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The majority of NOx in the atmosphere come from combustion-related human activities, including transportation, industrial boilers, power plants, domestic heating and municipal incineration (von Schneidemesser et al., 2015). The global emissions of NOx from these anthropogenic sources were estimated to be approximately 130 Tg (NO2) for the year 2014 (Hoesly et al., 2018). The close linkage between NOx and biogenic SOA formation is reflected in its ability to alter the SOA formation mechanism, composition and yield via affecting the gas-phase chemistry, gas–particle partitioning and particle-phase reactions, both during daytime and nighttime (Ma et al., 2012; Rollins et al., 2012).
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BVOC oxidation during daylight hours is dominated by ·OH (Ziemann and Atkinson, 2012). The initial addition or H-abstraction reaction between ·OH and BVOCs results in alkyl-type radicals (R·), most of which react rapidly with O2, leading to organic peroxy radicals (RO2·) (Atkinson, 2000). The general schematic of RO2· chemistry in SOA formation is summarized in Fig. 1. The influence of NOx is derived from its alteration of the fate of RO2·, which can either react with RO2·, hydroperoxy radicals (HO2·) or NOx under certain conditions. The different RO2· branches determine the distribution of oxidized products. For example, the reaction between RO2· and HO2· often produces hydroperoxides with low-volatility, RO2· self-reaction or reactions with other RO2· form alcohol or carbonyls, and the RO2· + NO reaction usually leads to organic nitrates as well as alkoxy radicals (RO·) that either undergo fragmentation or isomerization to form more volatile products (Ziemann and Atkinson, 2012; Sarrafzadeh et al., 2016). Since the fate of RO2· is highly related to the relative concentrations of NOx and VOCs in the urban atmosphere, laboratory chamber experiments often use the ratio of the initial BVOCs and NOx concentration ([BVOC]0/[NOx]0 or [NOx]0/[BVOC]0) to restrict the RO2· chemistry from the interpretation of NOx effects on new particle formation (NPF) and SOA yields (Pandis et al., 1991; Presto et al., 2005; Kim et al., 2012; Wildt et al., 2014; Xu et al., 2014; Stirnweis et al., 2017). It should be noted that attention must be paid to evaluate the O3-induced loss of BVOCs in the photooxidation system because O3 production and its effect would also vary with the [BVOC]0/[NOx]0 ratio, the relative rates of ozonolysis and ·OH oxidation and some other reaction conditions (Griffin et al., 1999). For the biogenic SOA formation in the presence of NOx listed in Table 1, the completely dominant role of ·OH oxidation in BVOC loss was estimated and thus O3 generation would not influence the NOx-dependent SOA yield.
Figure 1. General schematic picture of NOx effects on BVOC oxidation during daytime and nighttime. “Decom.” and “Isom.” represent decomposition and isomerization reactions, respectively.
BVOCs [BVOC]0
(ppb)[NOx]0/
[BVOC]0·OH
precursorsT (K) RH (%) Seed SOA mass
(μg m−3)Yield (%) Notes Reference Isoprene 91.4–114.6 0.7–7.3 H2O2 ~298 < 5 none 4.2–30.2 1.5–8.5 The SOA yield increased with initial NO/isoprene up to a ratio of 3, beyond which it decreases with increasing initial [NO]0/[isoprene]0 ratio. (Xu et al., 2014) 45±4 0–17 H2O2 ~301 < 10 ammonium sulfate 1.7–6.7 1.4–5.5 At high NOx (>200 ppb), the SOA yield decreased with increasing NOx. (Kroll et al., 2006) 26 0–1.9 H2O2 − 50 ammonium sulfate 1.9–7.6 2.7–11.6 The SOA yield was nearly constant at low NO until the [NO]0/[isoprene]0 ratio reached ~0.38). It further decreased with the increase of NO concentrations. (Liu et al., 2016) 50 0–0.5 H2O2 ~298 40±2 ammonium sulfate 0.5–1.2 0. 4–0.9 Higher [NOx]0/[isoprene]0 ratios produced lower aerosol yields. (King et al., 2010) 33–523 1.6–32 CH3ONO/
HONO296–298 9–11 ammonium sulfate 2.9–65.2 3.1–7.4 SOA yields were relevant to NO2/NO ratio under high NOx conditions. (Chan et al., 2010) 25–500 0.5–7.6 HONO 293–295 42–50 ammonium sulfate 0.7.–42.6 0.9–3 Higher [NOx]0/[isoprene]0 ratios produced lower aerosol yields. (Kroll et al., 2005) 180–2500 0.2–0.7 NOx 293 47–53 none 0.7–336 0.2–5.3 SOA yields first increased ([NOx]0/[isoprene]0 < 0.5) and then decreased with [NOx]0/[isoprene]0 ([NOx]0/[isoprene]0 > 0.5). (Dommen et al., 2006) α-pinene ~15 0–64.5 H2O2/HONO 296–299 3.3–6.4 ammonium sulfate 4.5–29.3 6.6–37.9 SOA yields were higher at lower initial [NOx]0/[α-pinene]0 ratios. (Ng et al., 2007b) 18.3–20.3 0.1–2.6 HONO 294–299 27–29 ammonium hydrogen sulfate and sulfuric acid 2.1–12 1.8–11.6 The yields at low [NOx]0/[α-pinene]0 ratios were in general higher compared to those at high [NOx]0/[α-pinene]0. (Stirnweis et al., 2017) 16.1–20.7 1.2–3.8 HONO 294–299 66–69 ammonium hydrogen sulfate and sulfuric acid 8.6–13.4 8.1–13.8 The yields at low [NOx]0/[α-pinene]0 ratios were in general higher compared to those at high [NOx]0/[α-pinene]0. (Stirnweis et al., 2017) 45–52.4 − H2O2/HONO/
CH3ONO293–298 < 10 ammonium sulfate 37.2–76.6 14.4–28.9 The SOA yield was suppressed under conditions of high NO. (Eddingsaas et al., 2012) 65–120 0.3–1.2 NOx 306–315 14–17 none 18–136 5.3–24 Aerosol yields should be higher at lower [NOx]0/[α-pinene]0 ratios. (Kim et al., 2012) 470–845 0.4–0.9 NOx 310–316 14–17 none 830–2100 34–68 SOA yields were higher at lower initial NOx/α-pinene ratios. (Kim et al., 2010) ~ 20 ~ 0–1 HONO 291–307 29–42 none − 0–10 Higher [NOx]0/[α-pinene]0 ratios produced lower aerosol yields. (Zhao et al., 2018b) β-pinene 37 0.01–3.9 HO2/NO 289±1 63±2 ammonium sulfate 14.3–38.1 8.2–20.0 SOA yields increased with increasing [NOx] at low-NOx conditions ([NOx]0 < 30 ppb, [NOx]0/[β-pinene]0 < 1 and decreased with [NOx] at high-NOx conditions ([NOx]0 >30 ppb, NOx/β-pinene ~1 to ~3.8). (Sarrafzadeh et al., 2016) 36–2000 0.2–19.6 NOx − − none − − Aerosol yields were small when [NOx]0/[β-pinene]0 was larger than 2, increased dramatically and reached maximum for the range of 0.7–1, then decreased slowly as the ratio decrease. (Pandis et al., 1991) 405–640 0.4–0.9 NOx 312–317 12–19 none 430–900 25–37 Higher [NOx]0/[β-pinene]0 ratios produced lower aerosol yields. (Kim et al., 2010) 32.3–96.5a ~2–10 NOx 308–313 ~5 ammonium sulfate 7.2–141.6 3.2–27.2 SOA yields were lower at higher NOx levels than at lower NOx levels.b (Griffin et al., 1999) Limonene 60–75 0.3–1.6 NOx 304–312 14–21 none 79.2–136 27–40 Higher [NOx]0/[limonene]0 ratios produced lower aerosol yields. (Kim et al., 2012) ~ 7 ~ 0–2.9 HONO 293–303 28–31 none − 0–5 Higher [NOx]0/[limonene]0 ratios produced lower aerosol yields. (Zhao et al., 2018b) 20.6–65.1a ~2–5 NOx 309–313 ~5 ammonium sulfate 9.5–120.2 8.7–34.4 SOA yields were lower at higher NOx levels than at lower NOx levels. b (Griffin et al., 1999) Sabinene 13.9–83.3a ~2–10 NOx 310–316 ~5 ammonium sulfate 2.5–14.5 1.9–65.2 SOA yields are lower at higher NOx levels than at lower NOx levels. b (Griffin et al., 1999) α-humulene 5–9.2a ~2–10 NOx 309–312 ~5 ammonium sulfate 12.9–59.2 31.9–84.5 The yields dependence on NOx levels is not obvious. b (Griffin et al., 1999) Longifolene ~4.3 0–131 H2O2/HONO 296–299 3.3–6.4 ammonium sulfate 28.5–51.6 84–157 SOA yields under high-NOx conditions exceed those under low-NOx conditions. (Ng et al., 2007b) Aromadendrene ~5 0– ~103 H2O2/HONO 296–299 3.3–6.4 ammonium sulfate 19.7–29.3 41.7–84.7 Aerosol yields increase with NOx concentrations. (Ng et al., 2007b) β-caryophyllene 3–32 0–1.7 H2O2/HONO 293±2 < 10 ammonium sulfate 8.4–311 19.3–137.8 SOA yields at low NOx conditions were lower than those at high NOx conditions. (Tasoglou and Pandis, 2015) 31.1–52.4 0.5–1.7 NOx ~298 ~70 none 35.6–66.2 9.5–19.9 The yields dependence on NOx levels was not obvious. (Alfarra et al., 2012) 5.9–12.9 ~2–5 NOx 309–312 ~5 ammonium sulfate 17.6–82.3 13.1–39.0 The yields dependence on NOx levels was not obvious. (Griffin et al., 1999) Notes: a Mixing ratios of BVOCs reacted due to the unavailable initial BVOC concentrations; b Effects of NOx on SOA yields are hypothesized if the reacted BVOCs are equal to the initial ones. Table 1. SOA formation from BVOC photooxidation in the presence of NOx.
The SOA yield is defined as the formed SOA mass concentration (ΔM, µg m−3) relative to the consumed parent hydrocarbon (ΔBVOC, µg m−3). The impact of NOx on SOA yields depends on the SOA mass production and is also parent hydrocarbon-specific (Table 1). For isoprene, the most abundant BVOC in the atmosphere (Kroll et al., 2006; Chan et al., 2010; Xu et al., 2014), the pathways of its reaction with RO2· under low and high NOx conditions are quite different (Fig. 2). Chamber studies have generally evidenced higher SOA yields at lower [NOx]0/[isoprene]0 ratios, and most of these studies have suggested that SOA yields first increase and then decrease with the increasing [NOx]0/[isoprene]0 ratios (Dommen et al., 2006; Kroll et al., 2006; King et al., 2010; Xu et al., 2014; Liu et al., 2016). The decrease of SOA yield with increasing NOx, more precisely with increasing NO, can generally be explained by the dominance of RO2· + NO reactions over RO2· + HO2· reactions, with the former producing more volatile products (such as organic nitrates) than the latter (hydroperoxides) (Kroll et al., 2006; Xu et al., 2014). Kroll et al. (2006) considered that the decline of the NO/HO2· ratio, which may lead to a switch from high-NOx to low-NOx conditions over the experimental process, might result in the complex SOA yield dependence under lower NOx conditions ([NOx]0/[isoprene]0 < 4.4). Xu et al. (2014) also observed similar nonlinear variation of aerosol volatility and oxidation state level with the [NO]0/[isoprene]0 ratio (0–7.3) as the SOA yield. They proposed that the presence of NO enhances the formation of methacrolein, the first generation product, whose further oxidation forms SOA-forming organics efficiently (Surratt et al., 2010), leading to increased SOA yield and decreased aerosol volatility when [NO]0/[isoprene]0 is lower than 3. In a more recent study focusing on a lower [NO]0/[isoprene]0 range (0–2), the SOA yield was nearly constant when the [NO]0/[isoprene]0 ratio was lower than ~0.38 (Liu et al., 2016). After this NO threshold level, the SOA yield decreased from 12% to 3% with a further increase of NOx, accompanied by a decrease of more highly oxygenated organic nitrates. These observations were explained by the suppression of NO on hydroxy hydroperoxide, which acts as the source of C5H11O6 peroxyl radicals and thus lowers the production of both second-generation multifunctional peroxides and multifunctional organic nitrates (Fig. 2). Similarly, with the composition analysis of isoprene SOA formed under low NOx in laboratory and aerosol samples collected from the isoprene-rich southeastern US environment, the none-IEPOX (isoprene epoxydiols) pathway under low NOx conditions was also suggested to contribute to notable highly oxidized compounds and SOA mass (Riva et al., 2016c).
Figure 2. Effects of NOx on isoprene SOA formation during daytime. Under high NOx conditions, isoprene RO2· primarily reacts with NO, forming methacrolein (MACR). The oxidation of MACR under high NO2/NO ratios forms methacryloylperoxynitrate (MPAN) while C4-hydroxynitrate peroxyacyl nitrate (C4-HN-PAN) is the main intermediate leading to SOA under high NOx conditions with low NO2/NO ratios. MPAN further reacts with ·OH to form methacrylic epoxide (MAE) and hydroxymethylmethyl-α-lactone (HMML). Acid-catalyzed reactions of MAE in the particle phase produce 2-methylglyceric acid, an organosulfate, and an oligomer. Under low NOx conditions, isoprene RO2· reacts predominantly with HO2·, leading to hydroxy hydroperoxide (ISOPOOH). ISOPOOH-derived epoxydiols (IEPOX) undergo multiphase acid-catalyzed chemistry to give various products in the particle phase. The non-IEPOX pathway that gives dihydroxy dihydroperoxides (ISOP(OOH)2) and organic nitrates (ISOP(OOH)N) is proposed to contribute to SOA formation without reactive aqueous seed particles. References for the non-IEPOX pathways are Liu et al. (2016) and Riva et al. (2016c), while for other pathways they are Lin et al. (2013b), Surratt et al. (2010), Lin et al. (2012) and Lin et al. (2013a).
Note that, although similar trends of the isoprene SOA yield response to NOx levels have been observed among different studies, the critical [NOx]0/[isoprene]0 points for the transition role of NOx are quite different [e.g., 4.4 (Kroll et al., 2006), 0.38 (Liu et al., 2016), and ~3 (Xu et al., 2014)]. It has been shown that, even under the same [NOx]0/[isoprene]0 ratios, the fate of RO2· radicals that are responsible for SOA formation can be quite different (Ng et al., 2007a). Recent studies have suggested that the composition of NOx itself is also a candidate for altering SOA formation pathways (Chan et al., 2010; Surratt et al., 2010). For example, oligoesters of dihydroxycarboxylic acids and hydroxynitrooxycarboxylic acids from isoprene photooxidation increased with increasing NO2/NO ratios (Chan et al., 2010). More recent studies show that SOA yields under high NOx conditions can be as high as those under low-NOx conditions because the NO2 + RO2· reaction can potentially yield substantial SOA mass (e.g., hydroxymethylmethyl-α-lactone, methacrylic acid) via the subsequent oxidation of methacryloylperoxynitrate, which is favorably formed from methacrolein (first-generation products of isoprene photooxidation) oxidation under high NO2/NO ratios (Fig. 2) (Chan et al., 2010; Surratt et al., 2010; Lin et al., 2012, 2013b; Pye et al., 2013; Nguyen et al., 2015). Besides NO2/NO ratios, the ·OH precursors, such as HONO, which strongly suppresses ISOPOOH chemistry and thus the formation of the second-generation organic nitrates, the chamber operation mode (flow or batch mode) and some other reaction conditions (e.g., seed particles), are potential factors that induce the differences in threshold [NOx]0/[isoprene]0 values, thus warranting further studies for more accurate model parametrization (Kroll et al., 2005; Xu et al., 2014; Liu et al., 2016; Shrivastava et al., 2017).
The effects of NOx on SOA formation from the photooxidation of monoterpenes, especially α-pinene, β-pinene and limonene, have also been characterized by chamber studies (Pandis et al., 1991; Zhang et al., 1992; Ng et al., 2007b; Eddingsaas et al., 2012; Kim et al., 2012; Wildt et al., 2014; Sarrafzadeh et al., 2016; Stirnweis et al., 2017; Zhao et al., 2018b). As summarized in Table 1, SOA yields are generally higher under low-NOx than high-NOx conditions when monoterpene ozonolysis is negligible. Besides the perturbation of NOx on RO2· chemistry, recent studies have found that NOx influence the SOA yield by altering the ·OH cycle and NPF (Wildt et al., 2014; Sarrafzadeh et al., 2016; Zhao et al., 2018b). Using realistic BVOC mixtures emitted directly by plants, Wildt et al. (2014) found that NPF was suppressed under high-NOx conditions ([BVOC]0/[NOx]0 < 7, [NOx]0 > 23 ppb). The self-reaction of higher-generation peroxy radical-like intermediates and their reaction with NO commonly limit the rate of NPF. More recently, a study focusing on β-pinene photooxidation showed that under low-NOx conditions ([β-pinene]0/[NOx]0 > 10 ppbC ppb−1) the increase in ·OH radicals through the reaction NO + HO2· → NO2 + ·OH was responsible for the increase in SOA yield with the increase in NOx (Sarrafzadeh et al., 2016). It was also evidenced that the ratio of NO/NO2 was correlated with the ·OH cycle and, thus, probably influenced SOA formation. Under high-NOx conditions ([β-pinene]0/[NOx]0 = ~10 to ~2.6 ppbC ppb−1), the decrease in SOA yield with NOx was attributed to NOx-triggered suppression of low-volatility products (such as hydroperoxides) that participated in NPF. The restrained NPF would further result in limited particle surfaces for the condensation of low-volatility species. Similarly, the suppression effect of NOx on NPF has been evidenced during the photooxidation of α-pinene and limonene (Zhao et al., 2018b).
Sesquiterpenes on a reacted mass basis have much higher SOA formation potential than isoprene and monoterpenes owing to their higher molecular weight and reactivity (Lee et al., 2006; Jaoui et al., 2013). As opposed to NOx effects on SOA formation from isoprene and monoterpenes photooxidation, SOA formed from longifolene, aromadendren and β-caryophyllene photooxidation under high-NOx conditions substantially exceeds that under low-NOx conditions (Ng et al., 2007b; Tasoglou and Pandis, 2015). The formation of less volatile products (e.g., large hydroxycarbonyls, multifunctional species) via isomerization instead of decomposition of large RO· and the relatively low-volatility organic nitrates were proposed to be responsible for this positive NOx effect. However, SOA yields from β-caryophyllene in the works of Griffin et al. (1999) and Alfarra et al. (2012) were less dependent on [NOx]0/[BVOC]0 ratios, probably due to the interference of other experimental conditions (e.g., OH precursors, the initial BVOC mixing ratios). Clearly, if the positive NOx effect on SOA formation observed by Ng et al. (2007b) can be extended to other sesquiterpenes, the contribution of sesquiterpenes to SOA in NOx-polluted air may be much higher (Ng et al., 2007b). A recent modeling study in the southeastern US showed underestimated SOA formation from monoterpenes and sesquiterpenes and argued that anthropogenic emissions would exert complex influences on biogenic SOA formation (Xu et al., 2018). Considering that studies on NOx effects only target a limited number of sesquiterpenes, a thorough evaluation of the effect of NOx on the photooxidation of a complete suite of sesquiterpenes is necessary for better constraint of their oxidation and contribution to ambient SOA.
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Nighttime biogenic SOA formation in the atmosphere is sensitive to NOx levels because of the changed radical (e.g., RO2·, HO2·, NO3·) chemistry and the oxidation capacity (Brown and Stutz, 2012; Ng et al., 2017). While ·OH dominates daytime BVOC oxidation, NO3·, which is mainly produced via the reaction between O3 and NO2, becomes one of the main oxidants at night (Fig. 1) (Wayne et al., 1991; Rollins et al., 2012; Edwards et al., 2017). The unsaturated and non-aromatic nature of BVOCs makes them particularly susceptible to oxidation by NO3· and O3 (Atkinson and Arey, 1998; Ayres et al., 2015). The competition between these two BVOC sinks is closely associated with the NOx level and composition because of the loss of NO3· through its reaction of NO and a decrease in its production through the reaction of NO2 and O3, as O3 is decreased by the reaction of O3 and NO (Rollins et al., 2012; Qin et al., 2018b; Wang et al., 2020a). The oxidation of BVOCs by NO3· occurs mainly via the addition of NO3· to the unsaturated bonds (another pathway is hydrogen abstraction, favored for aldehydic species), forming alkyl radicals that would either lose NO2 to form epoxides or further react with O2 to form RO2· (Fig. 1) (Ng et al., 2017; Fouqueau et al., 2020). RO2· would isomerize or react with HO2·, NO3· or RO2· to form various products such as organic nitrates that potentially generate SOA. NO3·–BVOCs chemistry is thus regarded as a prominent candidate for the generation of biogenic SOA and organic nitrates that are correlated with anthropogenic tracers (Fry et al., 2009; Kiendler-Scharr et al., 2016; Huang et al., 2019).
Such correlations have been evidenced in recent field observations around the world (Rollins et al., 2012; Brown et al., 2013; Kiendler-Scharr et al., 2016; Edwards et al., 2017; Fry et al., 2018; Yu et al., 2019). In a rural area in Southwest Germany, the contribution of organic nitrates to the increase of newly formed particles after sunset was observed to be 18%–25%. Considering both high BVOCs and NOx emissions in this area, the reactions between NO3· and BVOCs, especially monoterpenes, are responsible for organic nitrates and SOA formation (Huang et al., 2019). In some forest regions of the US, the concentration of organic nitrates was found to peak at night and its contribution to the total OA was up to 40% in Bakersfield owing to nighttime oxidation of BVOCs by NO3· (Rollins et al., 2012; Fry et al., 2013; Xu et al., 2015a). A substantial contribution of organic nitrates formed via nocturnal NO3·–BVOCs chemistry to particulate organic mass has also been observed in Europe and China (Kiendler-Scharr et al., 2016; Yu et al., 2019). Interestingly, the observation in the forest region of the western US showed that the concentration of nighttime aerosol organic nitrates was positively correlated with the product of the mixing ratios of NO2 and O3 instead of that of O3 alone (Fry et al., 2013). This indicates that NO3·-initiated oxidation of monoterpenes is related to the NOx level and is an important source of particle-phase organic nitrates at night.
The SOA formation potential of various BVOCs oxidized by NO3· has been investigated in many chamber studies [Ng et al. (2017) and references therein]. The reported SOA yields vary among different BVOCs, from nearly 0 for α-pinene, to 0.12 for isoprene, 0.33–0.44 for β-pinene, 0.44–0.57 for limonene, and up to 0.86 for β-caryophyllene at an atmospheric relevant aerosol mass loading of 10 μg m−3 (Fry et al., 2014). Except for α-pinene, these yield values are much higher than those from the ozonolysis of corresponding BVOCs (Song et al., 2007; von Hessberg et al., 2009; Saathoff et al., 2009; Tasoglou and Pandis, 2015). The relative importance of NO3· oxidation versus O3 is connected with the ratio of NO3· production to BVOC ozonolysis (Griffin et al., 1999). Considering, for example, 10 ppt NO3· and 30 ppb O3, the oxidation of these monoterpenes by NO3· proceeds 20–90 times faster than their ozonolysis, due to the much higher rate constants of the former reactions (Fry et al., 2014). The accelerated BVOC consumption by NO3· here is somewhat consistent with the field observations, which found NO3· + monoterpenes chemistry to be a significant nighttime aerosol source in regions with a high NOx level.
While most chamber studies have directly investigated NO3·-induced SOA under purified NO3· conditions (Griffin et al., 1999; Hallquist et al., 1999; Fry et al., 2014), some recent works have examined the biogenic SOA formation in the presence of Ox (O3 + NO2) (Table 2) (Presto et al., 2005; Perraud et al., 2012; Draper et al., 2015; Chen et al., 2017; Xu et al., 2020). The effects of NO2 on the dark ozonolysis of β-pinene, Δ3-carene, and limonene were examined by keeping the O3 mixing ratio constant while varying the NO2 mixing ratios ([O3]0/[NO2]0 = 2–0.5, [NO2]0/[BVOCs]0 = 0.5–1). It was found that, for β-pinene and Δ3-carene, SOA yields were comparable over the range of oxidation conditions. An increase of limonene SOA yield with increasing NO2 mixing ratio was observed and attributed to the increased fraction of oligomers and multifunctional organic nitrates in SOA through NO3· chemistry (Draper et al., 2015). More recently, the γ-terpinene SOA yield, as well as the contribution of organic nitrates to particle mass, were found to have both increased with increasing NO2 levels ([NO2]0/[O3]0 = 0–0.7, [NO2]0/[γ-terpinene]0 = 0–3), due to the change from O3-dominant to NO3·-dominant γ-terpinene oxidation, which yields organic nitrates as significant SOA components (Xu et al., 2020). Among the studied monoterpenes, α-pinene exhibited quite a different NO2 response during ozonolysis. Several studies have consistently found that SOA yields, as well as particle number concentrations, decreased with increasing NOx (Presto et al., 2005; Nøjgaard et al., 2006; Perraud et al., 2012; Draper et al., 2015). This is expected because the SOA yield from α-pinene ozonolysis is higher than that from NO3· oxidation, the latter process forming organic nitrates that have relative high volatility and are thus inefficient to nucleate (Perraud et al., 2012). In the real atmosphere, a good correlation between Ox and biogenic SOA tracers was also observed in a field campaign carried out in the Pearl River Delta, South China (Zhang et al., 2019b). With the elevation of Ox in the atmosphere, more observations focusing on the linkage between Ox and biogenic SOA are necessary but still limited. Altogether, these studies suggest that models should carefully handle the Ox effects on nocturnal SOA formation by capturing the detailed spatial distribution of BVOCs and Ox in order to reduce the uncertainty in the estimation of regional or global SOA budgets (Fry et al., 2014, 2018).
BVOCs [BVOC]0
(ppb)[NOx]0/
[BVOC]0[NOx]0/
[O3]0T(K) RH(%) Seed SOA mass
(μg m−3)Yield
(%)Notes Reference α-pinene 15–200 0.7–70 ~0.03–4 288–313 − none 1–346 0–0.29 The yields increase as NO2 concentrations decrease and reach an asymptote near [NOx]0/[BVOC]0 = 0.7. (Presto et al., 2005) 300–960 ~0–4.7 0–2.9 294–295 22–30 none − − The increase of [NO2]0 substantially depletes SOA formation. (Draper et al., 2015) 1000 0–6.3 ~0–4.5 − < 3 none − − Fewer particles are formed at higher NO2 conditions. (Perraud et al., 2012) 47±3 0–9.6 ~0–8.7 294±2 < 1 none − − Particle number concentration and volume were substantially reduced in the presence of NO2. (Nøjgaard et al., 2006) β-pinene 300–1100 ~0–6.7 0–4.2 295 23–40 none − − SOA yields are comparable over oxidant conditions studied. (Draper et al., 2015) Δ3-carene 220–650 ~0–3 0–1.9 294–295 27–38 none − − SOA yields are comparable over oxidant conditions studied. (Draper et al., 2015) limonene 150–159 0.2–0.4 0.5–75.9 295–297 9.2–9.9 none 30.3–157.3 0.27–0.73 The highest SOA yield occurred when [O3]/[NO] is around 1. (Chen et al., 2017) 300–560 ~0–3.3 0–2.2 295 20–31 none − − SOA formation was enhanced at higher NO2. (Draper et al., 2015) 51±3 0–6.9 0–7.1 294±2 < 1 none − − Particle number concentrations were lower at higher NOx conditions. (Nøjgaard et al., 2006) γ-terpene 152–154 0–2.9 0–0.7 297–301 24–30 none − 0.38–0.77 NOx enhance SOA yields and decrease particle number concentrations. (Xu et al., 2020) Table 2. SOA formation from BVOC ozonolysis in the presence of NOx.
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Many areas in China have been suffering severe haze events in the last few years (Li et al., 2017a; Zhao et al., 2018a; Lu et al., 2019). Although great efforts have been devoted to mitigating haze pollution by controlling various anthropogenic emissions (Xia et al., 2016), high mixing ratios of SO2, NOx and NH3 can still be observed to exceed 100 ppb and the contribution of POA to total submicron aerosol is up to 27% in regions like the North China Plain (Li et al., 2017a; Meng et al., 2018). While SOA derived from anthropogenic precursors, such as those VOCs emitted from traffic/coal burning, account for a significant fraction of fine particles, biogenic SOA also has a contribution and shows seasonal and regional dependence (Ding et al., 2014; Huang et al., 2014; Zhang et al., 2017b; Xing et al., 2019). Biogenic emissions in China were estimated to be 23.54 Tg yr−1 and contributed approximately for 70% of the total SOA in summer (Wu et al., 2020). Considering that anthropogenic emissions and BVOCs may coexist abundantly in regions like the Pearl River Delta, Yangtze River Delta, Sichuan Basin, and North China Plain, there is some evidence showing that anthropogenic–biogenic interactions are important in SOA formation in these regions, as Fig. 7 and Table 3 summarize (He et al., 2014, 2018; Hu et al., 2017; Zhang et al., 2019b; Wu et al., 2020).
Figure 7. Anthropogenic-biogenic interactions in China. The color-mapped annual emissions of total BVOCs in China, 2017, are adapted with permission from Wu et al. (2020). Copyright 2020 Elsevier. The observed correlations between anthropogenic pollutants and biogenic SOA are shown in red boxes and the modeled results are shown in yellow boxes. The pONSs, iONSs, iOSs, SOAI, SOAIE, SOAM, and SOAC refer to pinene-derived nitrooxyorganosulfates, isoprene-derived nitrooxyorganosulfates, isoprene-derived organosulfates, isoprene-derived SOA, IEPOX-derived SOA, monoterpene-derived SOA and β-caryophyllene-derived SOA, respectively; 2-MG and 2-MT are 2-methylglyceric acid and 2-methyltetrols derived from isoprene oxidation under high- and low-NOx conditions, respectively. a The modeled anthropogenic–biogenic interactions are taken from Qin et al. (2018). b–j The field-observed anthropogenic–biogenic interactions are taken from He et al. (2014), Zhang et al. (2019b), Zhang et al. (2017), Bryant et al. (2020), He et al. (2018), Ren et al. (2019), Ren et al. (2018), Li et al. (2013), and Wang et al. (2008), respectively.
Location Period T
(°C)RH
(%)SO2 a
(μg m−3)NOx a
(μg m−3)NO3− a
(μg m−3)SO42− a
(μg m−3)NH4+ a
(μg m−3)SOAI Tracers b ∑SOAI c
(ng m−3)SOAM Tracers d ∑SOAM e
(ng m−3)LWC f
(μg m−3)pHg References Guangzhou (urban) Year 2015 24.0 58 15.1 76.8 3.2 8.4 4.0 2-MT, 2-MG, C5-alkene triols, and 3-MeTHF-3,4-diols 22.6 cis-pinonic acid, pinic acid, 3-HGA, HDMGA, MBTCA 50.0 − − (Zhang et al., 2019b) Zhaoqing (urban) Year 2015 22.7 59 25.5 40.3 4.2 10.0 5.0 2-MT, 2-MG, C5-alkene triols, and 3-MeTHF-3,4-diols 49.3 cis-pinonic acid, pinic acid, 3-HGA, HDMGA, MBTCA 54.3 − − (Zhang et al., 2019b) Dongguan (urban) Year 2015 24.9 61 16.2 49.0 2.9 8.5 3.4 2-MT, 2-MG, C5-alkene triols, and 3-MeTHF-3,4-diols 16.0 cis-pinonic acid, pinic acid, 3-HGA, HDMGA, MBTCA 50.9 − − (Zhang et al., 2019b) Nansha (sub-urban) Year 2015 25.6 67 14.4 38.3 1.8 8.3 3.7 2-MT, 2-MG, C5-alkene triols, and 3-MeTHF-3,4-diols 17.0 cis-pinonic acid, pinic acid, 3-HGA, HDMGA, MBTCA 26.5 − − (Zhang et al., 2019b) Zhuhai (suburban) Year 2015 24.2 74 7.3 57.4 1.4 8.5 3.3 2-MT, 2-MG, C5-alkene triols, and 3-MeTHF-3,4-diols 10.8 cis-pinonic acid, pinic acid, 3-HGA, HDMGA, MBTCA 40.3 − − (Zhang et al., 2019b) Nanjing (urban) Summer 2013 32.4 59.7 128 i 39.3 h − 17 − 2-MT, 2-MG, and C5-alkene triols 0.3 − − 70 2.6 (Zhang et al., 2017c) Beijing (urban) Summer 2017 16–38 − − 225 i − − − 2-MT, 2-MT OSs, 2-MG, 2-MG OSs, glycolic acid sulfate, hydroxyacetone sulfate, lactic acid sulfate, cyclic, and 9 NOSs 107 − − − − (Bryant et al., 2020) Wanqingsha (forest) Summer 2010 29.6 79.7 29.4 42.4 h 2.8 9.1 3.1 2-MT sulfate ester, 2-MG sulfate ester 0.68 cis-pinonic acid, pinic acid, 3-HGA, HDMGA, MBTCA, NOSs (three isomers of MW 295) 75.9 − − (He et al., 2014) Fall 2010 21.6 69.1 45.1 37.5 10.4 18.6 8.8 0.66 205.4 − − Wanqingsha (forest) Summer 2008 29.0 66 − − 5.3 23.0 4.9 3-MeTHF-3,4-diols, 2-MT, C5-alkene triols, 2-MT sulfate ester, 2-MG, 2-MG sulfate ester 130.1 − − 24.5 0.5 (He et al., 2018) Fall 2008 22.6 47 − − 8.9 15.9 5.3 26.7 − 11.8 2.8 Wuyi Mountain Spring 2014 16 78 1.7 4.2 h − − − 3-MeTHF-3,4-diols, C5-alkene triols, 2-MT, 2-MG 6.6 cis-pinic acid, cis-pinonic acid, 3-HGA, MBTCA 26 9.7 0.2 (Ren et al., 2019) Summer 2014 23 79 0.9 1.7 h − − − 21 36 7.4 0.1 Autumn 2014 17 75 3.1 4 − − − 16 36 10.8 0.7 Winter 2014 6.4 64 6.7 6.2 − − − 3 20 7.2 1.6 Qinghai Lake Summer 2012 11 59 — − 0.4 2.2 0.4 3-MeTHF-3,4-diols, C5-alkene triols, 2-MT 3.8 cis-pinic acid, cis-pinonic acid, 3-HGA, MBTCA 16 − − (Ren et al., 2018) Winter 2012 −9 26 − − 0.8 2.2 0.1 0.6 1.3 − − Ürümqi (urban) Summer 2012 26 46 − − 3.4 6.4 0.4 3-MeTHF-3,4-diols, C5-alkene triols, 2-MT 10 cis-pinic acid, cis-pinonic acid, 3-HGA, MBTCA 44 − − (Ren et al., 2018) Winter 2012 −14 78 − − 19 65 21 1.9 6.6 − − Xi'an (urban) Summer 2012 24 78 − − 8.8 15 4.3 3-MeTHF-3,4-diols, C5-alkene triols, 2-MT 20 cis-pinic acid, cis-pinonic acid, 3-HGA, MBTCA 58 − − (Ren et al., 2018) Winter 2012 1 66 − − 26 36 13 2.1 22 − − Shanghai (urban) Summer 2012 28 78 − − 4.2 7.2 1.3 3-MeTHF-3,4-diols, C5-alkene triols, 2-MT 5.1 cis-pinic acid, cis-pinonic acid, 3-HGA, MBTCA 20 − − (Ren et al., 2018) Winter 2012 6 70 − − 16 16 6.1 2.5 16 − − Chengdu (urban) Summer 2012 25 81 − − 6.5 14 3.4 3-MeTHF-3,4-diols, C5-alkene triols, 2-MT 23 cis-pinic acid, cis-pinonic acid, 3-HGA, MBTCA 88 − − (Ren et al., 2018) Winter 2012 10 74 − − 26 32 12 5.9 17 − − Guangzhou (urban) Summer 2012 29 79 − − 3.3 6.2 0.8 3-MeTHF-3,4-diols, C5-alkene triols, 2-MT 10 cis-pinic acid, cis-pinonic acid, 3-HGA, MBTCA 43 − − (Ren et al., 2018) Winter 2012 17 73 − − 12 15 5.1 6 46 − − Tibetan Plateau (Qinghai Lake) Summer 2010 14.4 64.4 − − 0.8 3.9 0.6 C5-alkene triols, 2-MG, 2-MT 2.5 norpinic acid, pinonic acid, pinic acid, 3-HGA, MBTCA 3.0 5.8 −1.2 (Li et al., 2013) Changbai Mountain Summer 2007 25 59 5.35 1.3 − − − 2-MT, 2-MG, C5-alkene triols 53 pinic acid, norpinic acid, 3-HGA, MBTCA 31 − − (Wang et al., 2008) Chongming Island Summer 2006 29 68 25.9 40.9 j − − − 2-MT, 2-MG, C5-alkene triols 4.8 pinic acid, norpinic acid, 3-HGA, MBTCA 1.8 − − (Wang et al., 2008) Notes: a The mean concentration of tracers; b Isoprene-derived SOA (SOAI) tracers: 2-MG (2-methylglyceric acid), 2-MT (2-methyltetrols that represent the sum of 2-methylthreitol and 2-methylerythritol), 3-MeTHF-3,4-diols (the sum of trans-3-methyltetrahydrofuran-3,4-diol and cis-3-methyltetrahydrofuran-3,4-diol), C5-alkene triols (the sum of cis-2-methyl-1,3,4-trihydoxy-1-butane, trans-2-methyl-1,3,4-trihydoxy-1-butane, and 3-methyl-2,3,4-trihydoxy-1-butane), OSs (organosulfates), NOSs (nitrooxy organosulfates); c The sum of SOAI tracers; d Monoterpene-derived SOA (SOAM) tracers: 3-HGA (3-hydroxyglutaric acid), HDMGA (3-Hydroxy-4,4-dimethylglutaric acid), MBTCA (3-methyl-1,2,3-butanetricarboxylic acid); e The sum of SOAM tracers; f Aerosol liquid water content; g AIM-derived in situ pH of the aqueous phase on aerosols; h The concentration of NO2; i The max concentration. Table 3. Summary of gaseous and particulate species in different regions with anthropogenic–biogenic interactions in China.
Biogenic organosulfates in ambient particles, which are formed through the cross-reaction between BVOCs and anthropogenic pollutants, are important markers of anthropogenic–biogenic interactions. Quantification of organosulfates in fine particle samples collected in the central Pearl River Delta in 2010 showed nearly three times higher pinene-derived nitrooxyorganosulfates (MW = 295) in fall than in summer, probably due to the higher levels of sulfates and NOx in fall (He et al., 2014). 2-Methyltetrol sulfate ester produced via isoprene-derived IEPOX oxidation under low-NOx conditions showed low concentrations. The high NOx mixing ratio (daily 65 ppb and hourly 163 ppb) here could be the reason why IEPOX formation was suppressed. Other observations in this region also showed the Ox and sulfate dependence of isoprene-SOA tracers (He et al., 2018; Zhang et al., 2019b). Simultaneously, SOA tracers originating from β-caryophyllene and high-generation monoterpene oxidation were positively correlated with Ox and sulfate (Zhang et al., 2019b). Interestingly, the reduction of 50% Ox in this region was estimated to be more efficient in reducing biogenic SOA than that of sulfate. In eastern China, combining field measurements and model analysis, the depression of IEPOX SOA by high-NOx levels was confirmed as the reactive uptake of IEPOX and the ratio of IEPOX to isoprene high-NOx SOA precursors were lower than those observed in regions with abundant biogenic emissions, high particle acidity and low-NOx concentrations (Zhang et al., 2017c). Biogenic SOA formation in summer 2012 over China was simulated using the Community Multiscale Air Quality (CMAQ) model, which considers the reactive uptake of isoprene-derived intermediates, multigenerational oxidation and detailed monoterpene SOA production (Qin et al., 2018a). Isoprene SOA tracers showed high concentrations in southwestern China owing to the abundant IEPOX and high particle surface area provided by sulfate. Similar positive correlations between biogenic SOA tracers and sulfate were also observed in urban Ürümqi, Qinghai Lake and urban Xi’an, Beijing, Nanjing, Pearl River Delta, and -Wanqingsha (He et al., 2014, 2018; Zhang et al., 2017c; Ren et al., 2018; Zhang et al., 2019b; Bryant et al., 2020). The isoprene SOA formation pathway in some areas of the Yangtze River Delta Region and North China Plain was influenced by NOx emissions, as a high ratio of 2-methylglyceric acid and 2-methyltetrols (0.06–0.1 by the model and 0.58–0.78 in observations) showed in these regions (Qin et al., 2018a). We note that, although the simulated total biogenic SOA in summer 2012 in China accurately tracked the observed data (normalized mean bias of 1% and r2 of 0.59), CMAQ did not simulate the ratios of 2-methylglyceric acid and 2-methyltetrols well. The uncertainties in the fate of IEPOX, 2-methylglyceric acid reaction parameters and C5-alkene triols formation pathways could be possible reasons for the discrepancies between modeled and observed results. The linear correlations between SOA tracers of isoprene, monoterpenes and sesquiterpenes and anthropogenic pollutants, such as SO2 and NOx, were also observed at Wuyi Mountain and Changbai Mountain in southeastern and northeastern China, respectively, suggesting that SO2 and NOx can enhance biogenic SOA production in the remote mountain area through acid-catalyzed heterogeneous chemistry (Wang et al., 2008; Ren et al., 2019). For the more polluted urban Beijing, which is characterized by both local isoprene and anthropogenic pollutants, anthropogenic-influenced biogenic SOA formation in summer 2017 was also observed (Bryant et al., 2020). Isoprene-derived particulate organosulfates and nitrooxy-organosulfates, the formation of which is related to NOx and particulate SO42− levels, accounted for 0.62% and could be as high as ~3% on certain days.
For China as a whole, SOA formation in 2013 was modeled by incorporating updated two-product SOA yields and SOA formation from the reactive uptake of isoprene-derived IEPOX and methacrylic acid epoxide into the updated 3D air quality model (Hu et al., 2017). The enhancement effect of anthropogenic emissions on biogenic SOA was evidenced because the SOA concentration was less than 1 µg m−3 when solely considering biogenic emissions (Hu et al., 2017). Similar anthropogenic–biogenic interactions were found in a more recent study (Wu et al., 2020). With the modeled anthropogenic and biogenic emissions in China in 2016, the CMAQ model that includes updated POA aging, SOA properties and IEPOX organosulfates formation rate constants showed that removing all anthropogenic emissions while keeping biogenic emissions unchanged led to a 60% reduction of SOA formation. These studies suggest that, athough the emission of BVOCs is uncontrollable, biogenic SOA reduction can be achieved through controlling anthropogenic emissions. It should be noted that the modeled SOA concentrations have not been compared with the direct SOA measurements owing to data limitations. Many other studies show that current models usually underestimate or predict the SOA concentration with large uncertainties because of the missed SOA precursors, formation mechanism, components and complex atmospheric conditions (Shrivastava et al., 2017; Liu et al., 2018; Slade et al., 2019). With more detailed measurements of the particle composition and biogenic SOA tracer performed in many areas over China (Table 3) and the increased knowledge of the SOA formation mechanism by laboratory studies, models could be better constrained by the observed data and model performance could be better evaluated.
BVOCs | [BVOC]0 (ppb) | [NOx]0/ [BVOC]0 | ·OH precursors | T (K) | RH (%) | Seed | SOA mass (μg m−3) | Yield (%) | Notes | Reference |
Isoprene | 91.4–114.6 | 0.7–7.3 | H2O2 | ~298 | < 5 | none | 4.2–30.2 | 1.5–8.5 | The SOA yield increased with initial NO/isoprene up to a ratio of 3, beyond which it decreases with increasing initial [NO]0/[isoprene]0 ratio. | (Xu et al., 2014) |
45±4 | 0–17 | H2O2 | ~301 | < 10 | ammonium sulfate | 1.7–6.7 | 1.4–5.5 | At high NOx (>200 ppb), the SOA yield decreased with increasing NOx. | (Kroll et al., 2006) | |
26 | 0–1.9 | H2O2 | − | 50 | ammonium sulfate | 1.9–7.6 | 2.7–11.6 | The SOA yield was nearly constant at low NO until the [NO]0/[isoprene]0 ratio reached ~0.38). It further decreased with the increase of NO concentrations. | (Liu et al., 2016) | |
50 | 0–0.5 | H2O2 | ~298 | 40±2 | ammonium sulfate | 0.5–1.2 | 0. 4–0.9 | Higher [NOx]0/[isoprene]0 ratios produced lower aerosol yields. | (King et al., 2010) | |
33–523 | 1.6–32 | CH3ONO/ HONO | 296–298 | 9–11 | ammonium sulfate | 2.9–65.2 | 3.1–7.4 | SOA yields were relevant to NO2/NO ratio under high NOx conditions. | (Chan et al., 2010) | |
25–500 | 0.5–7.6 | HONO | 293–295 | 42–50 | ammonium sulfate | 0.7.–42.6 | 0.9–3 | Higher [NOx]0/[isoprene]0 ratios produced lower aerosol yields. | (Kroll et al., 2005) | |
180–2500 | 0.2–0.7 | NOx | 293 | 47–53 | none | 0.7–336 | 0.2–5.3 | SOA yields first increased ([NOx]0/[isoprene]0 < 0.5) and then decreased with [NOx]0/[isoprene]0 ([NOx]0/[isoprene]0 > 0.5). | (Dommen et al., 2006) | |
α-pinene | ~15 | 0–64.5 | H2O2/HONO | 296–299 | 3.3–6.4 | ammonium sulfate | 4.5–29.3 | 6.6–37.9 | SOA yields were higher at lower initial [NOx]0/[α-pinene]0 ratios. | (Ng et al., 2007b) |
18.3–20.3 | 0.1–2.6 | HONO | 294–299 | 27–29 | ammonium hydrogen sulfate and sulfuric acid | 2.1–12 | 1.8–11.6 | The yields at low [NOx]0/[α-pinene]0 ratios were in general higher compared to those at high [NOx]0/[α-pinene]0. | (Stirnweis et al., 2017) | |
16.1–20.7 | 1.2–3.8 | HONO | 294–299 | 66–69 | ammonium hydrogen sulfate and sulfuric acid | 8.6–13.4 | 8.1–13.8 | The yields at low [NOx]0/[α-pinene]0 ratios were in general higher compared to those at high [NOx]0/[α-pinene]0. | (Stirnweis et al., 2017) | |
45–52.4 | − | H2O2/HONO/ CH3ONO | 293–298 | < 10 | ammonium sulfate | 37.2–76.6 | 14.4–28.9 | The SOA yield was suppressed under conditions of high NO. | (Eddingsaas et al., 2012) | |
65–120 | 0.3–1.2 | NOx | 306–315 | 14–17 | none | 18–136 | 5.3–24 | Aerosol yields should be higher at lower [NOx]0/[α-pinene]0 ratios. | (Kim et al., 2012) | |
470–845 | 0.4–0.9 | NOx | 310–316 | 14–17 | none | 830–2100 | 34–68 | SOA yields were higher at lower initial NOx/α-pinene ratios. | (Kim et al., 2010) | |
~ 20 | ~ 0–1 | HONO | 291–307 | 29–42 | none | − | 0–10 | Higher [NOx]0/[α-pinene]0 ratios produced lower aerosol yields. | (Zhao et al., 2018b) | |
β-pinene | 37 | 0.01–3.9 | HO2/NO | 289±1 | 63±2 | ammonium sulfate | 14.3–38.1 | 8.2–20.0 | SOA yields increased with increasing [NOx] at low-NOx conditions ([NOx]0 < 30 ppb, [NOx]0/[β-pinene]0 < 1 and decreased with [NOx] at high-NOx conditions ([NOx]0 >30 ppb, NOx/β-pinene ~1 to ~3.8). | (Sarrafzadeh et al., 2016) |
36–2000 | 0.2–19.6 | NOx | − | − | none | − | − | Aerosol yields were small when [NOx]0/[β-pinene]0 was larger than 2, increased dramatically and reached maximum for the range of 0.7–1, then decreased slowly as the ratio decrease. | (Pandis et al., 1991) | |
405–640 | 0.4–0.9 | NOx | 312–317 | 12–19 | none | 430–900 | 25–37 | Higher [NOx]0/[β-pinene]0 ratios produced lower aerosol yields. | (Kim et al., 2010) | |
32.3–96.5a | ~2–10 | NOx | 308–313 | ~5 | ammonium sulfate | 7.2–141.6 | 3.2–27.2 | SOA yields were lower at higher NOx levels than at lower NOx levels.b | (Griffin et al., 1999) | |
Limonene | 60–75 | 0.3–1.6 | NOx | 304–312 | 14–21 | none | 79.2–136 | 27–40 | Higher [NOx]0/[limonene]0 ratios produced lower aerosol yields. | (Kim et al., 2012) |
~ 7 | ~ 0–2.9 | HONO | 293–303 | 28–31 | none | − | 0–5 | Higher [NOx]0/[limonene]0 ratios produced lower aerosol yields. | (Zhao et al., 2018b) | |
20.6–65.1a | ~2–5 | NOx | 309–313 | ~5 | ammonium sulfate | 9.5–120.2 | 8.7–34.4 | SOA yields were lower at higher NOx levels than at lower NOx levels. b | (Griffin et al., 1999) | |
Sabinene | 13.9–83.3a | ~2–10 | NOx | 310–316 | ~5 | ammonium sulfate | 2.5–14.5 | 1.9–65.2 | SOA yields are lower at higher NOx levels than at lower NOx levels. b | (Griffin et al., 1999) |
α-humulene | 5–9.2a | ~2–10 | NOx | 309–312 | ~5 | ammonium sulfate | 12.9–59.2 | 31.9–84.5 | The yields dependence on NOx levels is not obvious. b | (Griffin et al., 1999) |
Longifolene | ~4.3 | 0–131 | H2O2/HONO | 296–299 | 3.3–6.4 | ammonium sulfate | 28.5–51.6 | 84–157 | SOA yields under high-NOx conditions exceed those under low-NOx conditions. | (Ng et al., 2007b) |
Aromadendrene | ~5 | 0– ~103 | H2O2/HONO | 296–299 | 3.3–6.4 | ammonium sulfate | 19.7–29.3 | 41.7–84.7 | Aerosol yields increase with NOx concentrations. | (Ng et al., 2007b) |
β-caryophyllene | 3–32 | 0–1.7 | H2O2/HONO | 293±2 | < 10 | ammonium sulfate | 8.4–311 | 19.3–137.8 | SOA yields at low NOx conditions were lower than those at high NOx conditions. | (Tasoglou and Pandis, 2015) |
31.1–52.4 | 0.5–1.7 | NOx | ~298 | ~70 | none | 35.6–66.2 | 9.5–19.9 | The yields dependence on NOx levels was not obvious. | (Alfarra et al., 2012) | |
5.9–12.9 | ~2–5 | NOx | 309–312 | ~5 | ammonium sulfate | 17.6–82.3 | 13.1–39.0 | The yields dependence on NOx levels was not obvious. | (Griffin et al., 1999) | |
Notes: a Mixing ratios of BVOCs reacted due to the unavailable initial BVOC concentrations; b Effects of NOx on SOA yields are hypothesized if the reacted BVOCs are equal to the initial ones. |