Ecological Screening Assessment Report on Polybrominated Diphenyl Ethers: chapter 2

Fate, exposure and effects

A summary of selected physical and chemical properties of the commercial Polybrominated Diphenyl Ethers (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 Pentabromodiphenyl Ether (PeBDE) Octabromodiphenyl Ether (OBDE) Decabromodiphenyl Ether (DBDE)
Molecular weight 485.8 (tetrabromodiphenyl ether [tetraBDE])
564.7 (pentaBDE)
(WHO 1994)
643.6 (hexabromodiphenyl ether [hexaBDE])
722.3 (heptabromodiphenyl ether [heptaBDE])
801.4 (octaBDE)
(WHO 1994)
880.4 (nonabromdiphenyl ether [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)
10.9 (tetraBDE)
2.4 (pentaBDE)
(Stenzel and Markley 1997)
(CMABFRIP 1997b)
(CMABFRIP 1997f)
Log Kow 6.57
(MacGregor and Nixon 1997)
(CMABFRIP 1997c)
(Watanabe and Tatsukawa 1990)
(CMABFRIP 1997g)
(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)
(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.

Table 3: predicted partitioning of pentaBDE in the environment based on level III fugacity modeling
Release scenario Predicted partitioning (%)
Predicted partitioning (%)
Predicted partitioning (%)
Predicted partitioning (%)
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 (that is, TaPL3-2.10, ELPOS-1.1.1, Chemrange-2 and Globo-POP-1.1) and various physical and chemical properties (that is, 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 long-range atmospheric transport of PBDEs.

PBDEs have been detected in all environmental media as well as sewage sludge (see Tables 4 and table5), and there is evidence that their levels in the North American environment are increasing.

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, personal communication 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 tetra- to hexaBDEs ranged from approximately 1,350 to 1,900 µg/kg dw. BDEs 47, 99 and 209 were the 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).

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 Wakeford et al. 2002
Murre egg Northern Canada; 1975-1998 0.442-2.93 µg/kg ww Wakeford et al. 2002
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 trout Lake Ontario; 1997 95 µg/kg ww Luross et al. 2002
Lake trout Lake Erie; 1997 27 µg/kg ww Luross et al. 2002
Lake trout Lake Superior; 1997 56 µg/kg ww Luross et al. 2002
Lake trout Lake Huron; 1997 50 µg/kg ww Luross et al. 2002
Rainbow trout Spokane River, Washington, USA; 1999 297 µg/kg ww Johnson and Olson 2001
Mountain whitefish Spokane River, Washington, USA; 1999 1250 µg/kg ww Johnson and Olson 2001
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 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 to 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 (see Table 6). Tetra- to decaBDEs are predicted by AOPWIN (v1.90) to have air degradation half-lives which exceed 2 days (that is, 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 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 based on known empirical and predicted data, as well as structural similarities.

Table 6: persistence and bioaccumulation criteria as defined in CEPA Persistence and Bioaccumulation Regulations (Environment Canada 2000)
 Persistence: half-life Bioaccumulation
Air >= 2 days or is subject to atmospheric transport from its source to a remote area BAF >= 5000
Water >=182 days (>=6 months) BCF >= 5000
Sediment >=365 days (>=12 months) log Kow >= 5
Soil >=182 days (>=6 months)  

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 (for example, 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 half-life 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 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 (that is, 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 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 (for example, 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 (for example, 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 (for example, hepta- to decaBDEs) to lower brominated congeners (for example, 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 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 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 (for example, Alaee and Wenning 2002; de Wit 2002).

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

Table 7: summary of toxicity studies used in the derivation of critical toxicity value (CTVs) for the risk quotient analysis of PBDEs
Species, life stage Composition
of test material
Test duration Test concentrations Study design Effect level Reference
Daphnia magna
<24 hours old at test initiation
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 lowest-observed-effect concentration (LOEC) (mortality/immobility) = 20 µg/L

21-day no-observed-effect concentration (NOEC) (mortality/immobility) = 9.8 µg/L

96-hour EC50 (mortality/immobility) = 17 µg/L

7- to 21-day EC50 (mortality/immobility) = 14 µg/L

21-day EC50 (reproduction) = 14 µg/L

21-day LOEC (growth) = 9.8 µg/L

21-day NOEC (growth )= 5.3 µg/L

LOEC (overall study) = 9.8 µg/L

NOEC (overall study) = 5.3 µg/L

Lumbriculus variegatus
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/reproduction)= 6.3 mg/kg dw of sediment

28-day NOEC (survival/reproduction)= 3.1 mg/kg dw of sediment

28-day EC50 (survival/reproduction)> 50 mg/kg dw of sediment

growth (dry weights) not significantly different from solvent control and not concentration-dependent

Great Lakes Chemical Corporation 2000a
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
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-related effects on seedling emergence

21-day LC25, LC50 (seedling emergence) > 1000 mg/kg soil dw

mean shoot height significantly reduced at 250, 500 and 1000 mg/kg soil dw relative to controls

21-day EC25, EC50 (mean shoot height) > 1000 mg/kg soil dw

mean shoot weight significantly reduced at 62.5, 125, 250, 500 and 1000 mg/kg soil dw relative to controls

21-day EC25 (mean shoot weight) = 154 mg/kg soil dw

21-day EC50 (mean shoot weight) > 1000 mg/kg soil dw

21-day LOEC (mean shoot weight) = 62.5 mg/kg soil dw

21-day EC05 and (estimated) NOEC (mean shoot weight) = 16.0 mg/kg soil dw

Great Lakes Chemical Corporation 2000b
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 Sprague-Dawley CD rats per treatment

decreased food consumption and body weight, increased cholesterol, increased liver and urine porphyrins at 100 mg/kg bw dose

increased absolute and relative liver weights at 10 and 100 mg/kg bw, with return to normal ranges after 24-week recovery period

compound-related microscopic changes to thyroid and liver at all dosage levels

microscopic thyroid changes reversible after 24 weeks

microscopic liver changes still evident at all dosage levels after 24-week recovery period

liver cell degeneration and necrosis evident in females at all dosage levels after 24-week recovery

LOAEL (liver cell damage) = 2 mg/kg bw

No-observed-adverse-effect level (NOAEL) could not be determined, as a significant effect was observed at the lowest dose tested

Great Lakes Chemical Corporation 1984
Daphnia magna
<24 hours old at test initiation
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, reproduction, growth) > 2.0 µg/L (nominal) or 1.7 µg/L (measured)

21-day NOEC (survival, reproduction, growth) >= 2.0 µg/L (nominal) or 1.7 µg/L (measured)a.1

21-day EC50 (survival, reproduction, growth) > 2.0 µg/L (nominal) or 1.7 µg/L (measured)

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) > 1470 mg/kg dry soil

28-day NOEC (mortality) >= 1470 mg/kg dry soila.1

28-day EC10, EC50 (survival) > 1470 mg/kg dry soil

56-day LOEC (reproduction)> 1470 mg/kg dry soil

56-day NOEC (reproduction) >= 1470 mg/kg dry soila.1

56-day EC10, EC50 (reproduction) > 1470 mg/kg dry soil

Great Lakes Chemical Corporation 2001c
Lumbriculus variegatus
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 (ii) 5%OC:
<12.5, 90.7, *, *, 742 and 1272 mg/kg dw sediment

* 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%TOC, water holding capacity 9.3%, 23 ± 2°C, pH 7.6-8.4, DO >= 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

28-day LOEC (survival/reproduction, growth) > 1340 (2% OC) or 1272 (5% OC) mg/kg dw of sediment

28-day NOEC (survival/reproduction, growth) >= 1340 (2% OC) or 1272 (5% OC) mg/kg dw of sediment

28-day EC50 (survival/reproduction, growth) > 1340 (2% OC) or 1272 (5% OC) mg/kg dw of sediment

For 2% TOC study:

average individual dry weights for treatments statistically lower than in control; not considered treatment-related by authors, as average biomass in treatments comparable to control

Great Lakes Chemical Corporation 2001a,b
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 teratogenicity

LOAEL (maternal, increased liver weight, decreased body weight gain) = 15 mg/kg bw per day

NOAEL (maternal) = 5.0 mg/kg bw per day

LOAEL (fetal, delayed ossification of sternebrae) = 15 mg/kg bw per day

NOAEL (fetal) = 5.0 mg/kg bw per day

Breslin et al. 1989
Eisenia fetida
adult earthworm
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)> 4910 mg/kg dry soil (mean measured)

28-day NOEC (survival) >= 4910 mg/kg dry soil (mean measured)a.1

28-day EC10, EC50 (survival) > 4910 mg/kg dry soil (mean measured)

56-day LOEC (reproduction)> 4910 mg/kg dry soil (mean measured)

56-day NOEC (reproduction) >= 4910 mg/kg dry soil (mean measured)a.1

56-day EC10, EC50 (reproduction) > 4910 mg/kg dry soil (mean measured)

Lumbriculus variegatus
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. EPA OPPTS 850.1735

28-day NOEC (survival/reproduction, growth) >= 4536 (2.4% OC) or 3841 (5.9% OC) mg/kg dw of sedimental a.1

28-day LOEC (survival/reproduction, growth) > 4536 (2.4% OC) or 3841 (5.9% OC) mg/kg dw of sediment

28-day EC50 (survival/reproduction, growth) > 4536 (2.4% OC) or 3841 (5.9% OC) mg/kg dw of sediment

ACCBFRIP 2001a,b
Rat DBDE (Dow-FR-300-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, thyroid hyperplasia) = 80 mg/kg bw per day

NOAEL = 8 mg/kg bw per day

Norris et al.1974

Abbreviations used: ASTM = American Society for Testing and Materials; DL = detection limit; DO = dissolved oxygen; EC50 = median effective dose; EPA = Environmental Protection Agency; GLP = Good Laboratory Practice; LC50 = median lethal dose; LOAEL = Lowest-Observed-Adverse-Effect Level; LOEC = Lowest-Observed-Effect Concentration; NOAEL = No-Observed-Adverse-Effect Level; NOEC = No-Observed-Effect Concentration; OC = organic carbon; OECD = Organisation for Economic Co-operation and Development; OPPTS = Office of Prevention, Pesticides and Toxic Substances; TOC = total organic carbon; TOM = total organic matter; TSCA = Toxic Substances Control Act

a.1 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 or NOAEL are described as being greater than or equal to the highest concentration (or dose) tested.

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.

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