Introduction
Anthropogenic NOx (NO and NO2) emissions are
oxidized to nitrate in the atmosphere in the form of gaseous, wet or
particulate forms, HNO3 being one of the main precursors of
pNO3-. All these species may have detrimental effects on
human health and aquatic and terrestrial ecosystems through inhalation,
acidification and excess nitrogen deposition. In addition, aerosols may play
a significant role in regional climate dynamics as they interact with clouds
and solar radiation (e.g., IPCC, 2013). For these reasons, understanding the
chemical processes controlling the transport and fate of atmospheric reactive
N is required to help develop effective emission reduction strategies and
drive climate models (in the present article, we use nitrates to
collectively refer to pNO3, HNO3 and
wNO3).
Triple oxygen isotopes (δ18O and Δ17O) have
been used to decipher atmospheric oxidation pathways of NOx
leading to ambient nitrate. Michalski et al. (2003) performed the first
measurement of Δ17O values in atmospheric nitrate. The
authors found nitrate highly enriched in 18O and 17O,
likely due to the transfer of anomalous oxygen atoms from ozone
(O3) via the NOx–ozone photochemical cycle and
oxidation to nitrate. During its formation, O3 inherits abnormally
high δ18O and Δ17O values through mass
independent fractionation. The specific Δ17O departure from
the terrestrial mass-dependent fractionation line, named the 17O
anomaly, is often expressed as Δ17O=Δ17O-0.517×δ18O (Thiemens, 1999). Further investigations
suggested that the δ18O and Δ17O values of
wNO3- and pNO3- reflect several reactions
taking place after the atmospheric emission of NOx, i.e.,
atmospheric oxidation pathways transforming NOx into
secondary products (Hastings et al., 2003; Michalski et al., 2003, 2004;
Morin et al., 2007; Savarino et al., 2007; Alexander et al., 2009). Seasonal
δ18O differences in wNO3- samples (less
variable and lower values during summer) have been interpreted to be due to
changes in these chemical pathways (Hastings et al., 2003). Modelling and
validation based on sparse existing data provide hope regarding a global
understanding of atmospheric nitrate (Alexander et al., 2009); however,
further measurements need to be done on the ground, particularly at
mid-latitudes.
Additional studies dealing with triple oxygen isotope characterizations have
addressed questions of methodology (Kaiser et al., 2007; Smirnoff et al.,
2012), transfer of the ozone 17O anomaly to atmospheric nitrate
(Liang and Yung, 2007; Savarino et al., 2008; Michalski et al., 2014) or
sources and chemical pathways of high (Arctic) and low (Taiwan) latitude
nitrate (Morin et al., 2008; Guha et al., 2017, respectively). Triple oxygen
isotope characterizations of field NO3- samples are not yet
widespread. Also rare are the nitrate δ18O and Δ17O values of field samples downwind from
NOx-emitting sources at mid-latitudes (Kendall et al., 2007;
Proemse et al., 2013). The few existing studies have chiefly characterized
wNO3- or the sum of pNO3- and HNO3
(Michalski et al., 2004; Morin et al., 2007, 2008, 2009; Alexander et al.,
2009; Proemse et al., 2012; Guha et al., 2017), and suggested these
indicators would be useful to trace atmospheric nitrate in water (Kendall et
al., 2007; Tsunogai et al., 2010; Dahal and Hastings, 2016), or to apportion
the contribution of anthropogenic emissions to regional atmospheric nitrate
loads (Proemse et al., 2013).
In the past, due to sampling challenges, HNO3 and
pNO3- were generally collected together (without
differentiation). Therefore, no studies have separately and simultaneously
collected and analyzed the HNO3 and pNO3-δ18O and Δ17O values, and discussed these
isotopic characteristics of nitrate collected downwind of anthropogenic
emitters. While HNO3 and pNO3- can be in
equilibrium (e.g., if pNO3- is in the form of solid
NH4NO3), this is not always the case, for example, if nitrate is
bonded to calcium or dissolved in liquid water on a wet particle (see
Sect. 3.3). They have different lifetimes with respect to wet scavenging
(Cheng and Zhang, 2017) and dry deposition velocities (Zhang et al., 2009),
and may differ in their formation pathways as well. Therefore, investigating
the mass-independent and mass-dependent oxygen fractionations in nitrates collected
separately may help to identify their respective formation and loss
pathways, and provide additional constraints on processes controlling their
distribution.
Here we have characterized nitrate collected downwind of five emission
sources in central and southern Alberta, Canada, namely (1) coal-fired power
plants (CFPPs), (2) city traffic, (3) chemical industries and metal refining,
(4) fertilizer plant and oil refinery and (5) gas compressors plus cattle
and swine feedlots. To this end, we employed wind-sector-based active
samplers to collect HNO3 and pNO3- as well as
wNO3- downwind of the source types. The objective of this
work was to assess the atmospheric NOx reaction pathways and
determine processes responsible for the distribution of HNO3, w- and pNO3- in a mid-latitudinal region.
Methodology
Regional context
While national reported NOx emissions in Canada declined steadily from
2000 to 2015, emissions in the province of Alberta have remained relatively
constant since 2004 (Environment and Climate Change Canada, 2016).
Pioneering work was accomplished ,measuring nitrate on emitted PM2.5
(particulate matter less than 2.5 µm) and in bulk and throughfall
precipitation samples (wet and some dry deposition on ion exchange resin
collectors) collected at or downwind of the Athabasca oil sands mining
operations in northern Alberta (Proemse et al., 2012,
2013). However, the Edmonton area in central Alberta, known to generate the
highest NOx emissions in Canada, and the area of southern Alberta,
characterized by dense gas compressor station and agricultural emissions,
have never been investigated.
Aerial images showing sampling sites (green triangles) in central
and southern Alberta (a), and in the greater Edmonton area (b), along with
emissions of NOx as tonnes of NO2 reported in the National
Pollutant Release Inventory for 2013 (Environment and Climate Change
Canada, 2018b).
This research project investigated nitrates (pNO3-,
HNO3 and wNO3-) from two main emission source
areas: the Genesee and Edmonton areas of central Alberta, and the Vauxhall
area of southern Alberta (Fig. 1a). These areas experience a continental
climate, but the mean annual temperature at Vauxhall is slightly higher
(5.6 ∘C) and total annual precipitation lower (320 mm) than in
central Alberta (3.9 ∘C; 537 mm; Fig. S1 in the Supplement).
Fall is generally the wettest season and winter the driest. The sampling sites
were at altitudes between 645 and 820 m (above sea level), and in
continental regions devoid of the influence of marine air masses (negligible
halogen oxides).
The rural Vauxhall area was selected for collecting nitrates emitted from
multiple small gas compressor stations scattered throughout southern Alberta
and reduced N from cattle and swine feedlots. The other anthropogenic
emissions are from three sites in central Alberta (Fig. 1b): CFPPs at the Genesee site, 55 km southwest of Edmonton;
traffic-dominated emissions at Terrace Heights, a residential area near
downtown Edmonton; and an industrial area in Fort Saskatchewan, northeast of
Edmonton, where sampling two different wind sectors allowed separating
different industries. In Fort Saskatchewan, sampling in the northwest sector
targeted emissions from a mixture of sources of which the largest were a
chemical plant and metal refinery (referred to as chemical plus metal
industries; distance to sources of 3 to 7 km), while the north sector point
emissions were dominated by a fertilizer plant and an oil refinery (referred
to as fertilizers plus oil; distance to sources from 9 to 14 km). The
NOx emissions reported in the National Pollutant Release
Inventory (Environment and Climate Change Canada, 2018b) for 2013 from all
Alberta sources are also shown in Fig. 1.
Sampling protocols
Collection of nitrate samples took place between 30 September 2010 and 20 January 2014. Active air sampling was carried out using a modified version
of Environment Canada's CAPMoN (Canadian Air and Precipitation Monitoring
Network) sampling protocol, which is described in detail elsewhere
(Sirois and Fricke, 1992). Precipitation sampling also followed
CAPMoN wet-only protocols as described in the literature (Sirois and Vet,
1999). A conditional sampling method was employed to maximize the
collection of nitrogen compounds from the anthropogenic sources, in which
the sampling pumps and precipitation collector were activated when the site
wind vane registered winds faster than 0.55 m s-1 (2 km h-1) from the direction
of the targeted sources. The CAPMoN sampling system was installed and
operated at different sites, each at varying distances from the targeted
point (<1 to 35 km) and diffuse sources (3 to >125 km;
Table 1). Back trajectories run using the HYSPLIT model (Stein et al.,
2015; Rolph, 2017) for every hour of sampling verified that the conditional
sampling approach collected air masses that had primarily passed over or
near the targeted source (i.e., there was no landscape feature that
decoupled wind direction from back trajectories; see sample plot of back
trajectories from Genesee in Fig. S2).
Settings and conditions for wind-sector-based simultaneous sampling
of atmospheric nitrates.
Site
Distance
Sector
Sampling
(coordinates)
Sources
km (mean)
direction; opening
period;
n
Avg T (∘C)
Context
Genesee (114.14∘ W, 53.31∘ N)
Coal-fired power plants
7–35
NW, 35∘
30/09/2010–21/06/2011
6
11.7, 12.2,5.5, -9.8, -0.9, 12.2
3 plants
Vauxhall (112.11∘ W, 50.06∘ N)
Gas compres-sors and cattleand swinefeedlots
12–125+; 7.5–45+
W, 65∘
25/10/2011–13/12/2011
3
2.6, 0.7, -3.5
65+ compressors;200+ feedlots
Terrace Heights(113.44∘ W, 53.54∘ N)
Urban traffic
< 1–15 (4)
W, 150∘
24/07/2012–25/10/2012
4
20.3, 15.6,7.9, -1.8
Park in residentialarea, 3.5 km east of downtown core
Fort Saskatchewan(113.14∘ W, 53.72∘ N)
Chemicalindustries andmetal refining
3–7 (4)
NW, 88∘
12/04/2013–06/09/2013
4
4.3, 15.7,16.3, 17.7
Chemical plantand metal refi-nery largest NOxsources; fertilizerplant largest NH3source
Fort Saskatchewan(113.14∘ W, 53.72∘ N)
Fertilizer plantand oil refinery
9–14 (11)
N, 27∘
20/09/2013–20/01/2014
1
-8.1
Fertilizer plantlargest NH3 and NOx source, oilrefinery major NOx source
n: number of sampling sessions. Avg T: average temperature during each of
the consecutive sampling sessions.
Ambient air was pulled through a three-stage filter pack system to collect,
sequentially, particulate matter on a Teflon filter, gaseous nitric acid
(HNO3) on a Nylasorb nylon filter and gaseous ammonia on a citric
acid-coated Whatman 41 filter (all 47 mm). The Teflon–nylon filter method
for pNO3- and HNO3 has been extensively compared
and evaluated, and is currently used by national monitoring networks
targeting regional background sites, CAPMoN in Canada and CASTNet (Clean Air
Status and Trends Network) in the United States. Previous testing showed
negligible collection of HNO3 on the Teflon filter, <3 %
breakthrough of HNO3 from the nylon filter with loadings more than
3 times higher than reported here and blanks for pNO3-
and HNO3 of approximately 0.2 µg N per filter (Anlauf et
al., 1985, 1986). Intercomparisons with more labour-intensive methods, such as
tunable diode laser absorption spectroscopy and annular denuder–filter pack
systems, have shown evidence of some volatilization of ammonium nitrate from
the Teflon filter, leading to a negative bias in pNO3- and
positive bias in HNO3 under hot (>25 ∘C) and dry
conditions, particularly in high ambient concentrations (e.g., Appel et al.,
1981). However, other field studies have shown no significant differences in
HNO3 between filter packs and denuder and/or TDLAS systems (Anlauf
et al., 1986; Sickles Ii et al., 1990) or mixed results (Spicer et al., 1982;
Zhang et al., 2009). While those studies used short-duration sampling, a
comparison for weekly samples at a lower concentration site showed good
agreement between filter pack and denuder values for most of the study but
potential interference from HNO2 (nitrous acid) on the nylon filter
in two samples (Sickles Ii et al., 1999). Based on the conditions in Alberta,
we estimate that there is little or no volatilization of NH4NO3
for samples with mean temperatures below 5 ∘C, but there is a
possibility for nitrate loss of up to 30 % in the warmest sampling
periods.
After the first five sample periods, an experimental active sampling system
for NO2 and NOx was added downstream of the three-stage filter
pack. This system consisted of one or two custom cartridges packed with
Maxxam Analytics' resin to selectively collect NO2, and one- or
two-stage filter packs containing two identical Maxxam Analytics' impregnated
filters designed to collect NOx (mostly NO due to upstream collection
of NO2). Oxygen isotopes in NO2 and NOx were not measured
since we could not rule out oxygen isotope exchange during the extraction
process; however, concentrations meeting the quality control (QC) criteria (Savard et al., 2017) are presented for reference
in Table S1.
Here we report on oxygen isotopes in the simultaneously sampled
HNO3 and pNO3-, along with co-sampled
wNO3- in rain and snow samples. Note that precipitation
events did not occur regularly (see Fig. S1), so that the number of aqueous
samples collected was fewer than the gas and particulate samples. Both the
air and precipitation samplers were only active when the wind direction was
from the desired source sector and the wind speed was greater than
0.55 m s-1 (2 km h-1). Four identical air sampling systems
operated simultaneously at each site, with samples pooled when necessary to
provide sufficient filter loadings for isotope analysis and, when possible,
measured separately to estimate sampling precision. In contrast to the four
gas–particle sampling systems, there was a single precipitation collector
at each site, and therefore external precision was not determined for
precipitation samples. Individual sample deployment times ranged from 5 to
113 days, and total air sampling time within the wind direction sectors
ranged from 21 to 360 h (Table S1). The variable cumulative periods
reflected the frequency of the wind flow from the targeted source sectors and
the amount of time required to obtain sufficient mass loadings on the
filters.
Two or three replicate samples for most species were pooled at Genesee and
Vauxhall, the first two sampling sites, subject to the requirement that
sampled air volumes be within 15 % of each other, thereby eliminating
samples that experienced flow problems. Flow issues were primarily due to
pump failure, likely caused by cycling the pumps on and off frequently in
early samples. Therefore, for later samples the protocol was changed such
that the pumps remained on and valves were used to switch the pumps between
sampling lines and non-sampling tubing based on the wind sector. At the
sites sampled later in the Edmonton area, improvements to the laboratory
analytical procedure allowed for smaller sample amounts and eliminated the
need for sample pooling.
Analytical procedures
Nitric acid from nylon filters was extracted using 10 mL of 0.01 M solution
of NaCl. Particulate NO3 from Teflon filters was extracted in two
portions of 6 mL of ultrapure water (ELGA). To reduce possible evaporation,
filters were placed in an ultrasonic bath with ice. The extractions were
performed during 1 h and samples were left for 48 h in a fridge to
insure the complete extractions. The solutions were decanted and a small
portion (1–2 mL) was used to determine concentrations. The remaining
extracts were stored in the fridge for subsequent isotope analysis. The
blanks from both filters were treated the same way.
Concentration of nitrates in Teflon and Nylon filter extracts and in
precipitation samples were determined at the Institut national de la
recherche scientifique – Eau, Terre, Environnement (INRS-ETE). The
determinations used an automated QuikChem 8000 FIA+ analyzer (Lachat
Instruments), equipped with an ASX-260 series autosampler. The detection
limit for the method with sulfanilamide (31-107-04-1-A) was 0.03 µmol L-1 of NO3-–N. Nitrite concentrations were also measured in the
extracts. Nitrite concentrations above the detection limit (1.1 µmol L-1
of NO2-–N) were found in a handful of samples at Terrace Heights.
These samples were excluded from the reported data.
We characterized the Δ17O, δ18O and δ15N ratios of HNO3, wNO3- and
pNO3-, along with the δ15N values of
NH3, wNH4, pNH4 and NOx
(all N isotopic results are in Savard et al., 2017). The present article
deals solely with the δ18O and Δ17O values
obtained for the three nitrate species. We treated the samples using the
chemical conversion and thermal decomposition of N2O protocols,
providing the ability to simultaneously analyze low-concentration N- and
O-containing species (Smirnoff et al., 2012).
A notable challenge in the analysis of the filter-based atmospheric samples
is their small extraction volumes. Only 10–12 mL of extract solution was
normally available for the measurement of concentrations and isotopic
analysis. In addition, the concentrations of these low volume samples were
also low (7.1–21.4 µmol L-1 of NO3-–N). Therefore, not all
samples could be diluted to produce volumes sufficient for reduction of
NO3- to NO2 and subsequent conversion to N2O, the final
product before isotope analysis. Samples with an initial concentration below
2.3 µmol L-1 could not be treated individually and were combined to
produce volumes sufficient for analyses (same sampling period but
combination of collected parallel samples).
The preparation steps involved conversion of nitrate-containing samples into
nitrite (NO2-) using a custom-made cadmium column. The final
preparation step involved using sodium azide to ultimately produce
N2O (McIlvin and Altabet, 2005; Smirnoff et al., 2012). All
extracted N2O was analyzed using a pre-concentrator (PreCon, Thermo
Finnigan MAT) including a furnace with “gold” wires, online with an isotope
ratio mass spectrometer (Delta V Plus, Thermo Electron; Kaiser et al., 2007;
Smirnoff et al., 2012). The utilized approach allows the spectrum of δ15N, Δ17O and δ18O values from
O-bearing N-species to be determined in samples containing as little as
37.5 nmol of N (15 mL final solution). Extracts from filter blanks were
processed in the same way. The blanks from nylon filters were not detectable.
Peak heights from the blanks resulting from Teflon filters were detected and
always below 10 % of sample peaks, having a negligible effect (within the
analytical precision). The USGS-34, USGS-35 and USGS-32 nitrate reference
materials were used and processed exactly the same way as the samples, i.e.,
converted from nitrate to nitrite, then to N2O. The laboratory
analytical precision (average of replicates) determined during the present
study was 0.6 ‰ for δ18O and Δ17O
values in gaseous (n=12) and solid nitrates (n=20). For
wNO3, analytical replicates gave 0.6 ‰ and 0.5 ‰, for
δ18O (n=3) and Δ17O (n=4) values,
respectively. The Δ17O values are defined as ln(1+δ17O)-0.516×ln(1+δ18O), relative
to Vienna Standard Mean Ocean Water (VSMOW).
Isotopic reproducibility (modified median absolute deviation)
established using two to four parallel active CAPMoN sampling set-ups in seven
separate sampling periods, resulting in (n) total samples.
N compound (n)
δ18O/‰
Δ17O/‰
Teflon filters
pNO3- (19)
2
1
Nylon filters
HNO3 (18)
1
0.7
Results and interpretation
Isotopic reproducibility when using the CAPMoN filter pack sampling
system
Data obtained from at least two of the four identical CAPMoN sample
collection streams at each sampling site were used to calculate the
reproducibility of each isotopic value measured. With four or fewer samples
collected during each sampling period, a non-parametric approach was deemed
most appropriate. Therefore, for each of the 18 sampling periods a median
isotopic value was calculated, then the 2 to 4 absolute deviations from
this median were calculated (Tables 2, S1). Although there were 4
replicates in 18 periods, the pooling of simultaneously collected samples
and the QC steps described earlier reduced the total number of replicates
for each compound (Table 3). The median absolute deviation (MAD) for each
compound was then calculated from the 15–38 absolute deviations. Finally,
for comparability with the more familiar standard deviation, the MAD was
scaled using the standard 0.6745 divisor to give the modified median
absolute deviation (M.MAD), a scaled parameter that will be equal to the
standard deviation in the event of which the distribution is Gaussian
(Randles and Wolfe, 1979; Sirois and Vet, 1999). This suite of
parallel tests indicates that all measured species show coherent and
reproducible Δ17O and δ18O results, with the M.MAD
varying between 0.7 ‰ and 2 ‰ (Table 2). These estimations
encompass the precision of the entire method, including errors due to
sampling, chemical treatments and instrumental analysis.
Average oxygen isotopic ratios for NO3- sampled as gas
(HNO3),
w (precipitation) and p (particulate matter) relative to VSMOW.
Matrix
Gas
w
p
Gas
w
p
source
δ18O/‰
Δ17O/‰
Coal-fired power plants
69.7
66.1
70.7
25.1
25.4
26.6
(5)
(4)
(4)
(5)
(4)
(4)
Fertilizers plant and oil refinery
63.2
71.4
69.5
19.3
26.0
23.8
(1)
(1)
(1)
(1)
(1)
(1)
Chemical industries and metal refining
65.7
61.9
54.6
21.8
21.4
18.5
(4)
(2)
(4)
(4)
(2)
(4)
Gas compressors
65.0
–
63.1
24.5
–
26.4
(2)
(3)
(2)
(3)
City traffic
65.7
67.2
59.6
21.2
24.4
22.5
(3)
(2)
(3)
(3)
(2)
(3)
Mean
66.8
66.0
62.6
23.0
24.3
23.4
(n): number of sampling periods characterized.
A potential complication of the air sampling method can arise if there was
significant volatilization of NH4NO3 on the particle
filter into HNO3 and NH3, with subsequent collection on
the downstream gas filters. This could result in equilibrium isotopic
fractionation between the particle and gaseous components, which would become
artificially high and low, respectively, with more fractionation at higher
temperatures (summer) relative to lower temperatures (winter) when
volatilization is minimal (Keck and Wittmaack, 2005). We find the
pNO3- isotopic values (Δ17O and
δ18O) to be generally higher during winter than during
summer (see Sect. 3.4). Moreover, the pNO3-δ18O minus HNO3δ18O differences
are negative during summer (see Sect. 3.6), opposite to the expected isotopic
artefact if particulate volatilization were the dominant factor in
determining the particle–gas isotopic differences (the same was concluded for
the δ15N values in NH3 and NH4; Savard
et al., 2017). We therefore conclude that, while volatilization may occur in
the summer samples, other isotope effects must be larger in order to lead to
the observed differences. In addition, volatilization would cause
mass-dependent fractionation and would not affect the 17O anomaly;
therefore, Δ17O values remain robust indicators of chemical
pathways in this situation. Finding that the sampling protocols are adequate
for isotopic work is in agreement with a previous study using a comparable
method that found minimal fractionation for pNO3- and
HNO3 (Elliott et al., 2009).
Compilation of triple oxygen isotopic ranges obtained for
atmospheric and emitted nitrates.
δ18O/‰
Δ17O/‰
Regional context
Location
Authors
HNO3
62.4–81.7
19.3–29.0
Various contaminated sites
Alberta, Canada
This study
pNO3-
43–62
20–27
Coast, Trinidad Head
California, USA
Patris et al. (2007)
78–92
29.8–35.0
High Arctic (Alert, Ellesmere Is.)
Nunavut, Canada
Morin et al. (2007)
62–112
19–43
Coast
Antarctica
Savarino et al. (2007)
15.6–36.0
-0.2 to 1.8
Oil sands mining stacks, PM2.5
Alberta, Canada
Proemse et al. (2012)
49–86
19–27
Coast (onboard sampling)
California, USA
Vicars et al. (2013)
10.8–92.4
2.7–31.4
Mt. Lulin, partly polluted air masses
Central Taiwan
Guha et al. (2017)
48.4–83.2
13.8–30.5
Various contaminated sites
Alberta, Canada
This study
wNO3-
66.3–84.0
20.2–36.0
Shenandoah National Park
Virginia, USA
Coplen et al. (2004)
70–90
20–30
Bimonthly sampling across state
New England, USA
Kendall et al. (2007)
68–101
20.8–34.5
Rishiri Island, polluted air masses
Northern Japan
Tsunogai et al. (2010)
51.7–72.8
18.9–28.1
Highway traffic emissions
Ontario, Canada
Smirnoff et al. (2012)
35.0–80.7
15.7–32.0
Oil sands mining (with some dry dep.)
Alberta, Canada
Proemse et al. (2013)
57.4–74.4
19.2–30.1
Various contaminated sites
Alberta, Canada
This study
Undifferentiated and bulk NO3-
60–95
21–29
Polluted coastal area & Remote land
California, USA
Michalski et al. (2004)
57–79
22–32
High Arctic
Nunavut, Canada
Morin et al. (2008)
36–105
13–37
Marine boundary layer
65∘ S to 79∘ N Atlantic
Morin et al. (2009)
56.6–82.3∗
16.7–30.2∗
Various contaminated sites
Alberta, Canada
This study
Note: isotopic values rounded to the whole number are from published graphs
(except for precise O values in Morin et al., 2007). ∗ Calculated using weighted averages of HNO3 and
pNO3 isotopic results.
Concentrations and isotopic ratios of nitrates in Alberta
samples
The range of HNO3–N concentrations measured by the filters (from
0.01 to 0.15 µg m-3; average of 0.06) is slightly lower than
that of pNO3-–N (from 0.02 to 0.35 µg m-3;
average of 0.12). For context, the median concentrations at all CAPMoN sites,
which represent non-urban areas across Canada, range from 0.02 to
0.25 µg m-3 for HNO3–N and from 0.007 to
0.45 µg m-3 for pNO3-–N (Cheng and Zhang,
2017), with the higher values at sites affected by regional and transboundary
pollution. Background sites for this region are sparse, but concentrations at
Cree Lake in neighbouring Saskatchewan were the lowest in Canada reported up
to 2011 (Cheng and Zhang, 2017), and 2014–2016 measurements at Wood Buffalo
National Park on the northern Alberta border revealed similar average
concentrations of 0.02 µg m-3 of NO3-–N for both
HNO3 and pNO3- (preliminary internal data).
Therefore, the lowest concentrations in our samples approached average
background concentrations, while the highest were 20 or more times higher
than the regional background. The range of NO3-–N concentrations of
the wNO3- samples was 10.71–34.29 µmol L-1.
For comparison, volume-weighted mean annual concentrations of nitrate at the
remote CAPMoN site to the north (Snare Rapids) for 2011–2014 were
approximately 5.00 µmol L-1 of NO3-–N, while at
the most polluted site in southern Ontario (Longwoods) the volume-weighted
mean concentration was approximately 21.43 µmol L-1
(Environment and Climate Change Canada, 2018a). It should be pointed out that
precipitation ion concentrations vary significantly with precipitation
amount, so the short samples collected here are not necessarily
representative of annual volume-weighted means.
Triple O isotopic results obtained for simultaneously collected
atmospheric HNO3 (a), wNO3- (b) and
pNO3- (c), in Alberta, identified by sampling periods (cold
months – blue; warm months – red).
The average δ18O and Δ17O values of
HNO3 (gas), w- and pNO3- show no apparent
systematic ordering (Tables 3, S1 and S2), in
contrast to what was found for δ15N values in the same
samples (Savard et al., 2017). As expected, there is no systematic tendency
when looking at the samples collected from the anthropogenic sources: CFPPs
HNO3 and pNO3- have the highest
δ18O and Δ17O averages, but not the highest
delta values for wNO3- values; chemical industries show the
lowest δ18O and Δ17O averages for w- and
pNO3, but not for HNO3. Though the number of samples
was limited, wNO3-Δ17O values were roughly
correlated with the weighted average Δ17O values of
pNO3 and HNO3 in samples covering the same time
periods, consistent with scavenging of both HNO3 and
pNO3 by wet deposition. This observation indicates that the
oxygen isotopes in the three nitrate species are not predominantly source
dependent (see also Fig. S3), as previously suggested in the literature
(Michalski et al., 2003).
Considering all nitrate species, the Alberta δ18O and
Δ17O values range between +48.4 and +83.2 ‰,
and between 13.8 ‰ and 30.5 ‰, respectively (Tables 4, S1, Fig. S4).
These ranges indicate that ozone partly transferred its isotopic anomaly to
nitrates during NOx cycling and oxidation (nitrate derived
through combustion in O2 would show δ18O and
Δ17O values of 23.5 ‰ and 0 ‰, respectively). When
examining the existing δ18O and Δ17O data
for w- and pNO3- in the literature, the ranges for our
mid-latitude samples are within those previously reported (Table 4). The
worldwide compilation of documented data is broadening the
δ18O range of atmospheric NO3- previously
suggested to be between 60 ‰ and 95 ‰ (Hastings et al., 2003; Kendall
et al., 2007).
Main reactions producing atmospheric nitrates (Zel'dovich, 1946;
Lavoie et al., 1970; Erisman and Fowler, 2003; Michalski et al., 2003;
Morino et al., 2006; Morin et al., 2007; Stroud, 2008; Michalski et al.,
2014). Reactions (R1) and (R9–R12) can occur any time.
Daytime – summer
Night-time – winter
(R1) O2 + Q → O + O + Q; N2 + O → NO + N; N + O2 → NO + O
(R2) O + O2 + M → O3 ; NO + O3 → NO2 + O2
(R3) NO + RO2 → NO2 + RO
(R4) NO2 + O3 → NO3 + O2
(R5) NO2 + hv (sunlight) → NO + O
(R6) NO2 + OH + M → HNO3 + M
(R7) NO2 + NO3- ↔N2O5
(R8) N2O5 + H2O(surface) → 2HNO3 (aq)∗
(R9) HNO3(g) ↔HNO3(aq)∗ → NO3-(aq)∗ + H+(aq)
(R10) HNO3(g) + NH3(g) ↔NH4NO3(s)
(R11) HNO3(g) + CaCO3(s) → Ca(NO3)2(s) + HCO3
(R12) NO3 + HC;(CH3)2S → HNO3 + products
Q is a stable molecule of high energy; M is either O2 or N2; RO2 stands for both HO2 and alkyl peroxy.
HC stands for hydrocarbons.
∗This aqueous nitrate may be on a particle.
Previous studies that report triple isotope oxygen results in atmospheric
NO3- samples are scarce (Table 4). The HNO3 range
documented here is within the broad spectrum of pNO3- values
compiled for remote to contaminated sites. Elliott et al. (2009) reported
HNO3 oxygen results for δ18O values only, with a
range of +51.6 ‰ to +94.0 ‰ (mean of 77.4), with
simultaneously sampled pNO3-δ18O values
between +45.2 ‰ and +92.7 ‰ (mean of 75.2). Those ranges are
broader than the HNO3 and pNO3- values obtained in
the present study.
The δ18O and Δ17O trends in nitrates from cold
and warm sampling periods
The δ18O and Δ17O ranges for HNO3
identified by sampling period are narrower than those of the simultaneously
collected pNO3- (Fig. 2; Table S1), suggesting that there are
additional mechanisms affecting HNO3, or that pNO3 is
derived from different pathways with more variation in isotopic signatures.
Overall, the Δ17O and δ18O results for
HNO3, wNO3- and pNO3- clearly show
higher δ18O and Δ17O values during cold
periods relative to warm periods (Fig. 2), with the exception of
HNO3δ18O values, which were similar in cold and
warm periods. The collection of several samples lasted over periods
overlapping fall and winter and, in such cases, the results are labelled as
covering the two seasons; note that for many fall cases, the average sampling
temperatures were below 0 ∘C (Table S3). Nevertheless, plotting by
sampling period can be regarded as a general repartition of results between
warm and cold months, which show lower and higher isotopic values,
respectively, in both the w- and pNO3-.
A series of reactions listed in Table 5 summarizes the main atmospheric
processes taking place during the production of nitrates in contaminated air
masses. First, during anthropogenic combustion of fossil fuels, NOx (NO
and NO2) is produced through reactions of air N2 with atmospheric
O2 at high temperatures (Reaction R1; Table 5). Then, NOx cycles
between NO and NO2 through a series of reactions involving sunlight
(Reaction R5), O3 (Reactions R2, R4) and peroxy (HO2) or alkyl peroxy (RO2)
radicals (Reaction R3; Morin et al., 2007; Fang et al., 2011; Michalski et al.,
2014; here we use RO2 to refer collectively to HO2 and RO2).
The oxidation of NOx (specifically NO2) to
HNO3 further incorporates additional O atoms from different
oxidants (Reactions R6–R8; Table 5). Production of nitrate via Reaction (R6)
is restricted to daytime (since OH is generated through photochemistry),
whereas production through Reactions (R4), (R7) and (R8) dominates at night.
In addition, N2O5 is thermally unstable, so the contribution of
the R4-R7-R8 pathway is larger during winter than during summer.
Additionally, in the heterogeneous hydrolysis of N2O5
(Reaction R8), HNO3 is likely to be retained on the reaction
particle as pNO3- due to its hygroscopicity (Seinfeld and
Pandis, 2006). We have neglected contributions from BrO cycling due to the
location far from the coast, and from reactions of NO3 with
hydrocarbons (Reaction R12) since they are predicted to have a minimal
contribution to nitrate formation in this region (Alexander et al., 2009).
Finally, HNO3 in the gas phase can be irreversibly scavenged by wet
surfaces or precipitation (Reaction R9) and calcium carbonate on particles
(Reaction R11), and can equilibrate with solid ammonium nitrate where there
is excess ammonia available (Reaction R10).
It has been previously suggested that the δ18O and Δ17O values of w- and pNO3- formed during summer
are lower than those during winter due to a higher contribution from the
N2O5 path (Reactions R4, R7–R8) during that season (e.g.,
Hastings et al., 2003; Morin et al., 2008). As an early take on the data
identified by sampling periods, the w- and pNO3-δ18O and Δ17O data presented here follow the
same patterns for warm and cold months (Fig. 2). In contrast, the less
commonly studied HNO3 shows similar δ18O values
during warm and cold seasons, but summer Δ17O values mostly
lower than the fall–winter, fall and spring ones.
Correlation coefficients (r) of NO3- isotopic deltas with
meteorological parameters and concentration (or ratio) of co-contaminants.
In bold are the significant correlation coefficients, equal or above the 95 % significance value.
Relative
Daylight
humidity
Temperature
(fraction)
PM
SO2
O3
r
R2
r
R2
r
R2
r
r
r
R2
HNO3
δ18O
0.8
0.59
-0.4
-0.3
0.1
0.0
-0.29
n
8
15
15
13
13
13
Δ17O
0.6
-0.5
0.24
-0.4
0.4
0.3
-0.03
n
8
15
15
13
13
13
pNO3-
δ18O
0.9
0.79
-0.6
0.34
-0.6
0.35
0.1
0.5
-0.61
0.38
n
7
15
15
12
12
12
Δ17O
0.9
0.73
-0.6
0.34
-0.7
0.44
0.0
0.5
-0.47
n
7
15
15
12
12
12
Correlations with meteorological parameters and co-pollutants
The distribution and proportion of HNO3 and pNO3-
in polluted air masses can vary daily and seasonally with temperature,
relative humidity (RH) and concentration of co-contaminants (Morino et al.,
2006). For that reason, we compared the isotopic ratios of the HNO3
and pNO3 samples (n of wNO3 too low) with
meteorological and air quality parameters measured routinely at nearby
monitoring stations where available (Table S3). We found that the
pNO3- and HNO3δ18O and Δ17O values correlate with RH, with pNO3 values
showing stronger statistical links than HNO3 (Table 6). The
N2O5 hydrolysis reaction (Reaction R8) rate increases with
humidity (Kane et al., 2001), which may explain this positive correlation.
Significant inverse relationships exist between temperature and
pNO3-δ18O, pNO3-Δ17O and HNO3Δ17O. These
negative links likely arise since N2O5 is more stable under cold
conditions, leading to a higher contribution of R8. The stronger links with
pNO3- may be due to Reaction (R8) taking place on surfaces
(such as particles) with liquid water, which is likely to retain the
HNO3 as pNO3- rather than release it to the gas
phase. Therefore, in winter, R8 may contribute more to pNO3-
than to HNO3(g). Moreover, the highest δ18O and
Δ17O values for both pNO3- and
HNO3 were found for fall–winter samples collected at high RH
(76 %) and low temperature (-10 ∘C). In contrast, the lowest
pNO3- isotopic values were found for samples with similar
proportions of HNO3 and pNO3-, and sampled during
moderately humid (60–63 %) and warm (8–20 ∘C) periods. The
accompanying shift in δ18O and Δ17O
differences between pNO3 and HNO3 will help infer the
mechanisms that dominate during the cold and warm periods (Sect. 4.2).
Concentrations of oxidants, co-contaminants (e.g., SO4-
aerosols) and NOx influence the dominance and rates of the
discussed reactions (Brown et al., 2006; Michalski et al., 2014). However,
while temperature, RH and O3 are well captured within a 5 km
radius, other pollutant measurements like continuous SO2,
PM2.5 and NOx can have large gradients near sources;
therefore it is not surprising that no correlations were found with
SO2 or PM2.5 measured at sites 4–5 km away (Table 6).
Surprisingly, only the pNO3-Δ17O and
δ18O values correlated with the fraction of each sample
collected during daylight hours (i.e., between the sunrise and sunset times
on the day at the middle of each sampling period, either at Edmonton or
Lethbridge), which was expected for HNO3 as well due to the
daytime-only OH pathway. However, daylight hours do not take into account
light intensity, which can significantly influence the oxidation rate through
this pathway and consequently both the δ18O and
Δ17O values.
Comparison with high-latitude pNO3-
An interesting aspect of the Alberta pNO3- cold-period
Δ17O ranges is that they compare relatively well with the
range obtained for the Canadian Arctic (Fig. 4), during winter, when
night-time conditions and the N2O5 pathway prevail without
interruption (Morin et al., 2008; for comparison with HNO3 values
see Fig. S4). This observation supports the suggestion that the
N2O5 pathway produces around 90 % of nitrates during
mid-latitudinal cold months (Michalski et al., 2003; Sect. 4.1). The
δ18O ranges of cold months are similar in Alberta and in the
Arctic. This similarity goes against previous suggestions that at higher
latitudes, nitrate δ18O annual means should be higher than
at mid-latitudes due to local ambient conditions and atmospheric chemistry
affecting the proportions of species involved in producing nitrate (Morin et
al., 2009), namely, the sole influence of the N2O5 pathway
during the Arctic winter (Fang et al., 2011).
The Δ17O departure between the Alberta and Arctic winter
parallel lines is about 3 ‰. Such a difference is
slightly larger than the one calculated for winter NO3- at 80 and
40∘ N latitudes (about 2 ‰; Morin et al., 2008). In contrast, the warm months and summer data sets for
Alberta and the Arctic, respectively, show different isotopic ranges (Fig. 5), possibly due to the plume effects described later (Sect. 4.3).
Moreover, contrary to a previous suggestion, the winter–summer difference in
nitrate Δ17O values is similar at the mid- and high-latitudinal
sites (about 6 ‰ here, and 5 ‰ in Morin et al., 2008). This similarity is likely
coincidental as it may reflect the fact that within-plume chemistry may
lower the Δ17O values of NOx in the sampled anthropogenic
plumes in Alberta (see Sect. 4.3 for details), whereas the seasonal
departure in Arctic samples comes from the oxidation to nitrate through the
dominant OH and N2O5 pathways during summer and winter,
respectively. Finally, the Δ17O averages for the
Alberta summer and winter results approximately fit within ranges predicted
for the studied area by global modelling (Alexander et
al., 2009), suggesting that global modelling of nitrate distribution
worldwide is promising.
Isotopic differences between HNO3 and pNO3-
As far as the isotopic characteristics are concerned, an important feature to
keep in mind is that the HNO3 of central and southern Alberta has
distinct properties relative to simultaneously sampled pNO3-.
In practical terms, the relationships between the simultaneously sampled
HNO3 and pNO3- are of four types (Fig. 3):
(i) HNO3 δ18O and Δ17O are both
lower than pNO3-; (ii) HNO3 has lower
Δ17O but higher δ18O values than
pNO3-; (iii) HNO3 has higher δ18O
values and similar Δ17O ones relative to
pNO3-; and (iv) HNO3 has higher
δ18O and Δ17O values than
pNO3- (Fig. 3).
The fall–winter isotopic results belong to group (i), fall results to
groups (i), (ii) and (iii), and the spring and summer results to
groups (ii), (iii) and (iv) (Fig. 3). Elliott et al. (2009) reported
simultaneously sampled pNO3- and HNO3 in the northeastern United States with similar seasonal changes of δ18O
differences (no Δ17O measurement). The HNO3δ18O values were generally similar or lower than the
pNO3- values during winter and fall, and slightly to much higher during spring and summer, with the spring and fall
pNO3-–HNO3 relationships being roughly intermediate
between the winter and summer ones. The average δ18O
difference of pNO3- minus HNO3 reported between
winter and summer (15 ‰) by Elliott et al. (2009) agrees with the
difference for fall–winter and summer obtained here (12 ‰).
Line-connected δ18O and Δ17O
values for simultaneously collected HNO3 (empty symbols) and
pNO3- (solid symbols) from cold (blue) and warm (red)
sampling periods.
The marked shifts in isotopic differences between the separately analyzed
HNO3 and pNO3- reported here likely reflect
changes in the dominant reactions and processes leading to the production of
the two nitrates (see Sect. 4.2). Analyzing them separately provides
additional granularity that may be used to elucidate further details of the
production and loss of nitrate species downwind from a NOx
source.
Isotopic results for pNO3- identified by sampling
periods (solid lines), compared with summer and winter trends obtained for
Arctic sites (dashed lines; derived from ln (1+
δ) in Morin et al., 2008).
Discussion
Estimation of Δ17O values of NOx precursor to the studied
nitrates – highlighting oxidation mechanisms
In the present subsection, we estimate the Δ17O values of
NO2 involved during the production of the Alberta nitrates based on
the observed nitrate values and discuss the implications of these
estimations. Generally, winter to summer isotopic differences are thought to
be due to the high oxygen isotopic values of N2O5 due to
interaction with O3 (Johnston and Thiemens, 1997; Michalski et al.,
2003; Morin et al., 2008; Vicars et al., 2012) and low values of OH in
isotopic equilibrium with atmospheric H2O (Dubey et al., 1997).
According to Table 5, the first reaction pathway produces nitrates via
R4-R7-R8 with two-thirds of the O atoms coming from NO2, one-sixth from
O3 and one-sixth from H2O, while the second produces nitrates
via Reaction (R6) with two out of three O atoms coming from NO2 and
one-third from OH (e.g., Michalski et al., 2003). Using these proportions with the
Alberta Δ17O values of pNO3- and
HNO3 in weighted averages allows us to make a rough estimation of
the maximum and minimum Δ17O values of NO2
oxidized to nitrates in the air masses sampled. The calculations assume the O
from O3 contributes a signal of ∼39 ‰ as was
recently measured (Vicars and Savarino, 2014) and that Δ17O of OH
and H2O are zero. The estimated NO2 Δ17O
values for fall–winter (34 ‰–45 ‰ daytime, 25 ‰–36 ‰
night-time) and for summer (25 ‰–34 ‰ for daytime; 15 ‰–24 ‰
for night-time) represent the extremes, assuming daytime oxidation takes place
100 % through the OH pathway and night-time oxidation takes place entirely
through the N2O5 pathway. One should keep in mind that the
Alberta results are for nitrates collected during multi-week sampling
periods. Each nitrate sample therefore contains a mixture of O
from the pathways operating during daytime (Reaction R6) and night-time
(Reactions R4-R7-R8) a priori. Assuming a 50 % contribution from each pathway for
summer, we generate values ranging from 20 ‰ to 29 ‰. Alternatively,
assuming domination of the N2O5 pathway during winter (90 %;
Michalski et al., 2014), the range is 26 ‰–37 ‰. Fall and spring
values should fit between these summer and winter estimated ranges. The
estimated NO2Δ17O ranges indicate that the
potential parent NO2 had a smaller 17O anomaly than
O3 (39 ‰; Vicars and Savarino, 2014) or NO2 in
isotopic equilibrium with O3 alone (45 ‰; Michalski et
al., 2014) in all possible scenarios.
Two mechanisms could be responsible for the Δ17O differences
between these estimates and NO2 in isotopic equilibrium with O3.
One is the competition of Reaction (R3) with Reaction (R2) in oxidizing NO to NO2, since
RO2 will decrease the Δ17O values relative to an
ozone-only equilibrium. The relative reaction rates of Reactions (R2) and (R3) have
previously been presumed to control the NO2 isotopic composition
(e.g., Alexander et al., 2009) based on the
assumption of isotopic steady state. A larger contribution of RO2 is
expected in the NO2 precursors for summer relative to winter, since
biogenic VOCs that are major sources of RO2 radicals are much higher in
the summer (e.g., Fuentes and Wang, 1999). This
suggestion is consistent with the lower Δ17O ranges in summer
reported here. A second possibility is that the nitrates were formed from
some NOx that did not reach an isotopic steady state with O3,
retaining some of its original signature (assumed to be Δ17O=0 ‰). Most studies have assumed that
an isotopic steady state is established between O3 and NO2 within a
few minutes after emission of NOx from a combustion source – or at
least, that nitrate formation is negligible before NOx isotopic
equilibrium is reached. However, recent modelling by Michalski
et al. (2014) suggests that isotopic equilibration of NOx with O3
could take from several minutes up to a few hours at the relatively low O3
concentrations in rural Alberta. At the measured average wind speeds on site
of 8–19 km h-1, transit times from the nearest sources to observation
sites are estimated to be 9–55 min. While the fraction of NOx
converted to nitrate in this transit time may be small, these are large
sources of NOx in an area with very low background nitrates. For
example, a plume containing 10 nmol mol-1 of NO2 mixing with
background air with 0.1 pmol mol-1 of OH (Howell et al., 2014) would
produce HNO3 via Reaction (R6) at a rate of 0.011 µg m-3 min-1 of
NO3-–N at T=7 ∘C (Burkholder et al., 2015), or an
equivalent amount of a typical nitrate sample in 10 min (Table S1).
Even if equilibration with O3 is established within a few minutes, the
nitrate produced in the interim can constitute a substantial fraction of the
sample collected nearby. Therefore, the nitrates measured at our sites may
partly derive from NOx that had not yet reached an isotopic steady state
with O3. These two mechanisms are not exclusive and could both
contribute to lower NOx, and therefore nitrate, Δ17O
values.
Weighted Δ17O average for the sum of dry nitrates
as a function of NO2 concentration divided by pNO3
plus HNO3 concentrations, a ratio indicative of the maturity of a
plume.
An additional piece of evidence suggests that the NOx plumes
themselves, rather than ambient conditions, are the source of low-Δ17O nitrates in these samples. There is a strong correlation
between the total nitrate Δ17O values and the maturity of
the plume as expressed by the NO2 concentration divided by the sum of
HNO3 and pNO3- concentrations (Fig. 5). This
observation is consistent with the unequilibrated NO2 hypothesis.
However, it does not rule out the possible contribution of RO2,
since VOC releases from the NOx sources could lead to
elevated RO2 concentrations in the plume.
Causes of shifts in HNO3 to pNO3- isotopic
differences
A challenging question is as follows: why do the HNO3 to pNO3-
isotopic differences shift seasonally (Fig. 3)? One factor that may influence
the relationship between HNO3 and pNO3- is
the mass-dependent isotopic equilibrium between NH4NO3 and
HNO3 (Reaction R10); however, this mechanism would result in higher
δ18O in pNO3- and unchanged
Δ17O values and, therefore, cannot be solely responsible for
any of the observed patterns (Fig. 3). Alternately, the trend for cold months
(trend i) could be due to the fact that the heterogeneous
N2O5 pathway is likely to produce more pNO3-
than HNO3(g), which would result in a higher contribution from
ozone and explain why δ18O and Δ17O values
are both higher in pNO3-. A previous study addressing why
pNO3- on coarse particles is more enriched than on fine
particles invoked a similar explanation (Patris et al., 2007).
For some of the spring and summer samples, both δ18O and
Δ17O values were lower in pNO3- than in
HNO3 (trend iv). Therefore the mechanism above cannot
dominate the fractionation; nor can a mass-dependent process be responsible.
We suggest a different fractionation process because HNO3 dry-deposits to surfaces more rapidly than pNO3- (Zhang et al.,
2009; Benedict et al., 2013), which would create the discussed isotopic
shifts in the situation where NO2 has low Δ17O
values in a fresh plume. The first nitrates formed in the plume shortly after
emission from the NOx source have low δ18O
and Δ17O values, either because NOx has not
yet reached an isotopic steady state with O3 or because it reacted
with 17O-poor RO2 present in the plume due to VOC
emissions. Those nitrates that form as pNO3- or that
partition to pNO3- remain in the plume for longer than
HNO3, which is removed from the plume rapidly upon contact with
vegetation or other surfaces. As the plume travels, the NOx
becomes more enriched, and the newly formed nitrates take on higher
δ18O and Δ17O values. However,
pNO3- collected further downwind will derive from a mixture
of low-δ18O and -Δ17O pNO3-
formed earlier, plus high-δ18O and -Δ17O
pNO3- formed more recently, while HNO3 will have a
larger proportion formed more recently and will therefore have higher
δ18O and Δ17O values. The fact that we find
the lowest isotopic values in summer pNO3- samples collected
from various anthropogenic sources at distances less than 16 km supports
this suggestion (Table 1).
The above two mechanisms that we propose to explain the shifts in
HNO3 to pNO3 isotopic differences between cold and
warm sampling periods – a differential N2O5 contribution
resulting in higher Δ17O values in pNO3-
than in HNO3, and a differential deposition resulting in lower
Δ17O values in pNO3 – would essentially
compete against each other, with local conditions and chemistry influencing
the results. In winter, when the N2O5 pathway is most important,
the first mechanism dominates, as supported by the observation that
pNO3- concentrations are higher during that season (trend
i). Conversely, in summer, when the N2O5 pathway is
less important and dry deposition is likely faster due to absence of snow
cover, higher surface wetness and high leaf areas, the second mechanism is
more important (trend iv). The local reactant concentrations, wind
speeds and radiative fluxes (which control the time to reach an isotopic
equilibrium) would also be factors in the second mechanism. We find
intermediate trends (ii, iii) in the transitional seasons, as
expected. In addition to these non-mass-dependent fractionation processes,
mass-dependent fractionation in formation and loss of nitrate likely
contributes to the observed δ18O differences. For instance,
kinetic fractionation may be involved in the production of trend
iii.
Isotopic ratios for atmospheric pNO3-,
wNO3- and HNO3 samples in cold and warm periods
from central and southern Alberta (this study), compared with previously
published winter and summer bulk and throughfall deposition samples from the
oil sands (OS) region from northern Alberta (Proemse et
al., 2013), and pNO3- in-stack emissions data for an OS
upgrader located in the same region (Proemse et al.,
2012). The grey dotted line connects NOx from theoretical
combustion with O2 isotopic composition and at isotopic equilibrium
with tropospheric O3 (Michalski et al., 2014).
In summary, examining the isotopic relationship of HNO3 to
pNO3- (Fig. 3) reveals the complexity of anthropogenic
NOx oxidation mechanisms. The lower pNO3-
isotopic values relative to the HNO3 values during warm months may
reflect differential removal rates from plumes containing NO2
temporarily low in 17O.
Low δ18O and Δ17O trends in global w- and
pNO3- – implications for polluted air masses
Atmospheric nitrates measured in central and southern Alberta were sampled
downwind of well-identified anthropogenic sources to verify the potential
role of emitted NOx isotopic signals through to final nitrate isotopic
ratios (primarily N isotopes; see Savard et al.,
2017). As expected, the measured oxygen isotopes of the various nitrate
groups are consistent with exchange with O3 and oxidation through the
well-known OH and N2O5 oxidation paths. However, NO2 not in
isotopic equilibrium with O3, and/or NO reacted with RO2, may have
significantly influenced the overall results. Co-contaminants in the
emissions and sampling plumes at short distances from the sources may have
favoured these two mechanisms, and quantifying RO2 and/or HO2
would help distinguish between the two mechanisms. Meanwhile, our results
raise the following question: are these overall effects observable in triple oxygen
isotopes of nitrates from other polluted sites?
The full Δ17O and δ18O ranges for
pNO3-, wNO3- and HNO3 (between 13.8 ‰
and 20.5 ‰ and 48.4 ‰ and 83.2 ‰; Table 4) compare well with the
isotopic ranges obtained for bulk deposition NO3- samples
collected downwind from oil sands mining operations in the lower Athabasca
region farther north in Alberta (Proemse et al., 2013). Moreover, the
isotopic values in cold and warm months delineated here essentially overlap
with the data sets of winter and summer from the lower Athabasca region
(Fig. 6). This correspondence exists despite the slightly different climatic
conditions (Fig. S1) and very different sampling methods (bulk/throughfall
deposition samples using open ion exchange resin collectors vs. wind-sector-specific active sampling on filters and precipitation-only
collectors). Notably, many points carry relatively low δ18O
and Δ17O values.
Previous work in the Athabasca region reported very low δ18O
and near-zero Δ17O values for pNO3- sampled
directly within oil sands industrial stacks, i.e., in the emissions measured
in-stack and diluted with ambient air (Proemse et al., 2012). These values
are very close to those of O2. Similar isotopic signatures are very
likely produced in source emissions of NOx in the studied
Edmonton and Vauxhall areas (e.g., CFPPs, gas compressors, industries). This
source signature may persist into pNO3 collected close to the
sources. Within the first few hours in the atmosphere (less, in polluted
areas), the NOxδ18O and
Δ17O values rapidly increase due to isotope exchange with
O3 (Reactions R2, R3, R5 and O3 formation, Table 5;
Michalski et al., 2014) and reach isotopic equilibrium. Though the e-folding
lifetime for NOx oxidation to nitrates may be longer than
these few hours, depending on the NOx/VOC ratio,
only a fraction of the oxidized source NOx will create a
measurable contribution to the ambient nitrate where the background air is
very low in nitrate. This is likely the case in the oil sands region, where
Proemse et al. (2013) reported the lowest Δ17O values
within 12 km of the emission sites, and where direct stack emissions of
pNO3- were ∼5000 times lower than NOx
emissions (Wang et al., 2012).
In a methodological test study, we obtained low values for
wNO3- sampled near a high traffic volume highway in Ontario,
Canada (Smirnoff et al., 2012). Low δ18O and
Δ17O values in atmospheric nitrates during warm months (65
and 20 ‰ or less, respectively) have been reported for other parts
of the world as well (Table 4). Authors of these studies have invoked peroxy
radicals to account for low δ18O values in
wNO3- from a polluted city (Fang et al., 2011), in
pNO3- from Taiwan collected partly from air masses influenced
by pollutants (Guha et al., 2017) and from a polluted coastal site in
California (Michalski et al., 2004; Patris et al., 2007; Table 4). However,
sampling in these three other regions did not use collection restricted to
air masses transported from targeted anthropogenic sources. So uncertainties
persist regarding the ultimate sources of nitrates with low isotopic values.
Although a few low values are also reported for seemingly non-polluted areas
of the Arctic and Antarctic regions (unknown cause; Morin et al., 2008, 2009) and of coastal California (Patris
et al., 2007), the information from the literature integrated with the
interpretation proposed for the Alberta low δ18O and Δ17O values in summer nitrates may reflect the involvement of air
masses that include nitrates from oxidation of NO2 with light isotopes
in plumes. In such cases, while not ruling out a higher contribution from
RO2 oxidation of NO, it is also possible that significant portions of
the collected nitrate were formed before the NOx–O3 isotopic
equilibrium was reached (see Sect. 4.1). Keeping in mind that other
hydrocarbon and halogen pathways may play a role in determining the isotopic
nitrate characteristics in other parts of the world, we propose that, in
general, the warm periods' isotopic ranges appear to be lower in polluted
areas. Given these points, our nitrate δ18O and Δ17O may reflect relative proximity to anthropogenic N emitters in
general. Further research work on plume NOx to nitrates chemical
mechanisms may help to validate this suggestion. In the future, the
assumption of an NOx isotopic steady state with O3 should be
explored, given recent findings (Michalski et al., 2014), the
critical importance of NOx isotope characteristics on resulting nitrate
isotopic values (Alexander et al., 2009) and the
suggestion regarding the evolution of NOx–NO3- signals in
fresh anthropogenic plumes (present study).
Conclusion
The HNO3, wNO3 and pNO3 from
anthropogenic sources in central and southern Alberta, simultaneously
collected with wind-sector-based conditional sampling systems, produced
δ18O and Δ17O trends confirming the
previous contention that regional ambient conditions (e.g., light intensity,
oxidant concentrations, RH, temperature) dictate the triple isotopic
characteristics and oxidation pathways of nitrates.
The gaseous form of nitrate (HNO3), having distinct isotopic
characteristics relative to the wet and particulate forms, implies that
understanding nitrate formation and loss requires the nitrate
species to be characterized individually. Particulate NO3- in these samples
generally shows higher δ18O and Δ17O values
than HNO3 in the fall–winter period as the heterogeneous
N2O5 pathway favours the production of pNO3-. In
contrast, HNO3 has higher δ18O and
Δ17O values during warm periods, which we propose is due to
faster dry deposition rates relative to pNO3- in the event
that low-Δ17O NO2 is oxidized in the plume. The
mechanisms conferring nitrate with relatively low isotopic values, whether
oxidation before the NOx–O3 equilibrium is reached or
higher contributions from RO2, are likely to be observed in
anthropogenic polluted air masses. An interesting deduction arising from this
interpretation and from a comparison with nitrate isotopes from other
polluted areas of the world is that relatively low δ18O and
Δ17O values may reflect nitrates produced from
undifferentiated anthropogenic NOx emissions.
Future research should explore the assumption of NOx isotopic
equilibration with O3, given the critical importance of NOx
isotopes on resulting nitrate isotopic values. More field sampling,
including additional on-site oxidant data, and state-of-the-art isotopic
analyses of all tropospheric nitrate species as well as NOx are
required for refining our understanding of atmospheric nitrate worldwide.
This endeavour is fundamental for developing effective emission reduction
strategies towards improving future air quality.