Contributed by Mario J. Molina, October 9, 2016 (sent for review July 8, 2016; reviewed by Zhanqing Li and Sasha Madronich)
Exceedingly high levels of fine particulate matter (PM) occur frequently in China, but the mechanism of severe haze formation remains unclear. From atmospheric measurements in two Chinese megacities and laboratory experiments, we show that the oxidation of SO2 by NO2 occurs efficiently in aqueous media under two polluted conditions: first, during the formation of the 1952 London Fog via in-cloud oxidation; and second, on fine PM with NH3 neutralization during severe haze in China. We suggest that effective haze mitigation is achievable by intervening in the sulfate formation process with NH3 and NO2 emission control measures. Hence, our results explain the outstanding sulfur problem during the historic London Fog formation and elucidate the chemical mechanism of severe haze in China.
Sulfate aerosols exert profound impacts on human and ecosystem health, weather, and climate, but their formation mechanism remains uncertain. Atmospheric models consistently underpredict sulfate levels under diverse environmental conditions. From atmospheric measurements in two Chinese megacities and complementary laboratory experiments, we show that the aqueous oxidation of SO2 by NO2 is key to efficient sulfate formation but is only feasible under two atmospheric conditions: on fine aerosols with high relative humidity and NH3 neutralization or under cloud conditions. Under polluted environments, this SO2 oxidation process leads to large sulfate production rates and promotes formation of nitrate and organic matter on aqueous particles, exacerbating severe haze development. Effective haze mitigation is achievable by intervening in the sulfate formation process with enforced NH3 and NO2 control measures. In addition to explaining the polluted episodes currently occurring in China and during the 1952 London Fog, this sulfate production mechanism is widespread, and our results suggest a way to tackle this growing problem in China and much of the developing world.
Fine particulate matter (PM), which typically contains a complex mixture of inorganic and organic species, has important implications for several environmental issues (1⇓–3). Presently, the mechanisms leading to PM formation remain uncertain, particularly under highly polluted conditions, hindering efforts in developing effective mitigation policies to reduce their local, regional, and global impacts (1). It is well established, though, that sulfate (SO42−) is ubiquitous and is a key PM constituent in the atmosphere. Moreover, hygroscopic sulfate aerosols serve as efficient cloud condensation nuclei, affecting cloud formation, precipitation, and climate (4⇓⇓⇓–8). A major fraction of regional acid deposition is attributed to the sulfate content that exerts debilitating effects on acid-sensitive ecosystems (9). Furthermore, high levels of fine PM have been implicated in adverse human health issues (1), as exemplified by high fatality during the 1952 London Fog (1, 10). Sulfur compounds are emitted globally from many natural and anthropogenic sources (1⇓–3, 11), and there have been high SO2 emissions from combustion of coal and petroleum products in developing countries (such as China) spurred on by fast economic development (12).
Gaseous SO2 is converted to particulate sulfate through gas-phase oxidation or aqueous reactions, but the detailed chemical mechanisms remain controversial (1⇓–3, 13, 14). The gas-phase oxidation of SO2 is dominated by its reaction with the OH radical, with a lifetime of ∼1 wk at the typical tropospheric level of OH radicals. The aqueous pathways of SO2 oxidation may occur much faster, including reactions with dissolved ozone, hydrogen peroxide, organic peroxides, OH, and NO2 via catalytic or noncatalytic pathways involving mineral oxides (15⇓⇓⇓⇓–20). Most recently, an interfacial SO2 oxidation mechanism involving O2 on acidic microdroplets has been suggested (16).
It has been hypothesized that aqueous SO2 oxidation by NO2 can be an important pathway for sulfate formation under urban conditions and in the presence of sufficient neutralizing agents such as NH3 (2). Several earlier experimental studies, in which gaseous NO2 was exposed to bulk solutions containing sulfite (SO32−) and hydrogen sulfite (HSO3−) ions prepared from Na2SO3, investigated the aqueous sulfur oxidation by NO2; the measured rate constants differed by 1–2 orders of magnitude (17⇓–19). Typically, this aqueous oxidation has been neglected in atmospheric models because of limited water solubility of NO2 (1, 13, 20). A model simulation of dissolution of NO2 in cloud droplets under NOx-rich environments has shown enhanced regional wintertime sulfate by up to 20%, resulting in better agreement between simulations and observations (13). Also, atmospheric measurements have revealed high sulfate production during severe haze events in China (21⇓⇓⇓–25), which cannot be explained by current atmospheric models and suggests missing sulfur oxidation mechanisms (14). Typically, high sulfate levels during haze events in China occurred with concurrently elevated RH, NOx, and NH3 (24, 25), implicating an aqueous sulfur oxidation pathway. However, elucidation of the sulfur oxidation mechanisms from available atmospheric measurements remains challenging, particularly under polluted conditions because of multiple highly coemitted primary gaseous pollutants (1, 21). In this work we investigated the sulfur oxidation mechanism and its role in severe haze formation, by combining field measurements of gaseous pollutants and aerosol particle properties in two Chinese megacities (Xi’an and Beijing) and complementary laboratory experiments (Materials and Methods and SI Appendix).
The pollution episodes in Xi’an exhibit a periodic cycle of 4–5 d, which is reflected in the temporal evolutions of the mass concentrations of SO42− and PM smaller than 2.5 μm (PM2.5) (Fig. 1A and SI Appendix, Fig. S1A and Table S1). For each pollution episode, the SO42− mass concentration increases markedly from less than 10 μg m−3 (clean), 10–20 μg m−3 (transition), to greater than 20 μg m−3 (polluted), with the corresponding increases in the mean PM2.5 mass concentrations from 43, 139, to 250 μg m−3, respectively. Among the main nonrefractory PM2.5 species in Xi’an (Fig. 1B), organic matter (OM), nitrate (NO3−), and SO42− are most abundant throughout the pollution episode. The SO42− mass fraction increases during the transition and polluted (hazy) periods, whereas there is a slight decrease in the OM mass fraction. We quantified the molar ratio of SO42− to SO2, which reflects sulfur partitioning between the particle and gas phases. This ratio ranges from less than 0.1 at relative humidity (RH) <20% to 1.1 at RH >90% in Xi’an, exhibiting an exponential increase with RH (Fig. 1C). During the pollution development, there is increasing RH (Fig. 1D and SI Appendix, Fig. S1B), and the concentrations of SO2, NOx (NO + NO2), and NH3 are highly elevated (Fig. 1D and SI Appendix, Fig. S2 A–C). Clearly, the larger conversion of SO2 to SO42− during the hazy periods is responsible for the enhanced SO42− formation, i.e., with high mass concentrations and mass fractions (Fig. 1 A and B). Field measurements in Beijing also show a similar SO42− evolution. There are noticeable increases in SO42− and PM2.5 mass concentrations during the pollution development (Fig. 1E and SI Appendix, Fig. S3A and Table S2). The SO42− mass fraction increases from clean to polluted periods, in contrast to a decreasing OM mass fraction (Fig. 1F). During the hazy periods in Beijing, the molar ratio of SO42− to SO2 also exhibits an exponential increase with RH (Fig. 1G), and RH and the concentrations of SO2, NOx, and NH3 are high (Fig. 1H and SI Appendix, Figs. S3B and S4 A–C).
Sulfate production during pollution episodes in Xi’an and Beijing. (A–D) Measurements in Xi’an from 17 November to 12 December 2012, and the particle properties correspond to those in PM2.5. (E–H) Measurements in Beijing from 21 January to February 4, 2015, and the particle properties correspond to those in PM1 (particles smaller than 1 μm). In A and E, the dates on the x axis correspond to midnight local time. B and F show the mass fractions of the five main nonrefractory constituents from 5 to 12 December 2012 in Xi’an and from 21 January to 4 February 2015 in Beijing, respectively. The lines in C and G represent the exponential fits through the data, i.e., y = 0.07 + 1.0 × 10−4exp (x/11) with R2 = 0.60 in Xi’an and y = 0.05 + 7.0 × 10−3exp (x/15) with R2 = 0.88 in Beijing. Except for the colors in B and F depicting the aerosol compositions, the blue, orange, and black colors correspond to the SO42− mass concentrations of less than 10 μg m−3 (clean), 10–20 μg m−3 (transition), and greater than 20 μg m−3 (polluted), respectively. The top and bottom of the vertical line for each box in D and H correspond to the 95th and 5th percentiles, respectively, and the top, middle, and bottom horizontal lines of the box mark the 75th, 50th, and 25th percentiles of the data range. The white dot in each box represents the mean value.
Our field measurements demonstrate that efficient conversion of SO2 to SO42− occurs at high RH and concurrently with elevated concentrations of SO2, NOx, and NH3, implicating aqueous sulfate production from the participation of these species. Furthermore, the enhanced sulfate formation during the hazy periods is also accompanied by simultaneously increased formation of particulate NO3− and OM (SI Appendix, Fig. S5). The concentration of ozone is low during the hazy periods in both locations (i.e., a few parts per billion in SI Appendix, Figs. S2D and S4D and Tables S1 and S2), and the visibility is considerably reduced (SI Appendix, Figs. S1C and S3C and Tables S1 and S2), both indicating weak photochemical activity (26, 27). Further examination of the measurements in Beijing reveals markedly continuous growths in the PM2.5 mass concentration and the average particle size throughout the pollution episodes (SI Appendix, Fig. S6 A and C), which are attributable to efficient formation of SO42−, NO3−, and secondary organic aerosol (SOA) during the hazy periods. The considerably reduced photochemical activity during the hazy periods is also reflected in the measured decrease in the photolysis rate coefficient of NO2 (JNO2) (SI Appendix, Fig. S6B), evident from the anticorrelation between JNO2 and PM2.5. Clearly, the efficient PM mass and size growths at high RH and low photochemical activity during the hazy periods are indicative of an increasing importance of aqueous phase oxidation not only for SO42− but also for NO3− and SOA.
A comparison between the two field studies reveals some distinctions. For example, the ratio of SO42− to SO2 at 70% RH is 0.8 in Beijing, much larger than the corresponding value of 0.1 in Xi’an. In addition, the SO42− mass fraction and the total inorganic mass fraction in Beijing are larger than those in Xi’an, indicating that fine PM is more hygroscopic in Beijing (21). The measured contents of Fe and Mn of fine PM are small during the hazy periods in Xi’an (SI Appendix, Table S1), consistent with size-resolved composition measurements showing that the mineral elements are usually enriched in coarse particles, because of their dust origins in China (1, 21, 23). With negligibly low concentrations of water-soluble Fe and Mn (SI Appendix, Table S1), the catalytic capability of the mineral elements in fine PM is limited.
To evaluate the PM acidity during the field campaigns, we calculated the equivalent ratio of ammonium (NH4+) to the sum of SO42− and NO3− (SI Appendix, Fig. S7), because these species represent the dominant nonproton cations and anions in fine PM, respectively. During the hazy periods in Xi’an, this ratio remains near unity (Fig. 2A). Hence, SO42− and NO3− in fine PM are completely neutralized, because of the presence of high levels of gaseous ammonia (17–23 parts per billion, ppb) during the hazy periods. Further analysis of the PM2.5 chemical compositions reveals that the equivalent ratio of the total nonproton cations (NH4+, Na+, Ca2+, Mg2+, and K+) to anions (SO42−, NO3−, and Cl−) is also near unity (Fig. 2B and SI Appendix, Fig. S8A), with the mean values of 1.15 ± 0.14 and 1.06 ± 0.06 during the transition and polluted periods, respectively. The close balance between these cations and anions in Xi’an further confirms that fine PM exhibits negligible acidity. Similarly, the equivalent ratio of NH4+ to SO42− and NO3− is slightly larger than unity throughout the pollution episodes in Beijing (Fig. 2A and SI Appendix, Fig. S7B). When the chloride anion, which likely exists as NH4Cl in ambient PM, is included, the ratio is reduced to 1.09 ± 0.11 during the polluted period (Fig. 2B and SI Appendix, Fig. S8B). Hence, fine PM in both locations is effectively neutralized by ammonia with a calculated pH ∼7 (SI Appendix, Tables S1 and S2), when rapid sulfate production occurs during the polluted period. Interestingly, our results of fully neutralized fine PM in China are in contrast to a recent study showing highly acidic aerosols in the southeast United States, despite declining atmospheric sulfate concentrations over the past 15 years (28).
Neutralizing effect of ammonia on fine PM. (A) Equivalent ratio of NH4+ to the sum of SO42− and NO3− in Xi’an and Beijing. (B) Equivalent ratio of the total cations to anions in Xi’an and equivalent ratio of NH4+ to the sum of SO42−, NO3−, and Cl− in Beijing. The blue, orange, and black colors correspond to the clean, transition, and polluted periods, respectively, as defined in Fig. 1.
To elucidate the mechanism of SO2 oxidation and interpret the rapid sulfate production in our field measurements, we conducted a series of laboratory experiments by exposing pure water or ammonium (3 wt %) solutions under dark conditions to gaseous SO2 and NO2 in a reaction cell. Sulfate formation was quantified (SO42− at m/z = 96, SI Appendix, Fig. S9A) by thermal desorption-ion drift-chemical ionization mass spectrometry (TD-ID-CIMS) (29, 30). When the pure water or ammonium solutions were exposed simultaneously to SO2 and NO2 using either N2 or air as the buffer gas, significant SO42− production was detected, and the signal was higher in the ammonium solution than in pure water (SI Appendix, Table S3). In contrast, the SO42− production was absent for only SO2 exposure under similar conditions, indicating the oxidizing role of NO2. Also, there was little difference in the measured SO42− production between experiments using N2 and air, suggesting negligible SO2 oxidation by O2 molecules.
We performed additional experiments by exposing seed particles to gaseous SO2, NO2, and NH3 under dark and variable RH conditions in a reaction chamber (SI Appendix, Fig. S10). Size-selected oxalic acid particles, which were used to represent organic aerosols that dominate the early stages of haze development in China (21) (see also Fig. 1 B and F), were simultaneously exposed to SO2, NO2, and NH3, while the variation in the dry particle size was monitored. The evolution in particle size distributions measured after exposure to three different RH conditions is depicted in Fig. 3A: The size distribution remains unchanged at 30% RH (i.e., identical to that of the initially seeded particles), whereas exposures at 60% and 70% RH lead to dramatic shifts to larger size distributions. We also conducted experiments to analyze the chemical composition of exposed particles. SO42− production in collected particles after the exposure at high RH is clearly evident (Fig. 3B and SI Appendix, Fig. S9B).
Aqueous sulfate formation in the reaction chamber. (A) Evolution in the dry particle size distribution when sized selected oxalic acid particles are exposed to SO2, NO2, and NH3 under three different RH conditions in a 1-m3 reaction chamber. (B) Desorption spectra of particles collected by TD-ID-CIMS after exposure to SO2, NO2, and NH3 at 65% RH. (C) Particle growth factor after exposure to SO2, NO2, and NH3 as a function of RH. The exponential fit is y = 1.05 + 4.0 × 10−5exp (x/6.8) with R2 = 0.96. Each point corresponds to three measurements, and the error bar denotes the SD (1σ). (D) Integrated desorption peak areas of particles collected by TD-ID-CIMS after exposure to SO2, NO2, and NH3 as a function of RH. The exponential fit is y = 1.8 × 106 + 417exp (x/7.9) with R2 = 0.97. All experiments were performed under the dark condition and at temperature of 298 K. The exposure time was 60 min, and the initial particle size was 45 nm. The initial gas concentrations were 250 ppb for SO2 and NO2 and 1 ppm for NH3.
We quantified the growth of seed particles after exposure to SO2, NO2, and NH3, on the basis of the measured ratio of the dry particle sizes (Dp/Do), where Dp and Do are the values after and before the exposure, respectively. The growth factor increases with RH, with the values of near unity (no growth) at RH< 20% and 2.3 at 70% RH (Fig. 3C). The size growth after exposure is explained by SO42− production, as depicted in the corresponding increase of the integrated SO42− desorption peak areas (Fig. 3D). Both the particle size growth and the SO42− formation after the exposure to SO2, NO2, and NH3 display a similar exponential increase with RH, consistent with increasing hygroscopic growth of oxalic acid particles (31). We performed additional measurements when one or both of NO2 and NH3 were excluded from the exposure: In the absence of NO2, NH3, or both from the exposure, no observable particle growth or SO42− production was measured (SI Appendix, Table S4).
On the basis of our combined field and laboratory measurements, we establish the occurrence of an overall aqueous reaction between SO2 and NO2,
Our results indicate that aqueous SO2 oxidation by NO2 is favored in two atmospheric scenarios (Fig. 4), i.e., under cloud/fog conditions and on fine aerosols with high RH (> 60–70%) and sufficient neutralization (pH ∼ 7). In-cloud SO2 oxidation proceeds in the presence or absence of a neutralizing agent, when elevated levels of SO2 and NO2 coexist (Fig. 4A). Cloud droplets are large (exceeding several tens of micrometers in sizes), and the amount of sulfate formed is sufficiently diluted and does not appreciably alter the particle acidity (1⇓–3). Consequently, water evaporation from cloud droplets under unsaturated conditions leads to concentrated sulfuric acid particles (33), contributing to acid rain and regional acid deposition (9, 13). Also, in-cloud SO2 oxidation can be further enhanced in the presence of the basic species, as demonstrated in our laboratory work for NH3 (SI Appendix, Table S3). On the other hand, the oxidation on fine PM is inhibited by the acidity effect, and the presence of basic species (i.e., NH3) is necessary to maintain the oxidation (Fig. 4B),
Schematic of the sulfate formation mechanisms. Variations in temperature, RH, and particle size and acidity for the aqueous reactions between SO2 and NO2 leading to SO42− formation under in-cloud conditions (A) and on fine PM (B). The red, black, yellow, and green colors in A and B represent SO42−, H+, NH4+, and SOA, respectively. (C) Anticorrelation between the photochemical activity and aqueous chemistry during the severe haze evolution (i.e., from the clean, transition, to polluted periods) in China, displaying the central role of the SO2 to SO42− conversion in facilitating aqueous production of the major secondary constituents.
Hence, the acidity, hygroscopicity, and RH represent the key factors for sulfate formation on fine PM, explaining the differences in the various ambient measurements (14, 20⇓⇓⇓⇓–25). For example, the acidity effect on fine aerosols is effectively overcome by NH3 neutralization in Xi’an and Beijing (Fig. 2). Also, the noticeably earlier increase of the SO42− to SO2 ratio with RH in Beijing than in Xi’an (Fig. 1 C and G) is attributable to more hygroscopic aerosols, because of a larger inorganic mass fraction in Beijing (Fig. 1 B and F) (21).
We derived the equivalent SO2 uptake coefficient (γ) for sulfate production from our field and laboratory results in SI Appendix, Tables S5 and S6, respectively. The γ-values derived from the Beijing measurements are (2.1 ± 1.6) × 10−5 and (4.5 ± 1.1) × 10−5 during the transition (41% RH) and polluted (56% RH) periods, respectively, compared with (8.3 ± 5.7) × 10−5 at 65% RH and (3.9 ± 1.2) × 10−4 at 70% RH derived from the laboratory measurements. Hence, our laboratory experiments reproduce the rapid sulfate production measured under polluted ambient conditions, and these kinetic data are applicable for quantifying sulfate formation in atmospheric models (1, 13, 14).
Our results indicate that the formation of the various secondary organic and inorganic constituents in fine PM is mutually promoting and the severe haze development involves a transition from photochemical to aqueous phase processes (Fig. 4C). During the early stage, efficient photochemical oxidation of volatile organic compounds (VOCs) leads to SOA formation (Fig. 1 B and F), which provides an aqueous media for subsequent SO42− production. With high RH and low photochemical activity during the later hazy periods, continuously large PM growth (i.e., the SO42−, NO3−, and OM mass increases) is maintained by the aqueous chemistry (SI Appendix, Fig. S5). In particular, the SO42− production likely represents the most critical step in initializing the aqueous chemistry, because of increasing particle hygroscopicity. Efficient SO2 to SO42− conversion not only contributes to the high SO42− production rate, but also enhances formations of NO3− and SOA on aqueous particles, explaining the sustained high production of the major secondary constituents during the hazy periods in our current field measurements and those of the previous studies in China (1, 21, 40). For example, with reduced photochemistry during the hazy periods, the measured large NO3− mass concentration is attributable to an enhanced heterogeneous conversion of NOx to HNO3, because the hydrolysis reaction of N2O5 occurs efficiently on sulfate aerosols (41). Also, hydration and oligomerization reactions of glyoxal and methyglyoxal, which are produced with high yields by aromatic hydrocarbon oxidation from traffic emissions, are enhanced by sulfate formation, because these reactions are highly dependent on particle hygroscopicity (30, 32, 42). Furthermore, gaseous HONO formed from the aqueous SO2 oxidation with NO2 provides an additional photochemical OH source that enhances the atmospheric oxidizing capability during the hazy periods (43). It should also be pointed out that severe haze formation in China is characterized by a complex interplay between meteorological, thermodynamic, and chemical processes (1, 21, 44).
Atmospheric sulfur chemistry has remained an open problem (1, 13, 14). The formation of the 1952 London “Killer” Fog is still mysterious in terms of the detailed chemical mechanism for SO2 conversion to sulfate (1, 45). Our results indicate that the formation of London Fog was similar to in-cloud SO2 oxidation by NO2 (Fig. 4A), because both species were present in highly elevated levels as the coproducts of coal burning. The sulfate formation was greatly facilitated by high RH, low temperature, and the presence of large fog droplets (45), yielding elevated sulfuric acid levels that persisted throughout the event. The particle acidity was regulated by temperature, and water evaporation from fog droplets at warmer temperature resulted in concentrated sulfate acid particles (33), explaining the highly acidic nature of the London Fog (45).
Interestingly, we show that the same sulfur problem persists presently to contribute to severe haze formation in China, although the fine PM is mainly nonacidic. Major emission sources in China include industry (for SO2, VOCs, and NOx) and traffic (for VOCs and NOx), because of its fast-growing economy and urbanization (1, 23, 46, 47). Also, there has been a rapid increase in the production and use of nitrogen fertilizers in China, leading to high NH3 emissions (48). For example, the emissions of SO2, NOx, and NH3 in China are estimated to be about 22 Tg S y−1, 19 Tg N y−1, and 15 Tg N y−1 in 2010, respectively (48). In addition, traffic emissions have been suggested to represent an important urban NH3 source (49). High emissions of these organic and inorganic PM precursors result in large secondary production of SO42−, NO3−, NH4+, and SOA in China (Fig. 4C), via the combined atmospheric photochemical and aqueous processes (1, 21, 40, 50⇓⇓–53). Our results indicate that sulfate production is key to the formation of persistent severe haze in China (Fig. 4 B and C). Whereas current efforts have been focused primarily on minimizing SO2 emissions (1, 14, 21), significant haze reduction may only be achievable by disrupting this sulfate formation process. For example, controlled NH3 emissions may be important, because the acidity effect represents the key rate-limiting factor in sulfate production on fine PM. Also, because of the second-order nature of NO2 in the aqueous SO2 oxidation (i.e., reactions 1 and 2), reduction of the NOx level is likely effective in lowering sulfate formation. In light of large contributions to urban NOx, VOC, and NH3 levels from transportation (1, 21, 49), regulatory actions in minimizing traffic emissions may represent the critical step in mitigating severe haze in China. These measures are clearly supported by our experimental results, showing no particle growth or sulfate formation at high RH when oxalic acid particles were exposed to high levels of SO2 in the absence of NH3, NO2, or both (SI Appendix, Table S4).
In addition to polluted urban areas, efficient sulfate production is also expected in the proximity of power plant and biomass burning plumes and ship tracks (1⇓–3, 9, 13), where SO2 and NOx are coemitted. Because of increasingly high SO2, NOx, VOC, and basic species (NH3 and amine) emissions in many developing countries (1, 38, 48, 49), the synergetic sulfate formation pathway identified in our work is likely widespread globally, contributing not only to air quality problems but also to enhanced nitrogen (i.e., NH4+ or NO3−) or acid (in the absence of basic species) deposition, with major implications for the ecosystem vitality, greenhouse gas budgets, and biological diversity (48). Our results highlight the necessity for comprehensive understanding of the atmospheric aerosol chemistry in the development of effective pollution mitigation policies (1), to minimize the impacts of fine PM on visibility, human health, ecosystems, weather, and climate.
Field measurements of gaseous and PM pollutants were performed in Xi’an and Beijing. The sampling site in Xi’an (from 17 November to 12 December 2012) was located on the rooftop (around 10 m above the ground) of a three-story building on the campus of the Institute of Earth Environment of Chinese Academy of Science (CAS) in the southwest of the city (54). The sampling site in Beijing (from 21 January to 4 February 2015) was located on the campus of Peking University in northwestern Beijing (21). Gaseous species and PM properties were monitored by a suite of instrumentations and methods (54⇓⇓⇓⇓⇓⇓⇓⇓⇓–64). Laboratory experiments were performed to evaluate SO2 oxidation by NO2 on bulk solutions and aerosols under dark conditions (see also SI Appendix). Pure water or 3 wt % NH3 solution was exposed to SO2 and NO2 in N2 or pure air using a reaction cell. The exposed solution from the reaction cell was analyzed by TD-ID-CIMS for sulfate formation. To evaluate the conversion of SO2 into SO42− on aerosols under conditions relevant to the atmosphere, we conducted experiments by exposing seed particles to SO2, NO2, and NH3 and measuring the size variation and sulfate formation on the exposed particles in a 1-m3 Teflon reaction chamber covered with aluminum foil (SI Appendix, Fig. S10). Size-selected oxalic acid particles (45 nm) were used as model aerosols in the reaction chamber for the aqueous conversion of SO2 to sulfate, by exposing to SO2, NO2, and NH3 at variable RH. The variations in the dry particle sizes and sulfate formation were measured by a differential mobility analyzer and TD-ID-CIMS, respectively. Additional descriptions of the instrumentation and procedures of the field and laboratory measurements are provided in SI Appendix.
G.W. acknowledges the National Natural Science Foundation of China and the Strategic Priority Research Program of the CAS for financial support (Grants 41325014, XDA05100103, and XDB05020401). This work was partially supported by the Robert A. Welch Foundation (Grant A-1417), the Ministry of Science and Technology of China (Grant 2013CB955800), a collaborative research program by Texas A&M University and the National Natural Science Foundation of China. M.H. acknowledges the National Basic Research Program, China Ministry of Science and Technology (Grant 2013CB228503), National Natural Science Foundation of China (Grant 21190052) and the China Ministry of Environmental Protection’s Special Funds for Scientific Research on Public Welfare (Grant 20130916). W.M.-O. was supported by the National Science Foundation Graduate Research Fellowship Program, and B.P. was supported by the NASA Earth and Space Science Fellowship Program.
↵1Present address: Center for Excellence in Regional Atmospheric Environment, Institute of Urban Environment, Chinese Academy of Sciences, Xiamen 361021, China.
Author contributions: G.W. and R.Z. designed research; G.W., R.Z., M.E.G., L.Y., M.L.Z., M.H., Y. Lin, J.P., S.G., J.M., J.L., C.C., T.H., Y.R., Y.S.W., J.G., J.W., P.T., W.M.-O., J.S., Z.D., J.Z., Y.H., Y.W., Y.Z., Y. Li, J.H., B.P., Y.C., Y.J., F.Z., and M.J.M. performed research; G.W., R.Z., M.H., and M.J.M. contributed new reagents/analytic tools; G.W., R.Z., M.E.G., L.Y., M.L.Z., M.H., Y. Lin, J.P., J.C., Z.A., W.Z., G.L., D.S., L.Z., M.S., W.W., Y.W., L.C., D.R., P.S.L., R.A.D., C.E.K., and M.J.M. analyzed data; and G.W. and R.Z. wrote the paper.
Reviewers: Z.L., University of Maryland; and S.M., National Center for Atmospheric Research.
The authors declare no conflict of interest.
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