Ecological screening assessment report on perfluorooctane sulfonate, salts and precursors: chapter 2

2. Environmental Fate, Exposure and Effects

Environmental Fate of PFOS Precursors

PFOS precursors may be subject to atmospheric transport from their sources to remote areas in Canada. While exact transport mechanisms and pathways are currently unknown, the vapour pressures of PFOS precursors, such as N-EtFOSEA and N-MeFOSEA, may exceed 0.5 Pa (1000 times greater than that of PFOS) (Giesy and Kannan 2002). Several PFOS precursors are considered volatile, including N-EtFOSE alcohol, N-MeFOSE alcohol, N-MeFOSA and N-EtFOSA (US EPA OPPT AR226-0620). All of the above precursors have been predicted by CATABOL modeling and/or expert judgement to degrade to PFOS (Appendix 1). Two PFOS precursors, N-EtFOSE alcohol and N-MeFOSE alcohol, have been measured in air in Toronto and Long Point, Canada (Martin et al. 2002). For precursors released to the water compartment, the vapour pressure may be significant enough to allow the substance to enter into the atmosphere. For N-EtFOSE alcohol, the tendency to leave the water phase is indicated by its relatively high Henry’s law constant (1.9 × 103 Pa•m3.mol-1) (Hekster et al. 2002). It has been reported that when these PFOS precursors are present as residuals in products, they could evaporate into the atmosphere when the products containing them are sprayed and dried (US EPA OPPT AR226-0620). The volatility of certain PFOS precursors may lead to their long-range atmospheric transport (Martin et al. 2002). Although evidence of long-range transport of precursors is limited, it is suggested that this may be partially responsible for the ubiquitous presence of PFOS measured at a distance from significant sources.

It is predicted that the precursors identified in Appendix 1 will undergo degradation once released to the environment though transformation rates may vary widely. Precursors that reach a remote region through the atmosphere or other media may be subject to both abiotic and biotic degradation routes to PFOS (Giesy and Kannan 2002; Hekster et al. 2002). The mechanisms of this degradation are not well understood. When rats metabolize N-MeFOSE-based compounds, several metabolites have been confirmed in tissue samples, including PFOS and N-MeFOSE alcohol (3M Environmental Laboratory 2001a, 2001b). PFOS appears to be the final product of rat and probably other vertebrate metabolism of POSF-based substances. Precursors could be entering food chains by partitioning into biota and then undergoing degradation to PFOS somewhere along the food chain. Most available experimental environmental degradation rates of PFOS precursors are for N-MeFOSE alcohol, N-EtFOSE alcohol, N-MeFOSEA and N-EtFOSEA and are summarized in Table 2.

Table 2: Summary of Available Data on Transformation of PFOS and its Precursors
Substance Biodegradation Biotransformation Photolysis Hydrolysis
PFOS (K+) 0% N/AFootnote b 0% t½ > 41 years
N-MeFOSE alcohol N/A N/A N/A t½ = 6.3 years
N-EtFOSE alcohol To PFOS/PFOAFootnote a N/A 0%
t1/2= estimated 40 days at 25°C (indirect photolysis)Footnote c
t½ = 7.3 years 92% after 24 hours to PFOS (alkaline)
N-MeFOSEA N/A N/A N/A t½ = 99 days at pH 7, 25°C (extrapolated)
N-EtFOSEA N/A N/A N/A t½ = 35 days at pH 7, 25°C

Source: Hekster et al. (2002);

Some studies on photolysis show that this transformation mechanism will be of no importance in the breakdown of certain perfluorinated chemicals. Certain tests with PFOS, perfluorooctanoic acid (PFOA), POSF and N-EtFOSE alcohol show no photodegradation at all (Hekster et al. 2002; US EPA OPPT AR226-0184, AR226-1030a041). Aqueous photolytic screening studies carried out with N-EtFOSE alcohol, N-MeFOSE alcohol, N-EtFOSA and N-MeFOSA as well as on a surfactant and foamer product showed no direct photolysis, although some underwent indirect photolysis. The primary products were PFOA, perfluorooctane sulfonic acid (PFOSA) and N-EtFOSA (US EPA OPPT AR226-1030a073, AR226-1030a074, AR226-1030a080, and AR226-1030a106). A photolysis study on N-EtFOSE alcohol found that the primary products of indirect photolyis of this substance included PFOA, N-ethylperfluorooctane sulfonamide and perfluorooctane sulfonamide, with trace levels of additional substances including PFOS (US EPA OPPT AR226-1030a080). The study estimated an indirect photolysis half-life for N-EtFOSE alcohol of 40 days at 25°C, but noted environmental factors could lead to variation.


PFOS is resistant to hydrolysis, photolysis, aerobic and anaerobic biodegradation and metabolism by vertebrates. The perfluorinated moiety is known to be very resistant to degradation, a property attributed to the C-F bond, one of the strongest chemical bonds in nature (~110 kcal.mol-1) (US EPA OPPT AR226-0547). The perfluorinated chain provides exceptional resistance to thermal and chemical attack (US EPA OPPT AR 226-0547). Several biodegradation studies were reviewed by the OECD which indicated no biodegradation had taken place (OECD 2002a).

The estimated half-life for PFOS is reported as >41 years (Hekster et al. 2002), but may be significantly longer than 41 years. The persistent nature of PFOS is indicated in numerous studies (Key et al. 1997; Giesy and Kannan 2002; Hekster et al. 2002; OECD 2002a). In water, PFOS was observed to persist for more than 285 days in microcosms under natural conditions (Boudreau et al. 2003b). POSF, a precursor and analogue to PFOS, is resistant to atmospheric hydroxyl radical attack and is considered persistent in air, with an atmospheric half-life of 3.7 years (US EPA OPPT AR226-1030a104). PFOS and some of its precursors are considered to be persistent in the Canadian environment with the environmental half-life for PFOS exceeding the half-life criteria for persistence as defined by the Persistence and Bioaccumulation Regulations of CEPA 1999 (Government of Canada 2000).

Once PFOS is in the environment, it may enter the food chain or be further distributed at a distance from its source. PFOS has been detected in wildlife at remote sites far from known sources or manufacturing facilities (Martin et al. 2004a). This suggests that either PFOS or PFOS precursors may undergo long-range transport.

Predicting the environmental fate of PFOS can be difficult, given its physical and chemical characteristics. Due to the surface-active properties of PFOS, a meaningful log Kow value cannot be determined (OECD 2002a). Unlike the situation with most other hydrocarbons, hydrophobic and hydrophilic interactions are not the primary partitioning mechanisms, but electrostatic interactions may be more important. It has been suggested that PFOS adsorbs via chemisorption (Hekster et al. 2002). A soil adsorption/desorption study using various soil, sediment and sludge matrices found that PFOS adsorbed to all matrices tested (US EPA OPPT AR226-1107). River sediments displayed the most desorption, at 39% after 48 hours, whereas sludge samples did not desorb detectable amounts of PFOS. If PFOS does bind to particulate matter in the water column, then it may settle and reside in sediment. However, as noted, desorption may also occur.

While the vapour pressure of PFOS is similar to those of other globally distributed compounds (e.g., polychlorinated biphenyls [PCBs], dichlorodiphenyltrichloroethane [DDT]), its greater water solubility indicates that PFOS is less likely to partition to and be transported in air (Giesy and Kannan 2002). PFOS potassium salt has a water solubility value of 519 to 680 mg.L-1. This has been found to decrease significantly with increasing salt content (12.4 mg.L-1 in natural seawater at 22-23°C, and 20.0 mg.L-1 in a 3.5% NaCl solution at 22-24°C) (US EPA OPPT AR226-0620; Hekster et al. 2002; OECD 2002a). The OECD review of PFOS data suggested that any PFOS released to a water body would tend to remain in that medium, unless otherwise adsorbed onto particulate matter or taken up by organisms (OECD 2002a).


The use of log Kow and physical-chemical properties to predict the potential for bioaccumulation, in general, is based on the assumption that the hydrophobic and lipophilic interactions between compound and substrate are the main mechanisms governing partitioning. This assumption has been shown to hold for non-polar and slightly polar organic chemicals. However, this assumption may not be applicable for perfluorinated substances. Due to the perfluorination as described by Key et al. (1997), the hydrocarbon chains are oleophilic and hydrophobic and the perfluorinated chains are both oleophobic and hydrophobic. In addition, functional groups attached to the perfluorinated chain (e.g. a charged moiety such as sulfonic acid) can impart hydrophilicity to part of the molecule. Hydrophobicity is unlikely to be the sole driving force for the partitioning of perfluorinated substances to tissues because the oleophobic repellency opposes this partitioning process (Kannan et al., 2001). Perfluorinated substances are also intrinsically polar chemicals because fluorine, a highly electronegative element, imparts polarity. Thus, perfluorinated substances have combined properties of oleophobicity, hydrophobicity, and hydrophilicity over portions of a particular molecule. Based on the current scientific understanding, lipid normalizing of concentrations in organisms for perfluorinated substances may not be appropriate since these substances appear to preferentially bind to proteins in liver and blood rather than accumulating in lipids.

Measures of bioaccumulation (bioconcentration factors (BCFs), bioaccumulation factors (BAFs), and biomagnification factors (BMFs)) may be used as indicators of either direct toxicity to organisms that have accumulated PFOS or indirect toxicity to organisms that consume prey containing PFOS (via food chain transfer). Concerning the potential to cause direct toxicity, the critical body burden is the minimum concentration of a substance in an organism that causes an adverse effect. From a physiological perspective, it is the concentration of a substance at the site of toxic action within the organism that determines whether a response is observed, regardless of the external concentration. In the case of PFOS, the site of toxic action is often considered to be the liver.

Concerning the potential for toxicity to consumer organisms , it is the concentration in the whole body of a prey that is of interest since the prey is often completely consumed by the predator - including individual tissues and organs, such as the liver and blood. However, given the partitioning into liver and blood, most field measurements for perfluorinated substances have been performed for those individual organs and tissues especially for higher trophic level organisms (e.g. polar bear) where whole body analysis is not feasible due to either sampling or laboratory processing constraints. While it is feasible to measure whole body BAFsFootnote 3 on smaller, lower trophic level species, the lower trophic status of the organism would mean that, for perfluorinated substances, the estimated overall BAFs may be underestimated due to their trophic status.

Thus, from a toxicological perspective, BCFs, BAFs and BMFs based on concentrations in individual organs, such as the liver, may be more relevant when predicting potential for direct organ-specific toxicity (i.e., liver toxicity). However, BCFs and particularly BMFs based on concentrations in whole organisms may provide a useful measure of overall potential for food chain transfer. The ranges for whole body and tissue- and organ-specific BCFs/BAFs/BMFs are summarized as follows:

Table 3: Range of BCF/ BAF/ BMF Data for PFOS in Whole Body, Specific Tissues and Organs in Wildlife
  Whole Body Tissue Specific (blood or liver)
BCF 690 - 2 796 2 900 - 5 400
BAF None available 274 - 125 000
BMF 0.4 - 5.88 4.0 - 20

Estimated PFOS BCFs (assumed steady-state conditions) of 1100 (carcass), 5400 (liver) and 4300 (blood) have been reported for juvenile rainbow trout; the 12-day accumulation ratio was 690 (carcass), 3100 (blood), and 2900 (liver) (Martin et al., 2003a). A laboratory study with bluegill sunfish gave a whole body BCF of 2796 (US EPA OPPT AR226-1030a042). In addition to information on PFOS, the US Interagency Testing Committee estimated BCFs for N-EtFOSEA and N-MeFOSEA using structure-activity models to be 5543 and 26 000, respectively (Giesy and Kannan 2002). In the study by Kannan et al., (2005a), a BCF of 1000 (whole-body) was calculated in benthic invertebrates. Species differences for the elimination half-life of PFOS in biota have been determined to vary significantly: 15 days (fish); 100 days (rats), 200 days (monkeys) and years (humans) (OECD 2002a; Martin et al., 2003b).

In fish livers collected from 23 different species in Japan, PFOS BAFsFootnote 4 were calculated to range from 274 - 41 600 (mean 5500) (Taniyasu et al. 2003). Following an accidental release of fire fighting foam into Etobicoke Creek, Moody et al. (2002) calculated a BAF range of 6300-125 000 for PFOS, based on measured concentrations in common shiner liver and surface water (Moody et al., 2002).

Available data indicate certain fish species at specific life-stages (e.g. juvenile rainbow trout) with dietary exposure to PFAs would have BMFs lower than one and that biomagnification would not occur. Martin et al. (2003b) showed that PFOS did not biomagnify from food in juvenile troutFootnote 5. However, the authors suggest caution in extrapolating these results to larger fish, e.g. mature trout, as the half lives of other substances have been shown to increase by up to a factor of 10 times in mature fish compared to juveniles. A possible driving factor for this phenomenon is that the gill-surface-area to volume ratio and relative gill ventilation rate may decrease as fish mature, i.e., elimination via the gills may become less efficient and less significant. Another contributing factor could be growth dilution which is much more significant in relatively fast growing juveniles. Nevertheless, elimination through the gills is an important route for fish which is not available to birds, terrestrial mammals (e.g., mink, polar bear, Arctic foxes) and marine mammals (e.g., seals and whales). Furthermore, elimination from the lungs is expected to be much lower in view of the low vapour pressure and negative charge. High concentrations of PFOS have been found in the liver and blood of higher trophic level predators that consume fish (e.g., polar bears, mink and birds).

Moody et al., (2002) suggested that the BAFs in their study may be overestimated due to the metabolism of accumulated precursors to PFOS. The biotransformation of PFOS precursors (e.g. PFOS precursors in fire-fighting foam) is currently not well studied. However, it is possible that the transformation of precursors to PFOS within the organism could cause the total body burden of PFOS to exceed that which would be achieved by accumulation from water and diet alone. Since the water concentration used for the BAF calculation does not account for the proportion of precursors which could be transformed to PFOS within the organism, the calculated BAF may be artificially high. In the absence of established methods to account for precursor bioaccumulation and transformation, it can be argued that the results of the Moody et al. (2002) study provide a relevant expression of bioaccumulative potential and conservative BAF estimates especially given that the metabolic transformation of precursors to PFOS is an additional cause for concern.

There are three studies suggesting that PFOS biomagnifies in the Great Lakes and Arctic food webs. In the study by Kannan et al., (2005a), for the water-algae-zebra mussel (whole body)-round goby (whole body) -smallmouth bass (muscle tissue) -bald eagle (liver, muscle, or kidney tissue) food chain, a BMF of 10 to 20 was calculated in mink (liver) or bald eagles. It should be noted that comparison of PFOS concentrations between certain species was not always direct (i.e., whole body to whole body). Eggs of fish contained notable concentrations of PFOS, suggesting oviparous transfer of PFOS. In the study by Martin et al., (2004b), a benthic invertebrate/pelagic invertebrate-three forage fish (whole body analyses of alewife, slimy sculpin, rainbow smelt)-top predator fish (lake trout) food chain resulted in a multi-trophic level BMF of 5.88. Martin et al. (2004b) noted that the benthic invertebrate and its predator fish (sculpin) had higher concentrations of PFOS than the lake trout. Martin et al. (2004b) also suggested that bioaccumulation was occurring at the top of the food web for PFOS and all perfluoroalkyl substances (except for PFOA). Tomy et al., (2004) suggested that PFOS biomagnifies through the Arctic marine food web. Again, it is noted that comparison of PFOS concentrations between certain species was not always direct (i.e., whole body to whole body). The trophic level BMF for PFOS included walrus (liver) -clam (whole body) (4.6); narwhal (liver) -cod (whole body) (7.2); beluga (liver)-cod (whole body) (8.4); beluga (liver) -redfish (liver) (4.0); black-legged kittiwake (liver) -cod (whole body) (5.1); glaucous gull (liver) - cod (whole body) (9.0); and cod (whole body)-zooplankton (whole body) (0.4). Smithwick et al. 2005a) stated that polar bears as apex predators had high PFOS concentrations in their liver tissue suggesting food-chain accumulation. Concentrations of PFOS found in samples for East Greenland (mean = 2,470 ng.g-1 ww) were similar to Hudson Bay (mean = 2730 ng.g-1 as reported by Smithwick et al., (2005a); 3,100 ng.g-1 as reported by Martin et al., (2004a) and both populations had significantly greater concentrations than those reported for Alaska (350 ng.g -1 from Giesy et al., 2002).

The possibility of PFOS bioaccumulation in migratory birds is also a concern because migratory species, such as loons, ospreys, and cormorants could be exposed to higher concentrations of PFOS while wintering in the US before migrating to Canada where they could experience reproductive and other effects during the breeding season. It is assumed that the main route of exposure to PFOS for birds is through the diet. The dietary exposure route is particularly relevant because biomagnification of PFOS in bird tissues can occur this way. BMFs above one are reported for several bird species (eider duck, red-throated loon, razorbill, long-tailed duck) collected in the Gulf of Gdansk (Gulkowska et. al. 2005). In the water-algae-zebra mussel-round goby-smallmouth bass-bald eagle food chain Kannan et al., (2005a) suggested a PFOS BMF of 10 to 20 in bald eagles (relative to prey items). Tomy et al., (2004) suggested that PFOS biomagnifies through the Arctic marine food web: the trophic level BMF for PFOS included black-legged kittiwake-cod (5.1) and glaucous gull-cod (9.0). There is information indicating that PFOS has relatively shorter half-lives in blood and liver tissue in birds compared to mammals (Newsted et al., 2005). For example, the estimated elimination half-life for PFOS from serum is 13.6 days in male mallards whereas in male rats it is greater than 90 days. In addition, a recent study suggests that PFOS is excreted relatively rapidly from birds (Kannan et al., 2005). However, if birds are chronically exposed to PFOS in their diet, then biomagnification can still occur because, as pointed out in the Kannan study, the binding of perfluorinated compounds to proteins and retention by enterohepatic circulation are the major factors that determine accumulation and retention in biota.

Whole body aquatic BCFs are below 5000. However, the weight of evidence from laboratory and field-based whole body and tissue-specific BCFs and BAFs in conjunction with the field-based BMFs (avian and aquatic) indicates that PFOS is a bioaccumulative substance.

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