Introduction
Wintertime ozone air pollution has recently been observed in several North
American basins and currently represents one of the most severe air pollution
problems in the United States . It has been associated with
emissions from oil and gas operations coupled with meteorological conditions
that produce high surface albedo and temperature inversions, causing stable
stagnation events. As with more conventional summertime urban air pollution,
winter ozone production requires photochemistry of NOx
(= NO + NO2) and volatile organic compounds (VOCs). In polluted
areas, such as the Uintah Basin, NOx is emitted mainly from fossil fuel
combustion and can further oxidize to form reactive nitrogen species such as
HNO3, acyl peroxynitrates (PAN), N2O5, NO3,
ClNO2 and organic nitrates, which together with NOx make up
total reactive nitrogen (NOy). Oxidation of NOx occurs through
different reaction pathways during the day than at night, but both contribute
significantly to NOy speciation. Some of these species tend to be
permanent sinks of NOx, such as HNO3, whereas others such as
PAN or N2O5 can act as temporary sinks (reservoirs) and
revert to NOx via photo- or thermochemistry. Thus, an understanding of the
reactive nitrogen budget contributes to understanding ozone formation.
To study the conditions and precursors that cause these anomalous wintertime
ozone events, we deployed a suite of ground-based chemical, radiation, and
meteorological measurements as part of the Uintah Basin Winter Ozone Studies
(UBWOS) in 2012, 2013, and 2014. The UBWOS studies in 2012 and 2013
experienced very different meteorological conditions and yielded strikingly
different results. In 2012, the lack of snow cover and the associated shallow
inversions produced ozone with average values that showed distinct
photochemistry but did not approach the 75 ppbv 8 h National Ambient Air
Quality Standard (NAAQS), presenting a valuable baseline of chemical
concentrations for this oil- and gas-producing region . In
2013, however, the snow cover resulted in strong temperature inversions,
increased precursor concentrations, and increased photochemistry, which
brought about elevated ozone levels . The Horsepool
measurement site in the basin experienced exceedances of the ozone NAAQS on
20 out of the 28 days of measurement in 2013. In 2014 the conditions were
intermediate both meteorologically and chemically. A direct comparison of
2012 with 2013 provides valuable insight into the key elements that cause
high wintertime ozone. In this paper we focus on reactive nitrogen and its
partitioning during the two years to help explain the chemical processes that
cause high ozone.
Field campaigns and measurement techniques
The three successive campaigns were conducted on
15 January–27 February 2012, 23 January–21 February 2013, and
28 January–14 February 2014 at the Horsepool site near Vernal, Utah. The
site is located at 40.14370∘ N, 109.46718∘ W, 35 km south
of Vernal, Utah, the largest city in the basin. The basin is mostly rural,
with a total population of 50 000 concentrated mainly in three towns
(Vernal, Roosevelt, and Duchesne). Approximately 10 000 producing oil/gas
wells are spread throughout the basin, and the Horsepool measurement site is
situated within the predominantly natural-gas-producing wells in the eastern
half of the basin, as seen in Fig. .
Measurements of ambient gas-phase reactive nitrogen during
UBWOS 2012–2014. The method abbreviations are described in Sect. 2, and LOD
refers to the limit of detection. Not all measurements were used in this
analysis.
Species measured
Campaign year
Method
Accuracy
LOD
Reference
2012
2013
2014
%
pptv
NO,NO2,NO3,N2O5
x
x
x
CRDS
5–10
1–100
NOy
x
CL
20
10–100
NOy
x
x
TD-CRDS
10
20
HNO3,HONO
x
x
x
acid CIMS
30
10
Alkyl & peroxynitrates
x
TD-LIF
20
24–34
Acyl peroxynitrates
x
x
x
I-CIMS
20
10
ClNO2
x
x
x
I-CIMS
20
5
HO2NO2
x
I-CIMS
20
5
NO2,NO3,HONO
x
x
LP-DOAS
3–8
80, 2, 20
NO2,HONO
x
ACES
15
200
HONO
x
LoPAP
15
10
The suite of measurements over the three years varied but was very extensive
every year, and descriptions can be found in the final reports for the Uintah
Basin Ozone Studies on the website of the Utah Department of Environmental
Quality (www.deq.utah.gov/locations/U/uintahbasin/ozone/overview.htm).
A brief summary of the ambient gas-phase reactive nitrogen measurements is
given here. During all three years, NO, NO2, NO3, and
N2O5 were measured using cavity ring-down spectroscopy (CRDS), which
was also used in conjunction with thermal dissociation (TD-CRDS) to measure
NOy in 2013 and 2014 . In 2012, NOy was measured using
catalytic conversion to NO on a gold tube at 325 ∘C with
subsequent detection using chemiluminescence (CL) via the reaction with
O3. Nitric and nitrous acids were measured with an acetate ion
chemical ionization mass spectrometer (acid CIMS) all three years. Alkyl
nitrates and peroxynitrates were only measured in 2012, by thermally
dissociating them to NO2 and subsequently detecting them via
laser-induced fluorescence (TD-LIF). Acyl peroxynitrates (PANs) and
nitryl chloride (ClNO2) were measured all three years using an iodide
chemical ionization mass spectrometer (I-CIMS). Finally, there was extra
focus on HONO in 2014, which was measured by a long-path differential
optical absorption spectrometer (LP-DOAS), a broadband cavity-enhanced
spectrometer (ACES), and a long-path absorption photometer (LoPAP), as well
as the acid CIMS and the I-CIMS. The measurements and references for the
techniques are summarized in Table . Due to the overlap or lack of
some measurements in different years, not all the data were utilized in this
analysis.
Results
Ozone and reactive nitrogen levels
In this analysis we focus on analysis of diel profiles, averaged over the
duration of each field campaign. This method highlights the general
differences between the years but does not distinguish between different
meteorological conditions within a campaign. In Fig. , we
show whole-campaign diel averages of the ozone levels at the Horsepool ground
site for the winters of 2012, 2013, and 2014. The dotted line shows the NAAQS
level of 75 ppbv. On average, ozone levels were 2.5 times higher in 2013
than in 2012. Additionally, ozone production during midday (between the
dotted lines at 09:45 and 14:30 mountain standard time)
was 2.7 ppbv h-1 in 2012
and 6.9 ppbv h-1 in 2013, a factor of 2.6 higher. In 2014, the ozone
levels were intermediate, with the daily increase at 4.8 ppbv h-1.
Although the ozone increase is affected by both chemical production and
dilution due to the changing boundary layer, chemical ozone production
accounts for most of this increase at this site. For 2012, when atmospheric
conditions were least stable, chemical production was estimated to account
for 70–85 % of the observed average diel rise in surface O3. These
estimates were derived from comparison of the model to the measured surface
level rise and from measurements of the diel average O3 profile at
different heights up to 500 m from a tethered balloon. .
Diel averages of ozone mixing ratios during the campaigns in 2012
(45 days), 2013 (28 days), and 2014 (27 days), and the 75 ppbv NAAQS for
reference. Average ozone levels were 2.5 times higher in 2013 than 2012.
Linear fits to the midday ozone increase illustrate the difference in
average daily ozone production, plotted on the right.
The top plot in Fig. shows the diurnally averaged total
reactive nitrogen (NOy). The NOy in 2013 is on average a factor of 2.5
higher than 2012, with 2014 again at intermediate levels. However, the middle
plot of Fig. shows that the total NOx concentrations are
consistently similar for all three years, despite significantly different
meteorological conditions and ozone production rates. The bottom plot shows
the ratio NOx / NOy, a measure of the rate of oxidation of reactive
nitrogen independent of dilution, whereby a lower ratio implies more
oxidation. The large differences in this ratio (a factor of 2.6 on average
between 2012 and 2013) instead indicates large differences in levels of
NOx oxidation caused by changes in ambient chemistry, which caused the
similarity in NOx levels between the measurement years.
Diel averages of reactive nitrogen. Top: total NOy was a factor
of 2.5 times larger in 2013 than in 2012. Middle: the amount of
photochemically active NOx remained at similar levels all three years.
Bottom: the ratio of NOx/NOy, an inverse measure of the level of
oxidation of reactive nitrogen, was a factor of 2.6 smaller in 2013 than
2012.
Partitioning among reactive nitrogen species for 2012 and 2013,
shown as diel averages (left) as well as daytime and nighttime pie charts
(right). We take total NOz to be the sum of components in 2012, and the
difference between NOy and NOx in 2013. The missing NOz in 2013
(labeled “other” in the pie charts) is likely organic nitrates, for which we
do not have measurements in 2013. In 2012, daytime organic nitrates and
nighttime N2O5 and ClNO2 play an important role compared to
2013, where total PANs and HNO3 are the largest contributors
to NOz.
NOy partitioning and NOx oxidation
We examine the oxidation pathways and products in order to understand the
different levels of NOx oxidation for the various years.
Figure shows the partitioning of NOz
(≡ NOy - NOx) for 2012 and 2013. In 2012, since NOx
makes up approximately 80 % of NOy, the subtraction to calculate NOz
results in a noisy trace with large uncertainty relative to the amount of
NOz present, and we instead take the sum of components to define total
NOz. This is not the case in 2013, and the “missing” part of NOz is
likely organic nitrates (RONO2) for which we do not have a
measurement.
Ammonium nitrate might be measured partially in the acid CIMS and the NOy
instrument due to heated inlets, and its contribution to NOz has not been
included in this analysis. Measurements of aerosol nitrate, which would
include coarse-mode aerosol whose source might not be exclusively
photochemical, present an average upper limit of 0.4 ppbv in 2012 and 1 ppbv
in 2013. Nitrous acid, HONO, was measured as a small fraction
(2.4 %) of NOz in 2012. Its mixing ratio was measured by both the acid
CIMS and DOAS measurements, which both showed maximum values smaller than 120
pptv average at night and smaller during the day, with agreement to within a
factor of 2. During 2013, the acid CIMS was the only measurement available.
It showed very large signals at the mass normally interpreted as HONO
with a distinct, daytime maximum. As described in ,
HO2NO2 mixing ratios were observed to reach an average daytime
maximum of approximately 4 % of NOz. Unpublished laboratory results
suggest that a large fraction of the HO2NO2 is detected as
HONO using the acid CIMS, resulting in a positive daytime bias in the
2013 measurements. Based on the similarity of DOAS HONO measurements
in 2012 and 2014, HONO for 2013 was set equal to that from 2012. For
further details on comparisons of HONO measurements, please see
.
In 2012, N2O5 and ClNO2 make up about half of the total
NOz budget at night, whereas they form a small percentage in 2013. Nitric
acid (HNO3) and PAN, however, make up about 75 % of total
NOz throughout the whole diel cycle in 2013, with the inferred organic
nitrates making up most of the remainder. The major oxidation pathways that
produce these compounds during the day are
NO2+OH+M⟶HNO3+M,NO2+PA+M⟶PAN+M,NO+RO2+M⟶αRONO2+M,
where PA is the peroxyacetyl radical and includes all acyl peroxy
radicals, with CH3C(O)O2 being the most important. RO2 includes all
other organic peroxy radicals, and α is the temperature-dependent
yield of organic nitrates from the reaction of organic peroxy radical with
NO, where the majority of this reaction produces an alkoxy radical and
NO2 . At night, when NO3 is photochemically
stable, the main pathway for NOx oxidation is
NO2+O3⟶NO3+O2,NO3+NO2+M⟶N2O5+M.
This N2O5 can then further react heterogeneously to form nitric acid
and nitryl chloride.
N2O5+H2O⟶het2HNO3N2O5+HCl⟶hetHNO3+ClNO2
Calculating the reaction rates of Reactions (R1)–(R5) allows us to compare NOx loss rates (rates
of conversion to NOz) through these different pathways. The reaction rate
constants are known, and the concentrations of OH and PA are
supplied by a box-model simulation using the Master Chemical Mechanism (MCM),
as is the production rate of organic nitrates. The MCM utilizes greater than
104 reactions, and the base run accurately reproduces an ozone buildup
event in 2013 . Additionally, the OH concentrations
agree with OH inferred from VOC ratios with average
midday maximum OH levels calculated by the model to be approximately
1×106 cm-3. During the 2012 study, calculated midday OH
was 7×105 cm-3 . Although PAN can
thermally dissociate, the long lifetime at wintertime temperatures
(> 10 h below 10 ∘C) means we can effectively consider only the
forward reaction. The limiting step in Reactions (R4)–(R5) is the NO2+O3 reaction and we assume that the sequence of reactions in
Reactions (R4)–(R7) quantitatively converts NO2 to stable products,
mainly HNO3, at night in 2013 (we calculate N2O5 lifetimes
to be < 2 h; see below). The NOx loss rate due to Reaction (R4) is
doubled, because the sum of Reactions (R4) and (R5) would lead to NOx loss
at twice the rate of Reaction (R4). The reaction pathway to make
N2O5 is negligible during daylight hours due to photodissociation of
NO3 together with the fast reaction of NO3 with NO,
and has been set to zero. The resulting 2013 NOx loss rates due to
Reactions (R1)–(R5) are shown in Fig. .
Daytime and nighttime loss rates of NOx in 2013 through the major
oxidation pathways. Concentrations of OH and PA and the
production rate of organic nitrates (NO + RO2) were supplied by the
Master Chemical Mechanism box model used by . The daytime
NO3 production is set to zero because of the fast NO3
photolysis and reaction with photochemically generated NO, and doubled
at night due to Reaction (R5). The integrated nighttime loss toward
HNO3 is 5.9 times greater than during the day.
Separating the daytime and nighttime partitioning in Fig.
highlights the species that are long-lived at night and short-lived during
the day (N2O5 and ClNO2), demonstrating the role of the
nighttime species in reactive nitrogen chemistry. Nitric acid, PAN,
and organic nitrates, on the other hand, are long-lived compared to a diel
cycle, and we do not expect the nighttime or daytime average to reflect
chemical production that is restricted to these periods. It instead
represents an average not just over a diel cycle but over the whole campaign.
Integrating the diurnally averaged loss rates gives total daily calculated
production of the three major components of NOz, with the simplifying
assumption that all N2O5 is converted to nitric acid (we estimate
the ClNO2 yield for 2012 and 2013 to be 11 and 2 %, respectively).
In Fig. we compare the partitioning of these integrated
production rates with the measured partitioning of HNO3, PAN,
and inferred organic nitrates for 2013. Production rates and observed
concentrations should not necessarily be proportional, depending on the loss
mechanisms. For example, HNO3 will be lost via dry deposition to the
ground or snow surface such that its measured contribution to nitrogen
partitioning may be smaller than that inferred from its production rate.
However, the agreement between production rates and observations illustrates
that our methods of treating the reactive nitrogen in the current analysis
and in the MCM box model are self-consistent.
Comparison of the relative importance in 2013 of calculated oxidized
reactive nitrogen production rates to the measured NOz partitioning for
the three largest components of NOz. On the right chart, “other” refers
to the missing NOz which we attribute to the unmeasured organic nitrates.
Reactions (R1) and (R4)–(R7) result in formation of HNO3, which makes up
the bulk of NOz in 2013. Furthermore, the integrated nighttime loss toward
nitric acid is 5.9 times greater than during the day. Therefore much of the
difference in NOz between the low-ozone year of 2012 and the high-ozone
year of 2013 must be due to a large difference in nighttime N2O5
reactivity, which we analyze below.
N2O5 lifetimes
When the sinks of NO3 are small compared to those of N2O5,
and assuming an equilibrium state between NO2, NO3, and
N2O5, the ratio of the N2O5 concentration to the production
rate of NO3 equals the N2O5 lifetime
(τN2O5),
τN2O5=[N2O5]k⋅[NO2]⋅[O3],
where k is the rate coefficient for Reaction (R4) . An analysis of
the resulting lifetimes, which can be considered a measure of N2O5
reactivity, is shown with the solid lines in Fig. . Since
Eq. () assumes a steady state in NO3 and N2O5,
the relevant period when this lifetime interpretation will be most valid is
at the end of the night. However, a simple five-reaction chemical box model
including NO3 and N2O5 production and first-order loss
shows that it would take > 20 h to reach a steady state
in 2012. After the 14 h of night, we predict that the lifetime calculated
using Eq. () gives us 77 % of the actual lifetime. In 2013, the
model predicts that the system reaches 90 % of steady state in 1.8 h. The
lifetimes in 2012 are a factor of approximately 2 times longer than in 2013,
or 2.6 times if we use calculated equilibrium values. have
suggested an alternate method for lifetime analysis that explicitly takes the
time derivative of N2O5 into account to correct its lifetime for
failure to reach steady state. Figure also shows the
steady-state lifetime calculated using this method using a smooth fit
function for the N2O5 diel profile to calculate the derivative.
Since the reaction of N2O5 occurs heterogeneously via uptake onto
surfaces, the difference in lifetime between the two years could conceivably
be due to higher aerosol surface area or faster ground deposition. The
average value of the product of the NO3-N2O5 equilibrium constant,
Keq(T), and the NO2 concentration
(Keq[NO2]), equal to the predicted ratio of N2O5
to NO3, was 115 and 440 during nighttime hours in 2012 and 2013,
respectively. Late night average NO3 of 2.2 pptv agreed well with
the predicted equilibrium. Average predicted NO3 of less than
0.5 pptv in 2013 could not be accurately measured. The late night average
steady-state lifetime of NO3 in 2012 was approximately 100 s, while
in 2013 it was 13 s. Under these cold conditions, the very short NO3
lifetimes do not represent the reactivity of the NO3–N2O5
system, which is dominated by heterogeneous loss of N2O5, and we
provide them here for reference only.
Lifetimes of N2O5, calculated using the production rate of
NO3 (solid lines), the lifetime calculated using the method of
for 2012 (short dashed line; see text), and uptake to
aerosol using an uptake coefficient of γ=0.02 (dashed lines). In 2012
we expect that the calculation gives 77 % of the actual lifetime, due to
the system not reaching equilibrium at the end of the night. The McLaren
method, based on explicit inclusion of the time derivative for N2O5,
partially corrects for this effect, especially early in the night. An uptake
coefficient of γ=0.026 would bring the P(NO3) and aerosol
calculations in 2012 into agreement. The observed lifetimes from
P(NO3) include deposition, but the calculated curves do not.
Contributions to NO3 reactivity. In both years, formation of
N2O5 and consequent uptake to aerosol dominate NO3 loss, and
reactions with VOCs are primarily with alkanes. For comparison, the total
NO3 loss rate was 0.016 s-1 in 2012 and 0.118 s-1 in
2013.
Lifetimes due to aerosol can be calculated separately using measurements of
aerosol surface area and the equation for heterogeneous uptake, assuming no
limitation for gas phase diffusion (valid for small particle size and small
to moderate uptake coefficients, and consistent with conditions from both
2012 and 2013):
τN2O5=14γc¯SA-1,
where γ is the uptake coefficient, c¯ the mean molecular speed,
and SA the surface area density of the aerosol. The aerosol
surface area density was calculated from number size distributions measured
using a scanning mobility particle sizer for particles between 20 and 500 nm
geometric diameter, and a aerodynamic particle sizer for particles between
0.7 and 10.37 µm. Size distribution measurements were taken at
relative humidity < 25 %, and a hygroscopic growth factor was calculated
using measurements of ambient humidity and aerosol composition
. There are few determinations of N2O5 uptake
coefficients in winter. During winter measurements in Colorado,
determined an average γ=0.02 under similar conditions
of temperature and relative humidity, and at a site with nearly identical
latitude and elevation. Using γ=0.02, we calculate the lifetimes of
N2O5 due to aerosol uptake for 2012 and 2013, plotted as dashed
lines in Fig. . The 2012 lifetime includes a 10 %
correction from the contribution of losses due to VOCs (see below). On
average, lifetimes calculated from aerosol uptake were a factor of 4.1 higher
in 2012 than 2013, compared to the factor of 2.6 change in lifetime calculated
from the N2O5 steady state of Eq. () and the box model.
However, an uptake factor of γ=0.026 in 2012 would bring the lifetimes
calculated using these two methods into agreement. Since we did not perform
eddy covariance flux measurements, we do not know the deposition rate, and
the γ values derived from comparison to the steady-state lifetimes
thus represent an upper limit. Additionally, since the lifetime of
N2O5 is longer in 2012, the influence of deposition to the ground
surface might be greater if it were roughly constant relative to other sinks
that increased between 2012 and 2013. The change in aerosol uptake between
the two years is in part due to the higher relative humidity measured in
2013, which increased the aerosol surface area through hygroscopic growth.
The increased relative humidity in 2013 caused frequent and persistent fog. Due to the
difficulty in extrapolating a hygroscopic growth factor near saturation, data
during periods of relative humidity above 95 % have been excluded in this
analysis. Hygroscopic growth associated with the higher relative humidity
contributed a factor of approximately 1.3 to the difference in lifetime
between the two years.
One condition of Eq. () is that the major sink of NO3 is
through aerosol uptake via N2O5 instead of reactions with volatile
organic compounds (VOCs). Previous studies in regionally polluted areas have
shown that loss of NO3 and N2O5 can be dominated by
NO3–VOC reactions, N2O5 uptake, or a combination of the two
. Given the high VOC concentrations in the Uintah
Basin , we performed an analysis of NO3 reactivity to
quantify the contribution of NO3 chemistry to the lifetime of
N2O5. The loss due to VOC is simply the sum of all the
NO3–VOC rate constants (ki) times the measured VOC
concentrations
klossNO3=∑ikiVOCi.
This first-order loss rate coefficient for NO3 can be compared to the
first-order loss rate coefficient for uptake of N2O5 to aerosol by
dividing the former by the equilibrium ratio of N2O5 / NO3
. VOC measurements by proton transfer reaction mass
spectrometry and gas chromatography in 2012 provided measurements of a more
extensive VOC suite than the measurements in 2013, so VOC ratios from 2012
were used to estimate some compounds missing from 2013 measurements, as was
done by . The calculations show that with an N2O5
uptake coefficient of 0.02, NO3 losses due to reactions with VOCs
were approximately 10 times less than N2O5 uptake to aerosol in
2012, and approximately 40 times less in 2013. A lower N2O5 uptake
coefficient would increase the fraction of the NO3 and N2O5
reactivity attributable to NO3–VOC chemistry. However, the
comparisons of Fig. suggest that the average
N2O5 uptake coefficient is not appreciably smaller than 0.02.
Figure shows the relative loss rates, as well as the
breakdown of reactivity with different classes of VOCs. During both years,
reactivity with alkanes form the major part of NO3 loss to VOCs
(45–51 %). To our knowledge, this is the first instance in which alkanes
have been determined as the largest single component of NO3–VOC
reactivity in ambient air. For example, studies in other locations, such as
Houston, Texas, show that alkanes contribute approximately 1 % to ambient
NO3 reactivity . Despite their very slow rate
constants for reaction with NO3, alkanes make up an overwhelming
fraction of the measured VOC composition in the Uintah Basin, leading to an
unusually large contribution to NO3 reactivity. Isoprene and dimethyl
sulfide (DMS) are collectively labeled “biogenic” according to convention,
but due to winter conditions we anticipate no biogenic source for these
compounds. Rather, we assume both to be emissions from oil and gas
operations. For example, an anthropogenic source of isoprene may be emitted
in small quantities in vehicle exhaust , while DMS may be a
component of the reduced sulfur emissions from natural gas. In any case, the
measured concentrations of both compounds are small (2 and 0.7 pptv,
respectively, nighttime average in 2013), and their contribution to
NO3 reactivity represents the fast NO3 rate constant with
these species. It is possible that other highly reactive but unmeasured VOCs
contribute to the NO3 reactivity. For example,
report an important role for reduced sulfur species other than DMS in loss of
NO3 radicals near an oil refinery. Such measurements were unavailable
for the UBWOS studies.
Since N2O5 uptake to the ground can also affect lifetimes, one has
to consider differences in inlet height and ground composition between
different years. In 2012, N2O5 was measured from a scaffold tower at
a height of 11 m, whereas in 2013, the lack of such a tower limited the
sampling height to 4 m. To investigate a possible N2O5 gradient, we
alternately sampled from 14 and 1 m during the final weeks of the 2014
campaign, spanning the sample heights of the 2012 and 2013 inlets. In 2014,
the ground was snow-covered, and conditions generally resembled 2013 more
than 2012. The resulting lifetime calculations using NO3 production
rates (Eq. ) are shown in Fig. with black solid
and dotted lines. We measured roughly twice the N2O5 lifetime at the
high inlet as compared to the low inlet. This difference results solely from
differences in N2O5 concentrations; measurements of NO2,
O3, and aerosol surface area between 4 and 14 m did not show
significant differences at night and were assumed to be equal for the
lifetime calculation. Ground deposition of N2O5 can form an
important contribution to the lifetime , but the
year-to-year variability is a significantly larger effect than the measured
N2O5 gradient. This suggests that nighttime aerosol uptake of
N2O5 could play a major role in NOx oxidation and contributes to
keeping NOx levels similar between the three years.
The effect of inlet height on calculated lifetimes. Red and blue
lines are the same as in Fig. . Black lines are calculated
from 2014 measurements with the solid line from an inlet at 14 m and the
dashed line from an inlet at 1 m. These inlet heights span the inlets in
2012 at 11 m and 2013 at 4 m.
Sensitivity of NOx and O3 to NOx oxidation pathways
We again used the MCM box-model simulation to investigate the relative
sensitivities of nitrogen oxide loss and O3 production rates to some
of the different NOx oxidation pathways discussed above. We
increased/decreased the reaction rate constants of Reactions (R1)
(NO2+OH), (R2) (NO2+PA), and (R4) (NO2+O3) by a
factor of 2, keeping all else equal, and compared the resulting NOx and
ozone levels after the model stabilized to the base simulation results that
matched observations. The base simulation included a continuous source of
NOx, tuned to match observed levels . In the MCM, the
rate of Reaction (R6) was set empirically to match the observed N2O5
concentrations. The resulting rate was fast enough that Reaction (R4) was the
rate-limiting step in the reaction pathway Reactions (R4)–(R7), and was
therefore used to test the sensitivity of that pathway.
The effect on NOx and ozone concentration of changing the rates
of select reactions in a box-model simulation. The reaction NO2+O3
represents the nighttime reaction pathway to HNO3.
The results are shown in Fig. , with the left panel showing
the final day of the simulation, and the right panel comparing the final
day's 24 h averages. For Reactions (R1) and (R2), an increased/decreased rate
has very little effect on NOx once the model has stabilized. The nighttime
pathway has a much larger effect, however, and an doubled rate leads to a
28 % NOx reduction. Halving the rate causes a 43 % increase. During
the day, changing the rate of Reaction (R4) has no effect due to the fast
photodissociation of NO3. The response of O3 concentrations
is also shown, with the nighttime reactions having the greatest effect.
Changing PAN and HNO3 production have comparable effects on
ozone even though the effective NOx removal rates are approximately 4
times different. This may be because the OH + NO2 affects the
propagation of the HOx cycle directly with OH reacting with either
NO2 or a VOC. PAN production, on the other hand, has its
effect based on whether PA reacts with NO or NO2, which
scales as the ratio of PA loss to NO vs. loss to NO2.
Although organic nitrates are the largest photochemical pathway for nitrogen
loss, we did not perform an analogous simulation using Reaction (R3)
(NO + RO2). Since a comparable simulation involves changing all the
rate coefficients for a large number of reactions, performing these
simulations is beyond the scope of this paper. However, if we scale the
sensitivity of doubling/halving the reaction rates for organic nitrate
production to the sensitivity to daytime production of nitric acid (a factor
of 4.6), we get a change in NOx of approximately 7 % and a change in
O3 of approximately 17 %. The effect could be larger since NOx
is higher in the morning,
when the RO2 + NO rate is largest. Scaling it to PAN production
(Reaction R2) causes a change in NOx and O3 of approximately 3 and
6 %, respectively. If instead we were to scale α by a factor of 2,
the effect could be larger since there is no competition for the fate of
RO2; every RO2 reacts with NO. For example,
found that a 50 % increase in α results in a 7 ppb
decrease in ozone (at an ozone concentration of ∼ 60 ppbv), and they
estimate a 25 ppbv effect (at ∼ 140 ppbv ozone) for conditions with
higher J values and slower mixing. Thus, although organic nitrate
production should have the largest influence of the photochemical NOx loss
mechanisms on both NOx and O3, we anticipate that it still has a
smaller effect on NOx loss pathways than the nighttime chemistry in this
winter environment.
Winter O3 should be more sensitive to N2O5 chemistry because
it is predominant during winter conditions, with low primary radical
generation during daytime and longer duration of darkness. The majority of
polluted winter conditions do not produce O3 efficiently due to low
photochemical radical production rates. These systems are typically NOx-saturated . The result of
N2O5 chemistry in most of these situations would be to increase
O3 photochemistry during the daytime by reducing the NOx levels
overnight. In summertime urban environments, N2O5 chemistry should
have an effect, but it would be smaller because it will consume a smaller
fraction of reactive nitrogen compared especially to Reaction (R1) in more
typical summertime ozone photochemical systems. Its effect on O3 will
be highly sensitive to the O3–NOx sensitivity in any given region,
and would be difficult to generalize.
The influence of ClNO2 production from N2O5 is not
explicitly considered here, and was determined to be a small effect on NOx
due to its low yield. However, it may be an important effect on O3
production in other regions during both summer and winter, especially if
ClNO2 photolysis is a larger contribution to photochemical radicals
than was determined for the UBWOS 2013 study.