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OCR for page 53
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
OCR for page 54
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
OCR for page 55
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
OCR for page 56
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
, . .
OCR for page 57
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
OCR for page 58
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
.
OCR for page 59
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OCR for page 86
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
OCR for page 87
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
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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
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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.
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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 20°C, 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).
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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
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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.
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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.
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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
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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
20°C. Tokiwa et al.85 have reported extremely high mutagenic activity
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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 30°C 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 20°C 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
Representative terms from entire chapter:
polynuclear aromatic
