Ground-level ozone (O 3 ) is an oxidant that inflames airways and damages tissue in the respiratory tract, leading to increased coughing, wheezing, shortness of breath, and other asthmatic symptoms (US EPA, 2020b). Maximum daily average 8 h (MDA8) O 3 concentrations designed to protect public health are codified in the National Ambient Air Quality Standards (NAAQS) (US EPA, 2021) and the California Ambient Air Quality Standards (CAAQS) (California Air Resources Board, 2007). Seven of the 10 cities across the United States with the highest O 3 concentrations are located in California (American Lung Association, 2020), making O 3 pollution a continued public health threat for millions of California residents more than four decades after O 3 abatement efforts began.
O 3 levels are often described by the maximum daily average 8 h concentration. The annual fourth-highest MDA8 O 3 concentration averaged over 3 years has special regulatory significance. This design value determines whether the region containing the monitor complies with the O 3 NAAQS. O 3 design values in California decreased steadily between the years 1980 and 2019 (Fig. 1) due to the success of emissions control programs that reduced concentrations of precursors broadly divided into two groups: oxides of nitrogen (NO x ) and volatile organic compounds (VOCs) (Parrish et al., 2016;Simon et al., 2015). Continued progress after the year 2010 has been slower, and O 3 design values even increased in some air basins between the years 2015-2018 (Fig. 1). Multiple factors have been proposed to explain the lack of further reductions in O 3 concentrations in recent years. These potential factors include (i) growing importance of precursor VOC emissions not previously accounted for in the planning process as major sources such as transportation have been controlled (McDonald et al., 2018;Shah et al., 2020), (ii) an imbalance in the historical degree of NO x and VOC reductions (Cox et al., 2013;Parrish et al., 2016;Pollack et al., 2013a;Steiner et al., 2006), or (iii) more frequent heat waves (Jacob and Winner, 2009;Jing et al., 2017;Pusede et al., 2015;Rasmussen et al., 2013;Weaver et al., 2009) and wildfires (Jaffe et al., 2013;Lindaas et al., 2017;Lu et al., 2016;Singh et al., 2012) as a consequence of climate change. All these theories are supported to varying degrees by indirect measurements or model predictions, but there is an absence of strong direct evidence that identifies dominant factors contributing to the increased O 3 concentrations. The uncertainty that lingers over the recent O 3 trends suggests that fresh approaches are needed to directly verify the optimum emissions control path.
O 3 formation has been studied for decades in California, using both measurements and model simulations (Kroll et al., 2020). These past studies provide important background information about the effects of precursor NO x and VOC species and help build the foundation for new studies. Statistical analyses of long-term surface measurements have de- termined that lower NO x concentrations are associated with higher O 3 concentrations on weekends (Pollack et al., 2012;Pusede and Cohen, 2012) and higher temperatures are associated with increased VOC emissions and chemical reaction rates, leading to higher O 3 concentrations during warm stagnation events (Lafranchi et al., 2011;Nussbaumer and Cohen, 2020). These long-term studies suggest that VOCs are the limiting precursor for O 3 formation in the center of large cities, while NO x is the limiting precursor in downwind areas (Lafranchi et al., 2011;Pusede and Cohen, 2012). Neither long-term analysis method clearly explains the recent trend of increasing O 3 concentrations in Los Angeles.
O 3 sensitivity has also been analyzed over shorter timescales using ratios of photochemical indicator species including H 2 O 2 /HNO 3 and HCHO/NO 2 (Sillman, 1995;Tonnesen and Dennis, 2000). Satellite retrievals of HCHO/NO 2 from the Global Ozone Monitoring Experiment (GOME), the SCanning Imaging Absorption spectroMeter for Atmospheric CHartograpHY (SCIAMACHY), the Ozone Monitoring Instrument (OMI), and the TROPOspheric Monitoring Instrument (TROPOMI) have extended these O 3 sensitivity calculations over broad geographical regions (Chossière et al., 2021;Duncan et al., 2010;Jin et al., 2017;Martin et al., 2004;Schroeder et al., 2017a). The short-term measurements generally support the findings from the long-term studies but once again fail to identify the dominant factor(s) driving the recent increase in O 3 design values. Reactive chemical transport models (CTMs) have been used extensively to predict the effectiveness of candidate emissions control programs (Brown, 2018;California Air Resources Board, 2018;Meng et al., 1997;Sillman, 1999), and so one might expect that they would provide the most detailed explanation for recent O 3 trends. Models are necessarily incomplete approximations to highly complex real-world systems, and so they are often incapable of predicting subtle features in pollutant trends. No model calculation has been able to reproduce the observed increase in O 3 design values (Parrish et al., 2017). It is unclear whether this failure stems from a lack of accurate emissions trends, an incomplete description of atmospheric chemistry, or an incomplete representation of the effects of shifting climate on O 3 formation mechanisms.
Recent advances in measurement techniques provide new tools to study O 3 sensitivity directly. Mobile smog chambers bridge the gap between laboratory studies and the real atmosphere. Past studies have designed mobile smog chambers to measure the aging of secondary pollutants (i.e., O 3 , SOA) from certain emission sources (Howard et al., 2008(Howard et al., , 2010b;;Li et al., 2019;Platt et al., 2013;Presto et al., 2011). It is difficult to evaluate sensitivity of secondary pollutants formed from multiple sources using a single smog chamber. Recently, a mobile dual-smog-chamber system has been used to directly measure the SOA formation in ambient air (Jorga et al., 2020;Kaltsonoudis et al., 2019). The design used in the current study consists of three chambers that can simultaneously measure the non-linear response of O 3 formation to NO x and VOC perturbations. The automated valve and sampling incorporated into this design also allows long-term remote field measurements to evaluate the seasonal trends in O 3 sensitivity. At the same time, the satellite TROPOMI launched by the European Space Agency (ESA) in October 2017 provides measurements of HCHO and NO 2 tropospheric vertical column densities (TVCDs) with 3.5 km × 5.5 km spatial resolution that can start to resolve O 3 perturbations around major sources such as wildfires (Ialongo et al., 2020;Veefkind et al., 2012;Vigouroux et al., 2020b). The purpose of this study is to combine these two new measurement techniques into a detailed analysis of O 3 sensitivity to precursor NO x and VOC emissions spanning an entire spring-summer-fall cycle in California. Daily measurements from smog chamber perturbation experiments are analyzed for short-term trends (day of week) and long-term trends (seasonal variation) to reveal the effects of traffic, natural vegetation, and wildfires. The direct O 3 sensitivity measurements are then combined with TROPOMI HCHO/NO 2 ratios to extend our understanding of the O 3 sensitivity across the entire state of California.
Three identical transportable smog chambers were used to directly measure base case O 3 concentration and O 3 sensitivity to precursor NO x and VOC. Each chamber was constructed from fluorinated ethylene propylene (FEP) with a volume of 1 m 3 housed in an enclosure measuring 2.13 m H × 1.22 m L × 1.22 m W . UV lamp panels were placed on the floor and the roof of the chamber support frame. Each panel can hold up to six UV lamps (Sylvania, F40BL 40W T12) that emit at wavelengths between 280-400 nm. The lamp panels were configured to produce 50 W m -2 to replicate the mid-day photochemistry in California during the summer. The enclosure walls were constructed from polished aluminum with total reflectivity of ∼ 95 %. Figure S1 in the Supplement represents the crosssectional view of the transportable smog chamber system.
One cubic meter of ambient air was injected into the FEP chambers at the start of an experiment using a Teflon diaphragm pump (model DOA-V751-FB, Gas Manufacturing, Benton Harbor, MI, USA) operating at a flow rate of 10 L min -1 for each chamber. Solenoid valves were configured to inject perturbation gases NO x (8 ppb NO 2 ) and VOC surrogates (4.4 ppb ethylene, 2.8 ppb n-hexane, and 0.8 ppb m-xylene) respectively into chambers no. 1 and no. 3 for comparison to base case chamber no. 2. Perturbation gases were added halfway through the chamber filling operation so that they would be thoroughly mixed with the ambient air during the remainder of the chamber filling process. The composition of the VOC perturbation was based on the VOC mixture used to determine ozone formation potential (Carter et al., 1995). The magnitude of the perturbations was selected to be as small as possible while still generating an observable change in monitored ozone concentrations. A single set of monitors sequentially measured concentrations within each chamber to increase the precision of the inter-chamber comparisons. The current experiment includes measurements of NO x (model nCLD-855-Yh, Eco Physics, Dürnten, Switzerland), NO y (sum of all oxidized atmospheric odd-nitrogen species) (model 42i, Thermo Fisher, Franklin, MA, USA), O 3 (model 205, 2B technology, Boulder, CO, USA), and temperature-relative humidity sensor (model RH-USB, Omega Engineering, Norwalk, CT, USA). The total sample flow rate for all monitors was approximately 3 L min -1 . Seven measurements with a duration of 10 min were made from each chamber, resulting in a total sample volume of 210 L air, or approximately 21 % of the chamber volume (leaving 79 % of the total air in the chamber). The shape of the chambers was not greatly distorted at any point during the experiment. Chambers were drained at the conclusion of an experiment using a rotary vane vacuum pump (model 0523-101Q-G588DX, Gast Manufacturing, Benton Harbor, MI, USA). All chamber operations were controlled automatically using a program written in LabVIEW that interfaced with a customized set of data acquisition devices and solenoid valves (DAQ-SV).
The consistency of the O 3 formation rates across chambers was tested in a controlled laboratory environment prior to deployment in the field. All three 1 m 3 FEP chambers were filled with laboratory air and were perturbed by an equal mixture of both NO x and VOC prior to 180 min of UV exposure.
Several blank tests were performed by adding zero air (AI 0.0Z-K, Praxair) into all three chambers as part of the consistency tests to further develop confidence in the chamber measurements. Figure S2 in the Supplement illustrates the agreement between chamber no. 1 and no. 3 O 3 measurements vs. chamber no. 2 O 3 measurements during a typical QA/QC check. Uncertainty between chamber measurements is ∼ 1 % across a wide range of concentrations. The loss rate of O 3 to chamber walls was determined in the dark for all three 1 m 3 FEP chambers filled with identical NO x -VOC mixtures. Average loss rates of 5 % h -1 were calculated over the 3 h experiments. Loss rates were identical for all chambers in the system, and so this issue will not influence the comparisons between chambers in the current study.
To confirm that chamber measurements represent the behavior observed in the atmosphere, weekly-averaged O 3 concentrations in the base case chamber were compared to weekly-averaged ambient O 3 concentrations measured at the nearby monitoring station (marked in Fig. S5 in the Supplement) from April to December 2020 in Fig. S3 in the Supplement. The O 3 concentrations in the base case chamber at the start of each experiment were similar to the ambient O 3 concentrations, indicating that the gas-phase chemical composition related to O 3 formation was not changed while injecting ambient air into the chamber. The O 3 formation in the chamber generally reflects the O 3 chemical production from the in situ ambient air around 10:00-12:00 local time, while the ambient O 3 is influenced by chemical production, mixing, and deposition (Cazorla et al., 2012). As expected, the initial rate of O 3 formation in the chamber is therefore higher than the initial rate of change in the ambient O 3 concentrations. The current experiment is focused on measuring the response of this chemical production rate to changes in precursor NO x and VOC concentrations because this most closely approximates the local effects of potential emissions control programs.
VOC measurements are useful to help interpret O 3 formation trends and to identify the chemical regime on the NO x -VOC isopleth for O 3 (Seinfeld and Pandis, 2016). Ground-level daily VOC measurements from Photochemical Assessment Monitoring Stations (PAMS) are only available for a limited number of summer months, and so alternative indicator species were investigated. Baker (2008) found that non-methane hydrocarbon (NMHC) concentrations were correlated with CO concentrations in 28 US cities during the years 1999-2005. This may reflect situations where dominant sources that emit CO also emit large amounts of NMHC, or it may reflect situations where relatively constant sources of CO and NMHC are correlated because they are diluted by the same amount of atmospheric mixing. The success of emissions control programs targeting anthropogenic VOCs has increased the relative importance of residual biogenic VOCs in many urban atmospheres across the United States (US EPA, 2020a). Biogenic sources do not emit CO, but biogenic VOCs can react in the atmo-sphere to produce CO (Hudman et al., 2008). CO also acts as an indicator of atmospheric mixing that equally affects all primary sources. In an effort to improve the ability of CO to represent biogenic VOCs in the current study, an additional metric was calculated by multiplying the measured CO concentrations by the enhancement factor for isoprene emissions induced by temperature and relative humidity (Guenther et al., 1991). Figure S4 in the Supplement shows the correlation between measured VOC reactivity (VOCR) and CO × biogenic at Sacramento during summer months between 2010-2019. VOCR was calculated from PAMS measurements of VOC concentrations multiplied by their reaction rate constant with OH (Chen et al., 2010;Kleinman, 2005;Steiner et al., 2008). VOCR and CO × biogenic are reasonably well correlated (r = 0.6, p<0.001), while VOCR and CO were less correlated (r = 0.39). This analysis supports the preference for CO × biogenic as an approximate surrogate for VOCR in the current study, with the understanding that real-time measurements of VOCR would be highly preferred in future studies.
Tropospheric HCHO and NO 2 retrievals (level 2; unit: mol m -2 ) over California were obtained from the TROPOMI for February-October 2020. The TROPOMI is aboard the Sentinel-5 Precursor (Sentinel 5-P) satellite, which was launched by the ESA in October 2017. The polar-orbiting satellite enables quantitative information on trace gases to be retrieved approximately at 13:30 local sun time (ascending node) each day on a global scale. The retrieval algorithms for TROPOMI NO 2 data use the measurements of the earth's radiance in the visible absorption wavelengths (405-465 nm) made by the hyperspectral imaging spectrometer. The algorithms first derive the total slant column density of NO 2 using a differential optical absorption spectroscopy (DOAS) method. The total slant column NO 2 is then separated into stratospheric and tropospheric slant column densities of NO 2 while utilizing information from a data assimilation system. Finally, the tropospheric vertical column density of NO 2 is obtained by applying conversion factors, called air mass factors (AMFs), to the tropospheric slant column density of NO 2 . The retrievals of TROPOMI HCHO data apply a similar DOAS method to the ultraviolet (UV) wavelengths (328.5-359 nm) of the solar spectrum. Further details about the TROPOMI data are provided by Veefkind et al. (2012), Van Geffen et al. (2020), and De Smedt et al. (2018).
The spatial resolution of TROPOMI NO 2 and HCHO TVCDs is 3.5 km × 5.5 km, which is finer than that of the predecessor OMI (13 km × 24 km). Quality assurance (QA) values were obtained alongside the HCHO and NO 2 data, and only measurements with QA values > 0.50 were retained to ensure good data quality and sufficient data points when computing monthly averages (Van Geffen et al., 2021).
Correction factors were not applied to TROPOMI data in the current study. Verhoelst et al. (2021) and Vigouroux et al. (2020a) analyzed the accuracy of the TROPOMI data using ground-based measurement sites across the globe. Measurements were not made in California, but several of the evaluation sites had attributes similar to locations in California. Bias in daily TROPOMI NO 2 retrievals varied between -15 % and -56 % in moderately polluted areas with NO 2 column measurements between 3 × 10 15 -14 × 10 15 molec cm -2 (typical for moderate-sized cities in California). The bias in TROPOMI HCHO measurements ranged between +26 % ± 5 % at low HCHO levels and -30.8 % ± 1.4 % at high HCHO levels. HCHO levels measured in Sacramento (∼ 0.6 × 10 15 molec cm -2 ) had a bias of approximately zero. These results suggest that TROPOMI measurements over California almost certainly contain some amount of bias that could only be removed through a comparison to measurements from a ground-based network. Application of global-average bias correction factors would not change the trends in HCHO and NO 2 in time and space even if they would change the absolute magnitude of those values. The current analysis will therefore focus on trends in the TROPOMI measurements.
O 3 sensitivities to precursor NO x /VOC concentrations were measured in central Sacramento, CA (38.57 • N, 121.49 • W), from April-December 2020 (222 experiment days out of a total of 251 d). Sources in the vicinity of the site include commercial office buildings, restaurants, two major highways, freight and passenger rail lines, a shipping port, and suburban residences (see map in Fig. S5). Grab samples of ambient air were collected between 10:00 and 12:00 local time to characterize the daytime O 3 formation rates in the presence of variable atmospheric mixing and regional emissions. Sensitivities were based on perturbation concentrations of approximately 8 ppb of NO x injected into chamber no. 1 and 8 ppb of VOC surrogates injected into chamber no. 3. Initial gas concentrations were measured from the full chambers in the dark over a 30 min period (10 min for each chamber). The UV lamp panels were then illuminated for 180 min, and the chamber concentrations were measured in a continuous cycle of 10 min intervals over a total of seven cycles. Each active monitoring period lasted 210 min (= 30 min of dark measurements +180 min of light measurements). Measurements in different chambers are made at different times, making it difficult to compare chamber results at the conclusion of the experiment. It was noted that O 3 concentrations within each chamber averaged in each 10 min sampling interval increased linearly over the 180 min period when the UV lights were on. A linear regression model was therefore applied to extrapolate O 3 concentrations in each chamber to the end of the measurement period to facilitate direct comparisons between the base case chamber no. 2 and perturbed chambers no. 1 and no. 3. The difference of O 3 concentration after 3 h UV exposure was calculated between chamber no. 1 and chamber no. 2 ( O +NO x
) and chamber no. 3 to chamber no. 2 O +VOC 3 to quantify the O 3 sensitivity. An example of a typical day of O 3 results analysis is shown in SI.
A chamber model developed by Howard et al (2008Howard et al ( , 2010a, b) , b) was employed as a part of this analysis to quantify the sensitivity of the O 3 response to NO x perturbations under different experimental configurations. The chemical reaction system used by the chamber model is based on the SAPRC11 chemical mechanism (Carter and Heo, 2013) with wall loss rates based on the measured value of 5 % h -1 . The time integration procedures used to solve the set of differential equations that predict concentrations as a function of time are taken from the full University of California Davis and California Institute of Technology (UCD/CIT) chemical transport model (Venecek et al., 2018;Ying et al., 2007). Day-specific values of NO, NO 2 , and O 3 initial concentrations used in the chamber simulations are based on measurements near the study location. VOC initial concentrations used in the chamber simulations are based on UCD/CIT simulations over the study location. The seasonal profile of the simulated VOC concentrations matches the CO × biogenic trends illustrated in Fig. 2, but the amplitude of the simulated seasonal trend was damped. VOC initial concentrations used in the chamber simulations were therefore scaled to match the amplitude of the CO × biogenic factor. Section 4.1 presents a sensitivity study on the chamber measurement result using the chamber model described here. Figure 2a shows the monthly-averaged TROPOMI satellite NO 2 measurements and the boxplot of daily chamber NO 2 measurements at Sacramento between February-December 2020. NO 2 concentrations remained relatively stable between April and July and then sharply increased in August-September possibly due to increased wildfires in the late summer months. Enrichment of NO 2 and other pollutants in wildfire plumes has been noted in previous re- search (Jaffe and Wigder, 2012). The open boxes in Fig. 2a represent days within the months of August-November that were not influenced by wildfire smoke (Anderson and Kuwayama, 2022), leading to reduced NO 2 concentrations. The upward trend in NO 2 concentrations in October-December 2020 is likely associated with decreased boundary layer heights and increased fuel consumption for heating during the colder fall-winter season. This seasonal association can also be viewed in the decreasing TROPOMI NO 2 levels measured during the warmer spring season (February-April, 2020). TROPOMI column measurements will not directly depend on boundary layer height, but increased boundary layer heights are usually associated with higher average boundary layer wind speeds, leading to downwind advection and dispersion of pollutants. The effects of reduced transportation emissions in March-April 2020 caused by COVID-19 shelter-in-place orders are notably minor in the ambient NO 2 measurements. Although light-duty vehicle traffic decreased by as much as 50 % during this time period, heavyduty truck traffic was more constant (Liu et al., 2020;Parker et al., 2020). The ground-based measurement site is 0.8 and 1.8 km from two major freeways, but NO x concentrations at this site do not appear to be strongly influenced by the COVID-19 reduction in light-duty traffic activity. Increasing NO x emissions from residences and relatively quick recovery of the heavy-duty traffic compared to the light-duty traffic may also minimize COVID-19 effects on NO x concentrations (Liu et al., 2021). The seasonal pattern of NO x concentrations driven by wildfires, reduced boundary layer height, and increased residential fuel consumption appears to dominate at the urban Sacramento location.
Figure 2b shows the monthly-averaged TROPOMI satellite HCHO levels and the daily ground-based CO × biogenic concentration at the Sacramento sampling site. The agreement between the seasonal trend in the CO × biogenic and TROPOMI HCHO builds confidence in the use of CO × biogenic as a ground-based indicator of VOC concentrations at this location. Both indicators suggest that VOC concentrations increased from April-August 2020 and sharply declined in October 2020. Wildfires can emit large amounts of VOCs that can be transported to urban areas (Zhang et al., 2018). It is possible that wildfires contributed to the highest VOC concentrations observed between August and September 2020. Removing the days influenced by wildfires (open box) still leaves a strong seasonal trend with increasing VOC concentrations between April-August 2020, which is consistent with increasing VOC emissions from biogenic sources. Biogenic VOC (BVOC) emissions increase during warmer spring months and continue to increase as temperatures rise into summer (Guenther et al., 2006(Guenther et al., , 1991)). The CO × biogenic factor inherently incorporates this effect, but the strong agreement between the TROPOMI HCHO levels and the CO × biogenic metric in Fig. 2b suggests that the seasonal pattern of the biogenic emissions is a real feature of the dataset and not an artifact of how the CO × biogenic metric was constructed. Similarly, the declining VOC concentration observed in October 2020 and beyond matched the expected decrease in biogenic emissions during the colder fall and winter seasons when vegetation becomes dormant. The seasonal pattern illustrated in Fig. 2b suggests that BVOC is an important precursor of HCHO in Sacramento.
PAMS measurements of ground-level isoprene concentrations in Sacramento are shown as blue diamonds in Fig. 2b. Isoprene is highly reactive in the atmosphere, and so PAMSmeasured concentrations are lower than 4 ppb. The limited time period of available measurements makes it difficult to discern seasonal trends, but the slightly lower measured isoprene concentrations in July and slightly higher isoprene concentrations in August followed by decreasing (non-wildfire) isoprene concentrations in September generally match the VOC trends generated using both TROPOMI HCHO and CO × biogenic. Once again, the agreement between the three independent techniques builds confidence in the overall assessment of VOC seasonal trends.
Volatile chemical products (VCPs) are another important category of VOC emissions (McDonald et al., 2018). The expanded usage of spray disinfectant and sanitization products during the COVID-19 pandemic might have been a significant source of VOCs in the urban area, but the expected usage pattern of these products does not include a sharp decline in the fall period. The seasonal pattern of VOC concentrations increasing during spring-summer and decreasing during fall-winter is more consistent with a combination of biogenic sources and wildfires, as discussed above. 3a) did not reconcile the two measurements. The divergence of the ground-based measurements and satellite measurements in this month may reflect the presence of elevated plumes of wildfire smoke above the monitoring site that were detected by the satellite measurements (Jin et al., 2017). Cleaner air at the groundbased monitors, therefore, yielded O +NO x 3 values in a different chemical regime than the satellite measurements that are based on the tropospheric vertical column densities. This comparison suggests that ground-based measurements may be required to supplement satellite-based measurements to fully characterize the surface O 3 formation regime under special circumstances that generate concentrated pollution layers above the ground-level.
Removing the days influenced by wildfires from the chamber measurement (open box) and TROPOMI satellite measurement (open diamond) in Fig. 3a reduces both O +NO x 3 and TROPOMI HCHO/NO 2 . Figure S7 in the Supplement compares TROPOMI HCHO and NO 2 on wildfire days and non-wildfire days. Median TROPOMI HCHO measurements increased by 44 % and TROPOMI NO 2 measurements increased by 14 % on wildfire days. The comparison between wildfire vs. non-wildfire days implies that wildfires emit more VOC than NO x , which is in agreement with previous studies (Jaffe and Wigder, 2012). It is also notable that the decrease in O +NO x 3 is larger than the decrease in TROPOMI HCHO/NO 2 . This observation might once again reflect the fact that the wildfire identification algorithm (Anderson and Kuwayama, 2022) was based on ground-level measurements that do not flag all of the days with elevated plumes above the monitoring site that could differentially affect the satellite measurements.
Figure 3b shows the monthly variation of ground-based O +VOC 3 and TROPOMI satellite HCHO/NO 2 from February-December 2020 at the Sacramento sampling site.
O +VOC 3 (Fig. 3b) has an inverse time trend compared to O +NO x 3 and TROPOMI HCHO/NO 2 (Fig. 2a). The O +VOC 3 trend is well correlated to the TROPOMI HCHO/NO 2 trend plotted on a reversed axis between April-August 2020, but the two trends diverge in September-October 2020 when wildfires were prevalent. Removing the wildfire days from August to October (open box) increased the ground-based O +VOC 3 , once again suggesting that wildfires contributed more VOCs than NO x to the atmosphere (Altshuler et al., 2020). The divergence between the ground-based O +VOC 3 measurements and TROPOMI satellite HCHO/NO 2 measurements during the wildfire season once again reflects the presence of elevated plumes that were measured by the satellite but not by the ground-based monitors (Schroeder et al., 2017a). (Anderson and Kuwayama, 2022) to focus on the day-of-week patterns. Hypothesis tests were carried out to determine if weekday and weekend responses were similar in each month. The results indicate that weekend reductions in NO 2 concentrations were significant at a 90 % confidence level (or higher) before July. The similarity between weekday and weekend NO 2 concentrations after July may be associated with increased NO x emissions from wildfires in the late summer and space heating in the fall-winter since neither of these sources follows a weekday/weekend pattern. Although days directly affected by the wildfire smoke were removed from the analysis, residual emissions from smoldering fires and multi-day recirculation of air mass that have been affected by wildfire smoke may have contributed to elevated regional NO x concentrations through the formation of reactive nitrogen reservoir species such as peroxyacetyl nitrate (PAN) that can be transported over long distances (Lindaas et al., 2017). The CO × biogenic VOC surrogate did not display statistically significant differences between weekdays vs. weekends except in June and July. Extremely hot days (>35 • C) occurred on weekdays in June and weekends in July, driving the CO × biogenic factor higher.
Reduced NO x emissions on weekends are reflected in the O 3 sensitivity to precursors shown in Fig. 4c andd. The median O +NO x 3 sensitivity is higher on weekends for most months, indicating that the atmosphere was more NO xlimited. Large variability in the data makes the weekend vs. weekday O +NO x 3 response statistically significant at the 90 % (or higher) level only in April, September, and October. The large weekend reductions in median NO 2 concentrations detected in May and June did not lead to significantly higher weekend O +NO x 3 , possibly because of higher weekday median VOC concentrations in these months. Median O 3 sensitivity was NO x -limited ( O +NO x 3 >0) on both weekdays and weekends from June to August when BVOC emissions are expected to be highest. In spring and early fall (April, May, and September), the median weekday O 3 sensitivity is VOC-limited but the median weekend O 3 sensitivity is NO xlimited. In late fall and winter (October-November), the median O 3 sensitivity is VOC-limited on both weekends and weekdays. Weekend NO x reductions have an inverse effect on O +VOC
. The median O +VOC 3 is lower on weekends than weekdays because the O 3 formation is more NO x -limited on weekends. The mixture of daily data points (yellow to red points) shows the O 3 isopleth pattern where higher O 3 concentration (darker color) exists at higher NO x and VOC concentrations. The combination of the colors and the arrows illustrated in the isopleth diagram helps to define the measured ridgeline in the O 3 isopleth diagram that denotes the transition between VOC-limited chemistry and NO x -limited chemistry at Sacramento. Arrows in the upper left of the diagram point downwards (VOC-limited) towards the ridgeline, while arrows in the lower right of the diagram point upwards (NO x -limited) towards the ridgeline. The atmospheric system experiences a range of conditions throughout the 9month study period that moved the measurements around the O 3 isopleth diagram. The average seasonal cycle is illustrated in Fig. 5 using monthly-average points shown as blue circles with white month numbers. The monthly-average O 3 chemical regime traces an oval path through the isopleth diagram as NO x concentrations decrease and CO × biogenic (proxy of VOC) concentrations increase moving from spring to summer months. NO x concentrations increase rapidly in fall while CO × biogenic concentrations simultaneously decrease at the Sacramento sampling location, transitioning the O 3 chemistry to VOC-limited conditions. The pattern is expected to reverse for the months of January-March (not shown) to produce a repeatable annual cycle. The direct measurement of the seasonal pattern of the O 3 chemical regime clearly illustrates the effects of NO x and VOC emissions controls at different times of the year.
The days with the highest measured O 3 concentrations are of particular interest in the current study since emissions control programs are traditionally tailored to reduce the O 3 design value, which is determined by MDA8 O 3 concentration. Fig- Additional statistical analysis was carried out to characterize the extreme values in the O 3 sensitivity plots (Coles, 2001;Gilleland and Katz, 2016). Extreme value analysis characterizes high concentrations using return levels corre-sponding to a specified time period (T ). In the context of the current analysis, the return level is the O 3 perturbation response that is expected to be exceeded once during the specified time period. The probability of exceeding the return level is therefore 1/T . Figure 7
The consistency between the NO x and VOC measurements made using ground-based chambers and satellite observations enables a joint analysis to directly calculate the TROPOMI HCHO/NO 2 ratio at the transition between NO x and VOC-limited O 3 formation regimes. Three circular buffers (2.5, 5, and 7.5 km radii) centered on the monitoring location were used to generate the TROPOMI HCHO/NO 2 ratio that was then compared to the measured O +NO x 3 ratio at the monitoring site. The HCHO/NO 2 ratio generated using the 5 km radius buffer shows the best correlation with ground-based chamber results shown in Fig. 8a (results from other buffers are shown in Fig. S8 in the Supplement). Linear regression analysis between 1-week-averaged O +NO x 3 and HCHO/NO 2 with and without wildfires shows that removing the wildfires always improves the correlation coefficient (R), likely because the elevated wildfire plumes have different effects on surface vs. integrated column measurements. The regression carried out using a 5 km buffer radius with wildfires removed yielded a correlation coefficient R = 0.62 (p<0.001). The transition point between NO x -limited and VOC-limited conditions (corresponding to O +NO x 3 = 0) occurs when HCHO/NO 2 = 4.6 (95 % confidence interval: 4.39-5.90). When the TROPOMI satellite HCHO/NO 2 ratio fell below 4.6, then the ground-based measurement of O +NO x 3 was usually negative, and when the satellite HCHO/NO 2 ratio rose above 4.6, then the groundbased measurement of O +NO x 3 was usually positive. Ordinary lease squares (OLS) regression was used to estimate the transition point HCHO/NO 2 = 4.6 between chemical regimes. This approach does not account for uncertainty in chamber O +NO x
. Repeating the analysis using reduced major axis (RMA) regression that accounts for errors in both x and y yields an estimated transition point HCHO/NO 2 = 4.4 between chemical regimes (Fig. S9 in the Supplement).
The HCHO/NO 2 transition point directly measured in the current study is consistent with previous estimates constructed from the combination of satellite measurements and routine ground-based O 3 monitoring data (Jin et al., 2020). Other previous efforts to estimate HCHO/NO 2 value at the transition point between NO x -limited and VOC-limited regimes typically couple satellite HCHO/NO 2 measurements with O 3 sensitivity or O 3 sensitivity indicators (i.e., LNO x /LRO x ) predicted using reactive chemical transport models. These hybrid studies predict HCHO/NO 2 transition points lower than the value of 4.6 derived in the current study. Martin (2004) used HCHO/NO 2 from GOME to calculate the regime transition value HCHO/NO 2 = 1.0 for polluted areas across the globe. Duncan et al. (2010) used OMI to estimate the regime transition value HCHO/NO 2 = 1-2 across the continental United States. Schroeder (2017b) found the transition range could between HCHO/NO 2 = 1.3-5.0 during DISCOVER-AQ in Houston. These estimated HCHO/NO 2 transition values vary due to the different satel-
Figure 9 displays the monthly-average spatial distribution of TROPOMI HCHO/NO 2 ratios across California for the time period April-October 2020. Overall, TROPOMI HCHO/NO 2 was the lowest (mean (standard deviation) = 3.5 (1.2)) in April and the highest (mean (standard deviation) = 9.7 (3.2)) in July. The seasonal pattern of increasing NO x limitation during the summer months at Sacramento (Fig. 3a,b) is mirrored across most of California (Fig. 9), especially in the mountainous areas with dense vegetation. The majority of California is in the VOC-limited regime in April and May due to the low BVOC emissions. Only very remote regions with low NO x concentrations are still in the NO x -limited regime during these spring months. Most areas outside of major urban centers transition toward NO x -limited conditions between June and September as ambient temperature and BVOC emissions increase. These areas then transition back to the VOC-limited regime in the fall months beginning in October as temperatures decrease and vegetation becomes dormant.
Large urban centers including Los Angeles, San Diego, and the San Francisco Bay Area exhibit low HCHO/NO 2 ratios (VOC-limited conditions) throughout the study period. These urban areas contain less vegetation and larger numbers of NO x sources than outlying suburban and rural areas. Therefore, reducing NO x emissions in these urban centers may increase monthly-average O 3 concentrations throughout the year. The HCHO/NO 2 ratio in California's Central Valley is lower than the HCHO/NO 2 ratio in the surrounding mountainous area during all months of the study period. This spatial pattern reflects the high BVOC emissions from coniferous forests in the mountainous regions compared to the cropland in the Central Valley (Misztal et al., 2014).
Past studies have found that wildfire smoke plumes mixing with high urban NO x emissions can lead to enhanced urban O 3 concentrations (Jaffe and Wigder, 2012). The effects of wildfires during August-September 2020 can be observed in Fig. 9 as zones of reduced HCHO/NO 2 immedi-Figure 9. Spatial distribution of TROPOMI satellite (HCHO/NO 2 ) ratios in California for April-October 2020. TROPOMI NO 2 and HCHO data are re-gridded to 5 km resolution when calculating monthly-average ratios. The black bold line circles the burned area in each month detected by MODIS from Fire Information for Resource Management System (FIRMS). The NO x -limited conditions correspond to HCHO/NO 2 ratios above 4.6.
ately around the active burn areas followed by a larger halo zone of increased HCHO/NO 2 as the VOCs emitted from wildfires have time to react to form HCHO. This halo pattern is most obvious in October 2020, when the seasonal cycle of biogenic emissions declined sufficiently to shift the O 3 sensitivity back to the VOC-limited regime for the majority of the state except for the region surrounding a wildfire in the Sierra Nevada mountain range east of Fresno (near Yosemite National Park). VOCs emitted from the wildfire in October 2020 reacted to produce HCHO in the halo region, keeping the HCHO/NO 2 ratio in the NO x -limited regime. The extensive wildfires that occurred in 2020 appear to have extended the natural peak of the HCHO/NO 2 ratio from July into August, September, and even October 2020. It is unknown whether this satellite observation accurately represents conditions at ground level. The results at the Sacramento monitoring site in September 2020 (Fig. 3a andb) suggest that elevated smoke plumes can dominate the satellite observations, but they may not accurately represent conditions at ground level.
The seasonal variation of O 3 sensitivity can be observed over the entire state of California using the TROPOMI HCHO/NO 2 (Table S1). Figure 10a shows how the O 3 sensitivity seasonal pattern differs among different air basins. The air basins with the highest populations have suppressed seasonal variation of O 3 sensitivity because of the higher anthropogenic NO x emissions. The SoCAB had the lowest HCHO/NO 2 ratio among all the air basins in California during the study period. This is noteworthy since the So-CAB has the highest population and the highest O 3 concentrations. The San Francisco Bay Area and San Diego County, two other heavily populated areas in California, also have relatively low HCHO/NO 2 ratios compared to other air basins. Using HCHO/NO 2 = 4.6 as the transition point, even these highly urbanized air basins appear to transition from VOClimited to NO x -limited O 3 formation chemistry in summer 2020. It is noteworthy, however, that the urban cores of these regions remain VOC-limited across all months due to very high NO x emissions (see persistently green regions in Fig. 9). Figure 10b illustrates the TROPOMI HCHO/NO 2 monthly variation for different cities in SoCAB between February and October 2020. The cities inside/around the LA urban core have HCHO/NO 2 <4.6 throughout the entire year with a weak seasonal variation. This might be caused by reduced BVOC emissions in the urban center. The remote areas (darker colors in Fig. 10b) have greater seasonal variation and higher peak HCHO/NO 2 . The sharp increase in HCHO/NO 2 in summer leads to a shift in O 3 sensitivity from the NO x -saturated regime to the NO x -limited regime in the cities further away from the urban core. Due to the different seasonal variation of HCHO/NO 2 at different sites, the NO x -saturated region around the urban core will shrink in the summer and expand in the winter. Figure S10 in the Supplement shows this seasonal pattern of O 3 sensitivity regime distribution in Los Angeles as an example. Thus, the optimal emissions control strategy for the entire air basin may differ from the optimal emissions control strategy for urban cores areas.
The chamber measurements made in the current study capture the sensitivity of the O 3 chemical production term in response to the concentration of NO x and VOC. The experiment does not directly account for atmospheric processes such as mixing and deposition, but the chemical production term represents the dominant processes that determine how local emissions affect local O 3 concentrations. The good agreement between ground-based chamber measurements and satellite O 3 sensitivity measurements in the current study builds confidence in the reported seasonal trend of O 3 sensitivity. Limitations and uncertainties associated with the temperature, UV intensity, and NO x /VOC perturbations used in the ground-based chamber measurements are discussed in this section to build further confidence in the results. The potential of these issues to influence the results is analyzed through a combination of measurements and model calculations. The configuration of the chamber model is described in Sect. 2.4. Figure 11 shows that the chamber model can accurately predict the measured seasonal trends in O 3 sensitivity, providing a solid foundation for sensitivity tests.
The temperature in the reaction chambers was higher than the ambient temperature due to the heating effects of the UV lights. Figure S11 in the Supplement shows that the difference between the chamber gas temperature and the ambient temperature increased by 5-10 • C over the course of each experiment, with the exact temperature profile depending on the measurement month. Despite this temperature increase, all three chambers experience the same temperature profile, and so the comparison of O 3 formation between the chambers is not strongly biased by this issue. Figure 12a shows the calculated O +NO x 3 during each month of the experiment under the chamber and ambient temperature profiles. The difference between the chamber and ambient temperature has little effect on the O 3 sensitivity in each month. Temperature effects do not significantly modify the seasonal variation of the measured O 3 sensitivity in the current study. Similar behavior was shown in O +VOC 3 in Fig. S12a.
The UV intensity in the chambers was intentionally maintained at a constant level through all seasons so that the changes in O 3 sensitivity could be directly attributed to the changes in the ambient concentrations. A representative average UV intensity was selected for this purpose. As was the case with temperature, all chambers experience the same UV conditions, and so this factor is not expected to overly bias the comparison between chambers that acts as the core of the current study. The actual seasonal cycle of UV radiation would generate higher photolysis rates in the summer and lower photolysis rates in the winter that would further amplify the seasonal signal already detected by the measurements with constant UV intensity. SAPRC11 chamber model simulations were used to quantify the effect of seasonal variations in UV intensity. Simulations were carried out using the measured constant UV radiation in the chambers and using the clear-sky UV intensity calculated with the routines in the UCD/CIT CTM based on the latitude-longitude of the measurement site and the day of year. The calculations summarized in Fig. 12b show that the difference associated with the use of constant UV radiation does not change the seasonal pattern of O 3 sensitivity to NO x perturbations. The seasonal changes to UV intensity slightly amplify the magnitude of the seasonal trend in O 3 sensitivity (increase the absolute value of O +NO x
), but the overall seasonal pattern is unchanged. Similar behavior was shown in O +VOC 3 in Fig. S12b in the Supplement.
The constant 8 ppb NO 2 perturbations used in the current study are greater than or equal to ambient NO x concentrations during the summer season at Sacramento. O 3 forma- The size of the NO x and VOC perturbation used in the chamber experiments is most important when ambient conditions are close to the ridgeline on the O 3 isopleth diagram (spring and fall in the current experiment). An 8 ppb NO 2 perturbation may jump over the ridgeline in this case, suggesting that the chemistry is NO x -rich rather than NO x -limited. SAPRC11 chamber model simulations were used to quantify the effect of the 8 ppb NO 2 perturbation vs. a smaller 2 ppb NO 2 perturbation. As shown in Fig. 12c, the 8 ppb NO 2 perturbations used in the current study do not affect the shape of the seasonal trend in O 3 sensitivity measurement, but the 8 ppb NO 2 perturbation (open box in Fig. 12c) does affect the transition months when the atmospheric system changes to NO x -limited behavior. This issue may influence the estimated value of HCHO/NO 2 that characterizes the transition to NO x -limited behavior in Sect. 3.2, but it should be noted that the value of 4.6 derived in the current study is in good agreement with the value of 4.5 reported by Jin et al. (2020). The non-linearities associated with VOC chemistry are less severe, and so the size of the VOC perturbation does not complicate the interpretation of results (see O +VOC The O 3 sensitivity calculated with the combined effects of temperature, UV intensity, and lower perturbation size was compared to the base case calculated O 3 sensitivity in Figs. 12d and S12d (in the Supplement). The NO 2 perturbation size has the largest effect on the chamber O 3 sensitivity results, with relatively minor changes introduced by temperature and UV intensity. None of these issues changes the basic pattern of increasing NO x limitations during summer months transitioning to VOC limitations during winter months. It should be noted that operation of the mobile smog chamber system in cities with higher ambient NO x concentrations is expected to give O 3 sensitivity results that are even less dependent on the NO 2 perturbation size.
California's current O 3 control strategies mainly focus on NO x emissions from motor vehicles (Barcikowski et al., 2017) and especially heavy-duty trucks (Burke, 2020). Additional control strategies would require cleaner engines and zero/near-zero emission technologies (Brown, 2018;South Coast AQMD, 2021). VOC sources that dominate O 3 formation are still not clear due to the large numbers of activities that release VOCs and the complex reactions that VOCs undergo in the atmosphere. Controls on VOC emissions have been more effective than controls on NO x emissions over the past decades, mainly because of reduced emissions from large stationary sources (Barcikowski et al., 2017). The estimated VOC emissions decreased by a factor of 3 while NO x emission decreased by a factor of 1.5 between 1980 and 2010 according to the California inventory (Cox et al., 2013;Rasmussen et al., 2013). Long-term ambient measurements in the SoCAB confirm that ambient VOC concentrations decreased at an average rate of 7.5 % yr -1 , while ambient NO x concentrations decreased at an average rate of 2.6 % yr -1 between the years 1980 and 2010 (Pollack et al., 2013b;Warneke et al., 2012). These ambient measurements suggest emissions reduction by a factor of 10 for VOCs and a factor of 2.2 for NO x . Recent studies have shown that VOCs from consumer products are underestimated in the emission inventory (McDonald et al., 2018). However, the clear seasonal pattern in the measured O 3 sensitivity and the corre-sponding pattern for concentrations of VOC proxies (HCHO and CO × biogenic) suggest that BVOCs are also important. The 2016 California State Implementation Plan calls for a 34 % reduction in NO x emissions and a 30 % reduction in VOC emissions (California Air Resources Board, 2018), which will increase the VOC/NO x ratio. This will reduce peak O 3 concentrations in most areas across California that become NO x -limited in the middle of the summer. In contrast, the NO x emissions control program could cause a short-term increase in peak O 3 concentrations in the urban cores that are currently VOC-limited, and it could increase intermediate O 3 concentrations in late spring or early fall as regions transition back to VOC-limited conditions. These regions do not currently violate the O 3 NAAQS, but they could experience future violations depending on the timing of the transition to lower NO x concentrations. Despite these penalties, controls on NO x emissions may be the only alternative for long-term O 3 reductions in regions where VOC emissions are dominated by biogenic sources. As the NO x keeps decreasing, the O 3 photochemical regime will eventually transition back to NO x -limited conditions, and all further NO x reductions will yield decreasing O 3 concentrations. Previous studies have observed such a transition between VOC-limited and NO x -limited conditions in polluted urban areas with high NO x concentrations. Jin et al. (2020) observed a suppression of the NO x -limited area between 2013-2016 vs. 1996-2000 in Los Angeles by analyzing satellite HCHO/NO 2 ratios. Baidar et al. (2015) observed a weakening of the higher O 3 concentrations on weekends in the SoCAB between 1996 and 2014, reflecting a transition towards more NO x -limited conditions. These studies suggest that continued reductions in NO x emissions will eventually yield a transition to fully NO x -limited conditions in Los Angeles, albeit this transition may not be fully complete for decades.
Wildfires are an unpredictable factor that enhances O 3 formation in California. O 3 formation during wildfire events shifts towards more NO x -limited conditions, making reductions in NO x emissions attractive. The frequency and scale of wildfires in the western United States have increased over time due to the effects of drought and climate change (US-GCRP, 2017). Abatement strategies may focus on wildfire prevention as an effective way to reduce incidental O 3 concentrations.
Direct measurements of O 3 sensitivity to precursor NO x and VOC concentrations using a mobile smog chamber system in Sacramento, CA, from April to December 2020 show that O 3 sensitivity follows a seasonal cycle. O 3 formation is VOC-limited in the spring, is NO x -limited in the summer, and returns to VOC-limited in fall-winter. This seasonal pattern reflects higher emissions of reactive VOCs during the summer season and increased NO x concentra-tions during the other seasons. The most obvious potential source of increased VOC emissions during the summer season is biogenics. Comparing the ground-based chamber measurements to satellite measurements from TROPOMI suggests that the transition between NO x -limited and VOClimited chemical regimes for O 3 formation occurs at a TROPOMI HCHO/NO 2 ratio of 4.6. Monthly-averaged TROPOMI measurements show that O 3 sensitivity across most of California follows a seasonal cycle similar to Sacramento, but locations with higher population density are more VOC-limited. The urban cores of most large cities remain VOC-limited in all seasons even when the surrounding areas become NO x -limited in the middle of summer. The variability of the chemical regime for O 3 formation across space and time makes it difficult to design an emissions control strategy that will equitably reduce O 3 concentrations for all California residents currently living in air basins that violate the 8 h O 3 NAAQS. Reductions in NO x emissions will be the most efficient control strategy to reduce present-day peak O 3 concentrations, but this strategy will lead to increasing O 3 concentrations in urban cores during the middle of summer and increasing O 3 concentrations in surrounding regions during late spring and early fall. These penalties will persist until NO x emissions are reduced sufficiently to push the entire region into NO x -limited conditions sometime in the coming decades. VOC emissions reductions never cause increasing O 3 concentrations, and O 3 formation is VOC-limited during some seasons. It may be advisable to augment NO x emissions control programs with some amount of controls on volatile consumer products (VCPs) and mitigation of wildfires in an attempt to reduce any near-term increases in O 3 concentrations. Continued deep NO x emissions reductions should eventually transition all locations across California into the NO x -limited regime and will effectively push the state toward 8 h O 3 NAAQS attainment.
Data availability. Daily chamber measurement data including O 3 concentration, O 3 sensitivity, and NO x concentration can be accessed through https://datadryad.org/stash/share/ ktJh3AxAs0K7y8Iku8-VL3v7ZuGwBGQodYhRT-wHZ04 (Wu et al., 2022).
Supplement. The supplement related to this article is available online at: https://doi.org/10.5194/acp-22-4929-2022-supplement.
Author contributions. SW made field measurements and wrote the initial draft of each version of the manuscript. HJL processed TROPOMI data. AR analyzed wildfire vs. no-wildfire periods. SL provided project management. TK constructed the initial version of the chambers. JHS hosted initial measurements and helped revise the manuscript. MJK designed the experiment, directed the data analysis, coded the chamber model, and revised the manuscript.
Atmos. Chem. Phys., 22, 4929-4949, 2022 https://doi.org/10.5194/acp-22-4929-2022
https://doi.org/10.5194/acp-22-4929-2022 Atmos. Chem. Phys., 22, 4929-4949, 2022
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