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Committee on Problems of the Environment of the International Council of Scientific Unions, the International Oceanographic Commission, the Intergovernmental Working Group on the Global Environmental Monitoring System (the EARTHWATCH program) , the Integrated Global Ocean Station System, and the Experts on Scientific Aspects of Marine Pollution (supported by the United Nations Environment Program). Out of concern like that expressed at the 1974 symposium, there arose the Mussel Watch that uses mollusks (mussels, clams, and oysters) as biologic monitors of aquatic pollution. The 1980 NRC report The International Mussel Watch31 described the program and gave the two general aims of the "watch": to produce information on the contamination of coastal ecosystems and food resources and global data on the abundance of anthropogenic contaminants. 2-11

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TABLE 2-1 Polycyclic Aromatic Hydrocarbons in Coal, Oil-Shale, and Petroleum Isolatesa Relative Peak Valuesb Coal Coal Petroleum PAHSynthoil C Syncrude D Shale B Mix A Fluorene5.3 9.9 2.2 3.9 9-Methylfluorene2 .0 1. 4 0 . 7 2.6 1-Methylfluorene6.0 <4.7 3.2 5.6 Phenanthrene20.4 12.0 4.0 5.3 Anthracene-- 4.1 1.8 T 2-Methylanthracene2.4 3.0 1.1 - 1-Methylphen-<1.2 <5.1 3.6 8.1 anthrene 9-Methylanthracene-- -- T - Fluoranthene<1.9 <3.7 1.6 0.9 Pyrene35.0 14.2 3.7 4.3 Benzota~fluorene2.5 2.1 1.1 2.1 Benzo~bifluorene3.4 <1.5 1.2 2.2 1-Methylpyrene<8.0 c6.0 1.6 1.6 Benzo~ciphenanthrene<0.6 c2.2 1.3 3.1 Benzo~ghi~fluor-3.2 -- -- 0.3 anthene ~ Benz~ajanthracene<2.2 T 1.0 T Chrysene2.5 <1.5 1.5 2.6 Benzo~b, j, and/or<1.3 <0.5 0.5 0.2 k~fluoranthene Benzoteipyrene1.3 <1.2 0.3 0.3 Benzo~aipyrene<1.2 <0.5 0.4 0.2 PeryleneT <0.6 T 0.7 Dibenz~ac and/or ah~anthracene, indeno~l,2,3-cd] pyreneC PiceneC 3-Methylchol-cl.2 <1.3 T 0.70 anthrene o-Phenylenepyrene2~6 1.1 -- - Benzo~ghi~perylene6.6 4.3 -- - Anthanthrenec0.8 T -- - aAdapted from Guerin et al.12 bRelative peak values show a relationship of amounts present. Dash indicates none detected. T indicates trace quantity detected. < indicates known to contain co-eluting species. CPeaks identified, but relative values not sought. 2-12

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TABLE 2-2 PAHs Resulting from Gasification of Western Coal, Coal Blend, Wood, and Peata ~Concentration, 1lg/g of feed stock North WyomingMont ana I 11. #6 I 11. #6 N. C . Dakota Smith-Rosebud & CaCO & Wood Peat Compound Lignit e RolandbCoal BlendC Pel let s d Pe llet se Nap~thalene 1,520 3991,430 878 236 2,480 Acenaphthylene 322 182234 273 53.6 343 Fluorene 137 82.245.5 117 28.9 101 Phenanthrene 443 394375 412 94.0 542 Anthracene 81.4 65.949.0 224 32.7 108 Pyrene 221 253405 230 31.3 395 Benz~a]- 35.7 43.349.0 98.2 6.3 67.0 anthracene Chrysene 18.6 32.130.9 69.0 3.8 33.9 Benzotb]- 22.9 32.154.3 43.9 3.1 35.S fluoranthene Benzo~k]- 7.1 7.91.2 23.0 2.1 17.9 fluoranthene Benzo~aipyrene 17.1 23.752.5 37.6 3.S 34.9 Indenot1,2,3- 4.7 2.24.4 4.2 0.7 0.9 Cd J pyrene Dibenz[ah]- 5~2 11.033.1 16t? .? 11~3 er~ pene |90~0E8ht]- 7.q h'i4,0 t0.4 D.7 7,} ~eryi pn'F' ~Ad4~d frem NiBhole ~ ~.~4 bRubbitg~neq~ epal, 07Q; 0041~ ]~t limestOne. 4797 ~eal' at~ bsrdwood ~asitpLp, .~prth QaFeilns CO88~di peat (19: mai.~ure). 2-13

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TABLE 2-3 Polycyclic Aromatic Hydrocarbons in Used Motor Oil Identified by Gas Chromatography-Mass Spectrometrya Compound Methylbiphenyl Methylbiphenyl Me thy lbiphenyl Fluorene Methylbiphenyl Methylb iphenyl Me thy lbiphenyl Methylf luorene Methylf luorene Me thy 1 f luorene Phenanthrene Deuterated anthraceneC Methyl f luorene D ime thy 1 f 1uorened Anthrac ene Dimethy lf luorened flethylphenan tl:~rene Methyl pl~e nanthrene Me thylphenanthrene Trimethylfluorenee Me thy lphenanthrene Trimethylfluorenee rirnethylf luorenee Phenylnaph thalene Tri~nethylf luorenee Trime thyl f 1uorenee Dimethylphenanthrened D ime t hy lphe nanthrene d D ime thy 1 phenantl~rened Hethylanthracene D ime thy 1 phe nan th r ened D imethylphenanthrened Me thylanthracene D imethy lphenanthrened F luoranthene Methylanthracene Ethylcyclopenta [clef ~ phenanthreneC Ethylcyclopenta [clef ~ phenanthreneC Trime thylphenanthrenee Pyrene l 1 - ug/ml oi 1b 0.74 0.36 0.26 1.47 0.42 0.18 0.09 0.10 1.19 0.08 7.80 0.50 0.08 0.58 0.10 0.33 0.61 2.63 3.62 2.95 0.12 2.44 0.29 0.36 0.90 0.15 0.18 0.22 0.16 0.75 0.09 2.45 4.21 0.28 2.80 4.36 0.21 0.08 0.79 0.46 0.10 0.10 . . . 6.69 l 2-14 i , . .

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Table 2-3 (cont. ~ C ompound Ethylcyclopenta [clef ] phenanthrened Trimethylphenanthrenee Trimethylphenanthrenee Terphenyl Trimethylphenanthrenee D ime thy lanthracened Dimethylanthracened Trimethylphenanthrenee Trimethylphenanthrenee Dihydromethylpyrenef Trimethylphenanthrenee Benzo~a~fluorene Benzotb~fluorene Benzo~c~fluorene Methylpyrenef Trimethylanthracenee Dihydromethylpyrenef Me thy 1pyrenef Me thy lpyrenef Methylpyrenef Diethylphenanthreneg Dime thylpyrene f Diethylphenanthreneg Dimethylpyrened D ie thyl phenan threneg D ime thy lpy rene Dimethylpyrene Benzotciphenanthrene Dimethylpyrened~f Ethylmethylpyrened~f Benzo[~]anthracene Chrysene + triphenylene Cyclopentatcd~pyrene Methylbenzo~aJanthraceneh Methylbenzotmno~fluoranthene MethylbenzotaJanthraceneh Methylbenzo~a]anthraceneh Methylbenzotmno~fluoranthenet MethylbenzotaJanthraceneh MethylbenzotaJanthracene~ MethylbenzotaJanthraceneh Ethylbenzo~ajanthracene Ethylbenzo~aJanthracene Benzo~k~fluoranthene Benzo~eipyrene 2-15 g/ml oilb 0.17 0.39 0.08 0.29 0.12 1.32 0.08 0.10 2.72 0.48 0.13 1.16 0.93 1.38 0.44 1.19 0.51 0.32 1.14 1.14 0.78 0.18 0.13 0.14 0.70 0.34 0.28 0.33 O.12 0.27 0.14 0.87 0.22 2.48 0.78 1.68 0. 15 0.26 0.23 0.15 0.26 0.15 0.28 0.44 0.21 1.44 1.74

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Table 2-3 (cont.) Compound Benzo[a]pyrene Perylene Hethylbenzofluoranthene MethylbenzopyreneJ Benzo[ghi]perylene gtml oilb 0.36 0.13 0.18 0.41 0.32 1.67 aReprinted with permission from White et al.;49 copyright Ann Arbor Science Publishers, Inc. b~g/ml of oil based on the response of deuterated anthracene. CInternal standard. dCould be ethyl- or dimethyl-. eCould be ethylmethyl-, trimethyl-, or propyl fCould be a pyrene or fluoranthene. . Could be diethyl-, ethyldimethyl-, tetramethyl-, methylpropyl-' or butyl Could be a derivative of chrysene, triphenylene, benzotc] phenanthrene or benz~a]anthracene. Could be a derivative of benzotmno~fluoranthene or cyclopentatcd~pyrene. Compounds with molecular weight 276 can be any of the following: indeno~l,2,3-cd~pyrene, indenotl,2,3-cd~fluoranthene, cyclopentatcd]- perylene, phenanthro[10,-1,2,3-cdef~fluorene, acenaphth~l,2-a]- acenaphthylene, dibenzo~b,mno~fluoranthene, dibenzo~e,mno~fluoranthene, and dibenzo~f,mno~fluoranthene. Further possibilities are the benzo derivatives of cyclopentatcdipyrene and cyclopenta~cd~fluoranthene. 2-16 .

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23. McMahon, C. K., and S. N. Tsoukalas. Polynuclear aromatic hydrocarbons in forest fire smoke, pp. 61-73. In P. W. Jones and R. I. Freudenthal, Eds. Polynuclear Aromatic Hydrocarbons. Vol. 3. Carcinogenesis. New York: Raven Press, 1978. 24. Miguel, A. H., and L. M. S. Rubenich. Submicron size distribu- tions of particulate polycyclic aromatic hydrocarbons in combustion source emissions, pp. 1077-1083. In A. Bjorseth and A. J. Dennis, Eds. Polynuclear Aromatic Hydrocarbons: 4th Inter- national Symposium--Chemistry and Biological Effects. Columbus, Ohio: Battelle Press, 1980. 25. Murphy, D. J., R. M. Buchan, and D. G. Fox. Ambient particulate and benzo~a~pyrene concentrations from residential wood combustion in a mountain resort community, pp. 495-538. In J. A. Cooper and D. Malek, Eds. Residential Solid Fuels--Environmental Impacts and Solutions. Proceedings of a 1981 Conference. Beaverton, Ore: Oregon Graduate Center, 1981. 26. National Research Council, Committee on Biologic Effects of Atmos- pheric Pollutants. Particulate Polycyclic Organic Matter. Wash- ington, D.C.: National Academy of Sciences, 1972. 361 pp. 27. National Research Council, Committee on Energy and Materials Recovery from Solid Waste. The Recovery of Energy and Materials from Solid Waste. Washington, D.C.: National Academy Press, 1981. 120 pp. 28. National Research Council, Committee on Fire Research. Air Quality and Smoke from Urban and Forest Fires. Washington, D.C.: National Academy of Sciences, 1976. 381 pp. 29. National Research Council, Committee on Indoor Pollutants. Indoor Pol lutants . Washington, D. C.: National Academy Press , 1981. 537 pp. 30. National Research Council, Committee on Research Needs on the Health Effects of Fossil-fuel Combustion Products. Health Effects of Fo ss i 1- fue 1 Combus t ion Produc t s : Needed Research. Washington, D.C.: National Academy of Sciences, 1980. 73 pp. 31. National Research Council, Environmental Studies Board. The International Mussel Watch. Report of a Workshop. Washington, D.C.: National Research Council, 1980. 248 pp. 32. National Research Council, Ocean Affairs Board. Assessing Potential Ocean Pollutants. Washington, n. c.: National Academy of Sciences, 1975. 438 pp. Neff, J. M. Polycyclic Aromatic Hydrocarbons in the Aquatic Environment. Sources, Fates and Biological Effects. London: Applied Science Publishers, 1979. 34. Nichols, D.` G., S. K. Gangwal, and C. M. Sparacino. Analysis and assessment of PAR from coal combustion and gasification, pp. 397-407. In M. Cook and A. J. Dennis, Eds. Polynuclear Aromatic Hydrocarbons: 5th International Symposium--Chemical Analysis and Biological Fate. Columbus, Ohio: Battelle Press, 1981. 35. Peake, E., and K. Parker. Polynuclear aromatic hydrocarbons and the mutagenicity of used crankcase oils, pp. 1025-1039. In A. Bjorseth and A. J. Dennis ? Eds. Polynuclear Aromatic Hydrocarbons: 4th International Symposium--Chemistry and Biological Ef fects. Columbus, Ohio: Battelle Press, 1979. 33. 2-41

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3 ATMOSPHERIC TRANSFORMATIONS OF POLYCYCLIC AROMATIC HYDROCARBONS GENERAL CONSIDERATIONS: PERSISTENCE AND TRANSFORMATIONS OF PAHs - The atmospheric persistence of PAHs has received considerable attention in recent years and continues to be actively investigated. Two extreme situations can be envisioned. In the absence of any chemical interaction, the lifetime of PAHs adsorbed onto particles will depend solely on physical characteristics--the size of the carrier particle and scavenging processes, including wet and dry deposition. In addition, carrier-particle size is also critical with respect to the rate of deposition in (and clearance from) the human respiratory system and the rate of elusion from the carrier particle by the lung tissue. Because submicrometer particles have atmospheric residence times of several days, experimental evidence on long-range atmospheric transport of PAHs and their distribution in sediments appears to support a hypothesis of negligible chemical transformation of PAHs in the atmosphere. Given the same carrier-particle residence time, even relatively slow chemical reactions could compete effectively with physical processes, with respect to PAH removal from the atmosphere. A substantial body of experimental evidence has been accumulated on chemical reactions between PAHs and pollutant gases under laboratory conditions, with reaction times as short as a few hours. The products of these reactions are in some cases much more potent mutagens than the parent PAHs, thus warranting concern about the implications of these chemical transformations with respect to human exposure. The foregoing considerations suggest the format of this chapter. Pertinent information concerning the chemical and physical processes governing atmospheric persistence of PAHs is summarized first, followed by chemical reactions of PAHs, with an attempt to organize the somewhat conflicting published data according to reactant species and substrate, i.e., carrier particles and other substrates, including filters. PAH FORMATION: CHEMISTRY AND PHYSICS CHEMISTRY The exact synthetic chemistry that produces PAHs in a fuel-rich flame is not well known, but PAHs can be produced from almost any fuel burned under oxygen-deficient conditions.2 As an example of the PAH assemblage produced by combustion systems, Figure 3-1 (top) shows identified gas- chromatographic mass-spectrometry (GC/MS) peaks on PAHs produced by the 3-1

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combustion of kerosene.59 Note that fluoranthene (peak 22) and pyrene (peak 25) are present in about equal abundances; that the abundance of phenanthrene (peak 14) far exceeds that of anthracene (peak 15), a less stable compound; and that benzo[a]pyrene ( peak 39 ) is always found with its noncarcinogenic isomer benzo[e]pyrene (peak 38). A particularly interesting group of compounds in combustion effluents are those with a vinylic bridge' such as acenaphthylene (peak 4) and cyclopenteno[cd]pyrene (peak 32). Peak 23, although not labeled, has been positively identified as acephenanthrylene,54 which also has a vinylic bridge. We emphasize this structural feature because of its chemical reactivity (compared with that of the fully aromatic portions). This reactivity is important in considering the fate of PAHs in the atmosphere. The PAHs shown in Figure 3-1 (top) are typical of those produced from the combustion of various fuels. The combustion of almost any fuel will produce the mixture of compounds shown. The relative abundances, however, can be substantially different, depending on the temperature of combus- tion. In fact, the relative abundances of the alkyl homologues of PAHs depend heavily on the temperature at which the fuel is burned. Although Figure 3-l shows very modest amounts of alkyl homologues (see the region between peaks 25 and 30), other fuels burned under other conditions can show considerably greater abundances of alkyl PAHs. PHYSICS Adsorption PAHs are formed in almost ~11 combustion processes. As the effluent temperature decreases, PAHs initially present largely as vapors become adsorbed on condensing carriers, such as soot and fly ash.3 ,67 It is generally accepted that the adsorption process is virtually completed at or near the point of emission into the ambient air, and that PAHs in ambient air are adsorbed on carrier particles. Studies of the distribu- tion of selected PAHs between the gaseous and particulate phases in ambient airily have shown that, even though the smaller PAHs (e.g., with three and four aromatic rings) may have measurable gas-phase concentrations, these are, as a result of adsorption, lower by several orders of magnitude than those expected on the basis of the corresponding vapor pressures (see Table 3-l). Particle Size Din tribute; an once rams are adsorbed onto carrier particles, their size distribu- tion in the atmosphere is governed by aerosol dynamics, including co- agulation and condensation processes. Thus, carrier particles may evolve into substantially different "stable" size distributions. In many com- bustion processes, PAHs are emitted in the so-called nucleation mode, i.e., adsorbed on particles less than 0.1 Am in diameter. In diesel- engine exhaust, the carrier-particle distribution has mass median diameters of about 0.1-0.25 Am (National Research Council, unpublished 3-2

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manuscript). The contributions from various anthropogenic emission sources may have significant effects on the size distribution of airborne PAHs. In early studies of the PAH size distribution in urban air, DeMaio and Cornl5 reported that most of the benzo~aipyrene (BaP) was found to be associated with small particles (less than 2.3 Am in diameter). Kertesz-Saringer and co-workers47 reported that 50t of the BaP in Budapest air was found in particles smaller than 0.3 ~m. Size distribu- tion measurements were later extended to other PAHs, including benzo~k]- fluoranthene,1 8-PAM and 2-PAM quinones,72 4-azaarenes, and 3-alkyl- substituted PAHs.87 More recently, the application of new size- segregating sampling devices, such as the low-pressure impactor, has given more detailed information on the distribution of PAHs in the submicrometer range. Miguel and Friedlander64 reported on the distribution of BaP and coronene in Pasadena, California, ambient air (Table 3-2~. The largest concentration of both PAHs was found in particles with aerodynamic diameters between 0.075 and 0.12 ~m. The influence of particle size on human respiratory uptake has been the subject of a number of theoretical and experimental studies.5~16~68 In the specific case of PAHs, it has been conclusively shown that the rate of uptake by lung membranes is much higher for PAHs adsorbed on physio- logically inert carrier particles than for the same PAHs inhaled in the crystal state.12 356 In addition, simultaneous measurements of carrier- particle and BaP clearance from the respiratory tract of mice for small (0.5-1.0 ~m) and large (15-30 ~m) carbon particlesl2 showed that, even though smaller particles were cleared from the respiratory tract faster than larger ones, more BaP was eluted from small particles than from the large ones. These results are consistent with findings in tumorigenesis studies of Farrell and Davis, 18 in which the BaP-carbon combination of the smallest particles (0.5-1.0 ~m) was the most carcinogenic, and underline the importance of PAH size distribution for human toxicity. However, most respiratory deposition-clearance studies have been limited to two sizes (i.e., about 1 Am and over 10 ~m), and no information is available on the effect of carrier-particle size on PAH retention in the range of interest, i.e., less than 0.25 ~m. Such size resolution would provide valuable information not only on PAH retention from ambient particles, but also on the relative contribution of various emission sources to PAH uptake. PHYSICAL REMOVAL PROCESSES FOR ATMOSPHERIC PAHs Once PAHs are released from the combustion system and adsorbed on soot or fly ash, they are exposed to potential atmospheric degradation. In the absence of major photodecomposition or other chemical transformations PAHs would be removed from the atmosphere by dry and wet deposition.7 Dry deposition involves sedimentation, turbulence-induced collision with surface electrostatic deposition, and inertial impaction. Although settling velocities have apparently not been determined for PAHs, it is generally accepted that they are controlled by those of the carrier particle. 3-3

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Carrier-particle settling velocities can be estimated from Stokes's law, i.e., assuming that the settling velocity is proportional to the square of the particle diameter, to a term that includes the particle- to-fluid (air) density ratio, and to the reciprocal of the fluid viscosity. Thus, for a 1-pm particle with a density of about 2 g/cm3 in air at 20C, the settling velocity is about 6 x 10-5 m/s,90 in agreement with experimentally determined velocities of about 10 x 10-5 m/s for hum particles.80 For such a particle suspended in air at a height of 20 m and with an average wind speed of 4 m/s (about 9 mph), it would take 4 d to settle to the surface. Assuming a constant wind speed of 4 m/s and constant wind direction over the 4-d period, this atmospheric residence time is equivalent to atmospheric transport over a distance of 1,400 km. Experimental evidence of such regional- and subcontinental- scale transport of PAHs in the atmosphere is discussed below. A simple way in which to note the relative degradation suscepti- bility of the various PAHs is to compare the GC/MS data on PAHs coming from a combustion system (see Figure 3-l, top) with the PAH profile of atmospheric particles (Figure 3-1, middle). PAHs without vinylic bridges are still prevalent, the ratio of fluoranthene to pyrene is still about 1:1, and the ratio of phenanthrene to anthracene is about 10:1. Compounds with vinylic bridges (acenaphthylene, peak 14; acephenanthrylene, peak 23; and cyclopenteno[cd]pyrene, peak 32) have completely vanished from the PAH mixture found in the atmosphere. The increased chemical reactivity of the relatively localized double bond found in these compounds apparently makes them susceptible to photolytic oxidation. Assuming that most PAHs are stable in the atmosphere, what happens to these compounds after they are released from combustion systems throughout the world? Two types of data address this question: data on PAHs in marine and lacustrine sediments, presumably the ultimate environmental sinks of atmospheric PAHs; and data on PAHs in air sampled at remote locations . PAHs IN MARINE AND LACUSTRINE SEDIMENTS Many workers have observed significant concentrations of PAHs in aquatic sediments. For example, Figure 3-1 (bottom) shows a GC/MS analysis of PAHs in the sediment of the Charles River. In comparing the bottom and the middle of Figure 3-1, one sees considerable resemblance. The ratios of the major groups of compounds are the same; the PAHs with vinylic bridges are missing, as they were in the atmosphere; and the alkyl homologues are about as abundant as one might expect. Similar data have been obtained, but in a more quantitative fashion, on over 50 sediment samples from around the world.35 These data indicate that PAHs are ubiquitous and that they are found in almost all samples both near and remote from urban areas. The PAH pattern in all these samples, even the most remote, is similar to that shown in Figure 3-1 (bottom). 3-4

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Even though the relative distribution remains constant, the total amount of PAHs decreases dramatically with distance from urban centers. Figure 3-2 shows a plot of the total PAH abundance in five marine-sediment samples taken from Massachusetts Bay as a function of distance from Boston.90 There is a decrease by 3 orders of magnitude in the total abundance of PAHs within 100 km of Boston. At that point, the total PAH concentration is about 100 ppb; remarkably, that is what is seen in almost all other samples from areas remote from urban centers. On the basis of these and other measurements of PAH concentrations, the following scenario is suggested for the transport of PAHs. The various fuels that are burned in metropolitan areas produce airborne particulate matter (soot and fly ash) on which PAHs are adsorbed. These particles are transported by the prevailing wind for distances that depend heavily on particle diameter. The long-range airborne transport of small particles may account for the presence of PAHs in deep-ocean sediments. Larger airborne particles will settle back onto the urban area; rain then washes them from streets and buildings. The PAHs in this urban runoff eventually accumulate in local sinks. These highly contaminated sediments could be slowly transported by resuspension and currents to seaward locations, where the sediments accumulate in basins or in the deep ocean. The rapid decrease in PAH concentration to 100 ppb within 100 km of Boston (see Figure 3-2) indicates that this transport mode is a rather short-range effect. The stability of PAHs is also apparent when one examines sediment samples taken in such a way as to preserve the historical record. This can be done by carefully coring sediments, segmenting the core into 2 4-cm sections, and analyzing each section for PAHs quantitatively. An example of such data is shown in Figure 3-3; this represents a core from the Pettaquamscutt River in Rhode Island, an anoxic basin.35 The total PAH concentrations range from 14,000 ppb near the sediment surface to less than 120 ppb at the core bottom. Despite the range of concentrations, the relative distribution of the PAHs (excluding the natural products retene and perylene) is indicative of combustion. For example, the ratio of the C16Hlo isomers (nonalkylated) to their monoalkyl homologues (C17Hl2) is 3.0 + 0.4:1. In no case does this ratio become less than unity, which would be expected if the source were direct fossil-fuel contamination. The ratio of the C16Hlo isomers to the C18Hl2 isomers is 2.7 + 0.3:1, and the ratio of the C14Hlo isomers to the C20H12 isomers is 0.46 ~ 0.08:1. These ratios are consistent throughout the core and are indicative of combustion sources.55 Com bustion seems to have been the source of the PAHs in all sections of the core. With a reported deposition rate of 3 mm/yr, total PAHs (excluding retene and perylene) in the Pettaquamscutt core were plotted against year of deposition (see Figure 3-3). For comparison, the BaP data reported by Grimmer and Bohnke27 for a core from the Grosser Planer Sea are also plotted in Figure 3-3. The similarity between these two core profiles is 3-5

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remarkable. Both show rapid increases in PAH concentrations beginning around 1900. The increases could be due to the heavy industrialization that occurred at the turn of the century and the combustion associated with it. A slight decrease in total PAHs around 1930 is present in both cores (see Figure 3-3). It is intriguing to speculate that this reflects an event that occurred in both Europe and New England at this time. The Depression could be such an event. Durint the Depression, total U.S. -energy consumption decreased from 25 x 10 5 BTU (in 1929) to 18 x 1015 BTU (in 1932) before resuming its increasing trend.36 The Pettaquamscutt data are from a core deep enough to allow the assessment of the PAH burden before 1900. The PAH concentrations are low and constant (about 200 ppb) for the 50 yr before the turn of the century. That may be indicative of PAHs from natural combustion processes, such as forest fires. Contributions from natural processes appear to be insignificant in areas or periods of high anthropogenic activity. The decrease in PAHs after 1950 is interesting. It may reflect the change from coal to oil and natural gas as home heating fuels that occurred in the 1950s . During the period 1944-1961, the use of coal in the United States decreased by 40%, and the use of oil and gas increased by 207.36 Combustion of coal usually produces more PAHs than oil and gas, so the change in fuel would result in a decrease in PAH production during the same period. A return to coal as.a major energy source without stringent emission controls might therefore have an important effect on man's input of PAHs into the atmosphere and the sedimentary environment. In an effort to measure the deposition rates for PAHs from the atmosphere in both remote and urban locales, PAH concentrations in sediment cores from water bodies in several areas in the northeastern United States have been determined, and the corresponding atmospheric PAH fluxes to these sites have been calculated. In assessing flux information (rather than concentrations), many of the differences between sites are taken into account, thereby allowing useful comparisons. PAH fluxes calculated for lakes on islands and for remote high-altitude lakes were particularly interesting, in that these sites should reflect most accurately the atmospheric deposition of these combustion-derived pollutants. This background flux could then be compared with PAH inputs found nearer to urban centers, thereby showing the relative importance of long-range airborne vs. short-range runoff delivery of PAHs. With the observed PAH concentrations and information on the sedimen- tation rates and in situ dry densities, the fluxes of individual PAHs to five remote and three urban sites were calculated.32 Table 3-3 (top) shows the results of these flux calculations for core subsections reflecting PAH deposition in remote sites at present, in the interval including 1950, and at the turn of the century (1900). The first point to notice is that the average fluxes for most individual PAHs (except anthra- cene) to remote northeastern U.S. sites are 0.8-3 ng/cm2 at present. 3-6

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Where core subsampling resolution permits, it can be seen that atmospheric PAH fluxes approximately 30 yr ago were 2-3 times greater than and those around 1900 were one-tenth to one-fifth of the present flux rates. This historical PAH record clearly shows that man's activities over the last century resulted in an influx of PAHs to the environment and that coal-derived energy was a much greater source of polluting PAHs than energy derived from oil and gas. For comparison, similar flux estimates for three sites much closer to urban centers were calculated. The results are shown in Table 3-3. These locations all have much greater PAH fluxes than the remote locations. As suggested above, such locations probably receive most of their PAH contamination via water runoff from the watershed. This source of PAHs in sediment overwhelms the background atmospheric deposition rates seen at remote sites. In summary, several things are now apparent about the physical transport of PAHs from source to depot: o PAHs (except retene and perylene) in continental aquatic sediments originate largely in anthropogenic combustion. 0 The watershed runoff resuspension mechanism is of short range (about 100 km) and delivers a near-shore flux of about 35 ng/cm2 per yr for an individual PAH. o The airborne transport mechanism is of long range and delivers a- flux of about 1 ng/cm2 per ye for an individual PAH. O Anthropogenic sources of PAHs were first observed in sediments at concentrations significantly higher than natural background in around 1900, and the maximal depos ition was in about 1950. We need more information about the mechanisms of PAH transport to remote sediments. For example, what fraction of the PAH flux is delivered by aquatic transport mechanisms and what fraction by atmospheric fallout? How much, if any, is lost to the water column? PAHs IN AIR SAMPLED AT REMOTE LOCATIONS Experimental evidence of long-range transport of PAHs has been presented by Lunde, Bjorseth, and co-workers. ,61,62 They analyzed the PAH content of particulate samples collected in Norway with respect to air trajectory. As seen in Table 3-4, PAH concentrations in air masses originating in industrialized areas in western Europe were 20 times higher than those measured in air masses originating in Norway and were as high as those typically measured in urban and industrial areas. These results support the concept, at least for the 20 PAHs listed in Table 3-4 and collected during the winter (i.e., low temperature and low light intensity, resulting in little, if any, photochemical activity), of atmospheric transport of PARS over long distances from anthropogenic sources. 3-7

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CHEMICAL REMOVAL PROCESSES FOR ATMOSPHERIC PAHs Chemical reactions of PAHs in the atmosphere have received steady attention for about 30 yr, as have their implications for human health. In their classic work demonstrating the presence of BaP and other PAHs in the Los Angeles atmosphere and the carcinogenicity of atmospheric organic particulate matter in mice, Kotin et al.52 investigated the interactions between BaP deposited on filters and pollutant gases, including nitrogen dioxide (NO2) and ozone (O3). In this early study of the currently much investigated interactions between PAHs and oxides of nitrogen (NOx) and the health implications of nitro-PAH compounds, Kotin et al. reported a 60t loss of BaP deposited on filter paper when it was exposed to NO2. Later research has focused on BaP and a number of other PAHs; on photo- lysis and photooxidation, as well as on thermal reactions of PAHs with 03, NOX, and sulfur dioxide (SOLD; and on the influence of the physical and chemical nature of the substrate on the reactivity of adsorbed PAHs. The corresponding literature is somewhat conflicting, owing in part to the large number of characteristics that influence these complex and still only partially understood heterogeneous reactions. Thus, it is not surprising to note, even in the recent literature, statements to the effect that PAHs are not chemically reactive and are removed from the atmosphere by rain and sedimentation (e.g., Fishbein203. As discussed below, chemical reactions--including photooxidation, reactions with SO2 and NOX, and reactions with O3 and other oxidants--may, in fact, constitute.major pathways for removal of PAHs from the atmosphere. This discussion focuses on studies of the reactions of PAHs deposited or adsorbed on a variety of substrates (e.g- soot, silica gel, alumina, and glass-fiber filters). The corresponding literature concerning PAR chemistry in the bulk liquid phase (e.g., National Research Council66) is not included here, except for a few studies directly relevant to the chemistry of adsorbed PAHs. REACTION OF PAHs WITH OZONE Kotin et al.52 first reported on the reaction of pure BaP deposited on a filter and exposed to various pollutants and mixtures of pollutants, including 03, NO2, and 0' plus NO2. More recently, Lane and Katz,57 Pitts et al.,74, 5 and Katz et al.45 have reported on the chemical half-lives of PAHs exposed to O3 and on the nature and mutagenic activity of the products. In experiments conducted with BaP, benzotbifluoranthene (BbF), and benzo[k]fluoranthene (BkF) exposed to 0, (at 0.19-2.28 ppm) in air with and without irradiation, Lane and Katz5 reported half-lives of about 40 min for BaP exposed to O3 at 0.19 ppm in the dark. For the three PAHs studied, half-lives decreased with increasing O3 concentration and were further reduced by irradiation with quartzline lamps (Table 3-5~. Sub- stantial differences in reactivity were observed, with BbF and BkF being some 10 times more resistant than BaP to ozonolysis, both in the dark and 3-8

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under irradiation. Katz et al.45 extended this study to nine PAHs deposited on cellulose thin-layer chromotography (TLC) plates and exposed to O3 at 0.2 ppm in the dark, simulated sunlight, and both. The corresponding results (Table 3-5) show significant ozonolysis of some PAHs in the dark, with half-lives ranging from about 0.6 h for BaP and 1.2 h for anthracene to 7.6 h for BeP. Pyrene (half-life, about 16 h), BkF (35 h), and BbF (53 h) were more resistant to dark ozonolysis. For seven of the nine PAHs studied, half-lives were further reduced by irradiation. Pitts et al.74~75 also determined a half-life of about 1 h for BaP deposited on a glass-fiber filter and exposed to O3 at about 0.2 ppm. The results, including the direct comparison between adsorbed and liquid- phase data reported by Katz et al. ,45 clearly demonstrate that the reactivity of PAHs with O3 is much-greater for PAHs deposited on solid substrates than for PAHs in the bulk liquid phase. Products of the reactions between BaP and O3 have been analyzed. Katz _ al.45 identified the 1,6-, 3,6-, and 6,12-diones as major products and noted that all three BaP diones had been identified by Pierce and Katz71 in ambient Toronto air. Pitts et al.74 also reported the epoxide BaP 3,4-oxide as a reaction product. Van Vaeck et al.86 identified a variety of reaction pathways and tentative structures of oxygenated reaction products of the gas-phase ozonolysis of BaP as shown in Figure 3-4. With respect to health implications, Katz et al.45 stated that the BaP diones are direct-acting mutagens, but Pitts et al. 75 found these products inactive in the Salmonella/microsome assay. BaP 4,5-O4ide, a DNA-binding metabolite of BaP, is a strong, direct-acting mutagen. REACTIONS OF PAHs WITH OXIDES OF NITROGEN Kotin et al.52 reported substantial (60%) loss of BaP deposited on a filter and exposed to NO2. The high activity of nitro derivatives of PAHs--many of which are potent, direct-acting muLagens63~85--has prompted renewed interest in the possible formation of these compounds in the atmosphere by reaction of adsorbed PAHs with coemitted NOX. Recent studies discussed here include those of Jager,41 Gundel et al.,33 Pitts et al.,75 Hughes et al.~37 Jager and Hanus,42 Butler and Crossley,1-and Tokiwa et al.8 With the exception of Butler and . _ _ Crossley,~ who used a mixture of nitric oxide (NO) and NO2, all studies have focused on NO2. Hughes et al.37 reported no reaction between NO and PAHs adsorbed on coal fly ash, alumina, and silica gel. A list of the 14 PAHs studied to date is given in Table 3-6. The molecular structures of corresponding nitro-PAH products are shown in Figure 3-5. In all product studies cited above, exposure of adsorbed PAHs to NO2 at parts-per-million concentrations resulted in the formation of nitro-PAH derivatives. These are also listed in Table 3-6 and include mononitro as well as dinitro derivatives, the latter identified as nitration products of BaP42 and pyrene.37 In the study of Jager and Hanus,42 dinitro- BaP was readily produced under conditions relevant to air pollution, i.e., by exposure of BaP adsorbed on fly ash to NO2 at 1.33 ppm for 4 h at 20C. Tokiwa et al.85 have reported extremely high mutagenic activity 3-9

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for dinitro-PAH3 as direct mutagens in the Salmonella/microsome test (e.g., 192 x 10 revertants/nmol for 1,6-dinitropyrene with strain TA 98 without metabolic activation). Mononitro-PAHs, although not as potent mutagens as their diniggo homologues, also exhibit substantial activity as direct mutagens. 3~77' Two nitro-PAHs, 1-nitropyrene and 3-nitrofluoranthene, are carcinogenic in rats.69 The effect of substrate on the nitration of adsorbed PAHs has recently been investigated. Hughes et al.37 compared silica gel, alumina, and coal fly ash and noted that the extent of nitration of pyrene depends on the acidity of the substrate. Jager and Hanus42 discussed the effect of PAH structure, substrate chemical and physical characteristics, NO2 concentration, temperature, and exposure time on the yields of nitro products of pyrene and BaP. For both PAHs, the yields of nitro products were substrate-dependent according to the sequence silica gel > fly ash > alumina >carbon (soot), with silica gel-to-carbon yield ratios as large as a factor of 280 for nitropyrene (Table 3-7). As expected, nitration yields increased as a function of NO2 concentration and exposure time, but not necessarily in a straightforward manner, owing to such complex factors as the adsorption-desorption behavior of NO2 on the substrate. In view of the complex heterogeneous interactions involved, it is not surprising to note large differences in the nitro-PAH yields reported by several investigators. Tokiwa et al.85 prepared nitro derivatives by exposure of pyrene, phenanthrene' fluorene, chrysene ~ and f luoranthene deposited on Toyo #2 paper filters to NO2 for 24 h at 30C in the dark. Large yields were obtained with NO2 at 10 ppm, but yields of only a few percent were obtained at 1 ppm. These low yields are consistent with those reported by Jager and Hanus42 for BaP and pyrene exposure to NO2 at 1.33 ppm for 4 h at 20C with carbon as substrate. In con- trast, Pitts et al.75 reported 40% conversion of BaP deposited on glass-fiber filters and exposed to NO2 at 1 ppm for 8 h at ambient temperature, and a yield of 18% after exposure to NO2 at 0.25 ppm under the same conditions. The higher yields obtained on glass-fiber filters may be due to a greater catalytic effect of the glass-fiber filter than of soot (carbon) substrates. In the same way, reported PAH half-lives due to reaction with NO2 vary considerably with experimental conditions. From the above results, one can derive a half-life of lO h for BaP in the study of Pitts et al.,75 as opposed to half-lives of several days (or weeks) for several PAHs, including BaP, as investigated by Tokiwa et al.84 and Jager and Hanus.42 Butler and Crossley7 recently determined half-lives for 10 PAHs adsorbed on carbon (soot from a burner) and exposed to NO2 at 10 ppm for up to 50 d. Their results, listed in Table 3-6, indicate PAH half-lives ranging from 4-7 d for the more reactive PAHs (anthanthrene, BaP, and benzo~ghi] perylene) to about a month for the least reactive compounds (phenanthrene, fluoranthene, coronene, and chrysene). These half-41ives are consistent with those derived from the work of Jager and Hanus 2 and Tokiwa et al. 84 In view of the substrate used (combustion-generated soot), the results of Butler and Crossley7 and Jager and Hanus42 are probably applicable to heterogeneous nitration of PAHs in the atmosphere. 3-10