Info

"hi

1 2 3 4 5 6 7 Sampling event

1 2 3 4 5 6 7 Sampling event

FIGURE 10.16 Particle size vs activatable mutagenicity (mutagen density, rev m~3; TA98 +S9) for ambient particulate POM collected in seven week-long sampling events from November 15, 1994, to March 31, 1995. The site was "near a very busy road in Bologna, Italy" (adapted from Pagano et al, 1996).

source in that area (determined by the specific activities (rev /tig-1) and concentrations of the individual biologically active compounds associated with the particles), (b) the emission strengths of the particles (emission factors) from each source class, (c) the mutagenic potencies and concentrations of the secondary mutagenic PACs formed in atmospheric reactions of non-mutagenic PAHs (e.g., gas-phase reactions of naphthalene, phenanthrene, pyrene, and fluoranthene; see Section F), (d) the decay of reactive mutagenic/car-cinogenic PAHs during transport (e.g., of BaP and cyclopenta[c<i]pyrene; see Section E), and (e) the condensation of semivolatile mutagenic compounds on particles after their emission as gases.

Outdoor mutagenicity levels (rev m~3) are, of course, also influenced by seasonal, spatial, meteorological, and air pollution variables and may also reflect the sampling methodology. For example, in the mid-1980s, Alfheim and co-workers reported possible artifactual formation or degradation of mutagenic species when comparing ambient aerosols sampled with different types of Hi-Vol filters and XAD-2 resin (Alfheim and Lindskog, 1984a; Alfheim et al., 1985). In similar tests of sampling with glass fiber filters (GFF) and Teflon-coated fiberglass filters, Daisey and co-workers (1986a) concluded "the filter medium can significantly influence the extractable mass, chemical composition, and mutagenic activity of the organic fractions of the ambient aerosol."

However, in a study involving four kinds of Hi-Vol filters (glass, quartz, Teflon, and Teflon-impregnated glass fiber filters), Fitz and co-workers (1984) reported no large differences due to filter artifacts. Subsequently, de Raat and colleagues (1990) found that differences in sampling for particle-phase mutagenicity using glass fiber filters vs Teflon or Teflon-coated filters were too small to support one type of filter over the other.

Sampling for different periods of time, e.g., 3 h, 24 h, monthly, seasonal, and annual averages, conveys different types of information, chemically and toxicologi-cally. For example, mutagenic and chemical product evidence of the relative efficiencies of OH radical attack on PAH during daylight hours vs N03 radical attack at night can be deduced from 4- to f 2-h averages of direct mutagenicity (e.g., see Arey et al., 1989a) but would be masked in 24-h, monthly, seasonal, or yearly averages (see, e.g., discussion by Masclet et al. (1986) on the impacts of atmospheric reactions, sampling parameters, and air pollution/meteorological variables).

The location of a sampling site relative to the sources, meteorology, topography, etc. characteristic of that area is also clearly important. Thus, in some cases the sampling site is close to a strong source(s) of emissions that dominates the local air quality. Thus, in urban areas worldwide, with very different climates, meteorology, and topography, emissions of PAHs and PACs from diesel- and gasoline-powered light- and heavy-duty motor vehicles (e.g., diesel trucks and non-catalyst-equipped automobiles) have been, and continue to be, major contributors to the particle-phase promutagenicity of respirable ambient particulate POM (e.g., see Benner et al., 1989, and the review by van Houdt, 1990). This has been shown to be true, for example, in urban air in Oslo, Norway, and Stockholm, Sweden (Alfheim et al., 1983), Copenhagen, Denmark (Nielsen, 1996; Nielsen et al., 1996, 1999a, 1999b), the San Francisco Bay Area (Flessel et al., 1985; Kado et al., 1986; Flessel et al., 1991), southern California cities (Pitts et al., f982b, 1985a; Atkinson et al., 1988a; Venkataraman and Friedlander, f994b, 1994c; Schauer et al., f 996; Hannigan et al., 1994, f 996, f 997), Helsinki and Lahati, Finland (Tuominen et al., 1988), Athens, Greece (Viras et al., 1990), Padova, Italy (Nardine and Clonfero, 1992), The Netherlands (De Raat et al., 1991), and Sapporo, Japan (Matsumoto et al., 1998).

The impact of close proximity to a major source on the mutagenicity levels of ambient air particles (rev nT3) is seen in Fig. f0.17 for a site immediately adjacent to, and generally downwind from, a heavily traveled freeway in west Los Angeles (Pitts et al., f985a). Peak levels were -120 ( + S9) and 100 (-S9) rev nT3 during the midmorning rush hour and ~ 110 (±S9)

Pacific Standard Time (PST)

FIGURE 10.17 Diurnal variations in mutagen densities (rev m~\ TA98 + S9, TA98 - S9, and TA98NR) of 3-h averaged samples of ambient particles collected on TIGF filters at a site downwind from, and adjacent to, a heavily traveled freeway in west Los Angeles (March 9-10, 1983). Midmorning and early evening peaks reflect the two rush hour traffic periods and demonstrate the utility of short-time, 3-h vs 24-h average sampling periods (adapted from Pitts et al., 1985a).

Pacific Standard Time (PST)

FIGURE 10.17 Diurnal variations in mutagen densities (rev m~\ TA98 + S9, TA98 - S9, and TA98NR) of 3-h averaged samples of ambient particles collected on TIGF filters at a site downwind from, and adjacent to, a heavily traveled freeway in west Los Angeles (March 9-10, 1983). Midmorning and early evening peaks reflect the two rush hour traffic periods and demonstrate the utility of short-time, 3-h vs 24-h average sampling periods (adapted from Pitts et al., 1985a).

during the early evening traffic "jam," compared to ~60 rev m~3 in the early afternoon and ~40 rev m~3 in the early morning hours. The results for TA98NR were significantly lower, indicating the presence of ni-troarenes.

Alternatively, some sampling sites are at locations that are downwind from polluted urban centers, so that under common meteorological/weather conditions, one is sampling "aged" air parcels that have been subjected to the phenomena of transport and atmospheric reactions (e.g., Azusa and Riverside, California, east and usually downwind from central Los Angeles; see map, Fig. 10.23). Such transport can occur over long distances. For example, although the mutagenicity of atmospheric aerosols sampled at a "background site" in Birkenes, Norway, rarely exceeded 1 rev m~3, Alfheim and M0eller (1979) showed that a substantial fraction of it arose from long-range transported air pollution.

T. Nielsen and co-workers investigated, at a site in downtown Copenhagen, the mutagenicity of polluted air masses transported from the European continent compared to that originating from local sources. The direct mutagenicity was higher by a factor of 5-7 compared to "average" local levels; arrival of the air mass undergoing long-range transport was also associated with an increase in the local S02 level by a factor of -2.5 (Nielsen et al., 1999a, 1999b).

An interesting aspect of the long-range transport of mutagenic compounds is a field study by De Pollok and co-workers (1997), who collected and measured ambient air particles at three different levels of a TV tower (surface, <1 m; mid, 240 m, and top 433 m) near

Raleigh, North Carolina, for three periods prior to, during, and after Hurricane Gordon in November 1994. The surface samples were not mutagenic on strains YG1021 and YG1026 (vide supra), but the top- and mid-level samples showed significant mutagenicity for the "posthurricane, normal weather samples." The authors suggest this arises from long-range transport of mutagenic nitroarenes.

While mobile sources are generally very important, seasonal patterns of emissions from various stationary local sources combined with the particular meteorology (e.g., "tight" low-radiation inversions in the winter) can have a significant impact on the nature and extent of mutagenicity. For example, residential wood combustion plays a major wintertime role in many regions of the world such as Elverum, Norway, a small city in a heavily wooded region (Ramdahl et al., 1984a), Albuquerque, New Mexico (Lewis et al., 1988), Juneau, Alaska (Watts et al., 1988), and Christchurch, New Zealand (Cretney et al., 1985) where domestic wood and coal combustion contributes. Similarly, Pyysalo et al. (1987) reported seasonal variations (late spring vs early winter) in the genotoxicity of particulate and vapor phases of ambient air—and associated levels of PAHs and PACs—in a small industrial town in Finland impacted by emissions from domestic and industrial energy sources (coal, heavy oil, wood, waste wood materials, and peat).

Another seasonal combustion source with major local, and in some cases regional, impacts on particle loadings, PAH levels, and ambient mutagenicities in areas worldwide is large-scale, prescribed open burning of biomass to dispose of crop and forest residues, e.g., sugar cane, orchard prunings, wheat, barley, and rice straws, and Douglas fir and ponderosa pine slash (for emission factors, see Jenkins et al., 1996). For example, Mast and co-workers (1984) reported that in the early 1980s, some 2-3 million tons (U.S.) of these straws were generated annually in California alone, with virtually all of the rice straw burned in the field during October and November and early spring. In one study, they collected rice straw particulate matter both in the field during such a "prescribed burn" and in laboratory experiments in an instrumented burning tower. Particle extracts from all samples were active on strain TA98 with microsomal activation ( + S9). A wide range of PAHs and PACs were isolated and characterized (as well as other organics), including alkylated phenan-threnes, which they suggested contributed to the particle mutagenicity.

During wintertime, in most urban areas, mutagenicity levels are generally higher than during the summer. For example, Fig. 10.18 shows the results of Viras and co-workers (1990) for 24-h average mutagenicities (rev

Was this article helpful?

0 0

Post a comment