FIGURE 15.16 Annual average concentrations of PM2 5 indoors as a function of smoking in the home. The data are shown in percentiles as marked. (Adapted from Neas et al., 1994.)
solids in the water used in the humidifier. When an ultrasonic humidifier was operated in a closed single room using tap water with total dissolved solids of 303 mg L-1, concentrations of fine and coarse particles of 6307 and 771 /xg m~3, respectively, were generated. When distributed over the whole house, the corresponding levels were 593 and 65 ¡xg m~3, respectively. The whole-house values fell to 41 and 13 /xg m~3 when bottled water with total dissolved solids of 24 mg L_1 was used (Highsmith et al., 1992).
On the other hand, the use of evaporative ("swamp") coolers appears to decrease particulate matter indoors. For example, Quackenboss et al. (1989) report levels of PM25 and PMI0 in homes having such coolers (for both smokers and nonsmokers) that are about half that of homes without them.
Where indoor heating and cooking involves the use of coal or biomass, indoor particle concentrations can be extremely large. For example, Florig (1997) and Ando et al. (1996) report that in China typical indoor total suspended particle (TSP) concentrations can be in the range from 250 to 900 jug m-3 in homes using coal and 950-3500 /xg m~3 in those using biomass fuels. These levels can be compared to annual average outdoor concentrations of 250-4f0 ¡xg m~3. The high concentrations associated with coal burning combined with the mutagenic nature of the emissions have been suggested to be responsible for enhanced lung cancer in China (Mumford et al., f 987). Similarly, Davidson et al. (1986) measured TSP concentrations of 2900-42,000 ¡xg m~3 in homes in Nepal that used biomass fuels, compared to outdoor levels of 280 ¡xg m~3. For particles with diameters less than 4 /¿m, the levels ranged from 870 to 14,000 jag m"3.
Similar conclusions regarding the relative indoor and outdoor concentrations have been reached in studies of office and commercial buildings. For example, Ligocki et al. (1993) measured indoor and outdoor concentrations of particles and their components at five museums in southern California. The indoor-to-outdoor ratios of particle mass varied over a wide range, depending to a large extent on the ventilation and filtration systems in use. Ratios varied from 0.16 to 0.96 for particles with diameters less than 2.1 /¿m and from 0.06 to 0.3 for coarse particles with diameters greater than this.
The chemical composition of particles collected in the museums was also compared to that outdoors using a mass balance model. The results indicated that there were significant indoor sources of fine particle organics and that this source(s) was a significant fraction of the total indoor fine particle organic concentration (Ligocki et al., 1993). A similar conclusion was reached by Naik et al. (1991) in measurements made in a telephone switching office. The levels of the n-C27 through n-C33 alkanes in the fine particle fraction were all elevated, suggesting indoor sources such as waxes, polishes, and lubricants. Enhanced levels of dibutyl phthalate and di(2-ethylhexyl) phthalate were also observed in fine particles and attributed to plasticizers used in floor polishes and vinyl products.
Turk et al. (1989) measured particle concentrations in 38 buildings that had both smoking and nonsmoking areas. The average mass concentration of respirable particles in smoking areas was 70 /xg m"3 compared to 19 ¡xg m~3 in nonsmoking areas, whereas the outdoor concentrations were essentially the same, 19 ¡xg m~3.
In another study, Ott et al. (1996) measured particle concentrations inside and outside a tavern before and after a smoking ban was instated. Average respirable suspended particle mass concentrations were 57 ¡xg m~3 above the outdoor concentrations prior to the ban, compared to 6-13 /xg m~3 afterward. Cooking and resuspended dust also contributed to the indoor particle mass concentration, but at concentrations about 20-25% of that due to cigarette smoke.
In short, the indoor concentrations of particles depend on the outdoor levels, the ventilation system and exchange rates, and the presence of indoor sources such as cigarette smoke.
In many nonresidential buildings, deposition is of particular interest because of the potential for damage to materials in museums, offices, cultural objects, and industrial sites (e.g., Sinclair et al., 1988, 1990a,b; Nazaroff et al., 1990a; Salmon et al., 1994, 1995). Deposition onto both horizontal and vertical surfaces is of concern, but these can show different behaviors as a function of particle size. For example, deposition of particles to vertical and horizontal surfaces was measured inside five museums. Horizontal deposition velocities increased from ~10~6 to 10~3 m s_1 as particle size increased from about O.f to 30 ¡xm as expected for gravitational settling (see Chapter 9.A.3). However, the dependence on particle size of the uptake onto vertical surfaces, which is influenced by thermal and air flow fields (Nazaroff et al., 1990b), was variable, increasing with particle size in some cases, decreasing in others, and, in some, showing no dependence on particle size (Ligocki et al., 1990).
A number of models have been developed for particles indoors (e.g., Nazaroff and Cass, 1989a; Sinclair et al., 1990b; Nazaroff et al., f990a; Weschler et al., 1996; Wallace et al., 1996, and references therein). This is a complex problem, given the number of potential sources, different deposition velocities for particles of different sizes (e.g., see Chapter 9.A.3 and Nazaroff and Cass (f989b)), the different particle compositions, and the effects of outdoor concentrations and ventila tion rates. However, reasonably good agreement has been obtained in many cases between modeled and observed concentrations.
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