Introduction
Tropospheric ozone (O3) is detrimental to human health, impacting the
frequency of asthma attacks, cardiovascular disease, missed school days, and premature
deaths. Based on these impacts, the Environmental Protection Agency (EPA)
projects that reducing the O3 standard to the new 70 ppbv 8 h average will
result in health benefits of USD 6.4–13 billion yr-1 (EPA,
2014). O3 also damages plants, reducing agricultural yields (Tai et
al., 2014). Using crop yields and ambient O3 concentrations for 2000,
Avnery et al. (2011) estimate the loss of USD 11–18 billion yr-1
worldwide as a result of the reduction of staple crops worldwide (soybean,
maize, and wheat) from O3 damage. During summer months, the northern Front Range metropolitan area (NFRMA) of Colorado consistently violated the
pre-2016 US EPA National Ambient Air Quality Standard (NAAQS) of 75 ppbv fourth-highest daily maximum 8 h average (MDA8) ambient O3
concentration, despite proposed reductions in anthropogenic emissions
(CDPHE, 2014). The NFRMA has been an O3 non-attainment zone
since 2008 (CDPHE, 2009), prompting the Colorado Air Pollution Control
Division and the Regional Air Quality Council to develop the Colorado Ozone
Action Plan in 2008 to target key O3 precursors: volatile organic
compounds (VOCs) and NOx (NO+NO2) (CDPHE, 2008). Despite
these control efforts, 2013 was the NFRMA's fourth year in a row to exceed
the federal O3 standard (CDPHE, 2016), and the eight NFRMA
non-attainment counties, with their combined population > 3.5 million, exceeded the MDA8 75 ppbv O3 standard on 9–48 days between
2010 and 2012 (AMA, 2015). However, Colorado must comply with the new 70 ppbv MDA8 standard by 2018. In order to accurately design and implement
O3 reduction schemes, a thorough understanding of local O3 trends
and chemistry is required.
Ground-level or boundary layer O3 depends on local production,
transport, and meteorological parameters:
∂[O3]∂t=PO3+weO3-ud[O3]H-∇×(v[O3]),
where ∂ [O3]/∂t represents the time rate of
change of O3 concentration, P(O3) is the instantaneous net
photochemical O3 production rate (production-loss),
weO3-ud[O3]/H represents the entrainment rate (we) of
O3 in and deposition rate (ud) of O3 out of the mixing
layer height (H), and ∇×(v[O3]) describes the
advection of O3 mixing layer height. Briefly, ground-level O3 is
driven by a catalytic chain that is initiated by RO2 production from
VOC oxidation (Reaction R1) and propagated by local NOx emissions (Reactions R2, R3).
RH+OH+O2→RO2+H2O
Chain propagation occurs through reactions between HO2 or RO2
radicals with NO to form NO2 (Reactions R2a, b, R3), which is photolyzed (Reaction R4) and
leads to net O3 formation (Reaction R5). Reactions between NO and O3 also
produce NO2 (Reaction R6), leading to a null cycle with no net O3
production. Alkoxy (RO) radicals form carbonyl-containing compounds and
HO2 (Reaction R7).
RO2+NO→RO+NO2
RO2+NO→RONO2
HO2+NO→NO2+OH
NO2+hν→NO+O(3P)
O(3P)+O2→O3
NO+O3→NO2+O2
RO+O2→R′CHO+HO2
For every VOC that enters the cycle, approximately two NO2 radicals are
produced – but the resulting carbonyl-containing compounds and organic
nitrates can be repeatedly oxidized or photolyzed, further propagating the
P(O3) chain. Chain termination occurs through RO2 and HO2
self-reactions to form peroxides (dominant termination reactions in the
“NOx-limited regime”), OH and NO2 reactions to form HNO3
(“NOx-saturated” or “VOC-limited” regime), or RO2 and NOx
reactions to form organic nitrates (RONO2) or peroxyacyl nitrates
(RC(O)O2NO2). Formation of organic and peroxyacyl nitrates
suppresses P(O3) but does not shift the crossover point between
NOx-limited and NOx-saturated P(O3) regimes (Farmer et
al., 2011). This crossover point of maximum, or peak, O3 production is
controlled by the chain termination reactions and is sensitive to the
HOx production rate and thus VOC reactivity. Decreasing NOx is an
effective O3 control strategy in a NOx-limited system but will
increase O3 in a NOx-saturated system. Controls for
NOx-saturated systems often focus on reducing anthropogenic VOC
reactivity and/or suppressing NOx emissions sufficiently that the
system becomes NOx-limited.
Trends in O3 for 2000–2015 varied across the United States
(EPA, 2016a). Using the annual fourth maximum of daily 8 h
averages (MDA-8), the EPA reported a 17 % decrease in the aggregated
national average O3. However, regional trends deviated substantially
from the national average. For example, the EPA reported a 25 % decrease
in O3 throughout the southeast, while the northeast showed a 16 %
decrease. Smaller decreases in O3 occurred in the northern Rockies
(1 %), in the southwest (10 %), and on the west coast (4–10 %). These O3
reductions are concurrent with national reductions in O3 precursors
of 54 % for NOx, 21 % for VOCs, and 50 % for CO (EPA,
2016b). Due to the nonlinear behavior of O3 chemistry described above,
reductions in O3 precursors do not necessarily result in reductions of
ambient O3. Cooper et al. (2012) reported that 83, 66, and
20 % of rural eastern US sites exhibited statistically significant
decreases in summer O3 at the 95th, 50th, and 5th
percentiles (1990–2010). No increases in O3 occurred at any sites,
indicating that local emission reductions have been effective in those
regions. In contrast, O3 in the western US followed a very different
trend: only 8 % of western US sites exhibited decreased O3 at the
50th percentile; the 5th percentiles for O3 at 33 % of the
sites actually increased. These increases were larger for the lower
percentiles, indicating that, while local emissions reductions may have been
effective at some sites, increased background O3 offset the
improvement.
Lefohn et al. (2010) found that, when comparing O3 at the same sites
for a longer period of 1980–2008 and shorter period of 1994–2008, the
predominant pattern was a change from a negative trend (decreasing O3)
during the longer period to no trend (stagnant O3) in the shorter
period, indicating that O3 reductions had leveled off by the late
2000s. The leveling-off could be a result of either slowed precursor
emissions reductions, which is contrary to the EPA estimates, or, more
likely, shifting O3 chemistry regimes as precursor emissions are
changing. McDonald et al. (2013) report decreased VOC, CO, and NOx automobile emissions in major US urban centers and decreasing
VOC/NOx trends from 1990 to 2007 with a turnaround and small increase
after 2007. This will affect local O3 chemistry within the city and at
downwind receptor sites. Lefohn et al. (2010) reported that the
distributions of high and low hourly O3 values narrowed toward
mid-level values in the 12 cities studied, consistent with a reduction in
domestic O3 precursors and possibly increased transport of O3 precursors from east Asia. Modeling and measurement studies have also
reported increased baseline O3 in the western US due to the transport
of O3 precursors from east Asia (Cooper et al., 2010; Parrish et al.,
2004; Pfister et al., 2011; Weiss-Penzias et al., 2006). These studies
questioned the effectiveness of local precursor emission reductions in
controlling local O3 in impacted regions.
The intermountain west is an intriguing environment with potentially
increasing background O3 (Cooper et al., 2012). The NFRMA is of
particular interest due to the challenge in effective O3 regulation,
its growing population, and the dominantly anthropogenic sources of O3
precursors. VOCs have been well studied in the region, with a particular
focus on the Boulder Atmospheric Observatory (BAO) in Erie, CO (e.g.,
Gilman et al., 2013; McDuffie et al., 2016; Pétron et al., 2012; Swarthout
et al., 2013; Thompson et al., 2014). VOC composition in the NFRMA was
heavily influenced by oil and natural gas (ONG) sources, as well as traffic.
In winter 2011, ∼ 50 % of VOC reactivity was attributed to
ONG-related VOCs, and ∼ 10 % to traffic (Gilman et al.,
2013; Swarthout et al., 2013). Recent studies have shown that ONG and traffic
contributed up to 66 and 13 % of the VOC reactivity, respectively, at
BAO in mornings for both spring and summer 2015 but that biogenic isoprene
was a large, temperature-dependent component of VOC reactivity in the
summer, contributing up to 49 % of calculated daytime VOC reactivity
(Abeleira et al., 2017). We note that the anthropogenic VOCs were
typically lower in 2015 than previous measurements, pointing to the complex
roles of meteorology, transport, and local emissions. In contrast, observed
isoprene in summer 2012 was much lower than summer 2015, likely due to
shifting drought conditions. While temperatures across the two summers were
similar, 2012 was a widespread drought year in the region, and 2015 was not.
Drought is typically associated with suppressed biogenic VOC emissions
(Brilli et al., 2007; Fortunati et al., 2008; Guenther, 2006). Local
anthropogenic and biogenic sources are not the only VOC sources in the
region: longer-lived VOCs consistent with transport have also been observed
(21–44 % of afternoon reactivity in 2015), and smoke from both local and
long-distance wildfires impacted air quality in the NFRMA in punctuated
events. This smoke was sometimes, but not always, associated with elevated
O3 (Lindaas et al., 2017).
The impact of a changing climate on air quality is poorly understood due to
the complex climate–chemistry interactions and numerous feedbacks (Jacob
and Winner, 2009; Palut and Canziani, 2007). However, increasing temperature
is expected to increase O3 (Bloomer et al., 2009; Jacob and Winner,
2009; Palut and Canziani, 2007). The O3–temperature relationship is
attributed to (1) temperature-dependent biogenic VOC emissions that provide
a source of VOCs for OH oxidation leading to increased HOx cycling
(Guenther, 2006; Guenther et al., 1996), (2) thermal decomposition of
peroxyacetyl nitrate (PAN) to HOx and NOx (Fischer et al.,
2014; Singh and Hanst, 1981), and (3) increased likelihood of favorable
meteorological conditions for ozone formation (i.e., high insolation, stagnation,
circulating wind patterns) (Reddy and Pfister, 2016; Thompson et al.,
2001). In addition, increased temperatures and changing soil moisture could
alter soil emissions of NOx. Due to the nonlinearity of P(O3)
chemistry as a function of NOx, the increased VOC and NOx
emissions associated with warming can either increase or decrease P(O3)
depending on local NOx levels (i.e., NOx-limited vs.
NOx-saturated). Interactions between climate change and regional-scale
meteorology are complex and may also impact O3. High and low O3 in
the US is coupled to a variety of meteorological parameters, including
planetary boundary layer (PBL) heights (White et al., 2007; Reddy and
Pfister, 2016), surface temperatures (Bloomer et al., 2009),
stratospheric intrusions (Lin et al., 2015), soil moisture, and regional
winds (Davis et al., 2011; Thompson et al., 2001). PBL height is coupled
to increased temperatures, reduced cloud cover, stronger insolation, and
lighter circulating wind patterns, with higher 500 hPa heights correlating to
higher average July O3 in the NFRMA (Reddy and Pfister, 2016).
Summary of measurement sites used in this analysis. Note that
NO2∗ refers to the NO2 detected by the EPA Federal Reference
Method and thus includes a fraction of NOy species.
Site
Latitude
Longitude
Elevation (m)
Measurements
CAMP
39.7512
-104.988
1591
O3 & NO2∗
Welby
39.8382
-104.955
1554
O3 & NO2∗
Carriage
39.7518
-105.031
1619
O3
Fort Collins
40.5775
-105.079
1523
O3
Greeley
40.3864
-104.737
1476
O3
Rocky Flats
39.9128
-105.189
1784
O3
I-25
39.7321
-105.015
1586
NO2∗
La Casa
39.7795
-105.005
1601
O3 & NO2∗
In this paper, we used temperature, O3, and NO2 data from
2000 to 2015 at multiple sites in the NFRMA to investigate why O3 has not
decreased in the region despite decreases in NOx. We used a
weekend–weekday analysis to elucidate the NOx regime for P(O3) in
Denver and explored the temperature dependence of O3 and the role of
drought in influencing that relationship in the NFRMA.
Results and discussion
Long-term trends in O3 and NO2∗ in the
northern Front Range metropolitan area
Contrary to most other places in the US, O3 in the NFRMA was either
stagnant or increasing between 2000 and 2015, despite substantial decreases
in NOx emissions. At most sites in the eastern US and some on the
west coast, O3 was decreasing at all percentiles. In the NFRMA,
however, five out of six monitoring sites exhibited no change or increasing
O3 at the 50th and 95th percentiles in the 2000–2015
period (Fig. 2). The 5th percentile is often taken as background
O3, and studies have shown that background O3 in the western US
has increased (Cooper et al., 2010; Parrish et al., 2004; Pfister et al.,
2011; Weiss-Penzias et al., 2006). However, only the CAMP and Welby sites in
Denver exhibit significant increasing O3 with trends of 1.3 ± 1.0 and 1.1 ± 1.0 ppbv yr-1, respectively, at the
5th percentile, with significance determined by passing an F-test
(Sect. 2.2). The CAMP and Welby sites also exhibit statistically
significant increases at the 50th (CAMP: 1.2 ± 0.4; Welby: 0.7 ± 0.5 ppbv yr-1) and 95th (CAMP: 1.0 ± 0.9; Welby:
0.7 ± 0.5 ppbv yr-1) percentiles. Cooper et al. (2012) reported
that the Welby site exhibited no statistically significant increase in
O3 from 1990 to 2010, contrary to what we found for 2000–2015 at the
95th percentile, which could be a result of changing VOC and
NO2∗ emissions in the 2010–2015 period.
(a) Estimated yearly averaged natural gas withdrawals in Colorado
(US-EIA, 2017). (b) Yearly average number of active ONG well operations
(US-EIA, 2017). (c) Anthropogenic VOC emission estimates from the EPA
state average annual emissions trend for Colorado (EPA, 2016b). Emission sources are
separated by color and are added to give the total VOC emission estimates
for anthropogenic VOCs. Biogenic VOCs and VOCs from biomass burning
(controlled fires and wildfires) are not included.
The increasing O3 trends in the NFRMA occurred despite reductions in
NOx. NO2∗ at the CAMP site decreased significantly from 2000 at a
rate of -1.0 ± 0.6 and -1.4 ± 0.6 ppbv yr-1 for the
50th and 95th percentiles, respectively, for CAMP (Fig. 3). Welby exhibited a
non-significant decreasing NO2∗ trend at the 95th percentile of
-0.7 ± 0.8 ppbv yr-1 (Fig. 3). The increased O3 may be due to
increased summer temperatures in Colorado, increased regional baseline
O3, or increased local P(O3) from unknown emission sources
(Cooper et al., 2012). VOC emissions steadily increased in Colorado from
2000 to 2012 per the EPA state average annual emissions trend (Fig. 4). To
the best of our knowledge, the NFRMA does not have any long-term VOC
datasets, but the EPA state average annual emissions trend for Colorado
provided an estimate for yearly anthropogenic VOC (AVOC) emissions
(EPA, 2016b). All categories of AVOC emissions decreased slightly from
2000 to 2015, except for petroleum-related VOCs, which increased from 7.4×103 tons in 2000 to 2.6×105 tons in 2011 with a decrease to
1.5×105 tons in 2015 (Fig. 4). The US Energy Information
Administration (EIA) reports a twofold increase in active ONG wells from
∼ 25 000 to ∼ 40 000 from 2010 to 2012 (Fig. 4c)
(US-EIA, 2017). However, we note the state average annual emissions
are only an estimate and do not include biogenic sources of VOCs, which can
contribute substantially to VOC reactivity in the region but vary
substantially from year to year (Abeleira et al., 2017). The
increased O3 is thus unsurprising for the 2000–2015 time frame. The
long-term reduction in NOx with increasing VOC emissions concurrent
with an increase in O3 at both sites suggests that the downtown Denver
sites were in a NOx-saturated P(O3) regime and that, as NO2∗
decreased and VOC reactivity increased, P(O3) was increasing towards
peak production.
Weekend–weekday effect in Denver, CO
The “weekend–weekday effect” describes how anthropogenic emissions of
O3 precursors can be statistically different on weekdays versus
weekends, resulting in different secondary chemistry. This effect can be
used to elucidate information about local chemical regimes (i.e., CARB,
2003; Murphy et al., 2007; Fujita et al., 2003; Warneke et al., 2013; Pollack et
al., 2012; Cleveland et al., 1974; Heuss et al., 2003). Traffic patterns in
urban regions are different between weekends and weekdays from a decrease in
heavy-duty truck traffic on weekends (Marr and Harley, 2002). VOCs are
expected to be stable across the week, as major VOC sources do not vary by
day of week. Despite this reduction in heavy-duty trucking traffic, O3
can be higher on weekends than on weekdays if the system is in a
NOx-saturated regime because decreased NOx increases P(O3),
while decreased NO also reduces O3 titration to NO2 (Fujita et
al., 2003; Heuss et al., 2003; Marr and Harley, 2002; Murphy et al.,
2007; Pollack et al., 2012; Pusede and Cohen, 2012). Thus urban regions, which
are often NOx-saturated, tend to follow a day-of-week pattern in both
NOx and O3 (Fujita et al., 2003; Heuss et al., 2003; Pusede and
Cohen, 2012), while rural and semi-urban areas often experience no change in
NOx or O3 from weekdays to weekends. Rural regions have a lower
population density, less defined daily traffic patterns, and minimal or no
commercial trucking (Heuss et al., 2003). The weekend–weekday effect
typically relies on the assumption that the VOC reactivity and thus HOx
production are unchanged between the weekend and weekdays. However, this is
not always the case, as decreased weekend NOx reduces NOx+OH
reactions and thereby increases weekend OH and O3
(Warneke et al., 2013). Few studies of VOCs in the NFRMA exist, but our
previous work found no significant difference in measured VOC reactivity at
the BAO site between weekends and weekdays in summer 2015 (Abeleira
et al., 2017).
Weekend–weekday analysis (Sunday vs. Wednesday) for O3 (black with shading) and NO2∗ (blue) for the CAMP (a, squares), Welby
(b, circles), and La Casa (c, diamonds) sites in Denver. I-25 (d,
triangles) is limited to NO2∗ due to data availability. All
sites have plots for 2015, but only CAMP (a) and Welby (b) are additionally
plotted for 2007 and 2012 due to data availability. Wednesday is
representative of weekday NO2∗ and typically is not different
than the average of Tuesday, Wednesday, and Thursday at a 95 % confidence
for this dataset. Monday, Friday, and Saturday are considered carryover or
“mixed” days between weekdays and weekends and are ignored. Error bars
represent 95 % confidence intervals around the summertime mean of
Wednesday or Sunday O3 or NO2∗.
Weekday and weekend O3 versus NO2∗ for Welby (black)
and CAMP (blue) sites. Tethered symbols correspond to average Wednesday
values for weekdays and average Sunday values for weekends for each year
depending on data availability. The color shading corresponds to year, with
the lightest shade corresponding to the earliest year (2000 for Welby, 2005
for CAMP) and 2015 as the darkest shade. The 95 % confidence intervals for
each year are < 5 ppbv for O3 and
< 2.5 ppbv for NO2∗. The dashed blue line is a visual aid to guide the
reader's eye to the nonlinear O3 curve and was generated from the
simple analytic model described by Farmer et al. (2011).
In the NFRMA, long-term (i.e., 10+ years) NO2∗ datasets only existed
at the CAMP and Welby sites. Two sites in Denver added NO2∗
measurements in 2015: the I-25 and La Casa sites. The CAMP, I-25, and La
Casa sites are all located within a 6 km radius that straddles the I-25
motorway; are surrounded by a dense network of roads, businesses, and
industrial operations; and experience high traffic density. Welby is located
roughly 13 km northeast from the three other sites and borders a large
lake and the Rocky Mountain Arsenal open space. Welby is thus more
“suburban” than the other sites. Median NO2∗ at CAMP
decreased from 37 ppbv in 2003 to 13 ppbv in 2015. The median
weekday I-25 and La Casa NO2∗ mixing ratios in 2015 were similar to
CAMP in 2007 (Fig. 5), indicating that, although NO2∗ emission reductions
have been effective in the region, mixing ratios in Denver are very site
specific.
An observable weekend–weekday effect in NO2∗ existed for all years at
the CAMP site, and most years at the Welby site with intermittent years
that do not have a clear difference in weekday and weekend NO2∗. NO2∗ decreased by 20–50 % from weekdays to weekends. Assuming
that meteorology does not systematically change between weekends and
weekdays, we consider the weekend–weekday effect in O3 to be indicative
of changes in P(O3) due to lower NOx. Figure 6 follows the
analysis of Pusede and Cohen (2012), presenting summer average
weekday and weekend O3 values for Welby and CAMP with the values
tethered for each year. The values followed a curve similar to a modeled
P(O3) curve and indicates that reductions in NOx emissions from
2000 to 2015 have placed O3 production in the Denver region in a
transitional phase from NOx-saturated to peak P(O3). This analysis
suggests that continued reductions of NOx would shift the system to a
NOx-limited regime, in which changes in VOC reactivity due to shifting
anthropogenic or biogenic emissions would have little effect on O3.
(a) The change in O3 calculated as average weekend (Sunday)
minus weekday (Wednesday) O3 for the six NFRMA sites identified by
color and marker. The solid grey line is the average of the sites. The
inclusion of a site in the averaging for a given year was dependent on
available data for that year. The light-grey shading represents ± the
95 % confidence interval of all Wednesday and Sunday hourly values for
each year for sites with available data. (b) The change in NO2∗ is
calculated identically to O3 in (a) for the CAMP and Welby sites, and
the error bars represent the 95 % confidence interval of the averages.
The average change in O3 (ΔO3) and NO2∗ (ΔNO2∗) from weekend to weekday is plotted as a function of year for the
six available O3 NFRMA sites and the two NO2∗ sites
(Fig. 7a, b). A positive ΔO3 reflects a higher O3
concentration on the weekend than weekdays, consistent with a
NOx-saturated system. A negative ΔO3 is consistent with a
NOx-limited system in which O3 decreases when NOx decreases.
The weekend–weekday effect exhibits a non-significant decreasing trend from
2000 to 2015 for yearly averages of the six sites. This is consistent with
the decreased regional NOx emissions, which would move the system from
NOx-saturated to peak P(O3) in the absence of large changes in VOC
reactivity. The CAMP site was the exception and consistently had a larger
ΔO3 than the other sites. This was consistent with the CAMP
site's higher NO2∗ relative to Welby and the 30–50 % decrease in
NO2∗ from weekdays to the weekend. Measured NO2∗ decreased at both
CAMP and Welby (Fig. 3b), but with larger decreases at the CAMP site. The
ΔNO2∗ at Welby remained stable with an average value of -1.7±0.9 ppbv, while ΔNO2∗ at the CAMP exhibited a
statistically significant decrease of 0.6 ± 0.4 ΔNO2∗
ppbv yr-1. The decreasing ΔNO2∗ at the CAMP site
appears to be converging with the ΔNO2∗ at the Welby site. It
is unlikely that traffic patterns are assimilating between the two sites,
and a more plausible explanation is that emission control technologies on
heavy-duty commercial fleet vehicles are reducing the impact on emissions of
those specific vehicles and reducing the measurable ΔNO2∗
(Bishop et al., 2015). The ΔO3 decreased across
the NFRMA outside of the most highly trafficked regions in Denver, again consistent
with the hypothesis that the NFRMA P(O3) regime has transitioned from
NOx-saturated chemistry towards peak P(O3). Two sites, Greeley and
Rocky Flats, show negative ΔO3 values in recent years,
suggesting that those sites have, at least in those specific years,
transitioned to NOx-limited chemistry. Collectively, this
weekend–weekday analysis suggests that the region is NOx-saturated, but
transitioning to a NOx-limited region. Increases in O3 may thus be
due to a combination of decreasing NOx and increasing VOC emissions.
While the lack of long-term VOC measurements prevents identification and
quantification of those VOC sources, the state average annual emissions
suggested that petroleum-related VOCs have increased. However, we note that
large increases in VOC reactivity shift the transition point between
NOx-limited and NOx-saturated regions to higher NOx
concentrations. The clear regional decrease in the weekend–weekday effect,
as evidenced by the decreasing ΔO3 trend, indicates that the
region is transitioning and that any increases in VOC reactivity have not
been so large as to dramatically inhibit this effect.
The O3–temperature penalty in the NFRMA
Increasing temperature can increase P(O3) by enhancing biogenic and
evaporative VOC emissions but has variable impacts on the weekend–weekday
effect as a result of changing NOx emissions (Pusede et al.,
2014). We showed that while O3 increased with temperature in the
NFRMA, consistent with a NOx-saturated regime, this relationship was
variable year to year. Ambient O3 was correlated with increasing
temperature across the US (Bloomer et al., 2009; Jacob and
Winner, 2009; Pusede et al., 2014). While one study in the NFRMA from summer
2012 found that biogenic VOCs (i.e., isoprene) had a minor impact on VOC
reactivity at the BAO site (McDuffie et al., 2016), Abeleira et al. (2017) found that isoprene contributed up to 47 % of VOC reactivity on
average in the late afternoon in summer 2015. Studying the temperature
dependence of O3 allows us to investigate the extent to which biogenic
VOCs influenced P(O3) in the NFRMA and the interannual variability of
those temperature-dependent VOC sources, as well as the shift from a
NOx-saturated to NOx-limited P(O3) regime. NOx-saturated
regimes should be sensitive to changes in VOC reactivity, while
NOx-limited systems should not. We note that, while anthropogenic VOCs
such as solvents may be temperature dependent and contribute to this trend,
we only observed temperature trends in isoprene at the BAO site in 2015 – though we acknowledge that the observed VOC suite in that study was limited
(Abeleira et al., 2017).
(a) O3 versus temperature for CAMP, Fort Collins, and Rocky
Flats. Hourly O3 is binned by hourly temperature, with bins containing
51–110 points for O3 and temperature depending on data availability
at a site. The temperature bins typically contained 100–110 data points
(> 90 % of temperature bins for all sites in all available
years). Average O3 of each bin is plotted versus the average
temperature of each bin. Markers and colors represent yearly averages for
each site. Error bars were left off for visual clarity, but the 95 %
confidence intervals around the yearly bin averages are typically
< 8 ppbv. Years were selected based on availability of overlapping data for
multiple sites. (b) One-sided linear regressions of equal point temperature
bins for the 5th (red open diamond), 33rd (pink hash),
50th (green open triangle), 67th (blue open square), and
95th (black open circle) percentiles for the CAMP site for 2007
(left), 2012 (middle), and 2015 (right).
Slopes from one-sided linear regression of O3 versus
temperature (i.e., the temperature dependence of O3). Hourly O3 (10:00–16:00 LT) is binned by hourly temperature, with bins containing
51–110 points for O3 and temperature depending on data availability
at a site. The temperature bins typically contained 100–110 data points
(> 90 % of temperature bins for all sites in all available
years). The slopes of O3 versus temperature for the 5th,
50th, and 95th percentiles for the O3–temperature bins
are shown. Data are shown for CAMP (black squares), Welby (grey solid
circles), Carriage (blue open triangles), Fort Collins (green solid
squares), Greeley (teal Xs), and Rocky Flats (magenta open diamonds).
Shaded years correspond to Colorado summers with moderate to severe drought
conditions. Error bars are ±95 % confidence interval of the slopes.
Faint grey line across the 50th percentile is the average slope bounded
by the 95 % confidence interval for years excluding 2008, 2011, and 2012.
O3 in the NFRMA demonstrated a clear temperature dependence at all
percentiles for all sites, but with slopes that vary by site and year (Fig. 8, Fig. 9). The NFRMA appears to be NOx-saturated or near peak
P(O3) for all years, consistent with temperature-dependent biogenic
emissions impacting ambient O3. The variance in the O3–temperature
dependence was likely external to meteorological effects. High temperature
and linked meteorological parameters – such as high 500 hPa heights,
stagnant winds, or circulating wind patterns – do indeed correlate with
high-O3 events in Colorado (Reddy and Pfister, 2016), but those
parameters should not affect the O3–temperature relationship.
Figure 8a shows daytime summer O3 averaged in non-uniform temperature
bins with bin size dictated by maintaining an equal number of data points in
each temperature bin for CAMP, Fort Collins, and Rocky Flats for years in
which data were available at all sites. For every temperature bin, O3
was higher at Rocky Flats than at Fort Collins, and both were higher than at
CAMP. The Rocky Flats site was the most rural of the chosen sites adjacent
to the 1600 ha Rocky Flats Wildlife Refuge but was
< 24 km
from downtown Boulder. Rocky Flats likely had higher O3 because it was
downwind of both NOx (Boulder, Denver) and VOC sources (forested
regions in the neighboring foothills), had fewer nearby fresh NOx
sources and thus less NO+O3 titration, and experienced enhanced
P(O3) due to the region being near the crossover point between
NOx-saturated and NOx-limited chemical regimes (Fig. 6).
Bloomer et al. (2009) reported average O3–temperature relationships
of 2.2–2.4 ppbv ∘C-1 for the northeast, southeast, and
Great Lakes regions of the US across all O3 percentiles. In contrast,
the southwest region, including Colorado, had an average relationship of
1.4 ppbv ∘C-1 (Bloomer et al., 2009). We find that O3 was
indeed correlated with temperature at all NFRMA sites, with relationships
that ranged from 0.07 to 1.95 ppbv ∘C-1 with an average of
1.0 ± 0.4 ppbv ∘C-1 (Fig. 8) for all sites and years.
Quantitatively, this temperature dependence was low relative to other US
sites, consistent with previous findings that biogenic VOCs contribute to,
but do not dominate, VOC reactivity in the NFRMA (McDuffie et al.,
2016; Abeleira et al., 2017). However, the six NFRMA sites exhibited
significant variability in the 5th, 50th, and 95th
percentiles among the sites both within a given year and across years (Fig. 9). The 5th and 95th O3 percentiles showed greater
variability and larger uncertainties in the slopes than the 50th
percentile. This indicated that baseline O3 and high-O3 events in
the region were less dependent on temperature. Baseline O3 was likely
tied to the transport of O3 and O3 precursors from the west coast
(Cooper et al., 2012), while the high-O3 events were likely tied to
a combination of meteorological parameters, including 500 hPa heights and
stagnation events (Reddy and Pfister, 2016), stratospheric intrusions
(Lin et al., 2015), and local temperature-independent VOC emissions. In
contrast, the 50th percentile showed a clear temperature dependence at
all sites in most years (Fig. 8, Fig. 9), indicating that mean O3 was
typically influenced by local temperature-dependent, and likely biogenic,
VOC emissions.
Unlike for ambient O3 and the weekend-to-weekday ΔO3, we
noted no clear long-term trend in the O3–temperature relationship. The
O3–temperature relationships showed similar interannual patterns for
the six sites at the 50th percentile (Fig. 9). Specifically, years
2008 and 2011–2012 have suppressed O3–temperature slopes for the
50th percentile. Reddy and Pfister (2016) reported high 500 hPa
heights and O3 for 2002–2003, 2006, and 2012, while 2004 and 2009 had
low 500 hPa heights and low O3, so those exceptional years cannot be
explained solely by meteorology. However, those exceptional years (2008 and
2011–2012) did correspond to years in which Colorado was in moderate–severe
drought with little soil moisture (NOAA, 2017). Years 2002–2003 also
exhibited moderate to severe drought conditions in Colorado, and some but
not all sites exhibited suppressed O3–temperature slopes.
Drought in the NFRMA is connected to changes in mountain–plains circulation
and lower surface moisture, which reduces the surface latent heat flux and
causes increased surface temperature. These increased surface temperatures
lead to strong mountain–plains circulation, stagnant wind conditions, higher
PBLs, and 500 hPa heights, all of which are known to correlate with
high-O3 episodes (Reddy and Pfister, 2016; Ek and Holtslag, 2004; Zhou and
Geerts, 2013). Drought is also connected to reduced isoprene emissions
(Brilli et al., 2007; Fortunati et al., 2008; Guenther et al., 2006). Consistent
with this concept, Abeleira et al. (2017) noted that isoprene was
2–4 times higher at the Boulder Atmospheric Observatory site in summer 2015
(a non-drought year) than in summer 2012 (a drought year). Such a decrease
in biogenic isoprene emissions should also suppress the O3–temperature
dependence in NOx-saturated regimes, a trend that was observed in the
NFRMA (Fig. 9).
The suppressed O3–temperature relationship during drought years in the
NFRMA demonstrated the importance of temperature-dependent VOCs in driving
P(O3) in the region, particularly at the mid-range 50th percentile
– but not at the baseline 5th percentile. A standard t test showed
that the 50th- and 95th-percentile slopes (i.e., temperature
dependence of average and high O3 concentrations) are indeed different
between the drought and non-drought years at the 95 % confidence limit. If
NOx emissions continue to decrease, and the NFRMA continues its trend
towards a NOx-limited regime (Fig. 7), the O3–temperature
dependence should also decrease and temperature-dependent VOCs will play a
smaller role in driving O3 production. However, this would require
substantial decreases in NOx for the heavy-traffic region of Denver to
become fully NOx-limited, so temperature-dependent VOCs will likely
remain important in at least some regions of the NFRMA.