### Ecological State of the Science Re port on Decabromodiphenyl Ether (decaBDE)

Bioaccumulation and Transformation

Prepared by:

Environment Canada

August 2010ii 

### 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 Minist er of Health in the Canada Gazette , Part I. It was concluded that PBDEs—i.e., tetrabromodiphenyl ethe r (tetraBDE), pentabromodiphenyl ether (pentaBDE), hexabromodiphenyl ether (hexaB DE), heptabromodiphenyl ether (heptaBDE), octabromodiphenyl ether (octaBDE), nona bromodiphenyl ether (nonaBDE) and decabromodiphenyl ether (decaBDE)—whic h are found in commercial PentaBDE, OctaBDE and DecaBDE technical formulations, are entering the environment in a quantity

or concentration or under condi tions that have or may have an immediate or long-term harmful effect on the environment or its biologic al diversity, and thus meet the criteria set out in paragraph 64(a) of the Canadian E nvironmental Protection Act, 1999 (CEPA 1999). In addition, it was concluded th at all seven PBDEs homologues met criteria for persistence, but only tetra- to hexaBDE met the criter ia for bioaccumulation as defined in the Persistence and Bioaccumulation Regulations . The analysis also noted that the higher brominated diphenyl ethers, and decaBDE in pa rticular, could accumulate to some degree

in biota and debrominate to bioaccumula tive and persistent transformation products.

Since the completion of the Ecological Screening Assessm ent, a large amount of new information has been published regarding th e accumulation of decaBDE in biota and the potential transformation of decaBDE to pe rsistent and bioaccumulative products. The purpose of the present report is to provide an updated analysis of bioaccumulation and transformation of decaBDE, by summarizing evid ence considered in the original Screening Assessment, and then examining the relate d new science published up to August 25, 2009.

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

bioaccumulation as defined in the Persistence and Bioaccumulation Regulations . With regard to bioaccumulation, potential factors su ch as low assimilation efficiency and/or metabolic transformation appear to be im portant determinants of accumulation in organisms. Nevertheless, some recent studies ha ve 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, glauc ous gull, red fox, shark, harbour porpoise and whitebeaked dolphin, measured concentrations of decaBDE in tissues are interpreted as high. While trophic magnification or bioaccumula tion 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 consum ing contaminated refuse and/or inhabiting decaBDE hotspots close to industrialized areas.

This report also considers it reasonable to co nclude that decaBDE ma y 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 mamm als appear to have some capacity to metabolically break down decaBDE. In fis h, decaBDE appears to form hepta- to nonaBDEs, and potentially pent a- 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 a nd only a fraction (typical ly on the order of a few percent) of the total amount of decaBDE administered to the organism. However, some rodent studies have made inferences, base d on mass balance evalua tions, that rates of transformation may be higher, with one study suggesting that approximately 45% of theiii total dose of decaBDE was unaccounted for a nd may have been metabolized to other compounds (such as hydroxylated and hydr oxymethoxylated PBDEs) and/or bound as

inextricable residues.

The evaluation of transformation in the envir onment identified numerous laboratory studies that provide evidence that decaBDE may break down in the environment, particularly as a result of photodegradation and biodegradati on. Studies of photodegradation of decaBDE sorbed to solids in aqueous and dry systems have demonstrated transformation of decaBDE to tri- to nonaBDEs, tri- to octabrominat ed dibenzofurans (octaBDFs) and unidentified products. While relevant to the environment, the actual fraction of decaBDE exposed to sunlight adsorbed to atmospheric and aquati c particulates, or solids (anthropogenic or natural), would be a small fraction of the to tal amount of decaBDE in the environment. Biodegradation studies have also shown poten tial breakdown of decaBDE mainly to nona-, octa- and heptaBDEs, while transformation to triBDEs has also been shown under enhanced laboratory conditions. Ov erall, biodegradation appears to occur at a much slower rate than that of phototransforma tion, 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 (BAF s) and BMFs was conducted to estimate whether transformation products of decaBDE re sulting from processes in organisms and in the general environment could be bioaccumula tive. The evaluation found that many of the identified transformation products could be bi oaccumulative (i.e., have BAFs in excess of 5000) and some could biomagnify in food chai ns. The analysis also indicated potential transformation of decaBDE to products (i.e ., tetra- to hexaBDEs) that have been

established as bioaccumulativ e based on empirical evidence.

While laboratory studies on the transforma tion 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 studie s to occur in the environment. This suggests that the process of environmental transf ormation 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 whic h are dominated by congeners found in the commercial products. The fact that fewer studi es 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 similariti es in physical and chemical properties, persistence, transformation patterns, and accumulation in organisms. Based on the similarity in properties be tween 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 capaci ty to accumulate in wildlife and transform to potentially bioaccumulative products. Unders tanding the risk from alternatives generally will help to ensure that substitutions of flam e 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, Bioaccumula tion 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, Biomagnifi cation and Trophic Magnification Data.

Appendix C – Selection of Log K

ow for QSAR Modelling of DecaBDE.

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

Model Predictions for DecaBDE Transformation Products.

Appendix E – Summary of Envi ronmental Degradation and De bromination Studies for

DecaBDE.

Appendix F – Stuctural Diagrams for DecaBDE and DecaBDethanev 

### List of Tables

Table 1-1: Concentrations (%, w/w) of PBDE s in selected commercial Octa- and DecaBDE products (La Guardi a 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 concentratio ns of decaBDE in freshwater and marine biota from the Lumparn Estuary, Baltic Sea a nd 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 δ 15 N 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 Kierkega ard 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 d ecaBDE 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 M odel for decaBDE in the “Ontario Mixed Wooded Plain” region.......................................................................................................... 8 3 Table 3-2: Summary of decaBDE transformati on 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 frequenc y distributions of predicte d BAFs for metabolites of decaBDE (predictions for middle trophic level).................................................................. 61 Figure 1-2: Conceptual mode l of possible environmental transformation pathways for

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

### 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

E D 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

k 2 elimination rate constant

k E fecal egestion rate constant

k G growth rate constantvii kg/d kilograms per day

k M metabolic rate constant

K
oa octanol-air partition coefficient

KOAWIN model for prediction of octanol-air partition coefficient

K
ow octanol-water partition coefficient

k T 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 structur e-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 etherviii 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 plant1 

## 1 Introduction

The purpose of this report is to provide an updated analysis of the bioaccumulation and environmental transformation of decabromodiphe nyl ether (decaBDE), to be considered in the context of the information and analys es 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 indi vidual studies when forming a weight of evidence for persistence, bioaccumulation or i nherent 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 informati on and developed conclusions based on a weight-of-evidence approach as re quired 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, inhere nt toxicity, risk quotients and long-range transport to remo te regions like the Arctic). The PBDE screening assessment concluded that PBDE s—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 fo rmulations, 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 wa s considered for inclusion into the environmental screening assessment (Envir onment Canada 2006a; Environment Canada 2006b), while information received betw een November 2004 and October 2005 was reviewed but not generally added to the a ssessment, as these studies supported the conclusions of the draft assessment publis hed 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 bi oaccumulation and transformati on. In addition, a Notice of Objection, dated February 14, 2007, was submitted to the Minister of the Environment by2 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 esta blishment of a Board of Revi ew 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 in form a decision respecting the formation of a Board of Review, this detailed evaluation wa s prepared to consider the current science respecting the bioaccumulation and transf ormation of decaBDE to bioaccumulative products. This report consid ers 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 re levant information on bioaccumulation and

transformation examined in the PBDE scr eening assessment (Canada 2006; Environment Canada 2006b), and examines how the new scien ce builds upon existing lines of evidence respecting bioaccumulation and transformation. Ecological considerations rather than those relating to human health are presente d. 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 PB DEs so that levels do not reach a point where they could potentially harm the health of Canadians. The

evaluation does not consider decaBDE accumu lation 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 subs tantial internal and external written peer

review/consultation. While exte rnal comments were taken in to consideration, the final contents remain the responsibility of Envir onment Canada. Additionally, the draft of this evaluation was subject to a 60-day public co mment period. The critical information and considerations upon which the evaluati on is based are summarized below.



### 1.1.1 Composition of Commercial DecaBDE

In this evaluation, information on both the c ongener form of decaBDE (BDE209) and the commercial product also known as Decab romodiphenyl Ether (DecaBDE) are

considered. Since nona- and octaBDEs are al so found in DecaBDE formulations, these congeners are considered if appropriate.

According to the World Health Orga nization (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. For3 instance, FR-300-BA, produced in the 1970s and no longer commercially available, contained 77.4% decaBDE, 21.8% nonaBDE a nd 0.8% octaBDE (Norris et al. 1975).

La Guardia et al. (2006) analyzed compositi ons 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 (T able 1-1). They found that Saytex 102E and

Bromkal 82-ODE contained 96.8% and 91.6% BD E209, respectively. Both formulations contained nonaBDEs, with BDE206 in the highest quantity, followed by BDE207 then -208. In addition, Bromkal 82-ODE containe d 0.56% octaBDEs, with BDE196, -203 and

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

Commercial OctaBDE products (no longer in production) also contained decaBDE (La Guardia et al. 2006). For instance, Bromka l 79-8DE contained 49.6% decaBDE as well as significant nonaBDE (Table 1-1). In comp arison, DE-79 contained lesser amounts of deca- and nonaBDE (i.e., 1.31 and 13.07%, respectively).

Table 1-1: Concentrations (%, w/w) of PBDE s in selected commercial Octa- and DecaBDE products (La Guardia et al. 2006)

Commercial OctaBDE Products Commercial DecaBDE Products PBDE Congener DE-79 Bromkal 79-8DE Saytex 102E Bromkal 82-ODE

BDE154 1.07 0.04 nd 2 nd

BDE144 0.1 0.12 nd nd

hexaBDE 1 < 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 heptaBDE 1 < 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

octaBDE 1 < 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 detected4 

### 1.2 Definitions and Rationale

For the purposes of this document, it is im portant to define an d distinguish between bioaccumulation, bioconcentration and bioma gnification. 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 transformationof 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 detectab le 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 th e 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 beca use they tend to biomagnify, i.e., increase in concentration from prey to predator acros s 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 “bioaccumul ation” and the term

“bioaccumulative” as used in the Persistence and Bioaccumulation Regulations (Canada 2000), a regulation ma de under CEPA 1999 and as de scribed in Canada’s Toxic Substances Management Policy (C anada 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 intrinsi c properties of the substance, the ecosystem under considerati on and the conditions in the environment.” Thus, determining whether these criteria ar e met involves professional judgement which considers the intrinsic properties of the s ubstance and ecosystem under consideration.

These criteria were first propos ed as part of the federal Toxic Substances Management Policy (Canada 1995a, 1995b) and were intende d 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. Applica tion of a criterion of 5000 for bioaccumulation5 factors (BAFs) or bioconcentrati on factors (BCFs), or 5 for log K

ow (octanol-water partition coefficient), are recommended in the Policy. Substances th at 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 K

ow should be used with caution in predicting the bioconcentration and bioaccumulation potenti al of organic substances (Canada 1995b). For instance, many substances can be metabolized and this can function to decrease bioaccumulation potential of the parent com pound, and thus many substances with a very high log K

ow (e.g., >5) can have very low bioaccumu lation/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, “b ioaccumulation” is the process which may lead to a substance being considered “bio accumulative.” Evidence that a substance is both persistent and bioaccumulative when comb ined with evidence of toxicity and release or expected release into the environment prov ides a significant indi cation of its potential to cause ecological harm (Environment Canada 2007).

Gobas and Morrison (2000) provide a c oncise definition of the process of “biomagnification” as it is currently understood:

This is the process in which the chemical con centration 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 determin ed 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, bec ause 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,” “f ood web magnification” and “food web biomagnification” are often used interchang eably to refer to the same phenomenon. This is the phenomenon whereby chemical concentr ations in organisms increase with trophic level, resulting in higher concen trations in predators than in prey (i.e., biomagnification through successive trophic le vels). Gobas and Morrison (2 000) 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 pre dators 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 f ood 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 a nd 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 ofte n estimated by representing the relationship between the test organisms a nd 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 bi omagnification or tr ophic magnification. It

should be noted that BAFs measured or esti mated 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 subsequen tly the bioaccumulation potential. The BAF– QSAR model of Arnot and Gobas (2003) pr ovides bioaccumulation predictions for a generic aquatic food web (with lower, mi ddle and upper trophic levels) based on the current understanding of bioaccumulation pro cesses in aquatic organisms, accounting for the available fraction in water. Figure 1-1 i llustrates the model-predicted relationship for hydrophobic, non-metabolized substances between log K

ow , BAF, BCF and BMF for aquatic lower and upper trophic levels in the model. The model predictions suggest that as log K

ow increases toward and above 4–5, the BAF increases toward and above 5000. Also, in this range of log K

ow , 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 orga nisms, the BMF results from competing processes of chemical uptake (respiratory and dietary) and chemical elimination (respiratory and gastrointest inal), which depend on organism physiology and chemical hydrophobicity. For aquatic organisms, respirat ory elimination is significant for less hydrophobic chemicals, and elimination proces ses override gastroin testinal absorption processes. However, as hydr ophobicity increases (i.e., log K

ow > 5), respiratory elimination slows and gastrointestinal ab sorption becomes signi ficant, resulting in biomagnification (i.e., a BMF ex ceeding 1). At very high log K

ow , dietary chemical absorption efficiency tends to decline because chemicals are too strongly sorbed to the food matrix to be absorbed into the organi sm, resulting in a decline in both the BAF and the BMF (Arnot and Gobas 2003, 2004). Regardless of log K

ow , when sufficient,

chemical metabolism often sustains the tota l elimination rate at a high enough level to prevent chemical biomagnificat ion, even for higher log K

ow substances (Arnot and Gobas 2003; Kelly et al. 2004).7 This relationship between log K

ow , 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 criter ion of 5000 to indicate when bioaccumulation may occur. Note that the BCF is a less cons ervative endpoint since it does not increase as quickly with log K

ow as the BAF. However, a lower trophic level BCF of 5000 generally corresponds to a log K

ow of 5, which is consis tent with the log K

ow 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 K owlog BAF or BCF

0 1 2 3 4 5 6 7 8BMF 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 (sedim ent 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 cap acities 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 biomagnifi cation or trophic dilution). However, given the observed difference in sorptive capacity be tween lipid and organi c carbon (see Seth et al. 1999), BSAFs up to approximately 3 are still consistent with equilibrium conditions. Alternatively, the American Society for Te sting 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 f ood web biomagnification or an alternative magnification process is increa sing organism chemical concen trations above equilibrium. Although there are currently no guidelines fo r 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 ma gnification process is increasing organism chemical concentrations above equilibrium.8 The evaluation of bioaccumulati on 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. Concentr ations 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; howev er, this metabolic capacity can also result

in the formation of bioaccumulative tran sformation products. Full evaluation of BAFs, BCFs, BMFs and TMFs would consider the bi oaccumulation potential of both the parent substance and its metabolic product(s). Furthe r discussion on metabolic transformation is found later in this report.



## 2 Evidence for Bioaccumulation

The evaluation of new lines of evidence conducte d in this section reli es on the definitions of bioaccumulative and biomagnification summarized in the preceding section. Most importantly, the evaluation of whether these ne w lines of evidence indicate that decaBDE is bioaccumulative relies primarily on ratio-based methods for bioaccumulation

assessment. From the definitions in Secti on 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 suppor ting line of evidence for ratio-based measures of bioaccumulation. However, it is recognized th at the characterization of biota chemical concentrations as “high” is subjective a nd is based on professional judgement, and thus prone to alternativ e 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 concentr ations in predator organisms, there is a lack of knowledge regarding levels of exposur e. 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 magnifi cation, it could also be due to high geographically localized exposur e to a particular chemical, as would be th e case if the organism was inhabiting or scavenging from wa ste dumps. It is also possible that uptake

routes other than the food chain can be occurr ing, such as inhalati on, or direct ingestion of plastic. Further, the analysis of decaBDE can be challenging, affected by a number of critical factors that may contribute to th e overall accuracy of reported levels in the environmental samples. For instance, de Bo er and Wells (2006) note that particular attention is required during sampling and an alysis to avoid issues relevant to

− degradation due to sunlight, or UV light;9 − poor solubility;

− high background concentrations/potentia l 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 pr ecision of congener determination. They discuss a number of chromatographic inte rferences 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 temperatur e and analyzer temperature.

The analysis of decaBDE in environmental samples has apparently improved over the past decade, as demonstrated by a recent in ter-laboratory study undertaken to validate and harmonize the analytical methodology (L eonards and Duffek 2008). In total, ten routine laboratories participated in this evaluation of decaBDE in dust and sediment, including nine laboratories from Europe a nd one from Canada. Overall, the evaluation found acceptable accuracy in resu lts 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 sc reening assessment report (Can ada 2006; Environment Canada 2006b). This summary is followed by an analysis of new data available after 2004 and an interpretation of their signif icance in relation to bioaccumu lation 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 μ g/kg lipid measured in fat from Vancouver Island marmot ( Marmota vancouverensis ).

• European Communities (2002) a nd Law et al. (2003) summari zed the results of many analyses for decaBDE in fish and marine ma mmals from Europe and elsewhere. It was noted that decaBDE was detected only occas ionally, at a concentration close to the

detection limit of the method used.

• DecaBDE (i.e., BDE209) was detected on ce 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 concen trations in freshwater fish from the

northeastern United States and found that a ll 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 sample s 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 Holm an Island, Northwest Territories, between

1981 and 2000. The levels of decaBDE were e quivalent 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 sa mples showed concentrations of BDE209 below detection limits (ranged from approxi mately 0.8 to 9.0 μg/kg lipid); however, a few samples showed elevated levels in spec ific 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 origin ally 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 Screenin g Assessment of Polybrominated Diphenyl

Ethers (Environment Canada 2006b).



### 2.1.2 New Biota Concentration Data Recently, a large number of studies showing ti ssue concentrations of decaBDE in various global biota have been published. These stud ies are discussed belo w and are interpreted

with reference to bioaccumula tion and biomagnification as defi ned in earlier sections of this report.

Voorspoels et al. (2006a) reported de caBDE concentrations in red fox ( Vulpes vulpes ) from Belgium. Samples of liver (n=30), musc le (n=33) and adipose tissue (n=27) were collected from foxes which were being pro cessed 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, fr om < 3.9 to 290 ng/g lipid in muscle and from < 3.7 to 200 ng/g lipid for ad ipose tissue. The authors concluded that11 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, bi omagnification 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) re ported 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 (bel ow detection limit (B DL)—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 Voors poels 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 musc le samples from seven species of aquatic birds and terrestrial predatory bi rds 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 buzza rd. 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 speci es might be more exposed to decaBDE than aquatic species.

Christensen et al. (2005) repor ted 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 BDE 209) ranged from approximately 1 to 2.77 μ g/kg lipid in coastal bears and from approximately 0.5 to 41.71 μ g/kg lipid in interior bears (concentrations read from graph), sugges ting a higher exposure to decaBDE for bears feeding in interior environments where diets are predominantly terrestrial in nature. The

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

bioaccumulation of decaBDE in coastal and inte rior 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, resp ectively, 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 decaB DE was “bioaccumulative” to a limited degree. However, this claim was based sole ly on the occurrence of decaBDE in tissues

without an evaluation of whether the concentr ations 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. Th e authors noted a signi ficantly increasing temporal trend in decaBDE concentration in the falcon tissues. They also concluded that their results provided evidence of some upt ake and bioaccumulation of decaBDE (ruling

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

characterized decaBDE concentrat ions as low in falcon eggs.

S Ø rmo 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 dete cted in only 1 of 6 samples of ringed seal ( Pusa hispida ) blubber, at a concentrati on 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 co ncentration 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 mu scle, 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 li pid 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 es pecially elevated compared to other published reports and that this may be related to signifi cant production, usage or disposal13 of decaBDE-containing products in China. Th ey also suggested that exposure to decaBDE could be higher in terrestrial f ood 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, fo r 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 pr oduction 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 dete cted (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 blubbe r 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, re spectively, 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) conducte d monitoring of bird eggs an d 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 Niagar a rivers. BDE209 concen trations 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 concen trations 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 leve ls 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 quan tified) amounts in eggs of birds of prey from northern Norway (Herzk e 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 a nd plasma, but the authors speculated that the presence of detectable c oncentrations of octa- and nonaBDEs in the samples suggested accumulation and subsequent in vivo debromination of decaBDE. Although the detection limit for decaBDE was no t 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 concentrat ions 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 pe riod and these concentr ations 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 (approxim ately fourfold) in lake trout BDE209 occurred between 1998 and 2004. The study also found a significant incr ease in BDE209 concen trations 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, gill s, gastrointestinal tract, liver and muscle samples. However, the authors suggested that this detecti on frequency was likely underestimated due to the much higher anal ytical 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 sa mples of skin, gills, gastrointestinal tract, liver and muscles for all fish), respect ively. 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 gill s and gastrointestin al tract since the lipid-normalized BDE209 concentrations in gi lls and gastrointestin al tract from their15 study were actually slightly hi gher than those in liver a nd muscle. The study concluded

that BDE209 is accumulative in fish tissues under natural environments, and speculated that BDE209 may bioaccumulate significantly, particularly in biot a 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 c ontribution to total PBDEs measured in egg samples was 5.9% a nd 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 wa s 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 si gnature 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). Concentratio ns ranged from 1.4 to 420 ng/g ww for BDE209. The authors reported th at 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, BDE 209 concentrations were shown to increase significantly (r = 0.348, p < 0.005), with a doubli ng time of 5 years. In addition to BDE209, eight nona- and octaBDE congeners we re 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 leve ls and congener patterns in avian species by using stored historical ti ssue 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 samp les, 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 re latively higher propor tions of octa- and nona-BDE congeners were also f ound compared with other birds

.16 Additional studies are also de scribed in Section 2.2. These studies contained sufficient information for conducting a ratio-base d evaluation of bioaccumulation or biomagnification.



### 2.1.3 Synthesis of Biot a Concentration Data

The detected concentrations of decaBDE in a wide range of biot a provide confirmation that this substance can accumulate, at least to some degree, in organisms. However, these data provide little evidence that deca BDE may be bioaccumulative. The main shortcoming is that the data fail to compare th e measured concentratio ns 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 s ubjectively judge whether the observed concentrations in top predators appear hi gh, indicating potential trophic magnification.

Concentrations of decaBDE have been obs erved to be relatively high (e.g., exceeding 100 ng/g lipid) in some top predator species, including

• birds of prey, especially kestrel, spar rowhawk and owl from China; peregrine falcon, sparrowhawk and kestrel in th e UK and Sweden; peregrine falcon in Greenland; and buzzard in Belgium;

• red fox in Belgium;

• sharks of coastal Florida;

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

• some marine bird species, including her on from Sweden and glaucous gulls from Byoroya and Svalbard, Norway.

Of these, the Greenland peregrine falcon result s and the coastal Florida shark results may provide the strongest eviden ce 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 re sult, there is a potential that the organisms were exposed to locationally specific contamin ated environments that may explain their relatively high loading of decaBDE. A lthough high decaBDE concentrations are observed in Florida sharks, similarly high con centrations are not obser ved in teleost fish species from the same ecosystem, which woul d likely make up the prey species for shark. While trophic magnification is a potential expl anation 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 specie s inhabit decaBDE hotspots close to industrialized areas; and17 • 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 concentratio ns in some biota, overall the data from surveillance studies directly contradict a gene ralized conclusion that the concentration of decaBDE is relatively high in top pred ators. The surveillance data show

• relatively low and often non-detected concen trations of decaBDE in marine fish species, even though some of these speci es 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 c oncentrations of decaBDE in mammalian predators such as grizzly bear, polar bear and lynx which feed n ear the top of their respective food webs;

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

mammals which feed near the t op 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 concen trations 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 highe r 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 prov ide 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 re specting decaBDE bioaccumulation considered in the PBDE screening assessment report (Environment Canada 2006a; Environment Canada 2006b), then examines new data av ailable after 2004 and its significance.



### 2.2.1 Information Evaluated in the Screening Assessment

• The Japanese Ministry of International Tr ade 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 Commun ities (2002)). Given the exceptionally low water solubility limit of decaBDE it is not expected that this substance will be appreciably taken up from the wate r phase by aquatic organisms.

• Stapleton et al. (2004) expos ed juvenile common carp ( Cyprinus carpio ) to decaBDE (> 98% purity as reported by Cambridge Is otope Laboratories) amended in food on a daily basis for 60 days (d) (940 ng/d/fish) fo llowed 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 pe riods (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, s uggesting transformation of decaBDE. Their results suggested that at least 0.44% of decaB DE 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 (com position not determined in this study, but reported as 77.4% decaBDE, 21.8% nonaBD E 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, amounti ng to about 0.02–0.13% based on the muscle

concentrations of the total hexa- to decaB DEs 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 lo w bioaccumulation potential in mammalian species. For instance, Norris et al. (1973, 1974) dosed male and female rats with 1.0 mg

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

the administered 14 C-labelled DecaBDE was excreted in the feces within 24 hours (h), and by 48 h, all of the administered chemi cal had been excreted. Tissue accumulation

studies in which rats were fed diets of decaBDE at a ra te of 0.1 mg/kg body weight per day showed that bromine contents in various tissues were not signifi cantly 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 cont rols (Norris et al. 1973, 1974).

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

Additional studies were also reviewed in th e screening assessment but their findings did not allow estimates of bioaccumulation paramete rs. 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 publishe d 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 reporte d 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 we ight 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 peri od. Blood was sampled from individual fish at the initiation of exposure and at three time points duri ng the final 3 months of the19 study. Samples of blood serum, intestine, li ver and carcass at each sampling time were each analyzed for decaBDE and lower brominated BDE congeners.

The net uptake of decaBDE during the experi ment 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 d ecaBDE 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) a nd carcass (5.3 ng/g ww). The detection limit in this study was 1 ng/g ww. Several hept a-, octa- and nonaBDE congeners were also accumulated, potentially, as a result of decaB DE debromination (fish were dosed only with decaBDE, and background decaBDE concentr ations in the control fish ranged from < 0.5 ng/g to 0.5 ng/g). These debromination resu lts are discussed further in Section 3.1.

Using concentration data from day 112, the repo rted lipid content of the test fish, and the known decaBDE concentration/lip id 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)

Food 1 Carcass 1 Serum 1 Liver 1
DecaBDE concentration (ng/g lipid) 9307 204 3300 2 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 frac tion of decaBDE in total BDE burden.

It is expected that the highe r lipid content in food (10.1% li pid) 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, th e 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 transformati on 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 repor ted net uptake of neutral BDEs only would underestimate the actual total upt ake of decaBDE. If metabolites other than hepta-, octa-, and nonaBDEs were being formed and persistin g 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 Labor atories, all purities > 96%) plus DecaBDE (technical grade, Great Lakes Chemical Co rp., 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-expos ure treatments. For the low-exposure treatment, depuration of chemical during the elimination phase was non-detectable (slope not significant) and the absorp tion 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 BMF 1 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 estimat es because similar congeners were also present in the spiked food or potentially pr esent as impurities in the technical-grade DecaBDE used in the study. In addition, the st udy contained other uncertainties such as the use of fiberglass aquaria which may have re sulted in some adsorption of test material.

Ciparis and Hale (2005) examined the bioa vailability and accumulation of multiple PBDEs, including decaBDE, from sediments a nd 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 elimin ation phase. Following solvent extraction, clean-up in a size exclusion column, and furt her purification using solid-phase extraction columns, decaBDE was quantified from substr ates and tissues on a gas chromatography (GC) device equipped with a halogen-selective el ectrolytic conductivity detector with MDLs of 190 ng/g and 20 ng/g for tissu es and substrates, respectively. Although significant accumulation of lower brominat ed 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 th at the bioavailability of decaBDE was limited by its high log K

ow (suggesting that desorption from sediment particles is minimal) and large molecular size, whic h may impede its tr ansport across cell

membranes.

Burreau et al. (2004, 2006) reported the result s for three separate food web monitoring programs for PCBs and PBDEs. Burr eau 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 archipela go in the Baltic Sea

1 BMF = α F/k d where α is the absorption efficiency, F is the feeding rate on a lipid basis and k d is the total

elimination rate constant21 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 we re 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 concentratio ns of decaBDE in freshwater and marine biota from the Lumparn Estuary, Baltic Sea a nd 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 biomagni fication 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 invol ved a regression of lipid-normalized concentration vs. δ 15 N according to the following model:

2 The disparity between the n of samples and the n of anal ysis is explained by the fact that only the samples in which the congener was detected are reported.22 C = A• e (B• δ 15N)

Where C is the biota concentration (lipid-n ormalized), A is a constant representing δ 15 N at the base of the food chain and B repr esents the “biomagnification power” of the substance. A positive B-value indicates biomagnification while a negative B-value indicates trophic dilution of ch emical concentrations. B is similar to a TMF except that a TMF is expressed on an arithmetic, rather th an logarithmic basis, and the TMF is based on a regression with trophi c level (estimated from δ 15 N) rather than the δ 15 N content

itself. The B-values for both food webs were not significantly different from zero, indicating that biomagnification of decaBDE di d not appear to be occurring in these food

webs. Failure to detect decaBDE in salmon fr om 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 pred ator-prey combination, suggesting that biomagnification was taking place in some pr edator-prey combinations. However, it is important to consider that the exact feed ing 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 samp les create uncertainty in the overall23 biomagnification analysis. Currently, this app ears to be a common issue with field studies

of decaBDE in biota.

Shaw et al. (2009) studied the bioaccumulati on of PBDEs in northwest Atlantic marine

food webs. To evaluate the tran sfer 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 o ff the coast of Maine during the May–June 2006 Gulf of Maine Trawl Survey of commercia l 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 pr ior 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 standard s, a mixture of cyclohexane and dichloromethane was applied to the colu mn for extraction of PBDEs along with other lipophilic compounds and fat. The extract was washed, dried, and after solvent evaporation, gravimetric lipid determinat ion was performed. The final extract was evaporated by a stream of n itrogen to a final volume of 50 μ L containing C 13 labelled BDE139 as recovery standard. The measuremen ts were performed by high-resolution gas chromatography / high resolution mass spectrometry (HRGC/HRMS).

Total PBDE concentrations in fish range d 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 ra nged 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 mack erel) to 1.3 (white hake), which the authors suggested represented low biomagnificati ons potential. This contrasted with BMFs for the other PBDEs, which averaged from 17 to 76.5, indica ting high biomagnification in this marine food web. The authors suggested that the pr esence of higher brominated congeners, including BDE209, at measurab le levels in fish and s eal tissue, along with high biomagnification of BDE153, -155, and -154, su ggests 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 th e south basin of the lake ne ar Gimli, Manitoba. Muscle24 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 hor izontal 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 onl y the surficial 2 cm of sediment retained. Water was sampled in 2004 using a Teflon colu mn 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- μ m pore size) and then onto a

XAD-2 column.

All samples were analyzed for decaBDE (a nd several other chemicals) using GC/MS, with additional analyses of organic ca rbon (OC) for sediments and lipid and δ 15 N for

biota. The δ 15 N measurements were used to estimate trophic position. The detection limit of the analytical method used was 0.1 μ g/kg for biota and sedime nt 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 δ 15 N 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 pr edators). A regression of the lipid-normalized concentration of decaBDE vs. trophic level was used to estimate a TMF of 3.6 3 (r 2 =0.46p=0.0001) for decaBDE in the pelagic food web. Predator-prey BMFs (on a lipidnormalized basis) were also calculated usi ng 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-w orld indication of trophic magnification and biomagnification of decaBDE in an aquatic food web, it is important to consider some of the uncertainties asso ciated 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, biomagnificat ion was identified using lipid-normalized data; however, certain tissues were characteri zed 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 ba sis. When biomagnificat ion 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, pr otein binding is not expected. It is possible that decaBDE is subject to non-specific bindi ng in blood plasma (i.e., lipid s) (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, pref erential accumulation in liver could occur, but this has not been established definitively.

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

carbon content and δ 15 N in water, sediments and biota from Lake Winnipeg

Sample δ 15 N Trophic position Mean lipid 1 or organic

carbon 2 content Average decaBDE

concentration (ng/g lipid 1 ; ng/g dry weight (dw) 2 or pg/L 3 )

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 samples26 The United Kingdom (2007a) also highlighted th e following issues regarding the Law et al. (2006) study:

• There is some uncertainty as to how well the δ 15 N analysis characterizes trophic levels. The resulting trophic structure does not necessarily match that which would be expected based on species size and feeding ch aracteristics (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 sa mples represent the food web, since sample sizes were small in some cases, and they we re collected at varying times between 2000

and 2002.

• Information regarding the act ual 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 PB DE 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 ex tracted using a ball mill shaker with anhydrous sodium sulphate and hexane:dichl oromethane (DCM). Each cell was spiked with recovery internal standards, then shak en and left to stand 2–4 h before centrifuging, then decanted. The extraction was repeated tw ice, combining the decanted extracts. Fish, shrimp and clam tissues were homogenized wi th dry ice in a laboratory blender, then stored overnight in a –20°C freezer to allow for sublimation of the CO 2 . Thawed tissue was weighed and mixed with pelleted diatom aceous earth (Hydromatrix) then added to a cell along with the recovery internal standa rds RIS and extracted using an accelerated solvent extractor (ASE 300). Zooplankt on samples were weighed frozen and homogenized by directly mi xing with Hydromatrix prio r to accelerated solvent extraction. Void space was filled with sand. Af ter extraction, anhydrous sodium sulphate

was added to the collection bottles to remove water. Extracts were reduced in volume and filtered. Lipid content was determined gravimetri cally in an aliquot of extract, while lipid

was removed from the remainder of the extr act 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 volum e to 200 μL and instrument performance internal standards (10 μ L 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 so lution containing the BDE-mix (36 BDE congeners) and BDE-209 for quantification. Proc edural and instrument blanks were used, and all PBDE samples were blank-corrected. The relative trophic le vel of the organisms27 was determined by stable isotopes ( δ 15 N), 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 concentr ation 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. BDE 209 mean concentrati ons ranged from nondetect (walrus, beluga and narwhal) to 18.7 lipid ng/g (zooplankt on). BDE209 was found to contribute significantly to the body burden of total PBDEs in th e lower trophic level organisms: 60% in redfish and 75% in arctic cod, for example. Conversely, in the upper trophic level organisms like beluga and narwha l, BDE209 accounted for less than 2% of the total BDE burden, whereas BDE47 account s for over 40% in these samples. The authors suggested that elevated BDE209 leve ls in lower trophic level organisms may reflect greater exposure to this compound through zooplankton, and limited metabolic capabilities with respect to BDE209. BD E209 lipid-adjusted BMFs for individual predator–prey relationships were as follo ws: beluga (blubber) :cod (whole body) < 1; beluga (blubber):redfish (whole body) < 1; and cod (w hole body):zooplankton (whole body) < 1. BMFs were not calculated for narwha l:redfish and narwhal: cod. A statistically significant TMF of 0.3 was estimated for BD E209 (p=0.002; correlation coefficient R 2
=0.25). The authors concluded that BDE209 con centrations decreased with trophic level, suggesting metabolic depletion of this conge ner or reduced assimilation up the food web. The authors suggested that thes e results are consistent with other studies that have shown that BDE209 is not an abundant congener in higher trophic level or ganisms. 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 Arc tic 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 a nd Oceans Canada. The brominated compounds were measured in the blubber of ringed s eal and beluga, in the whole organism minus liver for the pelagic fish, and pooled composite s 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 13 C

12 - 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 intern al standard, respectivel y. Stable isotopes of nitrogen, expressed as δ 15 N, were analyzed at the stable isotope laboratory at the University of Winnipeg (Manitoba, Canada). BM Fs were determined as per Tomy et al. (2008) (see above study).

The relative trophic level status of the studied individual or ganisms was established using stable isotopes of δ 15 N. This analysis indicated the fo llowing food web: beluga whale > ringed seal > Arctic cod > P acific herring and Arctic cisco > pelagic amphipod > Arctic copepod. The total concentration of the se ven congeners analyzed (BDE47, -85, -99,28 -100, -153, -154 and -209) ranged from 2.6 ng/g lip id in ringed seal s to 205.4 ng/g lipid

in Arctic cod. The median concentrations of BDE 209 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: ci sco > Arctic cod > herring > beluga and ringed seal >

themisto > calanus. The researchers calcula ted 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 f ood chain, there is some uncertainty in the results. For instance, it appears that the orga nisms included in the study were collected at

different times and locations. None of th e PBDEs showed a sta tistically significant positive relationship with trophic level, and no statistically significant TMF was found

for BDE209.

Wu et al. (2009) evaluated the biomagni fication of PBDEs, including BDE209, in a highly contaminated freshwater food web from southern China. W ild aquatic species representing different trophi c 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 th e 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 we re measured, then stored at -20°C until further treatment. For chemical analysis, samples were thawed and whole-body homogenized, after which two sub-samples we re taken from each specimen: one for PBDEs determination, and one for nitrogen st able 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 hour s. 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 de viations (RSD) of sample triplicates were less than 24% for BDE209. Stable isotopes of nitrogen, expressed as δ 15 N, 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-norma lized 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 organism s, different environmental conditions (e.g., warmer water temperature), and the differe nt 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 K OW was found for PBDE s, and the authors29 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 com posite samples were analyzed for PBDEs. Samples were homogenized, extracted and an alyzed 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 tr eatment plant (WWTP) located in Roxboro, North Carolina. All samples were extract ed and purified us ing size-exclusion chromatography and then analyzed for PBDEs using GC/MS in electron capture negative ionization (ECNI) mode and electron ioniza tion (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 μ g/kg organic carbon (OC), 2880 μ g/kg lipid and 21 600 μ g/kg lipid, respectively. The much higher concentra tion in crayfish was attributed to the sediment-association of this species and the au thors 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 es timate sediment BSAFs of 0.0018 for sunfish and 0.013 for crayfish. Th ese are well below values suggested to potentially indicate biomagnifica tion (i.e., ~1.7 to 3; refer to Section 1.2). It is possible

that a combination of low sediment bioavail ability and/or metabolic transformation could be limiting the bioaccumulation and biomagni fication 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 efflue nt 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) and30 BDE209 were 6.33 and 237.01 μ g/kg dw in sediments. BDE 209 concentrations in lake water and effluent were below the analyti cal detection limit (not given for water; 1 μ g/kg (ww or dw unknown) for sediment and biota) . High concentrations of BDE209 were determined for lichen (1572 μ g/kg dw), march brown ( Limnodrilus hoffmeisteri ;
11.37 μ g/kg dw), coccid (114 μ g/kg dw) and the zooplankton Monia rectirostris, Monia

micrur and Monia macrocopa (151.9 μ g/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 μ g/kg dw.

Bioconcentration/bioaccumulation for BDE 209 was not identified. In addition, the authors found no obvious biomagnification of PBDEs when they analyzed the relationship between PBDE concentrati ons 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, th ey found non-detectable and measurable

BDE209 concentrations in al l biota species. Concentratio ns 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 li pid 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 appare ntly 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 th e 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 containi ng industries produci ng textiles and epoxy resins and involved with polymide polymeri zation. Using concentration measured in sediments and fish, the authors calculat ed 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, Br itish Columbia, Canada near the city of Victoria to evaluate and compare pattern s of PBDE and PCB bioaccumulation. Samples

were collected from 14 stations within 800 metr es of a municipal outfall, and from three reference locations. At each station, thr ee surface sediment samples (0–2 cm) and 1531 randomly selected mussels (> 50 mm length) were collected. Mussels were measured for shell length and width, total we ight, tissue weight, age and sex, prior to compositing and homogenizing tissue samples for chemical an alysis by station. PBDEs were analyzed using HRGC-HRMS. Methods for PBDE determ inations and QA/QC procedures were as per the U.S. Environmental Protection Agen cy (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 measur ed 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 BDE2 09 concentrations were highest near the wastewater outfall (2550 pg/g dw), the conge ner 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 range d from “not detected” to 5305 pg/g dw (wastewater outfall station), and were above th e 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 coul d 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 (referenc e). 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 th e effect of BDE209 on freshw ater bivalve zebra mussels ( Dreissena polymorpha ) under laboratory conditions. Th e primary objective of the study was to investigate the potent ial 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 ligh t/dark photoperiod, with consta nt temperature (20°C) and oxygen (> 90% saturation), and fed a suspension of Pseudokirchneriella subcapitata. Approximately 150 acclimated mussels per BD E209 concentration were selected for exposure tests. DecaBDE (98% purity) was di ssolved in isooctane/toluene (9:1 v:v) mixture, dissolved in dimethylsulfoxide (D MSO), 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 photodegra dation. Water was changed daily and mussels were fed 2 hours prior to water renewal. C ontrol groups of mussels receiving either freshwater or a solvent blank (D MSO) were also monitored.

The mussels reached a relatively constant con centration of decaBDE in their tissues after 48 hours of exposure, for each treatment concentration. The United Kingdom (2008) interpreted the graphical repres entations 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 presen ce of lower brominated congeners in the mussels after 168 hours exposure. These were not determined quantitatively but they32 were interpreted to have consistent GC an alysis 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 retardan ts (BFRs) from females to their eggs. The

tested BFRS included the PBDEs BDE 28, BDE183 and BDE209. The tested BDE209 compound was synthesized at the Department of Environmental Chemistry at Stockholm University, and labe lled standards (e.g., 13 C) 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 BF Rs 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 eva porate. 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 anal ysis. Eggs were coll ected 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 e ggs 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 lipi d). Maximum BDE209 c oncentrations in high-dose female zebrafish were measured af ter 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-dos e 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 BD E209 transfer) for the low-dose group. The authors suggested that the hi gh 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 formulati on (DE-83R 98.5% purity; 0.3 μ g/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 gr oups of three for tissu e analysis (liver, gastrointestinal tract, plasma, and remaini ng carcass) on days 0, 3, 7, 10, 14 and 21 of the33 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 subtra cted from the PBDE determin ations 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 anal yzed for a suite of PBDEs, including heptato decaBDE congeners.

Based on the analytical results, it was estimated that only 5% (or 3.6 μ g) of the total decaBDE dose was retained in rat tissues following the 21-d dosing period, while approximately 50% was excreted to feces dur ing this time. In addition to decaBDE, the authors concluded that one nonaBDE conge ner (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 fo r 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 concentrati on of decaBDE a “BCF” (analogous to a

BMF 4 ) of 0.05 (on a ww basis, justified by the fa ct 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 tissu es during the 21-d exposure pe riod 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 coul d 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 elimina tion 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-orde r decay equations were found to represent the data well, with relatively rapid di stribution phase half-lives of 0.7 and 1.2 d, respectively, but longer elimination phas e half-lives of 20.2 and 75.9 d, suggesting potential persistence of decaBDE in rat ti ssues 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 excr etion of PBDEs administered for 21 d as

either a dust reference material or as a co rn oil solution. The dust reference material (NIST Standard Reference Material 2585), c ontaining a characteri zed and homogeneous composition of PBDEs, was mixed with ra t 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 mixe d 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 hour s after their last feeding. To assess whether rats were approaching a stea dy-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 remaini ng 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 burde ns were reached, only epididymal fat was sampled and analyzed. This analysis showed th at 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 a nd 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 ge nerally differ between the dust and oil treatment groups. The study did not consiste ntly detect nona- and decaBDEs in the adipose tissues above that of the contro ls. Urine contained less than 0.3% of any congener. Fecal excretion was the major rout e of elimination and was described in the study as that component of the dose not adso rbed. 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 co mpound ranged from 28% to 31.9%. Metabolic transformation products could have accounted for some portion of these percentages. Derived BCFs (analogous to BMFs as define d in this study) for adipose tissues were inversely related to the degree of bromina tion, 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 ar e 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) re ported the findings of a 3-m onth feeding study with dairy cows. The study was originally undertaken to measure the long-term mass balance of PCBs; however, archived samples from two co ws and feed were subsequently analyzed for a range of PBDEs. Over the 13-week study period, the cows we re kept indoors and milk and feces were sampled once per week. The feed consisted of silage, concentrate and a mineral supplement which was not delibera tely spiked with PBDEs. As a result, the PBDE concentration in food generally repres ented “background” contamination levels.

The milk and feces samples were pooled according to the following scheme to obtain a series of 5 composite samples representing di screte portions of the 13-week study period: three 3-week composites and two 2-week comp osites. In addition, one of the cows was35 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 thr ee intervals during the 13-week study, and concentrate and minera l samples were anal yzed once during the study. All samples/composites of feed, milk, feces and tissue we re analyzed for hepta- to nonaBDEs using HRMS and were analyzed for decaBDE using LRMS in negative chemical ionization mode (l imit of quantificatio n = 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 influen ced 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 tissu es. Based on these observations, the authors proposed that the cows were in a state of PB DE elimination rather than accumulation and that the observed PBDE concentrations may ha ve 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 incr ease 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 aff ect the overall mass balance c onducted on the PBDEs in this study.

The Government of the United Kingdom (United Kingdom 2007a) conducted a critical analysis of the Kierkegaard et al. (20 07) 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. Th ey concluded that the high variation in silage PBDE con centration led to cons iderable uncertainty in the study results with respect to bioaccumulation. To es timate accumulation factors and BMFs on a lipid-normalized basis, they assumed a 4% li pid content in silage. Table 2-5 summarizes

the calculated accumulation factors and BM Fs for nona- and decaBDE congeners from the Kierkegaard et al. (2007) study.

The calculated BMFs and accumulation factor s were well below 1 for decaBDE and only exceeded 1 for BDE207. For BDE207, it is un certain whether the estimated values represent direct accumulation of BDE207 from food, or accumulation combined with

bioformation as a result of the debromina tion 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 fr om 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 pr esence of polar metabol ites, bound residues and36 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 cal culations by United Kingdom (2007a) using the findings of Kierkegaard et al. (2007) for one cow

Congener Parameter

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 herri ng continued (1–2.5 kg/ d), with all fish obtained from a single batch caught in the North Sea. The second month of the study37 involved a decaBDE exposure phase with the diet supplemented with 12 μ g decaBDE per day, dissolved in a cod liver oil capsule. For the final mont h 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 peri od, while blubber biopsies were taken and analyzed for PBDEs 3 times during the study (beginning, 3 d after in itiation of exposure phase, and end, after 29 d on a decaBDE-free diet).

The blood decaBDE concentration increased fr om 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 e nd of the exposure phase. Concentrations of decaBDE in blubber ranged from non-detectabl e 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 se als, 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 relativel y high (i.e., exceeding 1) BMF for blood and relatively low BMF for blubber. The rela tively high blood-based BMF suggests that significant magnification could be occurring from the diet to blood and related tissues, indicating relatively high bioaccumulati on potential. The blood BMFs reported by Thomas et al. (2005) may be somewhat uncertain because of other non-lipid constituents (proteins) that can be attr ibuted 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 di et) 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 thresh olds for chemical absorption. The relatively high apparent absorption efficiency was attributed to the following factors:

• Large fat reserves (i.e., blubber) which provi de a reservoir for lipophilic chemicals.

• High food absorption efficiency in homeothe rmic 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 degrada tion 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 μ g/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 potentia l formation of bound metabolites and/or the formation and subsequent excretion of pheno lic 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 th e formation of phenolic and methoxylated metabolites may be an important transf ormation pathway for accumulated decaBDE. Thus, if phenolic or methoxylated metabolites we re present but not analyzed in the feces, it is possible that the output flux was underestim ated, resulting in an overestimate of the absorption efficiency of decaBDE. Anothe r potential uncertainty relates to an inconsistency in iden tified dosage rate. While the hi gh proportion of total ingested decaBDE estimated to be present in blubbe r 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 concentrati ons 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 reas onable approximation for the BMF estimates

summarized in Table 2-6. Note that the bl ood concentrations do not appear to have reached steady state during the e xposure phase and it is likely that this also applies to the blubber concentrations. Furthermore, the blubbe r 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 underes timate the steady-state BMF to an unknown degree.39 Sellström et al. (2005) analyzed decaBDE in soil and earthworm (sp ecies not identified), samples collected in 2000 from three research stations (with reference plots and sewage sludge amended plots) and two farms (refere nce and amended/flooded soils) in Sweden.

Soil decaBDE concentrations at the vari ous 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 si te-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 de caBDE 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 fo r 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 ra tionale 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 di et increased. A positive bioaccumulation slope indicates trophic magnificati on while a negative bioaccumu lation slope indicates that trophic dilution was taking place. The authors indicated that the bioaccumulation slope for decaBDE was negative (albeit not significan tly different from 0) but did not provide the value. The lack of a significant positiv e bioaccumulation slope for decaBDE suggests that decaBDE was not undergoing trophic magnifi cation 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 pr edator and prey species. Unfortunately, the high number of non-detected concentrations pr ecluded 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 concentr ations 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. Th e BAF–QSAR model described by Arnot and Gobas (2003) has generic appli cability 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 mode l predicts both BAFs and BCFs for three representative fish trophic levels (low, middle and upper) in a generic aquatic food web40 based on a standard set of conditions found in the Canadian environment. For the terrestrial environment, Gobas et al. (2003) describe a terr estrial 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 K

oa (logarithm of octanol-air coefficient), chemical log K

ow , and a set of parameters desc ribing 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 ch emical elimination, allowing for metabolism correction of BAF, BCF and BMF predictions based on field or labor atory 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 K

ow values for decaBDE revealed that a log K

ow of 8.7 reported by Wania and Dugani (2003) is c onsidered to represent the most reliable

value available. For further rationale regarding the selection of log K

ow values, refer to Appendix C. Two prediction scenarios were c onducted: 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 (k T ) of approximately 0.027/d. The Tomy et al. (2004) value was used to

derive an in vivo -based metabolic rate constant (k M ) according to the method of Arnot et al. (2008b). In this method, when k T is available, k M is derived according to the following equation:

k M = k T - (k 2 + k E + k G )

where:

k M = the metabolic rate constant (1/d);

k 2 = the elimination rate constant (parameter ized using data from Arnot et al. 2008a);

k E = fecal egestion rate constant (parameteri zed using data from Ar not et al. 2008a); and

k G = 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 k M to account for error associated with the in vivo data (i.e., measurement variability, parameter estimation uncertain ty 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 k M was further normalized to 15 o C and then corrected for the body weight of the middle tr ophic level fish in the Arnot-Gobas model (184 g) (Arnot et al. 2008a). The middle trophi c level fish was used to represent overall model output and is most repres entative of fish weight likely to be consumed by an avian or terrestrial piscivore. After normalization routines, the k M 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 K

ow of approximately less than 4.5.

Table 2-7: BAF and BCF predictions for d ecaBDE using the Arnot-Gobas kinetic model (v1.11)

k M (metabolismcorrected; days) Log K
ow 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 me tabolic transformation, ranges from 4056 to 161 618 depending on the rate of metabolism. The predicted BAF for an average k M , 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 magn itude higher than the BAF calculated with average metabolic rate potential. These results demonstrate th e influence of both chemical partitioning behaviour (i.e., log K

ow ) and metabolic transformation on the bioaccumulation potential of decaBDE.

The metabolism-corrected BAFs probabl y provide the best estimate of the bioaccumulation potential of decaBDE since metabolic transformation of decaBDE has

been demonstrated or inferred in most laborato ry studies. It is importa nt to note that these

corrected BAFs could underestimate the tota l 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-parameter ized for dietary assimilation efficiency (E D )

and k M , the default settings of which are 90% and 0%, respectively. In addition to the range of potential log K

ow values described for the BAF–QSAR model in Appendix C, two log K

oa estimates were available: 15.27 (Tittlem ier et al. 2002) and 18.423 (predicted by the QSAR model, KOAWIN). Howeve r, the BMF predictions do not vary significantly for log K

oa or log K

ow values in these ranges, an d systemic variation of log

K
oa and log K

ow made little difference in the predic ted output. Four prediction scenarios

were conducted: the first w ith no correction for metabol ic transformation, and the remaining three corrected for metabolism base d on the laboratory observations of Huwe and Smith (2007a, 2007b). They determined a ra nge of potential half-lives for decaBDE

in rats based on a combination of first- order and second-order approximations to elimination data for carcass, liv er and blood plasma. For carcass, a first-order half-life of 8.6 d was found to best represent the elimin ation data. For blood plasma and liver, second-order models (i.e., with distribution and elimination phases) provided the better42 fit, with elimination phase half-lives (repr esenting slower elimin ation from residual body stores) of 75.9 and 20.2 d, respectively. Thes e half-lives were used to infer k M values using a similar method as with the BAF–QSAR model, resulting in k M 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 k M 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 decaB DE are summarized in Table 2-8. An E D of

~56% was calculated as outlined in Kelly et al. (2004) using the E D model for humans as

the best estimator for carnivores. In the ab sence 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 k M used in the model (Tab le 2-8). These metabolismcorrected and assimilation-efficiency-correct ed 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 transf ormation of decaBDE has been demonstrated or inferred in most laboratory and captiv e 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 unde restimate the total chemical biomagnification related to decaBDE since th ey 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 calculati ons, 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)

k M (wolf BW normalized) (days) Log K
ow used Log K
oa 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 av ailable regarding the bioaccumulation and biomagnification of decaB DE. It is intended to synthesize the existing state of the science regarding wh ether 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 cons iderations and interpretations on the capacity for decaBDE to bioaccumulate and/or biomagnify are also summarized in a third section.

1 . Equivocal 5 evidence with respect to whether d ecaBDE 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 rela tionships, suggesting that biomagnification may be occurring at these lower trophic leve ls in a western Arctic marine food web. However, BMFs for other feeding relations hips within the sa me food chain were below 1 for higher trophic-level feeding relationships and th e calculated TMF was also below 1. The study also has some uncer tainty since organisms were sampled at

different times and in different locations.

• DeBruyn et al. (2009) calculat ed 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 be low 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 BS AFs was calculated as 3.53, whic h suggests that, at this location, decaBDE may have bi omagnified 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 (lip id). However, the study was conducted as a genotoxicity study and many study attribut es relevant for an evaluation of bioconcentration were lacking (e.g., measur ed 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 fu ll 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 considere d, 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 ha d 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 we bs and found that biomagnification of decaBDE did not appear to be occurring. Fa ilure to detect decaBDE in salmon from the Atlantic Ocean precluded a similar an alysis for this food web. In addition, the relatively high levels of decaBDE in proce dural blanks and low concentrations of decaBDE in biota samples created uncer tainty 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. Th e 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 Chin a and Europe), red fox, shark, marine mammals and a few marine bird species, these high observe d 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 c oncentrations 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 a pproximately 4000 to 162 000 and reflect the uncertainty associated with estimating me tabolic 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 mode l, metabolism-corrected terrestrial BMF predictions can range from slightly above to slightly below 1 de pending on the rate of

metabolism assumed in the wolf model. Th e moderately high dietary efficiency of 56% contributes to a BMF greater than 1 wh en 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 we b. Any comparisons involve uncer tainties due to different species with different metabolic rates, a ssimilation efficiencies and body weights.45 2 . Evidence that does not support 6 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) obt ained a derived BMF of less than 1. A full accounting of parent decaBDE and meta bolites 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 a nd a very low BMF (0.03) for the polar cod / ice amphipod predator-prey relationship. Ot her BMFs could not be calculated because of the high frequency of samples in which decaBDE was not detected, suggesting a low potential for biomagnifi cation 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 BSAF s were well below the threshol d 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 sp iked sediments (Ciparis and Hale 2005).

• A study by Christensen et al. (2005) on griz zly bears suggested that biomagnification is not occurring.

• A feeding study by Tomy et al. (2004) usi ng juvenile lake trout determined a BMF of 0.3, using the kinetic method based on pare nt 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 solubi lity of decaBDE and the detection limit in tissues, and therefor e 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 (S tapleton et al. 2006; Kierkegaard et al. 1999).

6 Studies cited here indicate that decaBDE does not ha ve the potential to bioaccumu late or biomagnify in the environment; these studies include bo th 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 reten tion (generally less than 5%), low bioaccumulation potential in mammal ian species, and rapid excretion.

• Very low chemical assimilation efficien cies in fish suggest that assimilation efficiency in carnivorous mammals ma y be lower than predicted by the E D model for humans. However, the predicted E D value is based on a relationship using multiple halogenated organics and so is appropriate for modelling decaBDE. Fish do show a significantly lower E D than mammals at a log K

ow of 8.7 (conservatively estimated at 56% vs. ~0%) using the relationships summa rized in Kelly et al. 2004). Thus, high assimilation of halogenated aromatics by ma mmals 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 va lues could be offset by a high rate of gut

metabolism to lower brominated and hydroxylated bio-transformation products,

which would limit the E D vs. model predictions of E D

.



### 2.3.2 Conclusion Respecting Bioaccumulation The existing evidence for the bioaccumulati on of decaBDE does not support a conclusion

of “bioaccumulative” as defined in the current Persistence and Bioaccumulation Regulations under CEPA 1999. While most availabl e 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 mode lling undertaken to support this evaluation, however, shows uncertainty associated with metabolism in fish, as model-predicted aquatic BAFs range from below the 5 000 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 re levant than BAF or BMF, experimental

BCF measures are below the 5000 criterion. The substance is shown to be increasing in

concentrations in some wildlif e species, and some data suggest that decaBDE has reached concentrations in some organisms interprete d to be high. DecaBDE (i.e., BDE209) is also considered to be contributing to the bioaccu mulation potential of total PBDEs as a result of metabolism to lower brominated forms (t o 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 bioa ccumulative products in organisms (i.e., in

vivo by metabolism) and in the environment (i.e., by abiotic and biodegradation processes). The evidence provided by st udies 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 brominat ed 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 decaB DE. Thus, even if these impurities47 were present in the food (or the decaBDE test substance itself) at concentrations below the detection limit, they c ould still accumulate to measurable concentrations in the organism dur ing the course of the experiment. (b) Analytical challenges, which could infl uence the accuracy of the results (see Section 2). (c) Use of laboratory studies to make infere nces 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 th e Screening Assessment for PBDEs

The following summarizes the evidence regard ing debromination of decaBDE which was available during and prior to 2004 and consider ed in the screening assessment for PBDEs (Environment Canada 2006b).

In a decaBDE feeding study with juvenile car p, Stapleton and Baker (2003) and Stapleton et al. (2004) found that seven transforme d products of decaBDE accumulated in the

whole fish and liver tissues over the exposure period. These we re identified as penta- to octaBDEs with only BDE154 and -155 (i.e., hexaBDEs) being positively identified (personal communications from HM Staple ton to Environment Canada, July 2009; unreferenced). The unidentified hepta- and octaBDE congeners have recently been determined to be BDE188, -202 and -197 (Stapl eton 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, Staplet on 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 decaB DE (parent chemical plus metabolites) was estimated at 0.44%. This value could be highe r if other metabolites were present (not determined in these studies). It should be not ed that the alternative explanation provided by Stapleton and Baker (2003) for the transforma tion of decaBDE, i.e., that it could have

also been due to the action of intestinal mi croflora, has recently been refuted by Benedict et al. (2007), whose study confirmed that carp intestinal flora cannot debrominate

PBDEs.

In their feeding study with juve nile rainbow trout, Kierkega ard et al. (1999) detected several hexa-, hepta-, octa- and nonaBDEs in liv er and muscle tissues of the treated fish and some of these congeners were not meas ured in the food. It was not possible to determine in this study whether their presen ce was a result of decaBDE metabolism or the efficient adsorption of congeners initia lly 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 poten tial in mammalian species. For instance, Norris et al. (1973, 1974) dosed male and female rats with 1.0 mg of 14 C-labelled decaBDE as a suspension in corn oil and found that approximately 90.6% of the

administered 14 C-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 adipos e tissue of decaBDE-dosed rats was found to be significantly increased at the p<0.03 leve l 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) resp ecting the El Dareer et al. (1987) study is provided below (also cited as NTP 1986).

The National Toxicology Program (NTP 1986) a nd El Dareer et al. (1987) evaluated the adsorption of 14 C-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 e xposure, from 91.3 ± 4% to 101 ± 4% of the radioactivity was recovered in the excreted feces. In a fu rther study, the re searchers fed

both unlabelled and radiolabelled decaBDE (~22–25 mg/kg bw) in food (total food consumption 48 000 mg/d). The rats were fe d 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 c ontinued with an intravenous study in which rats were injected with 1.07 mg/kg 14 C-decaBDE and were sacrificed 72 hours after dosing. The

results revealed that 74% of the dose was f ound in the feces and gut contents, suggesting that biliary excretion was important. The aut hors 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 ove r 4 hours was evaluated after intravenous administration of 0.9 mg/kg of 14 C-decaBBDE to rats. The au thors 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 substan ce used in the study was synthesized by a multi-step process involving the bromination of 14 C-labelled phenol ultimately giving

14 C-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 a ssociated 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 “conve ntional” 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) design ed to maximize decaBDE solubility and absorption. Feces were collected for analysis from the conventional and cannulated rats at49 24-hour intervals, while bile wa s collected and analyzed from the cannulated rats at 4, 12, 24, 48 and 72 hours post-dosing. The conventional ra ts were sacrificed at day 3 post-dose (n=4) and day 7 post-dose (n=3) for analysis of tissues.

Radioactivity related to decaB DE was present in the feces, bile and tissues of exposed

rats as a combination of non-extractable residues, lipid-bound re sidues, water-soluble residues, phenolic metabolites and neutra l metabolites/parent compound, indicating the potential presence of parent decaBDE a nd 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 li pid-bound residues (decreasing from ~60% to ~11% in the three days post-dose) and wate r 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 14 C collected in urine and feces and subtracting that from 100%, the authors estimated that 9% remain ed 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; le vels in other tissues were low. The high absorption (relative to previous studies) of decaBDE was likely due to the use of the lutrol/soya phospholipon e carrier. Ten percent of the dos e 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-sol uble, lipid-bound, phenolic metabolites and parent compound/neutral metabol ites. Thus, there is some uncertainty that the 65% included some proportion of bound pa rent compound or metabolites.

To characterize decaBDE metabolites, the aut hors further manipulated and analyzed the phenolic and neutral metabol ites / parent compound fracti ons. Detection was conducted using GC/MS (ECNI). Based on their findings, the authors inferred, a lthough they did not positively determine, the presence of debrominated, phenolic and methoxylated

metabolites. Non-transformed decaBDE domi nated 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 monohydr oxylation followed by s econdary oxidation of the resulting phenol. The resulting ortho catech ol would then be me thylated, potentially by the action of catechol-O-methyltransf erase to form the observed guaiacol50 (hydroxymethoxylated) metabolites. The initia l debromination step would have to remove least 2 bromines from the hydroxylat ed/methoxylated ring but appeared to also remove bromines from the other ring.

The authors also speculated th at 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 reac tive 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 approxima tely 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 untransf ormed decaBDE accounting for the remainder (approximately 68%). In the liver, the burde n was primarily decaBDE with only a small fraction of lower brominated PBDEs (pri marily nonaBDEs). The predominance of BDE202 as a product of decaBDE debromina tion 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 c ongeners increased in concentration over the duration of the exposure. Base d on the total body burden of hept a- 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 th e 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 th e 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 tr out 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 tran sformed 65% of the decaBDE mass (30% was transformed to hexa BDE congeners). The liver microsomes of51 rainbow trout transformed as much as 22% of decaBDE to octa- and nonaBDEs in the

24-h incubation. The controls show ed no transformation of decaBDE.

The authors concluded that their results supported the hypothesi s that deiodinase enzymes were catalyzing debromination of decaBDE; however, they also cautioned that it was not possible to rule out the concurrent or alternat ive action of P450 enzymes. Stapleton’s work shows that removal of br omine atoms occurs preferentially from the meta- or para-substituted positions (pers onal communication from HM Stapleton to Environment Canada, Janua ry 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 he xaBDE, 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 orga nisms. However, it should be noted that each of the PBDE congeners dos ed in this study had purities of > 96%, and thus there is uncertainty rega rding whether these congeners we re in the dosed material as impurities. They suggested that the structural similarity of BDEs to thyroxine (T4) could mean that deiodinase enzymes were debrom inating higher brominated PBDEs to lower

brominated PBDEs. The authors concluded th at the degree biotransformation, especially for decaBDE, was likely to vary considerably between species, l eading to high potential interspecies variability in bioaccumulation.

La Guardia et al. (2007) ex amined the potential for in vivo debromination of decaBDE in aquatic organisms inhabiting the receiving en vironment of a WWTP located in Roxboro, North Carolina, which, based on releases repo rted by industry to the U.S. EPA’s Toxics Release Inventory, was determined to rece ive wastewater from a large plastics manufacturing facility. The PB DE 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) we re collected. Aquatic biota sampling involved the use of minnow tr aps at a location 15 m downstream of the WWTP outfall. All samples were extrac ted and purified us ing 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 s unfish (2880 μg/kg lipid) and crayfish (21 600 μg/kg lipid). The much higher concentration in cray fish was attributed to the sediment-association of this species and the au thors speculated that crayfish could form a link from sediments to pelagic organisms. Th e 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 closel y related to carp, wh ich have previously been demonstrated to have an enhanced cap ability to debrominate decaBDE (Stapleton et al. 2004). Two octa- and three heptaBDE congene rs 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 ef fects 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 Chemi cal Corp.) consisted of greater than 96% BDE209, with BDE153, -183 and -203 dete cted in amounts of between 0.00024 and

0.034% on a mass basis. Further characteriza tion revealed that as many as seven nonspecified PBDEs were qualitatively dete cted (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 w eek with capelin until th e 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 intera peritoneally with a dose of PCB126, and the other half were injected with corn oil al one. The PCB126 was injected to evaluate the impact of cytochrome P4501A (CYP1A) inducti on on the biotransformation of injected PBDEs contained in the decaBDE commercial pr oduct 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 w ith decaBDE (dose 400 ng/g fish; fish from the remaining tanks received a dose of corn oil al one) 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 fis h. These were also measured in the decaBDE product administered in the experiment, and thus, their presence cannot be attributed exclusively to biotransformation. Also, the pr esence of these substan ces could have been due to thermal degradation of the decaBDE product during analysis, or due to transformation in the tank due to fecal eg estion 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 th e 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 importa nce 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 c onsistent with that of Va lters et al. (2005), who found53 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 l east based on their study using BDE99) that it is doubtful that CYPs are responsible for debr omination of BDE99. In their evaluation of BDE99 debromination in carp microflora a nd microsomes, they concluded that an endogenous process that occurs with approxima tely equal activities in the intestine and liver microsomes appears responsible for de bromination 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 elucid ated from their study, but they noted that their work suggests that de bromination may involve thyr oid hormone deiodinases.

Nyholm et al. (2009) used zebrafish ( Danio rerio ) as a vertebrate model organism to study the uptake, metabolite formation and elim ination of 11 structurally diverse BFRs, including BDE28, BDE183 and BDE209 (purit y not reported). Solu tions containing a

mixture of BFRs were made from stock solu tions 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 adde d 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. Zebr afish were fed with BFR-treated or control chironomid feed for a 42-d exposure period a nd then with untreated feed for a 14-d elimination period. Samples were extr acted 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 concen trations were tenfold lower than those of BDE 28 (which was determined to have the highest concentration among BFRs analyzed). Uptake efficiency was lo west 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 K

ow values of the BFRs and the levels determined in fish. The study detected several compounds in fish tissues th at may have been metabolites of the dosed BFRs. Notably, several hexaBDEs were tent atively identified by full-scan GC/MS and these compounds may have been debromination products of BDE183 or BDE209; however, one cannot discount the possibility th at they were impurities in the material dosed to the organisms. No further discussi on 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-Dawle y 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 fo rmation in plasma were monitored over a 6-d period. One subset of 18 rats was dosed or ally 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 si ngle-dose study with radiolabelled decaBDE (Mörck et al. 2003) were analyzed to determ ine the total radioactiv ity associated with neutral and phenolic fractions in plas ma of rats exposed to decaBDE.54 The oral bioavailability was estimated to be approximately 26%, although the sampling methodology may have resulted in an artifici ally high determinati on of bioavailability. Oral bioavailability was defined as the fraction of administered parent compound reaching systemic circulation. The authors in dicated 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 phenol ic radioactivity in plasma at days 3 and 7 was 4 times that of the neutral fraction, suggesting si gnificant oxidative metabolism of neutral decaBDE, debrominated metabolites, and highe r exposure to phenolic metabolites than to neutral parent/metabolite compounds. The presence of phenolic metabolites compared with neutral compounds indica ted 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) ha lf-lives of 51 h (oral exposure) and 58 h (intravenous exposure).

The neutral compounds in plasma were identif ied as decaBDE (> 99.5%) as well as three nonaBDEs (< 0.5%). In addition to the ne utral compounds, 13 phenolic metabolites were determined in plasma with the major me tabolites inferred to consist of a hydroxyloctaBDE, a hydroxyl-nonaBDE and a guai acol-type hydroxymethoxy-hexaBDE. Based on the observed parent and metabolite com position, 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 vi a an arene oxide an d dihydrodiol, or via two sequential oxidation steps followed by methylation

The authors also described how hydroxyl-o ctaBDE and hydroxyl-nonaBDE could have affinity for transthyretin (TTR), which nor mally functions as a transport protein for thyroxine in plasma. The bi nding to TTR could explain th e high fraction of phenolic metabolites observed in plasma and may cause the phenolic metabolites to persist as bound residues in plasma. The binding of pheno lic metabolites could disrupt thyroxine transport in blood, resulting in adverse horm onal effects. This behaviour was reported by the authors to be similar to that reported for PCB metabolites.

Hakk and Letcher (2003) summarized the fi ndings of Orn and Kl asson-Wehler (1998), Hakk et al. (2002) and Mörck a nd Klasson-Wehler (2001) with respect to the fecal and biliary metabolites formed by rats dosed w ith 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 subse quent 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 compet e with thyroxine for binding to TTR.

• Guiacol metabolites could further oxidize to quinones, which are highly reactive and

would bind to cellular macromolecule s. 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 burde n in body tissues and feces of one nonaBDE congener (BDE207) and two octaBDE conge ners (BDE201 and BDE197) were higher than could be explained by the total dose of each congener associated with their background concentrations in ra t feed. The authors concluded that the elevated burdens

were the result of reductive debrominati on of decaBDE. The formation of BDE197 and

-207 from BDE209 results from meta de bromination(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 invo lved with the meta dehalogenation of the thyroid hormone thyroxine and are pr esent 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 debr omination was not the primary metabolic pathway or that the debrominated products rapidly underwent further metabolism. Overall, 45% of the dosed decaBDE was unacco unted for in rat tissues and feces and the authors speculated that the formation of bound and/or hydroxyl ated metabolites that were not included in their analysis was a likely e xplanation for the incomplete mass balance of

decaBDE.

Kierkegaard et al. (2007) repor ted the findings of a 3-month feeding study with two dairy cows. The congener profile in adipose and orga n 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. Traci ng of potential degradation of decaBDE during extraction and cleanup with ra diolabelled standards ruled out photolytic debromination as a source of these hepta- to nonaBDEs during extraction/cleanup. The authors also discounted higher di etary absorption of these cong eners as an explanation for the observed differences between feed and tissues since enhanced absorption was not apparent for similar congeners (i.e., BDE 183 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, followe d by meta debromination to BDE196, followed by meta debromination to BDE182

2. Meta debromination to BDE207, followe d by meta debromination to BDE197

In a study by Riu et al. (2008), Wistar ra ts were dosed orally with 99.8% pure [ 14 C]- radiolabelled decaBDE dispersed in peanut oi l, gestational days 16 to 19. Urine and feces56 were collected daily over this 4-d period. The animals were sacr ificed on day 20 of gestation (24 h afte r the last dose of 14 C-decaBDE) and the amounts of radioactivity, corresponding decaBDE levels, and metabolic pr ofiles were determined in various organs

and tissues. Extraction and an alytical methods were deve loped 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, demonstr ating relatively high absorp tion of decaBDE as well as

efficient distribution to tissues. The fecal route was the main excretion pathway, accounting for approximately 66.3% of the admi nistered 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 f eces and tissues, conducted using liquid chromatography/mass spectrometry (LC/MS) nuclear magnetic resonance (NMR), identified all 3 nonaBDEs, an unidentif ied octaBDE and a hydroxylated octaBDE derivative. Together (excreta + tissu es + 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 ru led out that the compound corresponded to a hydroxylated heptaBDE. Additional unidentifie d polar metabolites were detected in urine, with similar compounds found in the cont ents 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 ca lculated values were signifi cantly 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, th e highest residual le vels were found in endocrine glands: the adrenals (33 mg/kg) a nd ovaries (16 mg/kg). Both decaBDE and its metabolites were reported to cross the pl acental barrier in rats following oral administration (approximately 0.5% of the total dose was found in the fetuses).

Van den Steen et al. (2007) i nvestigated 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 unde r 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 euthan ized 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 subtra cted from values determined for biotic samples.57 Before implantation, BDE209 concentrati ons 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 eliminat ion. 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 musc le 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 f ound that various PBDEs were pr esent in both the control and exposed group muscle and liver tissues, w ith the differences be ing most pronounced for the nona- (BDE206, -207, and -208) and octaBD Es (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 transf ormation 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 addi tion to BDE209, eight nona- and octaBDE congeners frequently detected in their peregrin e falcon study have rarely been reported in wildlife. Together with BD E209, these congeners composed 16–57% of total PBDEs in urban eggs and 4.9–53% in rural eggs. The rese archers noted that th e 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 de tected BDE202 in 107 of 114 eggs. Since

BDE202 is not known to be found in Octa BDE or DecaBDE mixtures, they also speculated that its presence could be attri buted to metabolic tr ansformation of higher

brominated PBDEs.



### 3.1.3 Conceptual Models of Metabolism in Vivo

This section summarizes pathways of meta bolism based on those proposed in reviewed studies. This analysis respects that some studies contain uncerta inties (e.g., potential impurities in the test substance administered to organisms, lack of chemical specificity in analyses of metabolites). Based on the review ed 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 transfor mation products including hydroxylated and hydroxymethoxylated PBDEs and/or th e formation of bound residues.58 

### Fish
Based on currently available studies and inform ation reviewed in this report by Stapleton et al. (2004, 2006), personal comm unications from HM Staple ton to Environment Canada (January 2008 and July 2009) (unreferenced), Kierkegaard et al. ( 1999) and Tomy et al. (2004) (respecting the uncertainties in thes e 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 debrominat ed 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 preferen tially 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 Kierkega ard 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, poten tially via an arene oxide. This could involve the action of cyto chrome P450 enzymes.

4. The potentially formed hydroxylated BD Es may compete with thyroxine for binding to TTR, a thyroxine transpor t protein present in blood serum.

5. The catechols may then be methylated, pot entially by the action of catechol-Omethyltransferase, to form the observed guaicols.

6. The guiacol metabolites could further oxi dize 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 wate r-soluble metabolites which would be excreted via bile and feces, as was obse rved in conventional and cannulated rats.



### 3.1.4 Implications for Bioaccumulation Assessment

The existing evidence for the debrominat ion and metabolism of decaBDE suggests metabolic pathways which lead to the formation of the following:

1. Lower brominated PBDEs (down to he ptaBDE 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 metabolite s 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. (20 03). It is acknowledged

that in some cases some of the lower brom inated PBDEs may have field or laboratory measurements of BAFs or BMFs. However, field and laboratory bi oaccumulation data for the bioformed hydroxylat ed and hydroxymethoxylated BDE metabolites are generally lacking, and application of the models fo r all types of metabolites was seen as a consistent way to inform our knowledge regarding the bi oaccumulation and biomagnification potential of these metabolites. The PBDE screening assessment supporting working document (Environment Ca nada 2006b) provides a summary of the measured aquatic BCFs and BAFs for the Pe ntaBDE and OctaBDE co mmercial products.

These studies showed that BCFs and BAFs exceed 5000 for tetra-, penta- and hexaBDEs (this is consistent with the fi ndings of the modelling exercise conducted for this review).

For making BAF predictions with the BAF–QSAR model, the log K

ow of each metabolite was based on reported measurements wherev er possible. In the absence of reported measurements, log K

ow was estimated using the Experi mental Value Adjustment (EVA) method in the KOWWIN estimation progr am. The EVA uses a reference logK

ow from another structurally close substance (in this case decaBDE) and co rrects 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 k M for lower brominated PBDEs from laboratory data, reasonable approximations were made that took into c onsideration rates of metabolism for similar congeners. For hydroxylated and hydroxymethoxyl ated BDEs, there was no information

that would allow for estimation of k M , and a value of 0.026/day (i.e., equal to that estimated for decaBDE) normalized to the body weight of the m odel’s middle trophic

level fish at 15 o C (~0.02/day), was chosen for illustra tive purposes. These metabolites are expected to undergo further metabolism; how ever, the actual rates of metabolism are unknown. The chosen value and resulting BA Fs represent reasonable hypothetical predictions. Further details of log K

ow and k M estimates are provided in Appendix D.

Figure 3-1 provides frequency hi stograms of the pooled BAF predictions fo r all potential metabolites and for the middle trophic level. In the absence of meta bolism, the BAFs of all metabolites are predicted to exceed 5000. However, with consideration given to metabolic transformation (a more realistic s cenario), the predicted BAFs range from 295 to approximately 6 000 000. Based on the fit of a normal distributi on 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 nonaBD Es, hydroxy-hexaBDEs, and penta- to octaBDEs. The remaining 26% of BAF pred ictions that are below 5000 consist of hydroxyl-octa- to nonaBDEs a nd nonaBDEs. The large proportion of high BAFs, even

when metabolic transformation is consider ed, 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 metabo lites occurred much faster than approximately 0.02/d then it is possible that the BAFs would be much lowe r. 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 persis tence 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 k M values chosen for the model were ba sed on laboratory observations in most cases (the one exception was for nonaBDEs, wh ich 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 (m etabolism is relatively fast) compared to

the maximum half-lives observed in lab stud ies (refer to Appendix D). Therefore the predicted BAFs are less conser vative (i.e., lower) than if th e longest observed half-lives were chosen. Despite the use of less conser vative half-lives, th e 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 quantifie d 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 meta bolite 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 resu lting in increased exposure and risk to upper trophic level organisms.61 Figure 3-1: Combined frequency distributi ons 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 BAFFrequency 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 BAFFrequency 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 BAFFrequency 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 BAFFrequency Observed Frequency

Normal

Distribution BAF = 500062 BMF predictions for decaBDE metabolites in wolves were made using a spreadsheet version of the Gobas et al. (2003) model. Log K

ow values were chosen based on the

rationale described in A ppendix D, whereas log K

oa for each metabolite was estimated using KOAWIN and by i nputting corrected logK

ow 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 k M 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 k M 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 meta bolites are expected to undergo further metabolism; however, the actual rates are unknown.

Calculation of BMFs for the BDEs were co rrected for dietary assimilation efficiency according to the E D vs. log K

ow relationship described in Ke lly 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 K

ow and log K

oa estimated for decaBDE metabolites. However, with consideration given to me tabolic 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 relati vely high dietary assimilation efficiency

assumed for these substances.

In mammals, the actual rate of metabolism is unknown for the hydroxylated and hydroxymethoxy metabolites and th erefore the BMF predictions are uncertain and may be high. If the secondary transformation of th ese 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 sp ecies, the enzymes responsible for Phase II metabolism might be less developed or lack ing, resulting in persistence of these metabolites and much higher BMFs that approach the BMF values estimated without consideration given to me tabolic 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 biomagnifi cation potential. The predicted BMFs, which exceed 1 for all metabolites, raise concerns th at 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 increa sed 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 sc reening assessment for PBDEs (Environment Canada 2006a, Environment Canada 2006b).

• Norris et al. (1973, 1974) exposed solid de caBDE (98% decaBDE and 2% nonaBDE) to sunlight in water. The bromine concentration in water increased during the experiment, suggesting that photodegradati on 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 mi xture of hexane, benzene and acetone (8:1:1). After 16 h of exposure to ultrav iolet (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 photode gradation 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 la yer to silica gel, sand, soil or sediment. The toluene was eva porated off in the dark. Exposure was to natural sunlight and to UV light genera ted by three mercury lamps. Commercial DecaBDE applied to solid matrices was obs erved to transform following a pathway of reductive debromination with products in cluding primarily hexa- to nona-BDEs with some occurrence of penta and tetra produc ts (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), sedi ment (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 phototr ansformation products of decaBDE adsorbed to sand, sediment and soil. Based on the relative amounts of debromination products in the samples, th e 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 gla ss 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 c onditions and light used. The presence of unidentified compounds was inferred in some of the treatments. Further analysis of64 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), hydrat ed them with reagent-grade water, and irradiated the system with artificial sunli ght or with natural s unlight. After 60–72 h of irradiation, 29–56% of the ini tial decaBDE remained on the quartz surface, depending on the type of light and wave length used. Addition of humi c acid slowed the rate of decaBDE decay. Small amounts of nona- and octaBDEs were likely formed as degradation by-products. BDE47 and -99 were no t formed at detectable levels of the analytical method used (high-perfor mance liquid chromatography, HPLC).

• Palm et al. (2003) investigated the p hotodegradation of decaBDE dispersed in toluene, dichloromethane or a solvent mixt ure 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 w ith 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 percen t of decaBDE degradation followed a pathway of debromination, while the products of the remaining 25% could not be determined. In a follow-up study conducte d 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 artifi cial 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. Th e particles were suspended in water, atomized, dried and dispersed as an aeroso l in a smog chamber. The conditions used were assumed to maximize the degradation potential of decaBDE. Although the conditions may not replicate atmospheric transp ort, the particle size used (diameter of approximately 1μm in the aerosol) can be c onsidered environmentally relevant. When the aerosol was exposed to simulated sun light (fluorescent la mps) and/or hydroxyl radicals, the rate of degrada tion of decaBDE was just barely measurable, at a rate of

< 6 x 10 13 cm 3 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 th e other transformation products that were

formed.65 • Keum and Li (2005) investigated the debr omination of decaBDE in aqueous solution when in contact with the reducing agents zerovalent iron, iron sulphide and sodium sulphide. In the experiments with zerovale nt iron, decaBDE was rapidly transformed to lower brominated PBDEs. Approximately 90% of the parent was converted to mono- to hexaBDE congeners after 40 d. Du ring the initial reaction period (up to 5 d), decaBDE was predominantly transforme d into hexa- and heptaBDEs, but tetraand pentaBDEs were predominant after 14 d. The results demonstrated reductive debromination. The experiments with iron sulp hide and sodium sulphide also showed

transformation of decaBDE to lower brominat ed PBDEs, but the rate was slower than that found in the presence of the zerovalen t iron. A similar profile of transformation products was found to that determined in the experiment using zerovalent iron. The Government of the United Kingdom (Un ited 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 c ould influence reaction rate. Nevertheless,

similar reactants to those used in this study (e.g., iron-bearing mi nerals and sulphide ions) are present in sediments and so ils under anaerobic conditions, although it is

uncertain whether environmental conditions are appropriate to activate similar

reactions under na tural conditions.



### Biodegradation

The following bullets summarize the evidence re garding biodegradation of decaBDE that was considered in the screening assessm ent for PBDEs (Canada 2006; Environment Canada 2006b):

• The biodegradation of decaBDE was repor ted 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 14 C-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 14 C-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 14 CO 2 and 14 CH 4 , indicating that essentially no mineralization had occurred.

• Gerecke et al. (2005) undertook experi ments 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 spik ed 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 nonaBD Es (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 was66 0.005 nmol/sample). To stimulate degradation, primers (i.e., 4-bromobenzoic acid, 2,6-dibromobiphenyl, tetrabromobisphe nol A, hexabromocyclododecane and decabromobiphenyl) were added in the amount of 9–11 nmol/sample. Two other experiments were carried out with two nona BDE congeners to determine the fate of these potential decaBDE degradation products . Starch and yeast (50 g) were added immediately before the bottle was filled w ith 20 mL of freshly collected digested sewage sludge. The bottles were tightly ca pped and incubated at 37 ±1°C for up to 238 d in the dark. Analysis by GC/HRMS i ndicated 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 ps eudo-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 oc taBDEs. 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 de monstrated 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 Debrom ination in the Environment



### Abiotic Degradation
Bezares-Cruz et al. (2004) investigated the photodegradation of decaBDE (98% pure obtained from Great Lakes Chemical Compa ny) dissolved in hexane and exposed to natural sunlight. DecaBDE (i.e., BD E209) at a concentration of 2-5 μ M in hexane degraded rapidly to lower brominated PBDE co ngeners in the presence of sunlight, with up to 99% reduction in the decaBDE concentration in 30 minutes (exposure to summer sunlight). First-order degradation ra te 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 result s to a natural setting. The authors concluded that photodegradation in nature would be lim ited by sorption of decaB DE to particulates, light attenuation by humic materials and lower concentrations of hydrogen donor chemicals—which may also be less favour able hydrogen donors in the aquatic environment. Stapleton (2006a) also conclude d 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 ex tremely low water solubility. Since this study is not applicable to natu ral 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 artificial67 ultra-violet (UV) light was i nvestigated 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 cons tants 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 se ttings makes it difficult to ex trapolate these results to the environment. The results of this study we re therefore not considered relevant to

environmental conditions.

The water / humic acid mixture tested by Erik sson et al. (2004) can potentially provide a better simulation of potential phot odegradation in a natural se tting since humic acids are often present in aquatic systems and could play a role in photode gradation. To prepare the solution for this experiment, 20 mL of a sa turated solution of decaBDE in ethanol was

combined with 10 mL of ethanol containi ng 50 mg of humic acid. Approximately 10 mL of ethanol was evaporated using a nitroge n stream, and the remaining solution was

combined with 2 L of water. The solution wa s then heated and maintained at 80°C for 1 hour under a constant flow of nitrogen. Th e solution was cooled to room temperature and then transferred to cylindrical reaction vessels for the irradiation experiments. The Government of the United Kingdom (Unite d Kingdom 2007a) estimated a final humic substances concentration for this experime nt of approximately 25 mg/L and concluded that traces of ethanol may have been pr esent since it was uncl ear how much ethanol would have been lost by heating at 80°C. Th e experiments were replicated between 2 and 5 times but it is not clear how many replicat es 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 presen ted 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 lowe r brominated PBDEs (including 3 nonaBDEs, at least 7 octaBDEs, 8 heptaBDEs and sm all amounts of hexaBDEs), mono- topentabromodibenzofurans (mono- to pentaBDFs), and possibly brominatedmethoxylated-dibenzofurans. One key differen ce 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 pos sible that similar reactions could occur in the natural environment. However, the act ual environmental rates and degree of debromination are uncertain since this experime nt 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 cons equence, the findings of the water-only experiment were highly uncertain.

In their study respecting the effect of sewa ge sludge application on concentrations of PBDEs in soils and earthworms, Sellström et al. (2005) provide a br ief discussion of an investigation of photolytic debromination of BDE209 amended in soil. The BDE209 was68 amended with soil from Björketorp, placed in glass test tubes and e xposed 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 fr om the light. Ultimately, soils showed no evidence of photolytic breakdown. The authors c oncluded 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 formati on 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 2x10 6 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 PBDF s were made using HRGC/HRMS. The analysis technique was ab le to identify PBDFs with 6 or fewer bromines; hepta- and octaBDFs were not included in the analysis scheme. After

decaBDE:toluene solution was irradiated wi th 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 basi s and the authors suggested a trend toward greater transformation (but similar transf ormation products) at lower wavelengths. The

authors also concluded that the observed mono- to hexaBD Fs were likely the result of stepwise debromination of higher PBDFs sin ce mono-to hexaBDEs were not detected in any of the irradiated solutions. While thes e findings demonstrate the possibility of decaBDE transformation to PBDFs, they are co nsidered to have low applicability to the

natural environment.

Barcellos da Rosa et al. (2003) studied the photolysis of decaBDE di ssolved in toluene. DecaBDE (BDE209, Sigma-Aldrich, 98% purity) was dissolved in toluene at a concentration of 0.31 mM and exposed to li ght from a 500-watt (W) high-pressure xenon lamp. Following light exposure, samples of th e 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 ph otolysis rate constant of 3x10 -4 /s. Several debrominated congeners were observed in the decaBDE:toluene solutions and the degradation of decaBDE was inferred to proc eed via sequential debromination to nona-, octa- and heptaBDEs. The authors concluded that while their results demonstrated the potential for and pathway of photolytic de bromination, further work to confirm the environmental relevance of photol ysis in toluene was needed. Overall, the findings of this study are considered to have limited a pplicability to the environment.

Rahm et al. (2005) conducted a study to determine the relative susceptibility of a variety of compounds, including decaBDE, to hydrolys is via nucleophilic aromatic substitution.

When decaBDE was reacted with sodium meth oxide 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 r eaction 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 mine ral surfaces, enzymes, etc., would be the main reactant for hydrolysis, there is a high le vel 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 containi ng 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 base d on retention times. The photolysis products include d hepta- to nonaBDEs as well as tri- to hexabromodibenzofurans. Comparison with the chromatogram fo r 2,3,7,8-substituted

dibenzofurans indicated that these were not major degradation products. These findings have similar shortcomings to other studies us ing organic solvents a nd 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 Ocea n. DecaBDE (BDE 209; purity > 98.3%) was dissolved in isooctan e (250 ng/mL), the solution was placed in quartz GC-autosampler vials a nd 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 cont rol and an isooctane blank. The quantity of decaBDE was analyzed using GC–MS. The concentrat ion of decaBDE decreased according to first-order kinetics, with a ha lf-life of approximately 0.03 d.

Kuivikko et al. (2007) reported a quan tum yield for decaBDE of 0.28 ±0.04. The quantum yield was used in model simulations to predict the photodegr adation half-life in the Baltic Sea (both surface and 10 m mixing la yer) and the Atlantic Ocean (40 m mixing layer only). The model simulations assumed e nvironmentally relevant concentrations (3– 4 pg/L and 30–40 pg/L) in the water phase a nd typical summer sola r 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 repres ent only a very small amount of the total decaBDE present in the water co lumn. Therefore, there is unc ertainty as to whether the70 predicted half-lives would be a representative portray al of the persistence and degradation of decaBDE in the water column.

Ahn et al. (2006a) investigated the photoche mical debromination of decaBDE sorbed to major components of soil/sediment/mineral aerosols including clay minerals and noncrystalline metal oxides which are known to ha ve electron transferring capacities. The decaBDE used in the study had a purity of 98% . Test matrices included montmorillonite, kaolinite, organic-carbon-ri ch natural sediment (16.4% OC content), aluminum hydroxide, iron oxide and manga nese dioxide. Each of the decaBDE-amended test matrices (250 mg) was combined with 500 μ L water and the mixtures were then irradiated with artificial light or natural sunli ght. Irradiation with artif icial 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). Sample s of decaBDE were quantified using HPLC, while the debrominated products were quantif ied using GC – electron capture detector (GC–ECD).

The dark and light controls showed no signs of degradation alt hough the study reported minor peaks of nona- and octaBDEs in the da rk control which were likely impurities in

the decaBDE formulation. For the artific ial 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 a nd 990 d, respectively, with negligible degradation on aluminum hydroxi de (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 neglig ible 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 po tential for uncertainty in these estimates.

The half-lives for the natu ral light exposures suggest that decaBDE photodegradation occurs slowly in the natural setting.

The fastest degradation was observed on m ontmorillonite 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 (BDE 197 and -196), as well as trace amounts of tri- to heptaBDEs over longer sunlight exposure time s. Higher fractions of tri-to heptaBDEs

were also observed for the artificial li ght exposures. Several unidentified octaBDE products were also observed in the chroma tograms. The formation of the identified products was consistent with stepwise debrom ination, initially forming nona-, then octaand then heptaBDEs.

Ahn et al. (2006b) investigated metal oxide mediated debromination of decaBDE using birnessite (a naturally occurring mangane se 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 solution71 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 re move 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 (i i) the role of THF as the hydrogen donor for debromination of decaBDE, determined by m easuring the production of succinic acid in the system. All treatments were conducted in tr iplicate and shaken in the dark for a period of 24 h with subsampling for analysis at va rious times. A separate set of experiments investigated the degradation of decaBDE sorbed to birnessite in water systems, in the presence of the naturally occurring hydr ogen donor, catechol. Catechol was combined (0.003, 0.045 and 46.135 mmol) with 5 mL of wate r and mixed with 50 mg of decaBDEamended birnessite. All experimental treatment s were shaken in the dark for 23 d with subsampling at various times to track th e 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 re tention 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 expe riments determined that THF was reacting with the birnessite to form succinnic acid, a nd 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 depe nd on the presence of THF. Because THF is not normally present in the environment, the rate of decaBDE degradation is not considered realistic with resp ect 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 st epwise manner with decaBDE being initially degr aded 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 degradati on was observed over the 23-d period in the experiments with 0.003–0.045 mmol catechol. Howe ver, slow degradation was observed with 46 mmol catechol, with the mass of d ecaBDE in the reacti on vessels decreasing from approximately 0.1 μ mol to 0.085 μ mol 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 th e THF:water systems. Although th e 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) re ported 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 referen ce 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 certif ied 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 quan tity of decaBDE to yiel d a concentration of

2180 μ g/kg dw. Analysis of the “cleaned” dust pr ior to spiking conf irmed that decaBDE was not detected (< 0.2 μ g/kg). Samples (0.5 g) of each dus t 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 precipita tion 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 natura l 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 an alyzed for lower brominated degradation products. At the end of the exposure, a pproximately 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 predomin antly the three nonaBDE congeners, but also due to lesser amounts of octa- and heptaB DE 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 make s up a very small component (i.e., from below detection to 0.8%), and BDE203 is not detected.

This study provides reasonably strong evid ence that photodegradation of decaBDE on

dust can occur under natural conditions in th e environment and that lower brominated

PBDEs can be formed. The authors noted th at 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 s unlight, making debromination of decaBDE on dust in cars potentially significant.

The Government of the United Kingdom (U nited 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 temperature73 of 27.4°C. Over the 90-hour exposure period, decaBDE concentrati on in the house dust

decreased from 2180 μ g/kg dw to 1570 μ g/kg dw, indicating that approximately 28% of decaBDE degraded. The start and end concentr ations 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 bala nce of decaBDE and lower brominated PBDE congeners indicated that approxima tely 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 qua ntum yield of decaBDE on kaolinite and measured light penetration in this mineral in order to calculate th e photodegradation rate of decaBDE. The decaBDE used in the study ha d a purity of 98%. Light penetration into kaolinite was estimated using the Kube lka-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 prepar ed 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 de tail, 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 dr y kaolinite or wetted kaolinite. The spiked kaolinit e layers were irradiated w ith 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 pene tration in the kaolinite below 50 μ m and the authors concluded that only decaBDE sorbed to particle s at the very surface of soils would have the potential to undergo photolysis. In the e xperiments with decaBDE-spiked kaolinite,

half-lives of 76 and 73 minutes were determ ined 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 identif ied as lower brominated PBDEs, whereas for wet conditions, the products were not iden tified, suggesting that the products were not lower brominated PBDEs. These findings de monstrate that while photodegradation of decaBDE sorbed to mineral solids is possibl e in the environment, the rate and amount degraded is highly dependent on the penetration of light in to soil and mineral layers.

Nose et al. (2007) studied the degradation pathways of standard decaBDE (purity not noted, but purchased from Wako Pure Chemi cal 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 autoclav e chamber, which was sealed. Temperature was controlled, with the first heating of 25 mi n to 300oC and a pressure of 8 MPa. The experiment was repeated at different tim e 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%) af ter 10 min at 300°C. Debromination to nonaBDE was determined; however, debrominat ion 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 wi th other lower brominated PBDEs with similar findings. Thus, the authors concluded th at 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. Th e study also confirmed the formation of PBDD/DFs during the study. While having li mited applicability to the natural environment, the study appear s to confirm the findings of other studies respecting preferential meta- and para- position debromination (e.g., St apleton 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 phot olysis conditions, all PBDE s (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 debrom ination of decaBDE (DE-83R, > 97% purity

obtained from Great Lakes Chemical Co mpany) by resin-bound zerovalent iron nanoparticles. The study involved the use of a bout 50 test tubes prepared with decaBDE in acetone, mixed with an equal volume of di stilled water to form an 8-mL solution. Then 2 g of resin containing zerovale nt 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 result s 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 su bsequent formation of nona- to triBDEs in a sequential manner. All three nonaBDEs appeared in signifi cant amounts within 1 h, steadily increased for 8 h, th en decreased below detection after 24 h. The congener pattern dominance shifted to heptaBDEs af ter 2 d, and then hexaBDEs were most abundant. After 10 d, pentaBDEs were present in signi ficant amounts.

The authors noted that the id entification 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 availabl e chemical standards. Issues of co-elution were identified. However, all th ree nonaBDEs were identified with confidence and two of five octaBDE peaks were identified (i.e., BDE197 and -196). The he pta- to tetraBDEs that were confidently identif ied included BDE183, -153, -15 4, -99 and -47 (all contain both para-position bromines). Overall, th e presence of uniden tified peaks makes it75 difficult to conclusively confirm the positional preference of debromination pathways. Li et al. (2007) also conducted an identical study using PCB 209, which dechlorinated at a

much slower rate than the debrominati on of BDE209. After 10 d, only 21% of PCB209 was lost, with only the formation of nona- a nd 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 plas tics under natural sunlight. High-impact polystyrene (HIPS) wa s added to toluene containing 100 μ g/mL of DecaBDE and shaken overnight to achieve complete dissolution; toluene subsequently evaporated in the dark. Crushed televi sion 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 st ainless-steel sieve apparatus. The fine sieved powder (106–300 μ m) 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-washe d 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 irra diation experiment was performed.

The HIPS+DecaBDE samples showed a gradual disappearance of BDE209 upon exposure to sunlight. No degradation was found in the dark c ontrols throughout the experiment. After one week of exposure to sunlight, the concentration of BDE209 declined to approximately 50% of the initia l level, indicating prom pt photodegradation in the plastic matrix. Hydrated HIPS+DecaBDE sa mples showed more rapid degradation of BDE209 than non-hydrated samples. The aut hors suggested that this accelerated degradation might be explained by the fact th at water played a ro le as a hydrogen donor, promoting debromination.

Kajiwara et al. (2008) state that previous studies have focu sed on the photode gradation of PBDE dissolved in solvents or absorbed to pa rticles, and that thei r study is the first to show the photolysis of BDE209 compounded in to 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 fo r 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 show ed that concentrations of hexa- to nonaBDE congeners increased several-fold afte r one week of exposure, and the authors interpreted that debromination of BDE209 resu lted 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 d ecreased slightly, de spite the continuous decrease in BDE209. Congeners BDE47, -99 a nd -100 were not detected and stepwise debromination of PBDEs was not clearly obs erved 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 photolyt ic formation of tri- to octaBDFs in their HIPS+DecaBDE76 samples. The concentration of total PBDFs show ed an increase (greater than 40 times) on day 7 of exposure, while BDE 209 concentration decreased.

With respect to the TV casing samples, the study showed no disappearance of BDE209

and formation of lower brominated PBDE c ongeners throughout the 224 d of sunlight irradiation, in contrast to HIPS+DecaBDE sa mples. Two possible explanations for the differences in degradation profiles were suggested in the study: 1) th e difference in initial concentrations of BDE209 (HIPS+DecaBDE and the TV casing samples contained 0.1 and 10% BDE209, respectively) influenced degrad ation rates; and/or 2) the effects of the other plastic additives in TV casings, such as pigments, UV absorbers, and stabilizers, potentially affected light penetra tion depth in each plastic sample.

Raff and Hites (2007) examined the importan ce of photolysis in the removal of PBDEs from the atmosphere by comparing the photolys is frequencies of PBDE congeners to the

first-order removal rate constants for oxida nt reactions and dry and wet deposition. The study compared atmospheric removal processe s affecting PBDE (i.e., focused on BDE47, -99 and -209) fate using a box model simula ting the atmosphere above Lake Superior. The model results indicated that photolysis resu lts in the highest loss (90% of the total losses) of BDE-47 and -99 from the atmos phere above Lake Superior. However, for BDE209, more than 90% of the re moval 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 sa mples collected in the Arctic that showed significant depletion of BDE209 relative to tetra- and pent abrominated PBDE congeners. The main conclusion of the study is that de position processes contro l 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 e xperiments to their 2005 study. The anaerobic biodegradation of technical DecaBDE (98% purity) in digested sewage sludge was investigated using the same test sy stem 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 th e absence of a primer, the half-life was longer, up to 1400 d. The aut hors also conducted field monitoring of decaBDE in a sewage treatment plant to dete rmine 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-scal e digester. However, the authors cautioned that the relatively short residence time in th e reactor (28 d) and low level of replication (i.e., only one set of grab samples) meant that these re sults should be considered preliminary and not necessarily unequivocal.77 The biodegradation of decaBDE (> 98% purit y) 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 (T CE) 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 inoc ulation the test vesse ls were sealed to

ensure anaerobic conditions and th en autoclaved for 25 minutes at 121°C. To initiate each experiment, 5 mL of a decaBDE stock solution in trichloroethane (TCE, ~1 μ M decaBDE) was added to the test vessel for a final concentration of 1 mM TCE and

0.1 μ M 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 confir m 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 dechlo rination of TCE, but no degradation of decaBDE within the first weeks of incubation. After an additional two months of incubation without replacem ent of TCE, decaBDE was observed to degrade to non-detectable levels, while oc ta- and heptaBDEs became detectable in both replicates. No degradation was evident in the abiotic controls, indicating that the disappearance of decaBDE was due to anaerobi c degradation. These findings suggest that under appropriate conditions, certain bacteria may debrominate decaBDE. No degradation of decaBDE was seen in the expe riments using any of the other cultures.

Parsons et al. (2004) investig ated the potential for reductiv e debromination of decaBDE in anaerobic sediment suspensions. The e xperiments 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 suspension s were spiked with decaBDE (14 μ g/g sediment) and incubated anaerobically at room temperature in the dark (details of the anaerobic test system were not provided). Sterilized contro ls were also conducted (details of the sterilization procedure was not provided). At varying intervals over an approximately 205-day test period (read from graph), experi mental and control samples were extracted in hexane/acetone, subjected to a clean-up process and anal yzed using GC–low resolution mass spectrometry (LRMS).

Concentrations of decaBDE in the spiked samples decreased significantly over the first

two months of incubation a nd the decrease coincided with the appearance of new78 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 th at nonaBDEs were further debrominating to lower brominated PBDE congeners. There is significant uncertainty in these findings with respect to anaerobic sediment debromin ation 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 confir med by the production of methane upon addition of lactic, pyruvic and acetic ac ids 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 at tributed unequivocally to anaerobic degradation.

Parsons et al. (2007) report the findings of an additional investigation of reductive

debromination of decaBDE in anaerobic sedi ment microcosms. The experiments were conducted using sediment collected at Hansweer t 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 (c ontaining acetate, lact ate and pyruvate). The suspensions were then spiked with decaB DE or individual nonaBDE congeners and incubated at room temperature in the dar k. At varying intervals over an approximate 260-day test period (read from graph), experi mental 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 decaB DE in the active and control microcosms, although nonaBDEs were detected in the decaBDE-spiked samples at much higher concentrations than known backgr ound levels in the study sediments

Knoth et al. (2007) conducted a monitori ng 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 pr ocess—primary sludge, secondary excess sludge and (dewatered) digested sludge—were co llected 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 extracti on, and reduction to 100 μ L. 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 trea tment 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 rete ntion 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 envi ronment. Concentrations of decaBDE and79 lower brominated PBDEs were monitored in sl udge, sediments, and fish in the receiving environment of a WWTP located in Roxboro, No rth 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 congene rs were identified in 2005. The major congener in sludge was BDE209, accounting fo r 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 debromin ation in WWTP sludge. In receiving environment sediments the major PBDE c ongeners included BDE209, -206, -99 and -47, with BDE209 making up > 89% of the tota l 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 maximu m levels ranging from 3240 mg/kg OC to 2450 mg/kg OC. The lower brominated PBDE congeners were attributed to the PentaBDE technical formulation, suggesti ng minimal debromination of decaBDE in

surficial sediments.

Tokarz et al. (2008) cond ucted concurrent PBDE (i.e., BDE209, -99 and -47) experiments using anaerobic sediment microc osms and a cosolvent-enhanced biomimetic system. The sediment microcosms contained natural sediments with no detectable PBDEs collected from Celery Bog Park, West Lafaye tte, 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. Autocl aved control systems were implemented. The biomimetic experiment involved the use of Teflon-cappe d 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. Cont rols were used without titanium citrate. Debromination products were identified a nd quantified using GC–ECD and GC–MS.

Tokarz et al. (2008) notes that cobalamins, such as coenzyme vitamin B 12, have the ability to mediate reductive dehogenation of halorganics. Cobalamins from decaying cells have been iscolated from environmental samp les. In addition, they indicate that vitamin

B 12 has ubiquitous presence in the envir onment in anaerobic micro-organisms.

The biomimetic system demonstrated reducti ve 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 aver age of 14 years, based on observations over 3.5 years. After 8 months, BDE47 decreased approximately 30% without a consistent concurrent increase in daughter debromina tion products. While complete debromination to diphenyl ether was not ruled out, it appe ared 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 synthesized80 their data from both systems and proposed ma jor debromination pathways for sediment and biomimetic systems as follows: BDE209 > nonaBDEs (BDE206, -207 -208) > octaBDEs (BDE196, -197) > heptaBDE s (BDE191, -184, 2 unknown heptaBDEs) > hexaBDEs (BDE138, -128, -154, -153) > pe ntaBDEs (BDE119, -99) > tetraBDEs (BDE66, -47, -49) > triBDEs (BDE28, -17). Speci fically, at the end of the 3.5-year study, their analysis of BDE209 degradation in sediments identifie d BDE208, -197, -196, -191, -128, -184, -138, and -128, as well as 3 uni dentified octaBDEs and 2 unidentified heptaBDEs.

The findings of Tokarz et al. (2008) show ing prolonged persistence of decaBDE in sediment microcosms are supported by a fiel d 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 ra nged from 14.6 to 1082 ng/g dw at seven agricultural sites. Hi gh concentrations (i.e., 71.7 ng/g dw) were even found at one site that had not received sludge application for four years, i llustrating the persistence of

BDE209 in soils. Sells tröm et al. (2005) collected PBDE samples from sites that received past sewage sludge amendments. They sample d soil from three research stations (with reference plots and sewage-sludge-amende d plots) and two farms (reference and amended/flooded soils) in Sweden. They determ ined 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 auth ors identified all th ree 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 rapidl y with time, the increase of the octaBDE concentrations was slower. Congener patterns of the octa- and nona BDE 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 sp an. However, the authors noted different relative proportions of nona - and octaBDEs in the sediments in comparison to the commercial PBDE mixtures. Notably, they de tected 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. Al so, they noted that congener patterns of

octaBDE were dissimilar to those of the co mmercial PBDE mixtures, but similar to patterns in house dust and profiles of photodegradation products of decaBDE. Consequently, they suggested biotic and/or abiotic transf ormation processes are involved

between the release of the technical products into the e nvironment and their final residues in sediments that are responsible for the ch anges of congener pattern s that they observed

in the sediments of the Swiss lake.81 Zhou et al. (2007) evaluated the ability of wh ite 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, includ ing 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 applica tion of this technology is most suited for bioremediation applicati ons. Overall, this st udy 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 flas k systems. A 100-mL aliquot of aqueous medium (distilled water, malt extract, glucos e, peptone and yeast extract mixture) was added to each flask. White rot fungi were i noculated to the liquid culture and the test system was shaken for 10 d in the dark. Iden tical test systems were also implemented utilizing additions of Tween 80 and cycl odextrin 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 d ecrease of 42.2% over 10 d in the amount of BDE209 in the test system. The sterile cont rols showed no signifi cant degradation over

time. Tween 80 was found to enhance BD E209 degradation at an appropriate concentration (maximum degradation 96.5% ove r 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) presente d an ongoing study that may provide some key information

on the environmental transformation and a quatic bioaccumulation of decaBDE. Their field mesocosm study compares sediment and periphyton BDE209 concentrations from low (dose=0.039 g of DecaBDE), medi um (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 D ecaBDE 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 tr eatments, but not in the low-treatment or control mesocosms. DecaBDE was only observed in periphyton in the medium- and high-treatment mesocosms. NonaBDE was obs erved 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 Tran sformation in th e 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 degrad ation 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 (Webst er et al. 2004) was conducted and82 parameterized for the “Ontario Mixed W ooded Plain” region. The primary model parameter inputs included

• log K

ow = 8.7 (Wania and Dugani 2003; see Appendix C).

• log K

oa = 12 (maximum value allowed in m odel, 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 e xpected to be primarily sorbed to organic fractions. Table 3-1 summarizes the predicted fractions in each medi um for releases to air, water or soil. From Ta ble 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 sediment s (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 environm ent 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 th an 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 transforma tion 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 lo w since the dissolved/vapour fraction of decaBDE in the environment is pr edicted 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 av ailable information on debromination in the environment is how well the experimental conditions represent t hose that would be expected under natural settings . For example, in experiments that use organic solvents to disperse or dissolve decaBDE, extrapolati on 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 wate r is much lower than that of organic solvents. Furthermore, organic solvents are often mu ch stronger hydrogen donors than water or other naturally occurri ng substances, resulting in faster rates and potentially non-representative patterns of decaBDE transformation. In the environment, many different donor molecules would be pres ent, 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 par titioning properties of decaBDE.83 hydrogen is unlikely to be the sole route of tr ansformation. This creates uncertainty as to how well degradation studies that use orga nic solvents would represent expected degradation of decaBDE in the environment. Other experimental conditions also require scrutiny to ensure relevance to those that ma y be seen in the environment. For instance, experiments using natural sunlight or simula ted natural sunlight are likely to better represent natural conditions than those using ar tificial light. It is al so not likely that

materials such as zerovalent iron nanoparticles, specialized microbial cultures and fungal cultures will be found in great a bundance in the environment.

In general, laboratory studies examining decaBDE transformation were frequently

conducted under conditions that were favourab le for decaBDE debromination and are not necessarily representative of typical conditions in the environment. Thus, while some

transformation of decaBDE in the environmen t 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

water 1 1 kg/year to

soil 1 1 kg/year to

air 1

Mass in bulk air (kg) 1.27E-07 2.17E-07 2.23E-03

Fraction of total mass in bulk ai r 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 dissolv ed 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 bul k 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 av ailable information regard ing possible environmental debromination pathways for decaBDE in a c onceptual model while Table 3-2 provides a list of the products which are potentially fo rmed from each debromination process. A tabulated summary of the environmental de bromination information is provided in Appendix E.

Studies focusing on decaBDE sorbed to partic ulates or solid phases conducted in the absence of organic solvents, with natural su nlight, may provide a re alistic 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 partic ulates appears susceptible to relatively rapid transformation, with half-lives ranging from 76 min (decaBDE sorbed to a thin film of kaolinite; Gerecke 200 6) 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- an d pentaBDFs or unidentified products.

DecaBDE (i.e., BDE209) sorbed to partic ulates in aqueous systems appears to photodegrade at a generally slower rate, w ith half-lives rang ing 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 de monstrated significant degradation (up to 71%) over 60–72 h. The identified products incl uded hexa- to nonaBDEs and some lower brominated PBDEs. The potential for photoly tic transformation of the sorbed decaBDE appears to be modulated by the ability of s unlight to penetrate water and the sorbing matrix. Gerecke et al. (2006) determ ined that light penetrated only 50 μ m into kaolinite, suggesting that the photodegradation of decaB DE 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 photodegradati on on solid phases can oc cur at significant rates, only a very small fraction of the tota l 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 sorb ed to solids may also be subject to biodegradation processes in aquatic sediment s, 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 natura l sediment microcosms had a half-life that ranged from approximately 6 to 50 years, w ith 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-BDE s. 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 st udies of WWTP sludge (K noth et al. 2007; La Guardia et al. 2007) provide little evid ence of decaBDE debr omination in WWTPs, possibly because the residence time in WWTPs is too short for significant debromination to be observed. In addition, sediment micr ocosm studies by Parsons et al. (2004, 2007) fail to provide evidence of significant debr omination, although Parsons et al. (2007) did identify small amounts of nonaBDEs. Thus, wh ile 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 en vironmental transformation pathways for

decaBDE

A) Photodegradation on Particulates (Dus t, Soil, Aerosol and Sediment) in Dry and Aqueous Systems87 B) Biodegradation88 (C) Photodegradation in Solution (Natural and Organic Solvents)89 D) Abiotic Degradation (Oth er than Photodegradation)90 Table 3-2: Summary of decaBDE transformation products observed in laboratory studies with some relevance to environmental settings

DecaBDE Transformation ProcessCongener GroupFormedTargeted for VirtualElimination underCEPA 1999PhotodegradationSorbed to Solids inDry and AqueousSystems(references in brackets)PhotodegradationDissolved in NaturalSolvents(references in brackets)PhotodegradationDissolved in ArtificialOrganic 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 solv ents, these results are considered to have low relevance to the environment. ***Identification of debromination products wa s 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 bec ause the decaBDE degradation pr oducts were described as lower brominated PBDEs, with no indica tion 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. 200892 Although photodegradation experiments with d ecaBDE dissolved in water potentially represent environmentally relevant cond itions, 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 provide s a conceptual model of photodebromination pathways for decaBDE dissolved in water/solv ent systems. Two of the associated studies

are notable because the result s are potentially relevant to the natural environment. Eriksson et al. (2004) observed relatively ra pid transformation (half-life = 6.4 h) of

decaBDE in water/humic acid mixtures w ith hexa- to nonaBDEs formed. Because artificial light was used, there is some uncerta inty whether the rate would be as fast in natural sunlight. Kuivikko et al. (2006, 2007) observed a photodegrad ation 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 transforma tion of decaBDE provide support for a

conclusion that transformation to lower BD Es and BDFs should be occurring in the environment. However, there is a dichotomy between the findings of these studies and

the results of monito ring studies in terms of r ealistic support for significant

transformation of decaBDE in the environment. If these laboratory-based studies provide accurate depictions of transformation in th e environment, one would expect to find similarities in the patter ns observed in the environment and those described as transformation products of decaBDE in labo ratory 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 commerc ial mixtures. The United Kingdom (2007a)

notes that the available m onitoring data provide limited circumstantial evidence for a

correlation between concentrations of decaB DE and lower brominated PBDEs in some environmental matrices (e.g., citing the findi ngs 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 biodegradati on, decaBDE is susceptible to abiotic debromination in the absence of light e xposure as shown under laboratory conditions. Figure 3-2d summarizes the transformation path ways for these other abiotic degradation processes. Keum and Li (2005) observe d 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 sul phide. Li et al. (2007) showed complete disappearance after 8 h in a water-acetone sy stem containing nanos cale 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 tr ansformation and debromination of decaBDE through environmental processes such as photodegradation, abiotic degradation and biodegradation indicates potenti al 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 th e lower brominated PBDEs and PBDFs would, themselves, be susceptible to further degradation and the actual significance of the formation of these pr oducts is unknown. Palm et al. (2003) and Eriksson et al. (2004) have show n that the rate of transfor mation 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 sola r spectrum at ground level, and th us it is expected that they would be less susceptible to decay than decaBDE. Since lower brominated PBDEs demonstrate slower rates of phototransformati on, it could be speculate d 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 con tinued 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 degrada tion of PBDEs depends on their relative rates of formation and degradation in the envi ronment, both of which are not known. In addition, the magnitude of accumulation woul d depend on decaBDE loading rates to the environment, which are also difficu lt to predict (Un ited 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 la boratory studies discussed in this report are also occurring in the natural environmen t; however, it is acknowledged that this is a subject of uncertainty. As noted earlier, in the absence of me tabolism, the BAFs of hexato nonaBDEs are predicted to exceed 5000. When metabolic tr ansformation 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 informa tion that would allow for estimation of k M , and a value equal to that estimated for decaBDE was chosen for illustrative purposes. It is likely that PBDFs would undergo further me tabolism; however, the actual rates of metabolism are unknown. Although the chosen value and resulting BAFs represent reasonable hypothetical predicti ons, it is important to acknowl edge that these predictions are uncertain. Further detail of log K

ow and k M estimates are provided in Appendix D. In the absence of metabolism, th e BAFs for penta- and hexaBD Fs are predicted to exceed 5000 for the middle trophic level fish, while BA F predictions for tri- and tetraBDFs are94 below 5000. With consideration given to meta bolic transformation (a more realistic scenario), the status of th e 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 predic ted 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 tetr aBDFs were 8 and 45, respectively. When corrected for metabolism, the BMF valu es ranged from 4 to 6 (a similar k M value was used for all BDFs).

The modelling analysis of pot ential environmental transfor mation products of decaBDE predicted BAFs that predominantly exceed 5000 and BMFs exceeding 1 for practically all of the metabolites, because of relativel y high dietary assimila tion efficiency and optimal logK

ow and logK

oa values for bioaccumulati on via passive diffusion. This suggests concerns that the tr ansformation products formed in the environment as a result of the photodegradation, biodegradation, and possibly abiotic degradation, of decaBDE, are potentially bioaccumulat ive 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 bioaccumulati on of decaBDE does not support a conclusion

of “bioaccumulative” as set out in the Persistence and Bioaccumulation Regulations

. While most available data show that decaB DE 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 le vels that can be subjectively described as high. The modelling undertaken to support this evaluation, ho wever, shows uncertainty

associated with metabolism in fish as mode l-predicted aquatic BAFs range from below

the 5000 criterion to well above 5000. Predicte d terrestrial carnivore BMF values also range from below 1 to greater than 1 de pending on the rate of metabolism assumed. Although less relevant than BAF or BMF, e xperimental 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 effi ciency 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 PB DE product formed is typically very low95 (e.g., representing less than 1%, up to a fe w 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 gr oups, hexa- and pentaBDEs have been noted as bioformed in fish. However, the in terpretation of results from metabolic transformation studies is at times compli cated by incomplete characterization of impurities in the dosed material and/or dosage of many PBDEs concurrently. Some rodent studies have made inferences, base d on mass balance evalua tions, that rates of transformation may be higher, with one study suggesting that approximately 45% of the

total dose of decaBDE was unaccounted for a nd 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 Assessm ent of PBDEs (Environment Canada 2006a, 2006b), which identified photodegradation and biodegradation as lik ely mechanisms for

transformation in the environment. This revi ew 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 plausibl e 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 s ubject 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 si gnificant rates, only a very small fraction

of the total decaBDE reservoir in the e nvironment (e.g., that fraction adsorbed to particulates or on solid surfaces ) that has contact with sunlig ht 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 w ith potential exposure to sunlight. DecaBDE sorbed to solid surfaces (b oth anthropogenic and natural) could also be exposed to sunlight. Within these phases, light attenuation and ma trix 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 show n to be a very slow process, with halflives on the scale of a few y ears to several decades. Some studies have also not shown any obvious transformation in sediments or so il 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 patter ns dominated by the commercial PBDE products. As well, the infrequent historic an alysis of higher bromin ated PBDEs like octaand nonaBDEs could make it difficult to de tect 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- a nd hexaBDEs. One study also indicates trace96 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 deca BDE has also been shown to form tri- to

octaBDFs as well as un identified products.

Modelling of BAFs and BMFs played an impor tant 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 unc ertainty associated with the metabolism potential of decaBDE in fish and log K

ow determinations for this substance. Given the exceptionally low water solubility limit of deca BDE, it is not expected that this substance will be appreciably taken up from the wate r 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 transformati on products are also predicted to be very high; however, when assumptions were made respecting metabolic transformation, the predicted BMFs are significantly lower, but ar e still all greater than one. This analysis suggests that many possible decaBDE meta bolites/transformation products could be highly bioaccumulative and some metabolites may have the capacity to biomagnify in

food chains.

Overall, this review confirms that, base d 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 cons idered high. In addition, some equivocal evidence suggests poten tial biomagnification in food chains. Although uncertainties remain, it is reasonable to conclude that decaB DE may also contribute to the formation of bioaccumulative and/or potentially bioaccumu lative 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 conclusi ons are relevant to

alternative flame retardants with similar chemical structures and use patterns. For instance, decabromodiphenyl etha ne (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 in sulation 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 d ecaBD 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). Ba sed on structural similarities, the two substances likely have similar physical-chemical properties, characterist ics 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), ha s 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 E nvironment Agency recently published a detailed risk assessment for this substance (United Kingdom 2007b). While direct risks resulting from toxic effects of th is substance were considered low, concerns were identified over this substance’s pot ential 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 th is State of Science re port, the similarity in properties between decaBDE and decaBD et hane, the presence of the decaBD ethane in Canadian wildlife, and the potential for d ecaBD 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. Understa nding the risk from alternatives will help

to ensure that substitutions of flame reta rdants are made on an informed basis.



## 5 Key Uncertainties and Data Gaps

The data for bioaccumulation, biomagnificat ion 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 unders tanding of the bioaccumulation and biomagnification of decaBDE.

Overall, the relative contribut ion of environmental transf ormation 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 environmen t would help clarify the significance of in98 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 BM F determinations for decaBDE on the total burden of parent decaBDE and the metabol ites 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 pr esented in this report for decaBDE and known potential transformation products vary. These relied heavily on logK

ow logK

oa E D and k M values, which would benefit from refineme nt in order to improve the precision of the model predictions.

BAFs, BMFs and other indicators of bioaccumu lation or biomagnification have not been

measured for some of the known metabolite s and transformation products of decaBDE, especially their hydroxylated and methoxyhydr oxylated forms. Further study of the behaviour of these products would be useful.99 

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Chemosphere 70:172-177110Appendix A – Biota Concentration DataConcentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitBIRDS OF PREYChen et al. 2007 China 2004–2006Common kestrelFalco tinnunculusMuscle 6 21% 2150 ng/g MeanChen et al. 2007 China 2004–2006Common kestrelFalco tinnunculusLiver 6 21% 2870 ng/g MeanChen et al. 2007 China 2004–2006Common kestrelFalco tinnunculusKidney 6 21% 483 ng/g MeanChen et al. 2007 China 2004–2006SparrowhawkAccipiter nisusMuscle 11 21% 192 ng/g MeanChen et al. 2007 China 2004–2006SparrowhawkAccipiter nisusLiver 11 21% 249 ng/g MeanChen et al. 2007 China 2004–2006SparrowhawkAccipiter nisusKidney 11 21% 83 ng/g MeanChen et al. 2007 China 2004–2006Japanese sparrowhawkAccipiter gularisMuscle 6 21% 75 ng/g MeanChen et al. 2007 China 2004–2006Japanese sparrowhawkAccipiter gularisLiver 6 21% 78 ng/g MeanChen et al. 2007 China 2004–2006Japanese sparrowhawkAccipiter gularisKidney 6 21% 33 ng/g MeanChen et al. 2007 China 2004–2006Little owlAthene noctuaMuscle 6 21% 150 ng/g MeanChen et al. 2007 China 2004–2006Little owlAthene noctuaLiver 6 21% 96 ng/g MeanChen et al. 2007 China 2004–2006Little owlAthene noctuaKidney 6 21% 40 ng/g MeanChen et al. 2007 China 2004–2006Scops owlOtus suniaMuscle 6 21% 537 ng/g MeanChen et al. 2007 China 2004–2006Scops owlOtus suniaLiver 6 21% 174 ng/g MeanChen et al. 2007 China 2004–2006 Scopsowl Kidney 6 21% 59 ng/g Mean111Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitOtus suniaChen et al. 2007 China 2004–2006Long-eared owlAsio otusMuscle 6 21% 9.5 ng/g MeanChen et al. 2007 China 2004–2006Long-eared owlAsio otusLiver 6 21% 107 ng/g MeanChen et al. 2007 China 2004–2006Long-eared owlAsio otusKidney 6 21% 23 ng/g MeanChen et al. 2007 China 2004–2006Upland buzzardButeo hemilasiusMuscle 3 21% 21 ng/g MeanChen et al. 2007 China 2004–2006Upland buzzardButeo hemilasiusLiver 3 21% 58 ng/g MeanChen et al. 2007 China 2004–2006Upland buzzardButeo hemilasiusKidney 3 21% 29 ng/g MeanChen et al. 2007 China 2004–2006Common buzzardButeo buteoMuscle 3 21% 26 ng/g MeanChen et al. 2007 China 2004–2006Common buzzardButeo buteoLiver 3 21% 71 ng/g MeanChen et al. 2007 China 2004–2006Common buzzardButeo buteoKidney 3 21% 93 ng/g MeanChen et al. 2008 United States, NE 1996–2006Peregrine falconFalco peregrinusEggs 114 0 1.4–420 (26) ng/gRange(median inbrackets)de Boer et al. 2004 United Kingdom Not givenPeregrine falconFalco peregrinusLiver 6 2 of 6 < 5.7–181 < 0.17–6.7 μg/kg Range 0.08–9.8de Boer et al. 2004 United Kingdom Not givenPeregrine falconFalco peregrinusMuscle 5 None 53–344 1.8–9.5 μg/kg Rangede Boer et al. 2004 United Kingdom Not givenPeregrine falconFalco peregrinusEgg 6 4 of 6 < 1.8–108 < 0.08–7.5 μg/kg Range Lipid:de Boer et al. 2004 Sweden Not givenPeregrine falconFalco peregrinusEgg 20 9 of 20 < 4–412 < 0.3–21 μg/kg Range 1.5–200de Boer et al. 2004 Sweden Not givenSparrowhawkAccipiter nisusLiver4 All < 82 – < 200 < 3.2 – < 9.8 μg/kg Rangede Boer et al. 2004 Sweden Not givenSparrowhawkAccipiter nisusMuscle 5 None 13–275 0.26–2.2 μg/kg Rangede Boer et al. 2004 Sweden Not givenSparrowhawkAccipiter nisusEgg 5 2 of 5 < 2.1–38 < 0.16–1.5 μg/kg Range112Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitde Boer et al. 2004 Sweden Not givenKestrelFalco tinnunculusLiver 5 3 of 5 < 5.8–120 < 0.26–5.5 μg/kg Rangede Boer et al. 2004 Sweden Not givenKestrelFalco tinnunculusMuscle 5 4 of 5 < 4.2–10 < 0.11–0.29 μg/kg Rangede Boer et al. 2004 Sweden Not givenBarn owlTyto albaLiver 5 2 of 5 < 2.6–37 < 0.13–2.5 μg/kg Rangede Boer et al. 2004 Sweden Not givenBarn owlTyto albaMuscle 5 4 of 5 < 6.3–14 < 0.5–1.2 μg/kg Rangede Boer et al. 2004 Sweden Not givenBarn owlTyto albaEgg 4 1 of 4 < 20–30 < 2–1.7 μg/kg Rangede Boer et al. 2004 Sweden Not givenMontague's harrierCircus pygargusEgg 4 1 of 4 < 2.1–28 < 0.12–1.3 μg/kg Rangede Boer et al. 2004 Sweden Not givenMarsh harrierCircus aeruginosusEgg 2 All < 2 – < 2.4 < 0.09 μg/kg Rangede Boer et al. 2004 Sweden Not givenSea eagle Egg 1 None 6.2 0.48 μg/kg Rangede Boer et al. 2004 Sweden Not givenOspreyPandion haliaetusEgg 3 All < 4 – < 27 < 0.2 – < 1.5 μg/kg Rangede Boer et al. 2004 Sweden Not givenGolden eagleAquila chrysaetosEgg 5 All < 4.1 < 0.2 μg/kg Rangede Boer et al. 2004 Sweden Not givenMerlinFalco columbariusEgg 2 1 of 2 < 43–4.3 < 3.8–0.3 μg/kg RangeHerzke et al. 2005 Northern Norway 1993–2000Peregrine falconFalco peregrinusEgg 9 Detected ng/g MedianHerzke et al. 2005 Northern Norway 1992–2002Golden eagleAquila chrysaetosEgg 15 Detected ng/g MedianHerzke et al. 2005 Northern Norway 1995–2000MerlinFalco columbariusEgg 9 Detected ng/g MedianJaspers et al. 2006 Belgium 2003–2004Barn owlTyto albaLiver 7 Some 59 ng/g MedianJaspers et al. 2006 Belgium 2003–2004Barn owlTyto albaMuscle 7 Most 68 ng/g MedianJaspers et al. 2006 Belgium 2003–2004long-eared owlAsio otusLiver 6 Most 66 ng/g MedianJaspers et al. 2006 Belgium 2003–2004Long-eared owlAsio otusMuscle 6 All < 2.25 ng/gDetectionlimitJaspers et al. 2006 Belgium 2003–2004 Buzzard Liver 16 All < 2.25 ng/g Detection113Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitButeo buteolimitJaspers et al. 2006 Belgium 2003–2004BuzzardButeo buteoMuscle 16 All < 2.25 ng/gDetectionlimitJaspers et al. 2006 Belgium 2003–2004SparrowhawkAccipiter nisusLiver 5 Some 52 ng/g MedianJaspers et al. 2006 Belgium 2003–2004SparrowhawkAccipiter nisusMuscle 5 All < 2.25 ng/gDetectionlimitJaspers et al. 2006 Belgium 2003–2004KestrelFalco tinnunculusLiver 5 Most 85 ng/g MedianJaspers et al. 2006 Belgium 2003–2004KestrelFalco tinnunculusMuscle 5 All < 2.25 ng/gDetectionlimitKunisue et al. 2008 Japan (Coastal) 1994Steller’s sea-eagleHaliaeetus pelagicusMuscle 1 1 <0.5 ng/gDetectionlimitKunisue et al.2008 Japan (Inland) 1994–1995Golden eagleAquila chrysaetosMuscle 2 1 <0.5–10 ng/g RangeKunisue et al.2008 Japan (Inland) 1997GoshawkAccipiter gentilisMuscle 1 0 120 ng/gLindberg et al. 2004 Captive Sweden 1988–1999Peregrine falconFalco peregrinusEggs 4 2 < 7–9 (8.1) ng/gRange(geometricmean inbrackets)Lindberg et al. 2004 Southern Sweden 1992–1999Peregrine falconFalco peregrinusEggs 8 1 < 20–430 (86) ng/gRange(geometricmean inbrackets)Lindberg et al. 2004 Northern SwedenPeregrine falconFalco peregrinusEggs 8 0 28–190 (80) ng/gRange(geometricmean inbrackets)Lindberg et al. 2004 Northern Sweden 1991–1999Peregrine falconFalco peregrinusEggs 18 110 ng/g MeanPotter et al. 2009United States,Maryland andVirginia1993–2002Peregrine falconFalco peregrinusEggs 23 00.186–48.2(6.35)ng/gRange(median inbrackets)Voorspoels et al.2006bBelgium 2001–2003BuzzardButeo buteoBrain 29 All ng/gVoorspoels et al. Belgium 2001–2003 Buzzard Fat 16 All ng/g114Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimit2006bButeo buteoVoorspoels et al.2006bBelgium 2001–2003BuzzardButeo buteoLiver 3 Most 19–190 (79) ng/gRange(mean)Voorspoels et al.2006bBelgium 2001–2003BuzzardButeo buteoMuscle 29 All ng/gVoorspoels et al.2006bBelgium 2001–2003BuzzardButeo buteoSerum 20 Some 2–58 (15) ng/gRange(mean)Voorspoels et al.2006bBelgium 2001–2003OwlsBrain 8 All ng/gVoorspoels et al.2006bBelgium 2001–2003 Owls Fat 1 All ng/gVoorspoels et al.2006bBelgium 2001–2003 Owls Liver 8 All ng/gVoorspoels et al.2006bBelgium 2001–2003 Owls Muscle 8Allng/gVoorspoels et al.2006bBelgium 2001–2003 Owls Serum 3 Most 9 ng/g MeanVoorspoels et al.2006bBelgium 2001–2003SparrowhawkAccipiter nisusBrain 7 All ng/gVoorspoels et al.2006bBelgium 2001–2003SparrowhawkAccipiter nisusFat 2 All ng/gVoorspoels et al.2006bBelgium 2001–2003SparrowhawkAccipiter nisusLiver 7 Most 15–19 ng/g RangeVoorspoels et al.2006bBelgium 2001–2003SparrowhawkAccipiter nisusMuscle 7 All ng/gVoorspoels et al.2006bBelgium 2001–2003SparrowhawkAccipiter nisusSerum 2 None 16–36 ng/g RangeVorkamp et al. 2005 Greenland 1986–2003Peregrine falconFalco peregrinusEggs 37 0 3.8–250 (11) ng/gRange(median)de Boer et al. 2004 Sweden Not givenRed kiteMilvus milvusEggs 4 3 of 4 < 2.1–29.1 < 0.09–2.3 μg/kgFRESHWATER FISH ANDINVERTEBRATESHale et al. 2002United States; mid-Atlantic region2000BluegillLepomis macrochirusMuscle μg/kg115Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitHale et al. 2002United States; mid-Atlantic region2000BluegillLepomis macrochirusRemainderof bodyμg/kgEljarrat et al. 2007 SpainNov. 2004, Nov.2005CarpCyprinus carpioMuscle 2 None 63.7–95.3 ng/g Range 2–19 pg/g wwBarbelBarbus graellsiiMuscle 13 1 of 13 nd–707 ng/g Range 2–19 pg/g wwde Boer et al. 2003 The Netherlands 1999Freshwater musselsDreissena polymorpha< 4 – < 34 μg/kgDodder et al. 2002Northeastern UnitedStates; LakeSuperior1994SmeltOsmeru mordaxWhole body < 1.5 μg/kgDodder et al. 2002 Lake Ontario 1994SmeltOsmeru mordaxWhole body < 1.6 μg/kgDodder et al. 2002Lake of the Ozarks,MO1999White crappie; bluegillPomoxis annularis;Lepomis macrochirusWhole body < 1.4 μg/kgDodder et al. 2002 Hadley Lake, IN. 1999White crappie, bluegillPomoxis annularis; LepomismacrochirusWhole body < 1.4 μg/kgDodder et al. 2002 Hadley Lake, IN. 1999White crappie, bluegillPomoxis annularis;Lepomis macrochirusMuscle < 1.4, < 1.3 μg/kgGuo et al. 2009Pearl River Delta,S. ChinaBighead carp (farmed)Aristichthys nobilisSkin 86.78 ± 6.47ng/g dwMean (± SD)10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaBighead carp (farmed)Aristichthys nobilisGills 81.18 ± 1.28ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaBighead carp (farmed)Aristichthys nobilisGIT 82.09 ± 3.11ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaBighead carp (farmed)Aristichthys nobilisLiver 80.85 ± 2.07ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaBighead carp (farmed)Aristichthys nobilisMuscle 80.48 ± 0.56ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaMandarin fish (farmed)Siniperca chuatsiSkin 87.68 ± 5.46ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaMandarin fish (farmed)Siniperca chuatsiGills 81.34 ± 1.86ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009 Pearl River Delta, Mandarin fish (farmed) GIT 83.55 ± 4.60ng/g dw Mean (± SD) 10 ng/g dw116Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitS. ChinaSiniperca chuatsiGuo et al. 2009Pearl River Delta,S. ChinaMandarin fish (farmed)Siniperca chuatsiLiver 81.57 ± 3.05ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaMandarin fish (farmed)Siniperca chuatsiMuscle 80.22 ± 0.41ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaNorthern snakehead (farmed)Nemipterus virgatusSkin 81.61 ± 2.09ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaNorthern snakehead (farmed)Nemipterus virgatusGills 82.22 ± 2.75ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaNorthern snakehead (farmed)Nemipterus virgatusGIT 80.99 ± 2.80ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaNorthern snakehead (farmed)Nemipterus virgatusLiver 80.22 ± 0.59ng/g dwMean (± SD)10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaNorthern snakehead (farmed)Nemipterus virgatusMuscle 8NDng/g dwMean (± SD)10 ng/g dwIsmail et al. 2009 Lake Ontario 1979–2004Lake troutSalvelinus namaycushWhole body29 (4 or 5 peryear)None 2.3 (±0.5)–12 (±5.3)0.27 (± 0.30)–1.3 (± 0.8)ng/g Mean (± SE) 0.23 ng/g wwSellström et al. 1998 Sweden 1995PikeEsox luciusMuscle μg/kgSellström et al. 1998 Sweden 1995PikeEsox luciusMuscle not detected – trace μg/kgSellström et al. 1998 Sweden 1995PikeEsox luciusMuscle not detected – trace μg/kgWang et al. 2007 China Sept.–Dec. 2006 Lichen Whole 1 None 1572 ng/g dw 1 ng/g dwWang et al. 2008 China Sept.–Dec. 2006March brownLimnodrilus hoffmeisteriWhole1 None 11.37ng/g dw1 ng/g dwWang et al. 2008 China Sept.–Dec. 2006 CoccidWhole1 None 114ng/g dw1 ng/g dwWang et al. 2008 China Sept.–Dec. 2006ZooplanktonMonia rectirostris, Monia micrura,Monia macrocopaWhole1 None 151.9ng/g dw1 ng/g dwWang et al. 2008 China Sept.–Dec. 2006Common carpCyprinus carpioWhole4 75% 5ng/g dw1 ng/g dwWang et al. 2008 China Sept.–Dec. 2006Java tilapiaTilapia niloticaWhole3 100% < 1ng/g dw1 ng/g dwWang et al. 2008 China Sept.–Dec. 2006Leather catfishSilurus meridionalisWhole4 Some 19.32ng/g dwMean1 ng/g dw117Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitWang et al. 2008 China Sept.–Dec. 2006Crusian carpCarassius auratusWhole5 Some 1.79ng/g dwMean1 ng/g dwWang et al. 2008 China Sept.–Dec. 2006Chinese softshell turtleChinemys reevesiiWhole3 Some 2.63ng/g dwMean1 ng/g dwXiang et al. 2007 China Apr.–Aug. 2004Large yellow croakerPseudosiaena croceaMuscle 13Some117.4 ng/g MedianXiang et al. 2007 ChinaApr.–Aug. 2004 Silvery pomfretPlatycephalus argenteusMuscle 10Some24.4 ng/gMedianXiang et al. 2007 ChinaApr.–Aug. 2004 Flathead fishPlatycephalus indicusMuscle 17Some0 ng/gMedianXiang et al. 2007 ChinaApr.–Aug. 2004 Robust tongue fishCynoglossus robustusMuscle 8Some0 ng/gMedianXiang et al. 2007 ChinaApr.–Aug. 2004 Bombay duckHarpodon nehereusMuscle 9Some0 ng/gMedianXiang et al. 2007 ChinaApr.–Aug. 2004 Jinga shrimpMetapenaeus affinisSoft tissue 10Some0 ng/gMedianXiang et al. 2007 ChinaApr.–Aug. 2004 Greasy-back shrimpMetapenaeus croceaSoft tissue 10Some30.0 ng/gMedianXiang et al. 2007 ChinaApr.–Aug. 2004 Mantis shrimpOratosquilla oratoriaSoft tissue 9Some42.47 ng/gMedianMAMMAL – HERBIVORELichota et al. 2004Vancouver Island,Canada2001MarmotMarmota vancouverensisFat 0.5 μg/kgVoorspoels et al.2007Flanders, Belgium 2001Wood miceApodemus sylvaticusLiver andmuscle12 All < 17 ng/gVoorspoels et al.2007Flanders, Belgium 2001Bank volesClethrionomys glareolusLiver andmuscle12 All < 17 ng/gMAMMAL – PREDATORMariussen et al.2004Norway 1995MooseAlces alcesLiver Not reported 0.8 ng/g MeanChristensen et al.2005Coastal BritishColumbia2003Grizzly bearUrsus arctos horribilisFat ormuscle6 Not reported ~ 0.1–2.77 μg/kgRange (fromgraph)Christensen et al. Interior British2003 Grizzly bear Fat or 6 Not reported~0.5–41.71 μg/kg Range (from118Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimit2005 ColumbiaUrsus arctos horribilismusclegraph)Sørmo et al. 2006 Svalbard 2003Polar bearUrsus maritimusBlubber 4 0 0.03–0.16 (.09) ng/gRange(mean)Sørmo et al. 2006 Svalbard 2003Polar bearUrsus maritimusBlubber 4 00.075–0.04(0.0225)ng/gRange(mean)Mariussen et al.2004Norway 1993–1994LynxLynx lynxLiver 7 0.1 ng/g MeanMuir et al. 2006 SvalbardPolar bearUrsus maritimusAdipose 15 < 1 ng/g MeanMuir et al. 2006 SvalbardPolar bearUrsus maritimusAdipose 44 < 1 ng/g MeanSkaare 2004 Svalbard 2001–2003Polar bearUrsus maritimusAdipose 1 ng/g MeanVerrault et al. 2005 Svalbard, Norway 2002Polar bear (female)Ursus maritimusPlasma 15 93% < 0.06 –0.1 ng/g RangeVoorspoels et al.2006aBelgium 2003–2004Red foxVulpes vulpesLiver 30 60% < 9.1–760 ng/g RangeVoorspoels et al.2006aBelgium 2003–2004Red foxVulpes vulpesMuscle 33 79% < 3.9–290 ng/g RangeVoorspoels et al.2006aBelgium 2003–2004Red foxVulpes vulpesAdipose 27 85% < 3.7–200 ng/g RangeMARINE FISHAkutsu et al. 2001Inland Sea of Seto,Japan1998 Conger eel < 0.2–0.53< 0.020–0.029μg/kgAkutsu et al. 2001Inland Sea of Seto,Japan1998FlounderParalichthys olivaceus1.9–3.2 0.015–0.022 μg/kgAkutsu et al. 2001Inland Sea of Seto,Japan1998Gray mulletMugil cephalus< 0.2–0.25< 0.010–0.013μg/kgAkutsu et al. 2001Inland Sea of Seto,Japan1998 Horse mackerel < 0.2–1.4< 0.030–0.047μg/kgAkutsu et al. 2001Inland Sea of Seto,Japan1998Red sea breamPagrus major< 0.2–0.74< 0.010–0.020μg/kgAkutsu et al. 2001Inland Sea of Seto,Japan1998Sea bassLateolabrax japonicus0.40–0.81 0.0087–0.017 μg/kg119Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitBurreau et al. 2004 Lumparn Estuary Not GivenRoachRutilus rutilusMuscle 8 5 48 ng/g medianBurreau et al. 2004 Lumparn Estuary Not GivenPerchPerca fluviatilisMuscle 33 21 1.3 ng/g MedianBurreau et al. 2004 Lumparn Estuary Not GivenPikeEsox luciusMuscle 25 21 1.7 ng/g MedianBurreau et al. 2006 Baltic Sea 1998SpratSprattus sprattusMuscle 6 3 0.082 ng/g MedianBurreau et al. 2006 Baltic Sea 1998HerringClupea harengusMuscle 5 1 0.24 ng/g MedianBurreau et al. 2006 Baltic Sea 1998SalmonSalmo salarMuscle 10 7 0.41 ng/g MedianBurreau et al. 2006 North Atlantic Ocean 1999Small herringClupea harengusMuscle 6 0 0.31 ng/g MedianBurreau et al. 2006 North Atlantic Ocean 1999Large herringClupea harengusMuscle 6 3 0.039 ng/g MedianBurreau et al. 2006 North Atlantic Ocean 1999Atlantic salmonSalmo salarMuscle 10 0 ng/g Mediande Boer et al. 1998 Dutch Coast 1995MackerelScomber scombrusMuscle < 2 μg/kgde Boer et al. 2001 North Sea 1999HerringClupea harengusLiver < 4.6 – < 11< 0.300 –< 0.520μg/kgde Boer et al. 2001 North Sea 1999HerringClupea harengusFilet < 1.4 – < 2.6< 0.160 –< 0.250μg/kgde Boer et al. 2001 North Sea 1999HerringClupea harengusMilt andeggs< 1.9 – < 5.0< 0.130 –< 0.200μg/kgde Boer et al. 2001 North Sea 1999 Cod Liver < 0.22 – < 53< 0.110–0.380μg/kgde Boer et al. 2001 North Sea 1999 Cod Fillet < 8.1 – < 40< 0.049 –< 0.170μg/kgde Boer et al. 2001 North Sea 1999WhitingMerlangius merlangiusLiver < 0.4 – < 1.2< 0.280 –< 0.610μg/kgde Boer et al. 2001 North Sea 1999WhitingMerlangius merlangiusFillet < 6.5 – < 13< 0.042 –< 0.078μg/kgde Boer et al. 2001 Tees Estuary, UK 1999SpratSprattus sprattusWhole body < 10 < 0.5 μg/kg120Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitde Boer et al. 2001 Tees Estuary, UK 1999WhitingMerlangius merlangiusWhole body < 33 < 0.5 μg/kgde Boer et al. 2001 Tees Estuary, UK 1999WhitingMerlangius merlangiusMuscle < 33 – < 100 < 0.5 μg/kgde Boer et al. 2001 Tees Estuary, UK 1999WhitingMerlangius merlangiusLiver < 1.56 – μg/kgde Boer et al. 2001 Tees Estuary, UK 1999FlounderParalichthys olivaceusLiver < 2.7 < 0.5 μg/kgde Boer et al. 2001Western Scheldt,The Netherlands1999Greater sandeelHyperoplus lanceolatus< 18 – < 22 < 0.2 μg/kgde Boer et al. 2003 The Netherlands 1999FlounderParalichthys olivaceus< 0.2–< 6 μg/kgde Boer et al. 2003 The Netherlands 1999Bream< 0.03–< 21 μg/kgFjeld et al. 2004Svolvaer, Lofoten,Norway2002Atlantic codGadus morhuaLiver 21 0.99 ng/g MeanFjeld et al. 2004Varangerfjorden,Norway1998, 2002Atlantic codGadus morhuaLiver 21 0.98 ng/g MeanGuo et al. 2009Pearl River Delta,S. ChinaCrimson snapper (farmed)Lutjanus erythropterusSkin 81.47 ± 2.72ng/g dwMean (± SD)10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaCrimson snapper (farmed)Lutjanus erythropterusGills 80.30 ± 0.55ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaCrimson snapper (farmed)Lutjanus erythropterusGIT 81.06 ± 2.99ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaCrimson snapper (farmed)Lutjanus erythropterusLiver 80.59 ± 1.25ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaCrimson snapper (farmed)Lutjanus erythropterusMuscle 80.048 ± 0.14ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaGolden threadNemipterus virgatusSkin 815.54 ± 24.81ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaGolden threadNemipterus virgatusGills 82.69 ± 5.47ng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaGolden threadNemipterus virgatusGIT 8NDng/g dw Mean (± SD) 10 ng/g dwGuo et al. 2009Pearl River Delta,S. ChinaGolden threadNemipterus virgatusLiver 80.31 ± 0.83ng/g dw Mean (± SD) 10 ng/g dw121Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitGuo et al. 2009Pearl River Delta,S. ChinaGolden threadNemipterus virgatusMuscle 80.089 ± 0.25ng/g dw Mean (± SD) 10 ng/g dwSørmo et al. 2006 Svalbard 2003Polar codBoreogadus saidaWhole body 7 0 0.05–0.42 (0.2) ng/gRange(mean)Sørmo et al. 2006 Svalbard 2003Polar codBoreogadus saidaWhole body 7 00.006–0.05(0.024)ng/gRange(mean)Johnson-Restropo2005Coastal Florida 2004Silver perchBairdiella chrysouraMuscle 6 All < 0.022 ng/g MDLJohnson-Restropo2005Coastal Florida 2004Striped mulletMugil cephalusMuscle 6 All < 0.022 ng/g MDLJohnson-Restropo2005Coastal Florida 2004Spotted seatroutCynoscion nebulosusMuscle 7 All < 0.022 ng/g MDLJohnson-Restropo2005Coastal Florida 2004Red drumSciaenops ocellatusMuscle 11 All < 0.022 ng/g MDLJohnson-Restropo2005Coastal Florida 2004Hardhead catfishArius felisMuscle 8 4.5 ng/g MeanJohnson-Restropo2005Coastal Florida 2004Atlantic stingrayDasyatis SabinaMuscle 7 0.1 ng/g MeanJohnson-Restropo2005Coastal Florida 2004Spiny dogfishSqualus acanthiasMuscle 5 16.9 ng/g MeanJohnson-Restropo2005Coastal Florida 2004Atlantic sharpnose sharkRhizoprionodon terraenovaeMuscle 5 514.3 ng/g MeanJohnson-Restropo2005Coastal Florida 1993–1994Bull sharkCarcharhinus leucasMuscle 6 39 ng/g MeanJohnson-Restropo2005Coastal Florida 2002–2004Bull sharkCarcharhinus leucasMuscle 7 778 ng/g MeanWatanabe et al.1987Japan 1981–1985MulletMugilidae4 μg/kgWatanabe et al.1987Japan 1981–1985 Goby 2 μg/kgWatanabe et al.1987Japan 1981–1985 Sardine 2 μg/kgWatanabe et al.1987Japan 1981–1985 Sea bass 1 μg/kgWatanabe et al.1987Japan 1981–1985 Horse mackerel 1 μg/kg122Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitWatanabe et al.1987Japan 1981–1985MackerelScomber scombrus1 μg/kgWatanabe et al.1987Japan 1981–1985 Hairtail 1 μg/kgMARINE INVERTEBRATESBurreau et al. 2006 Baltic Sea 1999ZooplanktonCalanoid sp.Muscle 3 None 2.1 ng/g Mediande Boer et al. 2001 North Sea 1999StarfishAsteroideaPyloriccaeca< 2.0–8.9< 0.150–0.310μg/kgde Boer et al. 2001 North Sea 1999Hermit crabAbdomen < 1.1–7.0< 0.1600–0.470μg/kgde Boer et al. 2001 North Sea 1999Male whelkHaustellum brandarisWhole body < 1.9 – < 14< 0.052 –< 0.190μg/kgde Boer et al. 2001 North Sea 1999 Shrimp Whole body < 2.4 – < 5.6< 0.073 –< 0.081μg/kgde Boer et al. 2001 Tees Estuary, UK 1999Shrimp Whole body < 100 < 0.5 μg/kgde Boer et al. 2001 Tees Estuary, UK 1999 Polychaete worm Whole body < 50 < 0.5 μg/kgde Boer et al. 2001 Tees Estuary, UK 1999 StarfishWhole bodyandtentacle< 9 < 0.5 μg/kgde Boer et al. 2001Western Scheldt,The Netherlands1999 Mysid shrimp < 20 93 < 0.2–0.98 μg/kgde Boer et al. 2001Western Scheldt,The Netherlands1999 Copepod < 37 < 0.3 μg/kgde Boer et al. 2001Western Scheldt,The Netherlands1999 Gudgeon < 6.0 – < 17 < 0.1 – < 0.4 μg/kgde Boer et al. 2003 The Netherlands 1999 Marine mussels < 4–< 5 μg/kgFjeld et al. 2004Svolvaer, Lofoten,Norway2002Blue musselsMytilus edullisMuscle 20 33 ng/g MeanFjeld et al. 2004Varangerfjorden,Norway2002Blue musselsMytilus edullisMuscle 50 3.6 ng/g MeanSørmo et al. 2006 Svalbard 2003Ice amphipodGammarus wilkitzkiiWhole body1 (compositeof 5)0 7.22 ng/g MeanSørmo et al. 2006 Svalbard 2003 Ice amphipodWhole body 1 (composite 0 0.28 ng/g Mean123Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitGammarus wilkitzkiiof 5)Watanabe et al. 1987 Japan 1981–1985MusselsLampsilis radiata5 4 of 5 1.4 μg/kg 0.5MARINE MAMMALSde Boer et al. 1998 Dutch Coast 1995Minke whaleBalaenoptera acutorostrataBlubber < 1 μg/kgde Boer et al. 1998 Dutch Coast 1995Sperm whalePhyseter catodonBlubber < 3 – < 6 μg/kgde Boer et al. 1998 Dutch Coast 1995Sperm whalePhyseter catodonLiver < 3 μg/kgde Boer et al. 1998 Dutch Coast 1995White-beaked dolphinLagenorhynchus albirostrisBlubber < 10 μg/kgde Boer et al. 1998 Dutch Coast 1995White-beaked dolphinLagenorhynchus albirostrisLiver < 1 μg/kgde Boer et al. 1998 Dutch Coast 1995 Harbour seal Blubber < 10 – < 15 μg/kgde Boer et al. 1998 Dutch Coast 1995 Harbour seal Liver < 1 – < 2 μg/kgde Boer et al. 2001 North Sea 1999 Harbour porpoise Liver 1.2 – < 7.4 – μg/kgde Boer et al. 2001 North Sea 1999 Harbour porpoise Blubber < 2.7 – 26 –μg/kgμg/kgde Boer et al. 2001 North Sea 1999 Harbour seal Liver 9 < 6.5–160 μg/kgde Boer et al. 2001 North Sea 1999 Harbour seal Blubber < 1.8–16 μg/kgde Boer et al. 2001 North Sea 1999 White-beaked dolphin Blubber < 3.8 μg/kgde Boer et al. 2001 North Sea 1999 White-beaked dolphin Kidney < 8.2 μg/kgde Boer et al. 2001 North Sea 1999 White-beaked dolphin Heart < 4.9 μg/kgde Boer et al. 2001 North Sea 1999 White-beaked dolphin Lungs 35 μg/kgde Boer et al. 2001 North Sea 1999 White-beaked dolphin Brain < 0.8 μg/kgde Boer et al. 2001 North Sea 1999 White-beaked dolphin Liver 140–318 μg/kgde Boer et al. 2001 North Sea 1999 Bottlenose dolphin Blubber < 2.5 μg/kgde Boer et al. 2001 North Sea 1999 Bottlenose dolphin Liver < 6.4 μg/kgde Boer et al. 2001 North Sea 1999 Bottlenose dolphin Kidney < 3.0 μg/kg124Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitde Boer et al. 2001 North Sea 1999 Bottlenose dolphin Muscle < 7.7 μg/kgde Boer et al. 2001 Tees Estuary, UK 1999 Porpoise Blubber < 8.7 – < 9 < 7.5 μg/kgSørmo et al. 2006 Svalbard 2003Ringed sealPhoca hispidaBlubber 6 5 0.02 ng/gOnlydetectedconcentrationSørmo et al. 2006 Svalbard 2003Ringed sealPhoca hispidaBlubber 6 5 0.006 ng/gOnlydetectedconcentrationJohnson-Restropo2005Western coast ofFlorida1991–1996Bottlenose dolphinTursiops truncatusBlubber 0 ng/g MeanJohnson-Restropo2005Western coast ofFlorida2000–2001Bottlenose dolphinTursiops truncatusBlubber All < 0.022 ng/g MDLJohnson-Restropo2005Eastern coast ofFlorida2001–2004Bottlenose dolphinTursiops truncatusBlubber All < 0.022 ng/g MDLJohnson-Restropo2005Coastal Florida ??Striped dolphinStenella coeruleoalbaBlubber All < 0.022 ng/g MDLShaw et al. 2007(BFR)NW Atlantic coast(Maine to LongIsland)1991 – 2005Harbour sealPhoca vitulinaBlubber 4 2 of 4 nd – 7.4 ng/g RangeMARINE/AQUATIC BIRDSde Boer et al. 2001 Tees Estuary, UK 1999CormorantCorvus marinusLiver < 17 – < 25 < 0.5 μg/kgde Boer et al. 2001Western Scheldt,The Netherlands1999Common ternSterna hirundoEggs < 5.6–70 < 0.46–7.1 μg/kgde Boer et al. 2001Western Scheldt,The Netherlands1999Common ternSterna hirundoEggs 4.4–27 0.45–2.9 μg/kgde Boer et al. 2004 Sweden Not givenGreat crested grebePodiceps cristatusLiver 4 3 of 4 < 1.5–9.1 < 0.11–0.52 μg/kgde Boer et al. 2004 Sweden Not givenGreat crested grebePodiceps cristatusMuscle 3 2 of 3 < 8.1–31 < 0.4–1.2 μg/kgde Boer et al. 2004 Sweden Not given Heron Liver 4 All < 2.3 – < 5.7< 0.08 – <0.25μg/kgde Boer et al. 2004 Sweden Not given Heron Muscle 5 4 of 5 < 6.3–563 < 0.32–4.5 μg/kg125Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitde Boer et al. 2004 Sweden Not givenGannettMorus bassanusEggs 12 All < 4 – < 57 < 0.2 – < 2.2 μg/kgGauthier et al. 2008 Canada1982, 1987,1992, 1995–2006, April –MayHerring gullLarus argentatusEggs10–13eggs/site/year pooledhomogenatesNot reported 4.5–20 ng/g RangeJaspers et al. 2006 Belgium 2003–2004 Aquatic birds (heron/grebe) Multiple 12 All <2.25 ng/gKunisue et al. 2008Open sea (Asia) 1998–1999Black-footed albatrossDiomedea nigripesMuscle 5 All <0.5 ng/gDetectionlimitKunisue et al. 2008Open sea (Asia) 1998Laysan albatrossDiomedea immutabilisMuscle 5 All <0.5 ng/gDetectionlimitKunisue et al. 2008Open sea (Asia) 1999Northern fulmarFulmarus glacialisMuscle 5 0 <0.5_4.8 (1.9) ng/gRange(mean inbrackets)Kunisue et al. 2008 Japan (Coastal) 1999–2001Black-tailed gullLarus crassirostrisMuscle 5 All <0.5 ng/gDetectionlimitLam et al. 2007 China – Hong KongMarch – May2004Little egretEgretta garzettaEggs 5 Not reported2.6±1.6 (< 0.5–3.8)ng/gMean±standarddeviation(range)0.5Lam et al. 2007China – Hong KongMarch–May2004Black-crowned night heronNycticorax nycticoraxEggs 5 Not reported14±25 (0.62–59)ng/gMean±standarddeviation(range)0.5Lam et al. 2007China – XiamenMarch–May2004Little egretEgretta garzettaEggs 5 Not reported0.89±0.64 (0.42–2.0)ng/gMean±standarddeviation(range)0.5Lam et al. 2007 China – Xiamen March–May2004Black-crowned night heronNycticorax nycticoraxEggs 5 Not reported1.8±2.2 (< 0.5–5.2)ng/gMean±standarddeviation(range)0.5Lam et al. 2007 China – Xiamen March–May2004Chinese pond heronArdeola bacchusEggs 5 Not reported99±130 (3.1–290)ng/gMean±standarddeviation(range)0.5Lam et al. 2007 China – Xiamen March–May2004Cattle egretBubulcus ibisEggs 5 Not reported28±36 (< 0.5–75)ng/gMean±standard0.5126Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimitdeviation(range)Lam et al. 2007QuanzhouMarch–May2004Little egretEgretta garzettaEggs 5 Not reported1.6±0.55 (0.88–2.3)ng/gMean±standarddeviation(range)0.5Lam et al 2007QuanzhouMarch–May2004Black-crowned night heronNycticorax nycticoraxEggs 5 5 < 0.5 ng/gMean±standarddeviation(range)0.5Verrault et al. 2005 Svalbard, Norway 2004Glaucous gull (male)Larus hyperboreusPlasma 12 50% < 0.05–0.21 ng/g RangeVerrault et al. 2005 Svalbard, Norway 2004Glaucous gull (female)Larus hyperboreusPlasma 15 64% < 0.05–0.33 ng/g RangeVerreault et al. 2004 Bjoroya, Norway 2002–2004Glaucous gullLarus hyperboreusPlasma 89 410 ng/g MeanVerreault et al. 2004 Bjoroya, Norway 2002–2004Glaucous gullLarus hyperboreusEggs 4 pools 23–53 ng/g RangeVerreault et al. 2004 SvalbardGlaucous gullLarus hyperboreusLiver 20 nd–170 ng/g RangeOTHER BIRDSKunisue et al.2008 Japan (Inland) 2001Common cormorantPhalacrocorax carboMuscle 4 0 < 0.5–3.5 (1.4) ng/gRange(mean inbrackets)Kunisue et al.2008 Japan (Inland) 1998Jungle crowCorvus macrorhynchosMuscle 5 0 36–1800 (440) ng/gRange(mean inbrackets)Mariussen et al.2004Norway 1990–1993 Grouse Liver 0.5 ng/g MeanVoorspoels et al.2006bBelgium 2001, 2003 PasserinesLiver andmuscle25 All < 17 ng/gVEGETATIONSchlabach et al. Skoganvarre, 2002Moss n/a 1 0.025 ng/g Mean127Concentration DataStudy Authors Location Date Species Tissue Sampling N Not Detected Lipid Weight Wet Weight Units StatisticDetectionLimit2002 NorwaySchlabach et al.2002Valvik, Norway 2002 Moss n/a 1 0.12 ng/g Mean128Appendix B – Bioaccumulation, Biomagnification and Trophic Magnification DataSpecies, Food Web orPredator:PreyEcosystem Study Type Tissue Location Endpoint Result Statistic Reference NotesPelagic food web(zooplankton, whitefish, whitesucker, emerald shiner,goldeye, burbot, walleye)(Calanoid sp., Coregonusclupeaformis, Catostomuscommersoni, Notropisatherinoides, Hiodonalosoides, Lota lota,Stizostedion vitreum)Aquatic Field Whole body Lake WinnipegTMF (lipidnormalized)3.6RegressionslopeLaw et al. 2006Statistically significant slope(P < .001)Fish:fish Aquatic Field Whole body Lake WinnipegBMFs (lipidnormalized)0.1–6.8 Range Law et al. 2006Generally superseded by TMFfrom the same system but stilluseful supporting infoFish:invertebrates AquaticField Whole body Lake WinnipegBMFs (lipidnormalized)0.2–34 Range Law et al. 2006Generally superseded by TMFfrom the same system but stilluseful supporting infoPelagic food web(zooplankton, sprat, herring,salmon)(Sprattus sprattus, Clupeaharengus, Salmo salar)Aquatic Field Whole body Baltic SeaTMF (lipidnormalized)~0.6 (notsignificantlydifferent from 0)RegressionslopeBurreau et al.2006Lowest trophic level(zooplankton) had muchhigher concentrations thanfish; the feeding relationshipswere not describedPelagic food web (roach,perch, pike)(Rutilus rutilus, Percafluviatilis, Esox lucius)Aquatic Field Whole body Baltic SeaTMF (lipidnormalized)~0.80 (notsignificantlydifferent from 0)RegressionslopeBurreau et al.2004Lowest trophic level(zooplankton) had muchhigher concentrations thanfish; the feeding relationshipswere not describedFish:fish Aquatic FieldWhole body Baltic SeaBMFs (lipidnormalized)0.02–1.3 RangeBurreau et al.2004Fish:zooplankton Aquatic Field Whole body Baltic SeaBMFs (lipidnormalized)0.04–0.11 RangeBurreau et al.2006Unclear if fish feed on sampledzooplanktonFish:fish Aquatic FieldWhole body Baltic SeaBMFs (lipidnormalized)1.71–5 RangeBurreau et al.2006Juvenile Lake troutSalvelinus namaycushAquatic Lab Whole body naBMF (lipidnormalized)0.3 Mean Tomy et al. 2004It is unlikely that steady statewas reached during theexperiment and the BMF wasestimated using the kineticmethod129Species, Food Web orPredator:PreyEcosystem Study Type Tissue Location Endpoint Result Statistic Reference NotesCarpCyprinus carpioAquatic Lab Whole body n/a BCF (wet weight) < 3 000 MITI 1992Tissue MDL divided by watersolubility, high uncertaintyRainbow troutSalvelinus namaycushAquatic Lab Carcass n/aBMF (lipidnormalized)0.022 RangeStapleton et al.2006decaBDE onlyRainbow troutSalvelinus namaycushAquatic Lab Carcass n/aBMF (lipidnormalized)0.089 MeanStapleton et al.2006decaBDE + BDE metabolitesRainbow troutSalvelinus namaycushAquatic Lab Serum n/aBMF (lipidnormalized)0.35 MeanStapleton et al.2006decaBDE onlyRainbow troutSalvelinus namaycushAquatic Lab Serum n/aBMF (lipidnormalized)0.52 MeanStapleton et al.2006decaBDE + BDE metabolitesRainbow troutSalvelinus namaycushAquatic Lab Liver n/aBMF (lipidnormalized)1.28 MeanStapleton et al.2006decaBDE onlyRainbow troutSalvelinus namaycushAquatic Lab Liver n/aBMF (lipidnormalized)1.73 MeanStapleton et al.2006decaBDE + BDE metabolitesPolar cod:ice ampipodBoreogadus saida:Gammarus wilkitzkiiMarine Field Whole bodySvalbard,NorwayBMF (lipidnormalized)0.03 MeanSørmo et al.2005Grey sealsHalichoerus grypusMarinemammalCaptive Blood North SeaBMF (lipidnormalized)8.3–20.8 RangeThomas et al.2005Grey sealsHalichoerus grypusMarinemammalCaptive Blubber North SeaBMF (lipidnormalized)0.025–0.063 RangeThomas et al.2005Based on day 30concentrationGrey sealsHalichoerus grypusMarinemammalCaptive Blubber North SeaBMF (lipidnormalized)0.044–0.11 RangeThomas et al.2005Based on day 83concentration (unlikely torepresent steady-stateconditions)Grizzly bearUrsus arctos horribilisTerrestrial FieldFat andmuscleCoastal andinterior BritishColumbiaBioaccumulationslopenegative (notsignificantlydifferent from 0)RegressionslopeChristensen etal. 2005A positive bioaccumulationslope would provide evidenceof biomagnification or trophicmagnificationEarthwormPhylum annelidaTerrestrial Lab Whole bodySoil BSAF(lipidand OCnormalized)0.044–0.73 (0.3) Range (mean)Sellstrom et al.2005Rat Terrestrial Lab Whole body n/a BMF (wet weight) 0.05 MeanHuwe and Smith2007a, bCow Terrestrial Lab Adipose n/a BMF (lipid0.037 Mean Kierkegaard etBased on mean silage130Species, Food Web orPredator:PreyEcosystem Study Type Tissue Location Endpoint Result Statistic Reference Notestissue normalized) al. 2007 concentrationCow Terrestrial LabAdiposetissuen/aBMF (lipidnormalized)0.31 MeanKierkegaard etal. 2007Based on silage concentrationat end of experimentCow Terrestrial Lab Whole body n/aBMF (lipidnormalized)0.026 MeanKierkegaard etal. 2007Based on mean silageconcentrationCow Terrestrial Lab Whole body n/aBMF (lipidnormalized)0.21 MeanKierkegaard etal. 2007Based on silage concentrationat end of experimentSunfishLepomis gibbosusAquatic Field Whole bodyMarlowe Ck.,Roxboro, NCBSAF (lipid andOC normalized)0.0018 MeanLa Guardia et al.2007CrayfishCampbarus punctic ambarussp. CAquatic Field Whole bodyMarlowe Ck.,Roxboro, NCBSAF (lipid andOC normalized)0.013 MeanLa Guardia et al.2007Various (see comments) Aquatic Field Whole bodyPearl RiverDelta, ChinaBSAF (lipid andOC normalized)0 – 0.4 Range Xiang et al. 2007Large yellow croaker(omnivore), silvery pomfret(omnivore), flathead fish(benthivore), robust tonguefish(benthivore), Bombay duck(piscivore), jinga shrimp(benthic invertebrate), greasyback shrimp (benthicinvertebrate), mantis shrimp(benthic invertebrate)Fish (see comments) Aquatic Field Whole bodyRiver Vero,SpainBSAF (lipid andOC normalized)0.0011 to 0.0013 RangeEljarrat et al.2007Barbell and carpAquatic oligochaete Aquatic Lab Whole body n/a Accumulation MinimalCiparis and Hale2005Horse musselModiolus modiolusAquatic Field Whole bodyVancouverIsland, CanadaSediment BSAF(lipid and organiccarbonnormalized)0–3.53 RangeDeBruyn et al.2009Marine mammal: fish Marine FieldBlubber andwhole fishNorth Atlantic,coast of Maine,USABMF (lipidnormalized)0.67–1.3Range (ratio ofpredator tissueconcentration toprey tissueconcentration)Shaw et al. 2009Harbour seal(blubber):American plaice(whole body); Harbour seal(blubber):Atlantic mackerel(whole body); Harbour seal(blubber):White hake (wholebody)Marine food web: beluga Marine Field Blubber and Eastern TMF (lipid 0.3 (statistically Regression Tomy et al. 2008 Tissues sampled: blubber—131Species, Food Web orPredator:PreyEcosystem Study Type Tissue Location Endpoint Result Statistic Reference Noteswhale, narwhal, walrus, arcticcod, shrimp, clams,deepwater redfish, mixedzooplanktonDelphinapterus leucas,Monodon monoceros,Odobenus rosmarus,Boreogadus saida,Pandalusborealis,Hymenodaraglacialis,Mya truncata, andSerripes groenlandica,Sebastes mentella)wholeorganismCanadianArcticnormalized) significantnegative slope)slope beluga whale, narwhal andwalrus; whole organism—arctic cod, shrimp, clams,deepwater redfish, and mixedzooplanktonMarine mammal: fish Marine FieldBlubber andwholeorganismEasternCanadianArcticBMF (lipidnormalized)<1Range (trophiclevel adjustedratio of predatortissueconcentration toprey tissueconcentration)Tomy et al. 2008Beluga (blubber):cod (wholebody); Beluga (blubber):redfish(whole body)Fish: invertebrate Marine FieldBlubber andwholeorganismEasternCanadianArcticBMF (lipidnormalized)<1Trophic leveladjusted ratio ofpredator tissueconcentration toprey tissueconcentrationTomy et al. 2008Cod (whole body):zooplankton(whole body)Marine food web: belugawhale, ringed seal, arctic cod,Pacific herring, Arctic cisco,pelagic amphipod, and arcticcopepod (Delphinapterusleucas, Phoca hispida,Boreogadus saida, Clupeapallasi, Coregonusautumnalis, Thermistolibelluala, Calanushyperboreus)Marine FieldBlubber,liver, andwholeorganismWesternCanadianArcticTMF (lipidnormalized)Not statisticallysignificantRegressionslopeTomy et al. 2009Tissues sampled: blubber—Beluga whale, ringed seal;liver—arctic cod, Pacificherring, Arctic cisco; wholeorganism—pelagic amphipod,and arctic copepodMarine mammal: fish Marine FieldBlubber andliverWesternCanadianArcticBMF (lipidnormalized)0.3–0.9Range (trophiclevel adjustedratio of predatortissueconcentration toTomy et al. 2009Beluga whale (blubber):cod(liver); ringed seal(blubber):herring (liver);Beluga whale (blubber):Arcticcisco (liver)132Species, Food Web orPredator:PreyEcosystem Study Type Tissue Location Endpoint Result Statistic Reference Notesprey tissueconcentration)Fish: invertebrate Marine FieldLiver andwhole bodyWesternCanadianArcticBMF (lipidnormalized)4.8–12.7Range (trophiclevel adjustedratio of predatortissueconcentration toprey tissueconcentration)Tomy et al. 2009cod (liver):calanus (wholebody); cod (liver):thermisto(whole body)Freshwater food web: watersnake, northern snakehead,mud carp, common carp,crucian carp, prawn, andChinese mysterysnail(Enhydris chinensis, Channaargus, Cirrhinus molitorella,Cyprins carpio, Carassiusauratus, Macrobrachiumnipponense,Cipangopaludina chinensi)Aquatic Field Whole bodyReservoir nearQinyuan City,South ChinaTMF (lipidnormalized)0.26 (negativeslope, p=0.53)RegressionslopeWu et al. 2009Authors interpret result asstatistically significant133 Appendix C – Selection of Log K ow for QSAR Modelling of decaBDE

The log K

ow for commercial DecaBDE of 6.265 determined at 25°C by CMABFRIP (1997a) is frequently cited in the literatu re. This value was determined using the generator column method with a composite ma terial received from Great Lakes Chemical Corp., Albemarle Corp. and Bromine Co mpounds 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 con centration in octanol stock solution was 0.0738 g/L. This resulted in a K

ow value of 1.845x10 6 , or log K

ow of 6.265. It is of interest that the log K

ow 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 produc ts (6.59; MacGregor and Nixon 1997) using the same procedure. This is not expected, due to the increased hydrophobici ty and increased K

ow , which would be expected with increased brom ination. This also does not reflect available information respecting commercial DecaBDE’s solubility in water. CMABFRIP (1997c)

determined that the limit of wate r 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 pr edicted 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 K

ow by CMABFRIP (1997b). The European Communities (2002) also indicate that the presence of trace amounts of octanol in the water phas e of this study may have had an apparent effect on the water concentration and the log K

ow value. For these reasons, the log K

ow for

DecaBDE of 6.265 is considered too low and not appropriate for BAF and BMF modelling purposes.

Watanabe and Tatsukawa (1990) repor ted a significantly higher log K

ow of 9.97. Their approach, which applied to other PBDE s, showed a pattern of increased K

ow with increasing bromination using a re verse-phase HPLC method.

Overall, the range of reported and measured values of logK

ow for decaBDE indicate a significant degree of uncertainty and measurement error. Log K

ow 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.vccla b.org/lab/alogps/start.html). The comprehensive Syracuse Research Corporat ion database used for the KOWWIN program contains 11 135 values for log K

ow . The maximum measured value in this training set is 11.3 and the upper 99.5 percentile is 7.8 (pers onal communication from JA Arnot to Environment Canada 15 February 2008; unref erenced). Therefore, most of the model estimates for decaBDE are greater than presen tly measured values for all other organic chemicals. Analytical limitations and high leve ls of measurement error are the principle reasons for lack of log K

ow values generally above 8.0 (e.g., Wania and Dugani 2003). Wania and Dugani (2003) reported a log K

ow 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 the134 three solubility concept. This value was judged to represent a more realistic logK

ow value for modelling purposes and was used for bioaccumulation modelling in this report. The logK

ow value of 8.7 was further used to adjust the predicted logK

ow s of the expected metabolites of decaBDE using the Experi mental Value Adjustment (EVA) method outlined in the KOWWIN program. The EVA method takes into account structural differences between the metabolite and pa rent compound (decaBDE) and adjusts the preferred value of 8.7 accordingly.135 Appendix D – Model Input Paramete rs for BAF–QSAR and Terrestrial Biomagnification Model Predictions fo r DecaBDE Transformation Products

(a) BAF–QSAR input parameters for lower brominated PBDEs, hydroxylated BDEs and hydroxymethoxy-BDEs

Notes:

1 Estimated using KOWWIN using t he EVA method with a reference logK
ow of 8.7 for decaBDE.

2 Estimated using the KOAWIN estimation model and with input of corrected logK
ow from
KOWWIN EVA method.

3 Maximum measured value (Watanabe and Tatsukawa 1990).

4 k M based on Tomy et al. (2004) and normalized to weight of middle trophic level fish in Arnot and Gobas (2004) model and 15 o C. These metabolites are expected to undergo further metabolism; however, the actual ra tes of metabolism are unknown.

5 Information regarding the half-life and k M of nonaBDEs is lacking in the literature and a half-life of 26 d (equal to that of decaBDE) was theref ore assumed. Normalized to weight of middle trophic level fish in Arnot and Gobas (2004) model and 15 o 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 15 o 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 15 o 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 15 o 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 15 o C. Metabolite Log K
ow Median k M (1/d) (184 g fish at

15 o C

Hydroxymethoxy-octaBDEs 6.5 1 0.02 4

Hydroxymethoxy-hexaBDEs 4.7 1 0.02 4

Hydroxymethoxy-pentaBDEs 3.9 1 0.02 4

Hydroxymethoxy-heptaBDEs 5.6 1 0.02 4

Hydroxy-hexaBDEs 4.7 1 0.02 4

Hydroxy octaBDEs 6.4 1 0.02 4

Hydroxy-nonaBDEs 7.3 1 0.02 4

nonaBDEs 7.8 2 0.02 5

octaBDEs 6.9 3 0.03 6

heptaBDEs 6.0 2 0.01 7

hexaBDEs 5.1 2 0.001 8

pentaBDEs 4.3 1 0.001 9136 (b) Terrestrial biomagnification model input parameters for lower brominated PBDEs, hydroxylated BDEs and hydroxymethoxylated BDEs

Name and Structure Log K
ow Log K
oa E D 9 k M (1/d) (wolf BW

normalized 80 kg

Hydroxymethoxy octaBDE 6.5 1 17.2 2 99% 0.004 5

Hydroxymethoxy pentaBDE 3.9 1 13.4 2 99% 0.004 5

Hydroxymethoxy hexaBDE 4.7 1 14.6 2 99% 0.004 5

Hydroxymethoxy heptaBDE 5.6 1 15.9 2 99% 0.004 5

Hydroxy-hexaBDE 4.7 1 13.4 2 99% 0.004 5

hydroxy-octaBDE, 6.4 1 15.9 2 98% 0.004 5

hydroxy-nonaBDE 7.3 1 17.2 2 96% 0.004 5

NonaBDEs 7.8 3 13.7 2 90% 0.007 6

OctaBDEs 6.9 4 12.4 2 98% 0.003 7

HeptaBDEs 6.0 3 11.1 2 99% 0.004 8

HexaBDEs 5.1 3 9.8 2 99% 0.004 8

PentaBDEs 4.3 3 8.6 2 99% 0.004 8

1 Estimated using KOWWIN using t he EVA method with a reference logK
ow of 8.7 for decaBDE.

2 Estimated using KOAWIN and with input of corrected logK
ow 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 nonaBDE s 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 octaBD Es 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 hal f-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 E D for humans.

(c) BAF–QSAR and terrestrial biomagnification model input parameters for brominated

dibenzofurans

Product Smiles Log K
ow 1 Log K
oa 2 Aquatic

k M 3 (1/d) Mammal

k M 4 (1/d) E D

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 t he EVA method with a reference logK
ow of 8.7 for decaBDE.

2 Estimated using KOAWIN and with input of corrected logK
ow from KOWWIN EVA method.137 3 k M 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 and15 o C. These metabolites are expected to undergo furthe r 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 wei ght of wolf in Gobas model (80 kg); k M for these metabolites is illustrative only and is consis tent with a half-life of 26 d. These metabolites are expected to undergo further metabolism; howe ver, the actual rates of metabolism are

unknown.138Appendix E: Summary of Environmental Degradation and Debromination Studies for decaBDE(a) PhotodegradationStudy Authors Type of Study SubstancePurityDegradationMediatorExperimentalConditionsDegradation Rate DegradationProductsWatanabe andTatsukawa (1987)Photodegradation 97% decaBDEand 3%nonaBDEArtificial UV LightDissolved in organicsolventsNR Tri- to octaBDEs;possibly brominatedfuransSöderström et al.(2004)Photodegradation Traces of octaand nonaBDEsNatural sunlightand artificial UVlightSorbed to thin layer ofsilica gel, sand, soil orsedimentNR Reductivedebromination withproducts includingprimarily hexa- tononaBDEs; tetra- andpentaBDFsJafvert and Hua(2001)Photodegradation 98% Artificial andnatural sunlightSorbed to hydratedsurfaces of glass andsilica sand particles,humic acid-coated silicaparticles, and glasssurfaces in contact withaqueous solutions12–71% degradationover 60–72 hIdentification ofdebrominationproducts was largelyinconclusive, althoughthere was someevidence of theformation of hexa- tononaBDEs.Hua et al. (2003) Photodegradation NR Artificial sunlightor natural sunlightPrecipitated onto quartzglass, silica particlesand humic acid-coatedsilica particles; hydrated44–71% degradationover 60–72 h; humicacid slowed the rate ofdecaBDE decaySmall amounts ofnona- and octaBDEPalm et al. (2003) Photodegradation NR Artificial xenonlampsDispersed in organicsolventsHalf-life ~ 30 minutes Three isomers ofnonaBDE formed, thensix isomers ofoctaBDE, thendebromination toseveral isomers ofheptaBDE, and finallyto trace amounts ofhexaBDE.Approximately 75% ofdecaBDE degradationfollowed a pathway ofdebromination whilethe products of theremaining 25% couldnot be determined.139Study Authors Type of Study SubstancePurityDegradationMediatorExperimentalConditionsDegradation Rate DegradationProductsPalm et al. (2003) Photodegradation/oxidation NR Simulated sunlightor hydroxylradicalsSorbed to aerosol(silicon dioxide)< 6 x 1013cm3molecule1s-1Products notdeterminedBezares-Cruz etal. (2004)Photodegradation NR Natural sunlight decaBDE dissolved inhexane99% reduction in thedecaBDE concentrationin as little as 30 minutesProducts ranging fromtri- to nonaBDEs wereobserved.Eriksson et al.(2004).Photodegradation 98% decaBDE Artificial UV light Organic solvents /water mixtures4×10-4/s in themethanol/water mixture(half-life ~ 0.5 hr),6.5×10-4/s in methanol(half-life ~ 0.3 hr),8.3×10-4/s in puretetrahydrofuran (half-life~ 0.23 hr)3 nonaBDEs, at least 7octaBDEs, 8heptaBDEs and smallamounts of hexaBDEsEriksson et al.(2004).Photodegradation 98% decaBDE Artificial UV light Water / humic acidmixture3×10-5/s in water withhumic acid (half-life~ 6.4 hr3 nonaBDEs, at least 7octaBDEs, 8heptaBDEs and smallamounts of hexaBDEsGeller et al.(2006)Photodegradation 98% Artificial light Dissolved intetrahydrofuranNR Photolysis productsincluded hepta- tononaBDEs as well astri- to hexabrominateddibenzofuransKuivikko et al.(2006)Photodegradation >98.3% Natural sunlight Dissolved in iso-octane;modelling to determinerates in oceanHalf-life of 0.03 days iniso-octanePredicted half-lives inthe Baltic Sea of0.02 days (surface) and1.2 days (mixing layer)and in the AtlanticOcean of 0.09 days(mixing layer)NRKuivikko et al.(2006, 2007)Photodegradation > 98.3% Natural sunlight Dissolved in iso-octane;modelling to determinerates in oceanHalf-life of 0.03 days iniso-octane; mixing zonehalf-lives of 1.8 days(Baltic Sea) and0.4 days (AtlanticOcean)Mixing zone half-livesof 1.8 days (BalticSea) and 0.4 days(Atlantic Ocean),which were the samefor both decaBDEconcentrations140Study Authors Type of Study SubstancePurityDegradationMediatorExperimentalConditionsDegradation Rate DegradationProductsAhn et al. (2006a) Photodegradation 98% Artificial light andnatural sunlightSorbed tomontmorillonite,kaolinite, organiccarbon-rich naturalsediment (16.4% OCcontent), aluminumhydroxide, iron oxideand manganese dioxideFor artificiallight, halflives ranged from 36 to178 daysFor natural light, halflives for montmorillonite,kaolinite and sedimentwere 261, 408 and990 days, respectively(negligible degradationon other matrices)Identified products forkaolinite andmontmorilloniteexposed to sunlightincluded nonaBDEs(BDE208, -207 -206),octaBDES (BDE197,-196) as well as smallamounts of tri- toheptaBDEsStapleton andDodder (2006)Photodegradation NR Natural sunlight Either native decaBDEin house dust ordecaBDE spiked housedust2.3x10-3/hr in spikeddust and 1.7x10-3/hr innatural dust,corresponding tohalf-lives of 301 and408 h in sunlight,respectivelySpiked dust: lowerbrominated PBDEsincluding hepta-, octaand nonaBDEcongeners; 54% of thedegradation productswere not identifiedGerecke (2006) Photodegradation 98% Natural sunlight DecaBDE (BDE209)sorbed to kaolinite.Irradiated dry or inwaterHalf-lives of 76 and 73minutes weredetermined for dry andwet conditions,respectively; dependenton light penetrationLower brominatedPBDEs (under dryconditions);unidentified products(under wet conditions)Stapleton (2006b) Photodegradation NR Natural sunlight Either native decaBDEin house dust ordecaBDE spiked housedustHalf-life for decaBDEwas estimated at 216 hThree nonaBDEs, sixoctaBDEs and oneheptaBDE.Mass balance analysisof decaBDEdetermined thatapproximately 17% ofthe original decaBDEwas unaccounted for,suggesting theformation of alternative(unidentified) productsor volatilization oflower brominatedPBDEs141Study Authors Type of Study SubstancePurityDegradationMediatorExperimentalConditionsDegradation Rate DegradationProductsHagberg et al.(2006)Photodegradation NR Artificiallight DecaBDE (BDE209)dissolved in tolueneNR Mono- to hexasubstituted PBDFs; themajority of theproducts were tetra- tohexaBDFs; lowerbrominated PBDEs notmonitoredBarcellos daRosa et al. (2003)Photodegradation 98% Artificiallight DecaBDE (BDE209)dissolved in toluene3x10-4/s Hepta- to nonaBDEsKajiwara et al.(2008)Photodegradation NR Natural sunlight HIPS (high-impactpolystyrene) added totoluene and DecaBDE(BDE209) (100ug/L);toluene evaporated offin dark. A subset ofsamples hydrated withwater to examine effecton photolysis.Crushed TV casing alsotested.HIPS + DecaBDEsamples: 50% BDE209reduction after 1 week.Half life of BDE209 inHIPS estimated at 51days.Hydrated samplesshowed fasterdegradation thannonhydrated.No degradation in darkcontrols.No degradation ofBDE209 in crushed TVcasing samples during224 days of sunlightirradiation.HIPS + DecaBDEsamples: hexa to nonaBDE congenersincreased several-foldafter 1 week exposure,then levels remainedconstant or decreasedslightly.At end of exposure,proportion of BDE209of total PBDEs haddecreased from 90%to 44%.Study confirmedphotolytic formation oftri- to octaBDFs. TotalPBDFs increased (>40times) by day 7 ofexposure, withdecreasing BDE209(HIPS + DecaBDEsamples). Nodebrominationproducts measuredfrom crushed TVcasing samples.NR – not determined or not reported.142(b) Other Abiotic DegradationStudy Authors Type of Study SubstancePurityDegradationMediatorExperimentalConditionsDegradation Rate Degradation ProductsKeum and Li(2005)Reductive debromination NR DecaBDE (i.e.,BDE209) reducingagents – zerovalentiron, iron sulphide,and sodiumsulphide.Dissolved in ethylacetateUp to 90%transformation ofdecaBDE after 40 dStepwisedebromination withmono- to hexaBDEcongeners present after40 days (started withhigher brominatedPBDEs)Li et al. (2007) Reductivedebromination CommercialDE-83RDecaBDE(> 97%, GreatLakesChemical)Nanoscalezerovalent ironDissolved in acetone,distilled water added. 25± 0.5°CHalf-life was 2.5 h Stepwisedebromination to formtri- to nonaBDEsRahm et al. (2005) Hydrolysis (nucleophilic aromaticsubstitution)NR Reaction withsodium methoxideDissolved in methanol Half-life for thehydrolysis reaction was0.028 hNRAhn et al. (2006b) Metal oxide-mediateddebromination98% Birnessite Sorbed to birnessite inTHF:water andwater:catechol systemsTHF:water – >75%transformation ofdecaBDE in 24 hCatechol:water –degradation onlyobserved for highestconcentration ofcatechol (15%degradation over 23days)In THF:water –nonaBDEs (BDE207,-208), octaBDEs(BDE196, -197),heptaBDEs (BDE183,-190 and eightunknowns), hexaBDEs(BDE138, -153, -154and five unknowns),pentaBDEs (BDE99,-100, -118 and oneunknown), andtetraBDEs (BDE49,-47, -66).Not determined forwater:catecholNR – not determined or not reported143(c) BiodegradationStudy Authors Type of Study SubstancePurityDegradation Mediator ExperimentalConditionsDegradation Rate DegradationProductsMITI 1992 Biodegradation NR Activated sludge Aerobic conditionsNo degradation after 2weeksNRCMABFRIP(2001)Biodegradation 97.4%decaBDE,2.5%nonaBDE and0.04%octaBDESediment/water system Anaerobicconditions< 1% mineralizationobserved over 32weeksNRHe et al. (2006) Biodegradation 98%Anaerobic bacteria DecaBDE (i.e.,BDE209) dissolvedin TCE andinoculated withanaerobicculture/mediumDegradation onlyobserved with oneculture (withS.multivorans) in which0.1μM decaBDEdegraded to nondetectable levels over2-month experimentOcta- andheptaBDEs weredetectable at the endof the experimentKnoth et al.(2007)WWTP monitoring(Biodegradation)NR WWTP sludge (primary,secondary and digested)Field WWTP NR An increase in theproportion of lowerbrominated PBDEcongeners was notobserved, indicatingno transformation ofdecaBDE during thetotal WWTPretention timeGerecke et al.(2005)Biodegradation 97.9%decaBDE,2.1%nonaBDEsSewage sludge + primers (1 ormore of 4-bromobenzoic acid, 2,6-dibromobiphenyl,tetrabromobisphenol A,hexabromocyclododecane anddecabromobiphenyl)AnaerobicconditionsDecaBDE (i.e.,BDE209) decreasedby 30% within 238 dRate constant of 1 x10-3 d-12 nonaBDEs and 6octaBDEs. This wasindicative ofreductivedebromination; lossof bromine from thepara and metapositions144Study Authors Type of Study SubstancePurityDegradation Mediator ExperimentalConditionsDegradation Rate DegradationProductsGerecke et al.(2006)Biodegradation 97.9%decaBDE,2.1%nonaBDEsSewage sludge + single primers(either 2,6-dibromophenol or 4-bromobenzoic acid)AnaerobicconditionsHalf-lives of 700 days(with primer) and 1400days (without primer)were observed.Monitoring at anoperating WWTPfound that theconcentration ofDecaBDE in sludgedecreased betweenthe influent and outletstreams.In the experimentsusing a primer,decaBDEtransformed slowly toBDE208.La Guardia et al.(2007)WWTP monitoring(biodegradation)n/a WWTP Sludge NR Minimal evidence ofdebromination inWWTP sludge orsedimentsNRParsons et al.(2004)Reductive debromination insedimentsNR Anaerobic sedimentsAnaerobic sedimentsuspensions inanaerobic mediumSediments collectedfrom WesternScheldtSignificant decreasein decaBDE over 2months; however,results are highlyuncertain since similardecrease wasobserved in abioticcontrolNonaBDEs andpossibly lowerbrominated PBDEsParsons et al.(2007)Reductive debromination insedimentsNR Anaerobic sedimentsAnaerobic sedimentsuspensions inanaerobic mediumSediments collectedfrom WesternScheldtNo statisticallysignificant decrease indecaBDE over 260daysNonaBDEs detectedin decaBDE-spikedsamples, althoughdegradation ofdecaBDE was notsignificant145Study Authors Type of Study SubstancePurityDegradation Mediator ExperimentalConditionsDegradation Rate DegradationProductsTokarz III et al.(2008)Reductive debromination insedimentsNR Anaerobic sediments andcosolvent-enhanced biomimeticsystemNatural sedimentswith no detectablePBDEs collectedfrom Celery BogPark, WestLafayette, Indiana.PBDEs dissolved ina toluene solutionadded to sediments,then evaporated off.This mixture wasthen blended withwet sediments.Biomimeticexperiment involvedthe use of Tefloncapped glass vialswith 0.03 mM ofBDE209, -99 or -47mixed with 5.0 mMtitanium citrate and0.2 mM vitamin B12in 0.33 M TRIZMAbuffer solutioncontainingtetrahydrofuran.The biomimeticsystem demonstratedreductivedebromination atdecreasing rates withdecreasingbromination (e.g., halflife of 18 seconds forBDE209 and almost60 d for BDE47)In natural sedimentmicrocosms, the halflife for BDE209 wasestimated to rangefrom 6 to 50 years,with an average of 14years, based onobservations over 3.5yearsProposed pathwayfor both systemscombined: BDE209 >nonaBDEs (BDE206,-207 -208) >octaBDEs (BDE196,-197) > heptaBDEs(BDE191, -184, twounknownheptaBDEs) >hexaBDEs (BDE138,-128, -154, -153) >pentaBDEs(BDE119, -99) >tetraBDEs (BDE66,-47, -49) > triBDEs(BDE28, -17)Specifically, at theend of 3.5 years,their analysis ofBDE209 degradationin sedimentsidentified BDE208,-197, 196, -191,-128, -184, -184,138, and -128, aswell as threeunidentifiedoctaBDEs and twounidentifiedheptaBDEs146Study Authors Type of Study SubstancePurityDegradation Mediator ExperimentalConditionsDegradation Rate DegradationProductsKohler et al.(2008)Debromination in sediments NR Natural lake sediments Field – lakesediment coresCores showDecaBDEs appear inlayers correspondingto 1970s, >30 years.Despite highpersistence of Decain sediments, studysuggestsenvironmentaldebrominationoccurs, as shown bythe detection of ashift in congenerpatterns of octa- andnonaBDE insediments,compared to therespective congenerpatternsin technical PBDEproducts (e.g.presence ofBDE202). Suggestbiotic/abiotictransformationbetween the releaseof technical productsand final residues insediment.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