

indicator of bioaccumulation and biomagnification potential. The predicted BMFs, which 

exceed 1 for all metabolites, raise concerns that the metabolites formed in mammals as a 

result of the accumulation and metabolism of decaBDE have the potential to biomagnify 

in food webs, potentially resulting in increased exposure and risk to upper trophic level 

organisms. 
63

3.2 Debromination in the Environment 

Abiotic Degradation 

The following bullets summarize the evidence regarding abiotic debromination of 

decaBDE which was considered in the screening assessment for PBDEs (Environment 

Canada 2006a, Environment Canada 2006b). 

• Norris et al. (1973, 1974) exposed solid decaBDE (98% decaBDE and 2% nonaBDE) 

to sunlight in water. The bromine concentration in water increased during the 

experiment, suggesting that photodegradation was occurring. However, it was not 

possible to determine the relevance of this study to photolysis in the environment. 

• Watanabe and Tatsukawa (1987) carried out photolysis experiments with decaBDE 

(97% decaBDE and 3% nonaBDE) in a mixture of hexane, benzene and acetone 

(8:1:1). After 16 h of exposure to ultraviolet (UV) light, decaBDE was found to 

debrominate primarily to tri- to octaBDEs, but brominated furans with 1 to 6 bromine 

atoms were also formed. Under sunlight, similar patterns of transformation were 

determined. 

• Söderström et al. (2004) undertook photodegradation studies in which decaBDE 

(Dow FR-300 BA; exact composition was not provided, but contained traces of octaand nonaBDEs) was dissolved in toluene and applied as a thin layer to silica gel, 

sand, soil or sediment. The toluene was evaporated off in the dark. Exposure was to 

natural sunlight and to UV light generated by three mercury lamps. Commercial 

DecaBDE applied to solid matrices was observed to transform following a pathway of 

reductive debromination with products including primarily hexa- to nona-BDEs with 

some occurrence of penta and tetra products (i.e., BDEs 47, -100, -99, -119). Only 

trace amounts of lower brominated PBDEs were formed. Photodegradation occurred 

fastest in toluene and silica gel (estimated half-life of less than 15 min), followed by 

sand (half-life of 12–13 h), sediment (half-life of 30–60 h) and soil (half-life of 150–

200 h). In addition to lower brominated PBDEs, tetra- and pentabrominated 

dibenzofurans (tetra- and pentaBDFs) were detected as phototransformation products 

of decaBDE adsorbed to sand, sediment and soil. Based on the relative amounts of 

debromination products in the samples, the authors concluded that the origin of 

BDE47, -99, -100, -153 and -183, which are common in environmental samples, was 

“probably primarily from emissions of technical PentaBDE products and possibly 

from other degradation pathways of decaBDE.” 

• Jafvert and Hua (2001) conducted multiple photodegradation studies of decaBDE 

(98% purity) on hydrated surfaces of glass and silica sand particles, humic acidcoated silica particles, and glass surfaces in contact with aqueous solutions. The 

studies were conducted under simulated and natural sunlight. The amount of 

decaBDE remaining after light exposure for 60–72 h ranged from 29 to 88% 

depending on the type of surface, test conditions and light used. The presence of 

unidentified compounds was inferred in some of the treatments. Further analysis of 
64

the unidentified compounds was largely inconclusive, although there was some 

evidence of the formation of hexa- to nonaBDEs. 

• Hua et al. (2003) precipitated decaBDE onto surfaces (quartz glass, silica particles 

and humic acid-coated silica particles), hydrated them with reagent-grade water, and 

irradiated the system with artificial sunlight or with natural sunlight. After 60–72 h of 

irradiation, 29–56% of the initial decaBDE remained on the quartz surface, depending 

on the type of light and wavelength used. Addition of humic acid slowed the rate of 

decaBDE decay. Small amounts of nona- and octaBDEs were likely formed as 

degradation by-products. BDE47 and -99 were not formed at detectable levels of the 

analytical method used (high-performance liquid chromatography, HPLC). 

• Palm et al. (2003) investigated the photodegradation of decaBDE dispersed in 

toluene, dichloromethane or a solvent mixture of hexane:benzene:acetone (8:1:1) and 

then irradiated with filtered (300 nm) artificial xenon lamps. A half-life of 

approximately 30 minutes was determined for all the test systems used. Reductive 

debromination was found to occur, first with three isomers of nonaBDE formed, then 

six isomers of octaBDE, then to several isomers of heptaBDE, and finally to trace 

amounts of hexaBDE. Seventy-five percent of decaBDE degradation followed a 

pathway of debromination, while the products of the remaining 25% could not be 

determined. In a follow-up study conducted under natural sunlight over 2 d, the 

researchers found complete disappearance of decaBDE, with nona- to pentaBDEs 

formed. A similar pattern was observed when decaBDE was dispersed in 

tetrahydrofuran (THF) and exposed to artificial sunlight for 84 h; however, PBDEs 

with fewer than 5 bromine atoms per molecule were also formed. 

• Palm et al. (2003) also investigated the photochemical transformation of decaBDE in 

an aerosol smog chamber in which decaBDE was adsorbed to silicon dioxide at a 

concentration of about 1% by weight. The particles were suspended in water, 

atomized, dried and dispersed as an aerosol in a smog chamber. The conditions used 

were assumed to maximize the degradation potential of decaBDE. Although the 

conditions may not replicate atmospheric transport, the particle size used (diameter of 

approximately 1μm in the aerosol) can be considered environmentally relevant. When 

the aerosol was exposed to simulated sunlight (fluorescent lamps) and/or hydroxyl 

radicals, the rate of degradation of decaBDE was just barely measurable, at a rate of 

< 6 x 1013 cm3 molecule-1

s

-1 for reaction with hydroxyl radicals. The products of 

reaction were not determined. 

• Palm et al. (2004) provided early results of photolysis experiments with decaBDE 

adsorbed on pyrogenic silicon dioxide as a carrier material in aqueous suspension. 

The half-life of decaBDE exposed to UV light in aqueous suspension was reported as 

8.8 h, which was much slower than in non-aqueous solvents. Polybrominated 

dibenzofurans (PBDFs) were confirmed as short-lived intermediates in this study. 

There was no information available on the other transformation products that were 

formed. 
65

• Keum and Li (2005) investigated the debromination of decaBDE in aqueous solution 

when in contact with the reducing agents zerovalent iron, iron sulphide and sodium 

sulphide. In the experiments with zerovalent iron, decaBDE was rapidly transformed 

to lower brominated PBDEs. Approximately 90% of the parent was converted to 

mono- to hexaBDE congeners after 40 d. During the initial reaction period (up to 

5 d), decaBDE was predominantly transformed into hexa- and heptaBDEs, but tetraand pentaBDEs were predominant after 14 d. The results demonstrated reductive 

debromination. The experiments with iron sulphide and sodium sulphide also showed 

transformation of decaBDE to lower brominated PBDEs, but the rate was slower than 

that found in the presence of the zerovalent iron. A similar profile of transformation 

products was found to that determined in the experiment using zerovalent iron. The 

Government of the United Kingdom (United Kingdom 2005) points out that the 

actual conditions used in the Keum and Li (2005) experiment are not directly related 

to the environment and that concentrations of sulfides present in the environment are 

a few orders of magnitude lower than those used in this study. Thus, in the 

environment, other chemical parameters could influence reaction rate. Nevertheless, 

similar reactants to those used in this study (e.g., iron-bearing minerals and sulphide 

ions) are present in sediments and soils under anaerobic conditions, although it is 

uncertain whether environmental conditions are appropriate to activate similar 

reactions under natural conditions. 

Biodegradation 

The following bullets summarize the evidence regarding biodegradation of decaBDE that 

was considered in the screening assessment for PBDEs (Canada 2006; Environment 

Canada 2006b): 

• The biodegradation of decaBDE was reported by MITI (1992). Under aerobic 

conditions using activated sludge inoculum, no degradation was detected after two 

weeks of exposure as measured by biological oxygen demand. 

• The Chemical Manufacturers Association Brominated Flame Retardant Industry 

Panel (CMABFRIP 2001) investigated the anaerobic biodegradation of 14C-labelled 

decaBDE in a sediment-water system over 32 weeks. The decaBDE tested was a 

mixture of unlabelled substance (97.4% decaBDE, 2.5% nonaBDE and 0.04% 

octaBDE) with 14C-labelled decaBDE (radiolabelled decaBDE with a purity of 96%). 

Overall, decaBDE was found to be stable under the conditions tested. The study 

determined that less than 1% of the total radioactivity was found as 14CO2 and 14CH4, 

indicating that essentially no mineralization had occurred. 

• Gerecke et al. (2005) undertook experiments under anaerobic conditions using 

sewage sludge as inoculum in glass bottles (100 mL) filled with a 1-cm layer of glass 

beads. These containers were then spiked with technical DecaBDE (98% purity, 

10.0 nmol/sample) and incubated at 37 ±1°C in the dark for up to 238 d. The 

DecaBDE contained trace amounts of nonaBDEs (i.e., BDE 206 at approximately 2% 

on a molar basis, and BDEs 207 and 208 at approximately 0.04% each on a molar 

basis). No octaBDEs were detected in the sample (detection limit was 
66

0.005 nmol/sample). To stimulate degradation, primers (i.e., 4-bromobenzoic acid, 

2,6-dibromobiphenyl, tetrabromobisphenol A, hexabromocyclododecane and 

decabromobiphenyl) were added in the amount of 9–11 nmol/sample. Two other 

experiments were carried out with two nonaBDE congeners to determine the fate of 

these potential decaBDE degradation products. Starch and yeast (50 g) were added 

immediately before the bottle was filled with 20 mL of freshly collected digested 

sewage sludge. The bottles were tightly capped and incubated at 37 ±1°C for up to 

238 d in the dark. Analysis by GC/HRMS indicated that decaBDE decreased by 30% 

within 238 d in the experiments with the primers (i.e., from 11.2 to 7.9 nmol/bottle) 

and this corresponded to a pseudo-first-order degradation rate constant of 1 x 10-3 d-1. 

Without the primers, the rate constant was 50% lower. DecaBDE (BDE209) was 

observed to degrade to 2 nonaBDEs and 6 octaBDEs. This was indicative of reductive 

debromination. All three nonaBDE congeners in the separate bottles were also found 

to undergo reductive debromination. The mass balance between the loss of decaBDE 

and the formation of transformation products illustrated that 3 nmol disappeared, 

while only 0.5 mol of transformation products were identified. The researchers noted 

that this may have resulted from the formation of unidentified transformation 

products or bound (non-extractable) decaBDE residues, or from imprecisions in the 

analytical procedure used. The study demonstrated that the debromination of 

decaBDE proceeded most readily by the loss of bromine from the para and meta 

positions, as shown by the formation of BDE208 or -207. 

3.2.1 New Evidence of Debromination in the Environment 

Abiotic Degradation 

Bezares-Cruz et al. (2004) investigated the photodegradation of decaBDE (98% pure 

obtained from Great Lakes Chemical Company) dissolved in hexane and exposed to 

natural sunlight. DecaBDE (i.e., BDE209) at a concentration of 2-5 μM in hexane 

degraded rapidly to lower brominated PBDE congeners in the presence of sunlight, with 

up to 99% reduction in the decaBDE concentration in 30 minutes (exposure to summer 

sunlight). First-order degradation rate constants were estimated at 1.86×10-3/s. in July 

sunlight (half-life = 6.2 minutes) and 1.11×10-3/s in October sunlight 

(half-life = 10.4 minutes). Unfortunately, the use of hexane as a solvent in this study 

makes it difficult to extrapolate these results to a natural setting. The authors concluded 

that photodegradation in nature would be limited by sorption of decaBDE to particulates, 

light attenuation by humic materials and lower concentrations of hydrogen donor 

chemicals—which may also be less favourable hydrogen donors in the aquatic 

environment. Stapleton (2006a) also concluded that “degradation of decaBDE dissolved 

in water (or organic solvents) is not expected to be of environmental relevance” and this 

is supported by the fact that decaBDE has extremely low water solubility. Since this 

study is not applicable to natural settings, the results were not considered further in the 

assessment of whether decaBDE is undergoing significant debromination to lower 

brominated PBDE congeners in the environment. 

The photodegradation of decaBDE in an 80:20 methanol/water mixture, in pure 

methanol, in tetrahydrofuran and in water/humic acids mixtures exposed to artificial 
67

ultra-violet (UV) light was investigated by Eriksson et al. (2004). The decaBDE used in 

the experiments had a purity of approximately 98%. When exposed to UV light, 

decaBDE dissolved in the organic solvents or solvent:water mixture degraded rapidly, 

with degradation rate constants of approximately 4×10-4/s in the methanol/water mixture 

(half-life ~ 0.5 h), 6.5×10-4/s in methanol (half-life ~ 0.3 h) and 8.3×10-4/s in pure 

tetrahydrofuran (half-life ~ 0.23 h). The use of organic solvents and artificial conditions 

not representative of natural settings makes it difficult to extrapolate these results to the 

environment. The results of this study were therefore not considered relevant to 

environmental conditions. 

The water / humic acid mixture tested by Eriksson et al. (2004) can potentially provide a 

better simulation of potential photodegradation in a natural setting since humic acids are 

often present in aquatic systems and could play a role in photodegradation. To prepare 

the solution for this experiment, 20 mL of a saturated solution of decaBDE in ethanol was 

combined with 10 mL of ethanol containing 50 mg of humic acid. Approximately 10 mL 

of ethanol was evaporated using a nitrogen stream, and the remaining solution was 

combined with 2 L of water. The solution was then heated and maintained at 80°C for 

1 hour under a constant flow of nitrogen. The solution was cooled to room temperature 

and then transferred to cylindrical reaction vessels for the irradiation experiments. The 

Government of the United Kingdom (United Kingdom 2007a) estimated a final humic 

substances concentration for this experiment of approximately 25 mg/L and concluded 

that traces of ethanol may have been present since it was unclear how much ethanol 

would have been lost by heating at 80°C. The experiments were replicated between 2 and 

5 times but it is not clear how many replicates were conducted specifically for the water / 

humic acid treatment. 

The rate of degradation of decaBDE in water with humic acid was 3×10-5/s (half-life 

~ 6.4 h). Although the products were not presented in detail in the article, they were 

described as being almost identical to those determined for the methanol/water 

experiment which produced a range of lower brominated PBDEs (including 3 nonaBDEs, 

at least 7 octaBDEs, 8 heptaBDEs and small amounts of hexaBDEs), mono- to 

pentabromodibenzofurans (mono- to pentaBDFs), and possibly brominatedmethoxylated-dibenzofurans. One key difference with the water / humic acid experiment 

was the formation of a higher proportion of pentaBDFs. Because both water and humic 

acid are naturally occurring substances, it is possible that similar reactions could occur in 

the natural environment. However, the actual environmental rates and degree of 

debromination are uncertain since this experiment used artificial light and it is uncertain 

what fraction of decaBDE in the environment might be associated with humic acids, as 

opposed to particulates. Although the authors also reported results for the water-only 

treatment, there was a high level of difficulty with this experiment due to the extremely 

low water solubility of decaBDE. As a consequence, the findings of the water-only 

experiment were highly uncertain. 

In their study respecting the effect of sewage sludge application on concentrations of 

PBDEs in soils and earthworms, Sellström et al. (2005) provide a brief discussion of an 

investigation of photolytic debromination of BDE209 amended in soil. The BDE209 was 
68

amended with soil from Björketorp, placed in glass test tubes and exposed to artifical UV 

light on a “rocking/rolling” apparatus for 0, 7, 14 and 21 d. Controls were also exposed to 

the same agitation, but were protected from the light. Ultimately, soils showed no 

evidence of photolytic breakdown. The authors concluded that soil appears to encapsulate 

and shield contaminants so that they are less likely to break down when exposed to 

sunlight. 

Hagberg et al. (2006) observed the formation of PBDF congeners as a result of the 

photolytic decomposition of decaBDE in toluene. DecaBDE (i.e., BDE209) was 

dissolved in toluene to a concentration of 2x106

 μg/L and then irradiated with UV 

radiation from a fluorescent tube with or without filtering to generate UV-A (320–

400 nm), UV-AB (280–400 nm) or UV-ABC (250–400 nm) light. The light exposures 

were conducted in petri dishes for 2, 4, 8 and 16 h (parallel exposures) and a dark control 

was also conducted. All samples were subjected to a serial clean-up technique to separate 

PBDFs from PBDEs, and identification and quantification of PBDFs were made using 

HRGC/HRMS. The analysis technique was able to identify PBDFs with 6 or fewer 

bromines; hepta- and octaBDFs were not included in the analysis scheme. After 

decaBDE:toluene solution was irradiated with UV-A, UV-AB or UV-ABC, 27 mono- to 

hexaBDFs were detected, with the majority of products being tetra- to hexaBDFs. The 

PBDFs formed accounted for 0.31% (UV-A), 0.35% (UV-AB) and 1.2% (UV-ABC) of 

the initial amount of decaBDE on a molar basis and the authors suggested a trend toward 

greater transformation (but similar transformation products) at lower wavelengths. The 

authors also concluded that the observed mono- to hexaBDFs were likely the result of 

stepwise debromination of higher PBDFs since mono-to hexaBDEs were not detected in 

any of the irradiated solutions. While these findings demonstrate the possibility of 

decaBDE transformation to PBDFs, they are considered to have low applicability to the 

natural environment. 

Barcellos da Rosa et al. (2003) studied the photolysis of decaBDE dissolved in toluene. 

DecaBDE (BDE209, Sigma-Aldrich, 98% purity) was dissolved in toluene at a 

concentration of 0.31 mM and exposed to light from a 500-watt (W) high-pressure xenon 

lamp. Following light exposure, samples of the decaBDE:toluene solution were analyzed 

by GC – flame ionization detector (FID) to identify and quantify decaBDE and 

debrominated congeners down to hexaBDEs. The decaBDE congener was observed to 

undergo exponential decay with a photolysis rate constant of 3x10-4/s. Several 

debrominated congeners were observed in the decaBDE:toluene solutions and the 

degradation of decaBDE was inferred to proceed via sequential debromination to nona-, 

octa- and heptaBDEs. The authors concluded that while their results demonstrated the 

potential for and pathway of photolytic debromination, further work to confirm the 

environmental relevance of photolysis in toluene was needed. Overall, the findings of this 

study are considered to have limited applicability to the environment. 

Rahm et al. (2005) conducted a study to determine the relative susceptibility of a variety 

of compounds, including decaBDE, to hydrolysis via nucleophilic aromatic substitution. 

When decaBDE was reacted with sodium methoxide dissolved in methanol, the estimated 

half-life for the hydrolysis reaction was 0.028 h, indicating a rapid hydrolysis reaction. 
69

For lower brominated PBDEs, the rate of reaction decreased by roughly a factor of 10 for 

each bromine removed, relative to decaBDE. The authors concluded that decaBDE would 

be susceptible to hydrolysis by nucleophilic compounds in the environment. However, 

given that sodium methoxide is not normally present at significant concentrations in the 

environment and that reaction catalyzed by mineral surfaces, enzymes, etc., would be the 

main reactant for hydrolysis, there is a high level of uncertainty in extrapolating these 

results to the natural environment. 

Geller et al. (2006) report the findings of a photolysis experiment on decaBDE. 

DecaBDE (BDE 209 congener; 98% purity) was dissolved in 3 mL of tetrahydrofuran 

(THF) at saturation (10 g/L, also containing additional particulates of solid-phase 

decaBDE), and irradiated with four lamps for a period of up to 48 h. Following 

irradiation, samples were analyzed by HPLC and GC–MS using electron impact 

ionization with identification of degradation products based on retention times. The 

photolysis products included hepta- to nonaBDEs as well as tri- to 

hexabromodibenzofurans. Comparison with the chromatogram for 2,3,7,8-substituted 

dibenzofurans indicated that these were not major degradation products. These findings 

have similar shortcomings to other studies using organic solvents and are not considered 

environmentally relevant. 

Kuivikko et al. (2007) investigated the photodegradation of decaBDE dissolved in 

isooctane and combined the findings with a model to predict the photodegradation halflife in two marine systems—the Baltic Sea and the Atlantic Ocean. DecaBDE (BDE 209; 

purity > 98.3%) was dissolved in isooctane (250 ng/mL), the solution was placed in 

quartz GC-autosampler vials and the vials were exposed to natural light in Helsinki, 

Finland, in a shallow pool for 60 minutes on October 5. The concentration of parent 

decaBDE was analyzed four to five times during the irradiation; each time there were 

three replicates, a dark control and an isooctane blank. The quantity of decaBDE was 

analyzed using GC–MS. The concentration of decaBDE decreased according to 

first-order kinetics, with a half-life of approximately 0.03 d. 

Kuivikko et al. (2007) reported a quantum yield for decaBDE of 0.28 ±0.04. The 

quantum yield was used in model simulations to predict the photodegradation half-life in 

the Baltic Sea (both surface and 10 m mixing layer) and the Atlantic Ocean (40 m mixing 

layer only). The model simulations assumed environmentally relevant concentrations (3–

4 pg/L and 30–40 pg/L) in the water phase and typical summer solar radiation in the 

Baltic Sea, and also accounted for light attenuation by dissolved and particulate matter. 

Kuivikko et al. (2007) predicted mixing zone half-lives of 1.8 d (Baltic Sea) and 0.4 d 

(Atlantic Ocean), which were the same for both decaBDE concentrations. 

While the model appears to simulate natural conditions, it is important to consider that 

decaBDE in the water column would be primarily sorbed to suspended particulates rather 

than dissolved. Thus, although the degradation of the dissolved phase is predicted to be 

relatively rapid, this degradation would represent only a very small amount of the total 

decaBDE present in the water column. Therefore, there is uncertainty as to whether the 
70

predicted half-lives would be a representative portrayal of the persistence and 

degradation of decaBDE in the water column. 

Ahn et al. (2006a) investigated the photochemical debromination of decaBDE sorbed to 

major components of soil/sediment/mineral aerosols including clay minerals and 

noncrystalline metal oxides which are known to have electron transferring capacities. The 

decaBDE used in the study had a purity of 98%. Test matrices included montmorillonite, 

kaolinite, organic-carbon-rich natural sediment (16.4% OC content), aluminum 

hydroxide, iron oxide and manganese dioxide. Each of the decaBDE-amended test 

matrices (250 mg) was combined with 500 μL water and the mixtures were then 

irradiated with artificial light or natural sunlight. Irradiation with artificial light used four 

24-W lamps with peak output at 350 nm for a period of 14 d. Natural sunlight exposure 

was conducted between July and November 2004, in West Lafayette, Indiana, for a 

period of up to 101 d (further exposure in November/December did not result in 

additional degradation of decaBDE). Samples of decaBDE were quantified using HPLC, 

while the debrominated products were quantified using GC – electron capture detector 

(GC–ECD). 

The dark and light controls showed no signs of degradation although the study reported 

minor peaks of nona- and octaBDEs in the dark control which were likely impurities in 

the decaBDE formulation. For the artificial light exposures, half-lives for 

montmorillonite, kaolinite, sediment and aluminum hydroxide were 36, 44, 150 and 

178 d, respectively. In natural light, half-lives for decaBDE using montmorillonite, 

kaolinite and sediment were 261, 408 and 990 d, respectively, with negligible 

degradation on aluminum hydroxide (note that these half-lives represent days of 

continuous exposure to sunlight, as opposed to actual days including light and dark 

exposure). Degradation of decaBDE was negligible for iron oxide and manganese dioxide 

under either lighting scheme. It is important to note that all of these half-lives are longer 

than the light exposure times, suggesting the potential for uncertainty in these estimates. 

The half-lives for the natural light exposures suggest that decaBDE photodegradation 

occurs slowly in the natural setting. 

The fastest degradation was observed on montmorillonite and kaolinite and these 

matrices were the focus for identification of debromination products. Identified products 

for kaolinite and montmorillonite exposed to sunlight included nonaBDEs (i.e., BDE208, 

-207, -206) and octaBDEs (BDE197 and -196), as well as trace amounts of tri- to 

heptaBDEs over longer sunlight exposure times. Higher fractions of tri-to heptaBDEs 

were also observed for the artificial light exposures. Several unidentified octaBDE 

products were also observed in the chromatograms. The formation of the identified 

products was consistent with stepwise debromination, initially forming nona-, then octaand then heptaBDEs. 

Ahn et al. (2006b) investigated metal oxide mediated debromination of decaBDE using 

birnessite (a naturally occurring manganese oxide mineral) in THF:water and 

water:catechol reactor systems. The purity of the decaBDE used for the experiments was 

98%. The amended birnessite was prepared by combining 0.1 mL of a stock solution 
71

containing 1 mg/mL decaBDE dissolved in THF with 50 mg of birnessite in 15-mL test 

tubes and then air-drying for one day to remove the THF. The first set of experiments 

examined the degradation of decaBDE sorbed to birnessite in THF:water systems. For 

these experiments, the decaBDE-amended birnessite was mixed with 5 mL of THF:water 

solutions at ratios ranging from 0:10 to 10:0. Follow-up experiments examined (i) the 

reactivity of dissolved decaBDE in THF:water at a ratio of 7:3 by varying the amount of 

treated birnessite from 0 to 50 mg/mL, and (ii) the role of THF as the hydrogen donor for 

debromination of decaBDE, determined by measuring the production of succinic acid in 

the system. All treatments were conducted in triplicate and shaken in the dark for a period 

of 24 h with subsampling for analysis at various times. A separate set of experiments 

investigated the degradation of decaBDE sorbed to birnessite in water systems, in the 

presence of the naturally occurring hydrogen donor, catechol. Catechol was combined 

(0.003, 0.045 and 46.135 mmol) with 5 mL of water and mixed with 50 mg of decaBDEamended birnessite. All experimental treatments were shaken in the dark for 23 d with 

subsampling at various times to track the degradation of decaBDE. DecaBDE (i.e., 

BDE209) was extracted in THF and analyzed using HPLC while the potential products 

were quantified using GC–ECD and identified by matching peak retention times with 

known retention times for lower brominated PBDEs. 

The experiments using combined THF:water reactor systems produced rapid degradation 

of decaBDE with > 75% transformation of decaBDE over the 24-hour period for some 

THF:water ratios. However, the follow-up experiments determined that THF was reacting 

with the birnessite to form succinnic acid, and in the process acted as a hydrogen donor 

for the debromination of decaBDE sorbed to the birnessite. Thus, the relatively rapid 

observed degradation rates appeared to depend on the presence of THF. Because THF is 

not normally present in the environment, the rate of decaBDE degradation is not 

considered realistic with respect to conditions present in the natural environment. The 

rapid reaction rates in the THF:water experiments did allow for identification of 

debrominated products. The lower brominated products produced after 24 h included 

tetra- to nonaBDEs. The reaction appeared to proceed in a stepwise manner with

decaBDE being initially degraded rapidly to nonaBDEs, followed by further stepwise 

debromination to lower brominated PBDEs. 

The second set of experiments using water:catechol reaction systems were used to 

examine potential for decaBDE degradation in the presence of the naturally occurring 

substance catechol. No significant degradation was observed over the 23-d period in the 

experiments with 0.003–0.045 mmol catechol. However, slow degradation was observed 

with 46 mmol catechol, with the mass of decaBDE in the reaction vessels decreasing 

from approximately 0.1 μmol to 0.085 μmol during the 23-d experiment. Although the 

products were not quantified for the water:catechol systems, it is possible that they would 

follow a similar pattern to the THF:water systems. Although the rate of reaction was slow 

under the simulated natural conditions, it is possible that decaBDE may eventually 

debrominate over time in the presence of naturally occurring minerals and hydrogen 

donors, such as catechol. 
72

Stapleton and Dodder (2006) reported the findings of a study of photolytic debromination 

of decaBDE in house dust. The dust used for the study was a National Institute of 

Standards and Technology standard reference material (SRM) prepared from vacuum 

cleaner bag contents from homes, motels and hotels in the U.S. The dust is known to 

contain PBDE (including decaBDE), and certified PBDE concentrations were available 

for 15 congeners. Photolysis studies were conducted with both the SRM dust in its 

existing form and with SRM dust without PBDEs (these were removed by Soxhlet 

extraction) but then spiked with a known quantity of decaBDE to yield a concentration of 

2180 μg/kg dw. Analysis of the “cleaned” dust prior to spiking confirmed that decaBDE 

was not detected (< 0.2 μg/kg). Samples (0.5 g) of each dust material were placed in UV 

cuvettes and exposed to natural outdoor sunlight in Gaithersburg, Maryland, between 

9 a.m. and 5 p.m. on days when no precipitation was forecast, for a total of 200 h. Three 

replicates were included for each experiment along with three control samples for each 

dust material, which were covered in aluminum foil. 

The concentrations of decaBDE were found to decrease in both dust materials, with 

first-order degradation rates of 2.3x10-3/h in spiked dust and 1.7x10-3/h in natural dust, 

corresponding to half lives of 301 and 408 h in sunlight, respectively. The authors 

speculated that the longer half-life in natural dust might be due to matrix factors which 

attenuated the amount of light reaching the decaBDE molecules, or that sorbing materials 

present in the natural dust were removed by the extraction process, making the spiked 

decaBDE more available for photodegradation. An additional explanation was that solid 

decaBDE with a limited surface area for irradiation might be present in the natural house 

dust. The spiked dust samples were also analyzed for lower brominated degradation 

products. At the end of the exposure, approximately 38% of the initial decaBDE 

concentration had been lost or degraded. Part of the loss (i.e., equivalent to approximately 

13%) was due to debromination to predominantly the three nonaBDE congeners, but also 

due to lesser amounts of octa- and heptaBDE congeners. The remaining 25% of the 

original decaBDE concentration could not be accounted for and was lost to unknown 

pathways and/or products. The Stapleton and Dodder (2006) study suggested that the 

presence of BDE201 and -202 may provide a marker for debromination of BDE209. In 

the OctaBDE commercial mixture, BDE201 makes up a very small component (i.e., from 

below detection to 0.8%), and BDE203 is not detected. 

This study provides reasonably strong evidence that photodegradation of decaBDE on 

dust can occur under natural conditions in the environment and that lower brominated 

PBDEs can be formed. The authors noted that while the actual degree of sunlight 

exposure to household dust might be limited by windows and shading, dust in cars would 

be subjected to much higher levels of sunlight, making debromination of decaBDE on 

dust in cars potentially significant. 

The Government of the United Kingdom (United Kingdom 2007a) describes the findings 

of an additional study by Stapleton (2006b) on the debromination of decaBDE in house 

dust. The methods were similar to those of Stapleton and Dodder (2006) except that the 

exposure was carried out during the hours of 9 a.m. to 4 p.m. for up to a total of 90 h. 

Sunlight exposures were conducted between July and August 2004 at a mean temperature 
73

of 27.4°C. Over the 90-hour exposure period, decaBDE concentration in the house dust 

decreased from 2180 μg/kg dw to 1570 μg/kg dw, indicating that approximately 28% of 

decaBDE degraded. The start and end concentrations of decaBDE in the dark control 

were not statistically different. The half-life for decaBDE was estimated at 216 h 

(continuous sunlight exposure, or 27 d assuming 8 h sunlight per day). Lower brominated 

PBDEs were detected as degradation products and these included three nonaBDEs, six 

octaBDEs and one heptaBDE. The mass balance of decaBDE and lower brominated 

PBDE congeners indicated that approximately 17% of the original decaBDE was 

unaccounted for, suggesting the formation of alternative (unidentified) products or 

volatilization of lower brominated PBDEs. 

Gerecke (2006) determined the reaction quantum yield of decaBDE on kaolinite and 

measured light penetration in this mineral in order to calculate the photodegradation rate 

of decaBDE. The decaBDE used in the study had a purity of 98%. Light penetration into 

kaolinite was estimated using the Kubelka-Munk theory, which relies on the 

measurement of light absorption (k) and scattering (s). Thin layers of kaolinite were 

prepared on quartz glass with 10 different thicknesses prepared for estimating k and s. 

The kaolinite was spiked with decaBDE in an isooctane/toluene mixture (95/5, v/v). 

Although the spiking method is not provided in detail, it is presumed that after spiking 

the solvent was evaporated off, leaving a solid layer of decaBDE-spiked kaolinite on 

quartz glass, since the light exposures were conducted for either dry kaolinite or wetted 

kaolinite. The spiked kaolinite layers were irradiated with sunlight at noon on clear 

summer days in Dubendorf, Switzerland. The experiments were conducted in a water 

bath to maintain constant temperature and the study included dark controls as well as 

precautions to avoid light exposure outside of the specified times. The concentrations of 

decaBDE and potential products were analyzed using GC/MS. 

Sunlight had only very limited penetration in the kaolinite below 50 μm and the authors 

concluded that only decaBDE sorbed to particles at the very surface of soils would have 

the potential to undergo photolysis. In the experiments with decaBDE-spiked kaolinite, 

half-lives of 76 and 73 minutes were determined for dry and wet conditions. However, 

the observed degradation was non-exponential, likely due to the large difference in 

degradation rate between the upper and lower sides of the kaolinite layer. Under dry 

conditions, degradation products were identified as lower brominated PBDEs, whereas 

for wet conditions, the products were not identified, suggesting that the products were not 

lower brominated PBDEs. These findings demonstrate that while photodegradation of 

decaBDE sorbed to mineral solids is possible in the environment, the rate and amount 

degraded is highly dependent on the penetration of light into soil and mineral layers. 

Nose et al. (2007) studied the degradation pathways of standard decaBDE (purity not 

noted, but purchased from Wako Pure Chemical Industries, Ltd.) during hydrothermal 

treatment. Their evaluation was carried out in a micro autoclave made of stainless steel 

filled with 40 mL of distilled water. The decaBDE material was initially dissolved in 

toluene, and then spiked into the autoclave chamber, which was sealed. Temperature was 

controlled, with the first heating of 25 min to 300oC and a pressure of 8 MPa. The 

experiment was repeated at different time intervals (0, 10, 30, 60, 120, 240 and 360 min), 
74

defined as the processing time. The reaction time was defined as the sum of the heating 

time and the processing time. The chamber was then cooled down to 100oC using a fan, 

and then soaked in ice water for 20 min. 

The study found some decomposition (~45%) after approximately 12 min at 200°C, and 

almost complete breakdown (i.e. > 99%) after 10 min at 300°C. Debromination to 

nonaBDE was determined; however, debromination of meta and para positions to 

BDE208 and -207 occurred at a faster pace than debromination of the ortho position to 

BDE206. The experiment was also conducted with other lower brominated PBDEs with 

similar findings. Thus, the authors concluded that the reactivities of bromine on the para 

and meta positions were relatively high, while the reactivity of the ortho bromine was 

extremely low in hydrothermal treatment. The study also confirmed the formation of 

PBDD/DFs during the study. While having limited applicability to the natural 

environment, the study appears to confirm the findings of other studies respecting 

preferential meta- and para-position debromination (e.g., Stapleton et al. 2006; Gerecke 

et al 2005; Huwe and Smith 2007a, 2007b). As noted in the Canadian screening 

assessment on PBDEs (Canada 2006; Environment Canada 2006), under certain 

combustion/pyrolysis and photolysis conditions, all PBDEs (including DecaBDE) can 

form brominated dibenzofurans and dibenzo-p-dioxins. These transformation products 

are brominated analogues of the Toxic Substances Management Policy Track 1 

polychlorinated dibenzofurans and dibenzo-p-dioxins. Complete destruction of DecaBDE 

and any possible breakdown products appears to occur with exposure to temperatures of 

800°C and above for 2 seconds (European Communities 2002). 

Li et al. (2007) examined the debromination of decaBDE (DE-83R, > 97% purity 

obtained from Great Lakes Chemical Company) by resin-bound zerovalent iron 

nanoparticles. The study involved the use of about 50 test tubes prepared with decaBDE 

in acetone, mixed with an equal volume of distilled water to form an 8-mL solution. Then 

2 g of resin containing zerovalent iron was added. The test tubes were capped and shaken 

for 1 h to 10 d in a water bath (25 ±0.5°C). Analyses were performed using GC with an 

ECD detector, and HRGC/HRMS. The results demonstrated rapid debromination of 

BDE209 after about 8 h, which the authors determined to follow first-order 

transformation kinetics resulting in a rate constant of 0.28 ±0.04/h and half-life of 2.5 h. 

Disappearance of BDE209 occurred, with the subsequent formation of nona- to triBDEs 

in a sequential manner. All three nonaBDEs appeared in significant amounts within 1 h, 

steadily increased for 8 h, then decreased below detection after 24 h. The congener 

pattern dominance shifted to heptaBDEs after 2 d, and then hexaBDEs were most 

abundant. After 10 d, pentaBDEs were present in significant amounts. 

The authors noted that the identification of reaction products was challenging, mainly due 

to the lack of appropriate PBDE standards at the time of their experiments, with many 

peaks not matching any of those for 43 available chemical standards. Issues of co-elution 

were identified. However, all three nonaBDEs were identified with confidence and two of 

five octaBDE peaks were identified (i.e., BDE197 and -196). The hepta- to tetraBDEs 

that were confidently identified included BDE183, -153, -154, -99 and -47 (all contain 

both para-position bromines). Overall, the presence of unidentified peaks makes it 
75

difficult to conclusively confirm the positional preference of debromination pathways. Li 

et al. (2007) also conducted an identical study using PCB209, which dechlorinated at a 

much slower rate than the debromination of BDE209. After 10 d, only 21% of PCB209 

was lost, with only the formation of nona- and octaPCBs identified. Overall, the use of 

zerovalent iron nanoparticles means that this study has uncertain relevance to the 

environment. 

Kajiwara et al. (2008) examined DecaBDE photolysis in plastics under natural sunlight. 

High-impact polystyrene (HIPS) was added to toluene containing 100 μg/mL of 

DecaBDE and shaken overnight to achieve complete dissolution; toluene subsequently 

evaporated in the dark. Crushed television casing and solidified HIPS samples 

compounded with the technical mixtures were further pulverized in a liquid nitrogen 

chamber. Resultant materials were screened with a vibrating stainless-steel sieve 

apparatus. The fine sieved powder (106–300 μm) was used in the irradiation experiment 

(samples of HIPS+DecaBDE and TV casing). Aliquots of the powdered plastic were 

transferred to quartz tubes, sealed with Teflon plugs. A subset of HIPS+DecaBDE 

samples hydrated with 0.5 mL of hexane-washed water was also prepared to examine the 

effect of water on PBDE photolysis. Tubes were kept inside the laboratory in the dark at 

room temperature until the sunlight irradiation experiment was performed. 

The HIPS+DecaBDE samples showed a gradual disappearance of BDE209 upon 

exposure to sunlight. No degradation was found in the dark controls throughout the 

experiment. After one week of exposure to sunlight, the concentration of BDE209 

declined to approximately 50% of the initial level, indicating prompt photodegradation in 

the plastic matrix. Hydrated HIPS+DecaBDE samples showed more rapid degradation of 

BDE209 than non-hydrated samples. The authors suggested that this accelerated 

degradation might be explained by the fact that water played a role as a hydrogen donor, 

promoting debromination. 

Kajiwara et al. (2008) state that previous studies have focused on the photodegradation of 

PBDE dissolved in solvents or absorbed to particles, and that their study is the first to 

show the photolysis of BDE209 compounded into a polymer matrix. Assuming a firstorder reaction, they report a calculated half-life for BDE209 transformation in HIPS of 

51 d. The authors highlight that this half-life for plastic is longer than those measured in 

sand, sediments and soils but equivalent to the half-lives measured in house dust. 

During the study, the HIPS+DecaBDE samples showed that concentrations of hexa- to 

nonaBDE congeners increased several-fold after one week of exposure, and the authors 

interpreted that debromination of BDE209 resulted to some extent in the formation of 

lower brominated PBDE congeners. After this first week, the concentrations of hexa- to 

nonaBDE congeners remained constant or decreased slightly, despite the continuous 

decrease in BDE209. Congeners BDE47, -99 and -100 were not detected and stepwise 

debromination of PBDEs was not clearly observed in this study. At the end of the 

exposure period, the total PBDE concentration was less than 20% of the initial level, and 

the proportion of BDE209 to total PBDEs had decreased from 90% to 44%. Kajiwara et 

al. (2008) reported the photolytic formation of tri- to octaBDFs in their HIPS+DecaBDE 
76

samples. The concentration of total PBDFs showed an increase (greater than 40 times) on 

day 7 of exposure, while BDE209 concentration decreased. 

With respect to the TV casing samples, the study showed no disappearance of BDE209 

and formation of lower brominated PBDE congeners throughout the 224 d of sunlight 

irradiation, in contrast to HIPS+DecaBDE samples. Two possible explanations for the 

differences in degradation profiles were suggested in the study: 1) the difference in initial 

concentrations of BDE209 (HIPS+DecaBDE and the TV casing samples contained 0.1 

and 10% BDE209, respectively) influenced degradation rates; and/or 2) the effects of the 

other plastic additives in TV casings, such as pigments, UV absorbers, and stabilizers, 

potentially affected light penetration depth in each plastic sample. 

Raff and Hites (2007) examined the importance of photolysis in the removal of PBDEs 

from the atmosphere by comparing the photolysis frequencies of PBDE congeners to the 

first-order removal rate constants for oxidant reactions and dry and wet deposition. The 

study compared atmospheric removal processes affecting PBDE (i.e., focused on BDE47, 

-99 and -209) fate using a box model simulating the atmosphere above Lake Superior. 

The model results indicated that photolysis results in the highest loss (90% of the total 

losses) of BDE-47 and -99 from the atmosphere above Lake Superior. However, for 

BDE209, more than 90% of the removal is the result of precipitation events. The authors 

suggested that this explains why the PBDE congener profile found in sediments from 

Siskiwit Lake (located on a remote island in Lake Superior) and the Great Lakes appear 

to be enriched in BDE209 compared to other congeners. These results are consistent with 

congener profiles obtained from air-particle samples collected in the Arctic that showed 

significant depletion of BDE209 relative to tetra- and pentabrominated PBDE congeners. 

The main conclusion of the study is that deposition processes control the loss of BDE209 

from the atmosphere and are responsible for the enhancement of BDE209 found in 

sediment samples from the Great Lakes. 

 

Biodegradation 

Gerecke et al. (2006) report the findings of follow-up experiments to their 2005 study. 

The anaerobic biodegradation of technical DecaBDE (98% purity) in digested sewage 

sludge was investigated using the same test system as Gerecke et al. (2005) but with only 

single primers, either 2,6-dibromophenol or 4-bromobenzoic acid. The use of each primer 

resulted in debromination of decaBDE via loss of para-position bromine to BDE208 but 

the reaction was slow, with a half-life exceeding 700 d. In the absence of a primer, the 

half-life was longer, up to 1400 d. The authors also conducted field monitoring of 

decaBDE in a sewage treatment plant to determine whether biodegradation occurred in a 

full-scale anaerobic digester. Grab samples of influent, reactor and outlet sewage sludge 

were collected from a WWTP at Dubendorf, Switzerland. The concentration of decaBDE 

in sludge was observed to decrease between the influent and outlet streams, suggesting 

that transformation occurred in the full-scale digester. However, the authors cautioned 

that the relatively short residence time in the reactor (28 d) and low level of replication 

(i.e., only one set of grab samples) meant that these results should be considered 

preliminary and not necessarily unequivocal. 
77

The biodegradation of decaBDE (> 98% purity) by cultures of anaerobic bacteria was 

investigated by He et al. (2006). The cultures of anaerobic bacteria included 

Dehalococcoides ethenogenes 195, Sulfurospirillum multivorans and Dehalococcoides 

sp. strain BAV1 (each of which have been shown to dechlorinate organochlorine 

compounds), as well as an enriched autotrophic culture containing D. ethenogenes 195, 

and an enrichment containing a number of Dehalococcoides spp. All cultures were grown 

on ~500 mM trichlorethane (TCE) with the exception of Dehalococcoides sp. strain 

BAV1, which was grown on vinyl chloride. 

The experiments were conducted in 160-mL serum bottles containing a culture medium 

and carbon source appropriate for each organism. For D. ethenogenes (an autotrophic 

culture), no carbon source was used. Prior to inoculation the test vessels were sealed to 

ensure anaerobic conditions and then autoclaved for 25 minutes at 121°C. To initiate each 

experiment, 5 mL of a decaBDE stock solution in trichloroethane (TCE, ~1 μM 

decaBDE) was added to the test vessel for a final concentration of 1 mM TCE and 

0.1 μM decaBDE. The bottles were inoculated with one of the active cultures described 

above at 5% or 10% v/v. The test samples were incubated in the dark at 30°C for a period 

of up to 12 months, with weekly sampling of 1 mL of culture for the analysis of decaBDE 

and degradation products. Each experiment was conducted with two replicates per 

treatment and repeated at least once to confirm the results. Abiotic control samples were 

also included in the experimental design. PBDE congeners were detected using GC–

ECD. 

The experiment with S. multivorans exhibited rapid dechlorination of TCE, but no 

degradation of decaBDE within the first weeks of incubation. After an additional 

two months of incubation without replacement of TCE, decaBDE was observed to 

degrade to non-detectable levels, while octa- and heptaBDEs became detectable in both 

replicates. No degradation was evident in the abiotic controls, indicating that the 

disappearance of decaBDE was due to anaerobic degradation. These findings suggest that 

under appropriate conditions, certain bacteria may debrominate decaBDE. No 

degradation of decaBDE was seen in the experiments using any of the other cultures. 

Parsons et al. (2004) investigated the potential for reductive debromination of decaBDE 

in anaerobic sediment suspensions. The experiments were conducted using sediment 

collected at Hansweert in the Western Scheldt, which is known to contain high 

concentrations of decaBDE. A total of 20 g of the sediment were suspended in 20 mL of 

anaerobic medium and the suspensions were spiked with decaBDE (14 μg/g sediment) 

and incubated anaerobically at room temperature in the dark (details of the anaerobic test 

system were not provided). Sterilized controls were also conducted (details of the 

sterilization procedure was not provided). At varying intervals over an approximately 

205-day test period (read from graph), experimental and control samples were extracted 

in hexane/acetone, subjected to a clean-up process and analyzed using GC–low resolution 

mass spectrometry (LRMS). 

Concentrations of decaBDE in the spiked samples decreased significantly over the first 

two months of incubation and the decrease coincided with the appearance of new 
78

chromatogram peaks which were tentatively identified from their retention times and 

mass spectra as nonaBDEs. The concentrations of nonaBDEs were generally too low to 

be quantifiable and the authors speculated that nonaBDEs were further debrominating to 

lower brominated PBDE congeners. There is significant uncertainty in these findings 

with respect to anaerobic sediment debromination since a similar decrease in decaBDE 

was also found in the sterile controls. The authors suggested that this was due to 

incomplete control sterilization as confirmed by the production of methane upon addition 

of lactic, pyruvic and acetic acids to the control preparations. However, it is also possible 

that photodegradation of decaBDE occurred during sample handling, extraction and 

analysis. Therefore, although Parsons et al. (2004) found a decrease in decaBDE in 

anaerobic sediment suspensions, this decrease cannot be attributed unequivocally to 

anaerobic degradation. 

Parsons et al. (2007) report the findings of an additional investigation of reductive 

debromination of decaBDE in anaerobic sediment microcosms. The experiments were 

conducted using sediment collected at Hansweert in the Western Scheldt, which is known 

to contain high concentrations of decaBDE. A total of 10 g of the sediment were 

suspended in 50 mL of anaerobic medium (containing acetate, lactate and pyruvate). The 

suspensions were then spiked with decaBDE or individual nonaBDE congeners and 

incubated at room temperature in the dark. At varying intervals over an approximate 

260-day test period (read from graph), experimental and control samples were extracted 

in hexane/acetone, subjected to a clean-up process and analyzed using GC/MS in selected 

ion monitoring mode. 

There was no measurable decrease in decaBDE in the active and control microcosms, 

although nonaBDEs were detected in the decaBDE-spiked samples at much higher 

concentrations than known background levels in the study sediments 

Knoth et al. (2007) conducted a monitoring study of PBDEs, including decaBDE, in 

sewage sludge from 11 municipal WWTPs in Germany. A total of 39 sludge samples 

from different stages of the treatment process—primary sludge, secondary excess sludge 

and (dewatered) digested sludge—were collected from March 2002 to June 2003. The 

samples were subjected to a process of sterilization, freeze-drying, and spiking with 

stable isotope standards, followed by Soxhlet extraction in toluene, four-column clean-up 

of the toluene extraction, and reduction to 100 μL. The extracted and cleaned-up sample 

was analyzed by GC–select ion monitoring (electron ionization) (SIM(EI+))LRMS to 

quantify decaBDE and lower brominated PBDE congeners. 

The congener profiles in the various treatment plants and treatment stages were 

dominated by decaBDE, which ranged from 97.1 to 2217 ng/g dw. Debromination of 

decaBDE to lower brominated PBDEs during treatment was either not occurring or too 

slow to be detected during the total retention time of sludge (11–13 d) in German 

WWTPs. 

La Guardia et al. (2007) examined potential for in vivo and environmental debromination 

of decaBDE in a WWTP and its receiving environment. Concentrations of decaBDE and 
79

lower brominated PBDEs were monitored in sludge, sediments, and fish in the receiving 

environment of a WWTP located in Roxboro, North Carolina (further study details in 

Section 2.2). The PBDE congener profile was tracked from WWTP to receiving 

environment sediments and to biota in order to evaluate whether significant 

debromination was occurring. For the sludge samples, 17 PBDE congeners were 

identified in 2002 while 18 PBDE congeners were identified in 2005. The major 

congener in sludge was BDE209, accounting for 60% (58 800 μg/kg dw) of the total 

PBDE burden in 2002 and 87% (37 400 μg/kg dw) of the total PBDE burden in 2005. 

The sludge congener profiles were similar to the technical formulations (PentaBDE and 

DecaBDE), suggesting minimal debromination in WWTP sludge. In receiving 

environment sediments the major PBDE congeners included BDE209, -206, -99 and -47, 

with BDE209 making up > 89% of the total BDE burden in each of the sediment 

samples. The decaBDE concentration in sediments was highest between 1 and 6 km 

downstream of the outfall, with maximum levels ranging from 3240 mg/kg OC to 

2450 mg/kg OC. The lower brominated PBDE congeners were attributed to the 

PentaBDE technical formulation, suggesting minimal debromination of decaBDE in 

surficial sediments. 

Tokarz et al. (2008) conducted concurrent PBDE (i.e., BDE209, -99 and -47) 

experiments using anaerobic sediment microcosms and a cosolvent-enhanced biomimetic 

system. The sediment microcosms contained natural sediments with no detectable PBDEs 

collected from Celery Bog Park, West Lafayette, Indiana. The PBDEs were dissolved in a 

toluene solution and added to sediments, and then the solvent was evaporated off. This 

mixture was then blended with wet sediments to form a concentration of approximately 

5.0 μg/g and 3 μg/g for BDE99 and -47, and BDE29, respectively. The microcosms were 

fed methanol and dextrose to ensure the formation of anaerobic conditions, and to 

provide electron donors. Autoclaved control systems were implemented. The biomimetic 

experiment involved the use of Teflon-capped glass vials with 0.03 mM of BDE209, -99 

or -47 mixed with 5.0 mM titanium citrate and 0.2 mM vitamin B12 in 0.33 M TRIZMA 

buffer solution containing tetrahydrofuran. Controls were used without titanium citrate. 

Debromination products were identified and quantified using GC–ECD and GC–MS. 

Tokarz et al. (2008) notes that cobalamins, such as coenzyme vitamin B12, have the 

ability to mediate reductive dehogenation of halorganics. Cobalamins from decaying cells 

have been iscolated from environmental samples. In addition, they indicate that vitamin 

B12 has ubiquitous presence in the environment in anaerobic micro-organisms. 

The biomimetic system demonstrated reductive debromination at decreasing rates with 

decreasing bromination (e.g., half-life of 18 seconds for BDE209 and almost 60 d for 

BDE47). In natural sediment microcosms, the half-life for BDE209 was estimated to 

range from 6 to 50 years, with an average of 14 years, based on observations over 

3.5 years. After 8 months, BDE47 decreased approximately 30% without a consistent 

concurrent increase in daughter debromination products. While complete debromination 

to diphenyl ether was not ruled out, it appeared unlikely since no intermediate products 

were identified. The researchers speculated that there could have been formation of 

hydroxylated and methoxylated derivatives of tetraBDE. The researchers synthesized 
80

their data from both systems and proposed major debromination pathways for sediment 

and biomimetic systems as follows: BDE209 > nonaBDEs (BDE206, -207 -208) > 

octaBDEs (BDE196, -197) > heptaBDEs (BDE191, -184, 2 unknown heptaBDEs) > 

hexaBDEs (BDE138, -128, -154, -153) > pentaBDEs (BDE119, -99) > tetraBDEs 

(BDE66, -47, -49) > triBDEs (BDE28, -17). Specifically, at the end of the 3.5-year study, 

their analysis of BDE209 degradation in sediments identified BDE208, -197, -196, -191, 

-128, -184, -138, and -128, as well as 3 unidentified octaBDEs and 2 unidentified 

heptaBDEs. 

The findings of Tokarz et al. (2008) showing prolonged persistence of decaBDE in 

sediment microcosms are supported by a field study of Eljarrat et al. (2008) and 

Sellström et al. (2005). Eljarrat et al. (2008) examined the fate of PBDEs in sewage 

sludge from five municipal WWTPs after agricultural application in Spain in 2005. 

PBDE concentration in sewage sludge ranged from 197 to 1185 ng/g dw, with BDE209 

being the predominant congener, ranging in concentrations from 80.6 to 1083 ng/g dw. 

Concentrations of BDE209 in soils ranged from 14.6 to 1082 ng/g dw at seven 

agricultural sites. High concentrations (i.e., 71.7 ng/g dw) were even found at one site 

that had not received sludge application for four years, illustrating the persistence of 

BDE209 in soils. Sellström et al. (2005) collected PBDE samples from sites that received 

past sewage sludge amendments. They sampled soil from three research stations (with 

reference plots and sewage-sludge-amended plots) and two farms (reference and 

amended/flooded soils) in Sweden. They determined that decaBDE concentrations in soil 

ranged from 0.015 to 22 000 ng/g dw. Highest levels were noted at the farm site that had 

not received amendments for 20 years. 

Kohler et al. (2008) examined sediment concentrations and temporal trends of deca-, 

nona- and octaBDEs utilizing a sediment core from Greifensee, a small lake located near 

Zürich, Switzerland. In their study, they note that PBDEs first appeared in sediment 

layers corresponding to the mid-1970s. The authors identified all three nonaBDEs and at 

least seven octaBDEs (BDE202, -201, -197/204, -198/203, -196/200, -205 and -194) in 

the lake’s surface sediments using reference materials. While deca- and nonaBDE 

concentrations were found to increase rapidly with time, the increase of the octaBDE 

concentrations was slower. Congener patterns of the octa- and nonaBDE congeners in the 

sediments did not change with time, and thus transformation of PBDEs was not said to be 

observed in the sediments over the 30-year span. However, the authors noted different 

relative proportions of nona- and octaBDEs in the sediments in comparison to the 

commercial PBDE mixtures. Notably, they detected BDE202 which is apparently not 

reported in any of the PBDE mixture and they speculated that its presence could have 

been due to transformation of decaBDE. Also, they noted that congener patterns of 

octaBDE were dissimilar to those of the commercial PBDE mixtures, but similar to 

patterns in house dust and profiles of photodegradation products of decaBDE. 

Consequently, they suggested biotic and/or abiotic transformation processes are involved 

between the release of the technical products into the environment and their final residues 

in sediments that are responsible for the changes of congener patterns that they observed 

in the sediments of the Swiss lake. 
81

Zhou et al. (2007) evaluated the ability of white rot fungi to degrade BDE209 in a liquid 

culture medium, and the effects of Tween 80 and ß-cyclodextrin on the degradation of 

BDE209 by white rot fungi. White rot fungi have been shown to rapidly oxidize and 

mineralize many aromatic compounds, including PCBs. In order to improve BDE 

bioavailability, “solubilizing” agents such as the surfactant Tween 80 or cyclodextrin 

were added to their test systems. The authors note that the application of this technology 

is most suited for bioremediation applications. Overall, this study has unknown relevance 

to environmental conditions. 

Zhou et al. (2007) utilized 1 mL of decaBDE (98% pure) in dichloromethane added to 

250-mL flasks. The solvent was then allowed to evaporate, resulting in a total BDE209 

mass of 16 μg coating the bottom of the flask systems. A 100-mL aliquot of aqueous 

medium (distilled water, malt extract, glucose, peptone and yeast extract mixture) was 

added to each flask. White rot fungi were inoculated to the liquid culture and the test 

system was shaken for 10 d in the dark. Identical test systems were also implemented 

utilizing additions of Tween 80 and cyclodextrin concentrations ranging from 0 to 

900 mg/L. Analyses were conducted using HPLC with a UV detector. The test systems 

with only white rot fungi added showed a decrease of 42.2% over 10 d in the amount of 

BDE209 in the test system. The sterile controls showed no significant degradation over 

time. Tween 80 was found to enhance BDE209 degradation at an appropriate 

concentration (maximum degradation 96.5% over 10 d). Cyclodextrin was also shown to 

enhance BDE209 degradation (maximum degradation of 78.4% over 10 d). 

Transformation products were not identified in this study. 

Orihel et al. (2009) presented an ongoing study that may provide some key information 

on the environmental transformation and aquatic bioaccumulation of decaBDE. Their 

field mesocosm study compares sediment and periphyton BDE209 concentrations from 

low (dose=0.039 g of DecaBDE), medium (dose=0.28 g of DecaBDE) and high 

(dose=2.3 g of DecaBDE) treatments against a control mesocosm. Based on preliminary 

results, the authors suggest that DecaBDE breakdown had occurred since lower 

brominated PBDE products were observed in surface sediments as soon as 1 month after 

DecaBDE addition to mesocosm water. In their study, penta- to hepta-BDEs were 

observed in the medium- and high-mesocosm treatments, but not in the low-treatment or 

control mesocosms. DecaBDE was only observed in periphyton in the medium- and 

high-treatment mesocosms. NonaBDE was observed in periphyton in the high mesocosm 

after 1 month, but not after 4 months. Hexa-, hepta- and octaBDEs were not detected in 

periphyton after 1 month, but were after 4 months in the medium- and high-treatment 

mesocosms. In the coming years, the results will be finalized. 

3.2.2 Conceptual Model of Transformation in the Environment 

The likelihood, rates and potential products of decaBDE debromination will depend upon 

the medium in/on which it is present, and the rate of various degradation processes (e.g., 

photodegradation, abiotic degradation, biodegradation). To illustrate the partitioning 

properties of decaBDE under various possible release scenarios, Level III fugacity 

modelling using the CHEMCAN model (Webster et al. 2004) was conducted and 
82

parameterized for the “Ontario Mixed Wooded Plain” region. The primary model 

parameter inputs included 

• log Kow = 8.7 (Wania and Dugani 2003; see Appendix C). 

• log Koa = 12 (maximum value allowed in model, indicating that decaBDE is 

essentially non-volatile; estimated values for decaBDE exceed 15). 

• degradation rates = negligible in air, water, soil and sediment.7

Based on its chemical properties, decaBDE is expected to be primarily sorbed to organic 

fractions. Table 3-1 summarizes the predicted fractions in each medium for releases to 

air, water or soil. From Table 3-1 it is apparent that 

1. A large fraction (> 96%) of decaBDE in the environment is expected to be 

associated with either soils or sediments (depending on whether the release is to 

soil or the aquatic environment) and, within these bulk compartments, decaBDE 

is associated almost entirely with the solid phase. 

2. Although < 3.4% of decaBDE in the environment is expected to be associated 

with bulk air or bulk water phases, it is expected that of the mass of decaBDE 

present in bulk air or bulk water > 95% will be sorbed to either aerosol or 

suspended sediment. 

3. Regardless of the release scenario, less than 0.1% of decaBDE in the environment 

is predicted to be found dissolved in water or in the vapour phase. 

Thus, studies that examine the debromination of decaBDE present in soils and sediments 

or sorbed to particulate and/or biota phases in air or water likely give the best indication 

of the potential significance of a transformation process in the environment. For studies 

that examine the debromination of decaBDE in solution or in vapour phase, the 

significance of transformation is less certain and is likely low since the dissolved/vapour 

fraction of decaBDE in the environment is predicted to be extremely low. According to 

Stapleton (2006a), 

In the environment, it is expected that BDE 209 will be found primarily bound to solids in the 

water column, and bound to particles in the atmosphere. Therefore, degradation of BDE 209 

dissolved in water (or organic solvents) is not expected to be of environmental relevance.

A further consideration for evaluating the available information on debromination in the 

environment is how well the experimental conditions represent those that would be 

expected under natural settings. For example, in experiments that use organic solvents to 

disperse or dissolve decaBDE, extrapolation to normal environmental conditions is 

difficult since artificial organic solvents used in laboratory studies are not commonly 

found in the environment and the solubility of decaBDE in water is much lower than that 

of organic solvents. Furthermore, organic solvents are often much stronger hydrogen 

donors than water or other naturally occurring substances, resulting in faster rates and 

potentially non-representative patterns of decaBDE transformation. In the environment, 

many different donor molecules would be present, and thus substitution of bromine by 

 

7

 While some degree of transformation is likely to occur in the environment, the intent of the model 

simulation was to examine the partitioning properties of decaBDE. 
83

hydrogen is unlikely to be the sole route of transformation. This creates uncertainty as to 

how well degradation studies that use organic solvents would represent expected 

degradation of decaBDE in the environment. Other experimental conditions also require 

scrutiny to ensure relevance to those that may be seen in the environment. For instance, 

experiments using natural sunlight or simulated natural sunlight are likely to better 

represent natural conditions than those using artificial light. It is also not likely that 

materials such as zerovalent iron nanoparticles, specialized microbial cultures and fungal 

cultures will be found in great abundance in the environment. 

In general, laboratory studies examining decaBDE transformation were frequently 

conducted under conditions that were favourable for decaBDE debromination and are not 

necessarily representative of typical conditions in the environment. Thus, while some 

transformation of decaBDE in the environment is plausible, data showing accumulations 

of decaBDE in the environment (e.g., see Canada 2006; Environment Canada 2006b) also 

imply recalcitrance. 

Table 3-1: Predictions of the CHEMCAN Model for decaBDE in the “Ontario Mixed Wooded 

Plain” region 

1 kg/year to 

water1 1 kg/year to 

soil1 1 kg/year to air1

Mass in bulk air (kg) 1.27E-07 2.17E-07 2.23E-03 

Fraction of total mass in bulk air 1.46E-06% 7.64E-08% 3.70E-03% 

In bulk air, fraction as vapour 0.21% 0.21% 0.21% 

In bulk air, fraction sorbed to aerosol 99.79% 99.79% 99.80% 

Mass in bulk water (kg) 0.31 0.31 9.49E-02 

Fraction of total mass in bulk water 3.38% 0.11% 0.16% 

In bulk water, fraction dissolved in water 0.12% 0.01% 0.12% 

In bulk water, fraction sorbed to suspended 

particles 95.25% 95.35% 95.25% 

In bulk water, fraction accumulated in fish 4.63% 4.63% 4.64% 

Mass in bulk soil (kg) 3.32E-03 274.36 57.59 

Fraction of total mass in bulk soil 0.00% 96.80% 95.35% 

In bulk soil, fraction in gas phase in air 

spaces 0.00% 0.00% 0.00% 

In bulk soil, fraction dissolved in soil water 0.00% 0.00% 0.00% 

In bulk soil, fraction sorbed to soil solids 100.00% 100.00% 100.00% 

Mass in bulk sediment (kg) 8.78E+00 8.78 2.71 

Fraction of total mass in bulk sediment 96.60% 3.10% 4.49% 

In bulk sediments, fraction dissolved in 

porewater 0.00% 0.00% 0.00% 

In bulk sediments, fraction sorbed to 

sediment solids 100.00% 100.00% 100.00% 

Notes: 

1

 Releases to other media are assumed to equal zero. 
84

Figure 3-1 integrates the available information regarding possible environmental 

debromination pathways for decaBDE in a conceptual model while Table 3-2 provides a 

list of the products which are potentially formed from each debromination process. A 

tabulated summary of the environmental debromination information is provided in 

Appendix E. 

Studies focusing on decaBDE sorbed to particulates or solid phases conducted in the 

absence of organic solvents, with natural sunlight, may provide a realistic indication of 

decaBDE transformation in the environment. Figure 3-2a summarizes the 

photodegradation pathways for decaBDE sorbed to particulate or solid phases. DecaBDE 

(i.e., BDE209) sorbed to dust or other dry minerals and particulates appears susceptible to 

relatively rapid transformation, with half-lives ranging from 76 min (decaBDE sorbed to 

a thin film of kaolinite; Gerecke 2006) to 408 h (decaBDE sorbed to house dust; 

Stapleton and Dodder 2006). Transformation appears to follow stepwise reductive 

debromination to form hexa- to nonaBDEs. A few studies also indicated alternative 

reaction pathways to either tetra- and pentaBDFs or unidentified products. 

DecaBDE (i.e., BDE209) sorbed to particulates in aqueous systems appears to 

photodegrade at a generally slower rate, with half-lives ranging from 73 minutes 

(decaBDE sorbed to a thin film of kaolinite in presence of water; Gerecke et al. 2006) to 

990 d (decaBDE sorbed to sediment organic carbon and exposed to sunlight; Ahn et al. 

2006a). In addition, some studies have also demonstrated significant degradation (up to 

71%) over 60–72 h. The identified products included hexa- to nonaBDEs and some lower 

brominated PBDEs. The potential for photolytic transformation of the sorbed decaBDE 

appears to be modulated by the ability of sunlight to penetrate water and the sorbing 

matrix. Gerecke et al. (2006) determined that light penetrated only 50 μm into kaolinite, 

suggesting that the photodegradation of decaBDE would be mainly limited to that amount 

of the chemical present on exposed surfaces. In surface waters, sunlight reaching 

suspended and bottom sediments would also experience some level of attenuation. Thus, 

it can be concluded that while photodegradation on solid phases can occur at significant 

rates, only a very small fraction of the total decaBDE in the environment (i.e., adsorbed 

to particulates or on solid surfaces) which has contact with sunlight would be susceptible 

to photodegradation. Also, depending on matrix characteristics (e.g., shielding capacity) 

and level of exposure to sunlight, photodegradation will be limited. 

In addition to photodegradation, decaBDE sorbed to solids may also be subject to 

biodegradation processes in aquatic sediments, soils or WWTPs. Figure 3-2b summarizes 

biodegradation pathways for decaBDE. The results of biodegradation studies are 

somewhat mixed. Early studies (MITI 1992; CMABFRIP 2001) focused on decaBDE 

mineralization and these indicated very little, if any, degradation; however, these studies 

did not specifically determine the production of potential transformation products. Tokarz 

et al. (2008) showed that decaBDE in natural sediment microcosms had a half-life that 

ranged from approximately 6 to 50 years, with an average of 14 years. In laboratory 

studies using activated sludge, Gerecke et al. (2005, 2006) determined half-lives ranging 

from 693 to 1400 d depending on the presence/absence of primer, and identified 

debromination products as octa- and nona-BDEs. In a separate study, He et al. (2006) 
85

observed complete transformation of decaBDE to hepta- and octaBDEs over 2 months 

with one anaerobic culture (Sulfurospirillum multivorans) but negligible degradation with 

other anaerobic cultures. The monitoring studies of WWTP sludge (Knoth et al. 2007; La 

Guardia et al. 2007) provide little evidence of decaBDE debromination in WWTPs, 

possibly because the residence time in WWTPs is too short for significant debromination 

to be observed. In addition, sediment microcosm studies by Parsons et al. (2004, 2007) 

fail to provide evidence of significant debromination, although Parsons et al. (2007) did 

identify small amounts of nonaBDEs. Thus, while the experimental conditions of the 

activated sludge studies are environmentally relevant, it is possible that the rates of 

degradation are too slow, or the cultures used are too specific, for the observed 

debromination to be significant in the environment. Overall, it appears that 

photodegradation may be more significant than biodegradation for decaBDE sorbed to 

solids. 
86

Figure 3-2: Conceptual model of possible environmental transformation pathways for 

decaBDE 

A) Photodegradation on Particulates (Dust, Soil, Aerosol and Sediment) in 

Dry and Aqueous Systems

 
87

B) Biodegradation 
88

(C) Photodegradation in Solution (Natural and Organic Solvents)
89

D) Abiotic Degradation (Other than Photodegradation) 
90

Table 3-2: Summary of decaBDE transformation products observed in laboratory studies 

with some relevance to environmental settings 

 DecaBDE Transformation Process Congener Group Formed Targeted for Virtual Elimination under CEPA 1999 Photodegradation Sorbed to Solids in Dry and Aqueous Systems (references in brackets) Photodegradation Dissolved in Natural Solvents (references in brackets) Photodegradation Dissolved in Artificial Organic Solvents** (references in brackets) Biodegradation (references in brackets)

Abiotic Degradation (references in brackets)

nonaBDEs X (2,3***,4,9,10,11,

21) X (8) X (1,5,6,7,8,19) X (14,15,16, 21† & ††) X (12,13**,20**) 

octaBDEs X (2,3***,4,9,10,11,

21) X (8) X (1,5,6,7,8,19) X (14,15, 21† & ††) X (12,13**,20**) 

heptaBDEs X (2,3***,9,10,11, 

21) X (8) X (1,5,6,7,8,19) X (15,21† & ††) X (12,13**) 

hexaBDEs X X (2,3***,9,21) X (8) X (1,5*,6,7*) X (21† & ††) X (12,13**,

20**) 

pentaBDEs X X (9*) X (1,6) X (21††) X (12,13**,

20**) 

tetraBDEs X X (9*) X (1,6) X (21††) X (12,13**,

20**) 

triBDEs X (9*) X (1,6) X (21††) X (20**) 

pentaBDFs X (21) 

heptaBDFs X (21) 

hexaBDFs X (21) X (18) 

triBDFs (2,21) X (8,18) 

tetraBDFs X (2,21) X (8,18) 

triBDFs (21) X (8,18) 

diBDFs X (18*) 

monoBDFs X (18*) 

Unidentified 

products X (10,11,21,17) X (5) 

Notes:

* Indicates that only trace or small amounts were formed. 

** Due to the use of artificial solvents, these results are considered to have low relevance to the environment. 

***Identification of debromination products was largely inconclusive, although there was some evidence of the formation of 

hexa- to nonaBDEs. 

†In natural sediments. 

††In cosolvent enhanced biomemetic system. 

BDEs – brominated diphenyl ethers 

BDFs – brominated dibenzofurans 

The results of Gerecke (2006) have not been included because the decaBDE degradation products were described as 

lower brominated PBDEs, with no indication of which PBDEs were formed. 

The results of Parsons et al. (2004) have not been included because degradation to nona- and lower brominated PBDEs 

was observed in both the control and treatment groups. 
91

References

1. Watanabe and Tatsukawa 1987 

2. Söderstrom et al. 2004 

3. Jafvert and Hua 2001 

4. Hua et al. 2003 

5. Palm et al. 2003 

6. Bezares-Cruz et al. 2004 

7. Eriksson et al. 2004 

8. Geller et al. 2006 

9. Ahn et al. 2006a 

10. Stapleton and Dodder 2006 

11. Stapleton 2006b 

12. Keum and Li 2005 

13. Ahn et al. 2006b 

14. Gerecke et al. 2005 

15. He et al. 2006 

16. Gerecke et al. 2006 

17. Gerecke 2006 

18. Hagberg et al. 2006 

19. Barcellos da Rosa et al. 2003 

20. Tokarz et al. 2008 21. Kajiwara et al. 2008 
92

Although photodegradation experiments with decaBDE dissolved in water potentially 

represent environmentally relevant conditions, the significance of any observed 

debromination is uncertain since only a very low fraction of decaBDE would be present 

in the dissolved phase. Figure 3-2c provides a conceptual model of photodebromination 

pathways for decaBDE dissolved in water/solvent systems. Two of the associated studies 

are notable because the results are potentially relevant to the natural environment. 

Eriksson et al. (2004) observed relatively rapid transformation (half-life = 6.4 h) of 

decaBDE in water/humic acid mixtures with hexa- to nonaBDEs formed. Because 

artificial light was used, there is some uncertainty whether the rate would be as fast in 

natural sunlight. Kuivikko et al. (2006, 2007) observed a photodegradation half-life of 

0.3 d for decaBDE in isooctane but conducted modelling to estimate half-lives ranging 

from 0.2 to 1.8 d in Baltic Sea and Atlantic Ocean surface waters. 

Laboratory-based studies on the transformation of decaBDE provide support for a 

conclusion that transformation to lower BDEs and BDFs should be occurring in the 

environment. However, there is a dichotomy between the findings of these studies and 

the results of monitoring studies in terms of realistic support for significant 

transformation of decaBDE in the environment. If these laboratory-based studies provide 

accurate depictions of transformation in the environment, one would expect to find 

similarities in the patterns observed in the environment and those described as 

transformation products of decaBDE in laboratory studies. One would expect to see 

widespread occurrence of the decaBDE transformation products in various environmental 

media, such as sediments. The Canadian screening assessment on PBDEs (Canada 2006) 

showed how monitoring studies are dominated by congener profiles that are consistent 

with the PentaBDE and OctaBDE commercial mixtures. The United Kingdom (2007a) 

notes that the available monitoring data provide limited circumstantial evidence for a 

correlation between concentrations of decaBDE and lower brominated PBDEs in some 

environmental matrices (e.g., citing the findings of Voorspels et al. (2006a, 2007b). The 

United Kingdom (2007a) also notes that, in the study by Gerecke et al. (2005), 

biodegradation appeared to preferentially remove the para-position bromines from the 

BDE molecule, but that these PBDEs are not routinely observed in the environment. 

Besides photodegradation and biodegradation, decaBDE is susceptible to abiotic 

debromination in the absence of light exposure as shown under laboratory conditions. 

Figure 3-2d summarizes the transformation pathways for these other abiotic degradation 

processes. Keum and Li (2005) observed up to 90% transformation of decaBDE 

dissolved in deionized water over 40 d in the presence of reducing agents such as 

zerovalent iron, iron sulphide and sodium sulphide. Li et al. (2007) showed complete 

disappearance after 8 h in a water-acetone system containing nanoscale zerovalent iron, 

while Rahm et al. (2005) observed a half-life of 0.028 h for decaBDE dissolved in 

methanol and reacted with sodium methoxide. Similar to photodegradation studies using 

organic solvents, there is uncertainty as to how well these studies represent natural 

conditions and whether such processes may take place under natural conditions. 
93

3.2.3 Implications for Bioaccumulation Assessment 

The existing laboratory evidence for the transformation and debromination of decaBDE 

through environmental processes such as photodegradation, abiotic degradation and 

biodegradation indicates potential transformation pathways which lead to the formation 

of 

1. lower brominated PBDEs; 

2. brominated dibenzofurans; and 

3. unidentified degradation products. 

However, the laboratory studies also indicate that both the lower brominated PBDEs and 

PBDFs would, themselves, be susceptible to further degradation and the actual 

significance of the formation of these products is unknown. Palm et al. (2003) and 

Eriksson et al. (2004) have shown that the rate of transformation of PBDEs decreases as 

the number of bromine atoms/molecule decreases. With decreasing bromination, 

diphenyl ethers have a lessened overlap with absorption spectra of wavelengths >290 nm, 

which is the range of the solar spectrum at ground level, and thus it is expected that they 

would be less susceptible to decay than decaBDE. Since lower brominated PBDEs 

demonstrate slower rates of phototransformation, it could be speculated that in a system 

with a single input of decaBDE, a steady-state build-up of lower brominated PBDEs 

could occur. However, some level of continued photodegradation down to the biphenyl 

ether could also likely occur. The magnitude of the build-up would be difficult to predict 

since the formation and subsequent degradation of PBDEs depends on their relative rates 

of formation and degradation in the environment, both of which are not known. In 

addition, the magnitude of accumulation would depend on decaBDE loading rates to the 

environment, which are also difficult to predict (United Kingdom 2004). 

In order to evaluate whether these decaBDE transformation products have the potential to 

bioaccumulate or biomagnify in food webs, the BAF and BMF models from Section 3.1 

will be used (see Appendix D for model input values). This analysis assumes that the 

transformations of decaBDE observed in the laboratory studies discussed in this report 

are also occurring in the natural environment; however, it is acknowledged that this is a 

subject of uncertainty. As noted earlier, in the absence of metabolism, the BAFs of hexato nonaBDEs are predicted to exceed 5000. When metabolic transformation is considered 

(a more realistic scenario), BAF estimates for hexaBDEs to nonaBDEs still exceed 5000, 

but are orders of magnitude lower. 

BAF estimates for brominated dibenzofuran products were also undertaken. For these 

transformation products, there was no information that would allow for estimation of kM, 

and a value equal to that estimated for decaBDE was chosen for illustrative purposes. It is 

likely that PBDFs would undergo further metabolism; however, the actual rates of 

metabolism are unknown. Although the chosen value and resulting BAFs represent 

reasonable hypothetical predictions, it is important to acknowledge that these predictions 

are uncertain. Further detail of log Kow and kM estimates are provided in Appendix D. In 

the absence of metabolism, the BAFs for penta- and hexaBDFs are predicted to exceed 

5000 for the middle trophic level fish, while BAF predictions for tri- and tetraBDFs are 
94

below 5000. With consideration given to metabolic transformation (a more realistic 

scenario), the status of the BDFs with respect to the BAF criterion of 5000 remains 

unchanged, but values are somewhat lower overall. 

In the absence of metabolism, the BMFs of hexa- to nonaBDEs are predicted to be very 

high, ranging from ~130 to ~150. However, when metabolic transformation (a more 

realistic scenario) is considered, the predicted BMFs for hexa- to nonaBDEs ranged from 

5 to 6. The metabolism-corrected BMFs exceeded 1 for all metabolites, primarily due to a 

relatively high dietary assimilation efficiency calculated for these substances. For the 

BDF products, a similar BMF of 120 to 150 for penta- and hexaBDFs was predicted, 

while BMFs predictions for tri- and tetraBDFs were 8 and 45, respectively. When 

corrected for metabolism, the BMF values ranged from 4 to 6 (a similar kM value was 

used for all BDFs). 

The modelling analysis of potential environmental transformation products of decaBDE 

predicted BAFs that predominantly exceed 5000 and BMFs exceeding 1 for practically 

all of the metabolites, because of relatively high dietary assimilation efficiency and 

optimal logKow and logKoa values for bioaccumulation via passive diffusion. This 

suggests concerns that the transformation products formed in the environment as a result 

of the photodegradation, biodegradation, and possibly abiotic degradation, of decaBDE, 

are potentially bioaccumulative and biomagnifying in food webs, and may result in 

increased exposure and risk to upper trophic level organisms. 

 

4 Summary and Conclusions 

The existing evidence for the bioaccumulation of decaBDE does not support a conclusion 

of “bioaccumulative” as set out in the Persistence and Bioaccumulation Regulations. 

While most available data show that decaBDE has limited potential to bioaccumulate or 

biomagnify in the environment, some evidence suggests a higher BAF than previously 

considered for decaBDE, and some new data suggest possible biomagnification. Some 

studies also show that levels of decaBDE are steadily rising in some biota, and in some 

cases, measured concentrations have reached levels that can be subjectively described as 

high. The modelling undertaken to support this evaluation, however, shows uncertainty 

associated with metabolism in fish as model-predicted aquatic BAFs range from below 

the 5000 criterion to well above 5000. Predicted terrestrial carnivore BMF values also 

range from below 1 to greater than 1 depending on the rate of metabolism assumed. 

Although less relevant than BAF or BMF, experimental BCF measures are below the 

5000 criterion. The substance is shown to be increasing in concentrations in some 

wildlife species, and some data suggest that decaBDE has reached concentrations in some 

organisms interpreted to be high. Potential factors limiting the bioaccumulation potential 

of decaBDE include low assimilation efficiency and/or metabolic transformation. 

There is still some uncertainty respecting rates and pathways for metabolic 

transformation organisms. Debromination has been indicated in both mammalian and fish 

studies, but the amounts of debrominated PBDE product formed is typically very low 
95

(e.g., representing less than 1%, up to a few percent of the total dose of decaBDE 

depending on the study). Formation of nona-, octa- and heptaBDEs has been noted for 

mammals. In addition to these congener groups, hexa- and pentaBDEs have been noted 

as bioformed in fish. However, the interpretation of results from metabolic 

transformation studies is at times complicated by incomplete characterization of 

impurities in the dosed material and/or dosage of many PBDEs concurrently. Some 

rodent studies have made inferences, based on mass balance evaluations, that rates of 

transformation may be higher, with one study suggesting that approximately 45% of the 

total dose of decaBDE was unaccounted for and may have been metabolized to other 

compounds (like hydroxylated and hydroxymethoxylated PBDEs) and/or bound as 

inextricable residues. 

With respect to chemical transformation in the environment, this review supports the 

findings of the Ecological Screening Assessment of PBDEs (Environment Canada 2006a, 

2006b), which identified photodegradation and biodegradation as likely mechanisms for 

transformation in the environment. This review also identifies various new studies that 

quantify rates of degradation and propose chemical transformation pathways. Together, 

the new and existing studies provide evidence which makes it plausible to conclude that 

decaBDE is transforming in the environment. 

This evaluation found that decaBDE that is sorbed to dry minerals and particulates 

appears to undergo relatively rapid phototransformation in the presence of sunlight. In 

addition, decaBDE sorbed to solids may be subject to biodegradation; however, it appears 

that this process is occurring at a much slower rate than photodegradation. While 

photodegradation on solid phases can occur at significant rates, only a very small fraction 

of the total decaBDE reservoir in the environment (e.g., that fraction adsorbed to 

particulates or on solid surfaces) that has contact with sunlight would be susceptible to 

photodegradation. Based on fugacity modelling, < 3.4% of decaBDE in the environment 

is expected to be associated with bulk air or bulk water phases with potential exposure to 

sunlight. DecaBDE sorbed to solid surfaces (both anthropogenic and natural) could also 

be exposed to sunlight. Within these phases, light attenuation and matrix shielding would 

affect overall exposure to sunlight and potential for photodegradation. While most 

decaBDE in the environment would partition to sediment and soils (fugacity modelling 

predicts > 96%), biodegradation has been shown to be a very slow process, with halflives on the scale of a few years to several decades. Some studies have also not shown 

any obvious transformation in sediments or soil even after several decades. Hence, it is 

apparent that evidence of transformation in the environment could be very subtle and 

shielded by existing PBDE congener patterns dominated by the commercial PBDE 

products. As well, the infrequent historic analysis of higher brominated PBDEs like octaand nonaBDEs could make it difficult to detect or confirm transformation. 

Based on the available laboratory studies with conditions relevant to the environment, it 

is reasonable to expect that decaBDE may be transformed in the environment, leading to 

the formation of lower brominated PBDEs, PBDFs and other unknown products. A 

number of studies have now shown that decaBDE may transform by photodegradation or 

biodegradation to nona -, octa-, hepta- and hexaBDEs. One study also indicates trace 
96

formations of penta- and tetraBDEs by photodegradation, while another study also 

indicates biodegradation to penta- and tetraBDEs under enhanced biodegradation 

conditions. In addition, photodegradation of decaBDE has also been shown to form tri- to 

octaBDFs as well as unidentified products. 

Modelling of BAFs and BMFs played an important supplemental role in this review and 

was used to suggest whether decaBDE and its transformation products may be 

bioaccumulative or biomagnify in food chains. 

The metabolism-corrected model predicted aquatic BAFs for decaBDE ranged from 

below the 5000 criterion of the Persistence and Bioaccumulation Regulations to well 

above 5000. The predictions demonstrate the uncertainty associated with the metabolism 

potential of decaBDE in fish and log Kow determinations for this substance. Given the 

exceptionally low water solubility limit of decaBDE, it is not expected that this substance 

will be appreciably taken up from the water phase by aquatic organisms. Although less 

relevant than BAF or BMF, experimental BCF measures are below the 5000 criterion for 

decaBDE. With consideration given to decaBDE metabolism, terrestrial BMF predictions 

(based on a wolf food chain) indicate lack of or a low level of biomagnification of 

decaBDE. 

In the absence of metabolism, the BAFs of all potential metabolites and transformation 

products of decaBDE were predicted to exceed 5000. When assumptions were made 

respecting metabolic transformation (a more realistic scenario), almost all proposed 

transformation products still yielded BAFs exceeding 5000. In the absence of 

metabolism, the BMFs of potential transformation products are also predicted to be very 

high; however, when assumptions were made respecting metabolic transformation, the 

predicted BMFs are significantly lower, but are still all greater than one. This analysis 

suggests that many possible decaBDE metabolites/transformation products could be 

highly bioaccumulative and some metabolites may have the capacity to biomagnify in 

food chains. 

Overall, this review confirms that, based on the reviewed materials published up to 

August 25, 2009, decaBDE is not shown to meet bioaccumulation criteria as defined 

under the Persistence and Bioaccumulation Regulations under CEPA 1999. However, 

some studies show that levels of decaBDE are steadily rising in some biota, and in some 

cases, measured concentrations are considered high. In addition, some equivocal 

evidence suggests potential biomagnification in food chains. Although uncertainties 

remain, it is reasonable to conclude that decaBDE may also contribute to the formation of 

bioaccumulative and/or potentially bioaccumulative transformation products, such as 

lower brominated BDEs, in organisms and the environment. 

4.1 Consideration Regarding Related Products 

While this review has focused on decaBDE, its analyses and conclusions are relevant to 

alternative flame retardants with similar chemical structures and use patterns. For 

instance, decabromodiphenyl ethane (or decaBD ethane)—(1,2-
97

bis(pentabromodiphenyl)ethane; 1,1”-(ethane-, 1,2-diyl) bis[pentabromobenzene])—is a 

replacement for the decaBDE commercial product, having the same or similar 

applications. Both are additive flame retardants used in HIPS and in textiles used in the 

manufacture of television cabinets, cable insulation and adhesives (Kierkegaard 2007). In 

Japan, Watanabe and Sakai (2003) have shown that there has been a clear shift in 

consumption away from the decaBDE commercial product to decaBD ethane. 

The only structural difference between decaBD ethane and decaBDE is the carbon bond 

between the aromatic rings of decaBD ethane (for decaBDE, the aromatic rings are linked 

with an oxygen atom) (see Appendix F). Based on structural similarities, the two 

substances likely have similar physical-chemical properties, characteristics of persistence, 

transformation, and accumulation in organisms (Kierkegaard 2007). 

DecaBD ethane has been identified in sewage sludge, both in Canada (Konstantinov et al. 

2006) and in Spain (Eljarrat et al. 2005), has been measured in walleye and burbot in 

Lake Winnipeg (Law et al. 2006), and detected in herring gull eggs from the Great Lakes 

area of Canada (Letcher et al. 2007). The United Kingdom Environment Agency recently 

published a detailed risk assessment for this substance (United Kingdom 2007b). While 

direct risks resulting from toxic effects of this substance were considered low, concerns 

were identified over this substance’s potential to accumulate in wildlife and 

transformation to other chemical products. The Agency also identified a need for further 

work on decaBD ethane to confirm the findings of their assessment, particularly to 

provide more reliable measures of this substance’s potential to bioaccumulate and 

degrade in the environment. 

Based on concerns expressed for decaBDE in this State of Science report, the similarity 

in properties between decaBDE and decaBD ethane, the presence of the decaBD ethane 

in Canadian wildlife, and the potential for decaBD ethane to be used as a large-scale 

replacement for decaBDE, there is a need to further understand the potential risks from 

decaBD ethane in the environment and its capacity to accumulate in wildlife and 

transform to bioaccumulative products. Understanding the risk from alternatives will help 

to ensure that substitutions of flame retardants are made on an informed basis. 

 

 

5 Key Uncertainties and Data Gaps 

The data for bioaccumulation, biomagnification and trophic magnification of decaBDE 

are limited, especially for top predators in natural settings. Additional ratio-based 

assessments of decaBDE in top predators in natural settings, particularly in terrestrial 

systems, would provide better understanding of the bioaccumulation and 

biomagnification of decaBDE. 

Overall, the relative contribution of environmental transformation of decaBDE to the 

total loadings of lower brominated PBDEs (i.e., having nine or fewer bromine atoms) in 

the environment is not known. Studies which trace and/or “fingerprint” the source of 

lower brominated PBDEs in the environment would help clarify the significance of in 
98

vivo and environmental debromination of decaBDE, compared to that resulting from the 

use of the Penta- and OctaBDE formulations. 

It may be beneficial to have BAF or BMF determinations for decaBDE on the total 

burden of parent decaBDE and the metabolites that result from accumulation and 

transformation of the parent chemical (e.g., lower BDEs and PBDFs). In most of the 

studies conducted to date, this has been difficult to accomplish because (i) the lesser 

debrominated products were also present in the exposure medium (i.e., food or water) and 

their presence in tissue could thus have resulted from both accumulation and 

bioformation, or (ii) full identification and quantification of most metabolites has not 

been conducted. 

Model-estimated BCFs, BAFs and BMFs presented in this report for decaBDE and 

known potential transformation products vary. These relied heavily on logKow logKoa ED

and kM values, which would benefit from refinement in order to improve the precision of 

the model predictions. 

BAFs, BMFs and other indicators of bioaccumulation or biomagnification have not been 

measured for some of the known metabolites and transformation products of decaBDE, 

especially their hydroxylated and methoxyhydroxylated forms. Further study of the 

behaviour of these products would be useful. 
99

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110

Appendix A – Biota Concentration Data

Study Authors Location Date Species Tissue Sampling N Not Detected L

BIRDS OF PREY

Chen et al. 2007 China 2004–2006 Common kestrel Falco tinnunculus Muscle 6 21% 

Chen et al. 2007 China 2004–2006 Common kestrel Falco tinnunculus Liver 6 21% 

Chen et al. 2007 China 2004–2006 Common kestrel Falco tinnunculus Kidney 6 21% Chen et al. 2007 China 2004–2006 Sparrowhawk Accipiter nisus Muscle 11 21% 

Chen et al. 2007 China 2004–2006 Sparrowhawk Accipiter nisus Liver 11 21% 

Chen et al. 2007 China 2004–2006 Sparrowhawk Accipiter nisus Kidney 11 21% Chen et al. 2007 China 2004–2006 Japanese sparrowhawk Accipiter gularis Muscle 6 21% 

Chen et al. 2007 China 2004–2006 Japanese sparrowhawk Accipiter gularis Liver 6 21% 

Chen et al. 2007 China 2004–2006 Japanese sparrowhawk Accipiter gularis Kidney 6 21% Chen et al. 2007 China 2004–2006 Little owl Athene noctua Muscle 6 21% 

Chen et al. 2007 China 2004–2006 Little owl Athene noctua Liver 6 21% 

Chen et al. 2007 China 2004–2006 Little owl Athene noctua Kidney 6 21% 

Chen et al. 2007 China 2004–2006 Scops owl Otus sunia Muscle 6 21% 

Chen et al. 2007 China 2004–2006 Scops owl Otus sunia Liver 6 21% 

Chen et al. 2007 China 2004–2006 Scops owl Kidney 6 21% 
111

Study Authors Location Date Species Tissue Sampling N Not Detected L

Otus sunia

Chen et al. 2007 China 2004–2006 Long-eared owl Asio otus Muscle 6 21% 

Chen et al. 2007 China 2004–2006 Long-eared owl Asio otus Liver 6 21% 

Chen et al. 2007 China 2004–2006 Long-eared owl Asio otus Kidney 6 21% Chen et al. 2007 China 2004–2006 Upland buzzard Buteo hemilasius Muscle 3 21% 

Chen et al. 2007 China 2004–2006 Upland buzzard Buteo hemilasius Liver 3 21% 

Chen et al. 2007 China 2004–2006 Upland buzzard Buteo hemilasius Kidney 3 21% 

Chen et al. 2007 China 2004–2006 Common buzzard Buteo buteo Muscle 3 21% 

Chen et al. 2007 China 2004–2006 Common buzzard Buteo buteo Liver 3 21% Chen et al. 2007 China 2004–2006 Common buzzard Buteo buteo Kidney 3 21% Chen et al. 2008 United States, NE 1996–2006 Peregrine falcon Falco peregrinus Eggs 114 0 

de Boer et al. 2004 United Kingdom Not given Peregrine falcon Falco peregrinus Liver 6 2 of 6 

de Boer et al. 2004 United Kingdom Not given Peregrine falcon Falco peregrinus Muscle 5 None 

de Boer et al. 2004 United Kingdom Not given Peregrine falcon Falco peregrinus Egg 6 4 of 6 de Boer et al. 2004 Sweden Not given Peregrine falcon Falco peregrinus Egg 20 9 of 20 de Boer et al. 2004 Sweden Not given Sparrowhawk Accipiter nisus Liver 4 All < 

de Boer et al. 2004 Sweden Not given Sparrowhawk Accipiter nisus Muscle 5 None 

de Boer et al. 2004 Sweden Not given Sparrowhawk Accipiter nisus Egg 5 2 of 5 
112

Study Authors Location Date Species Tissue Sampling N Not Detected Lde Boer et al. 2004 Sweden Not given Kestrel Falco tinnunculus Liver 5 3 of 5 de Boer et al. 2004 Sweden Not given Kestrel Falco tinnunculus Muscle 5 4 of 5 

de Boer et al. 2004 Sweden Not given Barn owl Tyto alba Liver 5 2 of 5 

de Boer et al. 2004 Sweden Not given Barn owl Tyto alba Muscle 5 4 of 5 

de Boer et al. 2004 Sweden Not given Barn owl Tyto alba Egg 4 1 of 4 de Boer et al. 2004 Sweden Not given Montague's harrier Circus pygargus Egg 4 1 of 4 

de Boer et al. 2004 Sweden Not given Marsh harrier Circus aeruginosus Egg 2 All de Boer et al. 2004 Sweden Not given Sea eagle Egg 1 None de Boer et al. 2004 Sweden Not given Osprey Pandion haliaetus Egg 3 All de Boer et al. 2004 Sweden Not given Golden eagle Aquila chrysaetos Egg 5 All de Boer et al. 2004 Sweden Not given Merlin Falco columbarius Egg 2 1 of 2 

Herzke et al. 2005 Northern Norway 1993–2000 Peregrine falcon Falco peregrinus Egg 9 

Herzke et al. 2005 Northern Norway 1992–2002 Golden eagle Aquila chrysaetos Egg 15 Herzke et al. 2005 Northern Norway 1995–2000 Merlin Falco columbarius Egg 9 Jaspers et al. 2006 Belgium 2003–2004 Barn owl Tyto alba Liver 7 Some 

Jaspers et al. 2006 Belgium 2003–2004 Barn owl Tyto alba Muscle 7 Most 

Jaspers et al. 2006 Belgium 2003–2004 long-eared owl Asio otus Liver 6 Most 

Jaspers et al. 2006 Belgium 2003–2004 Long-eared owl Asio otus Muscle 6 All 

Jaspers et al. 2006 Belgium 2003–2004 Buzzard Liver 16 All 
113

Study Authors Location Date Species Tissue Sampling N Not Detected L

Buteo buteo 

Jaspers et al. 2006 Belgium 2003–2004 Buzzard Buteo buteo Muscle 16 All 

Jaspers et al. 2006 Belgium 2003–2004 Sparrowhawk Accipiter nisus Liver 5 Some 

Jaspers et al. 2006 Belgium 2003–2004 Sparrowhawk Accipiter nisus Muscle 5 All 

Jaspers et al. 2006 Belgium 2003–2004 Kestrel Falco tinnunculus Liver 5 Most 

Jaspers et al. 2006 Belgium 2003–2004 Kestrel Falco tinnunculus Muscle 5 All 

Kunisue et al. 2008 Japan (Coastal) 1994 Steller’s sea-eagle Haliaeetus pelagicus Muscle 1 1 Kunisue et al. 2008 Japan (Inland) 1994–1995 Golden eagle Aquila chrysaetos Muscle 2 1 

Kunisue et al. 2008 Japan (Inland) 1997 Goshawk Accipiter gentilis Muscle 1 0 Lindberg et al. 2004 Captive Sweden 1988–1999 Peregrine falcon Falco peregrinus Eggs 4 2 

Lindberg et al. 2004 Southern Sweden 1992–1999 Peregrine falcon Falco peregrinus Eggs 8 1 < Lindberg et al. 2004 Northern Sweden Peregrine falcon Falco peregrinus Eggs 8 0 2Lindberg et al. 2004 Northern Sweden 1991–1999 Peregrine falcon Falco peregrinus Eggs 18 

Potter et al. 2009 

United States, 

Maryland and 

Virginia 

1993–2002 Peregrine falcon Falco peregrinus Eggs 23 0 

Voorspoels et al. 2006b Belgium 2001–2003 Buzzard Buteo buteo Brain 29 All 

Voorspoels et al. Belgium 2001–2003 Buzzard Fat 16 All 
114

Study Authors Location Date Species Tissue Sampling N Not Detected L2006b Buteo buteoVoorspoels et al. 2006b Belgium 2001–2003 Buzzard Buteo buteo Liver 3 Most 1

Voorspoels et al. 2006b Belgium 2001–2003 Buzzard Buteo buteo Muscle 29 All 

Voorspoels et al. 2006b Belgium 2001–2003 Buzzard Buteo buteo Serum 20 Some 

Voorspoels et al. 2006b Belgium 2001–2003 Owls Brain 8 All Voorspoels et al. 2006b Belgium 2001–2003 Owls Fat 1 All Voorspoels et al. 2006b Belgium 2001–2003 Owls Liver 8 All Voorspoels et al. 2006b Belgium 2001–2003 Owls Muscle 8 All Voorspoels et al. 2006b Belgium 2001–2003 Owls Serum 3 Most Voorspoels et al. 2006b Belgium 2001–2003 Sparrowhawk Accipiter nisus Brain 7 All 

Voorspoels et al. 2006b Belgium 2001–2003 Sparrowhawk Accipiter nisus Fat 2 All 

Voorspoels et al. 2006b Belgium 2001–2003 Sparrowhawk Accipiter nisus Liver 7 Most 

Voorspoels et al. 2006b Belgium 2001–2003 Sparrowhawk Accipiter nisus Muscle 7 All 

Voorspoels et al. 2006b Belgium 2001–2003 Sparrowhawk Accipiter nisus Serum 2 None 

Vorkamp et al. 2005 Greenland 1986–2003 Peregrine falcon Falco peregrinus Eggs 37 0 3

de Boer et al. 2004 Sweden Not given Red kite Milvus milvus Eggs 4 3 of 4 

FRESHWATER FISH AND INVERTEBRATESHale et al. 2002 United States; mid- Atlantic region 2000 Bluegill Lepomis macrochirus Muscle 
115

Study Authors Location Date Species Tissue Sampling N Not Detected LHale et al. 2002 United States; mid- Atlantic region 2000 Bluegill Lepomis macrochirus Remainder of body Eljarrat et al. 2007 Spain Nov. 2004, Nov. 2005 Carp Cyprinus carpio Muscle 2 None 

Barbel 

Barbus graellsii Muscle 13 1 of 13 

de Boer et al. 2003 The Netherlands 1999 Freshwater mussels Dreissena polymorpha 

Dodder et al. 2002 

Northeastern United 

States; Lake 

Superior 

1994 Smelt Osmeru mordax Whole body 

Dodder et al. 2002 Lake Ontario 1994 Smelt Osmeru mordax Whole body 

Dodder et al. 2002 Lake of the Ozarks, MO 1999 White crappie; bluegill Pomoxis annularis; Lepomis macrochirus Whole body 

Dodder et al. 2002 Hadley Lake, IN. 1999 

White crappie, bluegill 

Pomoxis annularis; Lepomis 

macrochirus

Whole body 

Dodder et al. 2002 Hadley Lake, IN. 1999 

White crappie, bluegill 

Pomoxis annularis; 

Lepomis macrochirus

Muscle 

Guo et al. 2009 Pearl River Delta, S. China Bighead carp (farmed) Aristichthys nobilis Skin 8 

Guo et al. 2009 Pearl River Delta, S. China Bighead carp (farmed) Aristichthys nobilis Gills 8 Guo et al. 2009 Pearl River Delta, S. China Bighead carp (farmed) Aristichthys nobilis GIT 8 

Guo et al. 2009 Pearl River Delta, S. China Bighead carp (farmed) Aristichthys nobilis Liver 8 

Guo et al. 2009 Pearl River Delta, S. China Bighead carp (farmed) Aristichthys nobilis Muscle 8 

Guo et al. 2009 Pearl River Delta, S. China Mandarin fish (farmed) Siniperca chuatsi Skin 8 

Guo et al. 2009 Pearl River Delta, S. China Mandarin fish (farmed) Siniperca chuatsi Gills 8 Guo et al. 2009 Pearl River Delta, Mandarin fish (farmed) GIT 8 
116

Study Authors Location Date Species Tissue Sampling N Not Detected L

S. China Siniperca chuatsi

Guo et al. 2009 Pearl River Delta, S. China Mandarin fish (farmed) Siniperca chuatsi Liver 8 

Guo et al. 2009 Pearl River Delta, S. China Mandarin fish (farmed) Siniperca chuatsi Muscle 8 Guo et al. 2009 Pearl River Delta, S. China Northern snakehead (farmed) Nemipterus virgatus Skin 8 

Guo et al. 2009 Pearl River Delta, S. China Northern snakehead (farmed) Nemipterus virgatus Gills 8 

Guo et al. 2009 Pearl River Delta, S. China Northern snakehead (farmed) Nemipterus virgatus GIT 8 Guo et al. 2009 Pearl River Delta, S. China Northern snakehead (farmed) Nemipterus virgatus Liver 8 Guo et al. 2009 Pearl River Delta, S. China Northern snakehead (farmed) Nemipterus virgatus Muscle 8 

Ismail et al. 2009 Lake Ontario 1979–2004 Lake trout Salvelinus namaycush Whole body 29 (4 or 5 per year) None 2.3 (

Sellström et al. 1998 Sweden 1995 Pike Esox lucius Muscle 

Sellström et al. 1998 Sweden 1995 Pike Esox lucius Muscle not d

Sellström et al. 1998 Sweden 1995 Pike Esox lucius Muscle not d

Wang et al. 2007 China Sept.–Dec. 2006 Lichen Whole 1 None Wang et al. 2008 China Sept.–Dec. 2006 March brown Limnodrilus hoffmeisteri Whole 1 None 

Wang et al. 2008 China Sept.–Dec. 2006 Coccid Whole 1 None 

Wang et al. 2008 China Sept.–Dec. 2006 

Zooplankton 

Monia rectirostris, Monia micrura, 

Monia macrocopa

Whole 

1 None 

Wang et al. 2008 China Sept.–Dec. 2006 Common carp Cyprinus carpio Whole 4 75% 

Wang et al. 2008 China Sept.–Dec. 2006 Java tilapia Tilapia nilotica Whole 3 100% 

Wang et al. 2008 China Sept.–Dec. 2006 Leather catfish Silurus meridionalis Whole 4 Some 
117

Study Authors Location Date Species Tissue Sampling N Not Detected LWang et al. 2008 China Sept.–Dec. 2006 Crusian carp Carassius auratus Whole 5 Some Wang et al. 2008 China Sept.–Dec. 2006 Chinese softshell turtle Chinemys reevesii Whole 3 Some 

Xiang et al. 2007 China Apr.–Aug. 2004 Large yellow croaker Pseudosiaena crocea Muscle 13 Some 

Xiang et al. 2007 China Apr.–Aug. 2004 Silvery pomfret Platycephalus argenteus Muscle 10 Some 

Xiang et al. 2007 China Apr.–Aug. 2004 Flathead fish Platycephalus indicus Muscle 17 Some 

Xiang et al. 2007 China Apr.–Aug. 2004 Robust tongue fish Cynoglossus robustus Muscle 8 Some 

Xiang et al. 2007 China Apr.–Aug. 2004 Bombay duck Harpodon nehereus Muscle 9 Some 

Xiang et al. 2007 China Apr.–Aug. 2004 Jinga shrimp Metapenaeus affinis Soft tissue 10 Some 

Xiang et al. 2007 China Apr.–Aug. 2004 Greasy-back shrimp Metapenaeus crocea Soft tissue 10 Some 

Xiang et al. 2007 China Apr.–Aug. 2004 Mantis shrimp Oratosquilla oratoria Soft tissue 9 Some 

MAMMAL – HERBIVORE

Lichota et al. 2004 Vancouver Island, Canada 2001 Marmot Marmota vancouverensis Fat 

Voorspoels et al. 2007 Flanders, Belgium 2001 Wood mice Apodemus sylvaticus Liver and muscle 12 All 

Voorspoels et al. 2007 Flanders, Belgium 2001 Bank voles Clethrionomys glareolus Liver and muscle 12 All 

MAMMAL – PREDATORMariussen et al. 2004 Norway 1995 Moose Alces alces Liver Not reported 

Christensen et al. 2005 

Coastal British 

Columbia 2003 Grizzly bear Ursus arctos horribilis Fat or muscle 6 Not reported 

Christensen et al. Interior British 2003 Grizzly bear Fat or 6 Not reported 
118

Study Authors Location Date Species Tissue Sampling N Not Detected L2005 Columbia Ursus arctos horribilis muscle Sørmo et al. 2006 Svalbard 2003 Polar bear Ursus maritimus Blubber 4 0 0.0

Sørmo et al. 2006 Svalbard 2003 Polar bear Ursus maritimus Blubber 4 0 

Mariussen et al. 2004 Norway 1993–1994 Lynx Lynx lynx Liver 7 

Muir et al. 2006 Svalbard Polar bear Ursus maritimus Adipose 15 

Muir et al. 2006 Svalbard Polar bear Ursus maritimus Adipose 44 Skaare 2004 Svalbard 2001–2003 Polar bear Ursus maritimus Adipose Verrault et al. 2005 Svalbard, Norway 2002 Polar bear (female) Ursus maritimus Plasma 15 93% Voorspoels et al. 2006a Belgium 2003–2004 Red fox Vulpes vulpes Liver 30 60% 

Voorspoels et al. 2006a Belgium 2003–2004 Red fox Vulpes vulpes Muscle 33 79% 

Voorspoels et al. 2006a Belgium 2003–2004 Red fox Vulpes vulpes Adipose 27 85% 

MARINE FISHAkutsu et al. 2001 Inland Sea of Seto, Japan 1998 Conger eel 

Akutsu et al. 2001 Inland Sea of Seto, Japan 1998 Flounder Paralichthys olivaceus 

Akutsu et al. 2001 Inland Sea of Seto, Japan 1998 Gray mullet Mugil cephalus 

Akutsu et al. 2001 Inland Sea of Seto, Japan 1998 Horse mackerel Akutsu et al. 2001 Inland Sea of Seto, Japan 1998 Red sea bream Pagrus major 

Akutsu et al. 2001 Inland Sea of Seto, Japan 1998 Sea bass Lateolabrax japonicus 
119

Study Authors Location Date Species Tissue Sampling N Not Detected LBurreau et al. 2004 Lumparn Estuary Not Given Roach Rutilus rutilus Muscle 8 5 Burreau et al. 2004 Lumparn Estuary Not Given Perch Perca fluviatilis Muscle 33 21 

Burreau et al. 2004 Lumparn Estuary Not Given Pike Esox lucius Muscle 25 21 

Burreau et al. 2006 Baltic Sea 1998 Sprat Sprattus sprattus Muscle 6 3 Burreau et al. 2006 Baltic Sea 1998 Herring Clupea harengus Muscle 5 1 

Burreau et al. 2006 Baltic Sea 1998 Salmon Salmo salar Muscle 10 7 

Burreau et al. 2006 North Atlantic Ocean 1999 Small herring Clupea harengus Muscle 6 0 

Burreau et al. 2006 North Atlantic Ocean 1999 Large herring Clupea harengus Muscle 6 3 Burreau et al. 2006 North Atlantic Ocean 1999 Atlantic salmon Salmo salar Muscle 10 0 

de Boer et al. 1998 Dutch Coast 1995 Mackerel Scomber scombrus Muscle 

de Boer et al. 2001 North Sea 1999 Herring Clupea harengus Liver < 

de Boer et al. 2001 North Sea 1999 Herring Clupea harengus Filet < 

de Boer et al. 2001 North Sea 1999 Herring Clupea harengus Milt and eggs < 

de Boer et al. 2001 North Sea 1999 Cod Liver < 

de Boer et al. 2001 North Sea 1999 Cod Fillet < 

de Boer et al. 2001 North Sea 1999 Whiting Merlangius merlangius Liver < 

de Boer et al. 2001 North Sea 1999 Whiting Merlangius merlangius Fillet < 

de Boer et al. 2001 Tees Estuary, UK 1999 Sprat Sprattus sprattus Whole body 
120

Study Authors Location Date Species Tissue Sampling N Not Detected Lde Boer et al. 2001 Tees Estuary, UK 1999 Whiting Merlangius merlangius Whole body de Boer et al. 2001 Tees Estuary, UK 1999 Whiting Merlangius merlangius Muscle < 

de Boer et al. 2001 Tees Estuary, UK 1999 Whiting Merlangius merlangius Liver de Boer et al. 2001 Tees Estuary, UK 1999 Flounder Paralichthys olivaceus Liver 

de Boer et al. 2001 Western Scheldt, The Netherlands 1999 Greater sandeel Hyperoplus lanceolatus < 

de Boer et al. 2003 The Netherlands 1999 Flounder Paralichthys olivaceus 

de Boer et al. 2003 The Netherlands 1999 Bream 

Fjeld et al. 2004 Svolvaer, Lofoten, Norway 2002 Atlantic cod Gadus morhua Liver 21 

Fjeld et al. 2004 Varangerfjorden, Norway 1998, 2002 Atlantic cod Gadus morhua Liver 21 Guo et al. 2009 Pearl River Delta, S. China Crimson snapper (farmed) Lutjanus erythropterus Skin 8 

Guo et al. 2009 Pearl River Delta, S. China Crimson snapper (farmed) Lutjanus erythropterus Gills 8 

Guo et al. 2009 Pearl River Delta, S. China Crimson snapper (farmed) Lutjanus erythropterus GIT 8 

Guo et al. 2009 Pearl River Delta, S. China Crimson snapper (farmed) Lutjanus erythropterus Liver 8 

Guo et al. 2009 Pearl River Delta, S. China Crimson snapper (farmed) Lutjanus erythropterus Muscle 8 Guo et al. 2009 Pearl River Delta, S. China Golden thread Nemipterus virgatus Skin 8 

Guo et al. 2009 Pearl River Delta, S. China Golden thread Nemipterus virgatus Gills 8 

Guo et al. 2009 Pearl River Delta, S. China Golden thread Nemipterus virgatus GIT 8 

Guo et al. 2009 Pearl River Delta, S. China Golden thread Nemipterus virgatus Liver 8 
121

Study Authors Location Date Species Tissue Sampling N Not Detected LGuo et al. 2009 Pearl River Delta, S. China Golden thread Nemipterus virgatus Muscle 8 Sørmo et al. 2006 Svalbard 2003 Polar cod Boreogadus saida Whole body 7 0 0.0

Sørmo et al. 2006 Svalbard 2003 Polar cod Boreogadus saida Whole body 7 0 

Johnson-Restropo 2005 Coastal Florida 2004 Silver perch Bairdiella chrysoura Muscle 6 All 

Johnson-Restropo 2005 Coastal Florida 2004 Striped mullet Mugil cephalus Muscle 6 All 

Johnson-Restropo 2005 Coastal Florida 2004 Spotted seatrout Cynoscion nebulosus Muscle 7 All 

Johnson-Restropo 2005 Coastal Florida 2004 Red drum Sciaenops ocellatus Muscle 11 All 

Johnson-Restropo 2005 Coastal Florida 2004 Hardhead catfish Arius felis Muscle 8 

Johnson-Restropo 2005 Coastal Florida 2004 Atlantic stingray Dasyatis Sabina Muscle 7 

Johnson-Restropo 2005 Coastal Florida 2004 Spiny dogfish Squalus acanthias Muscle 5 

Johnson-Restropo 2005 Coastal Florida 2004 Atlantic sharpnose shark Rhizoprionodon terraenovae Muscle 5 

Johnson-Restropo 2005 Coastal Florida 1993–1994 Bull shark Carcharhinus leucas Muscle 6 

Johnson-Restropo 2005 Coastal Florida 2002–2004 Bull shark Carcharhinus leucas Muscle 7 Watanabe et al. 1987 Japan 1981–1985 Mullet Mugilidae 4 

Watanabe et al. 1987 Japan 1981–1985 Goby 2 Watanabe et al. 1987 Japan 1981–1985 Sardine 2 Watanabe et al. 1987 Japan 1981–1985 Sea bass 1 Watanabe et al. 1987 Japan 1981–1985 Horse mackerel 1 
122

Study Authors Location Date Species Tissue Sampling N Not Detected LWatanabe et al. 1987 Japan 1981–1985 Mackerel Scomber scombrus 1 Watanabe et al. 1987 Japan 1981–1985 Hairtail 1 MARINE INVERTEBRATES

Burreau et al. 2006 Baltic Sea 1999 Zooplankton Calanoid sp. Muscle 3 None 

de Boer et al. 2001 North Sea 1999 Starfish Asteroidea Pyloric caeca 

de Boer et al. 2001 North Sea 1999 Hermit crab Abdomen 

de Boer et al. 2001 North Sea 1999 Male whelk Haustellum brandaris Whole body < 

de Boer et al. 2001 North Sea 1999 Shrimp Whole body < 

de Boer et al. 2001 Tees Estuary, UK 1999 Shrimp Whole body de Boer et al. 2001 Tees Estuary, UK 1999 Polychaete worm Whole body de Boer et al. 2001 Tees Estuary, UK 1999 Starfish 

Whole body 

and 

tentacle 

 

de Boer et al. 2001 Western Scheldt, The Netherlands 1999 Mysid shrimp de Boer et al. 2001 Western Scheldt, The Netherlands 1999 Copepod de Boer et al. 2001 Western Scheldt, The Netherlands 1999 Gudgeon < de Boer et al. 2003 The Netherlands 1999 Marine mussels Fjeld et al. 2004 Svolvaer, Lofoten, Norway 2002 Blue mussels Mytilus edullis Muscle 20 

Fjeld et al. 2004 Varangerfjorden, Norway 2002 Blue mussels Mytilus edullis Muscle 50 Sørmo et al. 2006 Svalbard 2003 Ice amphipod Gammarus wilkitzkii Whole body 1 (composite of 5) 0 

Sørmo et al. 2006 Svalbard 2003 Ice amphipod Whole body 1 (composite 0 
123

Study Authors Location Date Species Tissue Sampling N Not Detected L

Gammarus wilkitzkii of 5) 

Watanabe et al. 1987 Japan 1981–1985 Mussels Lampsilis radiata 5 4 of 5 MARINE MAMMALSde Boer et al. 1998 Dutch Coast 1995 Minke whale Balaenoptera acutorostrata Blubber 

de Boer et al. 1998 Dutch Coast 1995 Sperm whale Physeter catodon Blubber 

de Boer et al. 1998 Dutch Coast 1995 Sperm whale Physeter catodon Liver 

de Boer et al. 1998 Dutch Coast 1995 White-beaked dolphin Lagenorhynchus albirostris Blubber 

de Boer et al. 1998 Dutch Coast 1995 White-beaked dolphin Lagenorhynchus albirostris Liver 

de Boer et al. 1998 Dutch Coast 1995 Harbour seal Blubber de Boer et al. 1998 Dutch Coast 1995 Harbour seal Liver de Boer et al. 2001 North Sea 1999 Harbour porpoise Liver de Boer et al. 2001 North Sea 1999 Harbour porpoise Blubber 

de Boer et al. 2001 North Sea 1999 Harbour seal Liver 9 de Boer et al. 2001 North Sea 1999 Harbour seal Blubber de Boer et al. 2001 North Sea 1999 White-beaked dolphin Blubber de Boer et al. 2001 North Sea 1999 White-beaked dolphin Kidney de Boer et al. 2001 North Sea 1999 White-beaked dolphin Heart de Boer et al. 2001 North Sea 1999 White-beaked dolphin Lungs de Boer et al. 2001 North Sea 1999 White-beaked dolphin Brain de Boer et al. 2001 North Sea 1999 White-beaked dolphin Liver de Boer et al. 2001 North Sea 1999 Bottlenose dolphin Blubber de Boer et al. 2001 North Sea 1999 Bottlenose dolphin Liver de Boer et al. 2001 North Sea 1999 Bottlenose dolphin Kidney 
124

Study Authors Location Date Species Tissue Sampling N Not Detected Lde Boer et al. 2001 North Sea 1999 Bottlenose dolphin Muscle de Boer et al. 2001 Tees Estuary, UK 1999 Porpoise Blubber Sørmo et al. 2006 Svalbard 2003 Ringed seal Phoca hispida Blubber 6 5 

Sørmo et al. 2006 Svalbard 2003 Ringed seal Phoca hispida Blubber 6 5 Johnson-Restropo 2005 

Western coast of 

Florida 1991–1996 Bottlenose dolphin Tursiops truncatus Blubber 

Johnson-Restropo 2005 

Western coast of 

Florida 2000–2001 Bottlenose dolphin Tursiops truncatus Blubber All 

Johnson-Restropo 2005 

Eastern coast of 

Florida 2001–2004 Bottlenose dolphin Tursiops truncatus Blubber All 

Johnson-Restropo 2005 Coastal Florida ?? Striped dolphin Stenella coeruleoalba Blubber All 

Shaw et al. 2007 (BFR) 

NW Atlantic coast 

(Maine to Long 

Island) 

1991 – 2005 Harbour seal Phoca vitulina Blubber 4 2 of 4 

MARINE/AQUATIC BIRDSde Boer et al. 2001 Tees Estuary, UK 1999 Cormorant Corvus marinus Liver < 

de Boer et al. 2001 Western Scheldt, The Netherlands 1999 Common tern Sterna hirundo Eggs 

de Boer et al. 2001 Western Scheldt, The Netherlands 1999 Common tern Sterna hirundo Eggs 

de Boer et al. 2004 Sweden Not given Great crested grebe Podiceps cristatus Liver 4 3 of 4 

de Boer et al. 2004 Sweden Not given Great crested grebe Podiceps cristatus Muscle 3 2 of 3 

de Boer et al. 2004 Sweden Not given Heron Liver 4 All < de Boer et al. 2004 Sweden Not given Heron Muscle 5 4 of 5 
125

Study Authors Location Date Species Tissue Sampling N Not Detected Lde Boer et al. 2004 Sweden Not given Gannett Morus bassanus Eggs 12 All Gauthier et al. 2008 Canada 

1982, 1987, 

1992, 1995–

2006, April – 

May 

Herring gull 

Larus argentatus Eggs 

10–13 

eggs/site/ 

year pooled 

homogenates 

Not reported 

Jaspers et al. 2006 Belgium 2003–2004 Aquatic birds (heron/grebe) Multiple 12 All Kunisue et al. 2008 Open sea (Asia) 1998–1999 Black-footed albatross Diomedea nigripes Muscle 5 All 

Kunisue et al. 2008 Open sea (Asia) 1998 Laysan albatross Diomedea immutabilis Muscle 5 All 

Kunisue et al. 2008 Open sea (Asia) 1999 Northern fulmar Fulmarus glacialis Muscle 5 0 <0

Kunisue et al. 2008 Japan (Coastal) 1999–2001 Black-tailed gull Larus crassirostris Muscle 5 All 

Lam et al. 2007 China – Hong Kong March – May 2004 Little egret Egretta garzetta Eggs 5 Not reported 2.6 ±

Lam et al. 2007 

China – Hong Kong 

March–May 

2004 Black-crowned night heron 

Nycticorax nycticorax Eggs 5 Not reported 14 

Lam et al. 2007 

China – Xiamen 

March–May 

2004 Little egret 

Egretta garzetta Eggs 5 Not reported 0.89 

Lam et al. 2007 China – Xiamen March–May 

2004 Black-crowned night heron 

Nycticorax nycticorax Eggs 5 Not reported 1.8 ±

Lam et al. 2007 China – Xiamen March–May 

2004 Chinese pond heron 

Ardeola bacchus Eggs 5 Not reported 99 ± 

Lam et al. 2007 China – Xiamen March–May 

2004 

Cattle egret 

Bubulcus ibis Eggs 5 Not reported 28 ± 
126

Study Authors Location Date Species Tissue Sampling N Not Detected LLam et al. 2007 

Quanzhou 

March–May 

2004 Little egret 

Egretta garzetta Eggs 5 Not reported 1.6 

Lam et al 2007 

Quanzhou 

March–May 

2004 Black-crowned night heron 

Nycticorax nycticorax Eggs 5 5 

Verrault et al. 2005 Svalbard, Norway 2004 Glaucous gull (male) Larus hyperboreus Plasma 12 50% 

Verrault et al. 2005 Svalbard, Norway 2004 Glaucous gull (female) Larus hyperboreus Plasma 15 64% Verreault et al. 2004 Bjoroya, Norway 2002–2004 Glaucous gull Larus hyperboreus Plasma 89 Verreault et al. 2004 Bjoroya, Norway 2002–2004 Glaucous gull Larus hyperboreus Eggs 4 pools Verreault et al. 2004 Svalbard Glaucous gull Larus hyperboreus Liver 20 

OTHER BIRDSKunisue et al. 2008 Japan (Inland) 2001 Common cormorant Phalacrocorax carbo Muscle 4 0 < 

Kunisue et al. 2008 Japan (Inland) 1998 Jungle crow Corvus macrorhynchos Muscle 5 0 36

Mariussen et al. 2004 Norway 1990–1993 Grouse Liver Voorspoels et al. 2006b Belgium 2001, 2003 Passerines Liver and muscle 25 All 

VEGETATIONSchlabach et al. Skoganvarre, 2002 Moss n/a 1 
127

Study Authors Location Date Species Tissue Sampling N Not Detected L2002 Norway Schlabach et al. 2002 Valvik, Norway 2002 Moss n/a 1 
128

Appendix B – Bioaccumulation, Biomagnification and Trophic Magnification Data

Species, Food Web or 

Predator:Prey Ecosystem Study Type Tissue Location Endpoint Result Sta

Pelagic food web 

(zooplankton, whitefish, white 

sucker, emerald shiner, 

goldeye, burbot, walleye) 

(Calanoid sp., Coregonus 

clupeaformis, Catostomus 

commersoni, Notropis 

atherinoides, Hiodon 

alosoides, Lota lota, 

Stizostedion vitreum) 

Aquatic Field Whole body Lake Winnipeg TMF (lipid normalized) 3.6 Regr

sl

Fish:fish Aquatic Field Whole body Lake Winnipeg BMFs (lipid normalized) 0.1–6.8 Ra

Fish:invertebrates Aquatic Field Whole body Lake Winnipeg BMFs (lipid normalized) 0.2–34 Ra

Pelagic food web 

(zooplankton, sprat, herring, 

salmon) 

(Sprattus sprattus, Clupea 

harengus, Salmo salar) 

Aquatic Field Whole body Baltic Sea TMF (lipid normalized) ~0.6 (not significantly different from 0) Regr

sl

Pelagic food web (roach, 

perch, pike) 

(Rutilus rutilus, Perca 

fluviatilis, Esox lucius) 

Aquatic Field Whole body Baltic Sea TMF (lipid normalized) ~0.80 (not significantly different from 0) Regr

sl

Fish:fish Aquatic Field Whole body Baltic Sea BMFs (lipid normalized) 0.02–1.3 Ra

Fish:zooplankton Aquatic Field Whole body Baltic Sea BMFs (lipid normalized) 0.04–0.11 Ra

Fish:fish Aquatic Field Whole body Baltic Sea BMFs (lipid normalized) 1.71–5 Ra

Juvenile Lake trout 

Salvelinus namaycush Aquatic Lab Whole body na BMF (lipid normalized) 0.3 M
129

Species, Food Web or 

Predator:Prey Ecosystem Study Type Tissue Location Endpoint Result Sta

Carp 

Cyprinus carpio Aquatic Lab Whole body n/a BCF (wet weight) < 3 000 

Rainbow trout 

Salvelinus namaycush Aquatic Lab Carcass n/a BMF (lipid normalized) 0.022 Ra

Rainbow trout 

Salvelinus namaycush Aquatic Lab Carcass n/a BMF (lipid normalized) 0.089 M

Rainbow trout 

Salvelinus namaycush Aquatic Lab Serum n/a BMF (lipid normalized) 0.35 M

Rainbow trout 

Salvelinus namaycush Aquatic Lab Serum n/a BMF (lipid normalized) 0.52 M

Rainbow trout 

Salvelinus namaycush Aquatic Lab Liver n/a BMF (lipid normalized) 1.28 M

Rainbow trout 

Salvelinus namaycush Aquatic Lab Liver n/a BMF (lipid normalized) 1.73 M

Polar cod:ice ampipod 

Boreogadus saida: 

Gammarus wilkitzkii

Marine Field Whole body Svalbard, Norway BMF (lipid normalized) 0.03 M

Grey seals 

Halichoerus grypus

Marine 

mammal Captive Blood North Sea BMF (lipid normalized) 8.3–20.8 Ra

Grey seals 

Halichoerus grypus

Marine 

mammal Captive Blubber North Sea BMF (lipid normalized) 0.025–0.063 Ra

Grey seals 

Halichoerus grypus

Marine 

mammal Captive Blubber North Sea BMF (lipid normalized) 0.044–0.11 Ra

Grizzly bear 

Ursus arctos horribilis Terrestrial Field Fat and muscle Coastal and interior British Columbia Bioaccumulation slope negative (not significantly different from 0) Regr

sl

Earthworm 

Phylum annelida Terrestrial Lab Whole body 

Soil BSAF(lipid 

and OC 

normalized) 

0.044–0.73 (0.3) Range 

Rat Terrestrial Lab Whole body n/a BMF (wet weight) 0.05 M

Cow Terrestrial Lab Adipose n/a BMF (lipid 0.037 M
130

Species, Food Web or 

Predator:Prey Ecosystem Study Type Tissue Location Endpoint Result Sta

tissue normalized) 

Cow Terrestrial Lab Adipose tissue n/a BMF (lipid normalized) 0.31 MCow Terrestrial Lab Whole body n/a BMF (lipid normalized) 0.026 MCow Terrestrial Lab Whole body n/a BMF (lipid normalized) 0.21 M

Sunfish 

Lepomis gibbosus Aquatic Field Whole body Marlowe Ck., Roxboro, NC BSAF (lipid and OC normalized) 0.0018 M

Crayfish 

Campbarus punctic ambarus 

sp. C

Aquatic Field Whole body Marlowe Ck., Roxboro, NC BSAF (lipid and OC normalized) 0.013 M

Various (see comments) Aquatic Field Whole body Pearl River Delta, China BSAF (lipid and OC normalized) 0 – 0.4 Ra

Fish (see comments) Aquatic Field Whole body River Vero, Spain BSAF (lipid and OC normalized) 0.0011 to 0.0013 Ra

Aquatic oligochaete Aquatic Lab Whole body n/a Accumulation Minimal 

Horse mussel 

Modiolus modiolus Aquatic Field Whole body Vancouver Island, Canada 

Sediment BSAF 

(lipid and organic 

carbon 

normalized) 

0–3.53 Ra

Marine mammal: fish Marine Field Blubber and whole fish North Atlantic, coast of Maine, USA BMF (lipid normalized) 0.67–1.3 

Range predat

concen

prey 

conce

Marine food web: beluga Marine Field Blubber and Eastern TMF (lipid 0.3 (statistically Regr
131

Species, Food Web or 

Predator:Prey Ecosystem Study Type Tissue Location Endpoint Result Sta

whale, narwhal, walrus, arctic 

cod, shrimp, clams, 

deepwater redfish, mixed 

zooplankton 

Delphinapterus leucas, 

Monodon monoceros, 

Odobenus rosmarus, 

Boreogadus saida, Pandalus 

borealis, Hymenodara 

glacialis, Mya truncata, and 

Serripes groenlandica, 

Sebastes mentella) 

whole 

organism 

Canadian 

Arctic 

normalized) significant 

negative slope) 

sl

Marine mammal: fish Marine Field 

Blubber and 

whole 

organism 

Eastern 

Canadian 

Arctic 

BMF (lipid 

normalized) <1 

Range 

level a

ratio of 

tis

concen

prey 

conce

Fish: invertebrate Marine Field 

Blubber and 

whole 

organism 

Eastern 

Canadian 

Arctic 

BMF (lipid 

normalized) <1 

Troph

adjuste

predat

concen

prey 

conce

Marine food web: beluga 

whale, ringed seal, arctic cod, 

Pacific herring, Arctic cisco, 

pelagic amphipod, and arctic 

copepod (Delphinapterus 

leucas, Phoca hispida, 

Boreogadus saida, Clupea 

pallasi, Coregonus 

autumnalis, Thermisto 

libelluala, Calanus 

hyperboreus) 

Marine Field 

Blubber, 

liver, and 

whole 

organism 

Western 

Canadian 

Arctic 

TMF (lipid 

normalized) 

Not statistically 

significant 

Regr

sl

Marine mammal: fish Marine Field Blubber and liver Western Canadian Arctic BMF (lipid normalized) 0.3–0.9 

Range 

level a

ratio of 

tis

concen
132

Species, Food Web or 

Predator:Prey Ecosystem Study Type Tissue Location Endpoint Result Sta

prey 

conce

Fish: invertebrate Marine Field Liver and whole body Western Canadian Arctic BMF (lipid normalized) 4.8–12.7 

Range 

level a

ratio of 

tis

concen

prey 

conce

Freshwater food web: water 

snake, northern snakehead, 

mud carp, common carp, 

crucian carp, prawn, and 

Chinese mysterysnail 

(Enhydris chinensis, Channa 

argus, Cirrhinus molitorella, 

Cyprins carpio, Carassius 

auratus, Macrobrachium 

nipponense, 

Cipangopaludina chinensi) 

Aquatic Field Whole body 

Reservoir near 

Qinyuan City, 

South China 

TMF (lipid 

normalized) 

0.26 (negative 

slope, p=0.53) 

Regr

sl
133

Appendix C – Selection of Log Kow for QSAR Modelling of decaBDE

The log Kow for commercial DecaBDE of 6.265 determined at 25°C by CMABFRIP 

(1997a) is frequently cited in the literature. This value was determined using the 

generator column method with a composite material received from Great Lakes Chemical 

Corp., Albemarle Corp. and Bromine Compounds Ltd. and consisting of 97.4% 

decaBDE, 2.5% nonaBDE and 0.04% octaBDE. The analytical method used was 

GC-ECD and, using this method, the mean concentration measured in the aqueous 

material was 0.04 μg/L, while the mean concentration in octanol stock solution was 

0.0738 g/L. This resulted in a Kow value of 1.845x106, or log Kow of 6.265. It is of interest 

that the log Kow determined for decaBDE by CMABFRIP was in fact lower than those 

which they determined for the commercial OctaBDE (6.29; CMABFRIP 1997b) and 

PentaBDE technical products (6.59; MacGregor and Nixon 1997) using the same 

procedure. This is not expected, due to the increased hydrophobicity and increased Kow, 

which would be expected with increased bromination. This also does not reflect available 

information respecting commercial DecaBDE’s solubility in water. CMABFRIP (1997c) 

determined that the limit of water solubility was < 0.1 μg/L at 25°C, while for 

commercial OctaBDE (CMABFRIP 1997d) it was determined to be 0.5 μg/L, and 

13.3 μg/L for commercial PentaBDE (Stenzel and Markly 1997). The available QSAR 

estimates for decaBDE indicate that it is predicted to be in the range of 2.841 x 10-8 to 

2.6078 x 10-4 μg/L at 25°C. Thus, one can expect that difficulties measuring substances 

with such low solubility in water could have affected the determination of log Kow by 

CMABFRIP (1997b). The European Communities (2002) also indicate that the presence 

of trace amounts of octanol in the water phase of this study may have had an apparent 

effect on the water concentration and the log Kow value. For these reasons, the log Kow for 

DecaBDE of 6.265 is considered too low and not appropriate for BAF and BMF 

modelling purposes. 

Watanabe and Tatsukawa (1990) reported a significantly higher log Kow of 9.97. Their 

approach, which applied to other PBDEs, showed a pattern of increased Kow with 

increasing bromination using a reverse-phase HPLC method. 

Overall, the range of reported and measured values of logKow for decaBDE indicate a 

significant degree of uncertainty and measurement error. Log Kow measurements and 

estimates in this report range from 6.3 to 12.9. Other available estimates include 7.68, 

10.37, 10.14, 9.44, 12.11 and 11.48 (from: www.vcclab.org/lab/alogps/start.html). The 

comprehensive Syracuse Research Corporation database used for the KOWWIN program 

contains 11 135 values for log Kow. The maximum measured value in this training set is 

11.3 and the upper 99.5 percentile is 7.8 (personal communication from JA Arnot to 

Environment Canada 15 February 2008; unreferenced). Therefore, most of the model 

estimates for decaBDE are greater than presently measured values for all other organic 

chemicals. Analytical limitations and high levels of measurement error are the principle 

reasons for lack of log Kow values generally above 8.0 (e.g., Wania and Dugani 2003). 

Wania and Dugani (2003) reported a log Kow value of 8.7 for decaBDE as the most 

reliable input for long-range transport modelling based on the establishment of a QSPR 

(quantitative structure-property relationship) for higher brominated PBDEs and using the 
134

three solubility concept. This value was judged to represent a more realistic logKow value 

for modelling purposes and was used for bioaccumulation modelling in this report. The 

logKow value of 8.7 was further used to adjust the predicted logKows of the expected 

metabolites of decaBDE using the Experimental Value Adjustment (EVA) method 

outlined in the KOWWIN program. The EVA method takes into account structural 

differences between the metabolite and parent compound (decaBDE) and adjusts the 

preferred value of 8.7 accordingly. 
135

Appendix D – Model Input Parameters for BAF–QSAR and Terrestrial 

Biomagnification Model Predictions for DecaBDE Transformation Products 

(a) BAF–QSAR input parameters for lower brominated PBDEs, hydroxylated BDEs and 

hydroxymethoxy-BDEs 

Notes: 

1

 Estimated using KOWWIN using the EVA method with a reference logKow of 8.7 for decaBDE. 

2

 Estimated using the KOAWIN estimation model and with input of corrected logKow from 

KOWWIN EVA method. 

3

 Maximum measured value (Watanabe and Tatsukawa 1990). 

4

 kM based on Tomy et al. (2004) and normalized to weight of middle trophic level fish in Arnot 

and Gobas (2004) model and 15o

C. These metabolites are expected to undergo further 

metabolism; however, the actual rates of metabolism are unknown. 

5

 Information regarding the half-life and kM of nonaBDEs is lacking in the literature and a half-life 

of 26 d (equal to that of decaBDE) was therefore assumed. Normalized to weight of middle 

trophic level fish in Arnot and Gobas (2004) model and 15o

C. 

6

 Based on a half-life of 19 d (Tomy et al. 2004) and normalized to weight of middle trophic level 

fish in Arnot and Gobas model and 15o

C. 7

 Based on a half-life of 38 d (Stapleton et al. 2004, Tomy et al. 2004—estimates ranged from 19 

to 346 d) and normalized to weight of middle trophic level fish in Arnot and Gobas (2004) model 

and 15o

C. 

8

 Based on a half-life of 58 d (Stapleton et al. 2004, Tomy et al. 2004—estimates ranged from 23 

to 231 d) and normalized to weight of middle trophic level fish in Arnot and Gobas (2004) model 

and 15o

C. 

9

 Based on half-life of 50.7 d (Stapleton et al. 2004, Tomy et al. 2004—estimates ranged from 43 

to 173 d) and normalized to weight of middle trophic level fish in Arnot and Gobas (2004) model 

and 15o

C. 

Metabolite Log Kow

Median kM (1/d) 

(184 g fish at 

15oC 

Hydroxymethoxy-octaBDEs 6.51 0.024

Hydroxymethoxy-hexaBDEs 4.71 0.024

Hydroxymethoxy-pentaBDEs 3.91 0.024

Hydroxymethoxy-heptaBDEs 5.61 0.024

Hydroxy-hexaBDEs 4.71 0.024

Hydroxy octaBDEs 6.41 0.024

Hydroxy-nonaBDEs 7.31 0.024

nonaBDEs 7.82 0.025

octaBDEs 6.93 0.036

heptaBDEs 6.02 0.017

hexaBDEs 5.12 0.0018

pentaBDEs 4.31 0.0019
136

(b) Terrestrial biomagnification model input parameters for lower brominated PBDEs, 

hydroxylated BDEs and hydroxymethoxylated BDEs

Name and Structure Log Kow Log Koa

ED

9 

kM (1/d) 

(wolf BW 

normalized 

80 kg 

Hydroxymethoxy octaBDE 6.51 17.22

99% 0.0045

Hydroxymethoxy pentaBDE 3.91 13.42

99% 0.0045

Hydroxymethoxy hexaBDE 4.71 14.62

99% 0.0045

Hydroxymethoxy heptaBDE 5.61 15.92

99% 0.0045

Hydroxy-hexaBDE 4.71 13.42

99% 0.0045

hydroxy-octaBDE, 6.41 15.92

98% 0.0045

hydroxy-nonaBDE 7.31 17.22

96% 0.0045

NonaBDEs 7.83 13.72

90% 0.0076

OctaBDEs 6.94 12.42

98% 0.0037

HeptaBDEs 6.03 11.12

99% 0.0048

HexaBDEs 5.13 9.82

99% 0.0048

PentaBDEs 4.33 8.62

99% 0.0048 

1

 Estimated using KOWWIN using the EVA method with a reference logKow of 8.7 for decaBDE. 

2

 Estimated using KOAWIN and with input of corrected logKow from KOWWIN EVA method. 

3

 Lowest QSAR estimates from the Supporting Working Document (Environment Canada 2006b). 4

 Maximum measured value (Watanabe and Tatsukawa 1990). 

5

 Average based on longest observed half-lives for decaBDE, nonaBDEs and octaBDEs by Huwe 

and Smith (2007) and normalized to BW of wolf in Gobas Model (80 kg). 

6

 Based on longest half-life of 22.6 d for nonaBDEs observed by Huwe and Smith (2007) and 

normalized to BW of wolf in Gobas Model (80 kg). 

7

 Based on longest half-life of 64.8 d for octaBDEs observed by Huwe and Smith (2007) and 

normalized to BW of wolf in Gobas Model (80 kg). 

8

 Based on half-life of 43.7 d (average of half-lives for nona- and octaBDEs) and normalized to 

BW of wolf in Gobas Model (80 kg). 

9

 Estimated according to Kelly et al. (2004) based on ED for humans. 

(c) BAF–QSAR and terrestrial biomagnification model input parameters for brominated 

dibenzofurans 

Product Smiles Log Kow

1

 Log Koa

2 Aquatic 

kM

3 (1/d) 

Mammal 

kM

4 (1/d) 

ED 

triBDBFs C12=CC(Br)=C(Br)C=C1OC3=C

2C=C(Br)C=C3 3.0 7.1 0.02 0.004 99% 

tetraBDBFs C12=CC(Br)=C(Br)C=C1OC3=C

2C=C(Br)C(Br)=C3 3.9 8.2 0.02 0.004 99% 

pentaBDBFs C12=CC(Br)=C(Br)C=C1OC3=C

2C=C(Br)C(Br)=C3Br 4.8 9.5 0.02 0.004 99% 

hexaBDBFs C12=C(Br)C(Br)=C(Br)C=C1OC

3=C2C=C(Br)C(Br)=C3Br 5.6 10.7 0.02 0.004 99% 

1

 Estimated using KOWWIN using the EVA method with a reference logKow of 8.7 for decaBDE. 

2

 Estimated using KOAWIN and with input of corrected logKow from KOWWIN EVA method. 
137

3

 kM for these metabolites is illustrative only and is consistent with a half-life of 26 d and 

normalized to weight of middle trophic level fish in Arnot and Gobas (2004) model and15oC. 

These metabolites are expected to undergo further metabolism; however, the actual rates of 

metabolism are unknown. 

4

 Average based on longest observed half-lives for decaBDE, nonaBDEs and octaBDEs by Huwe 

and Smith (2007a, 2007b) and normalized to body weight of wolf in Gobas model (80 kg); kM for 

these metabolites is illustrative only and is consistent with a half-life of 26 d. These metabolites 

are expected to undergo further metabolism; however, the actual rates of metabolism are 

unknown. 
138

Appendix E: Summary of Environmental Degradation and Debromination Studies for decaBDE (a) Photodegradation Study Authors Type of Study Substance 

Purity 

Degradation Mediator 

Experimental Conditions 

Degradation Rate 

Watanabe and Tatsukawa (1987) 

Photodegradation 97% decaBDE 

and 3% nonaBDE 

Artificial UV Light Dissolved in organic 

solvents 

NR 

Söderström et al. (2004) 

Photodegradation Traces of octaand nonaBDEs Natural sunlight and artificial UV 

light 

Sorbed to thin layer of silica gel, sand, soil or sediment 

NR 

Jafvert and Hua (2001) 

Photodegradation 98% Artificial and 

natural sunlight 

Sorbed to hydrated surfaces of glass and silica sand particles, humic acid-coated silica particles, and glass surfaces in contact with aqueous solutions 

12–71% degradation over 60–72 h 

Hua et al. (2003) Photodegradation NR Artificial sunlight 

or natural sunlight 

Precipitated onto quartz glass, silica particles and humic acid-coated silica particles; hydrated 

44–71% degradation over 60–72 h; humic acid slowed the rate decaBDE decay 

Palm et al. (2003) Photodegradation NR Artificial xenon 

lamps 

Dispersed in organic solvents 

Half-life ~ 30 minutes 
139

Study Authors Type of Study Substance 

Purity 

Degradation Mediator 

Experimental Conditions 

Degradation Rate 

Palm et al. (2003) Photodegradation/oxidation NR Simulated sunlight 

or hydroxyl radicals 

Sorbed to aerosol (silicon dioxide) 

< 6 x 1013 cm3 molec

1s-1 

Bezares-Cruz et al. (2004) 

Photodegradation NR Natural sunlight decaBDE dissolved in 

hexane 

99% reduction in the decaBDE concentratin as little as 30 minu

Eriksson et al. (2004). 

Photodegradation 98% decaBDE Artificial UV light Organic solvents / 

water mixtures 

4×10-4/s in the 

methanol/water mixt(half-life ~ 0.5 hr), 6.5×10-4/s in methan

(half-life ~ 0.3 hr), 8.3×10-4/s in pure tetrahydrofuran (half~ 0.23 hr) 

Eriksson et al. (2004). 

Photodegradation 98% decaBDE Artificial UV light Water / humic acid 

mixture 

3×10-5/s in water with 

humic acid (half-life ~ 6.4 hr 

Geller et al. (2006) 

Photodegradation 98% Artificial light Dissolved in 

tetrahydrofuran 

NR 

Kuivikko et al. (2006) 

Photodegradation >98.3% Natural sunlight Dissolved in iso-octane; 

modelling to determine rates in ocean 

Half-life of 0.03 days iso-octane Predicted half-lives ithe Baltic Sea of 0.02 days (surface) a1.2 days (mixing layeand in the Atlantic Ocean of 0.09 days (mixing layer) 

Kuivikko et al. (2006, 2007) 

Photodegradation > 98.3% Natural sunlight Dissolved in iso-octane; 

modelling to determine rates in ocean 

Half-life of 0.03 days iso-octane; mixing zohalf-lives of 1.8 days (Baltic Sea) and 0.4 days (Atlantic Ocean) 
140

Study Authors Type of Study Substance 

Purity 

Degradation Mediator 

Experimental Conditions 

Degradation Rate 

Ahn et al. (2006a) Photodegradation 98% Artificial light and 

natural sunlight 

Sorbed to montmorillonite, kaolinite, organiccarbon-rich natural sediment (16.4% OC content), aluminum hydroxide, iron oxide and manganese dioxide 

For artificial light, halives ranged from 36 178 days For natural light, halflives for montmorillonkaolinite and sedimewere 261, 408 and 990 days, respective(negligible degradation other matrices) 

Stapleton and Dodder (2006) 

Photodegradation NR Natural sunlight Either native decaBDE 

in house dust or decaBDE spiked house dust 

2.3x10-3/hr in spiked 

dust and 1.7x10-3/hr 

natural dust, corresponding to half-lives of 301 and 408 h in sunlight, respectively 

Gerecke (2006) Photodegradation 98% Natural sunlight DecaBDE (BDE209) 

sorbed to kaolinite. Irradiated dry or in water 

Half-lives of 76 and 7minutes were determined for dry awet conditions, respectively; dependon light penetration 

Stapleton (2006b) Photodegradation NR Natural sunlight Either native decaBDE 

in house dust or decaBDE spiked house dust 

Half-life for decaBDE was estimated at 216 
141

Study Authors Type of Study Substance 

Purity 

Degradation Mediator 

Experimental Conditions 

Degradation Rate 

Hagberg et al. (2006) 

Photodegradation NR Artificial light DecaBDE (BDE209) 

dissolved in toluene 

NR 

Barcellos da Rosa et al. (2003) 

Photodegradation 98% Artificial light DecaBDE (BDE209) 

dissolved in toluene 

3x10-4/s 

Kajiwara et al. (2008) 

Photodegradation NR Natural sunlight HIPS (high-impact 

polystyrene) added to toluene and DecaBDE (BDE209) (100ug/L); toluene evaporated off in dark. A subset of samples hydrated with water to examine effect on photolysis. Crushed TV casing also tested. 

HIPS + DecaBDE samples: 50% BDE2reduction after 1 weeHalf life of BDE209 iHIPS estimated at 5days. Hydrated samples showed faster degradation than nonhydrated. No degradation in dacontrols. No degradation of BDE209 in crushed Tcasing samples durin224 days of sunlight irradiation. 

NR – not determined or not reported. 
142

(b) Other Abiotic Degradation Study Authors Type of Study Substance 

Purity 

Degradation Mediator 

Experimental Conditions 

Degradation Rate 

Keum and Li (2005) 

Reductive debromination NR DecaBDE (i.e., 

BDE209) reducing agents – zerovalent iron, iron sulphide, and sodium sulphide. 

Dissolved in ethyl acetate 

Up to 90% transformation of decaBDE after 40 d 

Li et al. (2007) Reductive debromination Commercial 

DE-83R DecaBDE (> 97%, Great Lakes Chemical) 

Nanoscale zerovalent iron 

Dissolved in acetone, distilled water added. 25 ± 0.5°C 

Half-life was 2.5 h 

Rahm et al. (2005) Hydrolysis (nucleophilic aromatic 

substitution) 

NR Reaction with 

sodium methoxide 

Dissolved in methanol Half-life for the 

hydrolysis reaction wa0.028 h 

Ahn et al. (2006b) Metal oxide-mediated 

debromination 

98% Birnessite Sorbed to birnessite in 

THF:water and water:catechol systems 

THF:water – >75% transformation of decaBDE in 24 h Catechol:water – degradation only observed for highest concentration of catechol (15% degradation over 23 days) 

NR – not determined or not reported
143

(c) BiodegradationStudy Authors Type of Study Substance 

Purity 

Degradation Mediator Experimental 

Conditions 

Degradation Rat

MITI 1992 Biodegradation NR Activated sludge Aerobic conditions No degradation a

weeks 

CMABFRIP (2001) 

Biodegradation 97.4% 

decaBDE, 2.5% nonaBDE and 0.04% octaBDE 

Sediment/water system Anaerobic 

conditions 

< 1% mineralizatobserved over 32 weeks 

He et al. (2006) Biodegradation 98% Anaerobic bacteria DecaBDE (i.e., 

BDE209) dissolved in TCE and inoculated with anaerobic culture/medium 

Degradation only observed with onculture (with S. 

multivorans) in w

0.1 μM decaBDE 

degraded to nondetectable levels 2-month experim

Knoth et al. (2007) 

WWTP monitoring (Biodegradation) 

NR WWTP sludge (primary, 

secondary and digested) 

Field WWTP NR 

Gerecke et al. (2005) 

Biodegradation 97.9% 

decaBDE, 2.1% nonaBDEs 

Sewage sludge + primers (1 or more of 4-bromobenzoic acid, 2,6-dibromobiphenyl, tetrabromobisphenol A, hexabromocyclododecane and decabromobiphenyl) 

Anaerobic conditions 

DecaBDE (i.e., BDE209) decreasby 30% within 23Rate constant of 10-3 d-1 
144

Study Authors Type of Study Substance 

Purity 

Degradation Mediator Experimental 

Conditions 

Degradation Rat

Gerecke et al. (2006) 

Biodegradation 97.9% 

decaBDE, 2.1% nonaBDEs 

Sewage sludge + single primers (either 2,6-dibromophenol or 4-bromobenzoic acid) 

Anaerobic conditions 

Half-lives of 700 (with primer) and days (without primwere observed. Monitoring at an operating WWTP found that the concentration of DecaBDE in sluddecreased betwethe influent and ostreams. 

La Guardia et al. (2007) 

WWTP monitoring (biodegradation) 

n/a WWTP Sludge NR Minimal evidence 

debromination in WWTP sludge or sediments 

Parsons et al. (2004) 

Reductive debromination in sediments 

NR Anaerobic sediments Anaerobic sediment 

suspensions in anaerobic medium Sediments collected from Western Scheldt 

Significant decreain decaBDE over months; howeverresults are highly uncertain since sdecrease was observed in abiotcontrol 

Parsons et al. (2007) 

Reductive debromination in sediments 

NR Anaerobic sediments Anaerobic sediment 

suspensions in anaerobic medium Sediments collected from Western Scheldt 

No statistically significant decreadecaBDE over 26days 
145

Study Authors Type of Study Substance 

Purity 

Degradation Mediator Experimental 

Conditions 

Degradation Rat

Tokarz III et al. (2008) 

Reductive debromination in sediments 

NR Anaerobic sediments and 

cosolvent-enhanced biomimetic system 

Natural sediments with no detectable PBDEs collected from Celery Bog Park, West Lafayette, Indiana. PBDEs dissolved in a toluene solution added to sediments, then evaporated off. This mixture was then blended with wet sediments. Biomimetic experiment involved the use of Tefloncapped glass vials with 0.03 mM of BDE209, -99 or -47 mixed with 5.0 mM titanium citrate and 0.2 mM vitamin B12 in 0.33 M TRIZMA buffer solution containing tetrahydrofuran. 

The biomimetic system demonstrreductive debromination at decreasing rates decreasing bromination (e.g.life of 18 seconds BDE209 and alm60 d for BDE47) In natural sedimemicrocosms, the life for BDE209 westimated to rangfrom 6 to 50 yearwith an average oyears, based on observations oveyears 
146

Study Authors Type of Study Substance 

Purity 

Degradation Mediator Experimental 

Conditions 

Degradation Rat

Kohler et al. (2008) 

Debromination in sediments NR Natural lake sediments Field – lake 

sediment cores 

Cores show DecaBDEs appealayers correspondto 1970s, >30 ye

NR – not determined or not reported.
147

Appendix F: Structural Diagrams for DecaBDE and DecaBDethane 

Substance Name and CAS Number Structure 

Decabromodiphenyl Ether 

CAS: 1163–19–5 

Decabromodiphenyl Ethane 

CAS: 84852–53–9 

Br

Br Br

Br Br

Br Br

Br

Br Br