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Summer ozone in the Northern Front Range Metropolitan Area:
Weekend-weekday effects, temperature dependences and the
impact of drought
Andrew A. Abeleira1, Delphine K. Farmer1
1. Department of Chemistry, Colorado State University, Fort Collins, CO, 80523, USA
Correspondence to: Delphine K. Farmer (delphine.farmer@colostate.edu)
Abstract. Contrary to most regions in the U.S., ozone in the Northern Front Range Metropolitan Area (NFRMA) of
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Colorado was either stagnant or increasing between 2000 and 2015, despite substantial reductions in NOx emissions.
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We used available long-term ozone and NOx data in the NFRMA to investigate these trends. Ozone increased from
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weekdays to weekends for a number of sites in the NFRMA with weekend reductions in NO2 at two sites in downtown
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Denver, indicating that the region was in a NOx-saturated ozone production regime. The stagnation and increases in
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ozone in the NFRMA are likely the result of (1) decreasing NOx emissions in a NOx-saturated environment, and (2)
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increased anthropogenic VOC emissions in the NFRMA. Further investigation of the weekday-weekend effect showed
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that the region outside of the most heavily trafficked Denver area was transitioning to peak ozone production towards
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NOx-limited chemistry. This transition implies that continued NOx decreases will result in ozone being less sensitive
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to changes in either anthropogenic or biogenic VOC reactivity in the NFRMA. Biogenic VOCs are unlikely to have
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increased in the NFRMA between 2000 and 2015, but are temperature dependent and likely vary by drought year.
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Ozone in the NFRMA has a temperature dependence, consistent with biogenic VOC contributions to ozone production
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in the region. We show that while ozone increased with temperature in the NFRMA, which is consistent with a NOx-
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saturated regime, this relationship is suppressed in drought years. We attribute this drought year suppression to
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decreased biogenic isoprene emissions due to long-term drought stress.
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1. Introduction
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Tropospheric ozone (O3) is detrimental to human health, impacting asthma attacks, cardiovascular disease, missed
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school days, and premature deaths. Based on these impacts, the Environmental Protection Agency (EPA) projects that
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reducing the O3 standard to the new 70 ppbv 8-hour average will result in health benefits of $6.4-13 billion/yr (EPA,
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2014). O3 also damages plants, reducing agricultural yields (Tai et al., 2014). Using crop yields and ambient O3
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concentrations for 2000, Avnery et al. (2011) estimate the loss of $11-18 billion/yr worldwide as a result of the
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reduction of staple worldwide crops (soybean, maize, and wheat) from O3 damage. During summer months, the
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Northern Front Range Metropolitan Area (NFRMA) of Colorado consistently violated the pre-2016 U.S. EPA
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National Ambient Air Quality Standard (NAAQS) of 75 ppbv fourth-highest daily maximum 8-hour average (MDA8)
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ambient O3 concentration, despite proposed reductions in anthropogenic emissions (CDPHE, 2014). The NFRMA has
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been an O3 non-attainment zone since 2008 (CDPHE, 2009), prompting the Colorado Air Pollution Control Division
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and the Regional Air Quality Council to develop the Colorado Ozone Action Plan in 2008 to target key O3 precursors:
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volatile organic compounds (VOCs) and NOx (NO+NO2)(CDPHE, 2008). Despite these control efforts, 2013 was the
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NFRMA’s fourth year in a row to exceed the federal O3 standard (CDPHE, 2016), and the eight NFRMA non-
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attainment counties, with their combined population >3.5 million, exceeded the MDA8 75 ppbv O3 standard 9-48 days
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between 2010 and 2012 (AMA, 2015). However, Colorado must comply with the new 70 ppbv MDA8 standard by
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2018. In order to accurately design and implement O3 reduction schemes, a thorough understanding of local O3 trends
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and chemistry is required.
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Ground-level or boundary layer O3 depends on local production, transport, and meteorological parameters:
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(1)
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where / represents the time rate of change of O3 concentration, P(O3) is the instantaneous net photochemical
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O3 production rate (production – loss), / represents the entrainment rate (we) of O3 in and deposition
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rate (ud) of O3 out of the mixing layer height (H), and describes the advection of O3 mixing layer height.
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Briefly, ground-level O3 is driven by a catalytic chain that is initiated by RO2 production from VOC oxidation (R1),
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and propagated by local NOx emissions (R2,3).
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RH + OH + O2 RO2 + H2O (R1)
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Chain propagation occurs through reactions between HO2 or RO2 radicals with NO to form NO2 (R2a,b, R3), which
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is photolyzed (R4) and leads to net O3 formation (R5). Reactions between NO and O3 also produces NO2 (R6),
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leading to a null cycle with no net O3 production. Alkoxy (RO) radicals form carbonyl-containing compounds and
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HO2 (R7).
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RO2 + NO RO + NO2 (R2a)
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RO2 + NO RONO2 (R2b)
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HO2 + NO NO2 + OH (R3)
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NO2 + hν NO + O(3P) (R4)
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O(3P) + O2 O3 (R5)
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NO + O3 NO2 + O2 (R6)
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RO + O2 R’CHO + HO2 (R7)
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For every VOC that enters the cycle, approximately two NO2 radicals are produced – but the resulting carbonyl-
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containing compounds and organic nitrates can be repeatedly oxidized or photolyzed, further propagating the P(O3)
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chain. Chain termination occurs through RO2 and HO2 self-reactions to form peroxides (dominant termination
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reactions in the “NOx-limited regime”), OH and NO2 reactions to form HNO3 (“NOx-saturated”or “VOC-limited”
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regime), or RO2 and NOx reactions to form organic nitrates (RONO2) or peroxyacyl nitrates (RC(O)O2NO2).
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Formation of organic and peroxyacyl nitrates suppresses P(O3), but does not shift the cross-over point between NOx-
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limited and NOx-saturated P(O3) regimes (Farmer et al., 2011). This cross-over point of maximum, or peak, O3
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production is controlled by the chain termination reactions, and is sensitive to the HOx production rate and thus VOC
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reactivity. Decreasing NOx is an effective O3 control strategy in a NOx-limited system, but will increase O3 in a NOx-
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saturated system. Controls for NOx-saturated systems often focus on reducing anthropogenic VOC reactivity, and/or
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suppressing NOx emissions sufficiently that the system becomes NOx-limited.
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Trends in O3 for 2000 – 2015 varied across the United States (EPA, 2016a). Using the annual 4th maximum of daily
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8-hour averages (MDA-8), the EPA reported a 17% decrease in the aggregated national average O3. However, regional
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trends deviated substantially from the national average. For example, the EPA reported a 25% decrease in O3
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throughout the southeast, while the northeast shows a 16% decrease. Smaller decreases in O3 occurred in the northern
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Rockies (1%), the southwest (10%) and the west coast (4-10%). These O3 reductions are concurrent with national
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reductions in O3 precursors of 54% for NOx, 21 % for VOCs, and 50% for CO (EPA, 2016b). Due to the non-linear
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behavior of O3 chemistry described above, reductions in O3 precursors do not necessarily result in reductions of
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ambient O3. Cooper et al. (2012) reported that 83%, 66%, and 20% of rural eastern U.S. sites exhibited statistically
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significant decreases in summer O3 at the 95th, 50th, and 5th percentiles (1990-2010). No increases in O3 occurred at
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any sites, indicating that local emission reductions have been effective in those regions. In contrast, O3 in the western
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US followed a very different trend: only 8% of western U.S. sites exhibited decreased O3 at the 50th percentile; the 5th
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percentiles for O3 at 33% of the sites actually increased. These increases were larger for the lower percentiles,
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indicating that while local emissions reductions may have been effective at some sites, increased background O3 offset
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the improvement.
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Lefohn et al. (2010) found that O3 decreased across many U.S. sites at a less rapid pace during 1994-2008 than during
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1980-2008, indicating that O3 improvements had leveled off by the late 2000s. The leveling off could be a result of
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either slowed precursor emissions reductions, which is contrary to the EPA estimates, or, more likely, shifting O3
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chemistry regimes as precursor emissions are changing. Lefohn et al. (2010) reported that the distributions of high
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and low hourly O3 values narrowed toward mid-level values in the 12 cities studied, consistent with a reduction in
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domestic O3 precursors and possibly increased transport of O3 precursors from east Asia. A number of modeling and
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measurement studies have also reported increased baseline O3 in the western U.S. due to the transport of O3 precursors
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from east Asia (Cooper et al., 2010;Parrish et al., 2004;Pfister et al., 2011;Weiss‐Penzias et al., 2006). These studies
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questioned the effectiveness of local precursor emission reductions in controlling local O3 in impacted regions.
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Cooper et al. (2012) showed that the intermountain West is an intriguing environment with potentially increasing
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background O3. The NFRMA is of particular interest due to the challenge in effective O3 regulation, its growing
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population and the dominantly anthropogenic sources of O3 precursors. VOCs have been well-studied in the region,
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with a particular focus on the Boulder Atmospheric Observatory (BAO) in Erie, CO (e.g. Gilman et al., 2013;McDuffie
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et al., 2016;Pétron et al., 2012;Swarthout et al., 2013;Thompson et al., 2014). VOC composition in the NFRMA was
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heavily influenced by oil and natural gas (ONG) sources, as well as traffic. In winter 2011, ~50% of VOC reactivity
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was attributed to ONG-related VOCs and ~10% to traffic (Gilman et al., 2013;Swarthout et al., 2013). Recent studies
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have shown that ONG and traffic contributed up to 66% and 13% of the VOC reactivity respectively at BAO in
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mornings for both spring and summer 2015, but that biogenic isoprene was a large, temperature-dependent component
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of VOC reactivity in the summer, contributing up to 49% of calculated daytime VOC reactivity (Abeleira et al., 2017).
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We note that the anthropogenic VOCs were typically lower in 2015 than previous measurements, pointing to the
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complex roles of meteorology, transport and local emissions. In contrast, observed isoprene in summer 2012 was
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much lower than summer 2015, likely due to shifting drought conditions. While temperatures across the two summers
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were similar, 2012 was a widespread drought year in the region, and 2015 was not; drought is typically associated
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with suppressed biogenic VOC emissions. Local anthropogenic and biogenic sources are not the only VOC sources
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in the region: longer-lived VOCs consistent with transport have also been observed (21-44% of afternoon reactivity
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in 2015), and smoke from both local and long-distance wildfires impacted air quality in the NFRMA in punctuated
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events. This smoke was sometimes, but not always, associated with elevated O3 (Lindas et al., 2017).
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The impact of a changing climate on air quality is poorly understood due to the complex climate-chemistry interactions
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and numerous feedbacks (Jacob and Winner, 2009;Palut and Canziani, 2007). However, increasing temperature is
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expected to increase O3 (Bloomer et al., 2009;Jacob and Winner, 2009;Palut and Canziani, 2007). The O3-temperature
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relationship is attributed to (1) temperature-dependent biogenic VOC emissions that provide a source of VOCs for
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OH oxidation leading to increased HOx cycling (Guenther, 2006;Guenther et al., 1996), (2) thermal decomposition of
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peroxyacetylnitrate (PAN) to HOx and NOx (Fischer et al., 2014;Singh and Hanst, 1981), and (3) increased likelihood
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of favorable meteorological conditions for ozone formation (i.e. high insolation, stagnation, circulating wind patterns)
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(Reddy and Pfister, 2016;Thompson et al., 2001). In addition, increased temperatures and changing soil moisture could
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alter soil emissions of NOx. Due to the non-linearity of P(O3) chemistry as a function of NOx, the increased VOC and
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NOx emissions associated with warming can either increase or decrease P(O3) depending on local NOx concentrations
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(i.e. NOx-limited vs. NOx-saturated). Interactions between climate change and regional-scale meteorology are
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complex, and may also impact O3. High and low O3 in the U.S is coupled to a variety of meteorological parameters
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including planetary boundary layer (PBL) heights (White et al., 2007;Reddy and Pfister, 2016), surface temperatures
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(Bloomer et al., 2009), soil-moisture and regional winds (Davis et al., 2011;Thompson et al., 2001). PBL height is
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coupled to increased temperatures, reduced cloud cover, stronger insolation, and lighter circulating wind patterns with
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higher 500 hPa heights correlating to higher average July O3 in the NFRMA (Reddy and Pfister, 2016).
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In this paper, we used temperature, O3, and NO2 data from 2000-2015 at multiple sites in the NFRMA to investigate
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why O3 has not decreased in the region despite decreases in NOx. We used a weekday-weekend analysis to elucidate
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the NOx regime for P(O3) in Denver, and explored the temperature dependence of O3 and the role of drought in
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influencing that relationship in the NFRMA.
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2. Methods
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2.1 Measurement sites
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We used publicly available O3, NO2 and temperature data (https://aqs.epa.gov/aqsweb/documents/
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data_mart_welcome.html) from eight sites in the NFRMA (Fig. 1, Table 1). The CAMP site is 1 mile east of the I-25
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interstate highway in downtown Denver. O3 data was available for 2005 – 2007 and 2012 – 2015, while NO2 data was
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available for 2001 – 2007 and 2010 – 2015. Welby is roughly 8 miles northeast from the CAMP site, and is adjacent
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to a large lake and less than 1-mile west of the Rocky Mountain Arsenal open space. O3 data was available for 2000 –
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2009 and 2011 – 2015, while NO2 data was available for 2001 – 2002, 2004 – 2005, 2007 – 2008, and 2010 – 2015.
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The Carriage site is <1 mile west of the I-25 interstate at the same latitude as the CAMP site. O3 data was available
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for 2000 – 2012 for the Carriage site. The Fort Collins site is adjacent to Colorado State University near downtown
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Fort Collins. O3 data was available for 2000 – 2015. The Greeley site was located on the southeast side of Greeley and
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<1 mile south of CO state highway 34. O3 data was available for 2002 – 2015. The Rocky Flats site is in a rural area
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adjacent to the Rocky Flats Wildlife Refuge <15 miles south of Boulder. The I-25 site is adjacent to the I-25 interstate
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2-miles south of the Carriage and CAMP sites, and intercepts fresh NOx emissions directly from the I-25 interstate.
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NO2 data was available for 2015, but not O3. The La Casa site is <1 mile west of the I-70 and I-25 interstate junction.
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O3 and NO2 data were available for 2015. Temperature data was available for all sites for all years.
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2.2 Ozone and NO2 data treatment
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Ambient NO2 concentrations were measured by chemiluminescence monitors equipped with molybdenum oxide
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converters. These monitors are used as the EPA Federal Reference Method for monitoring ambient NO2
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concentrations, and have a known interference from nitric acid and organic nitrates (Dunlea et al., 2007). The true
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ambient NO2 mixing ratio is a component of the reported values. NO2* will be used in this manuscript to refer to the
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EPA NO2 measurements, which includes the interference, and can be considered to be a proxy for total reactive
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nitrogen oxides (NOy). While the absolute NO2* concentration will be greater than NO2 but less than NOy, trends in
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NO2* provided insight on trends in local NOx emissions. The O3 and NO2* mixing ratios are filtered to summer months
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(June 1 – August 31), and averaged to a daytime value (10:00 am – 4:00 pm local). A site was excluded for a given
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year when <50% of data is available for that summer.
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2.3 Trend analysis
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Following the analyses of Cooper et al. (2012), the statistical significance of the linear trends were tested with a
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standard F-test with the null hypothesis that there is no linear trend (R2 = 0). The null hypothesis was rejected with a
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confidence level ≥ 95% if the probability (p) associated with the F-statistics was low (p ≤ 0.05).
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3 Results and Discussion
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3.1 Long term trends in O3 and NO2* in the Northern Front Range Metropolitan Area
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Contrary to most other places in the U.S., O3 in the NFRMA was either stagnant or increasing between 2000 and 2015,
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despite substantial decreases in NOx emissions. At most sites in the eastern U.S. and some on the west coast, O3 was
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decreasing at all percentiles. In the NFRMA, however, five out of six monitoring sites exhibited no change or
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increasing O3 at the 50th and 95th percentiles in the 2000 – 2015 period (Fig. 2). The 5th percentile is often taken as
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background O3. With the exception of the Greeley site, the 5th percentile of O3 increased across the NFRMA between
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2000 and 2015. However, only the downtown Denver CAMP site had statistically significant increases in O3 of 2.6 ±
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0.9, 2.3 ± 0.3, and 1.8 ± 0.7 ppbv/yr for the 5th, 50th, and 95th percentiles, respectively. The Welby site had increases
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of 1.5 ± 0.5, 1.3 ± 0.4, and 1.4 ± 0.4 ppbv/yr from 2000 – 2015 (Fig. 2b), but with a statistical significance at only the
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95th percentile. Cooper et al. (2012) reported that the Welby site exhibited no statistically significant increase in O3
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from 1990 – 2010, contrary to what we found for 2000 – 2015 at the 95th percentile.
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The increasing O3 trends in the NFRMA occurred despite reductions in NOx. NO2* at the CAMP site decreased
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significantly from 2000 at a rate of 1.2 ± 0.2 and 1.5 ± 0.2 ppbv/yr for the 50th and 95th percentiles for CAMP (Fig. 3).
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Welby exhibited a non-significant decreasing NO2* trend at the 95th percentile of 0.5 ± 0.3 ppbv/yr (Fig. 3). The
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increased O3 may be due to increased summer temperatures in Colorado, increased regional baseline O3, or increased
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local P(O3) from unknown emission sources (Cooper et al., 2012). VOC emissions steadily increased in Colorado
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from 2000 to 2012 according to the EPA NEI-2014. To the best of our knowledge, the NFRMA does not have any
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long-term VOC datasets, but the EPA NEI-2014 for Colorado provided an estimate for yearly anthropogenic VOC
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(AVOC) emissions (EPA, 2016b). All categories of AVOC emissions decreased slightly from 2000 – 2015, except
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for petroleum related VOCs which increased from 7.4 x 103 tons in 2000 to 2.6 x 105 tons in 2011 with a decrease to
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1.5 x 105 tons in 2015 (Fig. 4). However, we note the NEI is only an estimate and does not include biogenic sources
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of VOCs, which can contribute substantially to VOC reactivity in the region, but vary substantially from year to year
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(Abeleira et al., 2017). The increased O3 is thus unsurprising for the 2000 – 2015 timeframe. The long-term reduction
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in NOx with increasing VOC emissions concurrent with an increase in O3 at both sites suggests that the downtown
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Denver sites were in a NOx-saturated P(O3) regime, and as NO2* decreases and VOC reactivity increases, P(O3) was
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increasing towards peak production.
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3.2 Weekday-Weekend effect in Denver, CO
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The ‘weekday-weekend effect’ describes how emissions can be statistically different on weekdays versus weekends,
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resulting in different secondary chemistry. This effect can be used to elucidate information about local chemical
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regimes (i.e. CARB, 2003;Murphy et al., 2007;Fujita et al., 2003;Warneke et al., 2013;Pollack et al., 2012;Cleveland
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et al., 1974;Heuss et al., 2003). Traffic patterns in urban regions are different between weekdays and weekends, with
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heavier traffic and thus higher NOx on weekdays due to rush-hour and commercial trucking patterns. VOCs are
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expected to be stable across the week, as major VOC sources do not vary by day-of-week. Despite this drop in traffic,
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O3 can be higher on weekends than on weekdays if the system is in a NOx-saturated regime because decreased NOx
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increases P(O3), while decreased NO also reduces O3 titration to NO2 (Fujita et al., 2003;Heuss et al., 2003;Marr and
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Harley, 2002;Murphy et al., 2007;Pollack et al., 2012;Pusede and Cohen, 2012). Thus urban regions, which are often
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NOx-saturated, tend to follow a day-of-week pattern in both NOx and O3 (Fujita et al., 2003;Heuss et al., 2003;Pusede
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and Cohen, 2012), while rural and semi-urban areas often experience no change in NOx or O3 from weekdays to
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weekends. Rural regions have a lower population density, less defined daily traffic patterns, and minimal or no
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commercial trucking (Heuss et al., 2003). The weekday-weekend effect typically relies on the assumption that the
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VOC reactivity and thus HOx production is unchanged between the weekend and weekday. However, this is not always
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the case, as decreased weekend NOx reduces NOx+OH reactions, and thereby increases weekend OH and increased
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O3 (Warneke et al., 2013). Few studies of VOCs in the NFRMA exist, but our previous work found no significant
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difference in measured VOC reactivity at the BAO site between weekends and weekdays in summer 2015 (Abeleira
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et al., 2017).
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In the NFRMA, long-term (i.e. 10+ years) NO2* datasets only existed at the CAMP and Welby sites. Two sites in
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Denver added NO2* measurements in 2015, the I-25 and La Casa sites. The CAMP, I-25, and La Casa sites are all
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located within a 4-mile radius that straddles the I-25 motorway; are surrounded by a dense network of roads,
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businesses, and industrial operations; and experience high traffic density. Welby was located roughly 8-miles northeast
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from the three other sites, and borders a large lake and the Rocky Mountain Arsenal open space. Welby was thus more
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‘suburban’ than the other sites. Median NO2* at CAMP has decreased from 37 ppbv in 2003 to 13 ppbv in 2015. The
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median weekday I-25 and La Casa NO2* mixing ratios in 2015 were similar to CAMP in 2007 (Fig. 5) indicating that
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although NO2* emission reductions have been effective in the region, mixing ratios in Denver are very site specific
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An observable weekday-weekend effect in NO2* existed for all sites with NO2* data in all years except for Welby in
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2007 (Fig. 5). NO2* decreased by 20-50% from weekdays to weekends. Assuming that meteorology doesn’t
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systematically change between weekends and weekdays, we consider the weekend-weekday effect in O3 to be
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indicative of changes in P(O3) due to lower NOx. Figure 6 follows the analysis of Pusede and Cohen (2012), presenting
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summer average weekday and weekend O3 values for Welby and CAMP with the values tethered for each year. The
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values followed a curve similar to a modeled P(O3) curve, and indicates that reductions in NOx emissions from 2000
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to 2015 have placed O3 production in the Denver region in a transitional phase from NOx-saturated to peak P(O3).
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Regions that have higher NOx should observe greater impacts from changing VOCs than those that are closer to the
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peak P(O3). This analysis also suggested that continued reduction of NOx would shift the system to a NOx-limited
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regime, in which changes in VOC reactivity due to shifting anthropogenic or biogenic emissions would have little
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effect on O3.
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The average change in O3 (ΔO3) and NO2* (ΔNO2*) from weekend to weekday is plotted as a function of year for the
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two available NFRMA sites (Fig. 7a, Fig. 7b). A positive ΔO3 reflects a higher O3 concentration on the weekend than
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weekday, consistent with a NOx-saturated system. The weekday-weekend effect decreased from 2000 to 2015 for five
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of the six sites, all with similar ΔO3. This is consistent with the decreased regional NOx emissions, which would move
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the system from NOx-saturated to peak P(O3). The CAMP site was the exception, and consistently had a larger ΔO3
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than the other sites. This was consistent with the CAMP site’s higher NO2* relative to Welby and the 30-50% decrease
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in NO2* from weekdays to weekend. Measured NO2* decreased at both CAMP and Welby (Fig. 3), although the
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ΔNO2* at CAMP and Welby was unchanged, with average NO2* of -11 ± 3 ppbv and -1.7 ± 0.9 ppbv, respectively.
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Thus while absolute NOx emissions have changed, weekly traffic patterns have not. Applying a one-sided linear
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regression to the five-site ΔO3 median for 2001-2015 yielded a statistically significant decreasing trend of -0.5 ± 0.1
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ppbv/yr with an r2 = 0.55. The ΔO3 decreased across the NFRMA outside of the highest traffic regions in Denver, again
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consistent with the hypothesis that the NFRMA P(O3) regime has transitioned from NOx-saturated chemistry towards
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peak P(O3). Two specific sites, Greeley and Rocky Flats, show negative ΔO3 values in recent years, suggesting that
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those sites have, at least in those specific years, transitioned to NOx-limited chemistry.
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Collectively, this weekend-weekday analysis suggested that the region is NOx-saturated, but transitioning to a NOx-
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limited region. Increases in O3 are likely due to a combination of decreasing NOx and increasing VOC emissions.
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While the lack of long-term VOC measurements prevents identification and quantification of those VOC sources, the
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NEI suggested that petroleum-related VOCs have increased.
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3.3 The O3-temperature penalty in the NFRMA
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Increasing temperature can increase P(O3) by enhancing biogenic and evaporative VOC emissions, but has variable
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impacts on the weekday-weekend effect as a result of changing NOx emissions (Pusede et al., 2014). We showed that
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while O3 increased with temperature in the NFRMA, consistent with a NOx-saturated regime, this relationship was
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variable year to year. Ambient O3 was correlated with increasing temperature across the U.S. (Bloomer et al.,
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2009;Jacob and Winner, 2009;Pusede et al., 2014). While one study in the NFRMA from summer 2012 found that
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biogenic VOCs (i.e. isoprene) had a minor impact on VOC reactivity (McDuffie et al., 2016), Abeleira et al. (2017)
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found that isoprene contributed up to 47% of VOC reactivity on average in the late afternoon in summer 2015.
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Studying the temperature dependence of O3 allowed us to investigate the extent to which biogenic VOCs influenced
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P(O3) in the NFRMA and the interannual variability in those temperature-dependent VOC sources, as well as the shift
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from a NOx-saturated to NOx-limited P(O3) regime. NOx-saturated regimes should be sensitive to changes in VOC
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reactivity, while NOx-limited systems should not. We note that while anthropogenic VOCs, such as solvents, may be
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temperature dependent and contribute to this trend, we only observed temperature trends in isoprene at the BAO site
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in 2015 – though we acknowledge that the observed VOC suite in that study was limited (Abeleira et al., 2017).
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O3 in the NFRMA demonstrated a clear temperature dependence at all percentiles for all sites, but with slopes that
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vary by site and year (Fig. 8, Fig. 9). The NFRMA appears to be NOx-saturated or near peak P(O3) for all years,
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consistent with temperature dependent biogenic emissions having large impacts on ambient local O3. The variance in
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the O3-temperature dependence was likely external to meteorological effects. High temperature and linked
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meteorological parameters such as high 500 hPa heights, and stagnant winds, or circulating wind patterns do indeed
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correlate with high O3 events in Colorado (Reddy and Pfister, 2016), but those parameters should not affect the O3
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temperature relationship.
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Figure 8a shows daytime, summer O3 averaged in 3°C temperature bins for CAMP, Fort Collins, and Rocky Flats for
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years in which data was available at all sites. For every temperature bin, O3 was higher at Rocky Flats than at Fort
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Collins, and both were higher than at CAMP. The Rocky Flats site was the most rural of the chosen sites adjacent to
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the 4,000 acre Rocky Flats Wildlife Refuge, but was <15 miles from downtown Boulder. Rocky Flats likely had higher
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O3 because it was downwind of both NOx (Boulder, Denver) and VOC sources (forested regions in the neighboring
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foothills), had fewer nearby fresh NOx sources and thus less NO+O3 titration, and experienced enhanced P(O3) due to
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the region being near at the cross-over point between NOx-saturated and NOx-limited (Fig. 6).
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Bloomer et al. (2009) reported average O3-temperature relationships of 2.2 – 2.4 ppbv/°C for the Northeast, Southeast,
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and Great Lakes regions of the U.S. across all O3 percentiles. In contrast, the Southwest region, including Colorado,
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had an average relationship of 1.4 ppbv/°C (Bloomer et al., 2009). We find that O3 was indeed correlated with
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temperature at all NFRMA sites, with relationships that ranged from 0.07 to 1.95 ppbv/°C with an average of 1.0 ± 0.4
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ppbv/°C (Fig. 8) for all sites and years. Quantitatively, this temperature dependence was low relative to other U.S.
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sites, consistent with previous findings that biogenic VOCs contribute to, but did not dominate the VOC reactivity in
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the NFRMA (McDuffie et al., 2016;Abeleira et al., 2017). However, the six NFRMA sites exhibited significant
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variability in the 5th, 50th, and 95th percentiles among the sites both within<