Nitrate radicals and biogenic volatile organic compounds: oxidation, mechanisms, and organic aerosol

Oxidation of biogenic volatile organic compounds (BVOC) by the nitrate radical (NO3) represents one of the important interactions between anthropogenic emissions related to combustion and natural emissions from the biosphere. This interaction has been recognized for more than 3 decades, during which time a large body of research has emerged from laboratory, field, and modeling studies. NO3-BVOC reactions influence air quality, climate and visibility through regional and global budgets for reactive nitrogen (particularly organic nitrates), ozone, and organic aerosol. Despite its long history of research and the significance of this topic in atmospheric chemistry, a number of important uncertainties remain. These include an incomplete understanding of the rates, mechanisms, and organic aerosol yields for NO3-BVOC reactions, lack of constraints on the role of heterogeneous oxidative processes associated with the NO3 radical, the difficulty of characterizing the spatial distributions of BVOC and NO3 within the poorly mixed nocturnal atmosphere, and the challenge of constructing appropriate boundary layer schemes and non-photochemical mechanisms for use in state-of-the-art chemical transport and chemistry–climate models. This review is the result of a workshop of the same title held at the Georgia Institute of Technology in June 2015. The first half of the review summarizes the current literature on NO3-BVOC chemistry, with a particular focus on recent advances in instrumentation and models, and in organic nitrate and secondary organic aerosol (SOA) formation chemistry. Building on this current understanding, the second half of the review outlines impacts of NO3-BVOC chemistry on air quality and climate, and suggests critical research needs to better constrain this interaction to improve the predictive capabilities of atmospheric models.


Introduction
The emission of hydrocarbons from the terrestrial biosphere represents a large natural input of chemically reactive compounds to Earth's atmosphere (Guenther et al., 1995;Goldstein and Galbally, 2007). Understanding the atmospheric degradation of these species is a critical area of current research that influences models of oxidants and aerosols on regional and global scales. Nitrogen oxides (NO x = NO + NO 2 ) arising from combustion and microbial action on fertilizer are one of the major anthropogenic inputs that perturb the chemistry of the atmosphere (Crutzen, 1973). Nitrogen oxides have long been understood to influence oxidation cycles of biogenic volatile organic compounds (BVOC), especially through photochemical reactions of organic and hydroperoxy radical intermediates (RO 2 and HO 2 ) with nitric oxide (NO) (Chameides, 1978).
The nitrate radical (NO 3 ) arises from the oxidation of nitrogen dioxide (NO 2 ) by ozone (O 3 ) and occurs principally in the nighttime atmosphere due to its rapid photolysis in sunlight and its reaction with NO (Wayne et al., 1991;Brown and Stutz, 2012). The nitrate radical is a strong oxidant, reacting with a wide variety of volatile organic compounds, including alkenes, aromatics, and oxygenates as well as with reduced sulfur compounds. Reactions of NO 3 are particularly rapid with unsaturated compounds (alkenes) (Atkinson and Arey, 2003). BVOC such as isoprene, monoterpenes, and sesquiterpenes typically have one or more unsaturated functionalities such that they are particularly susceptible to oxidation by O 3 and NO 3 .
The potential for NO 3 to serve as a large sink for BVOC was recognized more than 3 decades ago . Field studies since that time have shown that in any environment with moderate to large BVOC concentrations, a majority of the NO 3 radical oxidative reactions are with BVOC rather than VOC of anthropogenic origin (Brown and Stutz, 2012). This interaction gives rise to a mechanism that couples anthropogenic NO x emissions with natural BVOC emissions Xu et al., 2015a). Although it is one of several such anthropogenic-biogenic interactions (Hoyle et al., 2011), reactions of NO 3 with BVOC are an area of intense current interest and one whose study has proven challenging. These challenges arise from the more limited current database of laboratory data for NO 3 oxidation reactions relative to those of other common atmospheric oxidants such as hydroxyl radical (OH) and O 3 . The mixing state of the night-time atmosphere and the limitations it imposes for characterization of nocturnal oxidation chemistry during field measurements and within atmospheric models present a second challenge to this field of research. Figure 1 illustrates these features of nighttime NO 3 -BVOC chemistry.
Reactions of NO 3 with BVOC have received increased attention in the recent literature as a potential source of secondary organic aerosol (SOA) . This SOA source is intriguing for several reasons. First, although organics are now understood to comprise a large fraction of total aerosol mass, and although much of these organics are secondary, sources of SOA remain difficult to characterize, in part due to a large number of emission sources and potential chemical mechanisms (Zhang et al., 2007;Hallquist et al., 2009;Jimenez et al., 2009;Ng et al., 2010). Analysis of aerosol organic carbon shows that a large fraction is modern, arising either from biogenic hydrocarbon emissions or biomass burning sources (e.g., Schichtel et al., 2008;Hodzic et al., 2010). Conversely, field data in regionally polluted areas indicate strong correlations between tracers of anthropogenic emissions and SOA, which suggests that anthropogenic influences can lead to production of SOA from modern (i.e., non-fossil) carbon (e.g., Weber et al., 2007). Model studies confirm that global observations are best simulated with a biogenic carbon source in the presence of anthropogenic pollutants (Spracklen et al., 2011). Reactions of NO 3 with BVOC are one such mechanism that may lead to anthropogenically influenced biogenic SOA (Hoyle et al., 2007), and it is important to quantify the extent to which such reactions can explain sources of SOA.
Second, some laboratory and chamber studies suggest that SOA yields from NO 3 oxidation of common BVOC, such as isoprene and selected monoterpenes, are greater than that for OH or O 3 oxidation Griffin et al., 1999;Spittler et al., 2006;Ng et al., 2008;Fry et al., 2009;Rollins et al., 2009;. However, among the monoterpenes, the SOA yields may be much more variable for NO 3 oxidation than for other oxidants, with anomalously low SOA yields in some cases and high SOA yields in others (Draper et al., 2015;Nah et al., 2016b).
Third, not only is NO 3 -BVOC chemistry a potentially efficient SOA formation mechanism, it is also a major pathway for the production of organic nitrates (von Kuhlmann et al., 2004;Horowitz et al., 2007), a large component of oxidized reactive nitrogen that may serve as either a NO x reservoir or NO x sink. Results from recent field measurements have shown that organic nitrates are important components of ambient OA (Day et al., 2010;Rollins et al., 2012;Fry et al., 2013;Ayres et al., 2015;Xu et al., 2015a, b;Kiendler-Scharr et al., 2016;Lee et al., 2016). Furthermore, within the last several years, the capability to measure both total and speciated gas-phase and particle-phase organic nitrates has been demonstrated (Fry et al., , 2013Rollins et al., 2010Rollins et al., , 2013Lee et al., 2016;Nah et al., 2016b). The life-times of organic nitrates derived from BVOC-NO 3 reaction with respect to hydrolysis, photooxidation, and deposition play an important role in the NO x budget and formation of O 3 and SOA. These processes appear to depend strongly on the parent VOC and oxidation conditions and must be better constrained for understanding organic nitrate lifetimes in the atmosphere Hu et al., 2011;Liu et al., 2012b;Pye et al., 2015;Rindelaub et al., 2015;Lee et al., 2016;Nah et al., 2016b).
Fourth, incorporation of SOA yields for NO 3 -BVOC reactions into regional and global models indicates that these reactions could be a significant, or in some regions even dominant, SOA contributor (Hoyle et al., 2007;Pye et al., 2010Pye et al., , 2015Chung et al., 2012;Fry and Sackinger, 2012;Kiendler-Scharr et al., 2016). Model predictions of organic aerosol formation from NO 3 -BVOC until recently have been difficult to verify directly from field measurements. Recent progress in laboratory and field studies have provided some of the first opportunities to develop coupled gas and particle systems to describe mechanistically and predict SOA and organic nitrate formation from NO 3 -BVOC reactions .
Finally, analyses from several recent field studies examining diurnal variation in the organic and/or nitrate content of aerosols conclude that nighttime BVOC oxidation through NO 3 radicals constitutes a large organic aerosol source Fry et al., 2013;Xu et al., 2015a, b;Kiendler-Scharr et al., 2016). Although such analyses may correct their estimates of aerosol production for the variation in boundary layer depth, field measurements at surface level are necessarily limited in their ability to accurately assess the atmospheric chemistry in the overlying residual layer or even the gradients that may exist within the relatively shallow nocturnal boundary layer Brown et al., 2007b). Thus, although there is apparent consistency between recent results from both modeling and field studies, the vertically stratified structure of the nighttime atmosphere makes such comparisons difficult to evaluate critically. There is a limited database of nighttime aircraft measurements that has probed this vertical structure with sufficient chemical detail to assess NO 3 -BVOC reactions (Brown et al., 2007a;Brown et al., 2009), and some of these data show evidence for an OA source related to this chemistry, especially at low altitude . A larger database of aircraft and/or vertically resolved measurements is required, however, for comprehensive comparisons to model predictions.
The purpose of this article is to review the current literature on the chemistry of NO 3 and BVOC to critically assess the current state of the science. The review focuses on BVOC emitted from terrestrial vegetation. The importance of NO 3 reactions with reduced sulfur compounds, such as dimethyl sulfide in marine ecosystems, is well known (Platt et al., 1990;Yvon et al., 1996;Allan et al., 1999Allan et al., , 2000Vrekoussis et al., 2004;Stark et al., 2007;Osthoff et al., 2009) but is outside of the scope of this review. Key uncertainties include chemical mechanisms, yields of major reaction products such as SOA and organic nitrogen, the potential for NO 3 and BVOC to interact in the ambient atmosphere, and the implications of that interaction for current understanding of air quality and climate. The review stems from an International Global Atmospheric Chemistry (IGAC) and US National Science Foundation (NSF) sponsored workshop of the same name held in June 2015 at the Georgia Institute of Technology (Atlanta, GA, USA). Following this introduction, Sect. 2 of this article reviews the current literature in several areas relevant to the understanding of NO 3 -BVOC atmospheric chemistry. Section 3 outlines perspectives on the implications of this chemistry for understanding climate and air quality, its response to current emission trends, and its relevance to implementation of control strategies. Finally, the review concludes with an assessment of the impacts of NO 3 -BVOC reactions on air quality, visibility, and climate. reactions of RO 2 . More specifically, for all chamber experiments, constraining the fate and lifetime of RO 2 is required to attribute product and SOA yields to a specific pathway. As shown in Table 2 in Sect. 2.2, the nitrate yields and SOA yields for NO 3 -induced degradation of many VOC vary significantly between different studies. This is likely, in part, a result of each experiment having a different distribution of RO 2 fates, but may also arise from vapor wall losses.
In general, there are very few mechanistic studies for NO 3 relative to other oxidants. Furthermore, the elucidation of mechanisms is limited by the fact that most studies provide overall yields of organic nitrates (without individual identification of the species) and/or identification (without quantification) due to the lack of standards.

Organic aerosol yields, speciation, and particle-phase chemistry
Several papers have reported chamber studies to measure the organic aerosol yield and/or gaseous and aerosol-phase oxidation product distribution from NO 3 -BVOC reactions. These are summarized in Table 2. In general, these experimental results show that monoterpenes are efficient sources of SOA, with reported yields variable but consistently above 20 %, with the notable exception of α-pinene (yields 0-15 %). This anomalous monoterpene also has a much larger product yield of carbonyls instead of organic nitrates compared to the others. This difference among monoterpenes was investigated in the context of the competition between O 3 and NO 3 oxidation (Draper et al., 2015), in which shifting from O 3 -dominated to NO 3 -dominated oxidation was observed to suppress SOA formation from α-pinene, but not from β-pinene, Δ-carene, or limonene. The smaller isoprene has substantially lower SOA yields (2-24 %), and the only sesquiterpene studied, β-caryophyllene, has a much larger yield (86-150 %) than the monoterpenes.
In general, these chamber experiments are conducted under conditions that focus on firstgeneration oxidation only, but further oxidation can continue to change SOA loadings in the real atmosphere (e.g., Rollins et al., 2009;Chacon-Madrid et al., 2013). Recent experiments showed that particulate organic nitrates formed from β-pinene-NO 3 are resilient to photochemical aging, while those formed from α-pinene-NO 3 evaporate more readily (Nah et al., 2016b).
Other chamber studies have not reported SOA mass yields or gas-phase product measurements but have otherwise demonstrated the importance of NO 3 -BVOC reactions to SOA production. These studies have identified β-pinene and Δ-carene as particularly efficient sources of SOA upon NO 3 oxidation (Hoffmann et al., 1997), confirmed the greater aerosol-forming potential from β-pinene versus α-pinene (Bonn and Moortgat, 2002), and reported Fourier transform infrared spectroscopy (FTIR) and aerosol mass spectrometry (AMS) measurements of the composition of organic nitrates detected in aerosol formed from NO 3 -isoprene, α-pinene, β-pinene, Δ-carene, and limonene reactions .
Relative humidity (RH) can be an important parameter, as it affects the competition between NO 3 -BVOC reactions and heterogeneous uptake of N 2 O 5 . Among existing laboratory studies, only a few have focused on the effect of RH on SOA formation from NO 3 -initiated oxidation (Bonn and Moortgat, 2002;Spittler et al., 2006;Fry et al., 2009;. The impact of RH might be important, especially at night and during the early morning when RH near the surface is high and NO 3 radical chemistry is competitive with O 3 and OH reactions. However, observations of the effect of water on SOA formation originating from NO 3 oxidation hint at a varied role. Spittler et al. (2006) reported lower SOA yields under humid conditions, but other studies did not observe a significant effect (Bonn and Moortgat, 2002;Fry et al., 2009;. Among the important effects of water is its role as a medium for hydrolysis. In laboratory studies, primary and secondary organic nitrates were found to be less prone to aqueous hydrolysis than tertiary organic nitrates Hu et al., 2011). First-generation organic nitrates retaining double bonds may also hydrolyze relatively quickly, especially in the presence of acidity (Jacobs et al., 2014;Rindelaub et al., 2015). Depending on the relative amount of these different types of organic nitrates, the overall hydrolysis rate could be different for organic nitrates formed from NO 3 oxidation and photooxidation in the presence of NO x . Recently, there has been increasing evidence from field measurements that organic nitrates hydrolyze in the particle phase, producing HNO 3 (Liu et al., 2012b;Browne et al., 2013). This has been only a limited focus of chamber experiments to date . In addition to the effect of RH, particle-phase acidity is known to affect SOA formation from ozonolysis and OH reaction (e.g., Gao et al., 2004;Tolocka et al., 2004). Thus far, only one study has examined the effect of acidity on NO 3 -initiated SOA formation and found a negligible effect . Notably, an effect of acidity was observed for the hydrolysis of organic nitrates produced in photochemical reactions (Szmigielski et al., 2010;Rindelaub et al., 2015). While much organic nitrate aerosol is formed via NO 3 + BVOC reactions, some fraction can also form from RO 2 + NO chemistry. Rollins et al. (2010) observed the organic nitrate moiety in 6-15 % of total SOA mass generated from high-NO x photooxidation of limonene, α-pinene, Δ 3 -carene, and tridecane. A very recent study of Berkemeier et al. (2016) showed that organic nitrates accounted for ~ 40 % of SOA mass during initial particle formation in α-pinene oxidation by O 3 in the presence of NO, decreasing to ~ 15 % upon particle growth to the accumulation-mode size range. They also observed a tight correlation (R 2 = 0.98) between organic nitrate content and SOA particle number concentrations. This implies that organic nitrates may be among the extremely low volatility organic compounds (ELVOC) Tröstl et al., 2016) that play a critical role in nucleation and nanoparticle growth.

Heterogeneous and aqueous-phase NO 3 processes
The NO 3 radical is not only a key nighttime oxidant of organic (and especially biogenic) trace gases but it can also play an important role in the aqueous phase of tropospheric clouds and deliquesced particles (Chameides, 1978;Wayne et al., 1991;Herrmann and Zellner, 1998;Rudich et al., 1998). Whilst the reaction of NO 3 with organic particles and aqueous droplets in the atmosphere is believed to represent only an insignificant fraction of the overall loss rate for NO 3 , it can have a substantial impact on the chemical and physical properties of the particle by modifying its lifetime, oxidation state, viscosity, and hygroscopic properties and thus its propensity to act as a cloud condensation nucleus (Rudich, 2003).
Biogenic VOC include, but are not limited to the isoprenoids (isoprene, mono-, and sesquiterpenes) as well as alkanes, alkenes, carbonyls, alcohols, esters, ethers, and acids (Kesselmeier and Staudt, 1999). Recent measurements indicate that biogenic emissions of aromatic trace gases are also significant . The gas-phase degradation of BVOC leads to the formation of a complex mixture of organic trace gases including hydroxyl-and nitrate-substituted oxygenates which can transfer to the particle phase by condensation or dissolution. Our present understanding is that non-anthropogenic SOA has a large contribution from isoprenoid degradation.
As is generally the case for laboratory studies of heterogeneous processes, most of the experimental investigations on heterogeneous uptake of NO 3 to organic surfaces have dealt with single-component systems that act as surrogates for the considerably more complex mixtures found in atmospheric SOA. A further level of complexity arises when we consider that initially reactive systems, e.g., containing condensed or dissolved unsaturated hydrocarbons, can become deactivated as SOA ages, single bonds replace double bonds, and the oxygen-to-carbon ratio increases.
We summarize the results of the laboratory studies to provide a rough guide to NO 3 reactivity on different classes of organics which may be present in SOA and note that further studies of NO 3 uptake to biogenic SOA which was either generated and aged under welldefined conditions  or sampled from the atmosphere are required to confirm predictions of uptake efficiency based on the presently available database.

2.3.1
Heterogeneous processes-For some particle-phase organics, the reaction with NO 3 is at least as important as other atmospheric oxidants such as O 3 and OH (Shiraiwa et al., 2009;Kaiser et al., 2011). The lifetime (τ) of a single component, liquid organic particle with respect to loss by reaction with NO 3 at concentration [NO 3 ] is partially governed by the uptake coefficient (γ) : where D p is the particle diameter, ρ org and M org are the density and molecular weight of the organic component, respectively, c̄ is the mean molecular velocity of gas-phase NO 3 , and N A is Avogadro number. Thus, defined, τ is the time required for all the organic molecules in a spherical (i.e., liquid) particle to undergo a single reaction with NO 3 .
Recent studies have shown that organic aerosols can adopt semi-solid (highly viscous) or amorphous solid (crystalline or glass) phase states, depending on the composition and ambient conditions (Virtanen et al., 2010;Koop et al., 2011;Renbaum-Wolff et al., 2013). Typically, the bulk phase diffusion coefficients of NO 3 are ~ 10 −7 -10 −9 cm 2 s −1 in semisolid and ~ 10 −10 cm 2 s −1 in solids . Slow bulk diffusion of NO 3 in a viscous organic matrix can effectively limit the rate of uptake Shiraiwa et al., 2012). Similarly, the solubility may be different in a concentrated, organic medium. If bulk diffusion is slow, the reaction may be confined to the near-surface layers of the particle or bulk substrate. The presence of organic coatings on aqueous aerosols was found to suppress heterogeneous N 2 O 5 hydrolysis by providing a barrier through which N 2 O 5 needs to diffuse to undergo hydrolysis Cosman et al., 2008;Grifiths et al., 2009). Reactive uptake by organic aerosols is expected to exhibit a pronounced decrease at low RH and temperature, owing to a phase transition from viscous liquid to semi-solid or amorphous solid (Arangio et al., 2015). Therefore, the presence of a semi-solid matrix may effectively shield reactive organic compounds from chemical degradation in long-range transport in the free troposphere.
To get an estimate of the processing rate of BVOC-derived SOA, we have summarized the results of several laboratory studies to provide a rough guide to NO 3 reactivity on different classes of organics that may be present in SOA (Fig. 3). A rough estimate of the reactivity of NO 3 to freshly generated, isoprenoid-derived SOA, which still contains organics with double bonds (e.g., from diolefinic monoterpenes such as limonene), may be obtained by considering the data on alkenes and unsaturated acids, where the uptake coefficient is generally close to 0.1.
The classes of organics for which heterogeneous reactions with NO 3 have been examined are alkanoic/alkenoic acids, alkanes and alkenes, alcohols, aldehydes, polyaromatic hydrocarbons (PAHs), and secondary organic aerosols. Laboratory studies have used either pure organic substrates, with the organic of interest internally mixed in an aqueous particle; as a surface coating, with the reactive organic mixed in a nonreactive organic matrix; or in the form of self-assembling monolayers. The surrogate surface may be available as a macroscopic bulk liquid (or frozen liquid) or in particulate form and both gas-phase and particle-phase analyses have been used to derive kinetic parameters and investigate products formed.
In the gas phase, the NO 3 radical reacts slowly (by H-abstraction) with alkanes, more rapidly with aldehydes due to the weaker C-H bond of the carbonyl group, and most readily with alkenes and aromatics via electrophilic addition. This trend in reactivity is also observed in the condensed-phase reactions of NO 3 with organics so that long-chain organics, for which non-sterically hindered addition to a double bond is possible, and aromatics are the most reactive. In very general terms, uptake coefficients are in the range of 1-10 × 10 −3 for alkanes, alcohols, and acids without double bonds, 2-200 × 10 −3 for alkenes with varying numbers of double bonds, 3-1000 × 10 −3 for acids with double bonds again depending on the number of double bonds, and 100-500 × 10 −3 for aromatics. These trends are illustrated in Fig. 3 which plots the experimental data for the uptake of NO 3 to single-component organic surfaces belonging to different classes of condensable organics. Condensed-phase organic nitrates have been frequently observed following interaction of NO 3 with organic surfaces (see below).
Saturated hydrocarbons: Uptake of NO 3 to saturated hydrocarbons is relatively slow, with uptake coefficients close to 10 −3 . Moise et al. (2002) found that (for a solid sample) uptake to a branched-chain alkane was more efficient than for a straight-chain alkane, which is consistent with known trends in gas-phase reactivity of NO 3 . The slow surface reaction with alkanes enables both surface and bulk components of the reaction to operate in parallel. The observation of RONO 2 as product is explained (Knopf et al., 2006;) by processes similar to those proceeding in the gas phase, i.e., abstraction followed by formation of peroxy and alkoxy intermediates which react with NO 2 and NO 3 to form the organic nitrate.

Unsaturated hydrocarbons:
With exception of the data of Moise et al. (2002), the up-take of NO 3 to an unsaturated organic surface is found to be much more efficient than to the saturated analogue. The NO 3 uptake coefficient for, e.g., squalene, is at least an order of magnitude more efficient than for squalane Lee et al., 2013). The location of the double bond is also important and the larger value for γ found for a selfassembling monolayer of NO 3 + undec-10-ene-1-thiol compared to liquid, long-chain alkenes is due to the fact that the terminal double bond is located at the interface and is thus more accessible for a gas-phase reactant . NO 3 uptake to mixtures of unsaturated methyl oleate in a matrix of saturated organic was found to be consistent with either a surface or bulk reaction . The formation of condensedphase organic nitrates and simultaneous loss of the vinyl group indicates that the reaction proceeds, as in the gas phase, by addition of NO 3 to the double bond followed by reaction of NO 3 (or NO 2 ) with the resulting alkyl and peroxy radicals formed (Zhang et al., 2014b).

Saturated alcohols and carbonyls:
Consistent with reactivity trends for NO 3 in the gas phase, the weakening of some C-H bonds in oxidized, saturated organics results in a more efficient interaction of NO 3 than for the non-oxidized counterparts although, as far as the limited dataset allows trends to be deduced, the gas-phase reactivity trend of polyalcohol being greater than alkanoate appears to be reversed in the liquid phase . For multicomponent liquid particles, the uptake coefficient will also depend on the particle viscosity (Iannone et al., 2011) though it has not been clearly established if the reaction proceeds predominantly at the surface or throughout the particle (Iannone et al., 2011). The reaction products are expected to be formed via similar pathways as seen in the gas phase, i.e., abstraction of the aldehydic-H atom for aldehydes and abstraction of an H atom from either the O-H or adjacent α-CH 2 group for alcohols prior to reaction of NO 2 and NO 3 with the ensuing alkyl and peroxy radicals (Zhang and Morris, 2015).

Organic acids:
The efficiency of uptake of NO 3 to unsaturated acids is comparable to that found with other oxidized, saturated organics (Moise et al., 2002) suggesting that the reaction proceeds, as in the gas phase, via abstraction rather than addition. Significantly larger uptake coefficients have been observed for a range of unsaturated, long-chain acids, with γ often between 0.1 and 1 Knopf et al., 2011;Zhao et al., 2011a). γ depends on the number and position (steric factors) of the double bond. For example, the uptake coefficient for abietic acid is a factor of 100 lower than for linoleic acid . The condensed-phase products formed in the interaction of NO 3 with unsaturated acids are substituted carboxylic acids, including hydroxy nitrates, carbonyl nitrates, dinitrates, and hydroxy dinitrates (Hung et al., 2005;Docherty and Ziemann, 2006;McNeill et al., 2007;Zhao et al., 2011a).

Aromatics:
The interaction of NO 3 with condensed-phase aromatics and PAHs results in the formation of a large number of nitrated aromatics and nitro PAHs. Similar to the gas-phase mechanism, the reaction is initiated by addition of NO 3 to the aromatic ring, followed by breaking of an N-O bond to release NO 2 to the gas phase and forming a nitrooxycyclohexadienyl-type radical which can further react with O 2 , NO 2 , or undergo internal rearrangement to form hydroxyl species (Gross and Bertram, 2008;Lu et al., 2011). The uptake coefficients are large and comparable to those derived for the unsaturated fatty acids.
The literature results on the interaction of NO 3 with organic substrates are tabulated in Table  S1 in the Supplement, in which the uptake coefficient is listed (if available) along with the observed condensed-and gas-phase products. in aqueous particles is largely a result of transfer from the gas phase Tilgner et al., 2013). Concentrations of NO 3 in tropospheric aqueous solutions cannot be measured in situ, and literature values are based on multiphase model predictions (Herrmann et al., 2010). Model studies with the chemical aqueous-phase radical mechanism (CAPRAM; Herrmann et al., 2005;Tilgner et al., 2013) predict [NO 3 ] between 1 × 6 × 10 −16 and 2.7 × 10 −13 mol L −1 . High NO 3 concentration levels are associated with urban clouds, while in rural and marine clouds these levels are an order of magnitude lower. Since the NO 3 concentrations are related to the NO x budget, typically higher NO 3 concentrations are present under urban cloud conditions compared to rural and marine cloud regimes.

Aqueous
NO 3 radicals react with dissolved organic species via three different pathways: (i) by Hatom abstraction from saturated organic compounds, (ii) by electrophilic addition to double bonds within unsaturated organic compounds, and (iii) by electron transfer from dissociated organic acids (Huie, 1994;Herrmann and Zellner, 1998). For a detailed overview on aqueous-phase NO 3 radical kinetics, the reader is referred to several recent summaries (Neta et al., 1988;Herrmann and Zellner, 1998;Ross et al., 1998;Herrmann, 2003;Herrmann et al., 2010Herrmann et al., , 2015. Compared to the highly reactive and non-selective OH radical, the NO 3 radical is characterized by a lower reactivity and represents a more selective aqueous-phase oxidant. The available kinetic data indicate that the reactivity of NO 3 radicals with organic compounds in comparison to the two other key radicals (OH, SO 4 − ) is as follows: (Herrmann et al., 2015).
In Table S2, we list kinetic parameters for reaction of NO 3 with aliphatic organic compounds as presently incorporated in the CAPRAM database (Bräuer et al., 2017). Typical ranges of rate constants (in M −1 s −1 ) for reactions of NO 3 in the aqueous phase are 106-107 for saturated alcohols, carbonyls, and sugars; 10 4 -10 6 for protonated aliphatic mono-and dicarboxylic acids, with higher values for oxygenated acids; 10 6 -10 8 for deprotonated aliphatic mono-and dicarboxylic acids (higher values typically for oxygenated acids); 10 7 -10 9 for unsaturated aliphatic compounds; and 10 8 -2 × 10 9 for aromatic compounds (without nitro/acid functionality). The somewhat larger rate constants for deprotonated aliphatic mono-and dicarboxylic acids, unsaturated aliphatic compounds and aromatic compounds is related to the occurrence of electron transfer reactions and addition reaction pathways, which are often faster than H-abstraction reactions.
Many aqueous-phase NO 3 reaction rate constants, even for small oxygenated organic compounds, are not available in the literature and have to be estimated. In the absence of SARs for NO 3 radical reactions with organic compounds, Evans-Polanyi-type reactivity correlations are used to predict kinetic data for H-abstraction NO 3 radical reactions. The latest correlation for NO 3 where BDE is the bond dissociation energy (in kJ mol −1 ). The correlation is quite tight, with a correlation coefficient of R = 0.9.
A direct comparison of the aqueous-phase OH and NO 3 radical rate constants (k 298 K ) of organic compounds from different compound classes is presented in Fig. 4, which shows that the NO 3 radical reaction rate constants for many organic compounds are about 2 orders of magnitude smaller than respective OH rate constants. In contrast, deprotonated dicarboxylic acids can react with NO 3 via electron transfer and have similar rate constants for OH reaction. Rate constants for OH and NO 3 with alcohols and diols/polyols are well correlated (R 2 values are given in Table S3), whereas those rate constants for carbonyl compounds and diacids have a lower degree of correlation. Figure 4b shows a comparison of the modeled chemical turnovers of reactions of organic compounds with OH versus NO 3 radicals distinguished for different compound classes. The simulations were performed with the SPACCIM model (Wolke et al., 2005) for the urban summer CAPRAM scenario (see Tilgner et al., 2013 for details) using the master chemical mechanism (MCM) 3.2/CAPRAM 4.0 mechanism Bräuer et al., 2017) which has in total 862 NO 3 radical reactions with organic compounds.
Most of the data lie under the 1 : 1 line, indicating that, for most of the organic compounds considered, chemical degradation by OH is more important than by NO 3 , with a significant fraction of the data lying close to a 10 : 1 line, though OH fluxes sometimes exceed NO 3 fluxes by a factor of 10 3 -10 4 . Approximate relative flux ratios (NO 3 / OH) for different classes of organic are 10 −1 -10 −2 for alcohols (including diols and polyols) and carbonyl compounds, 10 −1 -10 −4 for undissociated monoacids and diacids, ~ 1 (or larger) for dissociated monoacids, 10 −2 -> 10 for dissociated diacids, and 10 −2 -1 for organic nitrates. For carboxylate ions, NO 3 -initiated electron transfer is thus the dominant oxidation pathway. As OH-initiated oxidation proceeds via an H-abstraction, high NO 3 -OH flux ratios can be observed for carboxylate ions but not for protonated carboxylic acids.
Overall, Fig. 4b shows that, over a 4-day summer cycle, NO 3 radical reactions can compete with OH radical reactions in particular for protonated carboxylic acids and multifunctional compounds. Nevertheless, aqueous NO 3 radical reactions with organics will become more important during winter or at higher latitudes, where photochemistry as the main source of OH is less important. Finally, it should be noted that NO 3 aqueous-phase nighttime chemistry will influence the concentration levels of many aqueous-phase reactants available for reaction during the next day.

Instrumental methods
Atmospheric models of the interaction of NO 3 with BVOC rely on experimental data gathered in both the laboratory and the field. These experimental data are used to define model parameters and to evaluate model performance by comparison to observed quantities. Instrumentation for measurements of nitrogen-containing species, oxidants, and organic compounds, including NO x , O 3 , NO 3 , BVOC, and oxidized reactive nitrogen compounds, are all important to understand the processes at work. Of particular importance to the subject of this review is the characterization of organic nitrates, which are now known to exist in both the gas and particle phases and whose atmospheric chemistry is complex. This section reviews historical and current experimental methods used for elucidating NO 3 -BVOC atmospheric chemistry.

Nitrate radical measurements-
Optical absorption spectroscopy has been the primary measurement technique for NO 3 . It usually makes use of two prominent absorption features of NO 3 near 623 and 662 nm. Note that the dissociation limit of the NO 3 molecule lies between the two absorption lines (Johnston et al., 1996); thus, illumination by measurement radiation at the longer wave-length band does not lead to photolysis of NO 3 . The room temperature absorption cross section of NO 3 at 662 nm is ~ 2 × 10 −17 cm 2 molec −1 and increases at lower temperature (Yokelson et al., 1994;Osthoff et al., 2007). Thus, at a typical minimum detectable optical density (reduction of the intensity compared to no absorption) and a light-path length of 5 km, a detection limit of 10 7 molec cm −3 or ~ 0.4 ppt (under standard conditions) is achieved.
Initial measurements of NO 3 in the atmosphere were long-path averages using light paths between either the sun or the moon (e.g., Noxon et al., 1978) and the receiving spectrometer (also called passive techniques because natural light sources were used) or between an artificial light source and spectrometer over a distance of several kilometers (active techniques; e.g., Platt et al., 1980). Passive techniques were later extended to yield NO 3 vertical profiles (e.g., Weaver et al., 1996). In recent years, resonator cavity techniques allowed construction of very compact instruments capable of performing in situ measurements of NO 3 with absorption spectroscopy (see in situ measurement techniques below).
An important distinction between the techniques is whether NO 3 can be deliberately or inadvertently removed from the absorption path as part of the observing strategy. Long-path absorption spectroscopy does not allow control over the sample for obtaining a zero background by removing NO 3 (Category 1). Resonator techniques (at least as long as the resonator is encased) allow deliberate removal of NO 3 from the absorption path as part of the measurement sequence and may also result in inadvertent removal during sampling (Category 2).
For instruments of Category 1, the intensity without absorber (I 0 ) cannot be easily detected. Therefore, the information about the absorption due to NO 3 (and any other trace gas) has to be determined from the structure of the absorption, which is usually done by using differential optical absorption spectroscopy (DOAS) (Platt and Stutz, 2008), which relies on the characteristic fingerprint of the NO 3 absorption structure in a finite wavelength range (about several 10 nm wide). Thus, a spectrometer of sufficient spectral range and resolution (around 0.5 nm) is required.
Instruments of Category 2 can determine the NO 3 concentration from the difference (or rather log of the ratio) of the intensity with and without NO 3 in the measurement volume. In this case, only an intensity measurement at a single wavelength (typically of a laser) is necessary, and specificity can be achieved through chemical titration with NO (Brown et al., 2001). However, enhanced specificity without chemical titration can be gained by combining resonator techniques with DOAS detection. It should be noted that the advantage of a closed cavity to be able to remove (or manipulate) NO 3 comes at the expense of potential wall losses, which have to be characterized. Such instruments have the advantage of being able to also detect N 2 O 5 , which is in thermal equilibrium with NO 3 and can be quantitatively converted to NO 3 by thermal dissociation (Brown et al., 2001. Another complication arises from the presence of water vapor and oxygen lines in the wavelength range of strong NO 3 absorptions. To compensate for these potential interferences in open-path measurements (where NO 3 cannot easily be removed), daytime measurements are frequently used as reference because NO 3 levels are typically very low (but not necessarily negligibly low) (Geyer et al., 2003). Thus, a good fraction of the reported NO 3 data (in particular, older data) represents day-night differences.
Passive long-path remote sensing techniques: Measurements of the NO 3 absorption structure using sunlight take advantage of the fact that NO 3 is very quickly photolyzed by sunlight (around 5 s lifetime during the day) allowing for vertically resolved measurements during twilight (e.g., Aliwell and Jones, 1998;Allan et al., 2002;Coe et al., 2002;von Friedeburg et al., 2002). The fact that the NO 3 concentration is nearly zero due to rapid photolysis in the directly sunlit atmosphere, while it is largely undisturbed in a shadowed area, can be used to determine NO 3 vertical concentration profiles during sunrise using the moon as a light source (Smith and Solomon, 1990;Smith et al., 1993;Weaver et al., 1996). Alternatively, the time series of the NO 3 column density derived from scattered sunlight originating from the zenith (or from a viewing direction away from the sun) during sunrise can be evaluated to yield NO 3 vertical profiles (Allan et al., 2002;Coe et al., 2002;von Friedeburg et al., 2002).
Nighttime NO 3 total column data have been derived by spectroscopy of moonlight and starlight (Naudet et al., 1981), the intensity of which is about 4-5 orders of magnitude lower than that of sunlight. Thus, photolysis of NO 3 by moonlight is negligible. A series of moonlight NO 3 measurements have been reported (Noxon et al., 1980;Noxon, 1983;Sanders et al., 1987;Solomon et al., 1989Solomon et al., , 1993Aliwell and Jones, 1996a, b;Wagner et al., 2000). These measurements yield total column data of NO 3 , the sum of tropospheric and stratospheric partial columns. Separation between stratospheric and tropospheric NO 3 can be accomplished (to some extent) by the Langley plot method (Noxon et al., 1980), which takes advantage of the different dependence of tropospheric and stratospheric NO 3 slant column density on the lunar zenith angle.
Active long-path techniques: A large number of NO 3 measurements have been made using the active long-path DOAS technique (Platt et al., 1980(Platt et al., , 1981(Platt et al., , 1984Pitts et al., 1984;Heintz et al., 1996;Allan et al., 2000;Martinez et al., 2000;Geyer et al., 2001aGeyer et al., , b, 2003Gölz et al., 2001;Stutz et al., 2002Stutz et al., , 2004Stutz et al., , 2010Asaf et al., 2009;McLaren et al., 2010;Crowley et al., 2011;Sobanski et al., 2016). Here, a searchlight-type light source is used to transmit a beam of light across a kilometer-long light path in the open atmosphere to a receiving telescope-spectrometer combination. The light source typically is a broadband thermal radiator (incandescent lamp, Xe arc lamp, laser-driven light source). More recently, LED light sources were also used (Kern et al., 2006). The telescope (around 0.2 m diameter) collects the radiation and transmits it, usually through an optical fiber, into the spectrometer, which produces the absorption spectrum. Modern instruments now almost exclusively use transmitter/receiver combinations at one end of the light path and retro-reflector arrays (e.g., cat-eye-like optical devices) at the other end. The great advantage of this approach is that power and optical adjustment is only required at one end of the light path while the other end (with the retro-reflector array) is fixed. In this way, several retro-reflector arrays, for instance, mounted at different altitudes, can be used sequentially with the same transmitter/ receiver unit allowing determination of vertical profiles of NO 3 (and other species measurable by DOAS) (Stutz et al., 2002. In situ measurement techniques: Cavity ring-down spectroscopy (CRDS) and cavityenhanced absorption spectroscopy (CEAS) are related techniques for in situ quantification of atmospheric trace gases such as NO 3 . These methods are characterized by high sensitivity, specificity, and acquisition speed (Table 3a), and they allow for spatially resolved measurements on mobile platforms.
In CRDS, laser light is "trapped" in a high-finesse stable optical cavity, which usually consists of a pair of highly reflective spherical mirrors in a near-confocal arrangement. The concentrations of the optical absorbers present within the resonator are derived from the Beer-Lambert law and the rate of light leaking from the cavity after the input beam has been switched off (O'Keefe and Deacon, 1988). CRDS instruments are inherently sensitive as they achieve long effective optical absorption paths (up to, or in some cases exceeding, 100 km) as the light decay is monitored for several 100 μs, and the absorption measurement is not affected by laser intensity fluctuations. For detection of NO 3 at 662 nm, pulsed laser sources such as Nd:YAG pumped dye lasers have been used because of the relative ease of coupling the laser beam to the optical cavity (Brown et al., , 2003Dubé et al., 2006). Relatively lower cost continuous-wave (cw) diode laser modules that are easily modulated also have been popular choices (e.g., King et al., 2000;Simpson, 2003;Ayers et al., 2005; Odame-Ankrah and Osthoff, 2011;Wagner et al., 2011).
In a CEAS instrument (also referred to as integrated cavity output spectroscopy, ICOS, or cavity-enhanced DOAS, CE-DOAS), the spectrum transmitted through a high-finesse optical cavity is recorded. Mixing ratios of the absorbing gases are derived using spectral retrieval routines similar to those used for open-path DOAS (e.g., O'Keefe, 1998O'Keefe, , 1999Ball et al., 2001;Fiedler et al., 2003;Platt et al., 2009;Schuster et al., 2009).
CRDS and CEAS are, in principle, absolute measurement techniques and do not need to rely on external calibration. In practice, however, chemical losses can occur on the inner walls of the inlet (even when constructed from inert materials such as Teflon) or at the aerosol filters necessary for CRDS instruments. Hence, the inlet transmission efficiencies have to be monitored for measurements to be accurate (Fuchs et al., 2008(Fuchs et al., , 2012Odame-Ankrah and Osthoff, 2011). On the other hand, a key advantage of in situ instruments over open-path instruments is that the sampled air can be manipulated. Deliberate addition of excess NO to the instrument's inlet titrates NO 3 and allows measurement of the instrument's zero level and separation of contributions to optical extinction from other species, such as NO 2 , O 3 , and H 2 O. Adding a heated section to the inlet (usually in a second detection channel) enables (parallel) detection of N 2 O 5 via the increase in the NO 3 signal (Brown et al., 2001;Simpson, 2003).
In addition, non-optical techniques have been used to detect and quantify NO 3 . Chemical ionization mass spectrometry (CIMS) is a powerful method for sensitive, selective, and fast quantification of a variety of atmospheric trace gases (Huey, 2007). NO 3 is readily detected after reaction with iodide reagent ion as the nitrate anion at m/z 62; at this mass, however, there are several known interferences, including dissociative generation from N 2 O 5 , HNO 3 , and HO 2 NO 2 (Slusher et al., 2004;Abida et al., 2011;Wang et al., 2014). There has been more success with the quantification of N 2 O 5 , usually as the iodide cluster ion at m/z 235 (Kercher et al., 2009), though accurate N 2 O 5 measurement at m/z 62 has been reported from recent aircraft measurements with a large N 2 O 5 signal (Le Breton et al., 2014).
Two groups have used laser-induced fluorescence (LIF) to quantify NO 3 (and N 2 O 5 through thermal dissociation) in ambient air Matsumoto et al., 2005a, b). The major drawback of this method is the relatively low fluorescence quantum yield of NO 3 , and hence the method has not gained wide use.
Another technique that was demonstrated to be capable of measuring NO 3 radicals at atmospheric concentration is matrix isolation electron spin resonance (MIESR) (Geyer et al., 1999). Although the technique allows simultaneous detection of other radicals (including HO 2 and NO 2 ), it has not been used extensively, probably because of its complexity.
Recently, a variety of in situ NO 3  and N 2 O 5 (Fuchs et al., 2012) measurement techniques were compared at the SAPHIR chamber in Jülich, Germany. All instruments measuring NO 3 were optically based (absorption or fluorescence). N 2 O 5 was detected as NO 3 after thermal decomposition in a heated inlet by either CRDS or LIF. Generally, agreement within the accuracy of instruments was found for all techniques detecting NO 3 and/or N 2 O 5 in this comparison exercise. This study showed excellent agreement between the instruments on the single-digit ppt NO 3 and N 2 O 5 levels with no noticeable interference due to NO 2 and water vapor for instruments based on cavity ringdown or cavity-enhanced spectroscopy. Because of the low sensitivity of LIF instruments, N 2 O 5 measurements by these instruments were significantly noisier compared to the measurements by cavity-enhanced methods. The agreement between instruments was less good in experiments with high aerosol mass loadings, specifically for N 2 O 5 , presumably due to enhanced, unaccounted loss of NO 3 and N 2 O 5 demonstrating the need for regular filter changes in closed-cavity instruments. Whereas differences between N 2 O 5 measurements were less than 20 % in the absence of aerosol, measurements differed up to a factor of 2.5 for the highest aerosol surface concentrations of 5 × 10 8 nm 2 cm −3 . Also, differences between NO 3 measurements showed an increasing trend (up to 50 %) with increasing aerosol surface concentration for some instruments. Perring et al. (2013). Sample collection techniques for organic nitrates include preconcentration on solid adsorbents (Atlas and Schauffler, 1991;Schneider and Ballschmiter, 1999;Grossenbacher et al., 2001), cryogenic trapping (Flocke et al., 1991) or collection in stainless steel canisters (Flocke et al., 1998;Blake et al., 1999), or direct sampling (Day et al., 2002;Beaver et al., 2012).

Gas-phase organic nitrate measurements-Analytical techniques to detect gaseous organic nitrates have been documented in a recent review by
The approaches to the analysis of the organic nitrates fall into three broad categories. First, one or more chemically speciated organic nitrates are measured by a variety of techniques including liquid chromatography (LC)  or gas chromatography (GC) with electron capture detection (Fischer et al., 2000), GC with electron impact or negativeion chemical ionization mass spectrometry (GC-MS) (Atlas, 1988;Luxenhofer et al., 1996;Blake et al., 1999Blake et al., , 2003aWorton et al., 2008), GC followed by conversion to NO and chemiluminescent detection (Flocke et al., 1991(Flocke et al., , 1998, GC followed by photoionization mass spectrometry (Takagi et al., 1981), GC followed by conversion of organic nitrates to NO 2 and luminol chemiluminescent detection (Hao et al., 1994), CIMS (Beaver et al., 2012;Paulot et al., 2012), and proton transfer reaction MS (PTR-MS) (Perring et al., 2009). Second, the sum of all organic nitrates can be measured directly by thermal dissociation to NO 2 , which is subsequently measured by LIF (TD-LIF) (Day et al., 2002), CRDS (TD-CRDS) (Paul et al., 2009;Thieser et al., 2016), or cavity-attenuated phase shift spectroscopy (TD-CAPS) (Sadanaga et al., 2016). Finally, the sum of all organic nitrates can be measured indirectly as the difference between all reactive NO x except for organic nitrates and total oxidized nitrogen (NO y ) (Parrish et al., 1993).
Recent advances in adduct ionization utilize detection of the charged cluster of the parent reagent ion with the compound of interest. This scheme is then coupled to high-resolution time-of-flight (HR-ToF) mass spectrometry. The combination of these methods allows the identification of molecular composition due to the soft ionization approach that minimizes fragmentation. Multifunctional organic nitrates resulting from the oxidation of BVOC have been detected using CF 3 O − Nguyen et al., 2015;Schwantes et al., 2015;Teng et al., 2015) and iodide as reagent ions (Lee et al., 2014aXiong et al., 2015Xiong et al., , 2016Nah et al., 2016b).

Online analysis of particulate matter-Total
(organic plus inorganic) mass of particulate nitrates is routinely quantified using online AMS (Jayne et al., 2000;Allan et al., 2004), from which the mass of organic nitrates can be obtained by three techniques. First, the NO + /NO 2 + ratio (or NO 2 + /NO + ratio) in the mass spectra is used to distinguish organic from inorganic nitrates (Fry et al., , 2013Farmer et al., 2010;Xu et al., 2015b;Kiendler-Scharr et al., 2016). It is noted that the NO 2 + /NO + approaches zero in the case of low or nonexistent NO 2 + signal, while NO + /NO 2 + gives large numbers. Second, positive matrix factorization (PMF) of data matrices including the NO + and NO 2 + ions in addition to organic ions (Sun et al., 2012;Hao et al., 2014;Xu et al., 2015b) is used. Third, the particulate inorganic nitrate concentration, as measured by an independent method such as ion chromatography, is subtracted from the total particulate nitrate concentration (Schlag et al., 2016;Xu et al., 2015a, b). A detailed comparison of these three methods is presented in Xu et al. (2015b). As the NO + /NO 2 + ratio in AMS data is dependent on instruments and the types of nitrates (inorganic and organic nitrates from different VOC oxidations), different strategies were developed when using this method to estimate particulate organic nitrates (Fry et al., 2013;Xu et al., 2015b).
A specialized inlet that selectively scrubs gaseous organic nitrates or collects particulate mass on a filter has been coupled to some of the techniques summarized in this section and utilized to observe particulate organic nitrates in the ambient atmosphere and laboratory studies. A TD-LIF equipped with a gas-scrubbing denuder (Rollins et al., 2010   The LC/(−)ESI-MS technique played a crucial role in relating the formation of organosulfates (OS) and nitrooxy-organosulfates (NOS) to NO 3 -initiated oxidation of BVOC in laboratory-generated and ambient SOA. Since these earlier works, a number of studies have reported the presence of OS and/or NOS compounds in ambient samples (Table  S4), though most studies do not connect these compounds explicitly to the NO 3 oxidation of BVOC. It should be noted that the direct infusion (−)ESI-MS technique rather than LC/ (−)ESI-MS is often used for the analysis of fog, rainwater, and cloud water samples as diluted liquid water samples can be injected into the ion source directly without a sample pretreatment procedure. However, caution is warranted for the direct infusion technique because it cannot separate isobaric isomers and it is susceptible to ion suppression, especially from the presence of inorganic ions in the samples.

Offline analysis of particulate matter-Owing
Whilst the LC or direct infusion (−)ESI-MS techniques have been successfully applied for the detection of the oxidation products from NO 3 -BVOC reactions, the techniques have been less successful in quantifying these compounds, mainly due to the lack of authentic standard compounds. The synthesis of these compounds should be a priority for future studies.
Finally, total organic nitrate functional groups within the particle phase have been quantified in ambient air using FTIR of particles collected on ZnSe impaction disks (low-pressure cascade impactor, size segregated) or Teflon filters (PM 1 ) (Mylonas et al., 1991;Garnes and Allen, 2002;Day et al., 2010). The organic nitrate content of particles can be quantified offline as well by collection on quartz fiber filters, extraction into solution (e.g., with wateracetonitrile mixtures), and analysis using standard wet chemistry techniques such as highpressure liquid chromatography coupled to electrospray ionization mass spectrometry (HPLC-ESI-MS) (Angove et al., 2006;Perraud et al., 2010;Draper et al., 2015).

Field observations
This section surveys the current literature on field observations of nitrate radicals and BVOC (Sect. 2.5.1), and organic nitrate aerosol attributable to NO 3 -BVOC chemistry (Sect. 2.5.2).

Nitrate radicals and BVOC-
A few years after the first measurement of tropospheric NO 3 (Noxon et al., 1980;Platt et al., 1980), it was recognized that the nitrate radical is a significant sink for BVOC, especially monoterpenes in terrestrial ecosystems and dimethyl sulfide (DMS) in maritime air influenced by continental NO x sources . The conclusion was based upon computer simulations using NO 3 concentrations measured in field studies in the western US and Europe, and measured rate constants of NO 3 with olefins. The scenarios in these simulations showed very low monoterpene concentrations in the early morning that were directly attributable to BVOC reactions with NO 3 . An analysis of NO 3 formation rates at several urban and rural sites in Scandinavia (Ljungström and Hallquist, 1996) resulted in the conclusion that while night-time urban loss of NO 3 is dominated by reaction with NO, the loss in rural regions is likely dominated by reactive hydrocarbons, especially monoterpenes.
Due to the fast reactions of NO 3 with BVOC, lifetimes of NO 3 in biogenically influenced environments can be very short, making simultaneous detection of VOC and NO 3 in biogenic regions very difficult. For this reason, several studies have inferred levels of NO 3 and its role in processing BVOC using observational analysis and supporting modeling. In particular, the rapid decay of isoprene after sunset has received considerable attention.
Measurements of BVOC ~ 1-2 m above canopy level in a loblolly pine plantation in Alabama during the 1990 ROSE program (Goldan et al., 1995) were used to infer a nighttime NO 3 mixing ratio of only 0.2 ppt and NO 3 lifetime of only 7 s due to high levels of monoterpenes. The 4 h decay time of isoprene after sun-set could not be accounted for by gas reactions with NO 3 and O 3 although the decrease in the α-/ β-pinene ratio at night was consistent with known NO 3 and O 3 chemistry. As part of the North American Research Strategy for Tropospheric Ozone -Canada East (NARSTO-CE) campaign, measurements of BVOC were made in Nova Scotia in a heavily forested region (Biesenthal et al., 1998). A box-model simulation based on the observational analysis found that the short lifetime of isoprene at night (τ = 1-3 h) could not be explained by the NO 3 radical, which was estimated to be 0.1 ppt maximum at night due to low NO x and O 3 levels and high monoterpene emissions. When OH yields from ozonolysis of BVOC were included in the model, this nighttime OH oxidant could partially account for the isoprene decay. During the Southern Oxidants Study (SOS) campaign in Nashville, TN (Starn et al., 1998), a chemical box model was used to show that rapid nighttime decays of isoprene were consistent with simulated NO 3 but only when the site was impacted by urban NO x emissions. During the PROPHET study, measurements of VOC were made in a mixed forest approximately 10 m above the canopy surface (Hurst et al., 2001). Isoprene decays at night had an average lifetime of ~ 2.7 h. Box modeling showed that O 3 reactions as well as dry deposition were insufficient to account for the decay, and that the NO 3 radical was a significant sink only after the majority of isoprene had already decayed. On some nights, oxidation by OH could account for all the decay but the decay rates were overpredicted. The authors concluded that vertical transport of isoprene-depleted air aloft contributes to the fast initial decay of isoprene, followed by nighttime OH, NO 3 , and O 3 chemistry decay. Steinbacher et al. (2005) reported on surface measurements in the Po Valley at a site 200-300 m from the closest edge of a deciduous forest. Bimodal diurnal cycles of isoprene were observed with morning and evening maxima that were reproduced by a Eulerian model. Isoprene decay lifetimes of 1-3 h were partially explained by NO 3 decay, although a dynamic influence on isoprene decrease seemed to be likely including horizontal and vertical dispersion. During the HOHenpeissenberg Photochemistry Experiment (HOHPEX) field campaign, BVOC were analyzed via 2-D GC at a site located on a hilltop above adjacent rural agricultural/forested area that is frequently in the residual layer at night (Bartenbach et al., 2007). For the reactive monoterpenes, a significant non-zero dependency of the concentration variability on lifetime was found, indicating that chemistry (as well as transport) was playing a role in determining the ambient VOC concentrations. The night-time analysis gave an estimate of the NO 3 mixing ratio of 6.2 ± 4.2 ppt, indicating it was a significant chemical factor in depletion of monoterpenes.
While the studies above made indirect conclusions about the role of NO 3 in BVOC processing, field studies including direct measurements of NO 3 are key to confirming the above findings. Golz et al. (2001) (Geyer et al., 2001b). Surface measurements at this semi-rural location close to forests found the NO 3 radical above detection limit (2.4 ppt) on 15 of 19 nights with a maximum of 70 ppt, a steady-state lifetime ranging from 20 to 540 s and N 2 O 5 ranging from 2 to 900 ppt. The two most significant losses of NO 3 were found to be its direct reaction with olefins (monoterpenes dominating) and indirect loss due to heterogeneous hydrolysis of N 2 O 5 . Over the study, it was possible for the first time to quantify the relative contribution of the NO 3 radical to oxidation of VOC as 28 (24 h) and 31 % for olefinic VOC (24 h) compared to the total oxidation via NO 3 , OH, and O 3 . As part of the 1999 SOS study, NO 3 , isoprene, and its oxidation products were measured at a suburban forested site in Nashville, TN (Stroud et al., 2002). The nitrate radical measured at multiple beam heights by DOAS had maximum mixing ratios of 100 ppt that were generally found to anticorrelate with isoprene levels with significant vertical gradients on some nights. Early evening losses of isoprene were attributable to reaction with the NO 3 radical.  (Liggio and Mclaren, 2003) during the study found that only monoterpene oxidation products pinonaldehyde and nopinone (not reported) were enhanced in aerosol filters collected at night, evidence of the role of NO 3 in nighttime oxidation of BVOC in the valley. In 2004 measurements of NO 3 and N 2 O 5 by CRDS, isoprene and its oxidation products were made on board the NOAA P-3 aircraft as part of the New England Air Quality Study (NEAQS) and International Consortium for Atmospheric Research on Transport and Transformation (ICARTT) campaigns in the northeast US . These studies found a very clear anticorrelation between isoprene levels after dark and NO 3 mixing ratios, which varied as high as 350 ppt when isoprene was absent from the air mass. The loss frequencies (i.e., first-order loss rate constants) of NO 3 were strongly correlated with the loss rate constant of NO 3 with isoprene for lifetimes less than 20 min, clearly showing that isoprene was the most important factor determining the lifetime of NO 3 . It was also shown that more than 20 % of emitted isoprene was oxidized at night and that 1-17 % of SOA was contributed by NO 3 -isoprene oxidation. A number of recent studies have also investigated the role of NO 3 + BVOC chemistry in more polluted areas. In many urban areas, the NO 3 + BVOC chemistry occurs in parallel to heterogeneous NO 3 / N 2 O 5 chemistry and reactions of NO 3 with anthropogenic VOC. Examples of such environments have been discussed in Brown et al. (2011Brown et al. ( , 2013 and Stutz et al. (2010) who presented observations in Houston, TX. Brown et al. (2011) and Stutz et al. (2010) found that up to 50 % of the NO 3 + VOC reactions in Houston are driven by isoprene, with the other VOC emitted by industrial sources. Surprisingly, heterogeneous NO 3 / N 2 O 5 chemistry plays a minor role in Houston. Brown et al. (2011) also point out that the nocturnal VOC oxidation by NO 3 dominates over that by ozone. Nocturnal NO 3 formation rates were rapid and comparable to those of OH during the day. Crowley et al. (2011) compared NO 3 chemistry in air masses of marine, continental, and urban origin at a field site in southern Spain. Under all conditions, NO 3 + BVOC reactions (predominately α-pinene and limonene) contributed to the overall NO 3 reactivity, confirming other observations that concluded that this chemistry is important in all environments where BVOC sources are present. In the southeastern US summer, this importance extends even through the daytime, when photolysis and NO reactions compete . The NO 3 + BVOC reaction rates observed in these studies imply a high production rate of SOA and organic peroxy radicals.

2.5.2
Organic nitrate aerosols-There are many factors that motivate understanding organic nitrate in the particulate phase through field deployment of a variety of instrumentation, much of which is described in other sections of this review. Nitrogencontaining organic fragments (not necessarily organic nitrates) have been identified in atmospheric particles using mass spectrometric techniques (Reemtsma et al., 2006;Farmer et al., 2010;O'Brien et al., 2014). Total atmospheric organic nitrates, as well as organic nitrates segregated by phase, also have been measured in the atmosphere using techniques such as TD-LIF, CIMS, etc. (Day et al., 2003;Beaver et al., 2012). Given these observations and the propensity of organic nitrate compounds to partition to the condensed phase to create SOA (Rollins et al., 2013), it is critical to determine the level of organic nitrates that reside specifically in the atmospheric aerosol phase under typical ambient conditions and to identify the chemical and physical processes that determine their concentrations. It is also important to note that formation of SOA that contains organic nitrate groups has the potential to sequester NO x , thereby influencing the cycling of atmospheric oxidants.
Organic nitrates in urban PM that were identified using functional group analyses such as FTIR spectroscopy have been attributed to emission of nitrogen-containing primary organic aerosol or to involvement of reactive nitrogen compounds in SOA formation chemistry (Mylonas et al., 1991;Garnes and Allen, 2002;Day et al., 2010). Other more advanced techniques, such as TD-LIF enhanced with the ability to separate phases or techniques to obtain high-resolution mass spectra (HR-ToF-AMS), have been utilized to quantify the amount of organic nitrate in particles in areas less likely to be influenced strongly by BVOC emissions, such as urban areas or areas influenced by oil and gas operations . Of specific interest here, however, are observations of organic nitrate PM in areas with a significant influence of BVOC, especially if co-located measurements allow for insight into the role that NO 3 plays in the initial BVOC oxidation step. As such, we focus here on online measurements and on measurements that allow specific attribution to BVOC-NO 3 reactions. Such measurements broadly can be categorized by region of sampling: the eastern United States (US), the western US, and Europe. Figure 5a summarizes average mass concentrations of submicrometer particulate organic nitrates (NO 3, org ) and particulate inorganic nitrates (NO 3,inorg ) in different months at multiple sites around the world. Figure  5b summarizes the corresponding percentage (by mass) of submicrometer particulate organic nitrate aerosols in ambient organic aerosols. Detailed information and measurements for each site are provided in Table S5.
Eastern United States: The first reports of aerosol organic nitrates in the southeastern (SE) US resulted from composition analysis of four daily PM filter samples from four Southeastern Aerosol Research and Characterization (SEARCH) network sites during summer 2004. Filters were analyzed for polar compounds, with particular focus on organosulfates, using offline chromatographic-MS methods (Gao et al., 2006;Surratt et al., 2007. Several nitrooxy organosulfates were identified, but the only one quantified (1-2 % of organic mass) was associated with α-pinene photooxidation or reaction with NO 3 .
Several of the nitrooxy organosulfates were likely the same as products from BVOCoxidant-NO x -seed systems based on comparison to spectra collected from chamber studies. Brown et al. (2013) (2013), an unprecedented suite of instruments quantified particle-phase organic nitrates using five different online methods: HR-ToF-AMS ( NO + /NO 2 + ), HR-ToF-AMS -PiLS (particle-into-liquid sampler) ion chromatography (2015) used a measurement-constrained model for nighttime that compared NO 3 production/ loss to total organic nitrate (HR-ToF-AMS NO + /NO 2 + , TD-LIF) formation to calculate a molar yield of aerosol-phase organic nitrates of 23-44 % (organic nitrate formed per NO 3 -BVOC reaction) that was dominated by monoterpene oxidation. They noted that the estimated yield was low compared to aggregated aerosol-phase organic nitrate yields, possibly due to rapid nitrate losses not considered in the model. Organic nitrate hydrolysis in the particle phase is one potential loss pathway, although recent laboratory studies suggest this process is slow for NO 3 + β-pinene SOA . Also, particle-phase organic nitrates were observed to contribute 30-45 % to the total NO y budget. Lee et al. (2016) quantified speciated particle-phase organic nitrates using iodide CIMS (88 individual C 4 -C 17 mono/dinitrates). A large fraction was highly functionalized, with six to eight oxygen atoms per molecule. Diurnal cycles of isoprene-derived organic nitrates generally peaked during daytime, and monoterpene-derived organic nitrates peaked at night or during early/mid-morning. Using an observationally constrained diurnal zero-dimensional model, they showed that the observations were consistent with fast gas-particle equilibrium and a short particle-phase lifetime (2-4 h), again possibly due to hydrolysis if the field-derived lifetimes for particle-phase organic nitrates can be reconciled with recent laboratory studies . The sum of the CIMS particle-phase organic nitrates (mass of nitrate functional groups only) was correlated with the two total aerosol organic nitrate AMS-based methods (R 2 = 0.52, 0.67) with slopes of 0.63 and 0.90 . The CIMS sum was also correlated with the total measured with the TD-LIF method (R 2 = 0.55); however, since the TD-LIF measurements were ~ 2-4 times higher (depending on period) than the AMS-based methods, the CIMS versus TD-LIF slope was substantially lower (0.19).
Reasons for the differences between the total organic nitrate measured by different methods have been investigated but remain unclear.
A seasonal and regional survey of particle-phase organic nitrates is reported by Xu et al. (2015b) using a HR-ToF-AMS and an aerosol chemical speciation monitor (ACSM) (Ng et al., 2011) at four rural and urban sites in the greater Atlanta area (2012-2013) and in Centreville, AL (summer 2013 only, SOAS). They show strong diurnal cycles during summer, peaking early/mid-morning, and cycles with similar timing but smaller magnitude during winter. The concentrations were slightly higher in summer, which was attributed to compensating effects of source strength and gas-particle partitioning. Shallower boundary layers during winter also may have played a role in making the summer and winter concentrations more similar (Kim et al., 2015). Fisher et al. (2016) report a broad regional survey of particle-phase (and gas-phase) organic nitrates (HR-ToF-AMS NO + /NO 2 + ) during summertime for the Studies of Emissions and Atmospheric Composition, Clouds and Climate Coupling by Regional Surveys (SEAC4RS) aircraft campaign (August-September, 2013, SE US only) as well as the ground-based SOAS measurements. A substantial vertical gradient was observed in particle-phase organic nitrates, with concentrations decreasing by several-fold from the boundary/residual layer into the free troposphere. Consistent with SOAS ground observations, 10-20 % of observed boundary layer total (gas plus particle) organic nitrates were in the particle phase for the aircraft measurements.
In addition to the measurements made in the SE US, characterization of aerosol organic nitrates has been performed in New England. As part of the New England Air Quality Study (NEAQS) in summer 2002, Zaveri et al. (2010) observed evolution of aerosols in the nocturnal residual layer with an airborne quadrupole (Q)-AMS in the Salem Harbor power plant plume. The aerosols were acidic and internally mixed, suggesting that the observed nitrate was in the form of organic nitrate and that the enhanced particulate organics in the plume were possibly formed from NO 3 -initiated oxidation of isoprene present in the residual layer. functionalities. On average, 21 % of ΣANs were in the particle phase and increased with OA, which was fit to a volatility basis set in which pΣANs / ΣANs increased from ~ 10 % at < 1 μg m −3 and plateaued at ~ 30 % by ~ 5 μg m −3 . At the same site, using PMF analysis of FTIR and HR-ToF-AMS measurements, Liu et al. (2012a) showed that the organic nitratecontaining biogenic SOA condensed onto 400 to 700 nm sized primary particles at night. As part of the Carbonaceous Aerosol and Radiative Effects Study (CARES) in June 2010, Setyan et al. (2012) observed enhanced SOA formation due to interactions between anthropogenic and biogenic emissions at a forest site in the foothills of the Sierra Nevada mountains, approximately 40 km downwind of Sacramento. While nitrate accounted for only ~ 4 % of the particle mass measured by a HR-ToF-AMS, it was attributed potentially to organic nitrates based on the much higher NO + /NO 2 + ion ratio than observed in pure ammonium nitrate. . Similar to studies in other forested environments, the organic nitrate concentration was found to peak at night. The organic nitrate concentration was positively correlated with the product of the nitrogen dioxide and O 3 mixing ratios but not with that of O 3 alone; this suggested nighttime NO 3 -initiated oxidation of monoterpenes as a significant source of nighttime aerosol organic nitrates. The gas-particle partitioning also showed a strong diurnal cycle, with the fraction in the particle phase peaking at ~ 30 % at night and decreasing to a broad minimum of ~ 5 % during daytime, which suggests a change in composition in addition to thermodynamic partitioning effects.
Europe: Iinuma et al. (2007) analyzed ambient aerosol samples collected on filters in a Norway spruce forest in northeastern Germany during the BEWA campaign (Regional biogenic emissions of reactive volatile organic compounds from forests: process studies, modeling, and validation experiments) and compared the results to those from chamber studies. The filter extracts were analyzed using LC-ESI-ToF-MS in parallel to ion trap MS. Several nitrooxy organosulfates with significant mass in the BEWA ambient samples were enhanced in the nighttime samples relative to the daytime samples. Their abundance in the nighttime samples strongly suggests that NO 3 -monoterpene chemistry in the presence of sulfate aerosols has an important role in the formation of these nitrooxy organosulfate aerosols.
A similar study by Gómez-González et al. (2008) focused on isoprene through LCmultidimensional MS (MS n ) analysis of filter samples from both chamber studies and ambient summer day/night PM 2.5 samples from K-Puszta, Hungary, a mixed deciduous/ coniferous forest site. Although not the focus of the study, they confirmed the presence of significant quantities of nitrooxy organosulfates that were enhanced in the nighttime samples over the daytime samples.
Initial online evidence of the production of organic nitrate aerosols in Europe was provided by Allan et al. (2006) when studying nucleation events driven by BVOC oxidation in Hyytiälä, a (boreal) forested region in Finland. The Q-AMS m/z 30 to m/z 46 ratio (the unit mass resolution approximation for NO + /NO 2 + ratio) was frequently found to be very high, ~ 10, for a distinct organic Aitken mode that became apparent late in the afternoon and increased at night. They hypothesized that the excess m/z 30 (NO + ) signal was associated with organic nitrates, although could not rule out amine contributions. During the same field study, Vaattovaara et al. (2009) applied two tandem differential mobility analyzer methods to study the evolution of the nucleation-and Aitken-mode particle compositions at this boreal forest site. The results showed a clear anthropogenic influence on the nucleation-and Aitken-mode-particle compositions during the events and suggested organic nitrate and organosulfate aerosol was generated from monoterpene oxidation. Also, it was shown that organic nitrate was enhanced in aerosol exposed to elevated temperatures, implying low volatility of organic nitrates (Häkkinen et al., 2012).
More recently, Hao et al. (2014) used a HR-ToF-AMS on a tower in Kuopio, Finland, 224 m above a lake surrounded by a mixed forest of mostly coniferous (pine and spruce) mixed with deciduous trees (mostly birch) to measure submicron aerosol composition. The site also was influenced by urban emissions. A particular focus of the study was to separate organic and inorganic nitrate using PMF. They found that ~ 37 % of the nitrate mass at this location and time could be allocated to organic nitrate factors, the rest being inorganic nitrate. The organic nitrate aerosol was segregated into two organic factors, less-oxidized OOA (LO-OOA), and more-oxidized OOA (MO-OOA) (previously called SV-and LV-OOA, respectively); the majority (74 %) of the organic nitrate was found to be in the more volatile LO-OOA factor. Based on meteorology, the air mass source of the organic nitrate aerosol was from a sector with residential and forested areas. Again, the organic nitrate aerosol showed a diurnal trend that was highest at night.
An analysis of AMS data taken across Europe within EU-CAARI and EMEP intensive measurement campaigns (Kulmala et al., 2011;Crippa et al., 2014) has recently shown high organic nitrate contributions to total measured PM 1 nitrate (

Summary of organic nitrate aerosol observations:
Taken together, the observations of particle-phase organic nitrates in the US and Europe suggest that particle-phase organic nitrates (formed substantially via NO 3 -BVOC chemistry) are ubiquitous, especially in, but not limited to, summer. Their formation appears to play an important role in SOA formation, which can potentially be underestimated due to short particle-phase lifetimes. Regions with widespread NO x and BVOC emissions and a humid climate may create optimal conditions for a rapid life cycle of particle-phase organic nitrates.

Models of NO 3 -BVOC chemistry
To understand the implications of NO 3 -BVOC chemistry on atmospheric chemistry as a whole, under both current and future scenarios, the physical and chemical processes, such as those reported in Sect. 2.1 through Sect. 2.3, must be parameterized in numerical models. In this section, we summarize how these reactions are represented in current air quality models (AQMs).

Chemical mechanisms-Organic
nitrates are produced from the reactions of VOC with OH followed by NO as well as with NO 3 , and both of these pathways are represented in chemical mechanisms albeit at varying levels of detail. The use of the term "model" below refers to the treatment of BVOC + NO 3 chemistry in lumped chemical mechanisms. The products formed from the OH-initiated (typically daytime) versus NO 3 -initiated (typically nighttime) chemistry may or may not be treated separately.
The NO 3 -BVOC reactions result in an RO 2 that reacts with NO 3 , other RO 2 , HO 2 , or NO. RO 2 -NO reactions for NO 3 -initiated chemistry are relatively unimportant due to rapid reaction of NO with NO 3 at night (Perring et al., 2009), but they are included in models. Unimolecular rearrangements of the NO 3 -initiated RO 2 radical are not currently considered in models (Crounse et al., 2011). The products of the initial NO 3 -BVOC reaction may retain the nitrate group, thus forming an organic nitrate or releasing nitrogen as NO 2 . The branching between organic nitrate formation and N recycling is parameterized in models. Table 4 summarizes the gas-phase organic nitrate yields for isoprene and monoterpene oxidation by NO 3 in a number of currently available chemical mechanisms. The yields represent the first-generation yields since products may react to form further organic nitrates or release NO 2 . The organic nitrate yield values span from 0 (e.g., SAPRC07 isoprene) to 100 % (e.g., MCM isoprene). Although GEOS-Chem v10-01 does not consider gas-phase monoterpene chemistry, the model has recently been updated to consider a 10-50 % yield of organic nitrates from the monoterpene-NO 3 reaction independent of the nitrate-RO 2 fate but dependent on monoterpene identity . Differences in the organic nitrate yield from NO 3 oxidation result from a number of causes including treatment of RO 2 fate, assumptions about decomposition versus retention, and prioritization of functional group identity.
Some models parameterize the yield of organic nitrates as a function of RO 2 fate while others, such as the carbon bond-based (CB) mechanisms, treat all RO 2 fates the same. The MCM v3.3.1 also considers the yield of isoprene organic nitrates to be independent of RO 2 fate, but monoterpene organic nitrate yields are variable between 0 and 100 % depending on RO 2 fate. Differences in organic nitrate formation, due to treating the organic nitrate yield as a function of RO 2 fate, may vary with atmospheric conditions. Reactions with both HO 2 and RO 2 are significant at night Pye et al., 2015). RO 2 -NO 3 may be important in urban areas or locations where BVOC concentrations are not so high as to deplete NO 3 .
Mechanisms differ in their assumptions about whether or not the organic nitrates from NO 3initiated chemistry release NO 2 or retain the nitrate group. An example of this difference in treatment of organic nitrates can been seen in the reactions of nitrated peroxy radicals with different radicals (NO, HO 2 , RO 2 ) predicted by SAPRC07 and MCM. MCM predicts greater loss of the nitrate group, while SAPRC tends to retain it, leading to either < 5 % (MCM) or > 50 % (SAPRC) organic nitrate yields.
In order to predict accurately the fates of RO 2 and yield of organic nitrates, models must also include information on RO 2 reaction rate constants. Some mechanisms use the same set of RO 2 rate constants for all hydrocarbons. However, the MCM (Jenkin et al., 1997;Saunders et al., 2003) indicates that the RO  AQMs and chemistry-climate models typically cannot handle the complexity associated with tracking each individual VOC and all its possible reaction products. As a result, surrogate species are often used to represent classes of compounds (e.g., CB05, which uses the designation NTR to indicate organic nitrates). This mapping can cause yields of organic nitrates to be falsely low in a mechanism if other functional groups are prioritized over nitrate in the mapping of predicted products to mechanism species. Compared to the other mechanisms in Table 4, SAPRC07 monoterpenes tend to have very low organic nitrate yields as a result of prioritization of peroxide and non-nitrate functional groups. If nitrate groups were prioritized, SAPRC07 would more closely resemble the "other monoterpene" yields from SAPRC07tic. In addition, the diversity across mechanisms in the RO 2 -HO 2 monoterpene organic nitrate yields would be reduced such that they would all indicate > 50 % organic nitrate yields and all but the CB mechanisms would predict a 100 % yield of organic nitrates from RO 2 -HO 2 . The RO 2 -HO 2 pathway is relatively unstudied in laboratory conditions due to difficulties in maintaining sufficient concentrations of both NO 3 and HO 2 radicals Schwantes et al., 2015).

2.6.2
Influence on organic aerosol-Nitrate radical oxidation can lead to significant amounts of SOA on global and regional scales. Due to a lack of information on the identity and volatility of later-generation BVOC + NO 3 products, most models parameterize SOA formation separately from gas-phase chemistry using either the Odum two-product (Odum et al., 1996) fit, volatility basis set (VBS)  fit, or fixed yield (Table 5). Based on the understanding of SOA pathways at the time, Hoyle et al. (2007) found that up to 21 % of the global average SOA burden may be due to NO 3 oxidation, and Pye et al.
(2010) predicted ~ 10 % of global SOA production was due to NO 3 . Regional contributions to SOA concentrations can be much higher (Hoyle et al., 2007;Pye et al., 2010). Nitrate radical reactions themselves are estimated to account for less than 3 % of isoprene oxidation and less than 2 % of sesquiterpene oxidation globally; however, they account for 26 % of bicyclic monoterpene oxidation . Representations of monoterpene-NO 3 SOA are more widespread in chemistry-climate models than other BVOC-NO 3 SOA parameterizations due to the relatively early recognition of its high yields (e.g., Griffin et al., 1999) and relative importance for SOA. Inclusion of SOA from isoprene-NO 3 is more variable as reflected in Table 5.
SOA from BVOC-NO 3 reactions traditionally has been parameterized on the initial hydrocarbon reaction assuming semivolatile products and an Odum two-product approach (e.g., Chung and Seinfeld, 2002). This treatment is often implemented in parallel to the gasphase chemistry, meaning that later-generation products leading to SOA are not identified. Information is still emerging on the fate of organic nitrates, and that information is just beginning to be included in models. Hydrolysis of particle-phase organic nitrates is one such process more recently considered with impacts for both O 3 and PM in models (Hildebrandt Ruiz and Yarwood, 2013;Browne et al., 2014;Pye et al., 2015;Fisher et al., 2016).

Influence on reactive nitrogen and ozone-
The influence of BVOC nighttime oxidation on the nitrogen budget remains unclear. Current modeling efforts have mainly focused on the nighttime oxidation of isoprene, which is dominated by isoprene-NO 3 reaction. This pathway is initialized via addition of NO 3 to one of the double bonds, as discussed in Sect. 2.1.2. Due to the additional stabilization from alkoxy radical and nitrate functional groups (Paulson and Seinfeld, 1992), the yield of first-generation organic nitrates is relatively high (62-78 %; Table 2); they may react with NO 3 again to produce secondary dinitrates (Perring et al., 2009;Rollins et al., 2009Rollins et al., , 2012. Assuming little NO x is recycled from these organic nitrates, most models suggest that nighttime oxidation of isoprene by NO 3 contributes significantly to the budget of organic nitrates (von Kuhlmann et al., 2004;Horowitz et al., 2007;Mao et al., 2013;Xie et al., 2013). Two recent studies (Suarez-Bertoa et al., 2012;Müller et al., 2014), however, suggest fast photolysis of carbonyl nitrates with high efficiency of NO x recycling, which could lead to release of NO x in the next day. Further modeling is required to investigate the importance of nighttime isoprene oxidation on the nitrogen budget.
Very little modeling effort has been dedicated to the influence of nighttime terpene oxidation on the nitrogen budget, mainly due to the lack of laboratory data on oxidation products and their fate. In contrast to isoprene, terpene emissions are temperature sensitive but not light sensitive (Guenther et al., 1995), leading to a significant portion of terpene emissions being released at night. The high yield of organic nitrates and SOA from the terpene-NO 3 reaction  provides an important sink for NO x at night, likely larger than for isoprene-NO 3 over the eastern US . Recent laboratory experiments suggest that aerosol organic nitrates can be either a permanent or temporary NO x sink depending on their monoterpene precursors (and hence nature of the resulting RO 2 ) as well as ambient RH Nah et al., 2016b). In order to understand the impact of terpenes on nighttime chemistry, a fully coupled model of terpene-NO x chemistry will be required, as monoterpenes can be the dominant loss process for NO 3 and N 2 O 5 at night .
While a significant portion of nitrogen is emitted at night , the impact of nighttime chemistry on the initiation of the following daytime chemistry has received little attention in regional and global models. Different treatments of NO 3 chemistry can result in 20 % change in the following daytime O 3 concentration, as shown by a 1-D model study  and box model simulations (Millet et al., 2016). This impact can be further complicated by uncertainty in emissions of BVOC and model resolutions. For example, a recent study by Millet et al. (2016) shows that in a city downwind of an isoprenerich forest, daytime O 3 can be largely modulated by the chemical removal of isoprene throughout the night. Such local-scale events may only be captured by a very high-resolution model with detailed characterization of emission sources. It is important to assess this impact on a global scale using 3-D chemistry models, owing to the profound coupling of boundary layer dynamics and chemistry. Quantifying the impact of BVOC-NO 3 chemistry on NO x fate is important given the long-standing problem in current global and regional AQMs of a large overestimate of O 3 over the eastern US in summer (Fiore et al., 2009).

Comparison of field data with air quality models-Recent field campaigns
(SOAS, SEAC4RS, EUCAARI, EMEP) have allowed for the attribution of SOA to NO 3 oxidation to provide model constraints not previously available. Pye et al. (2015) and Fisher et al. (2016) implemented updated BVOC + NO 3 chemistry in CMAQ and GEOS-Chem, respectively, to interpret data in the SE US during the summer of 2013 (SOAS and SEAC4RS). Model predictions of gas-phase monoterpene nitrates (primarily NO 3 derived) were higher than the sum of C 10 H 17 NO 4 and C 10 H 17 NO 5  by a factor of 2-3  and 7 , consistent with a significant fraction of the monoterpene nitrates being highly functionalized . The studies identified particle-phase hydrolysis as an important modulator of particulate organic-nitrate concentrations and organic nitrate lifetime. The GEOS-Chem simulation reproduced the particle-phase organic nitrate diurnal cycles (SOAS), boundary layer concentrations, and gas-particle partitioning reasonably well; however, it underestimated concentrations in the free troposphere, possibly due to measurement limitations and/or the implementation of rapid uptake followed by hydrolysis of all gas-phase organic nitrates in the model, which may not be valid for non-tertiary organic nitrates .

Perspectives and outlook
Section 3 outlines perspectives on the implications of NO 3 -BVOC atmospheric chemistry with respect to (1) aerosol optical and physical properties; (2) health effects; (3) trends in NO x emissions and organic aerosols and their implications for control strategies related to particulate matter; (4) critical needs for analytical methods; (5) critical needs for models; (6) field studies in the developing world and under-studied areas; and (7) critical issues to address in future field and laboratory measurements in light of current understanding of this chemistry and trends in emissions.

Aerosol optical and physical properties
The climatic effects of atmospheric aerosols depend on their various physical and chemical properties. Hygroscopicity, cloud condensation nuclei (CCN) activity, optical properties (namely light absorption and scattering), and ability to act as CCN and ice nuclei (IN) are the key aerosol properties that would determine their ability to affect climate. Additional properties such as aerosol number size distribution, chemical composition, mixing state, and morphology will determine whether the aerosols will be optically important or whether they would affect cloud properties. These aerosol properties depend on the sources, aging processes, and removal pathways that aerosols experience in the atmosphere (Boucher, 2013).
Absorption by aerosol may affect the cloud lifetime and altitude due to heating of the atmosphere (Mishra et al., 2014). They can also change the atmospheric lapse rate, which in turn can result in modification in aerosol microphysics in mixed-phase, ice, and convective clouds (Boucher, 2013). In addition to direct emissions of known absorbing particles (black carbon, mineral dust, biomass burning aerosols), SOA may also have absorption properties. The absorbing component of organic carbon (OC), namely "brown carbon" (BrC), is associated with OC found in both primary and secondary OC and has a spectral-dependent absorption that smoothly increases from short visible to UV wavelengths (Bond and Bergstrom, 2006). It has been suggested that BrC is a component of SOA that is composed of high molecular weight and multifunctional species such as humic-like substances, organic nitrates, and organosulfate species (Andreae and Gelencser, 2006;Bond and Bergstrom, 2006;Ramanathan et al., 2007b;Laskin et al., 2015;Moise et al., 2015). Many modeling studies often assume that BC and mineral dust are the only two significant types of lightabsorbing aerosols on the global scale. Therefore, they treat SOA as a purely scattering component that leads to climate cooling (Stier et al., 2007;Bond et al., 2011;Ma et al., 2012). However, observations suggest that BrC is widespread mostly around and downwind urban centers (Jacobson, 1999). In such places, BrC may have significant contribution, and in some cases it may dominate the total aerosol absorption at specific (short) wavelengths (Ramanathan et al., 2007a;Bahadur et al., 2012;Chung et al., 2012;Feng et al., 2013).
Based on observations, Chung et al. (2012) recently suggested that the direct radiative forcing of carbonaceous aerosols is +0.65 (0.5 to about 0.8) Wm −2 , comparable to that of methane, the second most important greenhouse gas. This study emphasizes the important role of BrC and calls for better measurements of the absorption properties of BrC, specifically at short wavelengths where the absorption is most significant. Many previous studies have concentrated on primary particulate matter, mostly from biomass burning. However, these studies often neglected contributions to absorption due to BrC in SOA. There is ample laboratory and field evidence for the formation of such absorbing material in SOA (Chung et al., 2012;Lack et al., 2012). This absorbing component is the least characterized component of the atmospheric absorbing aerosols and constitutes a major knowledge gap, calling for an urgent need to identify the optical properties of the organic (BrC) component in SOA, and the chemical pathways leading to its formation and losses Lin et al., 2015;Moise et al., 2015).
Recently, Washenfelder et al. (2015) measured aerosol optical extinction and absorption in rural Alabama during the SOAS campaign. While they found that the majority of BrC aerosol mass was associated with biomass burning, a smaller (but not negligible) contribution was attributed to biogenically derived SOA. This fraction reached a daily maximum at night and correlated with particle-phase organic nitrates and is associated with nighttime reactions between monoterpenes and the NO 3 radical (Xu et al., 2015a). Based on the above, it is concluded that SOA produced from reactions of NO 3 with BVOC can be a major source of SOA during the night that may affect daytime aerosol loading. This important fraction of NO 3 -derived SOA can contribute to the direct radiative effect of SOA through scattering and absorption of incoming solar radiation.
Nitration of aromatic compounds (oxidation via NO 2 , NO 3 , N 2 O 5 ) has a potential to form chromophores that can absorb solar radiation. Theoretical and experimental studies have shown that nitration of PAHs leads to nitro PAHs and their derivatives such as nitrophenols (Jacobson, 1999;Harrison et al., 2005;Lu et al., 2011). The nitro substituents on the aromatic ring in compounds enhance and shift the absorption to longer wavelengths (> 350 nm). Field studies report that nitrogen-containing mono-and polyaromatic SOA constituents absorb light at short (near-UV and visible) wave-lengths. The reaction products between NO 3 and BVOC have the potential to form effective chromophores. Multifunctional organic nitrates and organosulfate compounds formed during the nighttime suggest that the SOA produced from NO 3 reactions leads to formation of BrC that can absorb solar radiation (Iinuma et al., 2007).
Only a few studies have investigated optical properties of SOA partially composed of organic nitrates (Moise et al., 2015). Most existing literature on optical properties of organic nitrates in SOA has been focused on oxidation of anthropogenic precursor compounds (Jacobson, 1999;Nakayama et al., 2010;Lu et al., 2011;Liu et al., 2012b), while a few partially contradictory studies have examined SOA formed from NO 3 reaction with biogenic precursors (Song et al., 2013;Varma et al., 2013). The typically high mass absorption coefficient (MAC) that was observed for anthropogenic high-NO x SOA can be partially attributed to the presence of nitroaromatic groups, for example, via the nitration of PAHs (Jacobson, 1999;Lu et al., 2011). Song et al. (2013) examined optical properties of SOA formed by NO 3 + O 3 + α-pinene. With neutral seed aerosol, organic nitrates were present but observed to be non-absorbing; however, with acidic seed aerosol, SOA were strongly light absorbing, which the authors attributed to nitrooxy organosulfates formed via aldol condensation. Varma et al. (2013) measured absorption of NO 3 +β-pinene SOA and found a higher refractive index than when oxidation was via OH or O 3 , and attributed to the difference to the low HC / NO x ratio and presence of organic nitrates in the particle phase.
Laboratory and field studies suggest that SOA formed by nighttime chemistry can have profound regional and possible global climatic effects via their absorbing properties. However, the optical properties of NO 3 -containing SOA are not well known. Varma et al. (2013) measured a high value for the refractive index real part value of 1.61 (±0.03) at λ = 655-687 nm following reactions of NO 3 with β-pinene. This value is significantly higher than values observed following OH-and ozone-initiated terpene oxidation (Fig. 6) (Moise et al., 2015). This has been attributed to the high content (up to 45 %) of organic nitrates in the particle phase .
Key physical parameters of aerosols include particle size and number, volatility, viscosity, hygroscopicity, and CCN activity. While it is clear that atmospheric particle size increases through condensation of BVOC + NO 3 oxidation products, the effect of NO 3 oxidation on particle number is not usually studied in laboratory experiments. Very little is known about the volatility of SOA from NO 3 , with field studies from Hyytiälä indicating that organic nitrates may have low volatility (Häkkinen et al., 2012). Viscosity is not known. Few studies report the hygroscopicity and CCN activity of SOA from NO 3 oxidation of BVOC. A study by Suda et al. (2014) showed that organic compounds with nitrate functionality (compared to other functional groups such as hydroxyl, carbonyl, hydroperoxide) have the lowest hygroscopicity and CCN efficiency. Recently, Cerully et al. (2015) reported that the hygroscopicity of less-oxidized OOA (LO-OOA, mostly from BVOC + NO 3 ) is lower than other OA subtypes (MO-OOA and isoprene-OA) resolved by PMF analysis of AMS data from the SOAS campaign. As monoterpenes + NO 3 reactions can contribute ~ 50 % of nighttime OA production (Xu et al., 2015a), results from Cerully et al. (2015) suggested that it is possible that SOA formed from NO 3 oxidation of BVOC is less hygroscopic than OA formed from other oxidation pathways.

Health effects
Nitrated organic compounds also pose adverse health effects (Franze et al., 2003(Franze et al., , 2005Pöschl, 2005;Gruijthuijsen et al., 2006;Pöschl and Shiraiwa, 2015). In particular, several studies have reported that biological particles such as birch pollen protein can be nitrated by O 3 and NO 2 in polluted urban air (Franze et al., 2005;Reinmuth-Selzle et al., 2014). The mechanism of protein nitration involves the formation of long-lived reactive oxygen intermediates, which are most likely tyrosyl radicals (phenoxy radical derivatives of tyrosine) . The resulting organic nitrates were found to enhance the immune response and the allergenicity of proteins and biomedical data suggest strong links between protein nitration and various diseases (Gruijthuijsen et al., 2006). Inhalation and deposition of organic nitrates into lung lining fluid in the human respiratory tract may lead to hydrolysis of organic nitrates forming HNO 3 , which may reduce pulmonary functions (Koenig et al., 1989). Consequently, inhalation of aerosols partially composed of nitrated proteins or nitrating reagents might promote (i) immune reactions, (ii) the genesis of allergies, (iii) the intensity of allergic diseases, and (iv) airway inflammation. Toxicity of nitrated SOA compounds is still unclear. In the light of these observations and remaining uncertainties, the effect of organic nitrates present in biogenic SOA on human health should be a focus of future studies.
Formaldehyde is an important source of atmospheric radicals as well as a major hazardous air pollutant (HAP). It is a degradation product of almost every VOC in the atmosphere, and BVOC are known to contribute substantially to ambient concentrations of formaldehyde . The overall yield of formaldehyde from BVOC-NO 3 reactions is lower than from corresponding OH reactions, indicating that any changes in the relative distribution of oxidation routes will have a corresponding change in formaldehyde (and thus oxidant regeneration and HAP exposure).

Trends in NO x emissions and organic aerosols -implications for air quality control strategies
Nitrogen oxide emissions are converted to NO 3 and thus affect nitrate-derived SOA. In the United States, where NO x emissions are dominated by fuel combustion, regulatory actions have resulted in decreasing NO x levels after increases from 1940 to 1970(Nizich et al., 2000 and relatively stable levels between ~ 1970 and ~ 2000 (Richter et al., 2005). NO x emissions in the US are estimated to have decreased by roughly 30-40 % in the recent past (between 2005 and 2011/2012), as reflected in satellite-observed NO 2 , ground-based measurements, and the Environmental Protection Agency (EPA) National Emission Inventory (NEI) Xing et al., 2013Xing et al., , 2015Hidy et al., 2014;Tong et al., 2015). Recent decreases in NO x have been attributed to the mobile sector, and power plant controls including the EPA NO x State Implementation Plan Call implemented between 2003 and 2004 (Kim et al., 2006;Russell et al., 2012;Hidy et al., 2014;Foley et al., 2015;. In the United States, NO x emissions are expected to continue to decrease and reach 72 and 61 % of their 2011 levels in 2018 and 2025, respectively (Eyth et al., 2014). Furthermore, recent work indicates that NO x emissions may be overestimated in models for the United States  particularly for on-road gasoline vehicles (McDonald et al., 2012).
Globally, the Representative Concentration Pathway trajectories indicate that NO x emissions will decrease below year 2000 levels by the middle of the 21st century (Lamarque et al., 2011). Europe has experienced declines in NO x with NO 2 concentrations decreasing by 20 % over western Europe between 1996 and 2002 (Richter et al., 2005) and decreasing by an additional ~ 20 % in the more recent past (2004-2010) (Castellanos and Boersma, 2012). In contrast, NO x emissions in China have increased by large amounts since 1996 (Richter et al., 2005;Stavrakou et al., 2008;Verstraeten et al., 2015) with a more recent leveling out or decrease of NO 2 concentrations (Krotkov et al., 2016). NO 2 concentrations in India have continued to increase (Krotkov et al., 2016;Duncan et al., 2016).
These large past and expected future changes in anthropogenic NO x emissions indicate that analysis of historical data could reveal how NO x emissions affect organic aerosol formation and more specifically SOA from NO 3 -initiated chemistry. Long-term monitoring networks often measure NO x and OC, which could allow for correlation analysis. In addition, air quality trends in organic aerosol from traditionally less-sampled locations (e.g., Streets et al., 2008) and emissions for locations such as China have been characterized and could be used for analysis.
In addition to examining measurement data for relationships between NO 3 -derived SOA and NO x , chemical transport modeling with emission sensitivity simulations can be used to provide estimates of how various SOA pathways respond to changes in NO x emissions. For example, Carlton et al. (2010b) used the CMAQ model to determine that controllable NO x emissions were responsible for just over 20 % of total SOA in the United States based on the NO 3 -BVOC mechanism available at the time. Pye et al. (2015) predicted nitrate-derived SOA concentrations would decrease by 25 % due to a 25 % reduction in NO x emissions, but the overall change including all organic aerosol components would be only 9 % as a result of other less sensitive (or increasing) components. Other modeling studies (Lane et al., 2008;Zheng et al., 2015;Fisher et al., 2016) have shown that total organic aerosol or particlephase organic nitrates may not respond strongly to decreased NO x emissions, but significant spatial and composition changes can occur.

Organic nitrate standards
The CIMS technique allows for highly time-resolved, chemically speciated measurements of multifunctional organic nitrates (Beaver et al., 2012;Paulot et al., 2012;Lee et al., 2014a;Xiong et al., 2015). Synthesis, purification, and independent quantification of an individual, isomerically specific organic nitrate is, however, required for calibration because standards are not commercially available, except for a few monofunctional alkyl nitrates.
Most recently, three isomers of isoprene hydroxynitrates were synthesized (Lockwood et al., 2010;Lee et al., 2014b). As the precursor ingredient is an organic epoxide on which hydroxy and nitrate functional groups are attached, the same protocol (Nichols et al., 1953;Cavdar and Saracoglu, 2008) can be applied to synthesize hydroxynitrates of various VOC backbones assuming availability of precursor compounds. Oxidation of a single-parent compound can yield numerous isomerically unique byproducts possessing various functional groups, including one or more nitrates. As such, synthesis of and calibration for each nitrate rapidly become prohibitive. Given that multifunctional organic nitrates possessing more than four oxygen atoms, for which synthesis protocols currently do not exist, dominate the particulate nitrate mass of submicron particles , a more comprehensive calibration technique is needed. Three broad approaches are currently utilized. One is to cryogenically collect a suite of oxidation byproducts (present in the atmosphere, formed in a simulation chamber or flow tube, etc.) on a GC column. The desorbing eluent, separated in time by volatility/polarity as it is thermally desorbed, is measured simultaneously by CIMS and a quantitative instrument such as the TD-LIF (Day et al., 2002;Lee et al., 2014b). The corresponding eluting peaks detected by both instruments allow for calibration of each surviving, isobarically unique (at least for unit mass resolution spectrometers) organic nitrate Schwantes et al., 2015;Teng et al., 2015). The second approach employed for the iodide adduct ionization technique is to deduce the instrument response from a comparison of the binding energies of the numerous iodide organic nitrate clusters to those of compounds with known sensitivities by applying variable voltages in the ion molecule reaction region to break up charged clusters systematically. The rate at which the signal of an organic nitrate cluster decays with voltage is a function of its binding energy, which governs its transmission efficiency through the electric fields and thus its sensitivity . Lastly, quantum chemical calculations of specific compounds allow the determination of the sensitivity of their iodide adduct  and CF 3 O − (Kwan et al., 2012;Paulot et al., 2012) ionizations.

Robust and efficient representation of gas-phase chemistry-Previous
sections have detailed the reactions of BVOC with NO 3 and the need to include this chemistry to represent more accurately processes that control O 3 and SOA formation. But applying that information in a way that can be used for air quality studies presents a serious challenge. As highlighted in Sect. 2.6.1, the chemical mechanisms currently being used in AQMs are limited in their representation of NO 3 -BVOC chemistry, largely lumping all monoterpenes together, and with no agreement on yields. The lack of detail in current mechanisms is reflected in the variety of methods by which SOA formation from BVOC-NO 3 chemistry is estimated (Sect. 2.6.2).
Typically, the NO 3 -BVOC chemistry is implemented in AQMs into the existing system of organic and inorganic chemical reactions that occur in the atmosphere. Because there may be hundreds or thousands of different chemical reactions occurring simultaneously and the computational efforts required to solve those on a 3-D grid are onerous, the chemical mechanisms used in AQMs are typically condensed to a certain extent. The greatest challenges in modeling the reactions initiated by NO 3 and BVOC in AQMs are (1) deciding how much detail must be included to accurately represent the chemistry; (2) (4) validating the complete schemes against observational data.
Including all of the attack pathways and isomers that are formed in the reactions of NO 3 and BVOC and their subsequent products rapidly becomes an intractable problem, as the number of species and reactions produced from a VOC grows exponentially with the number of carbons in the compound (Aumont et al., 2005), resulting in an estimate of almost 400 million products from a single C 10 hydrocarbon. Even restricting the chemistry solely to the RO 2 formed from α-pinene, β-pinene, and limonene via addition of NO 3 to the double bond results in 861 unique product species and 2646 reactions as estimated from the MCM (http:// mcm.leeds.ac.uk/MCM-devel/home.htt; Saunders et al., 2003). In comparison, the chemical mechanisms used in AQMs typically consider a total of 100-200 species and less than 400 reactions to model the entire gas-phase chemistry occurring in the troposphere. One challenge is to find a balance between complexity and computational efficiency that involves both deriving complete mechanisms as well as condensing them to the extent possible.
The second major challenge is that many of the chemical pathways must be estimated given the limited experimental measurements of intermediate reaction rate constants and products. Structure-activity predictions have been used heavily in the past, but these have been formulated for a limited number of compounds. Their predictions become less accurate as the complexity of the molecule increases (Calvert et al., 2015). When heterogeneous reactions play a significant role in the transport and fate of reaction products, as they do in monoterpene chemistry, the challenge becomes even greater. With recent research, new product structures that contribute to SOA have been identified . However, these are not covered by existing predictive theory, and these new pathways must be characterized, including reaction rate constants, co-reactants, and products. Physical parameters of all of these new species, such as solubility, radiative properties, emission rates, and deposition velocities also are required, but data are often unavailable for these or even comparable species.
The last challenge is integrating the chemistry within the rest of the chemistry occurring in the atmosphere. The major chemical mechanisms used in AQMs today were developed primarily to address episodes of elevated O 3 under conditions of high NO x and have been evaluated for this purpose. Thus, the mechanisms often do not lend themselves well to predicting the chemistry of complex VOC or other air quality endpoints (Kaduwela et al., 2015). Minor pathways with respect to O 3 formation have been removed from the mechanisms to reduce the computational burden, but these pathways may be important for formation of SOA. In addition, the detailed chemistry of multistep alkoxy and peroxy radical chemistry is condensed into a single step in some mechanisms, but identifying whether these radicals react with NO x or HO x or isomerize is critical for predicting the types of organic molecules that are formed. As described in Sect. 2.6, existing mechanisms include the capability for a limited number of nitrates, and in many cases the links to facilitate expansion to more detailed representations are missing.
Significant work must be done to allow modelers to implement this new information in AQMs and thus use this updated knowledge to develop improved predictions of future air quality. One approach is to focus on key chemicals of interest, derive mechanisms that are suitable for specialized applications, and append these on to existing frameworks (for example, Xie et al., 2013). The longer-term view requires a more comprehensive approach that draws on the development of community archives that can better accommodate rapidly changing information and better represent the interactions of biogenic with anthropogenic chemistry. Here, we put forward our recommendations for future work in the following areas:

1.
Development of tools for the semi-automated production of the reaction pathways and products of later-generation products resulting from alternate pathways of radical reactions with BVOC. These tools should be able to incorporate experimental data when available. In conjunction with the automated development, we require advanced methods for condensing these large mechanisms into computationally feasible reaction schemes.

2.
Improvements in estimation techniques for uncertain pathways, including reaction rate constants for multifunctional stable compounds and radicals for which measurements are not available, and the quantification of the errors associated with these estimation methods.

3.
Development of theory and techniques for integrating gas-phase products with SOA production, in this case, describing the transformation of gas-phase organic nitrates to their SOA products.

4.
Development of more versatile base mechanisms that have the flexibility to accept increased detail in VOC description and the continuing validation of the complete tropospheric chemical mechanisms against observational data.
3.5.2 Improved techniques and protocols for evaluation of complex and reduced gas-phase mechanisms-Generally speaking, once detailed mechanisms are developed, they are evaluated through some form of benchmarking. Systematic strategies for mechanism evaluation include validation of highly detailed mechanisms unable to be run in 3-D models against benchmark data from well-characterized simulation chamber experiments (Jenkin et al., 1997;Aumont et al., 2005) and the incorporation of these mechanisms into box or 1-D models to validate radical and short-lived species against field campaign observations. Less-detailed air quality (AQ) mechanisms can then be compared to these reference mechanisms by way of sensitivity experiments in idealized modeling studies -often aimed at assessing the sensitivity in O 3 to changing NO x and VOC emissions (Archibald et al., 2010;Squire et al., 2015). AQ mechanisms are often also then re-evaluated against chamber and/or field experiment data before they are implemented into 3-D models and then undergo evaluation against extensive measurements in the residual layer.
One of the greatest challenges in the BVOC-NO 3 system is that current nighttime measurements are mainly collected from surface sites, which are confined to a shallow surface layer at night and not representative of the whole nighttime boundary layer. The impact of nighttime chemistry on daytime ozone and nitrogen/aerosol budget would require careful investigation of nighttime chemistry in the residual layer, which contains > 80 % of air masses at night.
Moreover, the benchmarking activities mentioned above and the development process discussed in Sect. 3.5.1 are not well aligned. A more unified approach that identifies some key mechanistic problems and identifies strategies to evaluate them is required in order to make improved progress on simulating the changing composition of the atmosphere.

Reduce uncertainties in sub-grid-scale processes-Uncertainties in AQM
predictions also arise from the representation of physical sub-grid-scale processes. The ones particularly relevant for the NO 3 -BVOC chemistry include, but are not limited to, the following.

Nighttime boundary layer mixing:
The spatial distribution of BVOC and NO x precursors is highly variable, but the current AQMs neglect these heterogeneities and assume perfect mixing within grid cells of typically 3-10 km in the horizontal. At those resolutions, models are unable to resolve the localized surface emission sources and the microscale structure of boundary layer turbulence, and therefore cannot resolve spatial heterogeneities in chemistry, partitioning, and mixing of chemicals, which are essential for predicting the concentrations of secondary pollutants.
Typically, the freshly emitted monoterpene species have a tendency to accumulate in the shallow nighttime boundary layer (typically < 200 m), and can react with NO 3 if available. However, often NO 3 is located in the residual layer that is decoupled from the nighttime boundary layer (NBL), and the BVOC + NO 3 reactions would depend on the model's ability to mix the two layers. Thus, mixing within and out of the boundary layer provides a key challenge for modeling the impacts of BVOC-NO 3 chemistry, as the measured gradients of NO 3 and BVOC are very strong in the vertical (e.g., Brown et al., 2007b;Fuentes et al., 2007).
A large focus on model evaluation has been on the impacts of higher horizontal resolution (Jang et al., 1995). It has been shown in several cases that owing to the complex interplay of chemical families, the sensitivity of the chemical system is not captured at lower resolution (e.g., Cohan et al., 2006). However, very little work has focused on the role of improvements in vertical resolution, despite the fact that inter-model differences in properties like the height of the boundary layer vary by over a factor of 2 in some cases (e.g., Hu et al., 2010). Moreover, the NBL is not well mixed, so evaluation of nocturnal physics requires more than just evaluating the NBL height.
Plume parameterizations: Typically, parameterizations have been applied to anthropogenic emission sources (e.g., aircraft plumes, urban plumes) and not to biogenic sources. Partly, this is a result of the differences in the source terms, anthropogenic emissions often being well represented as point sources in space, whereas biogenic emissions are often large area sources. However, as the emissions of BVOC are often very species specific, and observations highlight large spatial variability over small areas (e.g., Niinemets et al., 2010), the adoption of the anthropogenic plume parameterizations to BVOC emissions could lead to improvements in model performance.
One approach is the plume-in-grid (PiG) parameterization (Karamchandani et al., 2002). This aims to solve the problem of sub-grid-scale chemical processes by implementing ensembles of Gaussian puffs within the AQM (e.g., Vijayaraghavan et al., 2006). Other approaches include hybrid Eulerian-Lagrangian models (Alessandrini and Ferrero, 2009). These differ from the PiG models by simulating large numbers of stochastic trajectories that can make use of variable reactive volumes to simulate their diffusion into background air masses simulated on Eulerian grids.
Global models have generally used a different approach to the problem of plumes. Broadly, following one of the two paradigms (Paoli et al., 2011) to (i) modify the emissions of the reaction mix (using so-called effective emissions or applying emission conversion factors) and (ii) modify the rates of reaction (effective reaction rates).

Field studies in the developing world and under-studied areas
In light of the questions raised earlier in this review, assessing the role of NO 3 -BVOC chemistry will require field experiments over a wide range of ratios of isoprene to monoterpene emissions and of NO 3 to BVOC. Future studies of NO 3 -BVOC chemistry are in the planning stages for North America. These studies will provide access to environments with different NO x levels and over a modest range of isoprene and monoterpene emission rates. A wider range of these parameters can be accessed in countries where NO x emission controls are not as completely implemented and where BVOC emissions are abundant.
Bringing the state-of-the-art capabilities developed for study of NO 3 -BVOC chemistry to locations in China and India would allow insight not only into the role of that chemistry in those countries now but also into the role this chemistry played in Europe and the US prior to implementation of current emission standards. Experiments in the tropics potentially would allow observations of the confluence of BVOC and very low NO x to be explored, thus providing insight into BVOC-NO 3 as a sink of NO x .

Future needs for chamber studies
Field studies, by definition, include the entire complexity of the real atmosphere, so that the identification of single processes and quantification of their impact is challenging. Specific experiments in chambers allow investigating processes without effects from meteorology, which largely impacts observations in the real atmosphere specifically during nighttime, when the lower troposphere is not as well mixed as it is during daytime. In chamber experiments, specific compounds of interest can be isolated and studied under wellcontrolled oxidation environments, allowing a more detailed and direct characterization of the composition, chemical, and physical properties of aerosols. Because such laboratory chamber data provide the basic understanding for predicting SOA formation, it is important that the design of such experiments mimics the oxidation environments in the atmosphere to the greatest extent possible. Several important needs for understanding NO 3 -BVOC chemistry in chambers include (1) elucidation of kinetic and mechanistic information for NO 3 -BVOC reactions; (2) characterization of wall losses for low-volatility products in the NO 3 -BVOC system; (3) understanding the fate of peroxy radicals in the nighttime atmosphere and its influence on this chemistry; (4) hydrolysis and photooxidation of BVOCderived organic nitrates from specific BVOC plus specific oxidant pairs over a range of appropriate conditions; (5) optical properties of aerosol organic nitrate; and (6) intercomparison of instrumental methods for key species in the NO 3 -BVOC system.

Kinetic and mechanistic elucidation:
The number of chamber studies investigating NO 3 chemistry is small compared to the number of studies for photochemical oxidation and ozonolysis. In most of the studies, gas-phase oxidation products and SOA yields from the oxidation of BVOC have been measured. Studies include the investigation of SOA from monoterpenes Griffin et al., 1999;Hallquist et al., 1999;Fry et al., 2009Nah et al., 2016b), methyl butenol (Fantechi et al., 1998a, b), and isoprene Ng et al., 2010;. A few more studies investigated gas-phase reaction kinetics, including the reactions of NO 3 with aldehydes (Clifford et al., 2005;Bossmeyer et al., 2006), amines (Zhou and Wenger, 2013), or cresol (Olariu et al., 2013). As a consequence of the small number of studies, the oxidation mechanisms of organic compounds by NO 3 and the yields of oxidation products in the gas phase and particle phase have larger uncertainties. The wellcontrolled oxidation environments in chamber experiments, coupled with complimentary gas-phase and particle-phase measurements (online and offline), allow for elucidating detailed oxidation mechanisms under varying reaction conditions Schwantes et al., 2015). Identification of gas-and particle-phase reaction products from NO 3 -BVOC chemistry within controlled chamber environments can also greatly aid in the interpretation of field data in which multiple oxidants and BVOC are present. Future chamber experiments will naturally take advantage of new advanced gas-aerosol instrumentation and aim to constrain the formation yields of gas-phase oxidation products and establish a fundamental under-standing of aerosol formation mechanisms from NO 3 -BVOC under a wide range of oxidation conditions.

Wall losses:
Although chamber studies allow separating processes driven by chemistry and physics from transport processes that occur in the real atmosphere, careful characterization of the behavior of NO 3 in chambers as well as the organic products of the NO 3 oxidation remains a research priority. Yields of gas-phase oxidation products can be influenced by chamber-specific loss processes (surface loss on the chamber wall) and SOA yields can be impacted by both direct loss of particles and loss of species that can condense on particle or chamber wall surfaces (McMurry and Grosjean, 1985;Loza et al., 2010;Matsunaga and Ziemann, 2010;Yeh and Ziemann, 2014;Zhang et al., 2014aKrechmer et al., 2016;La et al., 2016;Nah et al., 2016a;Ye et al., 2016). The extent to which vapor wall loss affects SOA yields appears to be dependent on the VOC system, from relatively small effects to as high as a factor of 4 (Zhang et al., 2014a;Nah et al., 2016a). Studies on the effects of vapor loss on SOA formation from BVOC + NO 3 are limited. With minimal or no competing gas-particle partitioning processes, substantial vapor wall loss of organic nitrates has been observed in experiments not specific to NO 3 oxidation Krechmer et al., 2016). However, the use of excess oxidant concentrations and rapid SOA formation in BVOC + NO 3 experiments (hence, shorter experiments) could potentially mitigate the effects of vapor wall loss on SOA yields in chamber studies Nah et al., 2016a). In light of the developing understanding of this issue, an important consideration for the design of any future systematic chamber studies is the influence of vapor wall loss on SOA formation from nitrate radical oxidation under different reaction conditions, such as peroxy radical fates, relative humidity, seeds, oxidant level, chamber volume, etc.
Peroxy radical fate: As discussed above, the fate of peroxy radicals directly governs the product distribution in the NO 3 -BVOC system, including SOA yields and composition. Dark reactions of peroxy radicals differ significantly from their photochemical analogs, and are directly related to the development of mechanistic understanding in the NO 3 -BVOC system. There is a need to systematically investigate reaction products and SOA formation from NO 3 -BVOC reactions under different peroxy radical reaction regimes, but this aspect has only recently become a focus of chamber studies Schwantes et al., 2015). Rapid formation of highly oxygenated organic nitrates has been observed in laboratory studies of β-pinene + NO 3 and α-pinene + NO 3 ; these products could be formed by unimolecular isomerization of peroxy radicals or autoxidation (Nah et al., 2016b). The importance of this peroxy radical reaction channel in NO 3 -BVOC chemistry warrants further studies. Future chamber studies will need to be explicit in their specification of the peroxy radical chemistry regime that is investigated in a particular experiment, and will need to relate that regime to the conditions of ambient nighttime atmosphere.
Organic nitrate hydrolysis and photooxidation: Recent field studies have shown that organic nitrates formed from NO 3 -BVOC are important components of ambient OA. However, the reactivity in both gaseous and condensed phases of these biogenic nitrates, in particular of polyfunctional nitrates, has been subject to few studies and requires better characterization to evaluate the role of these compounds as reservoirs/sinks of NO x . Field results suggest that the fate of organic nitrates in both the gas and aerosol phase have variable lifetimes with respect to hydrolysis. The difference in the relative amount of primary/secondary/tertiary organic nitrates (which hydrolyze with different rates) from nitrate radical oxidation versus photochemical oxidation needs to be constrained. Most of the hydrolysis studies thus far are conducted in bulk, except for a few recent studies on monoterpene organic nitrates (e.g., Rindelaub et al., 2015). The solubility of multifunctional organic nitrates in water and the extent to which hydrolysis occurs in aerosol water warrant future studies. The effect of particle acidity on hydrolysis might also be important for organic nitrates formed in different BVOC systems.
While there are extensive studies on photochemical aging of ozonolysis SOA, studies on photochemical aging of NO 3 -initiated SOA and organic nitrates are extremely limited. A recent study shows that the particle-phase organic nitrates from NO 3 +β-pinene and NO 3 +αpinene reactions exhibit completely different behavior upon photochemical aging during the night-to-day transition, and act as permanent and temporary NO x sinks, respectively (Nah et al., 2016b). With the ~ 1-week lifetime of aerosols in the atmosphere and the majority of NO 3 -BVOC organic nitrates that are formed at night, the photochemical fates of these organic nitrates could impact next-day NO x cycling and ozone formation. Therefore, there is a critical need to understand the multigenerational chemistry and characterize the evolution of organic nitrates over its diurnal life cycle, including aging NO 3 -initiated SOA and organic nitrates by photolysis and/or OH radicals.

Aerosol optical properties:
The optical properties, especially in the short wavelength region, of NO 3 -derived SOA may be most conveniently measured during coordinated chamber studies that also include detailed measurements of gas-phase oxidation chemistry and aerosol composition. Such studies could also serve to isolate the specific optical properties of NO 3 -BVOC-derived aerosol to obtain better optical closure in the interpretation of field data. Field studies that include aerosol optical properties measurements in conjunction with other instrumentation can help quantify the bulk organic nitrate abundance and identify organic nitrate molecular composition in the SOA.

Instrument intercomparisons:
The discussion above shows that recent advances in analytical instrumentation are key to the developing science of NO 3 -BVOC chemistry. Chamber studies provide an excellent opportunity for the comparison and validation of such instrumentation. State-of-the-art and developing instruments for measurement of NO 3 and N 2 O 5 were compared approximately a decade ago (Fuchs et al., 2012;Dorn et al., 2013). These instruments have improved and proliferated since that time, and further validation studies are needed. Measurements of total and speciated gas and aerosol-phase organic nitrates, as well as other oxygenated compounds that result from NO 3 -BVOC reactions, have not been the subject of a specific intercomparison study. Their comparison and validation will be a priority in future coordinated chamber studies.
Utility of coordinated chamber studies: Because of the need for a better understanding of NO 3 oxidation and because of the challenges of chamber studies, investigating NO 3 chemistry, coordination between studies carried out in different chambers, and between chamber and field studies, can augment efforts of single or standalone chamber studies. Coordinated studies that would include several chambers could increase the accuracy and reliability of results and quantify realistic errors associated with product yield estimates. This can be achieved by determining the same quantities in similar experiments in different chambers. Studies could benefit from complementary capabilities and properties of chambers. Chambers that typically operate at higher concentration ranges, and therefore increased oxidation rates, are suitable to perform a larger number of experiments that are useful for screening experiments and a series of experiments with systematic variations of chemical conditions. Other chambers are suited to perform experiments at atmospheric reactant concentrations. Experiments in these chambers may take place on a longer timescale, for example, a scale characteristic of the duration of at least one night. Analytical instrumentation and capability also differs considerably among chambers, so that coordinated chamber studies can make use of the determination of complementary quantities such as product yields of different organic compounds and characterization of various properties of particles for the same chemical system. For instance, it would be invaluable to conduct coordinated studies where a variety of instrument techniques are used to measure total and speciated gas-and particle-phase organic nitrates, as well as aerosol physical and chemical properties in the same chamber.
Substantial insights into aerosol sources, formation, and processing can be gained from coordinated laboratory chamber and field studies. Laboratory chamber experiments provide the fundamental data to interpret field measurements. The analysis of field data in turn can provide important insights for constraining chamber experiment parameters so that the oxidation conditions in chambers can be as representative as possible of those in the atmosphere. Two recent sets of experiments serve as examples of this approach. Fundamental chamber studies on β-pinene+NO 3 in the Georgia Tech Environmental Chamber (GTEC) facility under conditions relevant to the SE US provided constraints on the contribution of monoterpenes + NO 3 to ambient OA during the 2013 SOAS campaign Xu et al., 2015a). The Focused Isoprene eXperiment at California Institute of Technology (FIXCIT) chamber study following SOAS advanced the understanding of isoprene oxidation chemistry relevant to the SE US . It is important not to consider fundamental laboratory studies as isolated efforts, but they should be an integrated part of field studies. Similarly, having the modeling community involved in early planning stages of laboratory and field studies will greatly aid in the identification of critically needed measurement data.

Impacts of NO 3 -BVOC chemistry on air quality
The previous sections have demonstrated that understanding how NO 3 reacts with BVOC, including the ultimate fate of products, encompasses all aspects of atmospheric physics, chemistry, and transport. These sections have raised numerous complex and fascinating scientific questions and highlighted the critical need for much more basic science to fill in unknown aspects of this system. However, "getting this system right" is not just an interesting scientific problem because it has direct implications for policy decisions that governments across the world are taking to protect citizens and ecosystems from harmful effects of air pollutants. Addressing the uncertainties raised in the previous sections is critical for developing efficient, accurate, and cost-effective strategies to reduce the harmful effects of air pollution.

EPA Author Manuscript
BVOC have long been predicted to be significant contributors to regional and global O 3 (e.g., Pierce et al., 1998;Curci et al., 2009) and PM 2.5 (Pandis et al., 1991), with NO 3 reactions providing a major pathway for loss of ambient BVOC Pye et al., 2010;Xie et al., 2013). If BVOC react with NO 3 instead of OH, the O 3 production of the BVOC can be reduced relative to reactions through OH, although in some instances they may slightly increase O 3 by reducing next-day NO x . For example, measurements in St.
Louis (Millet et al., 2016) demonstrate that nights with lower levels of NO 3 resulted in higher isoprene concentrations the following morning, producing higher and earlier O 3 peaks. Recent insights into the role of biogenic nitrates, which are produced in large quantities through the reactions of NO 3 with primary emitted BVOC and subsequent reactions of their stable products, demonstrate that these compounds can substantially alter the availability of NO x (Perring et al., 2013). This highlights the importance of accurate treatment of fates of organic nitrates that form from nighttime chemistry in models, which will impact the next-day NO x and ozone levels. Organic nitrates from BVOC + NO 3 also can contribute to nitrogen deposition , which adversely impacts ecosystems. The ways in which the patterns of deposition for biogenic nitrates affect inorganic nitrate deposition remain poorly understood.

Implications for spatial distribution of ozone and PM 2.5
While it is clear that NO 3 -BVOC reactions affect oxidant availability and SOA, it remains unclear how large that role is in the ambient atmosphere relative to other VOC and other oxidants and where it occurs. The extent of O 3 formation downwind of sources is influenced by the transport of NO y species, including organic nitrates, which can release NO x downwind, where O 3 may be formed more efficiently. Biogenically derived nitrates are the dominant organic nitrates in many places (Pratt et al., 2012). A variety of different organic nitrates are formed from different BVOC, with some being short lived (releasing NO 2 locally) and others being long lived (releasing NO 2 downwind unless they are removed in the meantime). Errors in our attribution of the lifetime of individual biogenic nitrate compounds can cause errors in predicted NO x redistributions regionally and globally, and modify the spatial distributions of O 3 (Perring et al., 2013). Updates to the chemistry of BVOC-NO 3 also could alter calculations of the relative role of biogenic species versus anthropogenic pollutants to O 3 and PM 2.5 formation.

Implications for control strategy development
Air quality models are used not only to understand the production of air pollutants in the current atmosphere but also to guide the development of strategies to reduce the future pollution burden. Uncertainties in the chemistry and removal of BVOC can contribute to uncertainties in the sensitivity of O 3 and PM to emission reduction strategies. This increases the risk of implementing expensive control strategies that are found later to be inefficient (more control specified than needed) or ineffective (do not meet the air quality goals for which they were developed). As noted by Millet et al. (2016), in urban areas downwind of high isoprene emissions, the loss of isoprene by NO 3 at night can produce the opposite O 3 -NO x behavior that would normally be expected in urban areas, potentially causing a reassessment of optimum control strategies. In addition, the early O 3 peaks noted on low NO 3 nights expands the high ozone time window, resulting in higher 8 h O 3 averages, on which regulatory compliance in the US is based.
The uncertainties in our understanding of NO 3 -BVOC chemistry propagate into chemical mechanisms, as described in Sect. 3. Past work has shown that vastly different chemical mechanisms may predict similar O 3 in current atmospheres but show huge differences for intermediate species (e.g., Luecken et al., 2008) and different potential responses to precursor reductions, including different indicators of O 3 sensitivity to VOC versus NO x reductions . The presence of large weekend effects in NO x makes identifying such errors more likely in current analyses.
Incorporating new information on biogenic chemistry within a chemical mechanism will impact the availability of NO x (e.g., Archibald et al., 2010;Xie et al., 2013) and modify the predicted effectiveness of anthropogenic NO x controls. Incorporating new chemical information into models can also impact PM 2.5 sensitivities to NO x reductions. In one example, organic PM 2.5 was almost twice as responsive to a NO x reduction than in older mechanisms . Because much of the NO x dependence of O 3 and aerosols from NO 3 -BVOC reactions is inadequately accounted for in models, the few examples we have hint that current NO x control strategies might result in more significant improvements to air quality than currently assumed. Retrospective analyses should focus on elucidating the elements of this hypothesis that are represented in the historical record.
The role of climate change in modifying air quality is also a highly uncertain issue and may be particularly sensitive to the characterization of BVOC. Biogenic emissions may increase or decrease in the future, depending on many factors including increased temperatures, changes in water availability, occurrence of biotic and abiotic stress (e.g., Kleist et al., 2012;Wu et al., 2015), CO 2 fertilization, CO 2 inhibition, and land use changes (Chen et al., 2009;Squire et al., 2014). Uncertainties in biogenic reactions may be amplified as they become a larger share of the VOC burden in some places. The predicted response of O 3 to future climate has been found to be especially sensitive to assumptions about the chemical pathways of BVOC reactions, in particular the treatment of nitrates. Mao et al. (2013) and several earlier researchers found that predictions of the O 3 response to NO x reductions change from negative to positive depending solely on how the isoprene chemistry was represented. Similarly, a comparison of several widely used chemical mechanisms with varied descriptions of BVOC-derived nitrates (Squire et al., 2015) found that description of BVOC chemistry significantly alters not only the amount of oxidant change predicted under future scenarios but also the direction of the change. Direct measurements of the key steps in isoprene oxidation should eliminate the ambiguity in such model calculations. Nonetheless, the exquisite sensitivity of model predictions of ozone trends to the representation of isoprene and NO x indicates that ambient observations of those trends are an excellent strategy for evaluating the accuracy of mechanisms.
The relative distribution of emissions among different types of BVOC may also shift as climate and land use changes, emphasizing the need to understand differences among terpenes in their chemistry, transport, and fate (Pratt et al., 2012). While most of the research to date has been done on isoprene, with some on α-pinene and β-pinene, little has been done on products or reaction parameters of other terpenes. The previous sections have demonstrated that different terpenoid structures can have vastly different atmospheric chemistry and physical properties, so it is unclear whether assuming one "representative" species or distribution, as is done in most chemical mechanisms, will adequately account for future impacts of BVOC on O 3 and PM.

Summary of impacts
This review has illustrated that accurate characterization of NO 3 -BVOC chemistry is critical to our understanding of both the air quality and climate impacts of NO x emissions. Our knowledge of the complexity of NO 3 -BVOC reaction pathways and multigenerational products has advanced rapidly, especially in the last decade. Despite the fact that much of that information is not yet in a form that can be included in current air quality models, we anticipate improved predictive capabilities in models in the coming years through sustained laboratory and field studies coupled to model development. While the current levels of uncertainty make it difficult to accurately quantify the impact of NO 3 -BVOC chemistry on air pollutant concentrations, we expect that developments in this field will improve the effectiveness of air pollution control strategies going forward. The limited studies available demonstrate that even small changes to BVOC chemistry modify the production of oxidants (NO 3 , OH, and O 3 ) and change the transport of NO y . Therefore, NO 3 -BVOC oxidation modifies the chemical regime in which additional BVOC oxidation occurs. Of most importance will be the studies that indicate changes in the direction of predicted future pollutant concentrations as chemical mechanisms of BVOC are updated. Emission control strategies and attainment of air quality goals rely on the best possible chemical models. Current and future laboratory and field research is critical to the improvement of chemical mechanisms that account for biogenic chemical processes and products which will augment efforts to reduce harmful air pollutants.

Supplementary Material
Refer to Web version on PubMed Central for supplementary material.

TD-CRDS
Thermal dissociation cavity ring-down spectroscopy  Condensed reaction mechanism for isoprene and β-pinene oxidation via NO 3 (adapted from Schwantes et al., 2015 and. For brevity, only products generated from the dominant peroxy radicals (RO 2 ) are shown. R′ represents an alkoxy radical, carbonyl compound, or hydroxy compound. Two of the largest uncertainties in β-pinene oxidation are shown in red: (1) quantification of product yields from the RO 2 + HO 2 channel and (2) identification of carbonyl products formed from RO 2 reaction with NO 3 , RO 2 , or HO 2 (see text for more details).   Table S1 in the Supplement.   (a) Average mass concentrations (in μg m −3 , ambient temperature and pressure) of submicrometer particulate organic nitrates (NO 3, org ) and particulate inorganic nitrates (NO 3,inorg ) in different months at multiple sites. The concentrations correspond to mass concentrations of -ONO 2 functionality. Note that the y axis is different for sites with total nitrates greater than 1 μg m −3 (shaded). Detailed information and measurements for each site are provided in Table S5. (b) Percentage (by mass; cyan) of submicrometer particulate organic nitrate aerosols in ambient organic aerosols in different months at multiple sites. Detailed information and measurements for each site are provided in Table S5.  The real part of refractive index (RI) (mr) for biogenic SOA compiled from several chamber studies. The legend specifies the precursor type and oxidation pathway as well as the reference. The figure is reprinted with permission from Moise et al. (2015).  Table 4 Gas-phase organic nitrate yields (in percent) from BVOC + NO 3 systems in current chemical mechanisms.

TD-LIF
Gas-phase organic nitrate yields depend on RO 2 fate as indicated in the ternary diagrams; clockwise from the top: RO 2 reacts with NO 3 , RO 2 , and HO 2 .