Introduction
The aerosol optical extinction coefficient (βext) represents the
attenuation of light due to aerosol absorption and scattering of solar
radiation. For a population of aerosol particles, βext depends
on aerosol size, composition, particle number concentration, shape, and
morphology (Bohren and Huffman, 1998). Atmospheric aerosols have important
implications on climate. They modify the Earth's radiative energy budget
directly through absorption and scattering of light and indirectly through
changing cloud characteristics (e.g., cloud droplet number concentration,
cloud droplet size, cloud reflectivity, or lifetime) (Ramanathan et al., 2001;
Seinfeld and Pandis, 2006; Langridge et al., 2011). In addition, aerosols
with diameters between 0.1 to 1 µm are the main contributors to
visibility degradation in anthropogenically polluted areas and on regional
scales due to their direct interactions with solar radiation (Malm, 1989;
Hobbs, 2000; Ying et al., 2004). For example, it has been observed that the
important anthropogenic contributors to light scattering in the Colorado
Rocky Mountains are particulate matter from the urban emissions (Levin et
al.,
2009). The Denver metropolitan area has also experienced seasonal air
pollution and visibility degradation in the past. The wintertime pollution in
Denver when trapped closer to the surface due to the low inversion layer
causes a greyish-brown cloud referenced to as the “Denver brown cloud.” The
composition of the Denver brown cloud and contribution of different chemical
species to the observed βext during the wintertime have been
investigated in the 1970s to late 1980s (Groblicki et al., 1981; Wolff et al.,
1981; Neff, 1997). These studies concluded that, among all the measured
aerosol species, elemental carbon, ammonium sulfate, and ammonium nitrate
contributed the most (37.7, 20.2, and 17.2 %, respectively) to wintertime
optical extinction in the visible range. Previous measurements of summertime
particle composition in the Colorado Front Range were conducted during the
northern Front Range Air Quality Study (NFRAQS) between 17 July and
31 August 1996 at several urban and rural sites. The major components of
PM2.5 mass were identified to be carbonaceous and inorganic aerosols,
with carbonaceous aerosols contributing about 46 % of PM2.5 mass.
The 24 h average measurements of PM2.5 organic carbon, nitrate, and
sulfate particles were observed to be 4.2, 0.9–1.2 µg m-3,
and 1.4–1.5, respectively (Watson et al., 1998). In response to the
wintertime haze episodes observed in the region, the State of Colorado has
implemented a visibility standard based on total optical extinction of
76 Mm-1 at 550 nm, averaged during a 4 h period when ambient relative humidity (RH) is
less than 70 % (Ely et al., 1993). Total optical extinction measurements
are provided by the Colorado Department of Public Health and Environment's
transmissometer installed in downtown Denver. The average total extinction
values of August 2001–2014, ranging from 40 to 80 Mm-1, reveal no
significant trend in summertime extinction and visibility in the region since
2001.
The meteorological influence on air quality and visibility in the Front Range
can also be important (e.g., Vu et al., 2016). Typically during the day,
easterly upslope flow transports emissions from local sources westward, while
during the night the flow reverses and downslope drainage flow through
Platte River valley sets in. Occasionally, a synoptic scale cyclone, called
the Denver cyclone, is established when drainage flow of air masses is
prevented due to propagation of a vortex that develops east of the Rocky
Mountains, contributing to transport and mixing of air masses in a cyclonic
flow pattern (Crook et al., 1990; Reddy et al., 1995).
With a twofold increase in natural-resource extraction wells since 2005 to
about 24 000 active oil and natural gas (O&G) production wells in 2012,
northeastern Colorado has experienced extensive fossil fuel production within
the past decade (Scamehorn, 2002; Pétron et al., 2014). This includes
increases in fossil fuel production from coal bed methane, tight sand and
shale natural gas, shale oil, and the associated gases. The emissions from
these processes have several environmental impacts, such as greenhouse
emissions of methane and emissions of non-methane hydrocarbons, that impair
air quality. Emissions from diesel trucks, drilling rigs, power generators,
phase separators, dehydrators, storage tanks, compressors, and pipes used in
O&G operations also contribute to the regional burden of volatile organic
compounds (VOCs), nitrogen oxides, and particulate matter (i.e., black carbon
and primary organic carbon) (Gilman et al., 2013). One of the major air
quality issues the Colorado Front Range faces is the exceedance of the 8 h
National Ambient Air Quality Standard (NAAQS) standard for ozone (75 ppbv)
during the summertime. In 2007, the Denver metropolitan area and the northern
parts of the Colorado Front Range were classified as nonattainment areas due
to the summertime elevated levels of ozone
(www.colorado.gov/cdphe/attainment). Summertime impacts of O&G
emissions on the formation of secondary pollutants and visibility reduction
in the Front Range have not been explored previously. In addition to local
point and area sources in the Front Range, biomass burning (BB) emissions from
wildfires in the region may also act as a source of aerosols, contributing to
regional haze (Park et al., 2003).
During July–August 2014, airborne measurements were conducted over the
Colorado Front Range as part of the Front Range Air Pollution and
Photochemistry Éxperiment (FRAPPÉ) to characterize the influence of
sources, photochemical processing, and transport on atmospheric gaseous and
aerosol pollutants in the area. This paper will discuss the role of local
aerosol sources in the Front Range and regional wildfires on aerosol optical
extinction in the absence of the Denver cyclone by investigating chemical and
optical properties of aerosols in different air masses.
C-130 flight tracks in the Colorado Front Range, color coded with
observed mixing ratios of (a) CO, (b) ethane, and
(c) ammonia. The yellow arrow indicates the Denver metropolitan
area. To the west of the Denver metropolitan area are the Rocky Mountain
foothills depicted by the topographic color scheme.
Measurements
FRAPPÉ field campaign
In situ measurements were conducted aboard the National Science
Foundation/National Center for Atmospheric Research (NSF/NCAR) C-130 aircraft
during 20 July–18 August 2014. Flight tracks of the C-130, color coded with
different trace gases, are presented in Fig. 1a–c. In the current analysis,
airborne data were limited to those obtained only in the boundary layer of
the Colorado Front Range (i.e., altitudes below 2500 m above sea level
(a.s.l.) east of the foothills and below 5500 m a.s.l. with easterly winds
over the foothills and the Continental Divide) to capture the influence of
local sources such as power plants, O&G, agriculture, livestock, and urban
emissions. Occasionally, when air masses with the mountain-valley circulation
patterns were sampled, data from higher altitudes (< 4000 m a.s.l.) over
the Denver metropolitan area were also considered.
Instrumentation and methodology
The NSF/NCAR C-130 aircraft carried an extensive collection of instruments
for the characterization of the diverse atmospheric pollutants in the
Colorado Front Range. The relevant instrumentations used in the current
analysis are described below. (The data produced by these instruments are
available at
http://www-air.larc.nasa.gov/cgi-bin/ArcView/discover-aq.co-2014?C130=1).
The extinction coefficient (βext) at 632 nm was measured using
a cavity attenuated phase shift particle light extinction monitor
(CAPS-PMex) (Aerodyne Research Inc., Billerica, MA). The
CAPS-PMex utilizes high-reflectivity mirrors at two ends of a
26 cm long, near-confocal cavity. Within the optical cell cavity, the highly
reflective mirrors create an effective path length of approximately 2 km.
In the particle-free sampling mode, the light-emitting diode (LED) light
output is directed to the first mirror, while a small fraction passes through
the second mirror to the photodiode detector, producing a slightly distorted
waveform of the square-wave modulated by the LED, whereas in the aerosol
sampling mode the detector detects a greater distorted waveform,
characterized by a phase shift. The vacuum photodiode, which is located
behind the second reflective mirror, detects and measures that phase shift
when the square wave becomes distorted due to interactions with sampled air
under a relatively long effective pathlength. The observed phase shift is
then related to aerosol βext as follows:
βext=(cotθ-cotθo)⋅(2πf/c),
where cotθo is the phase shift from the particle-free sampling
mode, cotθ is the phase shift during the aerosol sampling mode, f is
the frequency, and c is the speed of light. The estimated uncertainty in
βext is 10 %, while the 3σ detection limit for 1 s
data under the conditions of particle-free air encountered during FRAPPÉ was
∼ 1.5 Mm-1 (Massoli et al., 2010; Petzold et al., 2013).
Measurements of the baseline were conducted through the system's internal
filter unit regularly, at 5 min intervals. The filter period, which lasted
for 1 min, included 10 s of flush time, then 40 s of filter sampling, followed
by another 10 s of flush time. Although, for the majority (72 %) of the data,
consecutive baseline values had shifted by less than 0.5 Mm-1, baseline
values were interpolated for a more accurate estimation of optical
extinction. CAPS-PMex data, obtained at 1 Hz, were averaged to
the aerosol mass spectrometer's averaging time of 15 s. The measured
βext includes the combined effects of scattering and absorption
of light by aerosols at 632 nm; given relatively high single-scattering
albedo values of aerosols downwind of urban environments (Langridge et al.,
2012), βext is expected to be dominated by the scattering
coefficient. As discussed in Sects. 3.3 and 3.4, in urban or
biomass-burning-influenced air masses, contribution of absorption by black carbon (BC) to
βext could be more significant. It is worth mentioning that
anthropogenic gases such as nitrogen dioxide have minimal effect on the
measured βext at 632 nm since regular baseline corrections
based on sampled filtered air were applied to the data. Given the average
mixing ratio of NO2, the parameterization by Groblicki et al. (1981) for
estimating NO2 absorption at 550 nm, and the factor-of-10-smaller value
of NO2 absorption cross section at 632 nm compared to 550 nm
(Schneider et al., 1987), we estimated the average contribution of NO2
absorption at 632 nm to be ∼ 0.1 Mm-1, indicating a minor
contribution to total extinction at 632 nm.
The CAPS-PMex shared a common inlet with a compact aerosol mass
spectrometer (mAMS; Aerodyne Inc., Billerica, MA) coupled with a
time-of-flight (TOFwerk, Thun, Switzerland) detector to measure particle mass
distribution and non-refractory submicron aerosol composition (NR-PM1)
of organics, nitrate, sulfate, chloride, and ammonium. The mAMS inlet was
characterized to have a 50 % transmission of ∼ 800 nm (physical
diameter) particles. Aerosol concentrations from the mAMS were corrected for
vaporizer bounce using composition-dependent collection efficiencies
(Middlebrook et al., 2012). The estimated uncertainty for all aerosol species
is ∼ 30 % (Bahreini et al., 2009). Both instruments sampled
particles through a secondary diffuser mounted inside a NCAR HIAPER
(High-performance Instrumented Airborne Platform for Environmental Research) modular
inlet (HIMIL), mounted facing forward, under the C-130 aircraft. Given the
total flow rate within the inlet and assuming particle density of
1500 kg m-3, ambient temperature of 20 ∘C, and ambient
pressure of 70 KPa, 2 µm particles were estimated to be
transmitted by 50%, making the inlet nominally a PM2 inlet. Residence
time in the inlet was estimated to be ∼ 0.56 s. Ambient aerosol size
distributions were measured aboard the C-130 by a Passive Cavity Aerosol
Spectrometer Probe (PCASP). Estimated extinction values using Mie
calculations with a nominal refractive index of 1.5 and the measured PCASP
size distributions indicated that particles smaller than 800 nm captured
> 92 % of PM2 extinction values, confirming that the majority of
the extinction signal originated from aerosols in the size range of the mAMS.
We note that the calculated extinction coefficients were not highly sensitive
to the choice of refractive index; only a 4 % decrease in the slope of
scattering from PM0.8 vs. PM2 was observed by increasing the
refractive index from 1.48 to 1.52.
Based on the ambient RH and temperature and the
temperature within the CAPS-PMex extinction cell, and assuming that
aerosols had equilibrated to the conditions within the measurement cell, the
CAPS-PMex measurements during the flights discussed here represent
extinction values at an average RH of 20 ± 7 % (range of
15–30 %). Additionally, βext values were normalized for
standard temperature and pressure (273 K and 1 atm) conditions.
The relationship between the primarily emitted nitrogen oxides (NOx) and
the higher oxidized species of nitrogen captures the transformation of
NOx in the atmosphere upon aging (Kleinman et al., 2007; Langridge et
al., 2012). Thus, measurements of nitric oxide (NO), nitrogen dioxide
(NO2), nitric acid (HNO3), particulate phase nitrate
(NO3-), alkyl nitrates (ANs), peroxy acetyl nitrate (PAN), and peroxy
propionyl nitrate (PPN) were used to calculate the ratio of primary nitrogen
oxides (NOx = NO + NO2) to NOy(NOy
= NOx + HNO3 + NO3- + ANs + PAN + PPN)
in order to track the extent of photochemical aging in an air mass with
nonzero emissions of NOx (Kleinman et al., 2007; DeCarlo et al., 2010).
A ratio that yields a value close to 1 represents air masses that are
relatively fresh, whereas a ratio closer to 0 represents more-aged air
masses. NO and NO2 were measured using the NCAR two-channel
chemiluminescence instrument (Ridley and Grahek, 1990). A chemical ionization
mass spectrometer (CIMS) coupled with a quadrupole detector was operated to
measure HNO3, using SF5- as the reagent ion (Huey et al., 1998;
Huey, 2007). ANs were measured using thermal-dissociation laser-induced
fluorescence (TD-LIF) (Thornton et al., 2000; Day et al., 2002). PAN and PPN
species were measured using the CIMS Instrument by Georgia Tech and NCAR (PAN-CIGAR; Slusher et al., 2004;
Zheng et al., 2011), with I- as the reagent ion.
The impacts of different pollution sources on sampled air masses were
characterized by considering several auxiliary gas-phase tracers. Carbon
monoxide, the tracer for combustion emissions, was measured by
vacuum ultra-violet fluorescence with the estimated uncertainty of 3 % (Gerbig et al.,
1999). Ethane (C2H6), used to identify influence of O&G
emissions, was measured using the Compact Atmospheric Multispecies
Spectrometer (CAMS), employing infrared spectrometry (Richter et al., 2015).
The Aerodyne dual quantum cascade trace gas monitor for ammonia (NH3)
equipped with a mid-infrared absorption spectrometer was used to identify
emissions of agriculture and livestock operations (Ellis et al., 2010). The
influence of biomass burning was identified using the measurements of
hydrogen cyanide from the NCAR Trace Organic Gas Analyzer (TOGA), a fast gas
chromatograph coupled with a quadrupole mass spectrometer (GC-MS) set to
selected ion-monitoring mode for quantification (Apel et al., 2015) and
acetonitrile (CH3CN) by proton-transfer reaction mass spectrometry (PTR-MS), a
high-sensitivity instrument with fast time response that employs a quadrupole mass
spectrometer to measure volatile organic compounds (de Gouw and Warneke,
2007; Karl et al., 2009). A Passive Cavity Aerosol Spectrometer Probe (PCASP)
was available as the only instrument to measure ambient aerosol size
distributions in the size range of 0.1–3 µm (Rosenberg et al.,
2012).
Results and discussion
Urban aerosol optical extinction characterization in different
photochemical aging regimes
NOx / NOy ratios were observed to be highest over the city,
representing freshly emitted plumes from vehicular traffic (Fig. S1 in the Supplement). Away
from the city center, NOx / NOy values decrease, representing
the relative evolution of fresh air masses containing NOx. Figure 2
shows the scatterplot of βext vs. CO, color coded with the
NOx / NOy ratio, on 2–3, 7–8, 15–16, and 18 August (i.e.,
excluding days with the influences of the Denver cyclone and biomass burning
events). Measurements here focused on air masses impacted by urban sources
only, as defined by enhancement of CO over the background (105 ppbv, as
defined by the mode in the frequency distribution of CO in the Front Range
boundary layer) while ΔC2H6 / ΔCO < 20 pptv ppbv-1 (Warneke et al., 2007; Borbon et al., 2013).
The extinction enhancement ratios Δβext / ΔCO
in two aging regimes, categorized by NOx / NOy ratio values,
were analyzed by weighted linear orthogonal distance regression (ODR) fits,
with the slopes representing the enhancement ratios. In obtaining these fits,
weights represented standard deviations equal to the uncertainties in CO
(3 %) and βext (10 %). Uncertainties in the slope
values of ODR fits throughout the manuscript represent the estimated
propagated uncertainties, in this case the square root of the quadratic sum
of the relative uncertainties in the ODR fit (1σ), CO mixing ratio,
and βext coefficient. NOx / NOy values of
< 0.5 and > 0.5 represent relatively aged and fresh
NOx-containing air masses, respectively. Different trends in Δβext / ΔCO were seen in the two aging regimes,
with a lower value of 0.13 ± 0.014 Mm-1 ppbv-1 observed in
the fresh air masses. On the other hand, the relatively aged air masses
showed a higher Δβext / ΔCO value of
0.20 ± 0.025 Mm-1 ppbv-1, indicating about a 54 %
increase in the extinction enhancement ratio due to photochemical aging. The
correlation coefficient r values were 0.92 and 0.85 for relatively fresh
and aged air masses, respectively. The most dominant component of the
non-refractory aerosols in urban plumes was organic aerosol (OA), with a contribution
of 74 % of NR-PM1 mass. The high OA contribution combined with the observed
significant increase in the enhancement ratio of OA to CO with aging (from
0.021 ± 0.009 to 0.11 ± 0.01 µg m-3 ppbv-1) suggests that the bulk
of the aged urban aerosol mass during the daytime in the Front Range was secondary organic aerosol (SOA).
Although carrying out positive matrix factorization analysis on the measured
OA mass spectra during FRAPPÉ is beyond the scope of this paper, such
analysis in the future would be conducive to confirming the large
contribution of SOA to the measured OA. Since ΔNO3- / ΔCO and ΔSO42- / ΔCO
enhancement ratios did not increase with photochemical aging and demonstrated
poor overall correlation coefficients (r < 0.35 for ΔNO3- / ΔCO and r < 0.29 for ΔSO42- / ΔCO), the increase in the enhancement ratio of
the aerosol optical extinction coefficient with CO was likely also driven by SOA
formation.
Orthogonal distance linear regression fits to extinction (Mm-1)
vs. CO (ppbv) for fresh (blue fit line) and aged air masses (red fit
line).
Impacts of source and aerosol composition on aerosol optical
extinction
Analysis of the average composition of NR-PM1 in the Northern Front
Range, in the absence of the Denver cyclone, revealed significantly higher
concentrations of organic aerosols relative to inorganic anions in the urban
and urban + O&G-influenced air masses, with a fractional contribution
of ∼ 74 % (Fig. 3). On average, similar concentrations of
non-refractory aerosol sulfate and chloride were observed in the different
air masses, while concentration of nitrate aerosols increased by a factor of
∼ 2–3 in agriculturally influenced air masses compared to the other
air mass types with the exception of urban + O&G air masses.
Average chemical composition (µg m-3) of
non-refractory aerosols for different air mass sources.
Aerosol optical extinction values under the influence of different sources
were further analyzed using auxiliary gas-phase data. As mentioned in
Sect. 3.1, urban emissions were classified by enhancement of CO over the
background (105 ppbv, as defined by the mode in the frequency distribution
of CO in the Front Range boundary layer) while ΔC2H6 / ΔCO < 20 pptv ppbv-1.
O&G and agricultural emissions were classified using enhancements of C2H6
over 2500 pptv and those of ammonia over 5 ppbv, respectively,
while all other tracers were at background level. A fourth air mass classification used
in this analysis, urban + O&G, was based on air masses where both
urban and O&G classifications were satisfied. The background mixing ratios
for each gas tracer were determined by the mode of the frequency distribution
of the tracer's mixing ratio observed in each flight. The impacts of sources
and aerosol composition on extinction were explored by considering
correlation coefficients of linear least-squares regression fits to the
scatterplots of aerosol extinction vs. the mass concentration of the three
dominant aerosol species (OA, nitrate aerosols, and sulfate aerosols) in
urban, O&G-, urban + O&G-, and agriculturally influenced air masses.
Figure 4 shows the correlation coefficient (r) values of extinction vs.
aerosol species mass concentration, in different air mass types as
characterized above. The scatterplots of βext vs. OA under
conditions of urban, O&G, and urban + O&G air masses presented correlation
coefficients of r = 0.46, 0.72, and 0.46, respectively. This observation
suggests that O&G emissions are important for organic aerosol contribution
to βext. On the other hand, in urban plumes, the correlation
between βext and OA was lower than in O&G plumes, while, as
demonstrated in Fig. 2, βext and CO were strongly correlated
under conditions of both fresh and aged air masses. These observations suggest that species
other than OA, e.g., black carbon, that are co-emitted with CO are also
important in driving βext in urban air masses. The
correlation between βext vs. OA was weakest in plumes with
agricultural emissions (r=0.085), suggesting OA had little impact on
βext in these plumes. The correlation coefficients for
βext vs. aerosol nitrate mass were strongest under the
influence of O&G, urban + O&G, and agriculture and livestock
emissions (r=0.75, 0.75, and 0.90, respectively) and weakest in the urban
plumes (r=0.18). Aerosol nitrate formation depends on ambient conditions
(temperature and relative humidity), relative mixing ratios of nitric acid
and ammonia, and aerosol composition and pH (Seinfeld and Pandis,
2006; Weber et al., 2016). With uniform concentrations of sulfate aerosol and
small contribution of chloride and dust components to the Front Range fine
aerosol mass, variability in aerosol pH was not expected to be high.
Furthermore, there was no specific trend in temperature or relative humidity
in different plume types. On the other hand, mixing ratios of ammonia were
observed to be variable in the different air masses, with average values of
1.41 ± 1.2, 2.75 ± 1.88, 8.21 ± 2.06, and
5.47 ± 1.81 ppbv in urban, O&G, agriculture, and urban + O&G
plumes, respectively. These observations suggest that ammonia emissions that
are colocated with O&G-related activities in the Front Range play a
significant role in controlling βext in these air masses by
enhancing the partitioning of nitric acid to the condensed phase. In fact,
the average aerosol inorganic nitrate fraction over total inorganic nitrate
(aerosol nitrate / [HNO3 + aerosol nitrate]) in agriculture
and O&G plumes were 0.25 ± 0.09 and 0.11 ± 0.10, respectively,
compared to 0.070 ± 0.071 in urban plumes.
βext was poorly correlated with sulfate aerosols in the region
under the influence of all sources (r=0.30, 0.37, 0.07, 0.23 for urban,
O&G, agriculture, and urban + O&G, respectively), suggesting a low
impact of sulfate aerosol and its precursors on βext in the
region.
Correlations coefficients of βext vs. aerosol
composition for urban, O&G, agriculture, and urban + O&G emissions.
Due to the higher hygroscopicity of inorganic salts compared to organics,
contribution of sulfate and nitrate aerosols to the ambient
βext could be higher than what is discussed above. However,
under the average ambient conditions encountered during FRAPPÉ (average
RH ∼ 44 ± 17 %), the increase in ambient βext
due to aerosol hygroscopicity is not expected to be significant
(∼ 20 %) given the high organic fraction of 64–74 % in urban,
O&G-, or urban + O&G-influenced plumes (Massoli et al., 2009). In
agriculturally influenced plumes, the influence of nitrate aerosol on ambient
βext will be greater because of the lower organic fraction and
higher nitrate mass in these plumes, re-emphasizing the role of nitrate
aerosol in βext for such emissions.
Mass extinction efficiencies (MEEs) under (a) urban,
(b) O&G, (c) agriculture, and
(d) urban+O&G influence.
Mass extinction efficiency
Mass extinction efficiency (MEE) is a function of the diameter of the
particle, wavelength of attenuated light, and aerosol refractive index
(Seinfeld and Pandis, 2006). To further asses the impacts of aerosol sources
on βext, MEE values, i.e., the ratio of the observed
βext to NR-PM1 mass, in different air masses were
estimated. For this analysis, MEE values were determined as the slope of the
weighted linear ODR fits of βext against NR-PM1 mass, with
weights representing standard deviations equal to the uncertainties in
NR-PM1 mass (30 %) and βext (10 %). As indicated
in Fig. 5a–d, MEE values under the urban, O&G, agriculture, and
urban + O&G influence were
∼ 1.51 ± 0.49 m2 g-1 (r=0.40),
1.62 ± 0.51 m2 g-1 (r=0.79),
2.27 ± 0.83 m2 g-1 (r=0.83), and
2.14 ± 0.68 m2 g-1 (r=0.73), respectively. The highest
average MEE value was observed in agricultural plumes although, considering
the uncertainties in the fitted slopes, the MEE values were not significantly
different. The overall MEE value in the Front Range, i.e., MEE observed for
aerosols in all air mass types but in the absence of biomass burning, was
2.24 ± 0.71 m2 g-1 (r=0.80). Based on the values of the
intercepts of the ODR fits in Fig. 5, it appears that at background levels of
NR-PM1 mass there was a background extinction value of
∼ 2 Mm-1 in all, except agricultural, plumes. This observation
could be explained by optical extinction due to the presence of refractory
aerosol species, such as black carbon or dust, which are not accounted for in
NR-PM1 mass. A high degree of correlation between βext and
CO (Fig. 2) in urban plumes and low average concentrations of some of the
dust components (e.g., calcium and magnesium) throughout the region support
the non-negligible contribution of BC to βext in the Front
Range.
As seen in Fig. S2, different aerosol mass distributions were observed under
different air mass types. For the mass distribution analysis, dva
(vacuum aerodynamic diameter) was converted to dp (physical
diameter) by dividing dva by the overall mass-weighted effective
density (ρ), assuming ρ=1.25 g cm-3 for OA, assuming ρ=1.75 g cm-3 for ammonium sulfate and ammonium nitrate, and assuming
that particles sampled by the mAMS were internally mixed (Jayne et al., 2000;
Seinfeld and Pandis, 2006). Next we examine the similarity of MEE values
observed in the Colorado Front Range to previous measurements. MEE is the sum
of the mass absorption and scattering efficiencies (MAE and MSE,
respectively), which both depend on particle size, refractive index, and
wavelength of light (Seinfeld and Pandis, 2006). For typical urban air
masses, mass distributions were dominated by organic aerosols in the size
range of dp=150–500 nm (Fig. S2a). This is consistent with
previous observations for urban aerosol volume distributions with modes at
the size range of dp∼200–500 nm (Seinfeld and Pandis, 2006).
In O&G air masses (Fig. S2b), individual mass distributions were
broader, with modes for all species shifted to larger sizes (dp∼200–550 nm). In agriculturally influenced air masses nitrate aerosols
presented a significant mode in the size range of dp∼250–400 nm, while OA species were concentrated on smaller sizes
(dp∼100–200 nm; Fig. S2c). The mass distributions in urban
+ O&G plumes were more variable. Occasionally, the distribution was
dominated by OA in the smaller size range (∼ 90–110 nm), but it also
included contributions from sulfates and nitrates in the larger size range
(∼ 225–275 and ∼ 430–550 nm) (Fig. S2d), while other times the
mass distribution had significantly higher contribution from OA in the size
range of ∼ 225–350 nm, showing a clear shift and OA growth to larger
sizes (Fig. S2e).
Keeping in mind that in the presence of absorbing species, MEE is higher than
MSE, in the absence of estimates of MEE in other
regions we present estimates of MSE from previous studies for comparison
with the current MEE estimates in the Front Range. PM2.5 scattering
efficiencies at 550 nm in several ground-based studies at urban
commercial/residential sites have typically been measured to be in the range
of 2–3 m2 g-1 (e.g., Chow et al. 2002; Hand and Malm, 2007). In
such studies, the main aerosol sources contributing to the observed PM1
MSE were the automotive emissions and combustion processes. Although the
contribution of elemental or black carbon to PM1 mass during FRAPPÉ is
unknown, similar to these previous studies, OA contributed the most to the
NR-PM1 mass in the Front Range, and in comparison the observed average
MEE value (2.24 ± 0.71 m2 g-1) is consistent with the
previous estimates of MSE.
Impacts of biomass burning emissions on optical extinction
During 11 and 12 August, several wildfires were observed at Rocky Mountain
National Park, near Tonahutu Creek Trail, about 95 km NW of Denver and Grand
Mesa, Uncompahgre, and Gunnison National Forest. BB gas-phase markers, namely
hydrogen cyanide (HCN) and CH3CN from TOGA and PTR-MS
airborne data, respectively, were elevated in the boundary layer throughout
the flights on 11–12 August compared to non-biomass-burning days (26, 29,
and 31 July and 2–3, 7–8, 15–16, and 18 August). For example, during the BB days,
the HCN (CH3CN) mean mixing ratio in the boundary layer was
516 ± 58 pptv (201 ± 44 pptv), whereas the boundary layer mean
mixing ratio on non-BB days was 327 ± 59 pptv (148 ± 38 pptv).
Since elevated levels of HCN and CH3CN were not observed in individual
plumes but rather throughout the boundary layer on 11–12 August, a regional
influence of biomass burning emissions was suspected to be present in the
Front Range during this time. In addition, a 25 ppbv increase in CO
background values was observed on 11–12 August (Fig. S3) compared to non-BB
days. Ground-based measurements of PM2.5from Denver-La Casa
(39.78∘ N, -105.01∘ W), Denver-CAMP (39.75∘ N,
-104.99∘ W), and Denver-I25 (39.73∘ N,
-105.02∘ W) sites were analyzed to assess the regional impact of
wildfire emissions in the Front Range to PM2.5 during the BB and non-BB
days. The time series of PM2.5 mass concentrations at the sites
described above during days preceding and following the wildfires show
increases in mass concentration for PM2.5 during the days of BB
(Fig. S4). The mean PM2.5 mass concentrations during
09:00–19:00 local time at Denver, La Casa, Denver-CAMP, and Denver-I25
during non-BB days were 5.61 ± 2.02, 6.01 ± 3.52, and
7.28 ± 2.91 µg m-3, while mean mass concentrations
increased to 9.47 ± 2.05, 11.51 ± 3.04, and
14.08 ± 4.68 µg m-3, respectively, during the BB days.
As seen in Fig. 6, the average daytime PM2.5 mass concentration on BB
days increased by 75–98 % compared to the non-BB days, confirming the
regional influence of wildfires on the Front Range aerosol loadings.
Daily (09:00–19:00 local time) average PM2.5 mass
concentration for 3 monitoring sites for (a) non-biomass-burning
days of 26, 29, and 31 July and 2, 3, 7, 8, 15, 16, and 18 August and
(b) biomass burning days of 11 and 12 August. The whisker top,
whisker bottom, box top and box bottom represent the 90th, 10th, 75th, and
25th percentiles, respectively.
In addition to scattering of light by smoke particles, BB emissions of BC
and brown carbon (BrC) can lead to significant absorption of the
solar radiation in the visible and UV region; at 632 nm absorption by BrC is
minimal (Lack et al., 2012; May et al., 2014). MEE values were analyzed for
days with and without the BB influence, using weighted linear ODR fit
analysis, as explained previously. As seen in Fig. 7, average MEE on
11–12 August was ∼ 63 % greater compared to days without the
influence of BB (3.65 ± 1.16 m2 g-1 vs.
2.24 ± 0.71 m2 g-1). Additionally, during 11–12 August,
the background value of airborne βext was higher at
4.00 ± 0.71 Mm-1 compared to 0.25 ± 0.11 Mm-1 on days
without the BB influence, suggesting the additional contribution to
βext from the wildfires. Although the AMS does not detect
refractory materials such as BC due to the relatively low temperature of its
vaporizer (600 ∘C), it is likely that on 11–12 August BC emissions
from the fires had resulted in elevated extinction values on a regional
scale, resulting in higher MEE. The observed increase in MEE on 11–12 August
suggests that regional BB emissions have at least a comparable impact on
aerosol optical extinction and visibility in the Front Range relative to the
local sources.
Orthogonal distance linear regression fits to extinction (Mm-1)
vs. total NR-PM1 mass (µg m-3). Data points are color
coded with the average HCN mixing ratio for non-biomass-burning and biomass
burning days.