Objectives 
A major uncertainty in assessing the risks of PBTs to aquatic life is whether application of 
laboratory toxicological data adequately reflects complex exposure relationships and interactions 
important to responses in natural systems. Chemicals in one class of PBTs, PAHs, have been 
found in traditional laboratory tests to have relatively low toxicities, due both to a nonspecific 
mechanism of action and to reduced bioaccumulation in some organisms because of metabolic 
transformations of these chemicals. However, the toxicities of some PAHs to various aquatic 
organisms have been demonstrated to be greatly increased (by orders of magnitude) due to 
exposure to UV radiation (Bowling et al. 1983; Cody et al. 1984; Kagan et al. 1984, 1985; Oris 
and Giesy 1985, 1987; Newsted and Giesy 1987; Holst and Giesy 1989; Tilghman Hall and Oris 
1991; Huang et al. 1993; Buckler et al. 1994; Ankley et al. 1994, 1995, 1997; Boese et al. 1997; 
Erickson et al. 1999). General principles of dosimetry for this enhanced toxicity, based on PAH 
accumulation and UV intensity, have been described (Newsted and Giesy 1987; Ankley et al. 
1995, 1997; Erickson et al. 1999). 
An analysis of fuel contamination of the clear waters of Lake Tahoe concluded that 
photoactivated toxicity posed a significant risk to zooplankton (Oris et al. 1998) and current data 
suggest that ELS fish in PAH-contaminated littoral zones of the Great Lakes are at risk (Mount et 
al. 2001). However, this risk is uncertain due to several factors, some specifically related to 
photo-activated toxicity and some of more concern to PBTs in general: 1) most research to date 
has used laboratory UV light sources with spectra different from natural sunlight; 2) both the 
intensity and spectra of UV light in natural systems vary spatially and tempOTally, resulting in 
receptor organisms receiving widely varying exposures depending on their life habits and the 
properties of the system; 3) PAH exposure can also vary widely within natural systems, 
especially between sediments and overlying waters, so that PAH accumulation can also depend 
on organism attributes and system properties; 4) the accumulation, and thus the effects, of PAHs 
can vary between laboratory and natural systems due to food chain influences and maternal 
transfer to young organisms; 5) the relative accumulation of and sensitivity to PAHs of different 
life stages are poorly known; and 6) research has usually used individual compounds or simple 
mixtures of commercially-obtained PAHs, in contrast to complex mixtures of PAHs occurring in 
contaminated natural systems. However, current knowledge of a) the general levels of PAH 
contamination and the magnitude of UV light in natural systems and b) the sensitivity of many 
organisms to photo-activated toxicity in the laboratory indicate a potential for major impacts due 
to these interacting factors. 
Risks from PAHs will depend on a complex interaction among light, chemical, receptor 
organisms, and system characteristics. ELS fish are one group of organisms at potential risk. 
ELS fish potentially can have significant PAH accumulation due to maternal transfer to eggs, 
exposure of eggs to water and sediment, accumulation after hatching from food and water 
(especially when closely linked to sediments), and the absence of metabolic pathways which 
limit PAH accumulation in older fish. Provided that the fish behavior or the system attributes 
result in significant exposure to light, ELS fish might be particularly susceptible to UV-activated 
PAH toxicity because of their small size (i.e., large surface to volume ratios and short penetration 
distances) and lack of protective pigmentation and gill coverings. Past research has shown early 
life stage fish to be susceptible to photo-activated toxicity (Bowling et al. 1983; Oris and Giesy 
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