image
Home Journals Progress in Chemistry
Progress in Chemistry

Abbreviation (ISO4): Prog Chem      Editor in chief: Jincai ZHAO

About  /  Aim & scope  /  Editorial board  /  Indexed  /  Contact  / 
Online first  /  Current issue  /  All issues  /  Top access  /  RSS

Progress in Chemistry >

2023 , Vol. 35 >Issue 9: 1313 - 1326

DOI: https://doi.org/10.7536/PC221201

Review

Formation Mechanisms of Secondary Sulfate and Nitrate in PM2.5

  • Fangfang Guo ,
  • Shaodong Xie , *
Expand
  • College of Environmental Sciences and Engineering, Peking University,Beijing 100871, China
*e-mail:

Received date: 2022-12-28

  Revised date: 2023-08-16

  Online published: 2023-08-23

Supported by

The National Key Research and Development Program of China(2018YFC0214001)

Fold

Abstract

Secondary inorganic sulfate and nitrate, as the key chemical components of PM2.5, play important roles in the formation of severe regional haze events. The deteriorating sulfate and nitrate pollution has brought more serious challenges to the continuous improvement of air quality. Thus, elucidating the formation pathways and key factors of controlling the formation of inorganic sulfate and nitrate is crucial to eliminate PM2.5 pollution in the atmosphere. The formation of sulfate and nitrate involves complex chemical reactions, including gas- and aqueous-phase reactions and multi-phase reactions. Recent experimental and filed studies have revealed new reaction mechanisms and detailed reaction kinetics for the oxidation of SO2 and NO2 to form sulfate and nitrate. Merging new formation pathways of sulfate and nitrate with updating reaction kinetics based on laboratory measurements and field observations, air quality model performance is effectively improved to capture the spatial-temporal variations of sulfate and nitrate and identify their chemical formation pathways. This review provides a synthetic synopsis of recent advances in the fundamental mechanisms of sulfate and nitrate formation. In particular, the mechanisms and reaction kinetic results for a series of individual reaction pathways of current interest for the SO2 and NO2 oxidation are emphasized. The key factors affecting the SO2 and NO2 oxidation rates and significant challenges in laboratory studies of characterizing the reaction kinetics are also discussed. In addition, the sensitivity of nitrate to emission reductions of nitrogen oxides (NOx), volatile organic compounds (VOCs) and ammonia (NH3) is investigated. Finally, suggestions are put forward for the future research directions to improve the understanding of sulfate and nitrate formation.

Contents

1 Introduction

2 Mechanism of particulate sulfate formation

2.1 Gas-phase oxidation

2.2 Aqueous-phase oxidation

2.3 Heterogeneous oxidation

2.4 Multiphase photochemical oxidation

3 Mechanism of particulate nitrate formation

3.1 HNO3formation

3.2 HNO3-NO3- partitioning

4 Conclusion and outlook

Key words： nitrate; sulfate; formation mechanism; aerosol liquid water content; aerosol acidity

Cite this article

Fangfang Guo , Shaodong Xie . Formation Mechanisms of Secondary Sulfate and Nitrate in PM2.5[J]. Progress in Chemistry, 2023 , 35(9) : 1313 -1326 . DOI: 10.7536/PC221201

1 Introduction

Atmospheric fine particulate matter (e.g., PM2.5, aerodynamic diameter ≤ 2.5 μm) is a major driver of urban haze and has important impacts on air quality, climate, human health, and ecosystems[1⇓~3]. Sulfate ( S O 4 2 -) and nitrate ( N O 3 -) are the most important secondary inorganic components in fine particles. The aggravation of haze is often accompanied by the rapid rise of secondary sulfate and nitrate. During heavy haze, the hourly average concentration of sulfate can exceed 100μg·m-3, accounting for 25% of the mass concentration of PM2.5[4]. With the implementation of air pollution prevention and control measures in China, the emission structure of pollution sources has changed greatly, resulting in nitrate gradually surpassing sulfate in some areas and becoming the most important secondary inorganic aerosol in PM2.5[5][6,7]. At present, sulfate and nitrate aerosols are important target components for urban air quality control.Clarifying the formation mechanism and key influencing factors of sulfate and nitrate aerosols is the key to improve PM2.5 pollution.
The formation mechanism of sulfate and nitrate aerosols in the atmosphere is very complex, which is the result of photochemical reaction, liquid phase reaction, heterogeneous reaction and other chemical processes[8⇓~10]. The formation of both is affected by many factors, such as precursor concentration, solar radiation, atmospheric oxidation, relative humidity, temperature, aerosol pH and liquid water content (ALWC)[11⇓⇓~14,8]. In recent years, in addition to the traditional reaction pathway in which NO2 and SO2 molecules are completely dissolved in cloud or aerosol particles and converted to sulfate and nitrate,New reaction pathways (such as heterogeneous reactions on the surface of aerosol particles, photochemical reactions, etc.) And reaction kinetics have been used to explain the rapid accumulation of sulfate and nitrate under polluted conditions, and their use in air quality models can effectively improve the simulation of the spatial and temporal distribution characteristics of sulfate and nitrate in numerical models.
In this paper, the important reaction pathways of the oxidation of SO2 and NO2 to secondary sulfate and nitrate are summarized, the mechanisms of the related reactions, the kinetic results of the reactions, and the factors affecting the conversion of secondary sulfate and nitrate are discussed, and the key scientific issues to be solved in the conversion of SO2 and NO2 to sulfate and nitrate in future research are proposed.

2 Sulfate formation mechanism

S O 4 2 - is mainly produced by SO2 in the atmosphere through gas phase, liquid phase, heterogeneous phase and photochemical oxidation. Although there have been many studies on the conversion process of SO2 to SO 4 2 -, the specific chemical mechanism and the importance of different oxidation pathways are still controversial.

2.1 Gas phase oxidation

In the atmosphere, the gas-phase oxidation of SO2 mainly reacts with OH radicals, and the reaction equation is as follows:[15]
S O 2 + O H + M → H O S O 2 + M
H O S O 2 + O 2 → H O 2 + S O 3
S O 3 + H 2 O + M → H 2 S O 4 + M
The gaseous sulfuric acid (H2SO4(g)) formed by the reaction either nucleates to form new particles, attaches to existing aerosol particles, or reacts with NH3 or NaCl to form sulfate aerosol[15,16]. The gas-phase oxidation of SO2 by OH is a slow process at typical atmospheric OH radical levels, and the lifetime of SO2 is 5 – 10 days[10]. Therefore, the oxidation of SO2 to sulfate can occur during long distance transport[17]. The results of global scale simulation show that the gas-phase reaction pathway contributes about 25% to the formation of secondary sulfate[18]. However, during urban haze pollution, the decrease of solar radiation reduces photochemical activity, which inhibits the formation of OH free radicals, and then reduces the contribution of this pathway to sulfate formation[19].
Recent studies have shown that a stable Criegee intermediate (sCI, which can be generated during the ozonolysis of unsaturated hydrocarbons) may play an important role in the gas-phase oxidation of SO2[20⇓~22]. Welz et al. Measured the reaction kinetics of sCI with SO2, and the results showed that the reaction rate of sCI with SO2 was 50 to 10 000 times faster than previous estimates, and the sCI+SO2 reaction could produce as much SO 4 2 - as the OH+SO2 reaction[23]. The simulation results of Boy et al showed that the contribution of sCI and SO2 to the concentration of H2SO4 in the near-surface atmosphere could reach 33% ~ 46%[24].
sCI can react with SO2 to form a gaseous SO3(g), which then reacts with atmospheric water vapor to form sulfuric acid (H2SO4(g))[25,26]. sCI (such as (CH3)2COO, CH2OO, etc.) can also directly react with water vapor to reduce the formation of H2SO4 in the atmosphere. The reaction of sCI with water vapor depends largely on its own structure. The larger molecular weight of sCI can avoid the reaction with H2O and directly oxidize SO2[26][27]. The specific reaction equation of the gas phase reaction is as follows:
S O 2 + s C l → S O 3 ( g )
S O 3 + H 2 O → H 2 S O 4 ( g )

2.2 Liquid phase oxidation

Dissolved SO2(S(Ⅳ)) to sulfate (S (Ⅵ)) is the most important chemical transformation in the atmospheric liquid phase[15]. In urban atmospheres, sharply increased sulfate is often accompanied by high RH, indicating the importance of liquid-phase oxidation of SO2[28]. The liquid-phase oxidation of SO2 is usually faster than the gas-phase oxidation, which is mainly divided into two steps: first, SO2 and oxidant dissolve into cloud, fog and aerosol liquid water according to Henry's law, followed by the liquid-phase reaction[29]. S (Ⅳ) in the liquid phase has three existing forms: hydrated SO2(SO2·H2O), bisulfite ion ( HSO 3 -) and sulfite ion ( SO 3 2 -), and its main existing form varies with the acidity of the liquid phase. The liquid phase oxidants reacting with S (Ⅳ) mainly include hydrogen peroxide (H2O2), O3, NO2, O2 and hypohalous acid (HOX)[10]. The solubilities of SO2 and oxidant increase with decreasing temperature, so it is more favorable for liquid-phase oxidation of SO2 in winter.

2.2.1 H2O2

H2O2 is one of the most effective oxidizing agents for S (Ⅳ) to S (Ⅵ) in liquid phase[15]. The solubility of SO2 increases with pH, while the reaction rate constant of S(Ⅳ)+H2O2 is opposite, and the H2O2 pathway is less affected by pH because the dependence of the two on pH counteracts each other[30]. The high ionic strength of aerosol can significantly increase the oxidation rate of S (Ⅳ) to S (Ⅵ) by H2O2[31]. Studies have shown that the H2O2 oxidation of SO2 in liquid aerosols is the key way of sulfate formation in severe haze pollution events[32,33]. Liu et al. Found that the sulfate formation rate of the H2O2 reaction pathway was still higher than that of other reaction pathways when the pH value of aerosol was as high as 6.2 due to the high ionic strength of aerosol liquid water[33]. Since the reaction rate of this pathway is independent of aerosol pH, the pathway can maintain a high rate even if the aerosol pH is lower. Thus, the H2O2 oxidation pathway in aerosol liquid water is an important missing sulfate source during severe haze pollution.

2.2.2 O3

The solubility of O3 is 17 orders of magnitude lower than that of H2O2, so the concentration of O3 in droplets is much lower than that of H2O2[15]. Unlike the H2O2 oxidation pathway, the S(Ⅳ)+O3 reaction pathway is highly sensitive to liquid phase pH[15,34]. The ionic state of SO2 in solution at different pH is different, and the kinetics of its reaction with ozone is also different[15,34]. When the pH value changes by 1 unit, the reaction rate constant will change by 1 to 2 orders of magnitude[15]. The S(Ⅳ)+O3 pathway usually makes an important contribution to sulfate formation under alkaline conditions[35,15,36]. When pH > 6, the yield of S(Ⅳ)+O3 reaction pathway can reach several times of that of H2O2 oxidation pathway, which becomes the main pathway of sulfate formation in misty water[10]. Under the condition of sufficient NH3,The sulfate produced by the oxidation of SO2 by the mixture of O3 and NO2 is O3.And 2.0-3.5 times of the sum of sulfates generated by the single oxidation of the SO2 by NO2,The results show that the mixed oxidant has a strong synergistic effect on SO2 oxidation, which accelerates the formation of (NH4)2SO4 and NH4NO3, and is the reason for the explosive growth of PM2.5 in winter[37].

2.2.3 Active nitrogen

Reactive nitrogen species NO, NO2, and HNO2 are all possible oxidants of S (Ⅳ). The oxidation rate of S (Ⅳ) by NO is very slow, and NO2 and HONO are the main active nitrogen species oxidizing SO2 in the liquid phase[38]. At present, the HONO oxidation of SO2 is less studied, while the NO2 liquid-phase oxidation mechanism of S (Ⅳ) is still controversial. The low solubility of NO2 results in a slow rate of S(Ⅳ)+NO2 reaction in the liquid phase[8]. Recent studies have shown that the S(Ⅳ)+NO2 reaction in cloud/fog or aerosol liquid water is proved to be the key pathway for sulfate formation[39⇓~41,8,29,42]. However, this approach requires relatively high concentrations of NO2, aerosol liquid water content, and pH in the atmosphere. The aerosol liquid water content is typically three to five orders of magnitude lower than that of cloud/fog, but the NO2 reaction pathway still becomes important at such trace amounts of aerosol liquid water, primarily because of the high pH of the aerosol and the high concentration of NO2 in ambient air[43][39,44,8]. Alkaline substances are required to maintain oxidation because acidic aerosols can reduce the solubility of SO2 and the reaction rate of S(Ⅳ)+NO2, inhibiting the oxidation process in aerosols[45]. In areas with high NH3 emission, abundant NH3 can buffer the increase of aerosol acidity caused by sulfate formation and promote the NO2 oxidation of S (Ⅳ)[46,47,29]. Laboratory studies have shown that the reaction rate constant of S(Ⅳ)+NO2 increases with the increase of pH value of liquid phase[48]. The increase of aerosol pH value is beneficial to the increase of SO2 solubility and Henry's constant, so that more SO2 can be dissolved in aerosol liquid water, thus increasing the reaction rate[44]. The reaction rate of NO2 can be increased by an order of magnitude by increasing the pH value by one unit[8]. Some studies have shown that aerosol pH must be ~ 6 or higher for this reaction pathway to be significant[39,29,8]. However, Guo et al. And Liu et al. Considered that even in the case of high concentration of NH3, the pH value of aerosol could not be effectively raised to neutral, and the pH value was usually between 4.2 and 4.5, so the contribution of liquid phase oxidation of SO2 by NO2 to sulfate formation was very limited[49][50]. However, Wang et al. Considered that previous studies did not consider the influence of aerosol organic components, and the research method of using only inorganic components of particulate matter could not accurately infer the formation rate of aerosol acidity and sulfate[51]. For organic-based aerosols, the aerosol pH is sufficient to promote the efficient oxidation of SO2 by NO2 under sufficient NH3.
The mechanism by which NO2 oxidizes SO2 in aerosol particles is not fully understood. Studies have shown that more than 95% of NO2 dismutate on the surface of liquid sodium bisulfite (NaHSO3) (pH 3 ~ 6) to form HONO and HNO3, which is not the HSO 3 - of oxidative dissolution, and the oxidation of S (Ⅳ) is probably driven by HONO/ NO 2 -[41]. However, some scholars believe that in the case of sufficient NH3, NH3 enhances the atmospheric oxidation capacity by accelerating the formation of HONO, and the reaction between NO2 and SO2 in deliquescent aerosol particles exceeds the disproportionation reaction of NO2 itself[47].
Because the contribution of S(Ⅳ)+NO2 reaction pathway to sulfate production largely depends on the aerosol pH, the uncertainty of aerosol pH estimation results will affect the quantitative assessment of the contribution of different pathways to sulfate production. However, the high ionic strength of atmospheric aerosol particles makes the measurement of aerosol acidity difficult. At present, there are few studies on the direct determination of aerosol pH, mainly including Raman spectroscopy, probe detection and pH test paper[52⇓⇓~55]. However, direct measurement is labor-intensive and has poor time resolution. At present, indirect methods, such as ion balance method, molar ratio, phase partition and thermodynamic equilibrium model, are usually used to evaluate aerosol pH[56,57][58,59][60,61][8,49,62]. Methods such as ion balance and molar ratio can not effectively characterize aerosol pH because they ignore the influence of aerosol liquid water and partial dissolution of ions and acids on pH[60]. However, the application of phase partition method is limited in ammonia-poor atmospheric environment. Thermodynamic equilibrium models (such as ISORROPIA, E-AIM, etc.) Are commonly used to estimate aerosol pH, which is based on the concentration of water-soluble ions, humidity, temperature and other parameters. The uncertainty of model estimation is caused by many reasons, such as the use of models under non-equilibrium conditions, the existence of other phases (such as amorphous, crystalline or mixed States) of particulate matter besides liquid state, acid-base reactions involving organic acids and NH3/ amines, and the existence of organic components[51,63]. In addition, the kinetic experiments of liquid-phase oxidation of SO2 by NO2 used so far do not take into account the uptake process of the gas, which differs by more than two orders of magnitude[48,64]. Therefore, further development of aerosol acidity measurement technology and fine kinetic experimental study of SO2 liquid phase oxidation under atmospheric conditions are needed to clarify the formation mechanism of sulfate.

2.2.4 O2

The rate of direct reaction of S (Ⅳ) with O2 is very low, so the role of this reaction in sulfate formation is usually negligible[8]. Transition metal ions (TMI), such as Fe (Ⅲ) and Mn (Ⅱ), can significantly catalyze the oxidation of SO2 by O2 (abbreviated as TMI catalytic pathway)[65,66]. The TMI catalytic pathway in clouds has been considered to be the main pathway for sulfate formation[67,68]. However, recent observations based on stable sulfur isotopes in aerosols and model simulations suggest that this pathway dominates sulfate formation in aerosol liquid water during urban haze pollution events[65,66,69,70].
The rate of the TMI catalytic pathway is mainly affected by the dissolved Fe (Ⅲ) and Mn (Ⅱ) concentrations, the pH and ionic strength of the aerosol[71,8,10]. Higher dissolved concentrations of Fe (Ⅲ) and Mn (Ⅱ) can increase the reaction rate of TMI catalytic pathway by 4 ~ 5 orders of magnitude[71]. Model sensitivity analysis showed that the SO 4 2 - production rate of the TMI catalytic pathway was highly sensitive to aerosol pH, and the SO 4 2 - yield was higher at low pH[8]. After considering the ionic strength effect, the sulfate production rate of the TMI pathway may be significantly reduced under urban winter haze conditions. Liu et al. Studied the effect of ionic strength on TMI-catalyzed SO2 oxidation, and found that the rate of sulfate formation was inhibited by about 85 times at an ionic strength of 2.8 mol·kg-1 compared with dilute solution[33].

2.2.5 Hypohalous acid

Hypohalous acids HOX (where X = Cl, Br, and I) in marine aerosols are potentially important oxidants of S (Ⅳ). The reaction mainly proceeds via nucleophilic attack of the sulfur atom in HSO 3 - and/or SO 3 2 - on the halogen atom of HOX to form the XSO 3 - intermediate, followed by rapid hydrolysis of XSO 3 - to form $SO_{4}^{2-}$[72,73]. The oxidation of S (Ⅳ) by HOX in liquid phase is as follows:
H S O 3 - + H O X → X S O 3 - + H 2 O
S O 3 2 - + H O X → X S O 3 - + O H -
X S O 3 - + H 2 O → 2 H + + S O 4 2 - + X -
The reactivity of HOX followed the order HOI > HOBr > HOCl[38]. Assuming 1 ppt HOCl, the new kinetic parameters are applied to urban winter haze conditions, and HOCl has a potentially important role in sulfate formation. Considering that the concentration of HOCl can reach 300 ppt in eastern China, the heterogeneous oxidation of SO2 by HOCl in aerosol particles should be considered in the model study[74]. HOBr is less important than HOCl in the process of sulfate formation. This estimate may be subject to large uncertainties due to the lack of observations of HOBr concentrations[38].
Liquid-phase oxidation of SO2 by HOX in the marine boundary layer (MBL) is usually an important pathway for sulfate formation. Chen et al. Found that 33% to 50% of sulfate in MBL comes from this reaction through global model simulation[72]. In coastal cities, HOX liquid-phase oxidation of SO2 makes an important contribution to sulfate formation. Although the liquid phase oxygen of the HOX pair SO2 plays an important role in the formation of sulfate in MBL, the "S (Ⅳ) + HOX" reaction has not been included in most sulfur chemical models because of the low concentration of HOX in MBL and the large uncertainty of the reaction rate constants for HOX and HSO 3 - and the Henry constants for HOCl and HOBr.
To sum up, the contribution of different liquid-phase oxidation pathways of SO2 to sulfate formation is quite different, and its contribution mainly depends on factors such as oxidant concentration, aerosol liquid water content, acidity, and ionic strength. However, these factors vary greatly under different environmental conditions and affect the reaction rate of SO2, so it is impossible to accurately predict the sulfate formation rate in aerosol particles under all environmental conditions.

2.3 Heterogeneous oxidation

SO2 in the atmosphere can be heterogeneously oxidized on the surface of black carbon (BC), mineral dust, soot particles, and liquid aerosols[75][76⇓~78][79]. Zhang et al. Combined field observations, laboratory studies, and model simulations to determine the important role of BC-catalyzed SO2 oxidation in sulfate formation[75]. It has been shown that BC can effectively catalyze SO2 oxidation when NO2 and NH3 coexist, even at very low SO2 levels (as low as a few ppb) and at moderate relative humidity (30% – 70%). The mechanism is mainly the physical adsorption of NO2 on the BC surface, followed by the abstraction of allylic hydrogen by NO2 at the reducing surface site to form HONO, which oxidizes S (1V) to form sulfate.
The heterogeneous reaction of SO2 and NO2 on the surface of mineral dust is the current research focus. Mineral dust surface adsorbs or absorbs SO2 or sulfite, which is converted into sulfate under the action of NO2[76,78]. This process is affected by O 2 , NH3 and RH, and O2 is the main oxidant.NO2 and mineral oxides (e.g., Al2O3, ZnO, TiO2) act as catalysts to promote the conversion of SO2 to sulfate by the following reaction mechanism:[80,76][81][82,83][76]
2 N O 2 + S O 2 + M → M - S O 4 + 2 N O
2 N O + O 2 + M → 2 N O 2
M represents the mineral oxide surface. In addition, NH3 can promote the formation of sulfate on the acidic surface of α-Fe2O3 and α-Al2O3, while this effect is weaker on the surface of CaO and MgO, which may be affected by the alkaline characteristics of CaO and MgO[81]. The physical and chemical properties of substances used in laboratory studies may be completely different from those under ambient conditions, so the extrapolation of laboratory parameters to the atmospheric environment has great uncertainty.
Soot particles may also play an important role in the oxidation of SO2 by O2 to form sulfate[79,84]. He et al. Confirmed that soot particles can catalyze O2 to oxidize SO2 to form sulfate, which can also occur at night when photochemistry is weakened, and high RH can accelerate the reaction rate[85]. Water molecules on the surface of carbonaceous soot aerosol are decomposed into hydroxyl groups, which promotes the oxidation of SO2 to SO3 by O2[86]. This process is affected by the composition of the soot aerosol (e.g., the amount of aromatic C — H and oxygen-containing functional groups)[79]. However, some field observations and laboratory studies have shown that water molecules on the surface of carbonaceous soot promote the conversion of SO2 to sulfate under gas and liquid phase conditions[79,84]. Moreover, the catalytic oxidation pathway of H2O is not affected by aerosol pH, but mainly affected by environmental RH[86]. In the RH range of 6% ~ 70%, water molecules promote the adsorption of SO2 and the formation of sulfate; When the RH is higher than 80%, excessive water condensation on the soot inhibits the adsorption of SO2, which is not conducive to the formation of sulfate[79].
Recently, some studies have shown that the interfacial oxidation of SO2 on the surface of liquid aerosols, but not in aerosol liquid water, has an important contribution to sulfate formation[87⇓⇓~90,39]. In the process of haze pollution, the contribution of Mn catalytic pathway on aerosol surface to sulfate formation is as high as 92. 3% ± 3.5%, while the contribution of TMI catalytic pathway in aerosol liquid water to sulfate formation is negligible[91,88]. The reaction rate catalyzed by Mn on the aerosol surface is higher than that in the liquid phase mainly because (1) the high ionic strength of the aerosol promotes the ion-neutral molecule reaction on the surface, but inhibits the ion-ion reaction in the liquid phase; (2) The surface-to-volume ratio decreases with decreasing aerosol diameter, prompting interfacial chemistry to be more important than liquid-phase chemistry[88]. On the surface of acid aerosol droplets, SO2 can be uncatalyzed oxidized by O2, and the interfacial reaction rate is two orders of magnitude faster than that of H2O2 liquid phase oxidation[87,89,90][90]. Liu et al. Found that the reaction between SO2 and NO2 on the surface of deliquescent aerosol is mainly through the SO 3 2 - oxidation pathway, and the interfacial reaction rate is three orders of magnitude faster than that in solution, which is likely to be an important source of sulfate under haze pollution conditions[39].

2.4 Heterogeneous photochemical oxidation

Studies on heterogeneous photochemical oxidation of SO2 are mainly concerned with nitrate and ferric chloride photolysis in aerosol particles, photosensitization chemistry, and photochemical reactions on mineral dust surfaces[92⇓~94][95][96][97]. Particle-phase nitrates are photochemically active in the actinic range (e.g., λ > 290 nm) and can be photolyzed to highly reactive species such as NO2, OH, and N (Ⅲ) ( NO 2 -/HONO)[98,99]. Dissolved SO2 can be oxidized by N (Ⅲ) and OH to form $SO_{4}^{2-}$[41]. Nitrate photolysis rate constant is affected by particle composition, solar radiation intensity and other factors[98,100]. Nitrate photolysis can be enhanced in the presence of halide ions, such as in the presence of Cl-, Br- and I-, the rate constant of nitrate photolysis can be increased by 2.0, 1.7 and 3.7 times, respectively[101⇓~103][92]. Gen et al. Confirmed the importance of nitrate photolysis in aerosol particles to SO2 oxidation through experimental research, and the research results showed that this pathway contributed up to 60% to sulfate formation under the condition of pH 4 ~ 5 in haze events[94]. Zheng et al. added the nitrate photolysis reaction mechanism to the air quality model and evaluated the importance of this pathway to sulfate formation, and found that the importance of this mechanism mainly depends on the enhancement coefficient (1 to 3 orders of magnitude) of the nitrate photolysis rate constant in aerosol liquid water compared with that in the gas phase[93]. This pathway can explain 15% to 65% of the difference between simulated and observed sulfate concentrations during urban winter haze. In addition to nitrate photolysis, the photolysis of ferric chloride contained in the aerosol produces chlorine radicals (Cl*/ Cl 2 * -), which can oxidize SO2 to sulfate[95].
Photosensitive chemistry is a new research field of aerosol chemistry[38,104]. The photosensitized oxidation of S (Ⅳ) is mainly due to the fact that the photosensitizer molecule is excited to the triplet state by solar radiation, and then the SO2·H2O and bisulfite are oxidized to sulfate by energy conversion, electron transfer or hydrogen atom abstraction[96]. Experimental studies have shown that photosensitizers (e.g., brown carbon, humic substances) can promote the conversion of SO2 to sulfate[105,96]. Wang et al. Determined the reaction rate constants of S (Ⅳ) and photosensitizer molecules in different triplet excited States by pulsed laser experiments[96]. The rate constants vary widely depending on the type of photosensitizer, ranging from 6.9×107M-1·s-1 (reaction with 4-benzoylbenzoic acid photosensitizer) to 1.0×109M-1·s-1 (reaction with xanthone photosensitizer).
The mechanism of heterogeneous photochemical oxidation of SO2 on the surface of mineral dust is mainly based on the electron-hole pair theory. When a photon is absorbed by the metal oxide in the dust particle, the metal oxide electron is activated and enters the conduction band from the valence band to generate an electron-hole pair, which generates OH, hydrogen peroxide radical (HO2) and other radicals, and then reacts with SO2 to form sulfate[106⇓⇓~109]. The uptake coefficient ( γ S O 2) of SO2 on the surface of mineral dust is affected by meteorological conditions (humidity, temperature, and light intensity), air pollutants (NOx and O3), and mineral dust components[106,108,109]. For example, the surface γ S O 2 of Arizona Test Dust (ATD) is one order of magnitude higher under UV illumination than in the darkroom[106]. The value of γ S O 2 increases with increasing RH or when O3 and NO2 coexist, but the value of γ S O 2 on the surface of TiO2 decreases with increasing RH under UV irradiation[97,107][109].

3 Mechanism of nitrate formation

3.1 HNO3 generation

The formation of NO 3 - involves complex heterogeneous chemical reactions, which are mainly controlled by two processes, namely, the formation of HNO3 and the gas-particle partitioning of HNO3- NO 3 -. The generation of HNO3 mainly includes gas phase oxidation and heterogeneous reaction, as shown in Table 1. In gas phase oxidation, NO2 reacts with OH radicals to form gaseous HNO3(HNO3(g)), which is the main pathway for daytime HNO3 formation[15,110]. NO2 can also react with O3 to form NO3 radical, but NO3 is easily photolyzed and is reduced back to NO2 at high NO concentration[111]. Usually, NO3 reacts with hydrocarbons (HC) at night to produce HNO3(g)[112,113]. The results of the model study indicate that the contribution of the NO3+HC pathway to nitrate production is small (< 10%)[114⇓~116]. However, in some areas where VOCs are emitted in large quantities, the HC concentration in the atmosphere is high, which makes the contribution of NO3+HC reaction to NO 3 - increase. The isotope observation results of Wang et al. In Beijing area show that the contribution of NO3+HC pathway to nitrate is 34%, which may be related to the high VOCs emissions in some areas of Beijing[117]. In addition, NO can also react directly with HO2 or peroxyalkyl radical (RO2) to produce HNO3(g)[118].
表1 HNO3生成的主要化学反应

Table 1 The main reactions contributing to HNO3 formation

Type Reactions No.
Gas-phase
reactions		 N O 2 + O H → H N O 3 R1
N O 3 + H C → H N O 3 + p r o d u c t s R2
N O + H O 2 / R O 2 → H N O 3 ( g ) R3
Heterogeneous
reactions		 N O 2 + N O 3 → N 2 O 5 N 2 O 5 + H 2 O → 2 H N O 3 ( a q ) R4
N O 2 + N O 3 → N 2 O 5 N 2 O 5 + C l - → C l N O 2 + N O 3 - ( a q ) R5
2 N O 2 + H 2 O → H N O 3 + H O N O R6
N O 2 + X O → X N O 3 , X N O 3 + H 2 O → H N O 3 ( a q ) R7
N O 3 + H 2 O → H N O 3 ( a q ) R8
N O 3 + M T N / I S O P → R O N O 2 R O N O 2 + H 2 O → H N O 3 ( a q ) R9
N O + R O 2 → R O N O 2 R O N O 2 + H 2 O → H N O 3 ( a q ) R10
In the heterogeneous reaction, N2O5 hydrolysis is the main pathway for HNO3 generation at night[119,120]. N2O5 is produced by the reaction of NO3 with NO2, and NO3 is susceptible to photolysis, so N2O5 hydrolysis is usually more important at night. In addition, the reaction usually occurs on the surface of wet aerosol or in cloud droplets, and is slower in the gas phase[121]. The efficiency of heterogeneous hydrolysis of N2O5 to liquid HNO3(HNO3(aq)) is determined by factors such as the N2O5 uptake coefficient (γ(N2O5)), insolation and solar radiation[122]. The γ(N2O5) is mainly affected by the liquid water content, nitrate, chloride and organic matter concentration of aerosol[123,124]. γ(N2O5) is significantly positively correlated with aerosol liquid water content or relative humidity, so the hydrolysis rate of N2O5 on deliquescent or supersaturated aerosols is much higher than that on solid aerosols[14,125][125]. Higher NO 3 - content in aerosols can reverse the ionization process of N2O5, reduce the dissolution of N2O5, and allow N2O5 to diffuse out of aerosols (known as "nitrate inhibition" effect)[126]. In contrast, the presence of Cl- in the aerosol enhances N2O5 uptake because Cl- reacts with NO 2 + (R14) to form nitryl chloride (ClNO2), which shifts the equilibrium reaction (R12) to the right to counteract nitrate inhibition[127]. The presence of organics in aerosols can also hinder N2O5 uptake, mainly because organics reduce ALWC and/or limit aerosol surface activity[128]. In general, the γ(N2O5) vary greatly with the change of geographical and environmental factors and have great uncertainty, which has certain limitations in reflecting the real formation process of local atmospheric nitrate. Field observation and model simulation studies have found that parameterization schemes of γ(N2O5) based on experimental results usually overestimate the γ(N2O5) in the real atmosphere[129,14,130]. Yu et al. Modified the parameterization scheme of γ(N2O5) based on the field observation results of four stations under different atmospheric conditions in China, and found that the modified parameterization scheme could not only better predict the γ(N2O5) in the actual atmosphere, but also improve the simulation results of N2O5 and nitrate in the WRF-CMAQ model after introducing the parameter scheme into the model[131].
N 2 O 5 ( g ) ↔ N 2 O 5 ( a q )
N 2 O 5 ( a q ) ↔ N O 3 - ( a q ) + N O 2 + ( a q )
N O 2 + ( a q ) + H 2 O ( l ) → N O 3 - ( a q ) + 2 H + ( a q )
N O 2 + ( a q ) + C l - ( a q ) → C l N O 2 ( a q )
Heterogeneous hydrolysis of NO2 (R6) is also one of the important pathways for HNO3(aq) generation. The global scale simulation results show that the heterogeneous hydrolysis of NO2 contributes about 12.0% to the formation of NO 3 -, which is the third largest nitrate formation pathway after the gas phase reaction of NO2+OH and the heterogeneous hydrolysis of N2O5[114]. This reaction simultaneously produces HONO, which, as an important source of OH, can affect the concentration of OH in the atmosphere[132]. NO2 can also react with halogen oxides (XO) to form halogen nitrates (e.g., BrNO3, ClNO3, and INO3), which are then hydrolyzed to form HNO3(aq)[133]. The contribution of heterogeneous hydrolysis of NO3 to NO 3 - generation is small and negligible[134,135]. NO3 can also react with monoterpenes (MTN) and isoprene (ISOP) to form organic nitrates (RONO2),RONO2 can also be formed by the reaction of NO with RO2, and RONO2 is hydrolyzed to form HNO3(aq)[118]. While the contribution of these reaction pathways to NO 3 - generation is generally small.
The chemical conversion of NOx to HNO3 is highly correlated with atmospheric oxidizing capacity (AOC)[136]. Feng et al. Found that the increase of NO 3 - in recent years was mainly caused by the enhancement of AOC through WRF-Chem model[137]. OH and O3 are not only the key indicators of AOC, but also the important photochemical oxidants produced by HNO3 during the day and night[138,139]. The increase of photochemical oxidants (OH and O3) in winter in North China leads to persistent and serious nitrate pollution[135]. The high OH oxidation rate observed in the Beijing winter haze event indicates that photochemistry effectively increased NO 3 - generation[140]. Air quality improvement strategies aimed at reducing PM2.5 by reducing NOx emissions in recent years may lead to an increase in O3, which in turn promotes NO 3 - generation[141,135].
Since the formation of HNO3 depends on the level of OH and O3 photochemical oxidants, other processes that have some influence on atmospheric oxidation capacity, such as chlorine chemistry, HONO chemistry, and VOCs oxidation, can also affect the generation of NO 3 -[142⇓~144]. Wang et al. Studied the effect of anthropogenic chlorine emission on secondary inorganic aerosols, and found that anthropogenic chlorine emission led to a decrease in the annual average concentration of NO 3 - by 1.5μg·m-3, and had a negligible effect on the concentration of sulfate[145]. The main reason for the decrease of NO 3 - is that more N2O5 reacts with Cl- instead of with H2O.An increase in the ClNO2 of the product reduces the yield of the NO 3 -[144,146,147]. In addition, NO2 and chlorine atoms (Cl) are produced by photolysis of ClNO2 during the day[148]. Cl is a more active free radical than OH, which can affect the formation of O3 and OH, enhance atmospheric oxidation, and promote the oxidation of NO2 to NO 3 -[143,149,150]. Considering that ClNO2 is not only an important product of the heterogeneous reaction of N2O5, but also its photolytic products affect daytime photochemistry, the yield of ClNO2 (φ(ClNO2)) has an important effect on the formation of nitrate. φ(ClNO2) is affected by the concentration of chloride ion in aerosol, liquid water content and organic matter, and its value can not be directly measured by field observation at present, mainly through the measurement based on ClNO2 and other species, and then obtained by calculation or model[151,129][152]. At present, the parameterization method of φ(ClNO2) has greater uncertainty. Field observation studies have shown that the most commonly used parameterization schemes generally overestimate the φ(ClNO2)[129,151]. Wang et al. Proposed a new parameterization scheme to consider the effect of organic matter on φ(ClNO2), but the actual effect of the scheme remains to be verified[152]. Therefore, φ(ClNO2) is still a key problem to be solved. HONO photolysis is one of the main sources of OH, and its rapid photolysis makes the OH level in the atmosphere relatively high, which accelerates the NO2+OH gas-phase reaction and promotes the rapid formation of NO 3 -[153][142,154]. VOCs also affect the formation of NO 3 - due to their potential impact on AOC. VOCs are photolyzed or oxidized by Cl radicals to produce HO2 or RO2 radicals, which later drive the production of OH, resulting in increased HNO3 production[155,149]. Therefore, reducing VOCs emissions is beneficial to reducing the level of oxidants in the atmosphere and prolonging the life of NOx. The model sensitivity study shows that reducing VOCs emission by 30% – 50% can reduce NO 3 - in the North China Plain by about 10% in winter[156,135].
NOx is an important precursor of HNO3, and in most cases, reducing NOx emissions can reduce the concentration of NO 3 -[156⇓~158]. A 50% reduction in NOx emissions can reduce the NO 3 - concentration by 10.3 – 17.9%[158,156]. However, field observation and numerical modeling studies show that reducing NOx emissions may lead to an increase in NO 3 -[135][159,160]. Some scholars believe that it is mainly due to the reduction of NOx emissions and the increase of O3 caused by weak NOx titration, which is conducive to the formation of N2O5 and thus promotes the formation of nitrate[158]. Shah et al. Simulated the effects of simultaneous reduction of SO2 and NO2 emissions on sulfate and nitrate production in winter in the eastern United States, and found that sulfate production decreased, while nitrate production was almost unchanged, which was mainly related to the reduction of fine particulate matter acidity[160]. The above studies indicate a non-linear response relationship between the generation of NO 3 - and the amount of NOx emissions.

3.2 HNO3-NO3- gas-particle distribution

The HNO3 produced by the reaction can react with ammonia, dust and sea salt (NaCl) aerosol to form nitrate. High concentration of alkaline gas, high humidity and low temperature are beneficial to the transformation of HNO3 into NO 3 -[135]. NH3 play a key role in neutralizing HNO3 and generating NO 3 -. The NH4NO3 produced by the reaction between HNO3 and NH3 is the main form of particulate nitrate in urban atmosphere. NH4NO3 is semi-volatile, and its formation is affected by temperature, RH and aerosol acidity[15].
H N O 3 g + N H 3 g ↔ N H 4 N O 3 s
H N O 3 ( g ) + N H 3 ( g ) ↔ N H 4 + ( a q ) + N O 3 - ( a q )
RH determines whether the NH4NO3 exists in solid or liquid form, and the temperature determines the equilibrium constant (Kp) for this reversible reaction.
l n K p = 84.6 - ( 24220 / T ) - 6.1 l n ( T / 298 )
Studies have shown that reducing NH3 emissions is an effective way to mitigate NO 3 - and PM2.5 pollution[156,161⇓~163]. The model simulation results show that 20% – 50% emission reduction of NH3 can reduce the NO 3 - in the atmosphere by 6.85% – 47.0% in some areas of China[156,157,164]. It has also been shown that the conversion of NH3 to ammonium salt ( NH 4 +) under high concentration of NH3 is mainly driven by meteorological factors such as temperature and relative humidity. Conditions such as temperature below 10 ℃ and high RH of more than 70% are the best conditions for the conversion of gas-phase NH3 to ammonium salt, while under warm and dry conditions, ammonium salt is converted to gas-phase N H 3 again[165,166][167].
Calcium carbonate (CaCO3) is the most studied component of mineral dust[168,169]. The heterogeneous reaction between gaseous HNO3 and calcium carbonate dust produces soluble Ca(NO3)2, which is the main source of nitrate in dust weather[169,170]. The reaction is as follows:
C a C O 3 ( s ) + 2 H N O 3 → C a ( N O 3 ) 2 ( a q ) + C O 2 ( g ) + H 2 O
The reaction rate increases with RH[171]. Many laboratory studies and field observations have confirmed the importance of heterogeneous reactions between HNO3 and dust. For example, the conversion of calcium-rich dust particles into liquid phase is mainly due to the heterogeneous reaction between CaCO3 and HNO3(g), which produces hygroscopic Ca(NO3 ) 2 [170,172,173][172,170]. The acidity of urban fine particulate matter during the spring dust period in Xi'an, China, in 2009 was significantly higher than that during the non-dust period, which was also due to the rapid production of nitrate on the surface of dust[174,175].
In the atmosphere of coastal cities, the heterogeneous reaction between HNO3 and sea salt aerosol produces NaNO3 and HCl, and the HCl released by the reaction is one of the main sources of HCl in the troposphere[176][10]. The heterogeneous reaction of NaCl to NaNO3 particle conversion is as follows:
H N O 3 ( g ) + N a C l ( s o r a q ) → N a N O 3 ( s o r a q ) + H C l ( g )
The uptake coefficient of HNO3(g) on NaCl aerosol is affected by RH. Saul et al. Studied the uptake coefficient of gaseous HNO3 on NaCl aerosol in the range of 10% – 85% RH. When RH is higher than the relative humidity of NaCl deliquescence (DRH)/relative humidity of weathering (ERH) and decreases toward ERH, NaCl particles exist in liquid form, and the uptake coefficient of HNO3 increases from 0.05 at 85% RH to > 0.1 near ERH[177]. When RH is less than ERH, the NaCl particle is solid, and the uptake coefficient of HNO3 depends on the amount of water adsorbed on the surface of the NaCl particle. The reaction between HNO3 and sea salt aerosols leads to the coexistence of NaNO3 and NaCl, which changes the surface and internal composition of particles, and then changes their optical properties and cloud formation characteristics[178,179]. In addition, compared with pure NaCl particles, the NaCl-NaNO3 mixed aerosol can maintain liquid state in a larger RH range, which is beneficial to the occurrence of heterogeneous reactions[178]. For example, in the marine boundary layer, the RH is usually greater than 60%, and the N2O5 uptake coefficient on NaCl particles with a higher proportion of Cl-/ NO 3 - is higher, which promotes the heterogeneous reaction of N2O5[180].

4 Conclusion and prospect

At present, there are still many uncertainties in the chemical mechanism of sulfate and nitrate formation, which reflect that the understanding of the relevant chemical mechanism is not deep enough, and there is a lack of field observation or laboratory research on some key species or parameters. However, how to integrate the results of field observations and laboratory studies to establish the corresponding parameterization scheme remains to be further studied.
The following questions remain regarding several important pathways for conversion from SO2 to sulfate: The H2O2 oxidation SO2 pathway in aerosol particles is relatively important, but the H2O2 source is unclear. The reaction rates of O3 and NO2 oxidation pathways are highly dependent on aerosol pH, but it is still uncertain whether sufficient NH3 in the atmospheric environment can effectively change aerosol pH. The mechanism of liquid-phase oxidative SO2 of NO2 in aerosols is still unclear, and whether NO2 molecules form nitrite and nitrate through disproportionation or directly react with S (Ⅳ) remains to be tested. Experimental studies have shown that photochemical oxidation of SO2 is a new important pathway for sulfate formation. However, due to the large differences in the light absorption properties of molecules in laboratory photosensitizers and environmental aerosols, there is still a large uncertainty in applying the experimental results to atmospheric environmental conditions.
In order to improve the understanding of rapid sulfate formation during urban haze pollution, more laboratory and observational studies on the fine kinetics and mechanisms of heterogeneous chemistry are urgently needed in the future. Recommendations for future research are as follows:
(1) Determination of aerosol pH. The SO2 solubility and Heinz's law constants as well as many of the SO2 liquid-phase oxidation pathways involving sulfate formation, including O3, O2+TMI, and NO2, are strongly dependent on aerosol pH. When the pH value changes by 1 unit, the sulfate formation rate will change by 1 to 2 orders of magnitude. Aerosol pH direct measurement technology has not been widely used because of its poor time resolution and labor consumption. However, the further development of aerosol pH measurement technology is needed to clarify the formation mechanism of sulfate.
(2) to establish ionic strength-dependent kinetics. The effect of high ionic strength in aerosol liquid water on the reaction kinetics should be considered in the SO2 heterogeneous chemical model. Aerosol ionic strength can significantly affect the liquid phase oxidation rate of SO2, for example, high ionic strength can increase the oxidation rate of SO2 catalyzed by H2O2, but inhibit the oxidation rate catalyzed by TMI. At present, the yield calculation of sulfate reaction pathway is mainly based on the experimental data of low ionic strength solution, but the ionic strength of aerosol liquid water during haze is significantly higher than that of laboratory solution environment, so there is some uncertainty in the yield calculation based on the relevant parameters obtained from the current experimental research. The effect of high ionic strength of aerosol liquid water on SO2 oxidation needs to be further investigated in the future.
(3) To study the influence of particulate matter phase. The phase state of particulate matter can affect the heterogeneous chemical reaction process of SO2. At present, the observation of aerosol phase is relatively scarce, and the phase observation of particulate matter in different regions and environments needs to be carried out in the future.
(4) To carry out fine kinetics and mechanism studies under environment-matched conditions. Given that most laboratory model systems have not yet adequately matched the chemistry and speciation of aerosols under atmospheric environmental conditions, it is difficult to accurately describe the rate of sulfate formation in a polluted environment by finally identifying any one or set of reaction mechanisms. Therefore, laboratory research should be carried out under conditions matching the environment as far as possible, and the ionic strength, phase state and pH value of aerosols should be strictly limited.
As NO 3 - is gradually becoming the main component of PM2.5 in many regions of China, reducing NO 3 - has become a necessary step to reduce PM2.5 pollution. In the future, more laboratory studies and observation experiments should be carried out to deepen the understanding of the mechanism of nitrate formation. The main suggestions are as follows:
(1) Measurement of N2O5 and NO2 uptake coefficients in different environments. Influenced by different factors such as aerosol chemical composition, surface area and liquid water, the uptake coefficients of N2O5 and NO2 are different in different regions and have great uncertainty, which is also a key factor affecting the accuracy of nitrate simulation in numerical models. Therefore, extensive observations of γ(N2O5) and γ(NO2) in different environments need to be carried out in the future, and the relevant reaction mechanisms should be better parameterized into the model to improve the simulation of nitrate formation process.
(2) to quantify the level of AOC and its effect on NO 3 -. AOC plays a key role in the generation of NO 3 -, but the level and source of AOC in different regions under different meteorological conditions are still unclear. In the future, the level and source of AOC should be further studied to quantify the effect of AOC on the formation of NO 3 -.
(3) to quantify the impact of VOCs, NOx, and NH3 emissions on NO 3 -. The formation of NO 3 - is influenced not only by the precursor NOx, but also by the concentrations of AOC, O3 and NH3. AOC levels have a highly nonlinear relationship with NOx and VOCs. In order to reduce NO 3 - pollution, VOCs, NOx and NH3 should be coordinated to reduce emissions. Therefore, more studies are needed in the future to reveal the relationship between NO 3 - and NOx, VOCs, and NH3.

References

Publishing order | Descend order by publishing year | Descend order by cited within
[1]
Zhang Q, Zheng Y X, Tong D, Shao M, Wang S X, Zhang Y H, Xu X D, Wang J N, He H, Liu W Q, Ding Y H, Lei Y, Li J H, Wang Z F, Zhang X Y, Wang Y S, Cheng J, Liu Y, Shi Q R, Yan L, Geng G N, Hong C P, Li M, Liu F, Zheng B, Cao J J, Ding A J, Gao J, Fu Q Y, Huo J T, Liu B X, Liu Z R, Yang F M, He K B, Hao J M. Proc. Natl. Acad. Sci. U. S. A., 2019, 116(49): 24463.

[2]
Fuzzi S, Baltensperger U, Carslaw K, Decesari S, Denier Van Der Gon H, Facchini M C, Fowler D, Koren I, Langford B, Lohmann U, Nemitz E, Pandis S, Riipinen I, Rudich Y, Schaap M, Slowik J G, Spracklen D V, Vignati E, Wild M, Williams M, Gilardoni S. Atmos. Chem. Phys., 2015, 15(14): 8217.

[3]
Wang Y, Zhang R Y, Saravanan R. Nat. Commun., 2014, 5: 3098.

[4]
Zheng B, Zhang Q, Zhang Y, He K B, Wang K, Zheng G J, Duan F K, Ma Y L, Kimoto T. Atmos. Chem. Phys., 2015, 15(4): 2031.

[5]
Wang J D, Zhao B, Wang S X, Yang F M, Xing J, Morawska L, Ding A J, Kulmala M, Kerminen V M, Kujansuu J, Wang Z F, Ding D, Zhang X Y, Wang H B, Tian M, Petäjä T, Jiang J K, Hao J M. Sci. Total Environ., 2017, 584/585: 426.

[6]
Wen L, Xue L K, Wang X F, Xu C H, Chen T S, Yang L X, Wang T, Zhang Q Z, Wang W X. Atmos. Chem. Phys., 2018, 18(15): 11261.

[7]
Li H Y, Zhang Q, Zheng B, Chen C R, Wu N N, Guo H Y, Zhang Y X, Zheng Y X, Li X, He K B. Atmos. Chem. Phys., 2018, 18(8): 5293.

[8]
Cheng Y F, Zheng G J, Wei C, Mu Q, Zheng B, Wang Z B, Gao M, Zhang Q, He K B, Carmichael G, Pöschl U, Su H. Sci. Adv., 2016, 2(12): e1601530.

[9]
Quan J N, Liu Q, Li X, Gao Y, Jia X C, Sheng J J, Liu Y G. Atmos. Environ., 2015, 122: 306.

[10]
Zhang R Y, Wang G H, Guo S, Zamora M L, Ying Q, Lin Y, Wang W G, Hu M, Wang Y. Chem. Rev., 2015, 115(10): 3803.

[11]
Xie X D, Hu J L, Qin M M, Guo S, Hu M, Wang H L, Lou S R, Li J Y, Sun J J, Li X, Sheng L, Zhu J L, Chen G Y, Yin J J, Fu W X, Huang C, Zhang Y H. Environ. Int., 2022, 166: 107369.

[12]
Zang H, Zhao Y, Huo J T, Zhao Q B, Fu Q Y, Duan Y S, Shao J Y, Huang C, An J Y, Xue L K, Li Z Y, Li C X, Xiao H Y. Atmos. Chem. Phys., 2022, 22(7): 4355.

[13]
Yang J R, Wang S B, Zhang R Q, Yin S S. Environ. Pollut., 2022, 296: 118716.

[14]
Tham Y J, Wang Z, Li Q Y, Wang W H, Wang X F, Lu K D, Ma N, Yan C, Kecorius S, Wiedensohler A, Zhang Y H, Wang T. Atmos. Chem. Phys., 2018, 18(17): 13155.

[15]
Seinfeld J H, Pandis S N. Atmospheric Chemistry and Physics: from Air Pollution to Climate Change. 3rd. ed. Hoboken, New Jersey: John Wiley & Sons, 2016.

[16]
Zhang D, Zhang R Y. Environ. Sci. Technol., 2005, 39(15): 5722.

[17]
Rodhe H, Crutzen P, Vanderpol A. Tellus, 1981, 33(2): 132.

[18]
Faloona I. Atmos. Environ., 2009, 43(18): 2841.

[19]
Zheng G J, Duan F K, Su H, Ma Y L, Cheng Y, Zheng B, Zhang Q, Huang T, Kimoto T, Chang D, Pöschl U, Cheng Y F, He K B. Atmos. Chem. Phys., 2015, 15(6): 2969.

[20]
Sarwar G, Simon H, Fahey K, Mathur R, Goliff W S, Stockwell W R. Atmos. Environ., 2014, 85: 204.

[21]
Sarwar G, Fahey K, Kwok R, Gilliam R C, Roselle S J, Mathur R, Xue J, Yu J Z, Carter W P L. Atmos. Environ., 2013, 68: 186.

[22]
Mauldin R L, Berndt T, Sipilä M, Paasonen P, Petäjä T, Kim S, KurtÉn T, Stratmann F, Kerminen V M, Kulmala M. Nature, 2012, 488(7410): 193.

[23]
Welz O, Savee J D, Osborn D L, Vasu S S, Percival C J, Shallcross D E, Taatjes C A. Science, 2012, 335(6065): 204.

[24]
Boy M, Mogensen D, Smolander S, Zhou L, Nieminen T, Paasonen P, Plass-Dülmer C, Sipilä M, Petäjä T, Mauldin L, Berresheim H, Kulmala M. Atmos. Chem. Phys., 2013, 13(7): 3865.

[25]
Liu L, Bei N F, Wu J R, Liu S X, Zhou J M, Li X, Yang Q C, Feng T, Cao J J, Tie X X, Li G H. Atmos. Chem. Phys., 2019, 19(21): 13341.

[26]
Huang H L, Chao W, Lin J J M. Proc. Natl. Acad. Sci. U. S. A., 2015, 112(35): 10857.

[27]
Sipilä M, Jokinen T, Berndt T, Richters S, Makkonen R, Donahue N M, Mauldin R L, KurtÉn T, Paasonen P, Sarnela N, Ehn M, Junninen H, Rissanen M P, Thornton J, Stratmann F, Herrmann H, Worsnop D R, Kulmala M, Kerminen V M, Petäjä T. Atmos. Chem. Phys., 2014, 14(22): 12143.

[28]
Wang Y X, Zhang Q Q, Jiang J K, Zhou W, Wang B Y, He K B, Duan F K, Zhang Q, Philip S, Xie Y Y. J. Geophys. Res. Atmos., 2014, 119(17): 10425.

[29]
Wang G H, Zhang R Y, Gomez M E, Yang L X, Levy Zamora M, Hu M, Lin Y, Peng J F, Guo S, Meng J J, Li J J, Cheng C L, Hu T F, Ren Y Q, Wang Y S, Gao J, Cao J J, An Z S, Zhou W J, Li G H, Wang J Y, Tian P F, Marrero-Ortiz W, Secrest J, Du Z F, Zheng J, Shang D J, Zeng L M, Shao M, Wang W G, Huang Y, Wang Y, Zhu Y J, Li Y X, Hu J X, Pan B W, Cai L, Cheng Y T, Ji Y M, Zhang F, Rosenfeld D, Liss P S, Duce R A, Kolb C E, Molina M J. Proc. Natl. Acad. Sci. U. S. A., 2016, 113(48): 13630.

[30]
Weber R J, Guo H Y, Russell A G, Nenes A. Nat. Geosci., 2016, 9(4): 282.

[31]
Maaß F, Elias H, Wannowius K J. Atmos. Environ., 1999, 33(27): 4413.

[32]
Ye C, Chen H, Hoffmann E H, Mettke P, Tilgner A, He L, Mutzel A, Brüggemann M, Poulain L, Schaefer T, Heinold B, Ma Z B, Liu P F, Xue C Y, Zhao X X, Zhang C L, Zhang F, Sun H, Li Q, Wang L, Yang X, Wang J H, Liu C, Xing C Z, Mu Y J, Chen J M, Herrmann H. Environ. Sci. Technol., 2021, 55(12): 7818.

[33]
Liu T Y, Clegg S L, Abbatt J P D. Proc. Natl. Acad. Sci. U. S. A., 2020, 117(3): 1354.

[34]
Kahan T F, Ardura D, Donaldson D J. J. Phys. Chem. A, 2010, 114(5): 2164.

[35]
Fang Y H, Ye C X, Wang J X, Wu Y S, Hu M, Lin W L, Xu F F, Zhu T. Atmos. Chem. Phys., 2019, 19(19): 12295.

[36]
Sievering H. J. Geophys. Res., 2004, 109(D19): D19317.

[37]
Zhang S, Li D P, Ge S S, Liu S J, Wu C, Wang Y Q, Chen Y B, Lv S J, Wang F L, Meng J J, Wang G H. Sci. Total Environ., 2021, 772: 144897.

[38]
Liu T Y, Chan A W H, Abbatt J P D. Environ. Sci. Technol., 2021, 55(8): 4227.

[39]
Liu T Y, Abbatt J P D. Nat. Chem., 2021, 13(12): 1173.

[40]
Xue J, Yu X, Yuan Z B, Griffith S M, Lau A K H, Seinfeld J H, Yu J Z. Nat. Geosci., 2019, 12(12): 977.

[41]
Li L J, Hoffmann M R, Colussi A J. Environ. Sci. Technol., 2018, 52(5): 2686.

[42]
Guo S, Hu M, Zamora M L, Peng J F, Shang D J, Zheng J, Du Z F, Wu Z J, Shao M, Zeng L M, Molina M J, Zhang R Y. Proc. Natl. Acad. Sci. U. S. A., 2014, 111(49): 17373.

[43]
Herrmann H, Schaefer T, Tilgner A, Styler S A, Weller C, Teich M, Otto T. Chem. Rev., 2015, 115(10): 4259.

[44]
Chen T Z, Chu B W, Ge Y L, Zhang S P, Ma Q X, He H, Li S M. Environ. Pollut., 2019, 252: 236.

[45]
Turši? J, Berner A, Podkrajšek B, Grgić I. Atmos. Environ., 2004, 38(18): 2789.

[46]
Zheng G J, Su H, Wang S W, Andreae M O, Pöschl U, Cheng Y F. Science, 2020, 369(6509): 1374.

[47]
Ge S S, Wang G H, Zhang S, Li D P, Xie Y N, Wu C, Yuan Q, Chen J M, Zhang H L. Environ. Sci. Technol., 2019, 53(24): 14339.

[48]
Clifton C L, Altstein N, Huie R E. Environ. Sci. Technol., 1988, 22(5): 586.

[49]
Guo H Y, Weber R J, Nenes A. Sci. Rep., 2017, 7: 12109.

[50]
Liu M X, Song Y, Zhou T, Xu Z Y, Yan C Q, Zheng M, Wu Z J, Hu M, Wu Y S, Zhu T. Geophys. Res. Lett., 2017, 44(10): 5213.

[51]
Wang G H, Zhang F, Peng J F, Duan L, Ji Y M, Marrero-Ortiz W, Wang J Y, Li J J, Wu C, Cao C, Wang Y, Zheng J, Secrest J, Li Y X, Wang Y Y, Li H, Li N, Zhang R Y. Atmos. Chem. Phys., 2018, 18(14): 10123.

[52]
Rindelaub J D, Craig R L, Nandy L, Bondy A L, Dutcher C S, Shepson P B, Ault A P. J. Phys. Chem. A, 2016, 120(6): 911.

[53]
Craig R L, Nandy L, Axson J L, Dutcher C S, Ault A P. J. Phys. Chem. A, 2017, 121(30): 5690.

[54]
Wei H R, Vejerano E P, Leng W N, Huang Q S, Willner M R, Marr L C, Vikesland P J. Proc. Natl. Acad. Sci. U. S. A., 2018, 115(28): 7272.

[55]
Song Q P, Osada K. Atmos. Environ., 2021, 261: 118605.

[56]
Kerminen V M, Hillamo R, Teinilä K, Pakkanen T, Allegrini I, Sparapani R. Atmos. Environ., 2001, 35(31): 5255.

[57]
Metzger S, Mihalopoulos N, Lelieveld J. Atmos. Chem. Phys., 2006, 6(9): 2549.

[58]
Shi G L, Xu J, Peng X, Xiao Z M, Chen K, Tian Y Z, Guan X B, Feng Y C, Yu H F, Nenes A, Russell A G. Environ. Sci. Technol., 2017, 51(8): 4289.

[59]
Zhang T, Cao J J, Tie X X, Shen Z X, Liu S X, Ding H, Han Y M, Wang G H, Ho K F, Qiang J, Li W T. Atmos. Res., 2011, 102(1/2): 110.

[60]
Hennigan C J, Izumi J, Sullivan A P, Weber R J, Nenes A. Atmos. Chem. Phys., 2015, 15(5): 2775.

[61]
Trebs I, Metzger S, Meixner F X, Helas G, Hoffer A, Rudich Y, Falkovich A H, Moura M A L, Artaxo P, Slanina J, Andreae M O. J. Geophys. Res., 2005, 110(D7): D07303.

[62]
Song S J, Gao M, Xu W Q, Shao J Y, Shi G L, Wang S X, Wang Y X, Sun Y L, McElroy M B. Atmos. Chem. Phys., 2018, 18(10): 7423.

[63]
Qiu C, Zhang R Y. Phys. Chem. Chem. Phys., 2013, 15(16): 5738.

[64]
Lee Y N, Schwartz S E. Fourth International Conference on Precipitation Scavenging, Dry Deposition, and Resuspension, Santa Monica, California, 1982.

[65]
Angle K J, Neal E E, Grassian V H. Environ. Sci. Technol., 2021, 55(15): 10291.

[66]
Tao W, Su H, Zheng G J, Wang J D, Wei C, Liu L X, Ma N, Li M, Zhang Q, Pöschl U, Cheng Y F. Atmos. Chem. Phys., 2020, 20(20): 11729.

[67]
Harris E, Sinha B, van Pinxteren D, Tilgner A, Fomba K W, Schneider J, Roth A, Gnauk T, Fahlbusch B, Mertes S, Lee T, Collett J, Foley S, Borrmann S, Hoppe P, Herrmann H. Science, 2013, 340(6133): 727.

[68]
Alexander B, Park R J, Jacob D J, Gong S L. J. Geophys. Res., 2009, 114(D2): D02309.

[69]
Li J, Zhang Y L, Cao F, Zhang W Q, Fan M Y, Lee X H, Michalski G. Environ. Sci. Technol., 2020, 54(5): 2626.

[70]
Shao J Y, Chen Q J, Wang Y X, Lu X, He P Z, Sun Y L, Shah V, Martin R V, Philip S, Song S J, Zhao Y, Xie Z Q, Zhang L, Alexander B. Atmos. Chem. Phys., 2019, 19(9): 6107.

[71]
Yue F G, He P Z, Chi X Y, Wang L Q, Yu X W, Zhang P F, Xie Z Q. Atmos. Pollut. Res., 2020, 11(8): 1351.

[72]
Chen Q J, Geng L, Schmidt J A, Xie Z Q, Kang H, Dachs J, Cole-Dai J, Schauer A J, Camp M G, Alexander B. Atmos. Chem. Phys., 2016, 16(17): 11433.

[73]
Yiin B S, Margerum D W. Inorg. Chem., 1988, 27(10): 1670.

[74]
Xia M, Peng X, Wang W H, Yu C, Sun P, Li Y Y, Liu Y L, Xu Z N, Wang Z, Xu Z, Nie W, Ding A J, Wang T. Atmos. Chem. Phys., 2020, 20(10): 6147.

[75]
Zhang F, Wang Y, Peng J F, Chen L, Sun Y L, Duan L, Ge X L, Li Y X, Zhao J Y, Liu C, Zhang X C, Zhang G, Pan Y P, Wang Y S, Zhang A L, Ji Y M, Wang G H, Hu M, Molina M J, Zhang R Y. Proc. Natl. Acad. Sci. U. S. A., 2020, 117(8): 3960.

[76]
He H, Wang Y S, Ma Q X, Ma J Z, Chu B W, Ji D S, Tang G Q, Liu C, Zhang H X, Hao J M. Sci. Rep., 2014, 4: 4172.

[77]
Xie Y N, Ding A J, Nie W, Mao H T, Qi X M, Huang X, Xu Z, Kerminen V M, Petäjä T, Chi X G, Virkkula A, Boy M, Xue L K, Guo J, Sun J N, Yang X Q, Kulmala M, Fu C B. J. Geophys. Res. Atmos., 2015, 120(24): 12679.

[78]
Liu C, Ma Q X, Liu Y C, Ma J Z, He H. Phys. Chem. Chem. Phys., 2012, 14(5): 1668.

[79]
Zhao Y, Liu Y C, Ma J Z, Ma Q X, He H. Atmos. Environ., 2017, 152: 465.

[80]
Yu T, Zhao D F, Song X J, Zhu T. Atmos. Chem. Phys., 2018, 18(9): 6679.

[81]
Yang W W, Ma Q X, Liu Y C, Ma J Z, Chu B W, Wang L, He H. J. Phys. Chem. A, 2018, 122(30): 6311.

[82]
Zhang S P, Xing J, Sarwar G, Ge Y L, He H, Duan F K, Zhao Y, He K B, Zhu L D, Chu B W. Atmos. Environ., 2019, 208: 133.

[83]
Zhao D F, Song X J, Zhu T, Zhang Z F, Liu Y J, Shang J. Atmos. Chem. Phys., 2018, 18(4): 2481.

[84]
Brodzinsky R, Chang S G, Markowitz S S, Novakov T. J. Phys. Chem., 1980, 84(25): 3354.

[85]
He G Z, Ma J Z, He H. ACS Catal., 2018, 8(5): 3825.

[86]
He G Z, He H. Environ. Sci. Technol., 2020, 54(12): 7070.

[87]
Chen Z, Liu P, Wang W G, Cao X, Liu Y X, Zhang Y H, Ge M F. Environ. Sci. Technol., 2022, 56(12): 7637.

[88]
Wang W G, Liu M Y, Wang T T, Song Y, Zhou L, Cao J J, Hu J N, Tang G G, Chen Z, Li Z J, Xu Z Y, Peng C, Lian C F, Chen Y, Pan Y P, Zhang Y H, Sun Y L, Li W J, Zhu T, Tian H Z, Ge M F. Nat. Commun., 2021, 12: 1993.

[89]
Hung H M, Hoffmann M R. Environ. Sci. Technol., 2015, 49(23): 13768.

[90]
Hung H M, Hsu M N, Hoffmann M R. Environ. Sci. Technol., 2018, 52(16): 9079.

[91]
Wang T T, Liu M X, Liu M Y, Song Y, Xu Z Y, Shang F, Huang X, Liao W L, Wang W G, Ge M F, Cao J J, Hu J N, Tang G G, Pan Y P, Hu M, Zhu T. Environ. Sci. Technol., 2022, 56(12): 7771.

[92]
Zhang R F, Gen M S, Huang D D, Li Y J, Chan C K. Environ. Sci. Technol., 2020, 54(7): 3831.

[93]
Zheng H T, Song S J, Sarwar G, Gen M S, Wang S X, Ding D, Chang X, Zhang S P, Xing J, Sun Y L, Ji D S, Chan C K, Gao J, McElroy M B. Environ. Sci. Technol. Lett., 2020, 7(9): 632.

[94]
Gen M S, Zhang R F, Huang Dan Dan, Li Y J, Chan C K. Environ. Sci. Technol. Lett., 2019, 6(2): 86.

[95]
Gen M S, Zhang R F, Li Y J, Chan C K. Environ. Sci. Technol., 2020, 54(16): 9862.

[96]
Wang X K, Gemayel R, Hayeck N, Perrier S, Charbonnel N, Xu C H, Chen H, Zhu C, Zhang L W, Wang L, Nizkorodov S A, Wang X M, Wang Z, Wang T, Mellouki A, Riva M, Chen J M, George C. Environ. Sci. Technol., 2020, 54(6): 3114.

[97]
Yu Z C, Jang M. Atmos. Chem. Phys., 2018, 18(19): 14609.

[98]
Romer P S, Wooldridge P J, Crounse J D, Kim M J, Wennberg P O, Dibb J E, Scheuer E, Blake D R, Meinardi S, Brosius A L, Thames A B, Miller D O, Brune W H, Hall S R, Ryerson T B, Cohen R C. Environ. Sci. Technol., 2018, 52(23): 13738.

[99]
Benedict K B, McFall A S, Anastasio C. Environ. Sci. Technol., 2017, 51(8): 4387.

[100]
Ye C X, Zhang N, Gao H L, Zhou X L. Environ. Sci. Technol., 2017, 51(12): 6849.

[101]
Wingen L M, Moskun A C, Johnson S N, Thomas J L, Roeselová M, Tobias D J, Kleinman M T, Finlayson-Pitts B J. Phys. Chem. Chem. Phys., 2008, 10(37): 5668.

[102]
Richards N K, Wingen L M, Callahan K M, Nishino N, Kleinman M T, Tobias D J, Finlayson-Pitts B J. J. Phys. Chem. A, 2011, 115(23): 5810.

[103]
Richards N K, Finlayson-Pitts B J. Environ. Sci. Technol., 2012, 46(19): 10447.

[104]
George C, Ammann M, D’Anna B, Donaldson D J, Nizkorodov S A. Chem. Rev., 2015, 115(10): 4218.

[105]
Liu Y Y, Wang T, Fang X Z, Deng Y, Cheng H Y, Bacha A U R, Nabi I, Zhang L W. Sci. Total Environ., 2020, 734: 139415.

[106]
Park J, Jang M, Yu Z C. Environ. Sci. Technol., 2017, 51(17): 9605.

[107]
Park J Y, Jang M. RSC Adv., 2016, 6(63): 58617.

[108]
Dupart Y, King S M, Nekat B, Nowak A, Wiedensohler A, Herrmann H, David G, Thomas B, Miffre A, Rairoux P, D’Anna B, George C. Proc. Natl. Acad. Sci. U. S. A., 2012, 109(51): 20842.

[109]
Ma Q X, Wang L, Chu B W, Ma J Z, He H. J. Phys. Chem. A, 2019, 123(7): 1311.

[110]
Chen X R, Wang H C, Lu K D, Li C M, Zhai T Y, Tan Z F, Ma X F, Yang X P, Liu Y H, Chen S Y, Dong H B, Li X, Wu Z J, Hu M, Zeng L M, Zhang Y H. Environ. Sci. Technol., 2020, 54(15): 9243.

[111]
Wang H C, Lu K D, Guo S, Wu Z J, Shang D J, Tan Z F, Wang Y J, Le Breton M, Lou S R, Tang M J, Wu Y S, Zhu W F, Zheng J, Zeng L M, Hallquist M, Hu M, Zhang Y H. Atmos. Chem. Phys., 2018, 18(13): 9705.

[112]
Brown S S, DubÉ W P, Peischl J, Ryerson T B, Atlas E, Warneke C, de Gouw J A, te Lintel Hekkert S, Brock C A, Flocke F, Trainer M, Parrish D D, Feshenfeld F C, Ravishankara A R. J. Geophys. Res., 2011, 116(D24): D23405.

[113]
Heintz F, Platt U, Flentje H, Dubois R. J. Geophys. Res., 1996, 101(D17): 22891.

[114]
Alexander B, Sherwen T, Holmes C D, Fisher J A, Chen Q J, Evans M J, Kasibhatla P. Atmos. Chem. Phys., 2020, 20(6): 3859.

[115]
Alexander B, Hastings M G, Allman D J, Dachs J, Thornton J A, Kunasek S A. Atmos. Chem. Phys., 2009, 9(14): 5043.

[116]
Michalski G, Scott Z, Kabiling M, Thiemens M H. Geophys. Res. Lett., 2003, 30(16): 1870.

[117]
Wang Y L, Song W, Yang W, Sun X C, Tong Y D, Wang X M, Liu C Q, Bai Z P, Liu X Y. J. Geophys. Res. Atmos., 2019, 124(7): 4174.

[118]
Atkinson R. Atmos. Environ., 2000, 34(12/14): 2063.

[119]
He P Z, Xie Z Q, Chi X Y, Yu X W, Fan S D, Kang H, Liu C, Zhan H C. Atmos. Chem. Phys., 2018, 18(19): 14465.

[120]
Wang H C, Lu K D, Chen X R, Zhu Q D, Chen Q, Guo S, Jiang M Q, Li X, Shang D J, Tan Z F, Wu Y S, Wu Z J, Zou Q, Zheng Y, Zeng L M, Zhu T, Hu M, Zhang Y H. Environ. Sci. Technol. Lett., 2017, 4(10): 416.

[121]
Wahner A, Mentel T F, Sohn M. Geophys. Res. Lett., 1998, 25(12): 2169.

[122]
Fan M Y, Zhang Y L, Lin Y C, Hong Y H, Zhao Z Y, Xie F, Du W, Cao F, Sun Y L, Fu P Q. Environ. Sci. Technol., 2022, 56(11): 6870.

[123]
Bertram T H, Thornton J A. Atmos. Chem. Phys., 2009, 9(21): 8351.

[124]
Davis J M, Bhave P V, Foley K M. Atmos. Chem. Phys., 2008, 8(17): 5295.

[125]
Gržinić G, Bartels-Rausch T, Berkemeier T, Türler A, Ammann M. Atmos. Chem. Phys., 2015, 15(23): 13615.

[126]
Griffiths P T, Badger C L, Cox R A, Folkers M, Henk H H, Mentel T F. J. Phys. Chem. A, 2009, 113(17): 5082.

[127]
Bertram T H, Thornton J A, Riedel T P, Middlebrook A M, Bahreini R, Bates T S, Quinn P K, Coffman D J. Geophys. Res. Lett., 2009, 36(19): L19803.

[128]
Gaston C J, Thornton J A, Ng N L. Atmos. Chem. Phys., 2014, 14: 5693.

[129]
Wang Z, Wang W H, Tham Y J, Li Q Y, Wang H, Wen L, Wang X F, Wang T. Atmos. Chem. Phys., 2017, 17(20): 12361.

[130]
McDuffie E E, Fibiger D L, DubÉ W P, Lopez-Hilfiker F, Lee B H, Thornton J A, Shah V, JaeglÉ L, Guo H Y, Weber R J, Michael Reeves J, Weinheimer A J, Schroder J C, Campuzano-Jost P, Jimenez J L, Dibb J E, Veres P, Ebben C, Sparks T L, Wooldridge P J, Cohen R C, Hornbrook R S, Apel E C, Campos T, Hall S R, Ullmann K, Brown S S. J. Geophys. Res. Atmos., 2018, 123(8): 4345.

[131]
Yu C, Wang Z, Xia M, Fu X, Wang W H, Tham Y J, Chen T S, Zheng P G, Li H Y, Shan Y, Wang X F, Xue L K, Zhou Y, Yue D L, Ou Y B, Gao J, Lu K D, Brown S S, Zhang Y H, Wang T. Atmos. Chem. Phys., 2020, 20(7): 4367.

[132]
Lu K D, Hofzumahaus A, Holland F, Bohn B, Brauers T, Fuchs H, Hu M, Häseler R, Kita K, Kondo Y, Li X, Lou S R, Oebel A, Shao M, Zeng L M, Wahner A, Zhu T, Zhang Y H, Rohrer F. Atmos. Chem. Phys., 2013, 13(2): 1057.

[133]
Sherwen T, Schmidt J A, Evans M J, Carpenter L J, Großmann K, Eastham S D, Jacob D J, Dix B, Koenig T K, Sinreich R, Ortega I, Volkamer R, Saiz-Lopez A, Prados-Roman C, Mahajan A S, OrdÓñez C. Atmos. Chem. Phys., 2016, 16(18): 12239.

[134]
Chuang M T, Wu C F, Lin C Y, Lin W C, Chou C C K, Lee C T, Lin T H, Fu J S, Kong S S K. Atmos. Environ., 2022, 269: 118856.

[135]
Fu X, Wang T, Gao J, Wang P, Liu Y M, Wang S X, Zhao B, Xue L K. Environ. Sci. Technol., 2020, 54(7): 3881.

[136]
Qin M M, Hu A Q, Mao J J, Li X, Sheng L, Sun J J, Li J Y, Wang X S, Zhang Y H, Hu J L. Sci. Total Environ., 2022, 810: 152268.

[137]
Feng T, Zhao S Y, Zhang X, Wang Q Y, Liu L, Li G H, Tie X X. Sci. Total Environ., 2020, 745: 140961.

[138]
Jung D, de la Paz D, Notario A, Borge R. Sci. Total Environ., 2022, 827: 154126.

[139]
Liu T T, Hong Y W, Li M R, Xu L L, Chen J S, Bian Y H, Yang C, Dan Y B, Zhang Y N, Xue L K, Zhao M, Huang Z, Wang H. Atmos. Chem. Phys., 2022, 22(3): 2173.

[140]
Lu K D, Fuchs H, Hofzumahaus A, Tan Z F, Wang H C, Zhang L, Schmitt S H, Rohrer F, Bohn B, Broch S, Dong H B, Gkatzelis G I, Hohaus T, Holland F, Li X, Liu Y, Liu Y H, Ma X F, Novelli A, Schlag P, Shao M, Wu Y S, Wu Z J, Zeng L M, Hu M, Kiendler-Scharr A, Wahner A, Zhang Y H. Environ. Sci. Technol., 2019, 53(18): 10676.

[141]
Li K, Jacob D J, Liao H, Qiu Y L, Shen L, Zhai S X, Bates K H, Sulprizio M P, Song S J, Lu X, Zhang Q, Zheng B, Zhang Y L, Zhang J Q, Lee H C, Kuk S K. Proc. Natl. Acad. Sci. U. S. A., 2021, 118(10): e2015797118.

[142]
Fu X, Wang T, Zhang L, Li Q Y, Wang Z, Xia M, Yun H, Wang W H, Yu C, Yue D L, Zhou Y, Zheng J Y, Han R. Atmos. Chem. Phys., 2019, 19(1): 1.

[143]
Li J Y, Zhang N, Wang P, Choi M, Ying Q, Guo S, Lu K D, Qiu X H, Wang S X, Hu M, Zhang Y H, Hu J L. Environ. Pollut., 2021, 287: 117624.

[144]
Thornton J A, Kercher J P, Riedel T P, Wagner N L, Cozic J, Holloway J S, DubÉ W P, Wolfe G M, Quinn P K, Middlebrook A M, Alexander B, Brown S S. Nature, 2010, 464(7286): 271.

[145]
Wang X, Jacob D J, Fu X, Wang T, Le Breton M, Hallquist M, Liu Z R, McDuffie E E, Liao H. Environ. Sci. Technol., 2020, 54(16): 9908.

[146]
Qiu X H, Ying Q, Wang S X, Duan L, Zhao J, Xing J, Ding D, Sun Y L, Liu B X, Shi A J, Yan X, Xu Q C, Hao J M. Atmos. Chem. Phys., 2019, 19(10): 6737.

[147]
Sommariva R, Hollis L D J, Sherwen T, Baker A R, Ball S M, Bandy B J, Bell T G, Chowdhury M N, Cordell R L, Evans M J, Lee J D, Reed C, Reeves C E, Roberts J M, Yang M X, Monks P S. Atmos. Sci. Lett., 2018, 19(8): e844.

[148]
Ahern A T, Goldberger L, Jahl L, Thornton J, Sullivan R C. Environ. Sci. Technol., 2018, 52(2): 550.

[149]
Young C J, Washenfelder R A, Edwards P M, Parrish D D, Gilman J B, Kuster W C, Mielke L H, Osthoff H D, Tsai C, Pikelnaya O, Stutz J, Veres P R, Roberts J M, Griffith S, Dusanter S, Stevens P S, Flynn J, Grossberg N, Lefer B, Holloway J S, Peischl J, Ryerson T B, Atlas E L, Blake D R, Brown S S. Atmos. Chem. Phys., 2014, 14(7): 3427.

[150]
Young C J, Washenfelder R A, Roberts J M, Mielke L H, Osthoff H D, Tsai C, Pikelnaya O, Stutz J, Veres P R, Cochran A K, van den Boer T C, Flynn J, Grossberg N, Haman C L, Lefer B, Stark H, Graus M, de Gouw J, Gilman J B, Kuster W C, Brown S S. Environ. Sci. Technol., 2012, 46(20): 10965.

[151]
Wang X F, Wang H, Xue L K, Wang T, Wang L W, Gu R R, Wang W H, Tham Y J, Wang Z, Yang L X, Chen J M, Wang W X. Atmos. Environ., 2017, 156: 125.

[152]
Wang H C, Tang M J, Tan Z F, Peng C, Lu K D. Progress in Chemistry, 2020, 32(10): 1535.

(王海潮, 唐明金, 谭照峰, 彭超, 陆克定. 化学进展, 2020, 32(10): 1535.).

[153]
Alicke B. J. Geophys. Res., 2003, 108(D4): 8247.

[154]
Xue C Y, Zhang C L, Ye C, Liu P F, Catoire V, Krysztofiak G, Chen H, Ren Y G, Zhao X X, Wang J H, Zhang F, Zhang C X, Zhang J W, An J L, Wang T, Chen J M, Kleffmann J, Mellouki A, Mu Y J. Environ. Sci. Technol., 2020, 54(18): 11048.

[155]
Womack C C, McDuffie E E, Edwards P M, Bares R, Gouw J A, Docherty K S, DubÉ W P, Fibiger D L, Franchin A, Gilman J B, Goldberger L, Lee B H, Lin J C, Long R, Middlebrook A M, Millet D B, Moravek A, Murphy J G, Quinn P K, Riedel T P, Roberts J M, Thornton J A, Valin L C, Veres P R, Whitehill A R, Wild R J, Warneke C, Yuan B, Baasandorj M, Brown S S. Geophys. Res. Lett., 2019, 46(9): 4971.

[156]
Zhai S X, Jacob D J, Wang X, Liu Z R, Wen T X, Shah V, Li K, Moch J M, Bates K H, Song S J, Shen L, Zhang Y Z, Luo G, Yu F Q, Sun Y L, Wang L T, Qi M Y, Tao J, Gui K, Xu H H, Zhang Q, Zhao T L, Wang Y S, Lee H C, Choi H, Liao H. Nat. Geosci., 2021, 14(6): 389.

[157]
Lu M M, Tang X, Feng Y C, Wang Z F, Chen X S, Kong L, Ji D S, Liu Z R, Liu K X, Wu H J, Liang S W, Zhou H, Hu K. Sci. Total Environ., 2021, 788: 147747.

[158]
Li M M, Zhang Z H, Yao Q, Wang T J, Xie M, Li S, Zhuang B L, Han Y. Atmos. Chem. Phys., 2021, 21(19): 15135.

[159]
Dong X Y, Li J, Fu J S, Gao Y, Huang K, Zhuang G S. Sci. Total Environ., 2014, 481: 522.

[160]
Shah V, JaeglÉ L, Thornton J A, Lopez-Hilfiker F D, Lee B H, Schroder J C, Campuzano-Jost P, Jimenez J L, Guo H Y, Sullivan A P, Weber R J, Green J R, Fiddler M N, Bililign S, Campos T L, Stell M, Weinheimer A J, Montzka D D, Brown S S. Proc. Natl. Acad. Sci. U. S. A., 2018, 115(32): 8110.

[161]
Balamurugan V, Chen J, Qu Z, Bi X, Keutsch F N. Atmos. Chem. Phys., 2022, 22(11): 7105.

[162]
Han X, Zhu L Y, Liu M X, Song Y, Zhang M G. Atmos. Chem. Phys., 2020, 20(16): 9979.

[163]
Xu Z Y, Liu M X, Zhang M S, Song Y, Wang S X, Zhang L, Xu T T, Wang T T, Yan C Q, Zhou T, Sun Y L, Pan Y P, Hu M, Zheng M, Zhu T. Atmos. Chem. Phys., 2019, 19(8): 5605.

[164]
Ye Z L, Guo X R, Cheng L, Cheng S Y, Chen D S, Wang W L, Liu B. Atmos. Environ., 2019, 219: 116989.

[165]
Viatte C, Wang T Z, Van Damme M, Dammers E, Meleux F, Clarisse L, Shephard M W, Whitburn S, Coheur P F, Cady-Pereira K E, Clerbaux C. Atmos. Chem. Phys., 2020, 20(1): 577.

[166]
Wang S S, Nan J L, Shi C Z, Fu Q Y, Gao S, Wang D F, Cui H X, Saiz-Lopez A, Zhou B. Sci. Rep., 2015, 5: 15842.

[167]
Mozurkewich M. Atmos. Environ. A Gen. Top., 1993, 27(2): 261.

[168]
Wu C, Zhang S, Wang G H, Lv S J, Li D P, Liu L, Li J J, Liu S J, Du W, Meng J J, Qiao L P, Zhou M, Huang C, Wang H L. Environ. Sci. Technol., 2020, 54(24): 15622.

[169]
Tang M J, Cziczo D J, Grassian V H. Chem. Rev., 2016, 116(7): 4205.

[170]
Tobo Y, Zhang D Z, Matsuki A, Iwasaka Y. Proc. Natl. Acad. Sci. U. S. A., 2010, 107(42): 17905.

[171]
Liu Y, Gibson Cain, Wang H, Grassian, Laskin A. J. Phys. Chem. A, 2008, 112(7): 1561.

[172]
Li W J, Shao L Y, Shi Z B, Chen J M, Yang L X, Yuan Q, Yan C, Zhang X Y, Wang Y Q, Sun J Y, Zhang Y M, Shen X J, Wang Z F, Wang W X. J. Geophys. Res. Atmos., 2014, 119(2): 1044.

[173]
Goodman A L, Underwood G M, Grassian V H. J. Geophys. Res., 2000, 105(D23): 29053.

[174]
Wang G H, Cheng C L, Huang Y, Tao J, Ren Y Q, Wu F, Meng J J, Li J J, Cheng Y T, Cao J J, Liu S X, Zhang T, Zhang R, Chen Y B. Atmos. Chem. Phys., 2014, 14(21): 11571.

[175]
Wang G H, Zhou B H, Cheng C L, Cao J J, Li J J, Meng J J, Tao J, Zhang R J, Fu P Q. Atmos. Chem. Phys., 2013, 13(2): 819.

[176]
Gard E E, Kleeman M J, Gross D S, Hughes L S, Allen J O, Morrical B D, Fergenson D P, Dienes T, Gälli M E, Johnson R J, Cass G R, Prather K A. Science, 1998, 279(5354): 1184.

[177]
Saul T D, Tolocka M P, Johnston M V. J. Phys. Chem. A, 2006, 110(24): 7614.

[178]
Gupta D, Kim H, Park G, Li X, Eom H J, Ro C U. Atmos. Chem. Phys., 2015, 15(6): 3379.

[179]
Finlayson-Pitts B J. Proc. Natl. Acad. Sci. U. S. A., 2010, 107(15): 6566.

[180]
Ryder O S, Ault A P, Cahill J F, Guasco T L, Riedel T P, Cuadra-Rodriguez L A, Gaston C J, Fitzgerald E, Lee C, Prather K A, Bertram T H. Environ. Sci. Technol., 2014, 48(3): 1618.

Options
Outlines

/

〈 〉