Environ. Sci. Technol. 1992, 26, 179-184
Indoor Ozone and Nitrogen Dioxide: A Potential Pathway to the Generation of Nitrate Radicals, Dinitrogen Pentaoxide, and Nitric Acid Indoors Charles J. Weschler” Bell Communications Research, Red Bank, New Jersey 0770 1
Michael Brauer and Petros Koutrakls Harvard University, School of Public Health, Boston, Massachusetts 02 115
Outdoors, in the absence of direct sunlight, significant concentrations of the nitrate radical can result from the reaction between ozone and nitrogen dioxide. The nitrate radical will rapidly equilibrate with dinitrogen pentaoxide, and both of these species can further react to produce nitric acid. We suggest that similar chemistry can occur indoors. A recent study supports this suggestion. Measurements of acidic species in the Boston area indicate significant indoor sources of nitric acid during the summer, but not during the winter. A simple mass balance model (incorporating an 03/N02reaction sequence) yields indoor HN03 values in reasonable agreement with measured values. Whereas outdoors, at night, the major route to gas-phase nitric acid appears to be the homogeneous hydrolysis of Nz05,indoors the dominant path may often be NO3 abstraction of H atoms from volatile organic compounds (VOCs). Such chemistry may also be an important source of free radicals indoors-not just the nitrate radical, but also peroxy and hydroxyl radicals derived from reactions among the nitrate radical and VOCs.
Introduction Indoor ozone concentrations can be a significant fraction of outdoor concentrations (refs 1 and 2, and references therein). This is of concern because ozone can potentially affect health, materials, and processing. In this paper we call attention to the fact that ozone can also play a role in indoor air chemistry. We suggest that in indoor environments with elevated concentrations of both ozone and nitrogen dioxide, the following reaction sequence can occur:
O3 + NO,
2NO3 + O2
NO3 + NOz
Kd9
NzO5
Nz05 + H 2 0 -% 2”03 NO3
(1)
+ ORG 2HN03 + ORG’
(2) (3)
(4)
This chemistry occurs outdoors at night [see FinlaysonPitts and Pith (3)].Nitrate radical concentrations as high as 430 ppt have been measured in Riverside, CA, at a time when the O3 and NO2 levels were 79 and 35 ppb, respectively (4). During daylight hours, these same reactions are relatively unimportant outdoors because the nitrate radical is photolytically unstable. In indoor environments, with the absence of direct sunlight, conditions may approximate those that occur outdoors a t night.
Contrast between Outdoor Nighttime Chemistry and Indoor Chemistry Outdoors, reaction sequence eqs 1-3 are considered the dominant source of HN03. On the basis of simulations of 0013-936X/92/0926-0179$03.00/0
atmospheric chemistry, Russell et al. (5) have concluded that homogeneous hydrolysis of N205dominates nighttime HN03 formation in both an offshore oceanic environment and a dry desert environment. Tuazon and co-workers (6) have calculated a formation rate for HN03 of 0.3 ppb h-l during the night at 3 ppb NO,, 100 ppt NO3, and 50% relative humidity. Indoors, we suggest that the homogeneous chemistry can be dominated by reaction sequence eqs 1, 2, and 4, reflecting the fact that the total concentration of volatile organic compounds (VOCs) is frequently 1 or more orders of magnitude greater indoors than outdoors (7,8). Table I presents a comparison of rate constants for these two reaction pathways. The value of K , used in calculating 21t3[HzO]K,,[NOz] R i 3.95 X cm9 molecule-l (9). For the homogeneous hydrolysis of N205,we are using a rate cm3 molecule-l s-l. This value is constant of 2.7 X an upper limit recently determined by Sverdrup et al. (10). For further discussions of this reaction see Atkinson et al. (11)and Mozurkewich and Calvert (12). Hence, for the N206/Hz0pathway, the pseudo-first-order rate constant listed in Table I is an upper bound. The tabulated value is for 50% relative humidity and 20 ppb NO,; values at other relative humidities and NOz concentrations are proportional. For the N03/ORG reaction pathway, the pseudo-fmt-order rate constant, k4[ORG],is actually a sum of constants for each of the reactions of NO3 with an organic compound found in a given indoor environment. Hence, the pseudo-first-order rate constant for reaction 4 is given by k,[ORG] = Ck,i[ORGJ i
(5)
where kqi is the rate constant for NO3 abstraction of a hydrogen atom from an organic compound ‘‘ORG;. Table I contains a compilation of average indoor concentrations for 32 indoor pollutants. Most of these averages are taken from the national VOC data base assembled by Shah and Singh (8). For each of these compounds, the table also lists the reported (9) rate constant for its reaction with NO3and the resulting pseudo-first-order rate constant, k4i[ORG]. The total of these 32 values is “k4[0RG]”. Comparisons between 2k3[H20]Ke[NO,] and k,[ORG] indicate that, indoors, one would Irequently expect the H-atom-abstraction reaction of NO3 with VOCs to dominate the homogeneous hydrolysis of NzO5 (depending on the NO, concentration). Two of the compounds in Table I, dimethyl sulfide (DMS) and phenol, warrant special comment due to their relatively fast reaction with NO3 [both reactions proceed by H-atom abstraction (9)]. DMS is a common metabolic product of both humans and their pets. Other indoor sources include houseplants, meat cooking, garbage, and cardboard (13). We have been unable to find any indoor measurements of DMS. The indoor level listed in Table I (0.05 ppb) is a conservative estimate based on outdoor
0 1991 American Chemical Society
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Table I. Homogeneous Chemical Reactions Leading to HN03 Formation Indoors
humid.,
molecules cm-3
k3[Hz0],“
%
50
3.8 x 1017
1.0x 10-4
NO3 + ORG
-
HN03
2k3[H20~~eqep[NOz],b
S-1
3.8 x 10-3
d[HN03]/dt = K4[ORG][NO3]
reacting species mean indoor k4i, cm3 k4J0RG], conc,c ppbv mol-’ s-l S-l (ORG) 6.8 X 8.3 X formaldehyde 49.4 2.8 x 10-15 3.4 x 10-4 acetaldehyde 5.0 8.5 X 1.7 X 10” 8.0 acetone 0.8 3.0 x 10-17 5.9 x 10-7 chloroform 4.1 X 10” 3.2 X 5.2 benzene 1.8 3.0 X 1.3 X 10“ l,l,l-trichloroethane 8.0 X 1.8 X 10” methyl ethyl ketone 9.2 3.7 x 10-16 2.5 x 10-5 2.8 o-xylene 2.0 x 10-15 2.8 x 10-5 1,2,4-trimethylbenzene 0.6 7.0 X 5.0 X 10-5 2.9 ethylbenzene 2.5 x 10-15 9.9 x 10-5 1.6 benzaldehyde 3.4 x 10-16 7.3 x 10-5 m,p-xylene 8.7 7.0 X 6.9 X 4.0 1,4-dichlorobenzene 1.1 x 10-15 2.5 x 10-5 0.9 1,3,5-trimethylbenzene 6.9 x 10-17 1.3 x 10-5 7.4 toluene 1.3 X 4.4 X 10” 1.4 cyclohexane 1.1 X 1.8 X 10“ 0.7 n-C6 1.4 X 8.4 X 0.3 n-C7 1.8 X 3.9 X 10” 0.9 n-C8 6.8 X lo4 2.4 X 1.2 n-C9 2.7 X 5.2 X 10” 0.8 n-C10 5.5 X 10” 3.0 X 0.8 n-C11 6.1 X lo4 3.3 X 0.8 n-C12 3.6 X 1.3 X 1.5 n-C13 3.9 x 10-16 7.7 x 10” 0.8 n-C14 4.2 X 2.4 X 10” 0.2 n-C15 4.6 X 10” 9.0 X 0.2 2-butoxyethanol 6.9 X 7.0 X 4.0 1,3-dichlorobenzene 7.0 X 1.7 X 1.0 C4-alkylbenzene 6.0 x 10-16 1.2 x 10-4 8.1d ammonia 9.7 x 10-13 1.2 x 10-3 0.05“ dimethyl sulfide 2.3f 3.6 X lo-’’ 2.0 X lo-’ phenol 2.0 x 10-1 total (k4[ORG]) cm3 molecule-’ s-’, an “Calculated using a value of 2.7 X estimated upper limit (IO). *For NOp concentration of 20 ppbv. Unless otherwise noted, values are taken from the national VOC data base assembled by Shah and Singh (8). dReference 17. ‘Reference 14. f References 15 and 16.
measurements (0.0024.44 ppb) in continental air masses (14). Phenol is also a metabolic product. We are aware of only two indoor studies; the reported level averaged -2.4 ppb in a set of 4 measurements in a large lecture hall (15) and 2.3 ppb in 6 office buildings, 10 kindergartens, and 8 schools (16). In Table I, the rates of the NO3 reactions with DMS and phenol overwhelm the other rates. Without the contributions of the DMS and phenol reactions, the sum of the remaining N03/VOC reaction rate constants is 1.8 X s-l. Some reactions between NO3 and an indoor VOC proceed by addition rather than H-atom abstraction (9) and are unlikely to produce significant quantities of “OB. Common indoor chemicals that react with NO3 via an addition pathway include styrene, a-pinene, and limonene. Under certain conditions the indoor concentrations of these compounds are such that they will compete effectively for NO3 with the other organics listed in Table I. In either case, in indoor settings the reaction of NO3 with organics appears to frequently dominate the homogeneous hy180
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drolysis of N205. The heterogeneous hydrolysis of N205 on indoor surfaces may be fast enough to compete with the organic reactions. However, HN03 is unlikely to be emitted to the gas phase as a consequence of such processes.
Recent Suggestive Study
A recent study by researchers at the Harvard School of Public Health (17) supports the suggestion that ozone may react with nitrogen dioxide indoors to generate nitric acid. The researchers measured indoor and outdoor concentrations of inorganic acidic aerosols and gases during summer and winter periods in Boston and selected suburbs. During summer months they found a mean indoor nitric acid concentration of 0.84 ppb; during winter months the mean indoor concentration was 0.029 ppb. To some extent this reflects differences in outdoor nitric acid concentrations and differences in air-exchange rates between the summer period and the winter period. However, the indoor nitric acid levels during the summer are still higher than would be expected based on simply the outdoor levels and the air-exchange rate, suggesting an indoor source. Table I1 lists previously unpublished data from the Harvard study. Data are presented from three different residences where summer measurements were made. (Three additional residences were excluded from this analysis. One had an incomplete data set. The other two had unvented gas ranges, and a nitric oxide source complicates this discussion.) None of these homes operated air conditioners during the measurement period; the indoor relative humidities typically varied from 40 to 70%. For each residence the sampling dates are listed, as well as the indoor concentrations of HN03 and NO2, the outdoor concentrations of HNO,, and the indoor/outdoor values (I/O’s) for HN03 and SO2. A detailed examination of these data offers further circumstantial evidence for the chemistry described in reactions 1-4. Consider the I/O’s for nitric acid listed in Table 11. Consistent with its chemical reactivity, nitric acid is scavenged by interior surfaces at a relatively fast rate. Depending on near-wall airflow rates, the predicted deposition velocity is 0.04-0.29 cm s-l (18). For the summer period, the Harvard group (17) estimated average air-exchange rates of 3 air changes h-l (ach). Assuming no indoor sources, a deposition velocity of 0.07 cm s-l, a surface-tovolume ratio of 2.9 m-l and an air-exchange rate of 3 ach, the estimated (1)1/0 for nitric acid is 0.29. The tabulated I/O’s for nitric acid are, on average, much higher. Indeed, at the Newton (a Boston suburb) location, the 1/0 for nitric acid is greater than unity on one date (7/27-7/28) and close to unity on another date (7/25-7/26). Clearly, there are indoor sources of nitric acid. A further indication of this fact comes from a date by date comparison of the I/O’s for nitric acid and sulfur dioxide. If there were no indoor sources for either gas, one would expect a significantly smaller 1/0for nitric acid than for sulfur dioxide, since nitric acid is scavenged by indoor surfaces several times faster than sulfur dioxide (18-20). This reflects the fact that nitric acid is far more soluble in water than sulfur dioxide [2.1 X lo5vs 1.24 mol L-’ atm-l (3)] and is also a much stronger acid. Hence, nitric acid is more likely to be captured by indoor surfaces, which are commonly covered with an aqueous film and are often basic. At the Swampscott location, the I/O’s of nitric acid are close to those for sulfur dioxide for each of the five sampling dates; at Newton, the I/O’s for nitric acid are actually larger than those for sulfur dioxide; and at Boston the I/O’s for nitric acid tend to be slightly lower than those
Table 11. Indoor Concentrations (ppbv), Outdoor Concentrations (ppbv), and Indoor/Outdoor Ratios (I/O) for Selected Inorganic Gases Measured at Three Residences in the Greater Boston Area during the Summer of 1989
If0 site Swampscott Swampscott Swampscott Swampscott Swampscott Newton Newton Newton Newton Newton Boston Boston Boston Boston Boston
sampling date
indoor HN03
indoor NOz
outdoor HN03
HN03
SO2
7 130-7 f 31 7 131-8 f 1 8f 1-812 8f 2-8 f 3 813-8 f 4 7123-7 124 7124-7 f 25 7125-7 f 26 7 126-7 f 27 7 f 27-7 128 7f 23-7 f 24 1/24-7125 I f 25-7 126 7f 26-7 f 27 7f 27-7 f 28
0.15 0.32 0.85 0.79 0.78 0.39 0.69 1.77 0.61 1.86 1.43 0.87 1.17 0.83 1.21
11.6 11.6 11.6 14.5 14.5 26.8 28.5 19.9 21.6 19.6 28.7 28.7 15.9 15.9 15.9
0.44 0.74 1.46 2.14 2.25 0.67 1.10 1.85 1.12 1.75 3.02 2.58 3.02 2.17 2.68
0.34 0.43 0.58 0.37 0.35 0.58 0.63 0.96 0.54 1.06 0.47 0.34 0.39 0.38 0.45
0.38 0.44 0.60 0.54 0.50 0.51 0.41 0.56 0.43 0.56 0.49 0.41 0.54 0.53 0.51
for sulfur dioxide, but not as low as expected if there were no indoor source of nitric acid. The I/O's for nitric acid compared to sulfur dioxide offer further evidence for one or more indoor sources of nitric acid in the summer months. Reaction sequence 1-4 may be such an indoor source. For the summer period, the Harvard group (17)estimated average air-exchange rates of 3 ach; at such high air-exchange rates, indoor ozone levels could commonly exceed daily peak values of 40 ppb and daily mean values of 20 ppb (1). The reported mean indoor NO2 level in the summer was 18 ppb. Hence, the indoor levels of O3 and NO2are high enough for reactions 1-4 to be significant (see below). Ozone was not measured as part of the Harvard study (17). (The intent of the Harvard study was to characterize indoor and outdoor concentrations of acidic species, not to conduct a detailed examination of indoor air chemistry.) To estimate indoor ozone values at the residences listed in Table 11, we use outdoor ozone values measured at the nearest EPA monitoring site (Chelsea, MA) coupled with estimates of the 1/0for ozone. Table I11 lists 24-h-average ozone values measured outdoors at the Chelsea site for the relevant sampling dates. As an estimate of 1/0for ozone, at a given residence, on a given sampling date, we use the 1/0for SO2. [I/O for ozone or sulfur dioxide at a given residence, on a given sampling date, is a function of that residence's air-exchange rate and surface-to-volume ratio (1). For each residence, the surface-to-volume ratio will be the same for ozone and sulfur dioxide; hence, we are assuming that O3 and SOz have comparable deposition velocities-see Appendix A.] Table I11 lists indoor ozone concentrations obtained by multiplying the average outdoor ozone value for a given sampling period by the 1/0 of SO2 (see Table 11) for that same sampling period. To facilitate comparisons, the last column in Table I11 repeats the I/O's for HN03 from Table 11. It is apparent that the 1/0for HN03 tends to be highest on days when the indoor ozone is estimated to be high. Such a simple comparison is hindered by the fact that other factors such as the airexchange rate and the indoor concentration of NOz will also influence the indoor concentration of HNO,. The interplay among these factors is best addressed in a model. Mathematical Modeling The appropriate rate and equilibrium constants for reactions 1-4 are fairly well established and have recently been summarized (9,21). The second-order rate constant cm3molecule-I s-l (7.87 for reaction 1at 298 K is 3.2 X
Table 111. 24-h-Average Outdoor Ozone Concentrations (Chelsa, MA) and Estimated 24-h-Average Indoor Ozone Concentrations (See Text) at Three Residences in the Greater Boston Area during the Summer of 1989
site
sampling date
outdoor 03,ppbv
est indoor 03,ppbv
110 HN03
Swampscott Swampscott Swampscott Swampscott Swampscott Newton Newton Newton Newton Newton Boston Boston Boston Boston Boston
7130-7 f 31 7f 31-8 f 1 8f 1-8 f 2 8/2-8/3 813-8 f 4 7/23-7124 7124-7 f 25 7125-7 f 26 I f 26-7 f 27 I f 27-7 f 28 7123-7 124 7 124-7 f 25 7/25-7/26 7f 26-7 f 27 7 127-7 I 2 8
15 27 43 46 56 30 19 47 44 49 30 19 47 44 49
5.8 11.9 25.9 24.9
0.34 0.43 0.58 0.37 0.35 0.58 0.63 0.96 0.54 1.06 0.47 0.34 0.39 0.38 0.45
27.7 15.4 7.9 26.1 19.1 27.5 14.6 7.8 25.3 23.5 25.1 ~
X lo-' ppb-I SI). The equilibrium described in reaction 2 is achieved in less than 1min at 298 K and atmospheric pressure; Keg= 3.95 X lo-" cm3 molecule-l (9). Relevant rate constants for reactions 3 and 4 are presented in Table I. Under appropriate conditions (high relative humidity and/or high levels of VOCs), either reaction 3 or reaction 4 is fast relative to reaction 1. In calculating the indoor generation of HNO, via reactions 1-4, we assume that d[HN03]/dt = kl[03][N02]. If the dominant pathway for HN03production is the reaction sequence 1-3, and if reaction 3 is much faster than reaction 1, then the rate of HN03 production is equivalent to twice the rate of reaction 1 (i.e., 2kl[03][N02]). Even if reaction 3 is not faster than reaction 1, the rate of HNO, production is still twice the rate of reaction 1 if NO3 achieves steady-state levels (see Appendix B). On the other hand, if the dominant pathway for HNO, production is the reaction sequence 1,2,4 (more likely indoors), and if reaction 4 is much faster than reaction 1 or NO3 achieves steady state, then the rate of HN03 production is equivalent to the rate of reaction 1 (see Appendix B). Does the chemistry described in reactions 1-4 actually proceed at a fast enough rate to make a meaningful contribution to the indoor nitric acid levels? To answer this question, we use a simple, one-box mass balance model. Assume the two major sources of nitric acid are infiltration from outdoors and internal generation via reactions 1-4. Also assume the two major sinks are exfiltration and de-
Environ. Sci. Technol., Vol. 26, No. l , 1992 181
However, even in the case of these Newton data, the calculated values increase when the observed values increase, and the calculated values decrease when the observed values decrease. Equation 7 can also be used to estimate the relative contribution of reaction sequence 1-4 to the indoor level of HN03. For the 15 periods considered in Figure 1, the contribution of indoor chemistry varies from 8 to 31?& of the total. Despite the focus of this paper on indoor chemistry, a large fraction of the indoor nitric acid appears to have infiltrated from outdoors.
\'d"0'
? \so1, ' 3 0
Swampscott
q\pq,%;,%%qq,"~q\%;,%%q,l~,\4 ,db,,4 Icewtoi
Boston
SAMPL NG 3ATES Flgure 1. Comparison between indoor nitric acid concentrations calculated using eq 7 and the indoor concentrations measured in the Harvard study ( 1 7 ) .
position onto interior surfaces. Then, given the above discussion d[HN03in]/dt = E[HNO3,,J + ~ I [ O ~ I [ N O-E["O3inI ZI - kd["O3inl (A/V (6) where E is the air-exchange rate [air changes per second (acs), although typically expressed in air changes per hour (ach)]; k1 is the second-order rate constant for the reaction of ozone and nitrogen dioxide (ppb-' s-l), 7.87 X lo-' ppb-l s-l at 298 K (9); k d is the indoor deposition velocity for nitric acid (m s-l), 4 X 104-29 X lo4 m s-', depending on the airflow near the surfaces (18);we use the value 7 X 10" m s-l; and A / V is the surface-to-volume ratio within the residences (m2m-3); a reasonable value is 2.9 m-l, derived from the studies of Mueller et al. (22). Note that the infiltration term E[HN030ut]assumes a penetration efficiency (i.e., the fraction of nitric acid that passes through the infiltration paths) of unity. This is a conservative assumption in evaluating the importance of the internal generation of "OB. In actuality, the penetration efficiency is probably less than 1 (18), and the relative importance of "indoor chemistry" compared to "infiltration" is then even greater than that estimated in the following analysis. Under steady-state conditions
Using eq 7, we calculate indoor concentrations of nitric acid for each of the sampling intervals listed in Table 11. The appropriate values for indoor NO2 and outdoor HN03 are taken from Table 11. Estimated indoor O3levels are taken from Table 111. The air-exchange rate, E, is calculated (1) using the I/Os for SO2 (Table 11) and a deposition velocity m s-l (see Appendix A). Figure 1 (SO,) of 3.6 X compares indoor nitric acid levels, calculated from eq 7, with the measured indoor levels listed in Table 11. For each of the three residences, the calculated values track the observed values quite closely (Le., the values rise and fall together). The calculated values are similar to the observed values at Swampscott and Boston, but are smaller than the observed values at Newton. There are a number of possible explanations for this observation, including (i) underestimated air-exchange rates at the Newton residence due to higher than estimated surface-to-volume ratios or (ii) indoor generation of HN03 from a gas hot-water heater (note: gas appliances were not present at the other sites). 182
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Artifacts The objection might be raised that the high indoor nitric acid levels reported in the Harvard study are an artifact, and that oxidation of HONO by O3 is occurring during sampling (on the alkaline denuder surface) rather than in the indoor atmosphere. If such chemistry was responsible for a meaningful fraction of "OB, one would expect HN03 levels to correlate inversely with HONO levels, both indoors and outdoors. No such correlations were reported, indicating that the relationship was not significant at p < 0.01. Furthermore, such an artifact would be even more important outdoors than indoors, given the higher ozone levels outdoors. Detailed examination of the data sets presents no evidence that this is the case. Finally, the rate at which HONO is converted to HN03 on the denuder surface appears to be too slow for this process to make a significant contribution to the total amount of HNOB collected (17). In a similar fashion, one might suggest that oxidation of NO2by O3 is occurring during sampling (on the sodium carbonate coated denuder surface). However, as noted above for HONO, such an artifact would be more significant outdoors than indoors. The observations do not support this supposition. Furthermore, evidence from laboratory and field studies (23) indicates that reactions of NOz on sodium carbonate coated denuder surfaces are negligible. Other Studies Further support for indoor generation of nitric acid comes from a recent paper by Cass and co-workers at the California Institute of Technology (18). This paper reports nitric acid concentrations, measured indoors and outdoors, at five southern California museums. The authors compared observed I/O's for HN03 with theoretical values, calculated using a method outlined in their article. For each site, the calculation employs measured building parameters, appropriate to the season, and deposition velocities "computed theoretically from data on near-wall air velocities and temperatures" (see Figure 4 in ref 18). The calculation does not include any indoor chemical reactions. Of the five museums, only Sepulveda House has a summer air-exchange rate (with the outdoors) greater than 2 ach. This is also the only museum that has a measured summer 1/0 (k2a) for nitric acid that is larger than the theoretical I/O. Each of the other four locations has a measured summer 1/0 ( M a )that is less than or comparable to the theoretical value. These results may reflect higher indoor levels of O3 and NO2at Sepulveda House, consistent with its higher air-exchange rate, and internal generation of HN03 via the reaction sequence 1-4. In an earlier study (20), Nazaroff and Cass presented a mathematical model for predicting the concentrations of chemically reactive compounds indoors. They applied their model to pollutants inside a museum gallery. On the basis of 2 days when the measured indoor concentrations of O3 and NO2averaged 14 and 52 ppb, respectively, they estimated that the homogeneous reaction of O3 with NO2
indoors could account for 13% of the indoor nitric acid. At higher indoor O3concentrations, the contribution of this reaction would be expected to be even greater. Implications for Materials From previous studies (ref 1,and references therein) we know that indoor ozone levels can frequently be a large fraction of those outdoors. In heavily polluted areas, structures with high air-exchange rates can have indoor ozone concentrations in the range of 100 ppb. At such concentrations, with concomitant NO2levels of 20 ppb, the reaction sequence described in eqs 1-4 has the potential to generate indoor nitric acid at 5.7 ppb h-l. Even if there were no outdoor HN03 to infiltrate, this would result in a steady-state indoor HN03 concentration of 0.5 ppb (1.3 pg m-3) at 4 air changes h-l. At a deposition velocity of 7X m s-l, the accumulation rate on indoor surfaces pg rn-, s-l (7.9 X pg cm-2 day-l). would be 9.1 X Rapid conversion of Nz06to HN03 on moist surfaces (12) could lead to even larger accumulation rates. Nitric acid deposition on cultural objects can be very damaging (18,24). Nitric acid can also have devastating effects on electronic equipment. Nitrates have caused stress corrosion cracking in wire spring relays used in telephone switching equipment (25). Nitrates can contribute to current leakage and shorts in digital equipment (26). At an accumulation rate of 7.9 X pg cm-2 day-l, surface nitrates could threaten alloys sensitive to stress cracking in approximately 500 days (25) and could bridge conductive paths on circuit boards in -3000 days (26). These periods are much shorter than the intended life of the equipment and indicate the potential importance of removing ozone from ventilation air. Conclusions Outdoors, the nitrate radical is generated primarily through the reaction of O3 and NO2. A t night, this is the main reaction driving the nitric acid/peroxy radical chemistry. During the day, this chemistry is relatively unimportant since NO3 is rapidly photolyzed. We suggest that, analogous to outdoor nighttime chemistry, the reaction between O3 and NO2 can be a significant source of HN03 and peroxy radicals indoors. However, whereas in outdoor nighttime chemistry the major pathway to HN03 appears to be the homogeneous hydrolysis of N205 (5), indoors the major pathway to HN03 may often be NO3 abstracting an H atom from vapor-phase organic compounds. The potential dominance of this pathway reflects higher VOC concentrations indoors compared to outdoors (often more than 1 order of magnitude greater). A detailed examination of raw data (previously unpublished) from a study (17) of acidic species in the greater Boston area offers support for the above conjectures. A number of relationships in the data set suggest significant indoor source(s) of HN03 during the summer, but not during the winter: (i) on several summer dates, the I/O’s for HNO, are close to or greater than unity; (ii) the I/O’s for HN03 are frequently comparable to or even greater than the I/O’s for SO,; and (iii) estimated indoor O3 concentrations correlate with indoor HN03 concentrations. A simple model (incorporating a reaction sequence in which O3 initiates the oxidation of NO2, ultimately producing “0,) has been applied to the summer data. The calculated indoor HN03 values are in reasonable agreement with the measured values, with indoor oxidation by ozone accounting for 8-31 % of the indoor HN03. Additional studies (18,20) support the suggestion that indoor ozone can play an important role in indoor chemistry.
The processes discussed in this paper can result in nitrate accumulation on surfaces even in buildings that use highly efficient particulate filters, since such filters have little effect on O3 or NO2. Thus, indoor ozone not only directly threatens materials and equipment, but may also generate secondary pollutants, such as “OB, that are themselves harmful. In regions of the country with severe photochemical smog, it may be cost effective to remove ozone from the ventilation air of buildings that contain nitric acid sensitive objects or equipment. Free-radical chemistry is often considered of negligible importance indoors due to the absence of direct sunlight and the concomitant photochemistry. However, the nitrate radical, in addition to its own rich chemistry ( 9 , 2 1 ) ,is a route to both peroxy radicals (27) and hydroxyl radicals (9,27). In other words, this chemistry may be an important source of free radicals indoors. In selected indoor environments (elevated levels of both O3 and NOz), atmospheric transformations initiated by free radicals may be of potential significance to both health (11) and materials.
Acknowledgments We thank Jack G. Calvert (National Center for Atmospheric Research) for early discussions on this subject and Ashok Gadgil (Lawrence Berkeley Laboratory) for conversations regarding the factors that influence indoor deposition velocities.
Appendix A We have found no reports in the literature in which the indoor deposition velocities of SOz and O3 have been measured simultaneously. There is an attempt (28) to compare the deposition velocities of these two species based on experiments performed by different researchers in different settings; this comparison concludes that the deposition velocity of SOz is about one-third that for 03. However, such comparisons are somewhat tenuous. Factors such as the nature of the air flow and the surfaceto-volume ratios can vary greatly from one setting to another. Based on theoretical considerations, we would expect the deposition velocities of SO2 and O3 to be comparable. The justification for this assumption follows. In a typical indoor setting, flow in the boundary layer (the layer adjacent to a surface) tends to be parallel to the surface. Consequently, a molecule’s “approach” to the surface through the boundary layer is dominated by diffusion. The rate at which a molecule moves through the boundary layer will be proportional to its diffusion coefficient. Given their similar molecular weights (64 vs 48), one would expect SOz and O3to have comparable diffusion coefficients. A second factor that affects the deposition velocity is the likelihood that a molecule will come back off a surface after making “contact”. If every “contact” results in a “stick”, the concentration gradient through the boundary layer will be steeper than if some of the “contacts” result in “bounces”. Since indoor surfaces are frequently covered by an aqueous film, the aqueous solubilities of SO2and O3 are relevant to this discussion. SO2is 100 times more soluble in water than is O3 (1.24 vs 0.013 mol L-l atm-l). Potential reactions with the surface must also be considered. Ozone is a powerful oxidant, whereas sulfur dioxide can function as either a weak oxidant or a weak reductant. However, these redox reactions tend to have large activation energies and are relatively slow. Acid/base reactions are much faster (normally diffusion controlled). Sulfur dioxide is an acidic gas whereas ozone does not participate N
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in acidlbase reactions. Hence, one would expect that SOz is less likely to "leave" a basic surface (e.g., gypsum or concrete) than 03. To summarize, under identical airflow conditions, the deposition velocity of a given molecule indoors is influenced by both its diffusion coefficient and its tendency to come back off a surface that it has made contact with. Based on these factors, it seems unlikely that SOz has a deposition velocity that is one-third that of 03.Indeed, the argument could be made that the deposition velocity of SO2 in indoor environments is likely to be slightly greater than that of ozone. For the purposes of our analysis, we will simply assume that the deposition velocities of SO2 and O3 are comparable. Appendix B
Consider two extremes: either all the HN03 is produced by reaction 3 (case 1) or all the HNO, is produced by reaction 4 (case 2). In the first limiting case, the steady-state concentration of NO3 is given (11)by [NO,Iss = k1[03I /k&eq[H,OI (8) where Keq = [N2O5]/[NO,][NO3]. The rate of HN03 production is given by (94 d[HNO,]/dt = 2k3[Nz05][H&l = 2k&eq[N031 [NO21WZOI
6%)
substituting for [NO,] (eq 7 ) yields d[HNO,]/dt = 2k1[031[NO,]
(10)
In the second limiting case, the steady-state concentration of NO3 is given (11) by [N03lSs= ki[NOzl [oJ/kJORGI (11) In this case, the rate of HNO, production is given by d[HNOJ/dt = k4[NOs][ORG] (12) substituting for [NO,] (eq 10) yields dWNOJ/dt = ki[o3l[NOz1
(13)
Registry No. 03,10028-15-6; NOz, 10102-44-0; NO;', 12033-49-7; Nz05, 10102-03-1; "03, 7697-37-2.
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Received for review April 2, 1991. Revised manuscript received July 24, 1991. Accepted August 1, 1991.
