Temperature and Recent Trends in the Chemistry of Continental

Review pubs.acs.org/CR

Temperature and Recent Trends in the Chemistry of Continental Surface Ozone Sally E. Pusede,†,∥ Allison L. Steiner,§ and Ronald C. Cohen*,†,‡ †

Department of Chemistry, and ‡Department of Earth and Planetary Science, University of California Berkeley, Berkeley, California 94720, United States § Department of Atmospheric, Oceanic and Space Sciences, University of Michigan, Ann Arbor, Michigan 48109, United States References

1. INTRODUCTION Ozone (O3) is a harmful air pollutant1−5 produced in the atmosphere by complex, nonlinear, and temperature-dependent chemistry. In the future, global temperatures are projected to rise and the emissions of O3 forming precursors are expected to change.6 Therefore, a mechanistic understanding of the O3−temperature relationship is not only a key aspect of the basic science of tropospheric oxidation, it is also needed to effectively regulate O3 air quality over the coming decades. Reports of a correlation between high O3 and hot temperatures are ubiquitous; however, the temperature-driven chemical controls over O3 are not well constrained, even though the outline of chemical and physical mechanisms causing O3 in the troposphere is generally understood. The fundamental mixture of gas-phase organic molecules, nitrogen oxides (NOx ≡ NO + NO2), and sunlight that forms O3 was first recognized by Arie Haagen-Smit in the 1940s.7 Haagen-Smit was investigating a unique, regional crop damage, which correlated with local smog events and was only observed in Los Angeles, CA.8 Because the smog also triggered respiratory ailments and eye irritation in Los Angeles residents, soon after O3 and its precursors were identitfied,9 civic and political momentum forced county-level controls on O3 forming emissions.10,11 As air quality worsened in many U.S. cities, the 1970 federal Clean Air Act set nation-wide limits on ambient O3 concentrations (and other pollutants). National standards also required reductions of O3 precursors, and, as a result, NOx and anthropogenic organic emissions have fallen consistently across the country ever since.12−16 The large increases in emissions prior to mandated controls and subsequent regulatory decreases have altered the abundance of limiting reagents dictating atmospheric oxidation pathways and O3 formation. While O3 concentrations have been observed to decrease, increase, or stay the same in response to any specific policy change, the accumulated effects of sustained controls on precursors over multiple decades are dramatic reductions in the number of high O3 days in the U.S.13,17 Climate change will likely impact the effectiveness of current emission reduction strategies.18−20 It is predicted that

CONTENTS 1. Introduction 2. O3 Production and Temperature 3. Organic Reactivity 3.1. Anthropogenic Organic Reactivity 3.2. Biogenic Organic Reactivity 3.3. Organic Reactivity and Changes in Emissions 4. HOx Production and Concentration 4.1. HOx Precursors 4.2. PHOx and Changes in Emissions 5. NOx Emissions and Concentration 5.1. NOx Emissions 5.2. NOx Lifetime 5.3. NOx and Changes in Emissions 6. Peroxy Nitrates as a Control over NOx Concentrations 7. Alkyl Nitrates 7.1. αi 7.2. αnet 7.3. αnet and Changes in Emissions 8. Changes in O3 with Temperature 8.1. ∂O3/∂T 8.2. mO3−T 9. Regulating High O3 as a Function of Temperature 10. Conclusions and Open Questions Author Information Corresponding Author Present Address Notes Biographies Acknowledgments Abbreviations © XXXX American Chemical Society

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A B D D D E F F F G G G G H H H I J J J M N O P P P P P P Q

Special Issue: 2015 Chemistry in Climate Received: December 1, 2014

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radicals (RO2) (R1 and R2). R1 depicts OH reactions that proceed by H abstraction; however, if the RH contains a carbon−carbon double bond, oxidation occurs by OH addition. In the sequence of reactions leading to O3, RO2 then reacts with NO to yield RO and NO2, R3a. There is a branch of this reaction, R3b, that forms an alkyl nitrate (RONO2) and is chain terminating, but the dominant pathway is R3a (70−100%). Typically, the alkoxy radical (RO) formed in R3a reacts with O2 to produce HO2 and an aldehyde or ketone (R′O), R4. HO2 then reacts with NO to yield a second NO2 and regenerate OH, R5. Hence, the cycle is catalytic with respect to HOx (HOx ≡ OH + HO2 + RO2). The R′O can be further oxidized, restarting the cycle at R1.

surface O3 will be affected by warmer average surface temperatures,6 shifts in global circulation patterns,21 changes in the frequency of heat waves and frontal passages,6,18,22−26 altered surface mixed-layer depths (very uncertain),22,27 and variations in cloud cover, precipitation, and convection.24,28−30 Climate change will likely also modify patterns in fires, vegetation, and land use, which are all large sources of O3 precursors to the atmosphere.20,31−35 Here, we review one aspect of the scientific questions surrounding future surface O3, the O3 production (PO3)−temperature response. To do this, we describe the nonlinear chemistry of PO3, focusing on the underlying factors of this chemistry that vary with both temperature and precursor emission rates, relying mainly on observations to guide our thinking about future O3 trends. We are specifically interested in describing chemical mechanisms that lead to high O3 concentrations, with conditions conducive to high O3 typically occurring between late spring and early fall and found in inhabited areas and other locations where NOx emissions are large. Although urban areas comprise less than a few percent of the Earth’s surface, they are home to most of the world’s population. We begin by summarizing the mechanism of PO3 (section 2) and center our discussion on five chemical variables that play a role in determining the magnitude and temperature dependence of PO3. These are organic reactivity to the hydroxyl radical (section 3), HOx radical production and concentration (section 4), NOx (section 5), peroxy nitrates (section 6), and alkyl nitrates (section 7). Unlike in the laboratory where variables can be controlled one by one, changes in the atmosphere often involve many variables at the same time, challenging our ability to pinpoint causal chemical factors. To isolate the temperature-dependent effects of individual variables from the complex chemistry of PO3, we use examples from the literature where each chemical variable is the most important driver under local conditions. We also describe our understanding of the temperature-dependent impacts of these chemical drivers in the context of interannual emissions trends because, as precursor emissions change, so do the temperature relationships of chemical factors controlling PO3. In section 8, we review analyses investigating total temperature-driven effects on O3, with most of these studies having much less chemical detail than described in sections 2−7. We focus on the role of chemistry, linking results to the temperature dependence of PO3 and also discussing the slope of the correlation of O3 with temperature (mO3−T). In section 9, we describe implications of the temperature dependence of PO3 for the effectiveness of future regulatory actions to reduce high O3 and for chemical strategies for mitigating the impacts of warming temperatures. In section 10, we conclude by presenting directions for future research.

RH + OH → R + H 2O

(R1)

R + O2 + M → RO2 + M

(R2)

RO2 + NO → RO + NO2

(R3a)

RO2 + NO + M → RONO2 + M

(R3b)

RO + O2 → R′O + HO2

(R4)

HO2 + NO → OH + NO2

(R5)

RH + 2O2 + 2NO → R′O + H 2O + 2NO2

(series 1)

The net effect of R1−R5 via branch R3a is summed in series 1, where two NO2 molecules are produced. In the daylight, these NO2 photolyze to yield two new O3 (R6 and R7). NO2 + hν → NO + O(3P)

(R6)

O(3P) + O2 + M → O3 + M

(R7)

The importance of any given organic molecule in series 1 to PO3 results from both its concentration and its reactivity with OH. Although much of the engineering literature focuses on organic mass and O3 yields during multiple steps in the oxidation of individual organic precursors, the “organic reactivity” is a more useful construct from a chemical point of view. Equations 1 and 2 represent the organic reactivity for a single compound (RHOHi) and for the entire atmospheric organic mixture (RHOH), respectively. RHOHi = k OH + RHi[RHi] RHOH =

∑ k OH + RHi[RHi] i

(1)

(2)

Under conditions of rapid photochemistry and when a system is in steady-state, PO3 resulting from R1−R7 is summarized equivalently by eqs 3−5.36 In eq 3, PO3 is described by the oxidation rate of organic molecules and the number of O3 formed per oxidation. Primary emissions and secondary oxidation products, both represented by RHi, are counted separately in this formulation. The RONO2 branching ratio (αi) specifies the branching between R3a and R3b for any RHi and is equal to k3b/(k3a + k3b). The number of NO2, and hence O3, produced per cycle is γi. In series 1, all γi are two37 and all αi are zero. For a molecule like CO, which enters the HOx cycle as HO2, γi equals one. For an organic species with a carbon backbone of four or greater, γi may be greater than two, as the RO produced by R3a may isomerize to form a new R radical via intramolecular H-abstraction. When such a hydrogen shift occurs, the HOx cycle is reset to R2 and a second RO2 is formed. In eq 4, PO3 is described by the

2. O3 PRODUCTION AND TEMPERATURE The chemistry that produces O3 (PO3) is nonlinear, and the chemical terms driving PO3 vary nonlinearly with temperature. Here, we describe the PO3 mechanism, emphasizing variables that are important to understanding the response to both changes in temperature and variations in precursor emissions. We begin our discussion of PO3 at the oxidation of gas-phase organic molecules by the hydroxyl radical (OH), which is the usual atmospheric oxidant. Reactions between organic compounds (RH) and OH yield alkyl radicals (R), and these R’s immediately combine with O2 to give alkyl peroxy B

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HNO3 is much faster than PO3, and chemistry in this regime is known synonymously as “NOx-suppressed”, “NOx-saturated”, and “VOC-limited” (VOC is an abbreviation for volatile organic compound, or gas-phase RH). Under conditions of constant organic reactivity (along a y-axis in Figure 1), PO3 reaches a maximum near the transition point between the two chemical regimes. At constant NOx (along an x-axis in Figure 1), PO3 is almost unchanged in response to increasing organic reactivity in the NOx-limited regime and very sensitive to increasing organic reactivity in the VOClimited regime. The peak in PO3 occurs at NOx concentrations that are typical of suburban locations or, in recent years, urban downtown areas in the U.S. Peak PO3 shifts to higher NOx as organic reactivity increases, presenting a challenge for O3 mitigation policies, as most practical strategies rely on simultaneous reductions of both organic reactivity and NOx. The effectiveness of any specific control therefore varies with the relative atmospheric abundance of both precursors. To summarize, (a) organic reactivity determines the NOx concentration of peak PO3, as well as setting the absolute rate of PO3 at NOx concentrations higher than at peak PO3, and (b) NOx determines whether any precursor change, either organic reactivity or NOx, shifts O3 chemistry to or from peak PO3. PO3 is also a nonlinear function of the production rate of new HOx molecules (PHOx). PHOx is set by the abundance of HOx precursors and the photolysis rates of these species. Under chemical conditions that are VOC-limited or near peak PO3, increasing PHOx enhances PO3 linearly. If PO3 is NOxlimited, PO3 varies as the square root of PHOx. Decreasing PHOx also shifts peak PO3 to lower NOx, but the impact is smaller than from a decrease in organic reactivity. For example, at 10 s−1 organic reactivity and 1 ppt s−1 PHOx, a 50% reduction in organic reactivity lowers the NO x concentration of peak PO3 by ∼30%; a 50% reduction in PHOx moves the peak by 15%. If the dominant source of PHOx is photolysis of O3 to produce O(1D) followed by O(1D) + H2O, then changes in O3 will feedback on PO3. The production of peroxy nitrates (RO2NO2) (R13) and that of RONO2 (R3b) are key termination reactions in the PO3 mechanism. Characteristics of the chemistry of RO2NO2 and RONO2 that relate to their temperature-dependent impacts on PO3 are discussed in sections 6 and 7, respectively. The formation of these organic nitrates results in redistribution of NOx by meteorological processes. In the near field of NOx emissions, organic nitrate formation suppresses PO3 by removing NOx from the local atmosphere and storing it as organic reactivity. In downwind regions, these organic nitrates break down, returning NOx to an atmosphere with few local sources. Although this chemistry affects the concentration of NOx and the absolute rate of PO3, it does little to alter the functional form of PO3 as it is depicted in Figure 1.42,43 These five chemical factors of PO3, organic reactivity, HOx, NOx, RO2NO2, and RONO2, and their relationships with temperature play a large role in determining the nonlinear temperature dependence of PO3 in the continental boundary layer. In sections 3−7, we think about the impacts of these five drivers on PO3 individually. In sections 8−10, we reassemble the pieces.

number of NO2 molecules produced by peroxy radical oxidation of NO (as opposed to NO reaction with O3, which is O3 neutral).38,39 In eq 5, PO3 equals the number of new NO2, described as the difference between the total NO2 photolyzed at a rate jNO2 and the NO2 produced by reaction of NO with O3.40 Any term in these equations that varies with temperature will influence the temperature dependence of PO3. PO3 =

∑ (1 − αi)γik OH + RHi[RHi][OH] i

(3)

PO3 = kNO + RO2[NO][RO2 ] + kNO + HO2[NO][HO2 ] ≈ 2kNO + HO2[NO][HO2 ]

PO3 = jNO2 [NO2 ] − kNO + O3[NO][O3]

(4) (5)

The functional form of PO3 is nonlinear with respect to both organic reactivity and NOx (Figure 1). Nonlinearity arises

Figure 1. PO3 (contours have units ppb h−1) mapped as a function of NOx (ppb) and organic reactivity (s−1) for daytime conditions in a typical U.S. city under current emissions (2015). Organic reactivity is defined as in eq 2. PO3 contours are computed with an analytical model based on the assumptions that HOx is in steady state (PHOx equals the sum of R3b and R8−R13) and that [RO2] = [HO2].41

because HOx and NOx drive PO3, R1−R7, but also terminate this same free radical chemistry, R3b and R8−R13. OH + HO2 → H 2O + O2

(R8)

HO2 + HO2 → H 2O2 + O2

(R9)

RO2 + HO2 → ROOH + O2

(R10)

RO2 + RO2 → ROOR + O2

(R11)

OH + NO2 + M → HNO3 + M

(R12)

RO2 + NO2 + M → RO2 NO2 + M

(R13)

At low NOx, increasing NO enhances R3a and R5, and therefore increases the oxidation rate of RH to produce O3. The effect is approximately linear in NOx, and, consequently, this chemical regime is often described as being “NOxlimited”. Increases in organic reactivity have a small effect on PO3 when chemistry is NOx-limited. As the NOx concentration increases, the reaction of OH with NO2 (R12) begins to dominate the fate of HOx, slowing the rate of RH oxidation and decreasing PO3. At high NOx, the production of C

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Figure 2. Time series of mixing ratio (ppb) data (gray) for o-xylene (top left) and formic acid (bottom left) observed near Bakersfield, CA. Daily average (10 am−2 pm LT) mixing ratios are shown with “●”. The temperature dependence of the daily averages of each species is shown (top and bottom right). Adapted with permission from ref 46. Copyright 2014 Copernicus Publications.

3. ORGANIC REACTIVITY Organic reactivity (eqs 1 and 2) depends on temperature in two ways: (1) emissions of organic molecules exhibit some relationship with temperature, with processes dictating emission rates ranging from temperature independence to exponential temperature dependence, and (2) reaction rates between RHi and OH (kOH+RHi) also vary with temperature. The impact on PO3 of any specific temperature-driven or regulatory change in organic reactivity is determined by whether chemistry is NOx-limited or VOC-limited. If chemistry is VOC-limited, then increases in organic reactivity enhance PO3; if chemistry is NOx-limited, the same variation has little effect. Organic molecules are emitted to the atmosphere from many anthropogenic and biogenic sources that are both independent of and strongly dependent on temperature. In locations with largely temperature-dependent organic emissions, the contours of PO3 as a function of NOx and temperature have a shape similar to those as a function of NOx and organic reactivity. The direct mapping of PO3 and organic reactivity to PO3 and temperature has been borne out in locations with distinct and diverse organic reactivity compositions, for example, in Sacramento, CA,44 where organic emissions are mostly from nearby forests, and in Bakersfield, CA, where organic emissions are in large part from agricultural activities and energy development.45,46 To an important extent, a key variable determining the temperaturedependent impact of organic reactivity on PO3 is the emission rate as a function of temperature, not the molecular speciation of the reactivity (we revisit this statement in section 7). The organic reactivity metric also includes temperature’s influence on bimolecular reaction rates. Elementary bimolecular rates of reaction are Arrhenius in form, becoming more rapid with higher temperatures. This temperature dependence is incorporated into most atmospheric chemical transport models and has been shown to be important to reproduce observed O3 abundances versus temperature.47,48 Generally, the temperature dependence of reaction rates exerts a smaller influence over the organic reactivity than temperature-driven emission rates.49 In the remote atmosphere, where methane (CH4) dominates the organic reactivity, the temperature

dependence of the oxidation of CH4 by OH is more important than of variations in emission rates. 3.1. Anthropogenic Organic Reactivity

Major anthropogenic sources of organic reactivity are associated with transportation, energy generation, goods production, and agricultural activities. Tailpipe emissions are presently the biggest fraction of transportation-related emissions and are mostly independent of temperature.50−52 Evaporative emissions, such as volatized organics from gasoline, agriculture, and industrial activities (e.g., refining, chemical production, and energy development), increase with temperature according to the vapor pressures of individual compounds. For most molecules, at least in the near field of sources, temperature-driven changes in emissions cause parallel changes in ambient concentrations. This is exemplified in Figure 2, where o-xylene and formic acid are shown as a function of daytime temperature at a measurement site in the southern San Joaquin Valley city of Bakersfield.45 These observations suggest o-xylene is emitted to the atmosphere by a temperature-independent processes, consistent with other evidence that the main sources of xylenes are tailpipe emissions. By contrast, daytime formic acid concentrations increase exponentially with temperature. Formic acid sources are not well understood, but this behavior is consistent with an evaporative source, likely agricultural, and/or with photochemical production in the atmosphere by oxidation of a different temperature-dependent molecule.53 The overall importance of temperature-independent versus temperaturedependent reactivity to PO3 results from a sum over all organic compounds. As a consequence, which organic compounds drive organic reactivity varies as a function of temperature, with temperature-independent emissions dominating when it is cool and temperature-dependent emissions dominating when it is hot. 3.2. Biogenic Organic Reactivity

Temperature is a well-known control over biogenic emission rates.54−58 Globally, the biogenic species isoprene (C5H8) is the single most important source of organic reactivity in the continental boundary layer, dominating the reactivity on warm and hot days in many remote, rural, and even urban locations. D

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compounds than are possible to identify and quantify with current techniques.66,67 Analyses of field data have shown that even when the fluxes of individual molecules are small, the sum of many small fluxes equates to large reactivity fluxes.67 In addition, rapid oxidation occurs within forest canopies, converting source molecules into species that are more difficult to measure and affecting estimates of emissions and reactivity.68,69 In each case, these reactivity sources are expected to be temperature-dependent and are not included in O3 simulations with explicit chemical mechanisms.

Produced within plant leaves by a process linked to photosynthesis, isoprene emissions approach zero in the absence of photosynthetically active radiation. Enzymatic isoprene synthesis is temperature driven, and isoprene emissions increase with increasing ambient temperature.59 At temperatures >∼40 °C, isoprene synthase denatures, inhibiting emissions and slowing the otherwise exponential increase.55,60 The effect of suppressed isoprene at >∼40 °C on PO3 is large and has been indirectly observed in the atmosphere via measured concentrations of isoprene and O3.61 Other important classes of temperature-dependent biogenic reactivity are monoterpenes (C10H16) and sesquiterpenes (C15H24), which are emitted by conifers and many flowering species. Emissions occur from plant resins and surface pools, responding exponentially to temperature, and, in contrast to isoprene, usually do not require the presence of light.62 The combined temperature dependence of speciated and total organic reactivity was first reported by Di Carlo et al.63 for an isoprene-rich forested site in Michigan (Figure 3).

3.3. Organic Reactivity and Changes in Emissions

There are no long-term records of the total organic reactivity, and, as described above, the total organic reactivity is a sum of terms with distinct functional forms and varying relative contributions at different temperatures. Most of the existing speciated measurements are of molecules associated with gasoline-powered vehicle emissions (temperature independent). These data indicate organic vehicle emissions have responded to regulatory controls.70−76 In Los Angeles, such observations extend back to the 1960s, with measured species having fallen at an average rate of ∼7.5% year−1.14 Analyses in multiple U.S. cities suggest tailpipe organic reactivity and NOx emissions from gasoline-powered vehicles have decreased in tandem.77,78 These trends are not easily extrapolated to changes in the total organic reactivity. This is because the molecules routinely measured are the least reactive components of tailpipe emissions, because data are not typically reported in the literature as a function of temperature, and because observed species do not include the other major components of the total organic reactivity in many locations, both urban and rural. The combined effect is likely to bias the reactivity low on warm and hot days, which is usually when high O3 is most frequent. Techniques for the direct measurement of the total OH reactivity have only recently been developed,79−82 and observations are currently limited to multiweek field experiments preventing assessments of trends over time. To relate known interannual reductions in gasolinepowered vehicle emissions to changes in total organic reactivity over the past decade, Pusede et al.46 took advantage of the distinct temperature relationships of speciated organic reactivity. The authors sorted an observed suite of over 120 organic compounds measured near Bakersfield, CA, as either temperature-independent or temperature-dependent and combined this classification with documented activities of U.S. and state-level regulators. Emissions of temperatureindependent organic reactivity had decreased due to improvements in catalytic convertor technology at a rate of ∼6% year−1, which was determined using routine canister data in this region. At the same time, there was no evidence that emissions of temperature-dependent reactivity had also been regulated from any sector over the last 10 years. As a result, the controlled portion of the reactivity comprised a larger portion of the total reactivity at cool temperatures, while the uncontrolled portion was the dominant fraction at high temperatures. Pusede et al.46 estimated that the total organic reactivity had decreased from 2000 to 2010, but the percent decrease was much greater on days with maximum temperatures of 20 °C (50%) than on days with maximum temperatures of 40 °C (17%) (Figure 4). Similarly, continued controls on emissions are predicted to reduce organic

Figure 3. (A) Measured reactivity (○) and the total OH reactivity (s−1) calculated as the sum of speciated reactivities (■) versus temperature (K). (B) The total calculated OH reactivity (■) and the individual reactivities of isoprene (orange □), CO and NO2 (green ◇), formaldehyde (CH2O) (+), acetaldehyde (magenta ○), anthropogenic VOCs (red ▽), and terpenes (blue ×). Reprinted with permission from ref 63. Copyright 2004 American Association for the Advancement of Science.

Here, the isoprene reactivity increased by over 5 s−1 (more than a factor of 5) as temperature increased from 15 to 25 °C. A portion of the total reactivity was unaccounted for with observations of individual compounds, and this missing portion of reactivity was also temperature dependent, increasing from ∼1 s−1 at 15 °C to almost 4 s−1 at 25 °C. Because of this observed temperature dependence, the authors ruled out anthropogenic sources and speculated that the unknown reactivity was of origin similar to that of monoterpenes. Analytical advancements in the following decade have shown that the biosphere emits a variable and complex assortment of temperature-dependent organic species.64,65 Still, the within-canopy gaseous mixture includes more E

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oxidation and are important sources of HOx in the afternoon and at high altitudes.85−89 In the morning, the dominant contributors to PHOx may be HONO (R20) and ClNO2 (R21 and R22).90 Steep HONO concentration gradients are generally observed in the lowest tens of meters above the ground,91−98 and because HONO’s photolytic lifetime is very short, the morning PHOx from this source is localized very near the surface.90 In contrast, ClNO2 is formed in elevated nocturnal residual layers99−101 and is more important higher in the morning boundary layer.90 The temperature dependence of either HONO or ClNO2 has not yet been reported. Finally, the reaction between O3 and an alkene yields new HOx molecules (R23b). This source is not photolytic and has its greatest relative impact in the dark.

reactivity at low temperatures by another 50% by 2020, but only by 10% at 40 °C.

Figure 4. Effects of temperature-independent emissions reductions on the total organic reactivity as a function of the daily maximum temperature (°C). The temperature-dependent reactivity was assumed to be unchanged. The percent difference (higher) in the total organic reactivity in 2000, (2000−2010)/2000, is in black. The percent change (lower) in organic reactivity in 2020, (2020−2010)/ 2010, is in periwinkle. The turquoise line is at 0%. Reprinted with permission from ref 46. Copyright 2014 Copernicus Publications.

It is reasonable to expect that total organic reactivity is changing more slowly at high temperatures than at low and moderate temperatures, not just in Bakersfield, CA, but also in any location where a substantial portion of reactivity is temperature-dependent. The impact of temperature-varying organic reactivity on PO3 is therefore 2-fold: (1) changes in a single portion of the reactivity likely do not translate to changes in organic reactivity at all temperatures, and (2) the ratio of organic reactivity to NOx, which is control over the chemical sensitivity of PO3, is likely also temperature dependent. As a result, the impact of all changes in organic reactivity on PO3 will be temperature dependent.44,45,83,84

O3 + hν → O(1D) + O2

(R14)

O(1D) + H 2O → 2OH

(R15)

O(1D) + N2 , O2 → O(3P) + N2 , O2

(R16)

CH 2O + hν + 2O2 → 2HO2 + CO

(R17)

CH 2O + hν → H 2 + CO

(R18)

H 2O2 + hν → 2OH

(R19)

HONO + hν → OH + NO

(R20)

ClNO2 + hν → Cl + NO2

(R21)

Cl + RH → HCl + R

(R22)

RR + O3 → 2R′O

(R23a)

RR + O3 → HOx + 2R′O

(R23b)

The nonlinear dependence of PO3 on PHOx and NOx has been directly observed in the atmosphere.40,102,103 In the highbiogenic reactivity city of Nashville, TN, Thornton et al.40 used a detailed suite of observations to calculate the influence of variations in PHOx on PO3. The authors calculated faster PO3 with increasing PHOx at all NOx concentrations, but found that increases in PHOx had more than twice the impact when chemistry was VOC-limited. The two HOx precursors included in their analysis were O(1D) + H2O and CH2O, both of which under these conditions were likely to positively correlate with temperature. If PHOx is temperature dependent, the contours of PO3 as a function of NOx and temperature will approximate those as a function of NOx and PHOx.

4. HOX PRODUCTION AND CONCENTRATION The production of new HOx radicals (PHOx) varies with temperature if (1) the concentrations of HOx precursors are temperature dependent and (2) light levels, and hence photolysis rates, correlate with temperature. Temperaturedriven changes in PHOx impact PO3 as a function of NOx, with larger enhancements in PO3 occurring with increased PHOx under conditions that are VOC-limited.40 When chemistry is NOx-limited and the impact of PHOx on PO3 is at a minimum, reactions between O3 and H2O may result in a net O3 loss in response to temperature-driven increases in humidity.

4.2. PHOx and Changes in Emissions

There is little observational information about interannual trends in HOx or PHOx. Concentrations of OH and HO2 have been measured as part of field experiments over the past decade; yet, there have been few analyses of the temperaturedependent chemistry of these radicals,69,104,105 and there are no direct or inferred records of OH over multiyear time scales that are useful for interpreting trends in continental PO3. As O3 concentrations decline across the U.S., PHOx from O(1D) + H2O will also decrease, particularly in VOC-limited cities, where lower O3 concentrations positively feedback more strongly on PO3. Decadal trends in CH2O have been measured from space by the GOME and SCIAMACHY instruments (1997−2009) and shown to be downward in many cities in the Northeast U.S. (−1 to 5%) and upward in

4.1. HOx Precursors

The dominant HOx precursors in the near-surface atmosphere are typically O(1D) + H2O (R14−R16) and formaldehyde (CH2O) (R17 and R18), followed by nitrous acid (HONO), nitryl chloride (ClNO2), peroxides, and reactions between alkenes and O3. The importance of each HOx precursor varies with height in the atmosphere and with time of day. The O(1D) + H2O source is often the biggest source of HOx radicals at the Earth’s surface, and O3 and H2O both increase with increasing temperature. CH2O, and other carbonyls, and peroxides (R19) are secondary photoproducts of atmospheric F

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5.2. NOx Lifetime

major cities in California (2−5%), in China (2−4%), and in a number of other cities in Asia and Africa (1−3%).106 Changes of this magnitude are too small to significantly impact PO3, although changes in CH2O have a combined impact on organic reactivity as well. Trends in other PHOx sources are virtually unconstrained with observations, but are expected to have varied with controls on organic reactivity and NOx.

Temperature is expected to influence both the chemistry and the emissions that affect the NOx lifetime. Conceptually, the impacts of temperature on the NOx lifetime and thus the urban−rural NOx gradient of NOx are illustrated in Figure 5.

5. NOX EMISSIONS AND CONCENTRATION NOx is emitted to the atmosphere from a variety of sources including motor vehicles and power plants, microbial activity in soils, fires, and lightning. The lifetime of NOx in the summertime boundary layer is ∼6 h (1/e), and, as a result, NOx abundances are colocated with sources and urban−rural gradients are steep. Relationships between NOx concentration and temperature are dictated by the temperature-dependence of (1) NOx emissions and (2) the NOx lifetime. 5.1. NOx Emissions

NOx emissions from combustion depend on the temperature of the reaction vessel and not the ambient temperature; however, fuel usage and electrical demand, constituting 64% of the global burden,107 may change with temperature. Singh and Sloan108 quantified NOx emission changes from the frequency of cold engine starts and from vehicle air conditioning and found emissions decreased by ∼1.5% °C−1 between −10 and 20 °C and increased by ∼1% °C−1 over the range 20−35 °C. In the Eastern U.S., power plant NOx emissions have been shown to increase with increasing ambient temperature due to greater demand for air conditioning by ∼2−4% °C−1 over the temperature range 15−30 °C.109 He et al.109 attributed this increase in demand for air conditioning to an enhancement of 0.1−0.25 ppb NOx °C−1 at the surface in the urban Baltimore−Washington region. In cities in California, where there are no large power plants and mobile emissions are the dominant NOx source,110 NOx concentrations have been observed to be independent of temperature in the near-field of urban emissions in Sacramento104 and Bakersfield, CA.46 Outside of cities, microbial activity in soils is an important source of NOx, contributing 11% of emissions globally.107 Soil NOx emission rates depend on both temperature, generally increasing with increasing temperature until leveling off at high temperatures, and other variables influencing nitrification/denitrification that also vary with temperature, in particular soil wetness. Most chemical transport models include some temperature-dependent description of soil NOx emissions, calling different functions when soil is wet or dry.111−113 NOx emissions from fires and biofuel burning represent 14%107 of the global budget. These emissions vary with fuel type, combustion stage (flaming versus smoldering), and season. Biomass burning also emits organic reactivity, and the impacts of fire NOx emission are regional in scale. Fires occur regularly in many locations around the world, with evidence suggesting the number and size of fires vary with spring and summer temperatures, the timing of spring snowmelt, and the frequency precipitation.114−117 Lightning (10% of global NOx emissions)107 has its largest impact on NOx concentrations in the upper troposphere, away from the surface. This source may be temperaturedependent when storm activity correlates with surface temperature.

Figure 5. Hypothetical effects of temperature-driven increases in PHOx along a transect of a large NOx source or urban plume transported downwind. NOx chemical loss is illustrated at high (red) and low (blue) temperatures. Emissions from a point or area source are located only at the origin (x = 0) and advected downwind. Wind speed (plume dilution) is temperature-independent. Adapted with permission from ref 118. Copyright 2013 John Wiley and Sons.

The x-axis range adapts the same NO2 plume spatial extent as observed by Valin et al.118 To create Figure 5, we assume PO3 is VOC-limited within the urban core, that emissions only occur at the origin (x = 0), that wind speeds are temperature invariant, and that HOx and organic nitrate precursors are temperature dependent. The cartoon indicates a more rapid conversion of NOx to HNO3 associated with enhanced HOx, and the results are lower NO2 concentrations on warmer days and steeper initial NO2 declines. Thus, the size of the urban plume, its integrated abundance, and the sharpness of its edges are all predicted to be temperature dependent. This is a hypothetical discussion, as we have not quantified the change in NO2 (the y-axis has an arbitrary scale); however, we speculate that this effect should be observable using spacebased NO2 observations allowing a test of a full model calculation that would include the impacts of increased nearfield removal, long-range transport, and decomposition of RO2NO2 downwind. 5.3. NOx and Changes in Emissions

Over the past decade, large decreases in NOx concentrations have been observed in U.S. cities 15,119 and across Europe. 120−123 By contrast, NO x has increased in China.120,124,125 In the U.S., reductions in NOx emissions have primarily been achieved with improved three-way catalytic converters on gasoline-powered vehicles since 198113,74,126−128 and through a variety of engineering solutions on power plants.15,119,129 Diesel NOx emissions have been increasing both in magnitude and as a fraction of the total over the past three decades. This is due to decreases in gasoline-powered vehicle emissions, 3-fold faster growth in diesel fuel sales as compared to light-duty vehicles, and early engineering challenges in controlling diesel NOx emissions from trucks.130 In the coming decades, NOx emissions from diesel engines are poised to dramatically decline due to the introduction of G

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precursor species are expected to increase with temperature. LaFranchi et al.156 calculated the peroxyacetyl radical production rates as a function of temperature constrained with observations of organic precursors at a forested site downwind of Sacramento, CA. The sources of peroxyacetyl radical were found to be strongly dependent on temperature, following the temperature dependence of emitted biogenic precursors. The two temperature-dependent processes affecting RO2NO2 concentrations have opposite effects. Observations and models indicate that the increase in RO2 at warm temperatures usually outweighs the increase in the decomposition rate. As a result, NOx is more rapidly removed from the atmosphere at warm temperatures and more effectively transported downwind.105 The impact of this temperaturedependent chemistry is to increase PO3 by decreasing NOx from a system at the VOC-limited side of peak PO3. The release of this bound NOx in regions far removed from NOx sources, that is, into a system at the NOx-limited side of peak PO3, can lead to as many as ∼75 O3 molecules produced per molecule of NOx released.

selective catalytic reduction (SCR) technology and other after exhaust treatments along with the U.S. Environmental Protection Agency’s 2008 National Clean Diesel Campaign. Reductions in diesel NOx as great as 90% are expected.127,128 The effects will come first in states with active retrofit/engine replacement programs such as California, as otherwise diesel engines have multidecade service lives.127 The impacts of the new diesel controls have already been substantial at major ports in Los Angeles and Long Beach, CA, where engine retrofit/replacement requirements have been fast tracked.131 At the Port of Los Angeles, the average fleet age fell by 80% and NOx emissions decreased by 45%, a 4-fold greater reduction than observed in greater Los Angeles inland from the ports. Along with other regional, state, and district-level NOx control efforts, these diesel NOx regulations have the potential to transform urban PO3 in the U.S., shifting urban summertime PO3 from VOC-limited to mostly NOx-limited chemistry.

6. PEROXY NITRATES AS A CONTROL OVER NOX CONCENTRATIONS Peroxy nitrates are weakly bound organic nitrates of the form RO2NO2. They are produced by the reaction of RO2 and NO2 (R24) and represent an important, although often temporary, termination step in the chemical mechanism of PO3. A key role of RO2NO2 is to sequester NO2 in the near field of NOx emissions and transport bound NO2 to downwind regions, where RO2NO2 may thermally decompose and release NO2 back to the active pool of radicals.132−140 In this case, RO2NO2 are a source of NOx and HOx to NOxlimited areas. This input of NOx increases PO3, in effect transporting PO3 out of cities to rural and remote locations where more O3 is produced per unit NOx oxidized to HNO3.49,136,141−144

7. ALKYL NITRATES Alkyl and multifunctional nitrates (RONO2) are uniquely important to the chemistry of O3.43 Formed via the minor channel of reaction R3, R3b, the production of RONO2 directly decreases the number of O3 produced and terminates HOx radical cycling in the near field of NOx emissions. RONO2 are not as easily fragmented to return NOx to the active radical pool as RO2NO2, and it is generally thought that RONO2 are a less important mechanism for transporting NOx downwind. However, the chemical fate of RONO2 is poorly understood, and it is possible that they are more effective at transporting NOx than is currently known. We focus our discussion on the temperature-dependent impacts of RONO2 production on PO3, which are 2-fold: (1) RONO2 formation is an association reaction and so individual α (αi) are inversely proportional to temperature, and (2) α of the ambient organic mixture (αnet), defined as a sum over the organic reactivity (eq 6), is a function of temperature’s control over the organic speciation.

T

RO2 + NO2 + M ⇌ RO2 NO2

(R24)

There are two major temperature effects in the role of RO2NO2 as a control over NOx concentrations: (1) higher temperatures favor RO2NO2 decomposition (R24) and (2) temperature-driven emissions of isoprene and other RO2 precursors increase the storage of NOx as RO2NO2. The ROO−NO2 chemical bond is thermally unstable at temperatures common in the troposphere. Simple alkylperoxy nitrates, pernitric acid (HO2NO2) and methylperoxy nitrate (CH3O2NO2), are too weakly bound to accumulate near the surface and are most important at the low temperatures of the upper troposphere.145−147 Acylperoxy nitrates, the most important of which is peroxyacetyl nitrate (PAN),148 do accumulate and are stable enough to redistribute NOx away from large emission sources including cities,149,150 power plants,151 convection and lightning,152 and fires153 to suburban, rural, and remote atmospheres. Observed PAN concentrations often exceed concentrations of NO2,154−156 making their abundance and decomposition rate the strongest control over local NOx in some locations. The production rate of the peroxy radical is temperature dependent.156 Organic precursors of peroxyacetyl radicals include acetaldehyde and first-generation isoprene oxidation products, methyl glyoxal, methylvinyl ketone, and methacrolein. The relative importance of each precursor is dictated by the strength of the local acetaldehyde and isoprene sources,156−162 and the concentrations of these PAN

αnet = (∑ αi RHOHi)/(∑ RHOHi) i

i

(6)

7.1. αi

Laboratory studies have determined the RONO2 yield for many individual organic molecules (αi). The values range from 0.1% or less for CH4 and small oxygenates to as much as 35% for large molecules.163−166 RONO2 production begins with RO2 + NO forming a vibrationally excited adduct, ROONO*, followed by either dissociation back to reactants or rearrangement via a three-member ring transition state, which m ust be c ollisionally stabilized to give RONO2.163,167,168 Measured αi therefore decrease with higher temperatures, increase at higher pressures, and increase with longer carbon chain lengths.164−166,169−175 Experimental examples of the temperature dependence of αi are shown for a few alkanes in Figure 6.164,176,177 The solid lines are computed using an observationally derived parametrization developed by Carter and Atkinson177 based on these and other alkane oxidation experiments. An analogous dependence on temperature was observed for methyl and H

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αisoprene of 4−15%184−186 with more recent experiments narrowing the range to be 7−12%.187−189 If the αnet at low and moderate temperatures is small, largely set by CH4 and CO, then temperature-driven increases in biogenic emissions will increase αnet and decrease PO3. Likewise, in regions with substantial evaporative anthropogenic emissions, such as in areas of oil and gas development, higher temperatures will drive large, low vapor pressure alkanes into the gas phase. Large alkanes have among the highest αi (20−30%), almost certainly increasing αnet at high temperatures. The direction of the temperature dependence of αnet will vary by location. Two opposite effects of temperature on αnet are shown in Figure 7 for locations with distinct organic Figure 6. Temperature dependence of αi for various isomers of the RONO2 products heptyl nitrate (blue) and pentyl nitrate (red) and for 2,3-methylpropyl nitrate (yellow). 2-Pentyl nitrate (red ○) and 3pentyl nitrate (red ◇) are shown separately as are 2-heptyl nitrate (blue ○), 3-heptyl nitrate (blue ◇), and 4-heptyl nitrate (blue △). Markers are experimental data, and curves are the empirical fits using the empirically derived formulation in Carter and Atkinson.177 Dashed lines are extrapolated to temperatures below which laboratory data exist.

propyl nitrate by Scholtens et al.178 and Chow et al.,171 respectively. As shown in Figure 6, over the typical range of summertime temperatures, say 30−40 °C, αi decrease by ∼10% with increasing temperature. Over seasonal temperature differences, 0−30 °C, αi decrease by 35% with increasing temperature. Although no experiments have been published, similar dependence on temperature is expected for organic compounds other than alkanes. During a recent wintertime field experiment, Lee et al.179 investigated ambient temperature effects on αi. Their measurement site was located in Utah’s Uintah Basin, where the organic reactivity was dominated by alkanes due to the local oil and gas operations; therefore, the atmospheric mixture was ideal for comparing ambient RONO2 and RHi observations to the temperature-dependent αi parametrization developed by Carter and Atkinson.177 The authors constrained αnet using measurements of total RONO2 (all molecules with an −ONO2 functionality; ΣRONO2) made by thermal dissociation laser-induced fluorescence180,181 and speciated organic molecules. Lee et al.179 found that the ΣRONO2 observations were well described by αi only when the temperature dependence of αi was included. As a result, RONO2 production was in reality 30% higher than would be predicted using room temperature derived αi. Correspondingly, modeled PO3 was reduced by 20% when the dependence of αi was considered. To date, there has been no targeted field study on the impacts of the temperature dependence of αi on PO3 within the summer O3 season. The temperature dependence of αi is also not yet included in any widely used regional chemical transport models, despite a recommendation to do so made by Carter and Atkinson177 more than two decades ago.

Figure 7. Two different organic reactivity mixtures as a function of temperature and the corresponding αnet. (A) Carbon monoxide (black) and isoprene (green) reactivity in a forested location.63 (B) Light-duty vehicle emissions (black) and small oxygenate (green) reactivity in an agricultural location.46 (C and D) The αnet (%) and change in PO3 (%) corresponding to the above organic reactivity profiles.

reactivity profiles: a site heavily influenced by biogenic reactivity, adapted from the Michigan Forest data of Di Carlo et al.63 (Figure 7a), and a mixed-use urban/agricultural site in the southern San Joaquin Valley near Bakersfield, CA, adapted from the data of Pusede et al.46 (Figure 7b). The bottom panels show calculations of the αnet (left axis) and the percent change in PO3 (right axis). At the forested location, higher temperatures increase αnet and consequently decrease PO3. At the agricultural location, increased temperature is predicted to have the reverse effect, as the small oxygenates with low to zero αi associated with agricultural activities are more important to the organic reactivity at high temperatures. (With the full suite of organic data in Bakersfield, αnet was found to be approximately constant with temperature.46) αnet also affects O3 chemistry by altering the NOx lifetime, particularly under low to moderate NOx conditions where RONO2 production exceeds the formation of HNO3.190 Browne and Cohen191 examined impacts of variability in αnet, demonstrating that at NOx concentrations typical of rural/ remote regions (100 ppt), the lifetime of NOx was 27 h at αnet = 0%, but that this lifetime was dramatically decreased to 8 h at αnet = 5% and to only 5 h at αnet = 10%. Browne and

7.2. αnet

Temperature’s effect on organic reactivity also influences the production of RONO2. For example, temperature-driven isoprene and monoterpene emissions will increase their fractional contribution to αnet. The common monoterpenes, α-pinene, β-pinene, and limonene, have αi of 18%,182 24%, and 23%,183 respectively. Laboratory studies have reported I

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Cohen191 quantified changes to the O3 production efficiency, defined as PO3 normalized by the NOx loss rate, finding that increasing α from 0% to 10% decreased the O3 production efficiency by more than 80%. The range of variability in αnet explored by Browne and Cohen191 is comparable to the temperature-driven changes presented for two locations in Figure 7; it is feasible to expect variability of similar size from αnet−temperature effects. 7.3. αnet and Changes in Emissions

As a portion of the organic reactivity has decreased due to emission controls over the last four decades, αnet has likely also changed. The impact of a change in αnet on PO3 will be a function of the NOx concentration and will be largest, in terms of ppb O3 h−1, at the NOx abundance where PO3 is peak. Additionally, decreases or increases in RONO 2 formation per unit organic reactivity are calculated to have a large effect on PO3. αnet varies as a result of the local organic mixtures and ranges from 1% over the Pacific Ocean, 7% in Mexico City, Mexico, to 15−20% in the alkane-rich atmospheres of oil and gas extraction activities.43,179,192 A change in αnet of 7% to 2%, or vice versa, is a reasonable estimate of what might have occurred over the last three decades in urban areas with aggressive emission controls. Farmer et al.42 evaluated potential impacts of changes of this magnitude with observations collected in Mexico City, Mexico. Reductions in certain classes of organic emissions intended to decrease PO3, particularly large saturated hydrocarbons, could have inadvertently caused the opposite effect, resulting in an increase in PO3. For example, a 30% reduction in organic reactivity intended to reduce PO3 was calculated to have instead increased PO3 by 10% if it also reduced αnet from 7% to 3.5%.42 Returning to the temperature-dependent cases in Figure 7, we consider the impact to PO3 due to the change in αnet brought on by a 75% reduction in temperature-independent organic reactivity. We calculate PO3 with the same model used to generate Figure 1, using present-day organic reactivity before and after the control and assuming that αi and NOx and PHOx are temperature-independent and equal to 5 ppb and 0.5 ppt s−1, respectively, in both scenarios. If CO is reduced by 75% without any change in isoprene (Figure 7a), then after the control, αnet is calculated to be 65% larger at low temperatures (15 °C) and 7% larger at high temperatures (30 °C). The control’s effect from the change in αnet alone would be to decrease PO3 by 8% at low temperatures and by 4% at high temperatures. For the mixed-use/agricultural emission scenario (Figure 7b), if light-duty vehicle emissions were reduced by 75% but small oxygenates were unchanged, then after the control, αnet is calculated to be 40% smaller at low temperatures (15 °C) and 50% smaller at high temperatures (40 °C). Thus, this organic reactivity reduction would have decreased αnet and increased PO3 by 7%. Overall, this example highlights the complex role of temperature on the formation of RONO2 and subsequent impact on PO3.

Figure 8. PO3 (contours have units ppb h−1) mapped as a function of NOx (ppb) and daily maximum temperature (°C). PO3 was calculated with the same model as Figure 1, but constrained with the functional form of organic reactivity, PHOx, NOx concentration, and αnet versus temperature as observed in Bakersfield, CA. Adapted with permission from ref 46. Copyright 2014 Copernicus Publications.

PO3. The measurements were collected in Bakersfield, CA, a location where both total organic reactivity and PHOx were temperature-dependent, increasing from 2 to 12 s−1 and 0.2 ppt to 0.9 ppt s−1, respectively, with an increase in daily maximum temperature of 15 to 40 °C. At the same time, NOx and αnet were largely temperature-independent. The computed PO3 was strongly temperature-dependent at NOx concentrations greater than 2 ppb, but was mostly independent of temperature at lower NOx concentrations. The NOx threshold above which PO3 was temperature-dependent was also observed to vary with temperature. Even though the data set used to generate Figure 8 is specific to the local chemical conditions in Bakersfield, CA, by reducing each chemical variable to its functional form versus temperature, the spatial patterns in the contours of PO3 as a function of NOx and temperature extend qualitatively to other locations, as these relationships between temperature and PO3’s individual chemical terms are typical of many urban areas. PO3 was similarly found to depend on NOx and temperature in the biogenic reactivity-dominated location of Sacramento, CA.44 Comparable conclusions are reached by analyses of temperature’s integrated influence over PO3 using chemical transport models and/or O3 observations. Models allow control over the various individual chemical and meteorological terms that are difficult to distinguish using observations; however, all models lack some if not multiple aspects of PO3’s temperature-dependent chemistry. Modeling experiments typically investigate the total change in O3 due to a change in temperature (∂O3/∂T) or due to a change in any individual temperature-dependent term affecting PO3. This can be evaluated by perturbing emissions or meteorological values48,193−196 or by using projected changes in emissions and meteorology from global models.21,27,29,197−200 Simulations have the benefit of spanning days to decades in time and spatially covering kilometers to the entire globe. A complete account of modeling research on the O3−temperature relationship is beyond the scope of this Review, and interested readers are directed to the excellent reviews by Jacob and Winner,18 Weaver et al.,19 and Fiore et al.20 In this Review, we focus on the role of chemistry and summarize a subset of analyses that follow distinct chemical gradients.

8. CHANGES IN O3 WITH TEMPERATURE 8.1. ∂O3/∂T

In Figure 8, urban PO3 is drawn as a function of NOx and temperature.44−46 To create Figure 8, Pusede et al.46 reduced observations of most of the chemical variables discussed throughout sections 2−7 to their temperature-dependent forms, using these functions to constrain a calculation of J

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One of the first and most influential assessments of the role of temperature on O3 was by Sillman and Samson,49 which investigated observed and simulated O3 concentrations in both rural and urban locations. The authors tested three temperature-dependent variables individually, organic reactivity (isoprene), PHOx (water), and RO2NO2 decomposition, and compared the O3 computed under each scenario against a temperature-dependent simulation combining all three factors. At a rural location in Michigan, the full temperature dependence of modeled O3 could be approximated to within almost 90% by varying only RO2NO2 decomposition. However, in the city of Detroit, MI, the same scenario exhibited the greatest O3 bias, and the combined temperaturedependent simulation was best approximated when isoprene plus water were temperature-dependent. Sillman and Samson49 concluded that under the NOx-limited conditions at the rural site, RO2NO2 decomposition drove the temperature dependence of PO3, while under the VOC-limited conditions in Detroit, organic reactivity and PHOx exerted more influence over the temperature dependence of PO3. Measurements indicated that O3 was 40% more responsive to temperature in Detroit than in the rural locale. Steiner et al.193 tested ∂O3/∂T along gradients of NOxlimited to VOC-limited PO3. The authors used a regional chemical transport model to quantify the individual effects of temperature-dependent biogenic organic reactivity, absolute humidity, and remaining chemistry (including RO2NO2 decomposition) throughout central California. For each sensitivity test, ∂O3/∂T was calculated as the change in O3 concentration between present day and future due to a prescribed temperature change consistent with a 2 × CO2 scenario (corresponding to a 1−4 °C temperature increase). The authors found the relative impacts to ∂O3/∂T in the three experiments varied as a function of local PO3 chemical sensitivity. In the most VOC-limited region in the model domain, the San Francisco Bay Area, temperature-driven biogenic emissions increased ∂O3/∂T more than temperaturedependent chemistry including RO2NO2 decomposition. In the smaller city of Fresno, temperature-driven chemistry that included RO2NO2 decomposition enhanced ∂O3/∂T more than temperature-driven biogenic emissions. (Changes in biogenic reactivity in both locations were comparable.) For each individual driver and in the combined simulation, larger ∂O3/∂T were observed when PO3 was VOC-limited (in the Bay Area) than when PO3 was more NOx-limited (Fresno). Ito et al.105 tested ∂O3/∂T against variations in the fraction of NOx released (recycled) by the oxidation of isoprenederived RONO2 (Figure 9). The authors simulated three scenarios: (a) 0% RONO2 recycled to NOx, that is, all RONO2 were a permanent NOx sink; (b) 40% of RONO2 recycled to NOx; and (c) 100% of RONO2 recycled to NOx. Broadly, ∂O3/∂T was greater near urban areas and when 100% of isoprene-derived RONO2 were a temporary sink for NOx. Again, the simulations suggested that PO3 was more temperature-dependent under higher NOx conditions, reinforcing the two prior studies49,193 and consistent with our discussion of PO3 in sections 2−7, which argued that the response of O3 to temperature depends on the chemical regime. The dependence of ∂O3/∂T on O3 chemical sensitivity implies that modeled surface ∂O3/∂T not designed to capture steep urban photochemical gradients should be interpreted with care.105,118,201−203 Simulations with fine spatial resolution

Figure 9. Calculated change in July surface O3 (ppb) after a 5 °C increase in temperature, where organic reactivity and RO2NO2 decomposition were allowed to vary with temperature for 0% NOx recycling from isoprene-derived RONO2 (a), 40% NOx recycling from isoprene-derived RONO2 (b), and 100% NOx recycling from isoprene-derived RONO2 (c). Reprinted with permission from ref 105. Copyright 2009 John Wiley and Sons.

can represent distinct chemical regimes; however, coarsely resolved models artificially dilute urban emissions into the model grid and, as a result, often capture only the lower NOx range of local chemistry.118 Large-scale meteorological drivers, for example, trends in global circulation21 and air stagnation,204 are thus a larger relative contribution to modeled changes in O3; this may be especially true when simulations are averaged regionally, such that urban pixels (already diluted) represent a small fraction of the domain. For large urban areas early in the past decade (2001), coarsely resolved models diluted NOx such that PO3 was over predicted (chemistry was higher NOx than at peak PO3).201 Vertical level spacing also impacts simulated O3−temperature relationK

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Figure 10. (Top panel) Northern San Joaquin Valley (Stockton): 4-year medians of 8-h O3 exceedance probabilities versus NO2* tethering weekdays (circles) and weekends (diamonds) for 1999−2002 (black), 2003−2006 (brown), and 2007−2010 (green). Data are separated into high- (top) and moderate- (bottom) temperature regimes for the measurement sites of Stockton (a), Turlock (b), and Merced (c). (Bottom panel) Southern San Joaquin Valley (Bakersfield): data for Shafter (a), Bakersfield (b), and Arvin (c) sites. Uncertainties in the probability of exceedances of the CAAQS (by counting statistics) are typically less than ±0.04 (1σ) for weekdays and ±0.06 (1σ) for weekends. Curves (dashed gray lines) were included for visual aid and generated with an analytical model where only the organic reactivity was tuned and PO3 was then scaled to fit. Adapted with permission from ref 45. Copyright 2012 Copernicus Publications.

importantly, more NOx-limited at high rather than moderate temperatures. In this way, Pusede and Cohen45 recreated the decadal trend in the nonlinear impacts of controls on the total organic reactivity, expanding beyond current techniques that typically calculate the reactivity from the known portion of organic emissions. In Figure 10, adapted from their paper, the different dashed lines correspond to different estimates of organic reactivity. The authors showed that reductions in organic reactivity had been effective at reducing high O3 in and downwind of the northern San Joaquin Valley city of Stockton at all temperatures (Figure 10, top panel), as indicated by the need to invoke different curves to explain the change in frequency of O3 exceedances over time. However, in and around Bakersfield (Figure 10, bottom panel), the impacts of organic emission controls were mixed, with effects apparent at moderate but not high temperatures. The key evidence here is that there is no need for distinct curves with different organic reactivities to explain the trends in the frequency of O3 exceedances over time at the highest temperatures, while there is a need at moderate temperatures. This suggests two distinct sources of organic reactivity, a

ships, with potential biases in O3 chemistry and/or surface temperatures.76,201,203,205 There are some observational data sets that provide constraints on ∂O3/∂T calculated by models and point to conceptual deficiencies in their formulation. For example, in their analysis of the impact of emission controls on high O3 (represented as the frequency of exceedances of the 8-h O3 California ambient air quality standard (CAAQS) of 70 ppb), Pusede and Cohen45 exploited the temperature dependence of organic reactivity and PHOx to reconstruct the nonlinear impacts of organic and NOx emission controls on PO3 in three cities in the San Joaquin Valley of California (shown for two cities, Stockton and Bakersfield, in Figure 10). By combining interannual, weekday, and weekend trends in NOx, the authors showed that temperature was a surrogate for the total organic reactivity and PHOx. Weekday−weekend trends in NOx are useful because on weekends in the U.S., diesel trucks are mostly off the road. This substantially reduces NOx emissions15,127,206−208 but has little impact on organic reactivity.41,44,130,209−212 The applicability of Pusede and Cohen’s45 organic reactivity−temperature proxy is evident, as exceedances were both more frequent and, most L

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pollution and compromising existing air quality control strategies. Observed values of mO3−T have been termed the “climate penalty”,194,213,214 although “climate penalty” has also been used to describe the offset in O3 precursor reductions required to counteract effects from the O3−temperature relationship.29,215,216 Mechanisms controlling mO3−T are not well established; this is in part because mO3−T is useful phenomenologically, and, as discussed above, the temperaturedependent chemical drivers of PO3 are multifold. Observed mO3−T are also influenced by meteorology, as certain meteorological conditions or dynamic patterns associated with high temperatures are also conducive to high O3.18−21,195,203,217−220 A simple conceptual model of the controls over O3 is useful in understanding observed changes in mO3−T. The O3 concentration is a function of chemical terms; these are PO3, the topic of this Review, and O3 chemical loss, which is typically much slower than production (5−10 times). The O3 concentration is also a function of meteorological terms, including advective transport and the mixing of plumes laterally and across the top of the boundary layer. Meteorology has an indirect impact on chemistry, and high PO3 is typically associated with stagnation. Mixing can be thought of as a biexponential function with two time scales: (1) boundary layer dilution with free tropospheric air from above on a time scale of order approximately 6 h and (2) a longer time constant that represents the memory of the previous day’s chemical conditions. These two mixing time scales are also dependent on temperature and affect mO3−T. For example, transport of O3 is sensitive to temperature depending on location,220 and vertical mixing is driven by temperature-dependent, buoyance-driven turbulence. The main effect of mixing on O3 is the change in the concentration of precursors and the memory of the previous day; therefore, the temperature dependence of mixing amplifies the temperature dependence of PO3. Reported values of mO3−T are usually derived from daily maximum summertime O3 observations (1 or 8 h averages), with no consistent definition of summertime. These mO3−T range from 0.5−9 ppb O3 °C−1 and O3 variability about the slope is large, often ±10s of ppb O3.49,83,84,196,217,221−226 In the summer in most locations, the y-intercept of the O3− temperature correlation is the background O3 concentration, which itself is not strongly temperature-dependent. Recent work by Fu et al.220 introduced a new statistic, mΔO3−ΔT, which is the slope of the correlation of O3 and temperature anomalies. In their study using August 1988−2011 rural observations in the Southeastern U.S., correlations (r) improved from 0.55 for mO3−T (mO3−T = 3.9 ppb O3 °C−1) to 0.86 for mΔO3−ΔT (mΔO3−ΔT = 4.6 ppb O3 °C−1). Across the U.S., mO3−T has decreased alongside emission controls on multiyear time scales in populated areas (Figure 11).61,84,196 In Los Angeles, CA, mO3−T was 7.6 ppb O3 °C−1 in 1980−1989, 4.6 ppb O3 °C−1 in 1990−1999, and 3.4 ppb O3 °C−1 in 2000−2005, and in the San Joaquin Valley, CA, mO3−T was 3.3 ppb O3 °C−1 in 1980−1989, 2.6 ppb O3 °C−1 in 1990−1999, and 2.4 ppb O3 °C−1 in 2000−2005. Rasmussen et al.196 used a model to show that in NOxlimited regions, NOx control alone would reduce mO3−T, whereas in VOC-limited regions, reductions of both organic reactivity and NOx were required. The authors also found that in these two California air basins, modeled ∂O3/∂T had decreased similarly to observed mO3−T. In the Northeast U.S.,

temperature-independent source that had decreased over the past decade and an uncontrolled temperature-dependent source that dominated at high temperatures. The latter source has not been included in photochemical models, and thus the observed O3 response to emission controls in the region could not have been predicted. 8.2. mO3−T

Complex, temperature-dependent PO3 manifests as an approximately linear correlation of O3 with summertime temperature in almost all near-surface, continental atmospheres. This correlation is typically represented by its slope (mO3−T) and was first documented in the 1970s. A handful of published mO3−T have been reported along chemical gradients (Figure 11), and these data suggest that variability in PO3 can

Figure 11. A selection of mO3−T (ppb O3 °C−1) along chemical gradients. Observations are circles and model predictions are squares. mO3−T are slopes of the correlations of the daily maximum hourly O3 versus the daily maximum temperature unless otherwise noted. The pre- versus postregulation mO3−T are from Los Angeles (black) and the San Joaquin Valley (brown) for the 1980s (black filled), 1990s (gray filled), and 2000s (open) taken from Steiner et al.61 (June− October) and from the Northeast U.S. (purple) before (black filled) and after 2002 (open) taken from Bloomer et al.84 (May− September). The rural (blue, open) and urban (blue, black filled) mO3−T are summertime (April−September) averages in 1988 when daily maximum temperatures are greater than 27 °C from Sillman and Sampson.49 The cities are Phoenix, AZ, Detroit, MI, Atlanta, GA, and New York, NY (left to right). The weekend (open) and weekday (black filled) mO3−T are from Bakersfield (magenta), average of Edison and California Avenue stations, and Los Angeles (red), average of Upland and Pico Rivera stations, for the years 2000−2003 (May−October). The partial (open) to complete (black filled) RONO2 NOx recycling (green) are derived from modeled O3 concentrations (not daily maxima) by Ito et al.105

explain a considerable portion of the variability in mO3−T on seasonal to decadal time scales, at least in locations that have been examined in detail. Generally, trends in mO3−T suggest that (1) chemistry, not meteorology, dominates mO3−T in cities over the range of temperatures observed in a typical O3 season, and (2) in the future in the U.S., chemistry-driven urban mO3−T will reach a minimum due to sizable NOx reductions that are an imminent consequence of new regulations. Historically, empirical O3−temperature relationships were a tool to predict days with high O3. In recent years, there has been renewed attention on O3−temperature linkages in an attempt to quantify the likelihood that future climate-changedriven temperature increases will increase O3, worsening air M

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mO3−T were 3.2 and 2.2 ppb O3 °C−1 before and after a regional regulation targeting NOx emissions from power plants went into effect (2002), respectively. This control resulted in a 43% decrease in NOx emissions without large changes in organic reactivity.84 Bloomer et al.84 observed similar decreases (∼30%) in mO3−T in the Great Lakes and Mid-Atlantic U.S. at the same time. In U.S. cities, forthcoming NOx reductions will continue to transition urban PO3 to NOx-limited chemistry, essentially transforming cities such that their atmospheric chemistry matches rural chemical regimes. Sillman and Sampson49 compared mO3−T in rural and urban environments using summertime O3 data from the year 1988 and found mO3−T were on average greater under urban versus rural emissions scenarios. The average rural mO3−T was 2.4 ppb O3 °C−1, with three rural locations exhibiting mO3−T < 1 ppb O3 °C−1, as compared to the average urban mO3−T of 5.4 ppb O3 °C−1 from the cities of Phoenix, AZ (1.4 ppb O3 °C−1), a desert city with few sources of biogenic/agricultural organic reactivity, Detroit, MI (4.4 ppb O3 °C−1), Atlanta, GA (7.1 ppb O3 °C−1), and New York, NY (8.8 ppb O3 °C−1). These results are generally consistent with the generic model of (1) the nonlinear response of PO3 to organic reactivity and NOx (Figure 1) and of (2) increased PO3 sensitivity to temperature from low to high-NOx conditions, combined with temperature-dependent organic reactivity and HOx precursors. The contribution of PO3 to mO3−T appears when averaging mO3−T separately by day of week in the cities of Bakersfield and Los Angeles, CA. In Bakersfield, conditions leading to high O3 have been NOx-limited on weekends since the early 2000s (Figure 10).45,46 Thus, large decreases in NOx weekday to weekend (∼35%) reduce PO3 and lower mO3−T. By contrast, it is well documented that PO3 in Los Angeles is VOC-limited.211 In Figure 11, mO3−T in Los Angeles is shown to increase with decreasing NOx, as chemistry is right of, but reasonably near, peak PO3. As a final example, in their modeling study constrained to 2001 observations over a forested site in California, Ito et al.105 modeled mO3−T while varying the fraction of isoprene-derived RONO2 recycled back to NOx upon OH oxidation. The authors computed mO3−T was 2.0 ppb O3 °C−1 when 100% of RONO2-bound NO2 was recycled and 1.4 ppb O3 °C−1 if only 40% of RONO2 molecules were recycled to NO2. Taken together, chemistrydriven mO3−T are lowest when absolute rates of PO3 are low and when PO3 is least sensitive to temperature. From the published data, it is not possible to attribute the importance of the two effects separately, but it is apparent that variability in PO3 influences variability in mO3−T along the gradients in Figure 11. These trends in mO3−T raise questions about future chemistry-driven correlations between O3 and temperature in the U.S. and other countries that are successful in reducing O3 precursor emissions. As NOx decreases, mechanistic arguments also suggest the role of the temperature-dependent formation of organic nitrates that suppress mO3−T will increase, making the specific mix of organic reactivity and its αnet more influential; however, no direct observational evidence for this idea has been reported.

experiment, where one variable is altered holding all others constant. Instead, controls typically target emissions of both organic mass and NOx concurrently. Efforts to describe the chemical sensitivity of PO3 and the impacts of past controls on high O3 are motivated in part to accurately infer the response of local PO3 to future emissions regulations. Because of climate change-driven warming, understanding the temperature dependence of PO3 is of particular importance to prescribing effective policies. To predict future changes to PO3, several model227 and observation-based84,218,228 studies have disaggregated past O3 interannual trends by concentration distribution. Generally, this approach shows the largest O3 decreases have occurred in the top percentiles. Smaller differences and some positive trends are reported in the lowest percentiles, depending on which seasons are included in the analysis,214,228 geographic location and regional patterns in O3 transport,218,229 and whether the study area is urban.84 The reported highpercentile O3 decreases are consistent with high-temperature PO3 being more NOx-limited, and thus more responsive to the changes in NOx emissions over the past decade, if highpercentile data correlate with high-temperature days. For example, Bloomer et al.84 found that average O3 in the 95th percentile of observations fell by almost 30% in the urbanized Northeast U.S. after a 43% NOx reduction in 2002, while the fifth percentile of the distribution decreased by only 10%. Cooper et al.218 described similar trends in the eastern U.S., but found that western U.S. sites had fewer statistically significant decreases in O3 concentrations at the top percentiles. The authors attributed this observation to simultaneous changes in O3 transported to the surface from Asia and to greater exchange with the stratosphere at the high altitudes of the intermountain west. In each case, policy recommendations were derived in accordance with past trends in the top percentiles. LaFranchi et al.44 combined measurements and a model to tie the observed change in O3, after a 30% decrease in NOx from 2001−2008, to computed PO3 in the Sacramento, CA urban plume. Outflow from Sacramento is characterized by NOx emissions at the city center, which travel downwind across a band of isoprene emitting oak trees, ultimately reaching an evergreen region with high monoterpene emissions. The authors achieved good agreement between modeled PO3 and an observational analysis of trends in the change in O3 concentration (ΔO3) between monitoring stations (rather than absolute O3 abundances). Early in the record, PO3 was strongly coupled to temperature, implying the chemistry was VOC-limited. During this time, exceedances of the 1-h CAAQS of 90 ppb were most frequent but exhibited the largest decreases with time at high temperatures. The authors interpreted this to be the result of enhanced temperature-driven biogenic emissions, which caused PO3 chemistry to be more NOx-limited and hence more sensitive to the NOx controls that were taking place. LaFranchi et al.44 predicted that an additional 30% reduction in NOx (relative to 2007 levels) would essentially eliminate exceedances of 1-h CAAQS in and downwind of Sacramento. Sustained NOx reductions over the last seven years have occurred in the region, with current levels ∼30% below those in 2007. Monitors in the Sacramento area registered four or fewer exceedances of the 1-h standard in 2014, an almost 5-fold decrease in the frequency of high-O3 days since 2007.

9. REGULATING HIGH O3 AS A FUNCTION OF TEMPERATURE Quantifying the consequences of decadal-scale changes in O3 precursors for mechanisms of PO3 is challenging, as regulatory strategies are not implemented in the form of a scientific N

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In California’s southern San Joaquin Valley, Pusede et al.46 calculated PO3 specifically as a function of temperature, constraining an analytical model with the complete suite of observations from a 2010 field experiment. This model was then used to test the effects of four emission control strategies: a 50% reduction in NOx, a 50% reduction in temperature-independent organic reactivity, a 50% reduction in total known organic reactivity, and a 50% reduction in total reactivity (as observed from OH reactivity measurements). The outcome of the two most likely regulatory scenarios is shown versus temperature in Figure 12. On weekdays,

do not target these sources45 and routine monitoring networks do not observe them. The detailed observations used in the analysis of Pusede et al.46 were collected in the late spring and early summer (May 18−June 29). During this 6-week intensive, the average daily maximum temperature was 30 °C. The O3 season in the San Joaquin Valley spans May−October, and O3 is most severe in the hottest months of July−September. In Figure 12, it is apparent that policies to regulate O3 precursors based only on the conditions sampled during the May−June measurement experiment would have overestimated the impact of organic reactivity controls and incorrectly quantified the effects of NOx reductions on the entire O3 season. By constraining the temperature dependence of organic reactivity and PHOx, Pusede et al.46 used temperature as a coordinate for connecting the results of a short-duration field experiment to the complete O3 season.

10. CONCLUSIONS AND OPEN QUESTIONS Temperature is a fundamental variable affecting multiple factors that influence tropospheric PO3. In this Review, we describe the impact of temperature on the individual chemical drivers of PO3, organic reactivity, PHOx, NOx, RO2NO2, and α, and on the observables ∂O3/∂T and mO3−T. Future changes in O3 precursor emissions and warmer global temperatures due to greenhouse gas-driven climate change will alter PO3 in ways that are predictable but complex. In the U.S. and locations with similar O3 control policies, substantial NOx emission reductions currently underway will transform PO3 in cities such that urban chemistry looks much more like the chemistry of rural atmospheres. Appearing in the literature since the 1970s, the correlation between O3 and temperature is nearly axiomatic; however, the temperature-dependent chemistry of PO3 warrants renewed attention, as climate change has the potential to offset air quality improvements achieved from decades-long regulations on O3 precursor emissions. Additionally, accurately describing the temperature dependence of atmospheric oxidation mechanisms offers an informative test of our understanding of PO3 and HOx under a range of NOx conditions. We suggest the following broad questions to guide future work: Modeling experiments targeting the five chemical factors described here would be valuable to quantify the combined effects on PO3. These simulations could also be conducted with different explicit chemical mechanisms, which likely have different treatments of the temperature dependence of RONO2 formation, influencing the overall PO3 response. These simulations should be designed along distinct chemical gradients, as the temperature dependence of all five chemical factors is variable as a function of O3 precursor abundance and composition. What can be learned about the temperature dependence of PO3 through such a suite of experiments? O3 observations now span multiple decades and exist in locations with diverse emission sources. These long records are of sufficient length to tease apart temperature-driven chemical and meteorological effects. What can specifically be learned from large O3 data sets about trends in emissions, trends in oxidizing chemical mechanisms, and differences between locations with distinct emission source profiles? How can evaluations of temperature-dependent chemistry inform us on HOx chemical mechanisms? What is the full temperature dependence of the production, transport, and fate of the organic nitrates, RO2NO2 and

Figure 12. Predicted percent decrease in PO3 on weekends (dashed lines) and weekdays (solid lines) in the southern San Joaquin Valley in response to a 50% decrease in NOx (top panel) and to a 50% decrease in the temperature-independent component of the organic reactivity (bottom panel). The histogram along the top axis is the number of days in exceedance of the California 8-h O3 standard of 70 ppb in the southern San Joaquin Valley region (Kern County) also as a function of temperature. Adapted with permission from ref 46. Copyright 2014 Copernicus Publications.

reductions in NOx are shown to increase PO3 at temperatures below 29 °C, but to decrease PO3 on days with higher temperatures. On weekends, NOx reductions are effective at all summertime temperatures. Temperature-independent organic emission reductions would reduce PO3 at low temperatures, but not at high temperatures. By comparing these results with exceedances in 2010 of the California 8-h standard of 70 ppb in the region (Kern County), given along the top axis, it is clear that NOx controls are most beneficial when high O3 days are frequent. Even additional controls on the temperature-dependent portion of known reactivity are predicted to minimally decrease high-temperature PO3, with a 50% reduction in total organic reactivity, including reactivity from unidentified molecules, decreasing high-temperature (36.4 °C) PO3 by only 25% and 6% on weekdays and weekends, respectively. That said, organic reactivity reductions of unidentified emission sources necessitate the development of new analytical and policy approaches, as existing controls O

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RONO2? How have these organic nitrates and their impacts on PO3 changed over time? Regulations on NOx emissions have dramatically reduced NO x abundances across the U.S. Regions where PO 3 chemistry is NOx-limited are ideal for investigating how relationships between temperature and meteorology impact mO3−T. How can knowledge of chemical effects of PO3 be applied to disentangle chemical and meteorological influences on O3 in a changing climate? Most of the work discussed in this Review focused on the U.S. and, largely, on the state of California. This is in part because of the unique position Los Angeles holds in the history of O3 pollution research, the quality of the O3 observational records maintained by the California Air Resources Board, and continued frequent air quality exceedances in the state. Air pollution is a serious and growing issue in cities around the world affecting the health of hundreds of millions of people. Locations with high-NOx PO3 will likely see O3 air quality exacerbated under changing climate. Can the observational and modeling methodologies described in this Review be extended to explain O3 chemistry in cities where less is known a priori and where only a small subset of species are measured at the surface and from space?

Allison L. Steiner holds a B.S. in Chemical Engineering (1994) from Johns Hopkins University and Ph.D. in Atmospheric Science (2003) from Georgia Institute of Technology. She worked as a research fellow at the University of California Berkeley before joining the faculty at the University of Michigan in the Department of Atmospheric Oceanic and Space Sciences in 2006. Steiner is the recipient of an NSF CAREER grant award and the U.M. Henry Russel Award and was co-Chair of the 2014 Gordon Conference on Biogenic Hydrocarbons in the Atmosphere. Her research investigates the role that physical and biochemical processes in terrestrial vegetation play in understanding climate and air quality.

AUTHOR INFORMATION Corresponding Author

*E-mail: [email protected]. Present Address ∥

NASA Langley Research Center, Hampton, Virginia 23681, United States. Notes

The authors declare no competing financial interest. Biographies Ronald C. Cohen earned a B.A. with High Honors from Wesleyan University (1985) and a Ph.D. in Chemistry from the University of California Berkeley (1991). As a postdoctoral fellow and research associate at Harvard University (1991−1996) with Professor James G. Anderson, he studied the photochemistry of the stratosphere. He joined the U.C. Berkeley faculty in 1995. He is Professor and Vice Chair of Chemistry, Professor of Earth and Planetary Science, and Director of the Berkeley Atmospheric Science Center. He is a faculty scientist in the Energy and Environment Division at the Lawrence Berkeley National Laboratory. Cohen was elected a Fellow of the American Geophysical Union (2012) and received the Champion of Science award in the Educator category from the Chabot Space and Science Center (2013). Cohen’s research combines satellite, in situ, and laboratory observations aimed at understanding the role chemical reactions have in Earth’s climate and the chemistry producing unhealthy levels of ozone and fine particles.

Sally E. Pusede received her Ph.D. in Chemistry from the University of California Berkeley (2014). While at U.C. Berkeley, she studied the impacts of emission controls on ozone and aerosol chemistry using observations made at the surface and from onboard aircraft. She is currently a NASA Postdoctoral Fellow at NASA Langley Research Center where she researches urban greenhouse gas emissions.

ACKNOWLEDGMENTS Work on this Review at the University of California Berkeley was funded by NASA grant NNX10AR36G and by NOAA grant NA13OAR4310067. Work on this Review at the University of Michigan was funded by NSF grant 1242203. P

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This Review reflects perspectives developed over the years with support from NASA, NSF, NOAA, U.S. EPA, and the California Air Resources Board.

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DOI: 10.1021/cr5006815 Chem. Rev. XXXX, XXX, XXX−XXX