Canadian Environmental Protection Act, 1999 

Ecological Screening Assessment Report on 

Polybrominated Diphenyl Ethers (PBDEs) 

June 2006 

Environment Canada 

where x + y = 1 to 10 

Figure 1. PBDE structure 

Introduction 

The Canadian Environmental Protection Act, 1999 (CEPA 1999) requires the Minister of the 

Environment and the Minister of Health to conduct screening assessments of substances that meet the 

categorization criteria set out in the Act and Regulations to determine, in an expeditious manner, 

whether substances present or may present a risk to the environment or to human health. Based on the 

results of a screening assessment, the Ministers can propose taking no further action with respect to 

the substance, adding the substance to the Priority Substances List (PSL) for further assessment, or 

recommending that the substance be added to Schedule 1 of CEPA 1999 and, where applicable, the 

implementation of virtual elimination. 

A screening assessment involves an analysis of a substance using conservative assumptions to 

determine whether the substance meets the criteria as defined in section 64 of CEPA 1999. This 

ecological screening assessment examines various supporting information and develops conclusions 

based on a weight of evidence approach as required under Section 76.1 of CEPA 1999. The screening 

assessment does not represent an exhaustive review of all available data; rather, it presents the most 

critical studies and lines of evidence supporting the conclusions. One line of evidence includes 

consideration of risk quotients to identify potential for ecological effects. However, other concerns 

that affect current or potential risk, such as persistence, bioaccumulation, chemical transformation 

and trends in ambient concentrations, are also examined in this report. 

1 
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Environment Canada June 2006 

 

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Seven polybrominated diphenyl ethers (PBDEs) were identified in a pilot project list of 123 

substances for screening assessment under CEPA 1999, on the basis of their potential persistence 

and/or bioaccumulation in the environment and inherent toxicity to organisms. 

Data relevant to the ecological screening assessment of PBDEs were identified in original literature, 

review documents, and commercial and government databases and indices. 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 PBDEs. 

Ongoing scans were conducted of the open literature, conference proceedings and the Internet for 

relevant PBDE information. Information obtained as of October 2004 was considered for inclusion 

into this document, while that received between November 2004 and October 2005 was reviewed, but 

not generally added. The information obtained between November 2004 and October 2005 was 

found to support the conclusions of this report determined with information received up to October 

2004. In addition, an industry survey on PBDEs was conducted for the year 2000 through a Canada 

Gazette Notice issued pursuant to Section 71 of CEPA 1999. This survey collected data on the 

Canadian manufacture, import, uses and releases of PBDEs (Environment Canada 2003). 

Toxicological studies were also submitted by industry under Section 70 of CEPA 1999. 

This ecological screening assessment report and associated unpublished supporting working 

documentation was written by a team of Environment Canada evaluators at the Environmental 

Protection Branch, Pacific and Yukon Region, Vancouver, B.C., with the assistance of evaluators and 

management at the Existing Substances Branch, Gatineau, Quebec. The material in this report has 

been subjected to external review by Canadian and international experts selected from government 

and academia, including M. Alaee (Environment Canada, National Water Research Institute), L. 

Birnbaum (U.S. Environmental Protection Agency), C. de Wit (Stockholm University), S. Dungey 

(UK Environment Agency), R. Hale (College of William and Mary, Virginia), R. Law (UK Centre 

for Environmental, Fisheries and Aquaculture Science), F. Luckey (U.S. Environmental Protection 

Agency), J. Maguire (Environment Canada, National Water Research Institute), R. Norstrom 

(Environment Canada, National Wildlife Research Centre) and D. Stewart (Environment Canada, 

Ontario Region). 

The ecological and human health screening assessment reports were approved by the joint 

Environment Canada/Health Canada CEPA Management Committee. The supporting working 

documentation for the ecological assessment is available upon request by e-mail from 

ESB.DSE@ec.gc.ca. Information on ecological screening assessments under CEPA 1999 is available 

at http://www.ec.gc.ca/substances/ese. The supporting working documentation for the human health 

assessment is available upon request by e-mail from ExSD@hc-sc.gc.ca. Additional background 

information on health screening assessments conducted under this program is available at 

http://www.hc-sc.gc.ca/hecs-sesc/exsd/splash.htm. 

The critical information and considerations upon which the assessment is based are summarized 

below. 
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Environment Canada June 2006 

 

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Identity, Uses and Sources of Release 

PBDEs comprise a class of substances consisting of 209 possible congeners with 1–10 bromine 

atoms. The following seven PBDE homologues, present on the Domestic Substances list (DSL), were 

identified in the pilot project list of 123 substances and are considered in this assessment: 

• tetrabromodiphenyl ether (benzene, 1,1'-oxybis-, tetrabromo derivative; tetraBDE) (CAS No. 

40088-47-9); 

• pentabromodiphenyl ether (benzene, 1,1'- oxybis-, pentabromo derivative; pentaBDE) (CAS No. 

32534-81-9); 

• hexabromodiphenyl ether (benzene, 1,1'-oxybis-, hexabromo derivative; hexaBDE) (CAS No. 

36483-60-0); 

• heptabromodiphenyl ether (benzene, 1,1'-oxybis-, heptabromo derivative; heptaBDE) (CAS No. 

68928-80-3); 

• octabromodiphenyl ether (benzene, 1,1'-oxybis-, octabromo derivative; octaBDE) (CAS No. 

32536-52-0); 

• nonabromodiphenyl ether (benzene, 1,1'-oxybis-, nonabromo derivative; nonaBDE) (CAS No. 

63936-56-1); and 

• decabromodiphenyl ether; bis(pentabromophenyl) ether (benzene, 1,1'-oxybis[2,3,4,5,6-

pentabromo-; decaBDE) (CAS No. 1163-19-5). 

These PBDEs are found in three commercial mixtures, typically referred to as Pentabromodiphenyl 

Ether (PeBDE), Octabromodiphenyl Ether (OBDE) and Decabromodiphenyl Ether (DBDE). PeBDE 

is predominantly a mixture of pentaBDE, tetraBDE and hexaBDE congeners, but may also contain 

trace levels of heptaBDE and tribromodiphenyl ether (triBDE) congeners. OBDE is a mixture 

composed mainly of heptaBDE, octaBDE and hexaBDE, but may also contain small amounts of 

nonaBDE and decaBDE. Current formulations of DBDE are almost completely composed of 

decaBDE and a very small amount of nonaBDE. 

PBDEs are used mainly as additive flame retardants in polymer resins and plastics and, to a lesser 

extent, adhesives, sealants and coatings. Additive flame retardants are physically combined with the 

material being treated rather than chemically bonded as in reactive flame retardants; therefore, they 

are more susceptible, to a certain extent, to migration and loss from the polymer matrix. It has been 

estimated that approximately 90% or more of PeBDE produced globally is used in polyurethane 

foams in office and residential furniture, automotive upholstery, sound insulation and wood imitation 

products (WHO 1994; European Communities 2001; RPA Ltd. 2000). Most OBDE produced 

globally is added to polymers (mainly acrylonitrile butadiene styrene), which are then used to 

produce computers and business cabinets, pipes and fittings, automotive parts and appliances (WHO 

1994; European Communities 2003). DBDE is used as a flame retardant, to a large extent in highimpact polystyrene and other polymers, with broad use in computer and television cabinets and 

casings, general electrical/electronic components, cables and textile back coatings (OECD 1994;

European Communities 2002). 
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Environment Canada June 2006 

 

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The total worldwide market demand for PBDEs was about 67 390 tonnes in 2001, including 56 100 

tonnes of DBDE, 7500 tonnes of PeBDE and about 3790 tonnes of OBDE (BSEF 2003). There are 

significant differences in the usage of PBDEs by continent (see Table 1). The most apparent 

difference is that PeBDE is used almost exclusively in the Americas. 

Table 1. Market demand of PBDEs in 2001 (BSEF 2003) 

Americasa Europeb Asiac Commercial 

product Market 

demand 

Estimated 

consumption 

(tonnes) 

Market 

demand 

Estimated 

consumption 

(tonnes) 

Market 

demand 

Estimated 

consumption 

(tonnes) 

DBDE 44% 24 500 13% 7 600 43% 24 050 

OBDE 40% 1 500 16% 610 44% 1 680 

PeBDE 95% 7 100 2% 150 3% 250 a All countries in North, South and Central America were included. b

 All countries in Eastern and Western Europe were included. 

c

 Australia, New Zealand and the Indian subcontinent were included.

Results from a Section 71 Notice with Respect to Certain Substances on the Domestic Substances List 

(DSL) conducted for the year 2000 indicated that no PBDEs were manufactured in Canada, although 

approximately 1300 tonnes of PBDE commercial products (for manufacturing into finished articles) 

were imported into the country (Environment Canada 2003). Based on quantities reported, PeBDE 

was imported in the greatest volume, followed by DBDE. A very small amount of OBDE was 

imported into Canada in 2000. The volumes reported do not include quantities imported in finished 

articles. 

Various initiatives have resulted in significant changes in the global use of the PBDEs since 2001. 

The U.S. manufacturer of PeBDE and OBDE, Great Lakes Chemical Corporation voluntarily ceased 

its production of PeBDE and OBDE by December 31, 2004 (U.S. EPA 2005, Great Lakes Chemical 

Corp. 2005). ICL Industrial Products (2005) also announced complete termination of its production 

and sale of its OBDE product by the end of 2004. In addition, PeBDE and OBDE have been subject 

to a phase-out by the European Union (EU). In response to its risk assessments, the EU passed a 

Directive (2003/11/EC) which requires all member states to adopt laws that prohibit the marketing or 

use of any product containing more than 0.1% by mass of PeBDE or OBDE effective August 15, 

2004. While it is expected that these actions have resulted in significant changes in the global and 

Canadian use of PBDEs, many products currently in use will have been manufactured during or 

before 2004 using PeBDE and OBDE. 

PBDEs may be released to the environment during manufacturing and polymer processing 

operations, throughout the service life of articles containing them and at the end of article service life 

during disposal operations. 

Fate, Exposure and Effects 
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A summary of selected physical and chemical properties of the commercial PBDE products and their 

primary constituents is presented in Table 2. 

Table 2. Selected physical and chemical properties of commercial PBDEs and their constituents 

Property PeBDE OBDE DBDE 

Molecular weight 485.8 (tetraBDE) 

564.7 (pentaBDE) 

(WHO 1994) 

643.6 (hexaBDE) 

722.3 (heptaBDE) 

801.4 (octaBDE) 

(WHO 1994) 

880.4 (nonaBDE) 

959.2 (decaBDE) 

(WHO 1994) 

Physical state 

(20°C; 101.325 kPa) 

viscous liquid or semi-solid, 

white crystalline solid (pure 

isomers of pentaBDE) 

(European Communities 

2001) 

powder or flaked material 

(European Communities 

2003) 

crystalline powder 

(European Communities 

2002) 

Vapour pressure 

(21°C; Pa) 

4.69 × 10-5

(Stenzel and Nixon 1997) 

6.59 × 10-6

(CMABFRIP 1997a) 

1.58 x 10-6 - 4.68 x 10-7 

(hexa – heptaBDEs ; 25°C) 

Tittlemier et al. 2002) 

4.63 × 10-6

(CMABFRIP 1997e) 

2.95 x 10-9 

(estimated for decaBDE) 

(Wania and Dugani 2003) 

Water solubility 

(25°C; μg/L) 

13.3 

10.9 (tetraBDE) 

2.4 (pentaBDE) 

(Stenzel and Markley 1997) 

0.5 

(CMABFRIP 1997b) 

<0.1 

(CMABFRIP 1997f) 

Log Kow 6.57 

(MacGregor and Nixon 

1997) 

6.29 

(CMABFRIP 1997c) 

8.35-8.90 

(Watanabe and Tatsukawa 

1990)

6.27 

(CMABFRIP 1997g) 

9.97 

(Watanabe and Tatsukawa 

1990) 

Log Koa 10.53 - 11.31 

(tetra- and pentaBDEs) 

(Harner and Shoeib 2002) 

12.78 - 13.61 

(hepta- and octaBDEs) 

(Tittlemeier et al. 2002) 

14.44 - 15.27 

(estimated for nona- and 

decaBDE) 

(Tittlemier et al. 2002) 

Henry’s law constant 

(25°C; Pa·m3

/mol) 

11 

(European Communities 

2001) 

10.6 (estimated) 

(European Communities 

2003) 

>44 (estimated) 

(European Communities 

2002) 

With their low vapour pressures, very low water solubility and high octanol/water partition 

coefficient (log Kow) values, it is expected that PBDEs entering the environment will tend to bind to 

the organic fraction of particulate matter, soils and sediments. For instance, if it is assumed that equal 

quantities of pentaBDE are released to air, water and soil compartments, Level III fugacity modeling 

(EPI v. 3.10, Syracuse Research Corporation) indicates that much of the substance would be expected 

to partition to sediments and soils, with very little partitioning to water or air (see Table 3). If all 

pentaBDE is discharged to water, Level III fugacity modeling indicates that almost all of the 

substance would partition to sediments with only a very small proportion staying in the water 

column, or partitioning into air or soil compartments. If all pentaBDE were released to soil, the 

substance would remain almost exclusively in this environmental compartment. Partitioning 

characteristics for the other PBDEs subject to this assessment are expected to be very similar. 
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Table 3. Predicted partitioning of PentaBDE in the environment based on Level III Fugacity Modeling. 

Release scenario Predicted partitioning (%) 

Air Water Sediment Soil 

Equal quantities to air, 

water, soil 

0.2 1.2 59 40 

100% to air 1.07 0.4 21 77.5 

100% to water 8 x 10-5 1.93 98.1 0.006 

100% to soil 6.1 x 10-7 0.002 0.11 99.9 

The lower brominated PBDEs (tetra- to heptaBDEs) are slightly more soluble in water and have a 

greater propensity for volatilization and atmospheric transport than more highly brominated PBDEs. 

In the atmosphere, these homologues would tend to sorb to particulates. The higher brominated 

PBDEs are reported to have higher octanol-water (Log Kow) and air-water (Log Kaw) partition 

coefficients and a greater propensity to remain in solid form, and thus, transport would likely be in 

the form of particles. Researchers have noted that the transport of the lower brominated PBDEs may 

be characterized by a series of deposition/re-volatilization “hops” which are dependent on seasonally 

and diurnally fluctuating temperatures (Gouin and Harner 2003). 

Wania and Dugani (2003) examined the long-range transport potential of PBDEs using a number of 

models (i.e., TaPL3-2.10, ELPOS-1.1.1, Chemrange-2 and Globo-POP-1.1) and various physical and 

chemical properties (i.e., solubility in water, vapour pressure, log Kow, log Koa, log Kaw and estimated 

half-lives in different media). All models yielded comparable results, with tetraBDE showing the 

greatest potential for atmospheric transport and decaBDE the lowest transport potential. The 

researchers estimated a characteristic travel distance (CTD) ranging from 1,113 to 2,483 km for 

tetraBDE, 608 to 1,349 km for pentaBDE, and 480 to 735 km for decaBDE. The CTD was defined 

as the distance a parcel of air has traveled until 1/e or approximately 63% of the chemical has been 

removed by degradation or deposition processes (Gouin and Mackay 2002). 

In an earlier study, Dugani and Wania (2002) also used models to predict that of the various PBDE 

congeners, those with four to six bromine atoms would have a higher long-range transport potential 

than lower or higher brominated congeners. They found that the transport of lower brominated 

congeners is limited by their degradation in the atmosphere, while the transport of the more highly 

brominated congeners is limited by their low volatility. Atmospheric degradation is reduced at low 

temperatures, so some of the models may underestimate the long-range transport potential of the 

lighter congeners (Dugani and Wania 2002). 

As will be indicated later in this report, PBDE concentrations have increased exponentially in arctic 

biota over the past two decades and have been measured in Arctic air. This suggests efficient longrange atmospheric transport of PBDEs. 

PBDEs have been detected in all environmental media as well as sewage sludge (see Tables 4 and 5), 

and there is evidence that their levels in the North American environment are increasing. 
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Gouin et al. (2002) measured total PBDEs (sum of 21 congeners) ranging from 10 to 1300 pg/m3

 in 

air samples collected at a rural southern Ontario site in early spring of 2000. Total PBDEs (congeners 

not specified) up to 28 pg/m3 were detected in air samples from the Canadian Arctic collected over 

the period 1994-1995 (Alaee et al. 2000). 

Luckey et al. (2002) measured total (dissolved and particulate phases) PBDE (mono- to heptaBDE 

congeners) concentrations of approximately 6 pg/L in Lake Ontario surface waters in 1999. More 

than 60% of the total was composed of BDE47 (tetraBDE) and BDE99 (pentaBDE), with BDE100 

(pentaBDE) and BDEs 153 and 154 (heptaBDE congeners) each contributing approximately 5 to 8% 

of the total. Stapleton and Baker (2001) analyzed water samples from Lake Michigan in 1997, 1998 

and 1999 and found that total PBDE concentrations (BDEs 47, 99, 100, 153, 154 and 183) ranged 

from 31 to 158 pg/L. 

PBDEs have been detected in sediment and soil samples collected in North America, and high 

concentrations have been measured in sewage sludge. Kolic et al. (2004) determined levels of PBDEs 

in sediments from Lake Ontario tributaries flowing to Lake Ontario. The total PBDEs (tri-, tetra, 

penta-, hexa-, hepta- and decaBDEs) measured in sediment samples taken from fourteen tributary 

sites (6 reported) ranged from approximately 12 to 430 μg/kg dw. Of the reported sediment results, 

concentrations of tetra- to hexaBDEs ranged from approximately 5 to 49 μg/kg dw. Concentrations of 

BDE209 ranged from 6.9 to 400 μg/kg dw. BDE 47, 99 and 209 were the predominant congeners 

measured in sediments. Rayne et al. (2003a) measured PBDE concentrations (sum of 8 di- to 

pentaBDE congeners) ranging from 2.7 to 91 μg/kg OC in 11 surficial sediments collected in 2001 

from several sites along the Columbia River system in south eastern British Columbia. Domestic 

wastewaters arising from septic field inputs were identified as potentially dominant sources of 

PBDEs in the region. Dodder et al. (2002) reported concentrations of total tetra-, penta- and 

hexaBDEs ranging from approximately 5 to 38 μg/kg dw in sediment from a lake in the U.S. located 

near suspected PBDE sources. Preliminary results from a study by Muir et al. (2003) describe 

concentrations of BDE209 along a north-south transect from southern Ontario/upper New York state 

to Ellesmere Island. The highest concentrations of BDE209 (up to12 μg/kg dw) occurred in 

sediments collected from the western basin of Lake Ontario. However, sediments from two Arctic 

lakes in Nunavut Territory also had measurable concentrations of 0.075 and 0.042 μg BDE209/kg 

dw. One of the two Arctic lakes was located near an airport and so inputs of PBDEs from this source 

could not be ruled out. However, the second lake was completely isolated and was only visited for 

sampling purposes. The authors speculate that BDE209 was likely transported on particles to the 

Canadian Arctic due to its low vapour pressure and high octanol-water partition coefficient. Hale et 

al. (2002, 2003) reported concentrations of total PBDEs (tetra- and pentaBDE) of 76 μg/kg dw in soil 

near a polyurethane foam manufacturing facility in the United States, and 13.6 μg/kg dw in soil 

downwind from the facility. 

Kolic et al. (2004) determined levels of PBDEs in biosolids from southern Ontario municipal 

wastewater treatment plants (Reiner pers. comm.. 2004). They found total PBDEs (tri-, tetra-, penta-, 

hexa-, hepta- and decaBDEs) at five reported wastewater treatment facilities ranged from 

approximately 1,700 to 3,500 μg/kg dw. Of the reported biosolid results, total concentrations of tetrato hexaBDEs ranged from approximately 1,350 to 1,900 μg/kg dw. BDEs 47, 99 and 209 were the 
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predominant congeners measured in biosolid samples. Concentrations of BDE 209 in the samples 

ranged from 310 to 2000 μg/kg dw. La Guardia et al. (2001) analyzed 11 sludge samples before land 

application from a sewage treatment facility in the Toronto area and from10 facilities throughout the 

continental United States. Total PBDEs (sum of 11 tetra- to decaBDE congeners) in the samples of 

sewage sludge were 8280 μg/kg dw at the Toronto site, while those in the U.S. ranged from 730 to 

24, 900 μg/kg dw. Kolic et al. (2003) investigated PBDE levels in sewage sludge from 12 sites in 

southern Ontario and found concentrations of total PBDEs (21 mono- to decaBDE congeners) 

ranging from 1414 to 5545 μg/kg dw. Hale et al. (2002) measured total PBDEs (sum of BDEs 47, 99, 

100 and 209) of 3005 μg/kg dw in sludge samples collected in 2000 from a regional sewage 

treatment plant discharging to the Dan River in Virginia. 

Alaee et al. (1999) reported average concentrations in the blubber of marine mammals from the 

Canadian Arctic as 25.8 μg/kg lipid in female ringed seals (Phoca hispida), 50.0 μg/kg in the blubber 

of male ringed seals, 81.2 μg/kg lipid in female beluga (Delphinapterus leucus) and 160 μg/kg lipid 

in male beluga. In these samples, congeners of tetraBDE and pentaBDE were predominant. 

Ikonomou et al. (2000) reported PBDE concentrations in biota samples from the west coast and 

Northwest Territories of Canada. The highest concentration of total PBDE residues, 2269 μg/kg lipid, 

was found in the blubber of a harbour porpoise from the Vancouver area. With a concentration of 

about 1200 μg/kg lipid, a tetraBDE congener accounted for slightly more than half of the total PBDE 

in the sample. Ikonomou et al. (2002a,b) analyzed temporal trends in Arctic marine mammals by 

measuring PBDE levels in the blubber of Arctic male ringed seals over the period 1981–2000. Mean 

total PBDE concentrations increased exponentially from approximately 0.6 μg/kg lipid in 1981 to 6.0 

μg/kg lipid in 2000, a greater than 8-fold increase. TetraBDE was again predominant, followed by 

pentaBDE. A marked increase in tissue PBDE levels was also evident in blubber samples collected 

from San Francisco Bay harbour seals over the period 1989–1998 (She et al. 2002). Concentrations 

of total PBDEs (tetra-, penta- and hexaBDE) rose from 88 μg/kg lipid in 1989 to a maximum of 8325 

μg/kg lipid in 1998, a period of only 10 years. Stern and Ikonomou (2000) examined PBDE levels in 

the blubber of male southeast Baffin beluga whales over the period 1982–1997 and found that the 

levels of total PBDEs (tri- to hexaBDE) increased significantly. Mean total PBDE concentrations 

were about 2 μg/kg lipid in 1982 and reached a maximum value of about 15 μg/kg lipid in 1997. 

Total PBDE residues in the blubber of St. Lawrence estuary belugas sampled in 1997–1999 

amounted to 466 (± 230) μg/kg wet weight (ww) blubber in adult males and 665 (± 457) μg/kg ww 

blubber in adult females. These values were approximately 20 times higher than concentrations in 

beluga samples collected in 1988–1990 (Lebeuf et al. 2001). 
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Table 4. Measured concentrations of PBDEs in the North American ambient environment and sewage sludge 

Medium Location; year Total PBDEs Reference 

Air Alert, Canada; 1994–1995 1–28 pg/m3 Alaee et al. 2000 

Air Great Lakes; 1997–1999 5.5–52 pg/m3 Strandberg et al. 2001 

Air Southern Ontario; 2000 10–1300 pg/m3 Gouin et al. 2002 

Air Ontario; 2000 3.4–46 pg/m3 Harner et al. 2002 

Water Lake Michigan; 1997–1999 31–158 pg/L Stapleton and Baker 2001 

Water Lake Ontario; 1999 6 pg/L Luckey et al. 2002 

Sediment Lake Michigan; 1998 4.2 μg/kg dw Stapleton and Baker 2001 

Sediment British Columbia; 2001 2.7–91 μg/kg OC Rayne et al. 2003a 

Soil United States; 2000 <0.1–76 μg/kg dw Hale et al. 2002 

Sewage sludge Toronto, Canada 

United States 

8280 μg/kg dw 

730–24 900 μg/kg dw 

La Guardia et al. 2001 

Sewage sludge United States; 2000 3005 μg/kg dw Hale et al. 2002 

Sewage sludge Southern Ontario 1700-3500 μg/kg dw Kolic et al. 2004 

dw = dry weight; OC = organic carbon 

Table 5. Measured concentrations of PBDEs in North American biota 

Organism Location; year Total PBDEs Reference 

Dungeness crab 

hepatopancreas 

West coast, Canada; 1993–

1995 

4.2–480 μg/kg lipid Ikonomou et al. 2002b 

Mountain whitefish 

(muscle) 

Columbia River, British 

Columbia; 1992–2000 

0.726–131 μg/kg ww Rayne et al. 2003a 

Heron egg British Columbia; 1983–2000 1.308–288 μg/kg ww 

Murre egg Northern Canada; 1975–1998 0.442–2.93 μg/kg ww 

Fulmar egg Northern Canada; 1975–1998 0.212–2.37 μg/kg ww 

Wakeford et al. 2002 

Beluga whale blubber Canadian Arctic 81.2–160 μg/kg lipid Alaee et al. 1999 

Herring gull egg Great Lakes; 1981–2000 9.4–1544 μg/kg ww Norstrom et al. 2002 

Lake Ontario; 1997 95 μg/kg ww 

Lake Erie; 1997 27 μg/kg ww 

Lake Superior; 1997 56 μg/kg ww 

Lake trout 

Lake Huron; 1997 50 μg/kg ww 

Luross et al. 2002 

Rainbow trout 297 μg/kg ww 

Mountain whitefish 1250 μg/kg ww 

Largescale sucker 

Spokane River, Washington, 

USA; 1999 

105 μg/kg ww 

Johnson and Olson 

2001 

Carp Virginia, USA; 1998-1999 1140 μg/kg ww Hale et al. 2001 

These studies indicate that PBDE levels in Canadian biota are rising, with dramatic increases in 

tissue concentrations evident over the last two decades. The highest levels in biota are associated 

with industrialized regions; however, the increasing incidence of PBDEs in Arctic biota provides 

evidence for long-range atmospheric transport of these compounds (Stern and Ikonomou 2000). 

Although tetraBDE predominates in wildlife, there are recent indications of a shift in tissue congener 
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Environment Canada June 2006 

 

10

profiles. Ikonomou et al. (2002a) determined that over the period 1981–2000, penta- and hexaBDE 

levels in the blubber of Arctic ringed seals increased at rates that were roughly equivalent and about 

twice that of tetraBDE. 

There are indications from recent studies conducted in Europe that PBDE levels in some European 

biota may have peaked. Time trend analyses using Baltic guillemot (Uria aalge) eggs (Sellström 

1996; Sellström et al. 2003) and pike (Esox lucius) from Lake Bolmen in Sweden (Kierkegaard et al. 

2004) show a leveling off and possible decline in the concentrations of penta-like congeners 

beginning in the early 1990s. Any observed reduction in the concentrations of PBDEs in European 

biota may be a consequence of recent reductions in the production and use of commercial PeBDE 

throughout Europe. For further discussion of this issue, the reader should consult references such as 

de Wit (2002) and Law et al. (2003). 

An analysis of archived herring gull eggs (sampled in 1981, 1983, 1987, 1988, 1989, 1990, 1992, 

1993, 1996, 1998, 1999 and 2000) enabled Norstrom et al. (2002) to establish temporal trends in 

PBDE concentrations between 1981- 2000. At Lake Michigan, Lake Huron and Lake Ontario sites, 

concentrations of total tetra- and pentaBDEs increased 71 to 112 fold over the 1981 to 2000 period 

(from 4.7 to 400.5 μg/kg ww at Lake Ontario; from 8.3 to 927.3 μg/kg ww at Lake Michigan; from 

7.6 to 541.5 μg/kg ww at Lake Huron). These increases were found to be exponential at all three 

locations. Overall, the total PBDEs ranged from a low of 9.4 μg/kg ww in Lake Ontario to a high of 

1544 μg/kg ww Lake Michigan in 1998. These increases were largely due to the tetra- and 

pentaBDE congeners, but hexa- and heptaBDEs also increased during this period. 

Recent studies conducted in Europe provide evidence for the presence of decaBDE in biota. 

DecaBDE 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). De Boer et al. (2004) 

conducted sampling 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. In total, 124 samples from 

13 different species were analyzed. In addition, 10 peregrine falcon egg samples from the Swedish 

study by Lindberg et al. (2004) were re-analyzed. DecaBDE was detected in 10 of 28 liver samples 

(range < 1.5 to 181 μg/kg lipid weight), 14 of 28 muscle samples (range < 4.2 to 563 μg/kg lipid 

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

peregrine falcon eggs re-analyzed in the study were all within 30% of those originally determined by 

Lindberg et al. (2004). Highest concentrations of decaBDE were measured in muscle tissue samples 

collected from United Kingdom heron and peregrine falcon, and eggs from Swedish peregrine falcon. 

Empirical and predicted data indicate that all PBDEs subject to this ecological screening assessment 

are highly persistent, and each satisfies the requirements for persistence as defined by the Persistence 

and Bioaccumulation Regulations under CEPA 1999 (see Table 6). Tetra- to decaBDEs are predicted 

by AOPWIN (v1.90) to have air degradation half-lives which exceed 2 days (i.e., ranging from 7.14 

to 317.53 days). Further, tetra-, penta-, hexa-, hepta- and decaBDEs have been measured in the Arctic 

environment in spite of their very low vapour pressures, providing evidence that they are subject to 

long-range atmospheric transport. It has been shown that BDE 47 and DBDE are not subject to 

statistically significant anaerobic biodegradation over a period of 32 weeks. Neither PeBDE, OBDE 

nor DBDE are readily biodegradable based on short-term studies conducted under aerobic conditions 
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using an activated sludge inoculum. However, decaBDE is susceptible to some biodegradation under 

certain anaerobic conditions using sludge inoculum as described by Gerecke et al. (2005). Tetra- to 

decaBDEs are predicted by BIOWIN (v.4.00) to be recalcitrant with respect to biodegradation. It is 

reasonable to conclude that all PBDEs subject to this assessment meet the criteria for persistence as 

defined by CEPA 1999 based on known empirical and predicted data, as well as structural 

similarities. 

Table 6. Persistence and bioaccumulation criteria as defined in CEPA 1999 Persistence and Bioaccumulation Regulations 

(Environment Canada 2000) 

Persistencea

Medium Half-life 

Bioaccumulationb

Air 

Water 

Sediment 

Soil 

≥2 days or is subject to atmospheric transport from its source to a remote area 

≥182 days (≥6 months) 

≥365 days (≥12 months) 

≥182 days (≥6 months) 

BAF ≥ 5000 

BCF ≥ 5000 

log Kow ≥ 5 

a

 A substance is persistent when at least one criterion is met in any one medium. 

b

 When the bioaccumulation factor (BAF) of a substance cannot be determined in accordance with generally recognized 

methods, then the bioconcentration factor (BCF) of a substance will be considered; however, if neither its BAF nor its 

BCF can be determined with recognized methods, then the log Kow will be considered. 

Although all PBDEs subject to this assessment are considered to be persistent, evidence shows that 

PBDEs are susceptible to some degree of abiotic and biotic transformation under certain laboratory 

conditions. 

The predominant phototransformation pathway for decaBDE in organic solvents appears to be 

reductive debromination, with nona- to triBDEs and polybrominated dibenzofurans (PBDFs) 

identified as possible phototransformation products (e.g., Norris et al. 1973, 1974; Watanabe and 

Tatsukawa 1987; Eriksson et al. 2001; Palm et al. 2003; Herrmann et al. 2003; Hua et al. 2003; 

Peterman et al. 2003). Researchers also report the formation of other as yet unidentified 

photodegradation products. The relevance of these studies, which disperse PBDEs in organic solvents 

such as hexane and octanol, to conditions existing in the environment is still uncertain. 

Studies using more environmentally relevant media have also been conducted. Söderström et al. 

(2004) undertook photodegradation studies in which DBDE (exact composition not provided, but 

contained traces of octa- and nonaBDEs) was dissolved in toluene and then applied as a thin layer to 

silica gel, sand, soil or sediment substrates. The toluene solvent was evaporated off in the dark prior 

to exposure of the substrates to ultraviolet (UV) light or natural sunlight. Prior to light exposure, a 

small amount of water was added to the sediment in order to more closely emulate natural conditions. 

DBDE applied to silica gel decayed quickly under artificial and natural lighting, with an estimated 

half life of less than 15 min. The half-life of DBDE on sand was 12 and 13 h under UV and natural 

sunlight, respectively, while that of DBDE on sediment was 40-60 and 30 h under UV and sunlight, 

respectively. Overall, decay proceeded slowest with DBDE on soil exposed to UV light, with a halflife of 150-200 hours. The researchers concluded from their experiments that the photodegradation of 

decaBDE, at least initially, seems to follow a stepwise debromination process. They noted that as 

decaBDE disappeared, lower brominated DEs (nona- to hexaBDEs) were formed, but that after the 
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maximum occurrence of hexaBDEs, only minor amounts of lesser brominated DEs (tetra- and 

pentaBDEs) were formed, resulting in a discontinued mass balance. This suggested that other 

unknown compounds were also being formed, but that these were lost during the sample clean-up. In 

addition to the identified PBDEs, tetra- and pentaBDFs were also detected as transformation products 

of DBDE adsorbed to sand, sediment and soil. 

Jafvert and Hua (2001) conducted photodegradation studies of DBDE adsorbed to solid matrices 

(sand and quartz surfaces) with water and/or humic acid and irradiated with natural or artificial 

sunlight. Their studies showed that some photodegradation of DBDE occurred under natural or 

artificial sunlight (over time periods up to 240 h loss of decaBDE varied up to 71%). Although 

Jafvert and Hua (2001) did not conclude that lower brominated DEs were produced, the European 

Communities (2002), based on their review of the decaBDE humic acid coated sand exposure, noted 

that there were indications that lower brominated DEs (particularly hexaBDE) were formed. Palm et 

al. (2003) irradiated decaBDE adsorbed onto silicon dioxide in aqueous suspension with artificial 

sunlight. They also found that approximately 50% of the initial decaBDE concentration was lost after 

about 360 min. Details regarding the degradation products were not provided; however, Palm et al. 

(2004) notes that PBDFs were confirmed as short-lived trace intermediates. 

Keum and Li (2005) investigated the debromination of PBDEs (including decaBDE) in contact with 

the reducing agents, zerovalent iron, iron sulphide and sodium sulphide. In the experiments with 

zerovalent iron, decaBDE was rapidly transformed to lower BDEs. Approximately 90% of the parent 

was converted to mono- to hexaBDEs after 40 d. During the initial reaction period (up to 5 d), BDE 

209 was predominantly transformed into hexa- and heptaBDEs, but tetra and pentaBDEs were 

predominant after 14 d. The results demonstrated that decaBDE undergoes reductive debromination 

in the presence of zerovalent iron. The experiments with sodium sulphide also showed transformation 

of decaBDE to lower brominated DEs, but the rate was slower than that determined in the presence of 

zerovalent iron. A similar profile of transformation products was found to that determined in the 

experiment with zerovalent iron. Experiments were also conducted with BDEs 28, 47, 66 and 100 in 

the presence of zerovalent iron. These also showed that debromination had occurred but that the rate 

of reaction decreased with a decreasing number of bromines. Although the conditions of this study 

are not directly related to those common in the natural environment, it is possible that similar 

reactions maybe taking place in the environment (United Kingdom 2005). 

Gerecke et al. (2005) conducted experiments to determine the rates of degradation of decaBDE and 

nonaBDEs under anaerobic conditions conducted in the dark at 37 °C using sewage sludge as 

inoculum. The researchers found that BDE 209 decreased by 30% within 238 d in experiments with 

primers added (i.e., 4-bromobenzoic acid, 2,6-dibromobiphenyl, tetrabromobisphenol A and 

hexabromocyclododecane) and this corresponded to a pseudo-first-order degradation rate constant of 

1 x 10-3 d-1, statistically significant at the 95% confidence level. The sample with decaBDE was 

observed to form two nonaBDE and six octaBDE congeners. The rate of decaBDE decay without 

primers added was about one-half that of the experiments with primers. The study demonstrated that 

the debromination of decaBDE proceeded most readily by the loss of bromine from the para- and 

meta-positions. The United Kingdom (2005) notes that the conditions in themselves are not 

representative of sewage sludge treatment processes, or those typical in the natural environment. 

However, such conditions could occur in landfill sites which are anaerobic, methanogenic and have 
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13

high temperatures. The study provides evidence that decaBDE could be susceptible to some level of 

slow degradation under conditions of anaerobic biodegradation. 

PBDE congener patterns found in the environment are sometimes reported to resemble those of the 

PeBDE and OBDE commercial products, leading some researchers (e.g., Song et al. 2004) to propose 

that these products are the primary sources of PBDEs into the environment. Rayne and Ikonomou 

(2002) placed semipermeable membrane devices (SPMD) in the Fraser River, BC and analyzed the 

resultant SPMD samples for 36 PBDEs (mono- to hexa- congeners). They found that the congener 

patterns observed in the SPMD samples differed significantly from those of the commercial PeBDE 

and OBDE mixtures. They then applied modeling and calculation procedures and found that the 

reconstructed congener patterns more closely approximated those of the technical mixtures. These 

analyses lead the researchers to suggest that the PBDEs present in the region arose primarily from 

PeBDE and OBDE mixtures.. 

Söderström et al. (2004) also concluded that the lower brominated DEs (e.g., BDE 47, 154 and 183) 

found in the environment probably originate mainly as emissions from the commercial PeBDE and 

OBDE mixtures rather than DBDE phototransformation. In their studies they note that the most 

commonly found PBDEs in environmental samples (BDE 47, 99 and 100) were only formed to a very 

minor degree during their photolysis studies. However, it should be noted that most monitoring 

studies to date have only investigated PBDEs for which standards are available. These PBDEs are 

also the main components of the commercial products. As a result, one can expect that the results 

reported for environmental samples would predominantly be for PBDEs present in the commercial 

products. Analytical standards are not available for all congeners, and thus, it may be that studies 

conducted to date have not investigated all congeners present in environmental samples, including 

those occurring as photodegradation products of decaBDE. 

Studies have shown the transformation of higher brominated PBDEs (e.g., hepta- to decaBDEs) to 

lower brominated congeners (e.g., tetra- to hexaBDEs), which are associated with high levels of 

bioaccumulation. A dietary exposure study has shown that congeners of heptaBDE and pentaBDE 

rapidly biotransform in the gut of carp (Cyprinus carpio), and at least 10–12% is debrominated to 

congeners of hexaBDE and tetraBDE, respectively (Stapleton et al. 2004b,c; Stapleton and Baker 

2003). These transformation products then accumulate in the tissues of the carp. Carp have also 

demonstrated a limited ability to biotransform decaBDE when exposed via food, producing various 

penta- to octaBDE congeners. In a study described by Stapleton et al. (2004a) and Stapleton and 

Baker (2003), approximately 0.4% of consumed decaBDE was shown to biotransform in carp to form 

penta- to octaBDEs. The researchers note that while this amount may appear insignificant, high 

concentrations of decaBDE reported in river sediment and land-applied sludge could lead to 

appreciable accumulation in organisms exposed to such material (Stapleton et al. 2004a). 

 

While conditions of laboratory experiments showing that decaBDE will transform to lesser 

brominated DEs are not completely representative of those in the natural environment or sewage 

treatment facilities, they indicate that some degree of transformation cannot be ruled out. Globally, 

DBDE has become the most used technical PBDE product (see Table 1).There is a weight of 

evidence suggesting that highly brominated PBDEs such as decaBDE are precursors of the more 

toxic, bioaccumulative and persistent lower brominated PBDEs. While the degree to which this 
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14

phenomenon adds to the overall risk presented to organisms from formation of the more toxic and 

persistent tetra- to hexaBDE congeners is not known, there is sufficient evidence to warrant concern. 

Measured data indicate that tetra-, penta- and hexaBDE are highly bioaccumulative, with 

bioconcentration factors (BCFs) exceeding 5000 for aquatic species; thus, they satisfy the criteria for 

bioaccumulation as described in the CEPA 1999 Persistence and Bioaccumulation Regulations (see 

Table 6). A BCF of about 27 400 L/kg for PeBDE was reported by European Communities (2001), 

based on a recalculation of data contained in a study by CITI (1982), in which carp were exposed for 

8 weeks to PeBDE at 10 or 100 μg/L. This BCF for the commercial product was driven by a high 

BCF calculated for the tetraBDE component. The recalculated BCFs for the various components 

were 66 700 L/kg for tetraBDE, 17 700 and 1440 L/kg for separate pentaBDE congeners (identities 

not provided) and 5640 and 2580 L/kg for separate hexaBDE congeners (identities not provided). A 

bioaccumulation factor (BAF) of 1.4 × 106

 was reported for PeBDE in blue mussels (Mytilus edulis) 

exposed for 44 days (Gustafsson et al. 1999). The same study reported BCFs of 1.3 × 106

 for 

tetraBDE and 2.2 × 105

 for hexaBDE in these organisms. High rates of accumulation in biota are 

supported by high log Kow values for PBDEs and reports of biomagnification of tetraBDE and 

pentaBDE in aquatic food chains (e.g., Alaee and Wenning 2002; de Wit 2002). 

Key studies of toxicity to organisms in different environmental media are presented in Table 7. Since 

testing is frequently carried out using commercial mixtures, effects must frequently be best 

considered in relation to the total exposures to all congeners involved (see below). 

Risk Characterization 

The approach taken in this ecological screening assessment was to examine various supporting 

information and develop conclusions based on a weight of evidence approach as required under 

Section 76.1 of CEPA 1999. Particular consideration was given to risk quotient analyses and 

persistence, bioaccumulation, chemical transformation and trends in environmental concentrations. 

This assessment has used data corresponding to commercial products, individual congeners and 

homologues/isomer groups. The presentation of data and the risk quotient analyses have been 

structured around the PBDE commercial products since a great deal of empirical data which are 

central to this assessment (e.g., relevant to environmental toxicity) have been determined using the 

commercial products. Nonetheless, the risk analysis and scientific evidence presented in this report 

relate to all congeners found in the commercial products, PeBDE, OBDE and DBDE. 

The risk determined for each commercial product is a result of the combined activity of the various 

co-occurring PBDEs, adding complexity to the interpretation of the results. Due to these reasons, 

their common chemical structure, and due to issues relating to their chemical transformation, PeBDE, 

OBDE, DBDE and their brominated constituents are assessed as a group. 

Risk quotient analyses, integrating known or potential exposures with known or potential adverse 

environmental effects, were performed for each of the commercial PBDE products subject to this 
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15

assessment. An analysis of exposure pathways and subsequent identification of sensitive receptors 

were used to select ecological assessment endpoints (e.g., adverse reproductive effects on sensitive 

fish species in a community). For each endpoint, a conservative Estimated Exposure Value (EEV) 

was selected based on empirical data from monitoring studies. Where monitoring data were not 

available, the EEVs were based on simple calculation procedures taking into account some degree of 

local environmental conditions, but largely relying on generic environmental parameters. Chemical 

concentrations from the Canadian and North American environment were used preferentially for 

EEVs; however, data from other regions in the world were used in the absence of sufficient Canadian 

data of satisfactory quality or to provide a weight of evidence. EEVs usually represented worse-case 

scenarios, as an indication of the potential for these substances to reach concentrations of concern 

and to identify areas where those concerns would be most likely. 

An Estimated No-Effects Value (ENEV) was also determined by dividing a Critical Toxicity Value 

(CTV) by an application factor. CTVs typically represented the lowest ecotoxicity value from an 

available and acceptable data set. Preference was generally for chronic toxicity data, as long-term 

exposure was a concern. Where these data were not available, the following were used in order of 

preference: acute data, analogue data, quantitative structure–activity relationship (QSAR) data and 

data derived from equilibrium partitioning methods. 

Application factors were derived using a multiplicative approach, which uses 10-fold factors to 

account for various sources of uncertainty associated with making extrapolations and inferences 

related to the following: intra- and interspecies variations; differently sensitive biological endpoints; 

laboratory-to-field impact extrapolation required to extrapolate from single-species tests to 

ecosystems; and potential effects from concurrent presence of other substances. For substances that 

meet the persistence and bioaccumulation criteria as outlined in the CEPA 1999 Regulations (see 

Table 6), an additional application factor of 10 is applied to the CTV. 

Risk quotients derived for PBDEs are summarized in Table 8. Exposure data used as EEVs can be 

found in Tables 4 and 5 or are summarized in the notes to Table 8. Toxicity data used to determine 

CTVs and ENEVs are summarized in Table 7. 

The risk quotient analysis indicates that the greatest potential for risk from PBDEs in the Canadian 

environment is due to the secondary poisoning of wildlife from the consumption of prey containing 

elevated PeBDE and OBDE congener concentrations. Elevated concentrations of components of 

PeBDE in sediments may present risk to benthic organisms. HexaBDE is a component of both 

PeBDE and OBDE and could be a product of heptaBDE, octaBDE, nonaBDE or decaBDE 

transformation. Therefore, risk associated with components of PeBDE may be due to the use of 

OBDE or debromination of highly brominated PBDEs, in addition to the use of PeBDE itself. The 

risk analysis for soil organisms indicates that risk quotients were below 1 for PeBDE, OBDE and 

DBDE; however, the lack of data characterizing PBDE concentrations in soil and sewage sludge 

applied to soil indicates the need for further research. PeBDE, OBDE and DBDE would present low 

potential for risk as a result of direct toxicity to pelagic organisms due to their very low water 

solubility. In the water column, risk associated with components of PeBDE and OBDE (tetra-, pentaand hexaBDE congeners) may be due to bioaccumulation and toxicity to secondary consumers. 
Screening Assessment Report — Environment 

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There is a lack of data characterizing the toxicity of PBDEs to wildlife. Recent studies using rodents 

provide evidence that exposure to PBDEs may lead to behavioural disturbances, disruptions in 

normal thyroid hormone activity and liver effects (e.g., Eriksson et al. 2002, Zhou et al. 2001 and 

2002, Great Lakes Chemical Corporation 1984). The relationship of these studies to potential effects 

from accumulation in the wild is not clear at this time. 

There are a variety of data indicating that all PBDE congeners subject to this assessment are highly 

persistent and each satisfies the requirements for persistence as defined by CEPA 1999 Persistence 

and Bioaccumulation Regulations. 

 

Although uncertainty regarding the possible transformation products of decaBDE exists, there is 

sufficient evidence to conclude that some level of decaBDE phototransformation likely occurs in the 

environment and that lower brominated PBDEs are being formed during this process. These products 

are likely to be more bioaccumulative than the parent compound and could be considered persistent 

and may be directly toxic to organisms. There is limited information available on the relative rates of 

lower BDE formation, and the rates by which these products subsequently degrade in the 

environment. In addition, results from some studies suggest that other as yet unidentified products are 

also being formed as well as PBDFs. It is expected that decaBDE in the environment would mainly 

sequester into sediment or soil and this could limit the amount available for photodegradation, but it 

could make some amount available for transformation via other processes such as anaerobic 

biodegradation or reaction with reducing agents. Overall, it is very difficult to determine the extent to 

which the transformation of decaBDE in the environment may contribute to the potential 

accumulation of lower BDEs and other products. Nevertheless, it is reasonable to consider that 

various transformation processes could contribute to the formation of at least some amount of lower 

brominated PBDEs and PBDFs. Future monitoring would help to clarify whether and the degree to 

which decaBDE transformation contributes to the overall risk presented by the lower brominated DEs 

such as tetra- to hexaBDEs. 

DBDE has become the prevalent commercial PBDE product used in North America and the world. In 

North America and Europe, it is often found in concentrations which exceed those of other PBDEs in 

sewage sludge and sediments. Concentrations of DBDE are now exceeding mg/kg dw levels in North 

American sewage sludge. High accumulation of DBDE in the environment and evidence of 

debromination has led researchers to note that even slight and very long term degradation to lower 

brominated diphenyl ethers over periods spanning several decades could have serious ecological 

consequences. Thus, while current concentrations measured in the environment for homologues 

found in commercial DBDE do not appear to exceed known effect thresholds, their overall 

persistence and potential transformation to bioaccumulative forms, and observed commercial and 

environmental trends, indicate environmental concerns. 

Measured data indicate that tetra-, penta- and hexaBDE are highly bioaccumulative and satisfy the 

criteria for bioaccumulation in the CEPA 1999 regulations. Concentrations of PBDEs in herring gull 

eggs have increased exponentially between 1981 and 2000 at Lake Ontario, Huron and Michigan 

sampling sites. Concentrations of PBDEs (predominantly tetra- and pentaBDE congeners) have also 

increased exponentially between 1981 and 2000 in Arctic male ringed seals. 
Screening Assessment Report — Environment 

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Pyrolysis and extreme heating can cause all PBDEs to form brominated dibenzo-p-dioxins and 

dibenzofurans (European Communities 2001, 2002, 2003). These transformation products are 

considered brominated analogues of the TSMP Track 1 polychlorinated dibenzo-p-dioxins and 

dibenzofurans. 

The PBDEs subject to this assessment have low vapour pressures and low Henry’s Law constants 

(see Table 2) and are not expected to partition significantly into the atmosphere. As such, they are 

considered to present a negligible risk with respect to atmospheric processes such as global warming, 

stratospheric ozone depletion and ground-level ozone formation; however, they do reside in the 

atmosphere adsorbed to suspended particulates and can be transported over long distances. 

Conclusion for the Environment 

It is therefore concluded that tetraBDE, pentaBDE, hexaBDE, heptaBDE, octaBDE, nonaBDE and 

decaBDE, which are found in commercial PeBDE, OBDE and DBDE, 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 meets the criteria under 

Paragraph 64(a) of CEPA 1999. Based on considerations of potential contribution to atmospheric 

processes, it is concluded that PBDEs are not entering the environment in a quantity or concentration 

or under conditions that constitute or may constitute a danger to the environment on which life 

depends, and thus do not meet the criteria under Paragraph 64(b) of CEPA 1999. 

The available data regarding persistence and bioaccumulation of tetraBDE, pentaBDE and hexaBDE 

indicate that they satisfy the criteria outlined in the Persistence and Bioaccumulation Regulations of 

CEPA 1999. Their presence in the environment results primarily from human activity, and they are 

not naturally occurring radionuclides or naturally occurring inorganic substances. 

. 
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Table 7. Summary of toxicity studies used in the derivation of CTVs for the risk quotient analysis of PBDEs 

Species, life stage 

Composition of test material 

Test duration 

Test concentrations Study design Effect level 

Daphnia magna<24 hours old at test initiation 

PeBDE: 33.7% tetraBDE 54.6% pentaBDE 11.7% hexaBDE 

21 days Nominal: 0, 1.9, 3.8, 7.5, 

15 and 30 μg/L 

Measured: 0, 1.4, 2.6, 5.3, 9.8 and 20 μg/L 

• flow-through using well water • 20 ± 1°C, pH 7.9–8.3, DO ≥ 76% 

saturation, hardness 128–136 mg/L as CaCO3, alkalinity 174–176 mg/L as 

CaCO3, conductance 310–315 μmhos/cm 

• 40 animals per treatment • GLP, protocol based on OECD 202, 

TSCA Title 40 and ASTM E1193-87 

• 21-day LOEC (mortality/im

μg/L 

• 21-day NOEC (mortality/i

μg/L 

• 96-hour EC50 (mortality/im

• 7- to 21-day EC50 (mortalit

μg/L 

• 21-day EC50 (reproduction

• 21-day LOEC (growth) = 9• 21-day NOEC (growth ) = • LOEC (overall study) = 9.• NOEC (overall study) = 5.

Lumbriculus variegatus adult

PeBDE: 0.23% triBDE 36.02% tetraBDE 55.10% pentaBDE 8.58% hexaBDE (Great Lakes Chemical Corporation 2000c) 

28 days Nominal: 0, 3.1, 6.3, 13, 

25 and 50 mg/kg dw of sediment 

Analysis of test concentrations at days 0, 7 and 28 indicated they were well maintained throughout the test. Results based on nominal concentrations. 

• flow-through using filtered well water • 23 ± 2°C, pH 7.9–8.6, DO 6.0–8.2 

mg/L, hardness 130 mg/L as CaCO3

• artificial sediment: pH 6.6, water 

holding capacity 11%, mean organic matter <2%, 83% sand, 11% clay, 6% silt 

• 80 animals per treatment • GLP, protocol based on Phipps et al. 

(1993), ASTM E1706-95b and U.S. EPA OPPTS No. 850.1735 

• 28-day LOEC (survival/rep

mg/kg dw of sediment 

• 28-day NOEC (survival/re

mg/kg dw of sediment 

• 28-day EC50 (survival/repr

mg/kg dw of sediment 

• growth (dry weights) not s

different from solvent contconcentration-dependent 

Zea mays corn PeBDE: 

0.23% triBDE 36.02% tetraBDE 55.10% pentaBDE 8.58% hexaBDE (Great Lakes Chemical Corporation 2000c) 

21 days Nominal: 0, 62.5, 125, 

250, 500 and 1000 mg/kg soil dw 

or 0, 50.0, 100, 200, 400 and 800 mg/kg soil ww, assuming 20% soil moisture content Analysis of test concentrations indicated they were well maintained throughout the test. Results reported based on nominal concentrations. 

• artificial soil: 92% sand, 8% clay and 

0% silt, pH 7.5, organic matter content 2.9% 

• watering with well water using 

subirrigation, 14:10 light:dark photoperiod, 16.0–39.9°C, relative 

humidity 19–85% 

• 40 seeds per treatment • GLP, protocol based on U.S. EPA 

OPPTS Nos. 850.4100 and 850.4225 and OECD 208 (based on 1998 proposed revision) 

• no apparent treatment-rela

seedling emergence 

• 21-day LC25, LC50 (seedlin

1000 mg/kg soil dw 

• mean shoot height signific

250, 500 and 1000 mg/kg controls 

• 21-day EC25, EC50 (mean s

mg/kg soil dw 

• mean shoot weight signific

62.5, 125, 250, 500 and 10relative to controls 

• 21-day EC25 (mean shoot w

soil dw 

• 21-day EC50 (mean shoot w

mg/kg soil dw 

• 21-day LOEC (mean shoo

mg/kg soil dw 

• 21-day EC05 and (estimate

shoot weight) = 16.0 mg/k
Screening Assessment Report — Environment 

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Species, life stage 

Composition of test material 

Test duration 

Test concentrations Study design Effect level 

Rat PeBDE (DE-71): 

45–58.1% pentaBDE 24.6–35% tetraBDE (Sjodin 2000; Zhou et al. 2001) 

90 days maximum exposure with recovery periods of 6 and 24 weeks 

In diet: 0, 2, 10 and 100 mg/kg bw per day (doses adjusted weekly based on mean body weight of animals) 

• 30 male and 30 female SpragueDawley CD rats per treatment • decreased food consumptio

increased cholesterol, increporphyrins at 100 mg/kg b

• increased absolute and rela

10 and 100 mg/kg bw, with ranges after 24-week recov

• compound-related microsc

thyroid and liver at all dos

• microscopic thyroid chang

24 weeks 

• microscopic liver changes 

dosage levels after 24-wee

• liver cell degeneration and 

females at all dosage levelrecovery 

• LOAEL (liver cell damage• NOAEL could not be deter

significant effect was obsedose tested 

Daphnia magna<24 hours old at test initiation 

OBDE: 5.5% hexaBDE 42.3% heptaBDE 36.1% octaBDE 13.9% nonaBDE 2.1% decaBDE (European Communities 2003) 

21 days Nominal: 0, 0.13, 0.25, 

0.5, 1.0 and 2.0 μg/L 

Measured: 0, *, *, 0.54, 0.83 and 1.7 μg/L * two lowest concentrations could not be measured 

• flow-through using filtered well water • 20 ± 1°C, pH 8.2–8.5, DO ≥ 77% 

saturation, hardness 132–136 mg/L as CaCO3

• 20 animals per treatment • GLP, protocol based on OECD 202, 

ASTM E1193-87 and TSCA Title 40 

• 21-day LOEC (survival, re

> 2.0 μg/L (nominal) or 1.

• 21-day NOEC (survival, re

≥ 2.0 μg/L (nominal) or 1.

• 21-day EC50 (survival, rep

2.0 μg/L (nominal) or 1.7 μ

Eisenia fetida adult earthworm 

OBDE (DE-79): 78.6% bromine content 

56 days Nominal: 0, 94.0, 188, 

375, 750 and 1500 mg/kg dry soil 

Measured: 0 , 84.9, 166, 361, 698 and 1470 mg/kg dry soil 

• artificial soil: sandy loam, 69% sand, 

18% silt, 13% clay, 8.0% organic matter (4.7% carbon), pH 6.0 ± 0.5 

• 17–21°C, 16:8 light:dark photoperiod, 

pH 5.9–6.8, soil moisture 22.0–33.5% 

• 40 animals per treatment • GLP, protocol based on U.S. EPA 

OPPTS 850.6200, OECD 207 and proposed OECD (2000) guideline 

• 28-day LOEC (mortality) > 

soil 

• 28-day NOEC (mortality) 

soila

• 28-day EC10, EC50 (surviva

dry soil 

• 56-day LOEC (reproductio

dry soil 

• 56-day NOEC (reproductio

dry soila

• 56-day EC10, EC50 (reprod

mg/kg dry soil 

Lumbriculus variegatus adult 

OBDE (DE-79): 78.6% bromine content. 

28 days Nominal: 0, 94, 188, 375, 

750 and 1500 mg/kg dw of sediment 

Measured: (i) 2% OC: <0.354, 76.7, *, *, 755 and 1340 mg/kg dw sediment 

• 80 animals per treatment • flow-through using filtered well water, 

hardness 128–132 mg/L as CaCO3

• Two trials with different artificial 

sediments: (i) 6% silt, 9% clay, 85% sand, 2% TOC, water holding capacity 9.3%, 23 ± 2°C, pH 7.6–8.4, DO ≥

• 28-day LOEC (survival/rep

> 1340 (2% OC) or 1272 (of sediment 

• 28-day NOEC (survival/re

≥ 1340 (2% OC) or 1272 (of sedimenta

• 28-day EC50 (survival/repr
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Species, life stage 

Composition of test material 

Test duration 

Test concentrations Study design Effect level (ii) 5% OC: <12.5, 90.7, *, *, 742 and 1272 mg/kg dw sediment * concentrations were not measured 

45% saturation (3.8 mg/L); (ii) 6% clay, 14% silt, 80% sand, 5% TOC, water holding capacity 13.9%, 23 ±2°C, pH 7.5–8.3, DO ≥ 64% saturation 

(5.4 mg/L) 

• GLP, protocol based on Phipps et al. 

(1993), ASTM E1706-95b and U.S. EPA OPPTS 850.1735 

1340 (2% OC) or 1272 (5% sediment 

For 2% TOC study: • average individual dry wei

statistically lower than in cconsidered treatment-relateaverage biomass in treatmecontrol 

Rabbit OBDE (Saytex 

111): 

0.2% pentaBDE 8.6% hexaBDE 45.0% heptaBDE 33.5% octaBDE 11.2% nonaBDE 1.4% decaBDE (Breslin et al. 1989) 

Days 7–19 of gestation 

By gavage: 0, 2.0, 5.0 and 15 mg/kg bw per day 

• 26 New Zealand White rabbits per 

treatment 

• offspring examined on day 28 of 

gestation 

• no evidence of teratogenic• LOAEL (maternal, increas

decreased body weight gaiper day 

• NOAEL (maternal) = 5.0 m• LOAEL (fetal, delayed oss

sternebrae) = 15 mg/kg bw 

• NOAEL (fetal) = 5.0 mg/k

Eisenia fetida adult earthworm 

DBDE: 97.90% decaBDE 

28 and 56 days 

Nominal soil concentrations: 0, 312, 650, 1260, 2500 and 5000 mg/kg soil dw Mean measured concentrations: <DL, 320, 668, 1240, 2480 and 4910 mg/kg dw 

• artificial sandy loam soil: 69% sand, 

18% silt, 13% clay, 8% TOM, 4.7% TOC, pH adjusted to 6.0 ± 0.5, 60% moisture content, 26% water holding capacity 

• 28-day LOEC (survival) > 

soil (mean measured) 

• 28-day NOEC (survival) ≥ 

soil (mean measured)a

• 28-day EC10, EC50 (surviva

dry soil (mean measured) 

• 56-day LOEC (reproductio

dry soil (mean measured) 

• 56-day NOEC (reproductio

dry soil (mean measured)a 

• 56-day EC10, EC50 (reprod

mg/kg dry soil (mean meas

Lumbriculus variegatus adult 

DBDE: 97.3% decaBDE 2.7% other (not specified) (composite from three manufacturers) 

28 days Nominal: 0, 313, 625, 

1250, 2500 and 5000 mg/kg dw of sediment 

Mean measured: (i) 2.4% OC: <1.16, 291, *, *, 2360 and 4536 mg/kg dw; (ii) 5.9% OC: <DL, 258, *, *, 2034 and 3841 mg/kg dw * 625 and 1250 mg/kg concentrations were not measured 

• 80 animals per treatment • flow-through using filtered well water, 

hardness 128–132 mg/L as CaCO3

• two trials with different artificial 

sediments: (i) 6% silt, 9% clay, 85% sand, 2.4% TOC, water holding capacity 9.3%, 23 ± 2°C, pH 7.7–8.6, 

DO ≥ 36% saturation (3.1 mg/L); (ii) 6% clay, 14% silt, 80% sand, 5.9% TOC, water holding capacity 13.9%, 23 ± 2°C, pH 7.7–8.6, DO ≥ 56% 

saturation (4.8 mg/L) 

• gentle aeration from day 7 to test end • GLP, protocol based on Phipps et al. 

(1993), ASTM E1706-95b and U.S. 

• 28-day NOEC (survival/re

≥ 4536 (2.4% OC) or 3841 dw of sedimenta

• 28-day LOEC (survival/rep

> 4536 (2.4% OC) or 3841 dw of sediment 

• 28-day EC50 (survival/repr

4536 (2.4% OC) or 3841 (dw of sediment 
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Species, life stage 

Composition of test material 

Test duration 

Test concentrations Study design Effect level 

EPA OPPTS 850.1735 

Rat DBDE (Dow-FR300-BA): 

77.4% decaBDE 21.8% nonaBDE 0.8% octaBDE 

30 days In diet: 0, 0.01, 0.1 and 

1.0% (nominal or measured not specified) 

Dosage approximately equivalent to 0, 8, 80 and 800 mg/kg bw per day 

• 5 male Sprague Dawley rats per 

treatment 

• LOAEL (enlarged liver, th

80 mg/kg bw per day 

• NOAEL = 8 mg/kg bw per 

Abbreviations used: ASTM = American Society for Testing and Materials; DL = detection limit; DO = dissolved oxygen; EC50 = median effective dose; EPA = Environm

Good Laboratory Practice; LC50 = median lethal dose; LOAEL = Lowest-Observed-Adverse-Effect Level; LOEC = Lowest-Observed-Effect Concentration; NOAEL = N

Level; NOEC = No-Observed-Effect Concentration; OC = organic carbon; OECD = Organisation for Economic Co-operation and Development; OPPTS = Office of PrevSubstances; TOC = total organic carbon; TOM = total organic matter; TSCA = Toxic Substances Control Act

a Study identified that the highest concentration (or dose) tested did not result in statistically significant results. Since the NOEC or NOAEL could be higher, the NOEC o

greater than or equal to the highest concentration (or dose) tested.
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Table 8. Summary of data used in risk quotient (Q) analysis of PBDEs Commercial Pelagic organisms Benthic organisms Soil organisms product EEVa

(μg/L) 

CTVb

(μg/L) 

AFc ENEV 

(μg/L) 

Q 

(EEV/ 

ENEV)

EEVd

(mg/kg 

dw) 

CTVe

(mg/kg 

dw) 

AFc ENEV 

(mg/kg 

dw) 

Q 

(EEV/ 

ENEV)

EEVf

(mg/kg 

dw) 

CTVg

(mg/kg 

dw) 

AFc ENEV 

(mg/kg 

dw) 

Q 

(EEV/ 

ENEV)

EEVh

(mg/kg 

ww) 

PeBDE 2 × 10-4 5.3 100 0.053 4 × 10-3 1.4 3.1 100 0.031 45.2 0.035–

0.070 

16 100 0.27m 0.13–

0.26 

1.250 

OBDE 2 × 10-4 1.7 100 0.017 0.01 3.03 1340 100 9.1l 0.33 0.03–

0.06 

1470 100 6.3m 0.005–

0.01 

0.325 

DBDE NAk NA NA NA NA 3.19 4536 100 76l 0.04 0.31–

0.62 

4910 100 21m 0.02–

0.03 

0.03 

a Stapleton and Baker (2001). b CMABFRIP (1997d, 1998). c AF (application factors): 10 applied for extrapolation from laboratory to field conditions, intraspecies and interspecies variations in sensitivity; 10 applied because com

are bioaccumulative and persistent. 

d PeBDE: Due to a lack of empirical data characterizing PeBDE sediment concentrations in Canada and due to uncertainty in concentrations throughout North America

surrogate for Canadian data. Concentrations of PeBDE-related components (tetraBDE and pentaBDE) totalled 1.4 mg/kg dw in sediments from Sweden in a heavily in

from a polymer processing site involved with the production of circuit boards (Sellström 1996). This value is used as the EEV. Although climate and local hydrologicatwo countries, polymer processing facilities also exist in Canada. The European Union risk assessment of PeBDE also used this value to assess local risk from a polyu

Communities 2001). 

OBDE: PBDEs found in OBDE are very poorly characterized in North America. Therefore, measured OBDE concentrations from Europe were used as a surrogate for 

OBDE up to 3.03 mg/kg dw have been reported for sediments in the UK downstream of a warehouse facility. This value is used as the EEV (Environment Agency 199

2003). DBDE: There has been insufficient sampling conducted to properly characterize DBDE concentrations in sediments in North America. Concentrations of DBDE in UK 

were determined, with the highest concentration located near a foam manufacturer downstream of a wastewater treatment plant (Law et al. 1996; Allchin et al. 1999). Aenvironment, this value is taken as the EEV.

e Great Lakes Chemical Corporation (2000a, 2001a,b); ACCBFRIP (2001a,b). f Due to the lack of measured data, the EEVs were estimated for tilled agricultural soil and pastureland based on the equation (Bonnell Environmental Consulting 2001)

 EEVsoil = (Csludge × ARsludge × T) / (Dsoil × BDsoil) 

 where:  EEVsoil = EEV for soil (mg/kg); 

 Csludge = concentration in sludge (mg/kg); 

 ARsludge = application rate to soils (kg/m2 per year, default value = 0.5); 

 Dsoil = sludge is mixed in soil to a depth of 0.2 m (depth of tillage) in agricultural soils and 0.1 m in pastureland (European Communities 1994); 

 BDsoil = bulk density of soil (kg/m3, default value = 1700); and  T = number of years sludge is applied to soils (assumed 10 years). 

This equation assumes the following: • no PBDE loss due to erosion; • no PBDE transformation (including transformation of highly brominated PBDEs to tetra- to hexaBDE congeners); 

• no PBDE input from atmospheric deposition; and • no background PBDE accumulation in the soil. In order to calculate the EEVs for PeBDE, a concentration of 2.380 mg/kg dw (total tetraBDE, pentaBDE and hexaBDE) reported in biosolids from a California waste(La Guardia et al. 2001). The EEVs for OBDE were calculated using measured PBDE concentrations (total of hexaBDE, heptaBDE and octaBDE) of 2.08 mg/kg dw i

Guardia et al. (2001). This biosolids sample was taken from a Massachusetts wastewater treatment facility. To calculate the EEVs for DBDE, a PBDE concentration o

and decaBDE) in biosolids was used. This concentration was also reported for a Massachusetts wastewater treatment facility biosolid sample (La Guardia et al. 2001). 

g Great Lakes Chemical Corporation (2000b, 2001c); ACCBFRIP (2001c). 
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h Johnson and Olson (2001); Allchin et al. (1999); Sellström et al. (2001); Lindberg et al. (2004). 

 PeBDE: Johnson and Olson (2001) measured a total PBDE (i.e., BDEs 47, 99, 100, 153 and 154) concentration of 1250 μg/kg ww in mountain whitefish from the Spo

drainage from urbanized areas. No sources other than those typically associated with urbanization (e.g., sewage discharge and urban runoff) are known to exist upstrea

pers. comm. 2003). Although these data are from the United States, such a scenario could exist in Canada, and therefore, the concentration 1250 μg/kg ww in mountai

OBDE: Due to very limited sampling for PBDEs found in OBDE in Canadian biota, the concentration of OBDE of 325 μg/kg ww in dab from the River Tees, UK, wa1999). Although this concentration was determined in liver tissues, it was assumed to equal the concentration of OBDE on a whole body basis. DBDE: There is also a similar lack of data characterizing PBDEs found in DBDE in Canadian biota. DBDE was detected in 18 of 21 analyzed eggs of peregrine falcon

Sweden, at concentrations from 28 to 430 μg/kg lipid weight (lw) (Sellström et al. 2001; Lindberg et al. 2004). The value 430 μg/kg lw (or 0.43 mg/kg lw) will be use

content of these 21 eggs was 5.94% (de Wit 2003), the EEV is converted to 0.03 mg/kg ww. 

i Studies reporting dietary or oral exposure were used for the evaluation of secondary poisoning. The results of these studies are usually expressed as a concentration in 

body weight [bw] per day) causing low or no observed effects. For derivation of a CTVfood and ENEVfood, the results were expressed as a concentration in food (in unit

information on the effect level (CTVtotal daily intake, mg/kg bw per day) in units of daily food intake (DFI, kg ww/day) and body weight (bw, kg ww) for the receptor spec

CTVfood = (CTVtotal daily intake × bw) / DFI. This equation assumes that all substance is exposed via food, and that the substance is completely bioavailable for uptake by t

available data characterizing the toxicity of PBDEs to wildlife species; therefore, data derived using rodents and rabbits were used as surrogates. Interspecies scaling u

was used to extrapolate to determine a food concentration protective of this species. This calculation involved the use of a typical adult body weight (i.e., 0.6 kg) and d

ww/day) of a female American mink (Mustela vison) (CCME 1998). References for toxicity data used in the calculation of the CTVfood include Great Lakes Chemical 

(1989) and Norris et al. (1974). It is noted that Norris et al. (1974) used the product, Dow FR-300-BA, an older DBDE formulation which was composed of 77.4% dec0.8% octaBDE. This product is no longer produced and current formulations of DBDE are composed of a much higher proportion of decaBDE (e.g., usually > 97%). Tin food is nevertheless considered appropriate for use as a CTV for DBDE since the subject study is of acceptable quality and represents a conservative measured endpolder DBDE formulation, its constituents represent homologue groups (predominantly nonaBDEs and decaBDE) subject to this assessment. This assessment is not lim

products, but rather PBDEs in the homologue groups with four to 10 bromine atoms/molecule. Thus, this study is deemed appropriate for use in the quotient analysis, mixture with a greater proportion of nonaBDE (and a small fraction of octaBDE) than current DBDE formulations. 

j To derive the ENEVs, the CTVs were divided by a factor of 10 to account for extrapolation from laboratory to field conditions, a factor of 10 to extrapolate from a rod

further factor of 10 since components of PeBDE and OBDE are bioaccumulative and persistent, and DBDE congeners are persistent and there is a weight of evidence bioaccumulative PBDEs. 

k Not applicable. An ENEV was not derived for pelagic organisms and a risk quotient analysis was not conducted. Based on the available DBDE studies and the toxicity 

it was considered very unlikely that effects for DBDE will be observed in aquatic organisms up to the substance’s water solubility limit. 

l Adjusted to 4% organic carbon. m Adjusted to 2% organic carbon.