Introduction
The Canadian Air and Precipitation Monitoring Network (CAPMoN) measures
trace gas concentrations and particulate inorganic ion concentrations in air
and precipitation at rural locations across Canada. Since 1983, the network
has been collecting filter and precipitation samples, and the number of sites
has expanded to 33 as of 2010. CAPMoN was developed to monitor trends in
atmospheric pollutants contributing to smog and acid rain, and the data were
later used to assess the impacts of environmental policies in the
Canada–US Air Quality Agreement. This bilateral agreement signed in 1991
recognizes the impacts of transboundary pollution and sets objectives to
reduce SO2 and NOx emissions.
In this study, the focus is on the particulate base cations (Ca2+,
Mg2+, K+, Na+, NH4+), and acidic anions (Cl-,
NO3-, and SO42-), nitric acid, and sulfur dioxide that
have direct impacts on acid rain. Nitrates and sulfates in acid rain reduce
soil quality by causing the depletion of base cations, which are plant
nutrients and are also involved in neutralizing acids. Base cations in soil
can be replenished by mineral weathering, deposition, wind erosion,
agricultural tilling, and forest fires (Hedin et al., 1994; Driscoll et al.,
2001). However, when acidic deposition exceeds the supply of base cations,
soil acidification occurs. Soil acidity has consequently increased the
leaching of inorganic aluminum (Al) monomers, which is a toxic form of Al to
plants and animals (Driscoll et al., 2001). Trees (e.g., red spruces and
sugar maples) experienced damage to foliage, decreased adaptability to cold
climates, slower growth, and mortality during 1960s–1980s from direct and
indirect impacts of acid rain (Driscoll et al., 2001; Watmough and Dillon,
2003). Acid rain and runoff of acidic soil also increased nitrates, sulfates,
and inorganic Al and reduced pH in surface waters of Atlantic Canada,
south-central Ontario, and the northeastern USA (Clair et al., 2002; Driscoll et
al., 2003; Jeffries et al., 2003). Lake acidification has led to detrimental
effects including mortality on zooplankton and fish (Driscoll et al., 2001,
and references therein). Terrestrial birds are also impacted because when
calcium is depleted from soil, less calcium-rich insects are available for
birds to consume (Hames et al., 2002). Calcium deficiency in birds can cause
eggshell thinning and other reproductive consequences (Hames et al., 2002).
Assessments of lake acidification in the previous decade indicate declines
in nitrates and sulfates in surface water, some improvements to pH and acid
neutralizing capacity, and conversion to less toxic organic Al (Clair et
al., 2002; Driscoll et al., 2003; Jeffries et al., 2003; Kothawala et al.,
2011; Strock et al., 2014). Although nitrate and sulfate deposition have
been decreasing, surface water conditions have not improved at the same rate
because nitrates and sulfates that have accumulated in soil and wetlands
over a long period of time is gradually releasing to surface waters
(Stoddard et al., 1999; Driscoll et al., 2001; Clair et al., 2002; Jeffries
et al., 2003). A recent assessment by Lawrence et al. (2015) indicates no
additional soil acidification and that acid deposition effects on soil have
started to diminish in the northeastern USA and eastern Canada according to
indicators, such as exchangeable Ca and Al, base cations, and pH levels.
Considering the role of inorganic ions on acid deposition effects on biota,
it is important to continually study the wet deposition of inorganic ions.
The wet deposition of particulate base cations and acidic anions depends on
the particulate concentrations of these inorganic ions in air and some trace
gases, such as nitric acid and sulfur dioxide. This simplified relationship
is the premise behind the scavenging ratio, defined as a ratio of a
pollutant's concentration in precipitation to that in air. In reality, wet
deposition is a very complex process that involves an understanding of cloud
and precipitation processes and aqueous-phase chemistry, which are considered
to be the major sources of uncertainty in wet deposition modeling (Tost et
al., 2007; Kajino and Aikawa, 2015). Scavenging ratios can be considered a
measure of the wet scavenging efficiency of air pollutants, since they have
been used to compare the precipitation removal of different pollutants in
previous studies (Galloway et al., 1993; Guerzoni et al., 1995; Tuncel and
Ungör, 1996; Shrestha et al., 2002; Hicks, 2005; Kulshrestha et al.,
2009; Bourcier et al., 2012; Zhang et al., 2015). These studies demonstrated
that scavenging ratios vary according to particle size distribution similar
to the particle size dependency of scavenging coefficients typically used in
wet deposition modeling (Wang et al., 2014). Thus, scavenging ratios of
particulate-phase pollutants have been used as a surrogate for other
particulate-phase pollutants with similar particle sizes (Cadle et al., 1990;
Sakata and Asakura, 2007; Cheng et al., 2015).
Site and data descriptions. NA indicates no available data.
Short-term data were not analyzed. Refer to Fig. S1 for a map of the sites.
Site name
Province
Latitude
Longitude
Elevation
Coastal/inland
Land use
Air data
Wet deposition data
(m)
Saturna
BC
48.78
-123.13
178
Coastal
Forest
Dec 1990–Dec 2010
Jan 1990–Dec 2011
Snare Rapids
NT
63.52
-116.00
240
Inland
Forest
NA
Jan 1989–Dec 2011
Esther
AB
51.67
-110.20
707
Inland
Agricultural
Oct 1991–Mar 2003
Jan 1987–Dec 2002,
Jan 2009–Dec 2011
Cree Lake
SK
57.35
-107.13
499
Inland
Forest
Jul 1982–May 1993
Jan 1984–Dec 1992
Bratt's Lake
SK
50.20
-104.71
600
Inland
NA
Aug 2001–Dec 2010
Jan 2001–Dec 2011
McCreary
MB
50.71
-99.53
335
Inland
Agricultural
NA
Jan 1984–Dec 1995
Island Lake
MB
53.87
-94.67
245
Inland
Forest
NA
Jan 1984–Dec 1997
Experimental Lakes
ON
49.66
-93.72
369
Inland
Forest
Jan 1979–Dec 2010
Jan 1984–Dec 2011
Area (ELA)
Pickle Lake B
ON
51.45
-90.22
370
Inland
Forest
NA
Jan 2003–Dec 2011
Algoma
ON
47.04
-84.38
411
Inland
Forest
Oct 1980–Dec 2010
Jan 1985–Dec 2011
Burnt Island
ON
45.82
-82.95
185
Inland
Forest
NA
Jan 1992–Dec 2011
Bonner Lake
ON
49.39
-82.12
245
Inland
Forest
short-term
Jun 1985–Dec 2011
Longwoods
ON
42.88
-81.48
239
Inland
Agricultural
Jan 1983–Dec 2010
Jan 1984–Dec 2011
Priceville
ON
44.17
-80.66
475
Inland
Agricultural
NA
Jan 1985–Dec 1994
Egbert
ON
44.23
-79.78
253
Inland
Agricultural
Jul 1988–Dec 2010
Jan 1989–Dec 2011
Egbert-2
ON
44.23
-79.78
253
Inland
Agricultural
short-term
Jan 1997–Dec 2011
Sprucedale
ON
45.42
-79.49
350
Inland
Agricultural
May 2002–Dec 2010
Jan 2003–Dec 2011
Warsaw Caves
ON
44.46
-78.13
230
Inland
Agricultural
NA
Jan 1986–Dec 2011
Chalk River
ON
46.06
-77.41
184
Inland
Forest
Jan 1979–Dec 2010
Jan 1984–Dec 2011
Chapais
QC
49.82
-74.98
381
Inland
Forest
Jun 1988–Dec 2010
Jan 1988–Dec 2011
Frelighsburg
QC
45.05
-72.86
203
Inland
Agricultural
Nov 2001–Dec 2010
Jan 2002–Dec 2011
Sutton
QC
45.08
-72.68
243
Inland
Forest
Jan 1986–Mar 2002
Jan 1984–Dec 2001
Lac Edouard
QC
47.68
-72.44
243
Inland
Forest
Jan 2002–Dec 2010
Jan 2002–Dec 2011
Montmorency
QC
47.32
-71.15
640
Coastal
Forest
Dec 1980–Jan 1997
Jan 1984–Dec 1996
Harcourt
NB
46.50
-65.27
37
Coastal
Forest
NA
Jan 1984–Dec 2011
Kejimkujik
NS
44.43
-65.21
127
Coastal
Forest
May 1979–Dec 2010
Jan 1984–Dec 2011
Mingan
QC
50.27
-64.22
14
Coastal
Forest
short-term
Jan 1994–Dec 2011
Jackson
NS
45.59
-63.84
90
Coastal
Forest
NA
Jan 1984–Dec 2011
Goose Bay
NL
53.31
-60.36
39
Coastal
Forest
NA
Jan 1984–Dec 2011
Goose Bay B
NL
53.29
-60.39
39
Coastal
Forest
NA
Jan 1989–Dec 2007
Bay d'Espoir
NL
47.99
-55.81
190
Coastal
Forest
short-term
Jan 1984–Dec 2011
The objectives of this study were to (1) analyze long-term geographical
patterns and temporal trends in nitric acid and sulfur dioxide
concentrations and the air concentrations and wet deposition of base cations
and acidic anions; (2) examine geographical and temporal trends in aerosol
acidity and acid rain; (3) determine scavenging ratios for particulate
inorganic ions using precipitation and air concentrations; and (4) develop
an approach for estimating particulate and gaseous species wet scavenging
contributions to total nitrate, ammonium, and sulfate wet deposition and
their scavenging ratios.
Results and discussion
Air concentrations
Geographical patterns
Air concentration statistics and geographical patterns of eight particulate
inorganic ions, SO2, and HNO3 are plotted in Fig. 1. The data were
divided into two time periods from 1983 to 1996 and 1997 to 2010 to examine
potential changes in concentrations due to NOx and SO2 emission
changes. The range in concentrations (based on the 5th percentile to
95th percentile concentration) from all daily samples at all locations
was 0.009–2.9 µg m-3 for Ca2+, 0.002–0.5 µg m-3
for Mg2+, and 0.006–0.2 µg m-3 for K+. Larger
variability was observed in Ca2+, likely because of the variability in
soil emissions depending on land use and wind. Large variability is also
expected for Na+ and Cl- because coastal sites are more frequently
impacted by sea-salt aerosols than continental sites. The Na+ and
Cl- air concentrations ranged from 0.005 to 1.4 and
0.003 to 1.9 µg m-3, respectively. The range in concentrations
was 0.018–5.8 µg m-3 for NH4+, 0.009–8.7 µg m-3 for NO3-, and 0.07–14.5 µg m-3 for
SO42-. HNO3 and SO2 ranged from 0.014 to 5.0 µg m-3 and 0.011 to 25.2 µg m-3, respectively. These ions and
trace gases are likely to have larger variability in air concentrations than
base cations because some sites may be impacted more by anthropogenic
emissions, which form secondary pollutants such as SO42-,
HNO3, NH4+, and NO3-, than other sites.
The geographical patterns in air concentrations were examined in greater
detail based on the median concentration at each location. Long-term median
Ca2+ concentrations among the sites ranged from 0.03 to 0.6 µg m-3. The highest median during both time periods were observed at
Longwoods and Egbert, which are the lowest latitude and most inland air
concentration sites. Longwoods and Egbert are also predominantly agriculture
sites. The median Mg2+ concentrations ranged from 0.01 to 0.09 µg m-3. The highest median was also observed at Longwoods and Egbert.
Higher median concentrations were also found at several western Canada sites,
including at Bratt's Lake, Esther, and Saturna in the post-1997 period. The
median concentrations ranged from 0.02 to 0.06 µg m-3 for K+,
which is the ion with the least spatial variability in the air
concentration. The highest concentrations were observed at Longwoods as well
as at Bratt's Lake post-1997. The median Na+ and Cl- concentrations ranged from 0.02 to 0.5 and 0.007 to 0.3 µg m-3, respectively. As expected, the highest median for both
ions were observed at the two coastal locations, Saturna and Kejimkujik, due
to the proximity to sea-salt aerosol emissions from the ocean. These two
sites are the farthest west and east air sampling locations respectively.
Na+ and Cl- concentrations at Saturna were larger and had
greater variability than at Kejimkujik likely because of the higher
frequency of marine airflows arriving at Saturna (68 % of winds from N and
W directions) than at Kejimkujik (31 % of winds from E and S directions).
The median NH4+ and NO3- concentrations ranged from
0.1 to 1.7 and 0.03 to 2.0 µg m-3, respectively
(Fig. 1). Compared to pre-1997 period, the median concentrations of
NH4+ and NO3- were lower in the post-1997 period. The
highest concentrations for both ions were observed at Longwoods and Egbert.
Higher concentrations were also found at Sutton, Esther, and Frelighsburg
post-1997. The majority of these sites except for Sutton are agriculture
sites located in southern Ontario and Québec, which implies that higher
ammonia emissions from agricultural regions may react with acidic gases in
the atmosphere to form particulate ammonium (Pitchford et al., 2009). Acidic
gases, such as H2SO4 and HNO3, are produced from the
oxidation of SO2 and NOx, respectively, and are primarily emitted
from industrial and urban areas. The proximity of these lower-latitude air
sampling sites to major industrial areas in Ohio and Pennsylvania, USA, could
result in higher acidic gas concentrations at these sites. This is evident
in the air concentration plots for HNO3 that show higher concentrations
of HNO3 at sites having higher NO3-. The median HNO3
concentrations ranged from 0.07 to 1.1 µg m-3. Southerly winds also
impacted Longwoods, Egbert, and Frelighsburg/Sutton approximately 20,
32, and 34 % of the time, respectively. The median SO42-
concentrations among the air sampling sites ranged from 0.6 to 3.5 µg m-3. The concentrations were lower during the post-1997 than during the
pre-1997 period. The highest median concentration was observed at Longwoods.
Higher median concentrations were found at several southern Ontario and
Québec sites, including Egbert, Sutton, Frelighsburg, Sprucedale, and Chalk
River. Larger variability in the concentrations was generally observed
across sites in southern and eastern Canada. This pattern is likely
attributed to the proximity of the sites to combustion and industrial
sources in southern Ontario and Québec. The southern Canada and Atlantic
Canada sites (e.g. Kejimkujik) are downwind of combustion and industrial
areas in Ohio and Pennsylvania. In contrast, SO42- concentrations
at sites located in western and central Canada (e.g., Saturna, Esther, Cree
Lake, Bratt's Lake, and ELA) were at or below the overall median
concentration of all the sites and had smaller variability.
Rate of change in annual air concentrations. Slope refers to the
seasonal Kendall slope (ng m-3 a-1); C.I. refers to the 90 %
confidence interval of the slope; ns indicates no significant trend; na
indicates no available data.
Site
Ca2+
Mg2+
K+
Na+
Cl-
Slope
C.I.
Slope
C.I.
Slope
C.I.
Slope
C.I.
Slope
C.I.
Saturna
-0.9
-1.3 to -0.6
-0.8
-1.1 to -0.4
-0.7
-1 to -0.5
-4.5
-6.5 to -1.8
1.0
ns
Esther
9.9
4.8 to 15.8
1.5
0.9 to 2.5
1.3
0.8 to 1.8
2.1
0.8 to 3.6
-0.4
-0.7 to 0
Cree Lake
na
na
na
na
-1.3
-2 to -0.5
0.5
ns
-0.5
-1 to -0.1
Bratt's Lake
5.5
ns
2.4
ns
-3.8
-5.3 to -2.4
-2.4
-4.2 to -1.3
-0.6
-1.3 to -0.2
ELA
0.5
ns
-0.2
ns
-0.5
-0.6 to -0.3
-0.7
-0.8 to -0.5
-0.2
-0.4 to -0.2
Algoma
-0.3
ns
-0.2
-0.5 to 0
-0.6
-0.8 to -0.5
-0.3
-0.5 to -0.2
-0.1
-0.2 to 0
Longwoods
-15.9
-21.1 to -10
-1.1
-2.1 to -0.4
-0.6
-0.9 to -0.3
-0.5
-0.8 to -0.2
-0.8
-1 to -0.5
Egbert
-9.7
-14.8 to -3.6
0.1
ns
0.02
ns
-0.2
ns
0.2
ns
Sprucedale
-2.7
ns
-0.9
ns
-1.2
-1.8 to -0.6
-1.3
-1.9 to -0.4
-0.2
ns
Chalk River
-0.3
ns
-0.6
-0.8 to -0.4
-0.8
-1 to -0.6
-0.5
-0.8 to -0.3
0.03
ns
Chapais
0.4
ns
-0.03
ns
-0.5
-0.6 to -0.4
-0.9
-1.2 to -0.7
0.3
0 to 0.6
Frelighsburg
-0.8
ns
0.3
ns
-1.1
-1.8 to -0.7
-2.6
-4.4 to -0.7
-0.01
ns
Sutton
1.5
ns
0.2
ns
-1.8
-2.3 to -1.4
-1.8
-2.3 to -1.4
-0.2
ns
Lac Edouard
-0.01
ns
-0.2
ns
-1.0
-1.4 to -0.6
-1.4
-2.1 to -0.5
-0.4
ns
Montmorency
na
na
na
na
-0.6
-1.1 to 0
1.4
0.8 to 2.1
0.1
ns
Kejimkujik
0.4
0 to 0.8
-0.5
-0.8 to -0.1
-0.5
-0.7 to -0.4
1.8
0.4 to 3
4.9
3.7 to 6.6
Site
NH4+
NO3-
SO42-
SO2
HNO3
Slope
C.I.
Slope
C.I.
Slope
C.I.
Slope
C.I.
Slope
C.I.
Saturna
-7.6
-9.1 to -6.2
-9.3
-12.8 to -5.6
-28.8
-32.6 to -25.2
-62.6
-72.5 to -52.5
-17.1
-20.7 to -13.6
Esther
7.9
2 to 15.6
40.1
32.6 to 50.4
7.0
ns
-6.5
ns
12.0
2.2 to 21.8
Cree Lake
3.7
1.1 to 6.2
-1.1
ns
0.0
ns
8.4
ns
8.9
6 to 11.4
Bratt's Lake
-16.5
-23.4 to -9.7
-21.4
-34.3 to -10.5
-30.4
-51.1 to -11.8
1.6
ns
-19.4
-27.3 to -13.3
ELA
-4.4
-6 to -2.6
3.5
1.7 to 5.2
-28.4
-32.1 to -24.4
-14.6
-17.7 to -12.2
-2.3
-3.7 to -0.8
Algoma
-7.0
-9.9 to -3.7
6.4
4.4 to 8.4
-46.8
-55 to -39
-77.4
-87.2 to -67.3
-3.8
-7.1 to -0.9
Longwoods
-28.8
-34.4 to -23.1
-15.2
-22.9 to -6.5
-97.4
-110.2 to -87.4
-221.5
-240.2 to -202.1
-27.6
-32.7 to -23.3
Egbert
-41.8
-49.3 to -35.5
-34.2
-43.2 to -23.6
-93.9
-108.9 to -81.5
-199.4
-221.6 to -177.3
-25.1
-31.2 to -19.5
Sprucedale
-32.8
-52.4 to -12.6
-29.5
-39.9 to -16.6
-99.0
-150.5 to -36.8
-140.0
-188.7 to -97.2
-54.4
-74.5 to -36.3
Chalk River
-8.8
-11.6 to -6.4
2.8
1.5 to 4.3
-59.8
-66.9 to -52.1
-104.7
-115.3 to -94.1
-8.2
-10.8 to -5.7
Chapais
-5.7
-7 to -4.2
1.3
0.8 to 2
-41.3
-46.3 to -36
-45.5
-52.3 to -39.6
-4.3
-5.8 to -3.1
Frelighsburg
-58.2
-69 to -45.4
-53.4
-68.6 to -38.9
-108.6
-139.5 to -84
-160.8
-206 to -119.7
-81.7
-91.2 to -68.9
Sutton
-4.6
ns
18.8
13.9 to 24.2
-64.2
-80.4 to -51.2
-90.1
-112.8 to -67.3
-12.7
-18.9 to -4
Lac Edouard
-20.3
-27.9 to -13.1
-11.0
-14.3 to -6.4
-64.8
-88 to -47.6
-55.8
-76.2 to -34.1
-35.9
-42.4 to -27.1
Montmorency
9.6
6.1 to 14.6
5.9
4.5 to 8.2
-1.0
ns
-20.1
-40 to -3.2
21.0
14.9 to 29
Kejimkujik
-5.3
-6.5 to -4
2.5
1.4 to 3.8
-53.1
-59 to -47.5
-35.9
-41.2 to -30
-6.5
-7.9 to -4.8
The median SO2 concentrations ranged from 0.4 to 6.4 µg m-3
during the 1983–1996 period and from 0.6 to 2.3 µg m-3 post-1997
(Fig. 1). There was also a reduction in the variability in the
concentrations in the post-1997 period. The lower-latitude southern Ontario
and Québec sites had higher SO2 concentrations, whereas western and
central Canada sites had much lower concentrations. This geographical
pattern is similar to that of SO42-. One exception was that the
SO2 concentrations in eastern Canada were similar to or even lower than
in western Canada, whereas SO42- concentrations in eastern Canada
were slightly higher than in western Canada. Eastern Canada sites including
Montmorency, Lac Edouard, and Kejimkujik are remote sites; therefore,
SO2 concentrations are likely not elevated by the time it arrives at
these remote locations since SO2 can undergo deposition or transform
to SO42- during transport. The slightly higher SO42- in
eastern Canada could be from sea-salt sulfate due to the proximity to the
Atlantic Ocean.
Temporal patterns
The Kendall slopes and confidence interval in Table 2 shows the annual rate
of change in the concentrations of particulate ions and trace gases at the
air sampling sites for all years with available data. A significant temporal
trend in Ca2+ was observed at 5 of 16 sites with no significant changes
in the concentrations observed at the remaining sites. Decreasing trends
were observed at Saturna, Longwoods, and Egbert, while increasing trends
were observed at Esther and Kejimkujik. The largest decline in Ca2+ was -16 ng m-3 a-1 at Longwoods. Overall the temporal changes in
Ca2+ are small. For Mg2+, a significant decreasing trend was found
at Saturna, Longwoods, Chalk River, and Kejimkujik. However, the rate of
decline was very small ranging only from -0.5 to -1.1 ng m-3 a-1.
The rate of decline for K+ ranged from -0.5 to -3.8 ng m-3 a-1 and was observed at 14 of 16 sites. A plot of the annual average
K+ for 10 of the active air sampling sites are shown in Fig. 2a for
the 1983–2010 period, which illustrates a gradual decline in K+.
Significant decreases in Na+ were found at 11 of 16 sites, with
magnitudes ranging from -0.3 to -4.5 ng m-3 a-1. The steepest
decline was observed at a coastal site in western Canada. However,
increasing trends were observed at two coastal sites in eastern Canada
(Montmorency and Kejimkujik). For Cl-, decreasing temporal trends were
found at only 6 of 16 sites, suggesting that the temporal trends are not
necessarily related to sea-salt emissions. The declines in Cl-, ranging
from -0.1 to -0.8 ng m-3 a-1, were found at sites in western and
central Canada and at Algoma and Longwoods.
Temporal trends of annual average atmospheric K+ (a)
and atmospheric NH4+ and annual ammonia emissions (b). In
panel (b) between 2002 and 2010, atmospheric NH4+ at
agricultural sites decreased by 6.3 % while ammonia emissions in Ontario
and Québec decreased by 2.4 and 1.1 %, respectively.
NH4+ concentrations have been decreasing at 12 of 16 air sampling
sites (Table 2). This result is consistent with the widespread decrease in
NH4+ at CAPMoN sites during 1988–2007 (Zbieranowski and Aherne,
2011). The rate of decrease ranged from -4 to
-58 ng m-3 a-1. The largest declines as shown in Fig. 2b were
observed at Longwoods, Egbert, Sprucedale, and Frelighsburg, which are
agriculture sites located in southern Ontario and Québec. The annual
decrease was 7.3 times greater than other sites based on the linear
regression slopes. The decreasing trend in NH4+ corresponds to the
decreasing trend in ammonia emissions in Ontario and Québec particularly
in the post-2002 period (Fig. 2b) (ECCC, 2014). Aside from its relationship
to ammonia, the negative trend in NH4+ was also strongly tied to
trends in NO3- and SO42-. There was an even split in the
number of sites with increasing trends and decreasing trends in NO3-.
The rate of increase ranged from 1.3 to 6.4 ng m-3 a-1 among
active sites, whereas the annual trend was 6–40 ng m-3 a-1
among inactive sites (e.g. Esther, Sutton, Montmorency). The annual trend in
NO3- decreased from -9.3 to -53 ng m-3 a-1 at other
sites (Table 2). Larger declines were observed at the agriculture sites
located in southern Ontario and Québec. Differences in temporal trends
were also observed during different time periods. At 9 of 16 sites, an
increasing trend was found between 1991 and 2001 which was followed by a
decreasing trend from 2001 to 2010 (Fig. 3). The difference in NO3-
trends between the two decades was also reported in previous analysis of
CAPMoN sites (Zbieranowski and Aherne, 2011). The change in NO3-
temporal trends closely resembled that of NOx emissions in Canada.
Between 1991 and 1997, NOx emissions in Canada increased annually and
only began to decrease from 1997 to 2010 (Fig. 3) (ECCC, 2014). In the USA,
NOx emissions were constant over the 1991–1994 period and only began to
decrease after 1994 (USEPA, 2017). Reductions in NOx emissions were
implemented following the introduction of the Canada–US Air Quality
Agreement and the US Acid Rain Program and Clean Air Interstate Rule. The
decrease in NOx emissions were largely attributed to lower emissions
from stationary fuel combustion and transportation sectors (Lloret and
Valiela, 2016).
Temporal trends of annual atmospheric NO3- and annual
NOx emissions. Slope 1 refers to the regression line for NO3-
between 1991 and 2001 (positive trend, 3.6 % increase), while Slope 2 is
for the period between 2001 and 2010 (negative trend, 6.5 % decrease).
Between 1991 and 1997, NOx emissions in Canada (Cdn) increased by
8.5 %. Between 1997 and 2010, Cdn NOx emissions decreased by
25.8 %.
SO42- decreased at a rate of -28 to -109 ng m-3 a-1
depending on the location (Table 2). The steepest annual declines were
observed in the southern Ontario and Québec region as shown in Fig. 4.
The slope of the linear regression equation for the southern Ontario and
Québec sites in Fig. 4 was 2 times greater than that of other air
sampling sites, which are coastal or higher latitude sites distant from major
industrial and urban areas. The geographical patterns in the temporal trends
of SO42- were also similar to those of SO2 and HNO3. The
annual trends for SO2 and HNO3 in the southern Ontario and
Québec region declined 3.8 and 4.9 times faster, respectively, than other
air sampling sites across Canada based on the linear regression slopes.
Negative trends for SO42- and SO2 concentrations followed the
decreasing trend in SO2 emissions in both Canada and the USA since 1990
(ECCC, 2014; USEPA, 2017) (Fig. 4), corresponding to the period of the
Canada–US Air Quality Agreement and the US Acid Rain Program and Clean Air
Interstate Rule. Note the steeper decline in SO2 emissions in Ontario in
recent years, which is potentially attributed to the phase-out of coal use in
Ontario power plants beginning in 2005 (MOE, 2015).
Temporal trends of annual average atmospheric SO4- and
annual SO2 emissions. ON and QC refer to the province of Ontario and
Québec, respectively. Between 1990 and 2010, atmospheric SO4-
decreased by 4.5 % at southern Ontario (ON) and Québec (QC) sites and by
3.6 % at other sites. Over the same period, SO2 emissions decreased
by 1.8 % in Canada (Cdn) and 4.3 % in the USA.
Aerosol acidity (c / a)
The median c / a ranged from 0.97 to 1.6 and an overall median c / a of 1.07 was
observed across all sites (Fig. 5a). These values are based on inorganic ion
contributions to aerosol acidity. Organic acids do not contribute
significantly to aerosol acidity compared to the strong inorganic acids
(Zhang et al., 2007; Ziemba et al., 2007; He et al., 2012); however, they
have been included in the measure of aerosol acidity in some studies
(Hennigan et al., 2015, and references therein). Among the sites, the highest
aerosol acidity was observed at Kejimkujik because of the higher equivalent
anion concentrations relative to cations (Table S2). Higher aerosol acidity
was prevalent generally in eastern Canada and central Ontario regions. In
contrast to Kejimkujik, the majority of the sites had higher cation than
anion concentrations or near equivalent cation and anion concentrations
(Table S2). Even though locations like Longwoods and Egbert had a greater
amount of anions than other locations due to higher NO3- and
SO42-, there were sufficient amounts of cations, mainly
NH4+, to neutralize the acidic species. While the overall median
c / a was close to 1 for all Canadian sites, more than half of the daily c / a
data were below 1 at eight locations (Fig. 5a). This suggests there was a
substantial amount of time between 1994 and 2010, when aerosols were acidic
at some Canadian sites.
Geographical patterns of cation / anion (c / a) ratio (a)
and temporal trends of annual average c / a ratio and precipitation
pH (b). In panel (a), the blue line indicates the median;
the red dot indicates the mean; the box and whiskers include the
interquartile range and the 5th to 95th percentile range, respectively; and the
dotted line is the overall median among the sites.
Geographical patterns of the annual wet deposition of ions. Sites
are arranged in order from western to eastern Canada. The blue line indicates
the median; the red dot indicates the mean; the box and whiskers include the
interquartile range and the 5th to 95th percentile range, respectively; and the
dotted line is the overall median among the sites.
A significant increasing trend in the c / a was observed between 1994 and
2010, which indicates a widespread decline in aerosol acidity (Fig. 5b). The
rate of decrease in aerosol acidity was small and fairly uniform spatially
based on the Kendall slope results. According to Table S2 and Fig. 5b, the
annual average cation and anion concentrations and c / a at most of the sites
(data combined from 13 of 15 sites) were relatively constant between 1994
and 2000. Since 2001, the annual average cation and anion concentrations and
aerosol acidity have been on a slight decline (Table S2). The decrease in
cation concentrations appeared consistent with the declining trend in
NH4+ discussed earlier, since NH4+ is the largest
contributor to cation concentrations. For anions, NO3- and
SO42- are the predominant ions. Thus, the decline in anions was
also consistent with the decreasing rates for NO3- and
SO42-. As mentioned earlier, the decreasing trends in
NH4+, NO3-, and SO42- were consistent with the
reductions in ammonia, NOx, and SO2 emissions.
Wet deposition
Geographical patterns
The annual wet deposition statistics for the various ions at the 31 locations are shown in Fig. 6. The annual wet deposition was based on all
years with complete data because the annual flux was determined by summing
the daily fluxes. The range in the annual wet deposition based on the
5th and 95th percentile annual wet deposition rate (kg ha-1 a-1) was 0.08–3.6 for Ca2+, 0.02–1.6 for Mg2+, 0.01–0.7 for
K+, 0.03–12.0 for Na+, 0.06–23.0 for Cl-, 0.1–6.4 for
NH4+, 0.4–26.5 for NO3-, and 0.5–32.7 for
SO42-. The lowest annual wet deposition rates for Ca2+,
Mg2+, Na+, NH4+, NO3-, and SO42-
were observed at Snare Rapids, which is a remote site in the Northwest
Territories of Canada. The highest annual wet deposition recorded at the
most eastern coastal site in Bay d'Espoir, Newfoundland were from ions
related to sea-salt aerosols, including Na+, Cl-, and Mg2+.
For Ca2+ and ions derived from anthropogenic sources (e.g.,
NH4+, NO3-, and SO42-), the highest annual wet
deposition rates were observed at Priceville and Longwoods, which are the
two most southern wet deposition sites in Canada and closest to urban and
industrial areas.
Long-term median annual wet deposition of Ca2+ ranged from 0.1 to 2.8 kg ha-1 a-1 among the wet deposition sites. The highest annual wet
deposition was observed at Priceville and Longwoods (Fig. 6). Higher median
wet deposition was also found at Algoma, Egbert, and Warsaw Caves. The
majority of these sites are agriculture sites. The lowest median wet
deposition was observed at the western and eastern coastal sites. The median
annual wet deposition of Mg2+ ranged from 0.03 to 1.0 kg ha-1 a-1. The highest wet deposition was found at the western and eastern
coastal sites with the exception of Goose Bay, which is a higher latitude
coastal site located in Labrador. It had the lowest annual precipitation
amount among coastal sites in eastern Canada (Fig. S3). Annual Mg2+ wet deposition at inland sites was typically at or below the overall
median wet deposition for all sites; however, they are slightly higher at
Longwoods and Priceville. The median annual wet deposition of K+ ranged
from 0.05 to 0.4 kg ha-1 a-1. Similar to Mg2+, higher annual wet
deposition of K+ was observed at eastern coastal locations with the
exception of Goose Bay. Annual wet deposition for inland sites was around
the overall median annual wet deposition for all sites except at Longwoods,
Algoma and Priceville. The median annual wet deposition of Na+ and
Cl- ranged from 0.05 to 7.5 and 0.1 to 13.6 kg ha-1 a-1, respectively. The geographical patterns in the annual
wet deposition were similar between Na+ and Cl-. Annual wet
deposition at western and eastern coastal sites was higher and had greater
variability than inland locations.
Median annual NH4+ wet deposition ranged from 0.2 to 5.8 kg ha-1 a-1 (Fig. 6). Higher annual wet deposition was observed at
lower-latitude continental sites. Higher latitude continental locations (e.g.
Cree Lake, Island Lake, Pickle Lake, Bonner Lake, Chapais) and coastal
locations were well below the overall median annual wet deposition. The
median annual wet deposition ranged from 0.8 to 23.3 kg ha-1 a-1 for
NO3- and 0.8 to 26.6 kg ha-1 a-1 for SO42-. These
two ions have similar spatial patterns in annual wet deposition. Higher
median annual wet deposition occurred at southern Ontario and Québec sites.
Lower median annual wet deposition was observed in eastern Canada. The
lowest median annual wet deposition for NO3- and
SO42-, which were well below the overall median annual wet
deposition for all sites, was recorded in western and central Canada. These
results are consistent with Vet et al. (2014), which observed higher sulfur
wet deposition around Lake Ontario and Lake Erie and much lower sulfur wet
deposition in western North America than eastern North America.
Rate of change in annual wet deposition of Ca2+, Mg2+,
K+, Na+, and Cl-. Slope refers to the Sen's slope
(kg ha-1 a-1); C.I. refers to the 90 % confidence interval of
the slope; ns indicates no significant trend.
Site
Ca2+
Mg2+
K+
Na+
Cl-
Slope
C.I.
Slope
C.I.
Slope
C.I.
Slope
C.I.
Slope
C.I.
Saturna
0.001
ns
-0.0003
ns
-0.0002
ns
0.001
ns
0.002
ns
Esther
0.010
ns
0.003
ns
0.002
ns
0.002
ns
-0.001
ns
Cree Lake
-0.014
ns
-0.001
ns
-0.003
ns
0.001
ns
-0.002
ns
Bratt's Lake
0.036
ns
0.009
ns
0.005
ns
0.002
ns
0.001
ns
ELA
0.008
ns
0.002
ns
0.002
ns
-0.002
-0.004 to
-0.001
ns
-0.0002
Algoma
-0.009
ns
-0.003
-0.0064 to
-0.003
-0.0059 to
-0.004
ns
-0.015
-0.025 to
-0.0005
-0.0003
-0.01
Longwoods
-0.017
ns
-0.002
ns
0.006
0.003 to
0.004
ns
-0.013
-0.023 to
0.011
-0.003
Egbert
0.003
ns
-0.001
ns
-0.0005
ns
0.004
0.002 to
-0.008
ns
0.009
Sprucedale
-0.024
ns
-0.008
-0.015 to
-0.002
ns
-0.005
ns
-0.037
ns
-0.003
Chalk River
0.001
ns
-0.001
ns
-0.001
ns
0.001
ns
-0.008
-0.013 to
-0.004
Chapais
0.004
ns
-0.001
-0.0022 to
-0.0002
ns
-0.003
-0.006 to
-0.007
-0.012 to
-0.0003
-0.001
-0.002
Frelighsburg
0.020
ns
0.001
ns
0.001
ns
0.002
ns
-0.012
ns
Sutton
-0.029
-0.058 to
-0.002
ns
-0.001
ns
0.002
ns
-0.006
ns
-0.005
Lac Edouard
-0.004
ns
-0.002
ns
-0.002
ns
-0.014
-0.022 to
-0.025
-0.046 to
-0.001
-0.008
Montmorency
-0.023
ns
-0.002
ns
0.001
ns
0.009
0.0009 to
0.006
ns
0.016
Kejimkujik
0.007
ns
0.003
ns
-0.0003
ns
0.024
ns
0.056
ns
Rate of change in annual wet deposition of NH4+,
NO3-, SO42-, nss-SO42-, precipitation amount and pH.
Slope refers to the Sen's slope (kg ha-1 a-1); C.I. refers to the
90 % confidence interval of the slope; ns indicates no significant trend;
na indicates no available data.
Site
NH4+
NO3-
SO42-
nss-SO42-
Annual precip. (mm a-1)
pH (a-1)
Slope
C.I.
Slope
C.I.
Slope
C.I.
Slope
C.I.
Slope
C.I.
Slope
C.I.
Saturna
-0.01
ns
-0.07
-0.13 to -0.01
-0.13
-0.17 to -0.06
-0.12
-0.16 to -0.07
-1.5
ns
0.012
0.009 to 0.016
Esther
0.02
ns
0.04
ns
-0.02
ns
na
na
-1.4
ns
-0.012
ns
Cree Lake
-0.02
ns
-0.05
ns
-0.09
ns
na
na
-2.0
ns
-0.010
ns
Bratt's Lake
0.07
ns
0.02
ns
0.10
ns
na
na
25.9
ns
-0.009
ns
ELA
0.05
0.03 to 0.07
0.02
ns
-0.04
ns
na
na
6.0
ns
0.009
0.003 to 0.014
Algoma
-0.06
-0.09 to -0.02
-0.38
-0.51 to -0.26
-0.60
-0.71 to -0.47
na
na
-7.8
-14.5 to -2.7
0.020
0.016 to 0.025
Longwoods
0.03
ns
-0.33
-0.48 to -0.17
-0.55
-0.7 to -0.36
na
na
3.9
ns
0.023
0.019 to 0.028
Egbert
-0.001
ns
-0.31
-0.45 to -0.18
-0.45
-0.57 to -0.33
na
na
3.8
ns
0.027
0.022 to 0.03
Sprucedale
-0.03
ns
-1.01
-1.4 to -0.48
-1.00
-1.46 to -0.36
na
na
9.1
ns
0.030
0.018 to 0.042
Chalk River
0.01
ns
-0.19
-0.24 to -0.11
-0.36
-0.43 to -0.29
na
na
5.3
2.1 to 8.3
0.020
0.018 to 0.022
Chapais
-0.01
ns
-0.21
-0.28 to -0.11
-0.34
-0.45 to -0.25
na
na
-0.9
ns
0.015
0.013 to 0.018
Frelighsburg
-0.02
ns
-0.93
-1.29 to -0.64
-0.93
-1.57 to -0.16
na
na
13.5
ns
0.056
0.036 to 0.069
Sutton
0.05
0.01 to 0.09
0.01
ns
-0.57
-0.9 to -0.23
na
na
4.0
ns
0.017
0.01 to 0.026
Lac Edouard
-0.07
ns
-0.69
-0.91 to -0.38
-0.46
-0.77 to -0.23
na
na
14.6
ns
0.027
0.018 to 0.045
Montmorency
0.05
ns
0.20
ns
-0.27
ns
-0.27
ns
27.8
6.1 to 45.7
0.008
0.001 to 0.014
Kejimkujik
0.01
ns
-0.12
-0.18 to -0.06
-0.27
-0.37 to -0.2
-0.28
-0.36 to -0.2
10.9
3 to 17.2
0.014
0.011 to 0.017
The geographical patterns in the wet deposition were predominantly affected
by the air concentrations. For example, the higher Na+ and Cl- wet deposition at coastal locations can be traced back to the higher
Na+ and Cl- air concentrations. Similarly, the higher
NH4+, NO3- and SO42- wet deposition occurring
at southern Ontario and Québec locations was consistent with the
geographical patterns in the air concentrations. Although precipitation
amount is used to determine wet deposition, only the wet deposition patterns
of Mg2+ and K+ were potentially influenced by precipitation
amount. As shown in Fig. S3, the annual precipitation amount generally
increases from western to eastern sites. Only the Mg2+ and K+ wet deposition were higher at eastern Canada locations.
Temporal trends
Long-term temporal trends in the annual wet deposition of ions were analyzed
using Sen's slope (Tables 3 and 4) and linear regression analysis. The annual
wet deposition of Ca2+ and K+ has not changed significantly at
almost all locations. For Mg2+and Na+, there were no significant
changes in the annual wet deposition rate at most of the sites, while a very
small statistically significant decline in the annual wet deposition was
observed at other sites. Of these sites, the rate of decline ranged from
-0.001 to -0.008 kg ha-1 a-1 for Mg2+ and -0.002 to -0.02 kg ha-1 a-1 for Na+. Decline in base cations has been reported
at other Canadian sites during the 1990s (Watmough et al., 2005). The small
decrease or lack of change in base cation wet deposition is expected because
the major source of base cations at rural Canadian sites are from natural
emissions (Watmough et al., 2005). Monitoring the wet deposition Ca2+,
Mg2+, K+, and Na+ trends are important because these ions
neutralize soil acidity and mitigate further harmful impacts to plants and
wildlife. A declining trend in the Cl- wet deposition was observed at 5
of 16 sites (Fig. 7a), and the magnitude ranged from -0.007 to -0.03 kg ha-1 a-1 (Table 3).
Significant trends in the annual wet deposition of NH4+ were
observed at only 3 of 16 locations. There were two sites with increasing
temporal trend (0.05 kg ha-1 a-1), whereas a decreasing trend was
found at one location (Table 4). A lack of an overall consistent temporal
trend in precipitation NH4+ was also found in previous analysis of
CAPMoN sites (Zbieranowski and Aherne, 2011). These trends were in contrast
to those in the USA, which has seen an increase in precipitation
NH4+ at 64 % of the wet deposition sites between 1985 and 2004
(Lehmann et al., 2007). While increasing NH4+ in precipitation
helps to increase precipitation pH and promote plant growth, a
counter-effect is that soils can become acidic when NH4+ undergoes nitrification (Vogt et al., 2006).
Declining trends in NO3- wet deposition was observed at 10 of
16 sites (data combined in Fig. 7a), while no significant trend was found at
other locations. The rate of decrease ranged from -0.07 to
-1.0 kg ha-1 a-1 and was largest at the southern Ontario and
Québec sites (Algoma, Longwoods, Egbert, Sprucedale, Frelighsburg, Lac
Edouard). One exception was the non-significant trend at Sutton (Table 4)
even though this site is only 15 km from Frelighsburg. The discrepancy in
the temporal trends at these two sites can be partially attributed to the
different measurement periods. Measurements of wet deposition at Sutton ended
in 2002, whereas Frelighsburg has been actively measuring wet deposition
since 2001. This suggests the rate of decrease in NO3- wet deposition
was more rapid in the period after 2001 and corresponds with the decline in
NO3- air concentrations over the same time period discussed earlier.
In the US northeast, the decline in precipitation NO3- was observed
at only 25 % of the sites in that region during 1985–2004 (Lehmann et
al., 2005). A recent study of NO3- wet deposition from 1985 to 2011
across North America indicates a 40–50 % decrease in eastern North
America after 2000 (Lloret and Valiela, 2016). Similar to the air
concentrations of SO42-, a decline in SO42- wet deposition was
also prevalent throughout Canadian sites. This finding is consistent with the
decrease in precipitation SO42- reported at other Canadian sites
during the 1990s (Watmough et al., 2005) and at 89 % of the wet
deposition sites in the USA between 1985 and 2004 (Lehmann et al., 2007).
Decreasing temporal trends in SO42- wet deposition was found at 11 of
16 sites with magnitudes ranging from -0.1 to
-1.0 kg ha-1 a-1 depending on the location (Table 4). The
overall rate of decline was ∼ 2 times higher at the southern Ontario
and Québec sites relative to other locations (Fig. 7b), which is consistent
with the patterns in the air concentrations of SO42- described
earlier. The large declining trends in the wet deposition of nitrogen and
sulfur-containing species in conjunction with the relatively smaller declines
or lack of change in base cations indicate that acid rain has attenuated over
time. This is largely attributed to policies controlling NOx and
SO2 emissions.
Temporal trends of annual wet deposition of Cl- and
NO3- (a) and SO42- (b).
Geographical patterns of precipitation pH. See Fig. 6 caption.
Daily wet deposition of ions was correlated with their respective daily
particulate matter concentrations and daily precipitation amount to gain
insight into factors influencing the temporal trends in wet deposition.
Moderate correlations (r=0.38–0.41, p < 0.05) between daily wet
deposition and particulate matter concentration were found for
SO42-, Na+, and Cl-, whereas only weak correlations were
found for other ions. This result partly explains the prevalent decline in
wet deposition of SO42- and Cl-. Decreasing NO3-
wet deposition was also widespread; however, it did not strongly correlate
with particulate NO3- concentrations (r=0.21, p < 0.05). This is potentially because both gaseous and particulate nitrogen
species can contribute to NO3- wet deposition. This is supported
by the slightly higher correlation (r=0.33, p < 0.05) between
daily NO3- wet deposition and HNO3. For SO42- wet
deposition which can be attributed to the precipitation scavenging of
particulate SO42- and SO2, only a weak correlation was
found between daily SO42- wet deposition and SO2 (r=0.13, p < 0.05). Further analysis on the relative contributions of
gaseous and particulate species to NO3- and SO42- wet deposition will be discussed in Sect. 3.4. Moderate correlations
between daily wet deposition and daily precipitation were found for
NH4+, NO3-, and SO42- (r=0.50–0.56,
p < 0.05), while weaker correlations were found for other ions. Some
correlation was expected because wet deposition is determined from
precipitation concentration and precipitation amount. On an annual basis,
there has been no significant change to the annual precipitation amount at
most of the sites, except for significant increases at Bratt's Lake, ELA,
Chalk River, and Kejimkujik (Table 4). The lack of change to the annual
precipitation is inconsistent with the decreasing trends in SO42-,
NO3-, and Cl- annual wet deposition found at the majority of
the sites. Thus, long-term wet deposition of ions was not strongly
influenced by long-term precipitation trends between 1983 and 2010.
Acid rain
The geographical patterns in acid rain as measured by precipitation pH are
shown in Fig. 8. Precipitation pH is slightly acidic by nature due to the
presence of carbonic acid formed by the dissolution of CO2. A pH below 5
is considered to be acidic precipitation (Lehmann et al., 2007). According to
Fig. 8, the median pH in daily precipitation samples across the sites ranged
from 4.4 to 5.7. Between 1983 and 2011, acidic precipitation was observed in
more than 50 % of the daily precipitation samples at 19 of 31 or 61 %
of the sites. Acidic precipitation was prevalent in southern Ontario and some
parts of eastern Canada, whereas pH was above 5 in western and central
Canada. Similarly in regions close to southern Ontario and eastern Canada,
higher occurrences of acid rain have been observed in the US northeast region
during the 1994–1996 and 2002–2004 periods (Lehmann et al., 2007). Acid
rain has contributed to the acidification of soil and lakes (Stoddard et al.,
1999; Driscoll et al., 2001; Clair et al., 2002; Jeffries et al., 2003;
Watmough and Dillon, 2003). In the southern Ontario and eastern Canada
region, the acid deposition effects are more concerning because the soil is
naturally slightly acidic and shallow and the underlying bedrock provides
insufficient acid buffering capacity (Clair et al., 2002; Watmough and
Dillon, 2003). The geographical patterns in precipitation pH were consistent
with those of aerosol acidity discussed earlier. The correlation between the
median pH and aerosol acidity (as measured by c / a ratio) among
15 locations was 0.68, which suggests acidic particles partially contributed
to acid rain at various Canadian sites.
Between 1983 and 2011, an increasing trend in pH was observed at 13 of 16 sites, while no significant change in pH was found at the remaining three
sites (Table 4). The overall decline in acidic precipitation in Canada is
consistent with the trends in the USA (Lehmann et al., 2007) and H+
wet deposition trends between the 2000–2002 and 2005–2007 periods over most of
North America, Europe, and Africa (Vet et al., 2014). The rate of increase in
pH was slightly higher at southern Ontario and Québec sites (Table 4).
Recent studies indicate there has been a gradual improvement to soil and
surface water conditions due to decreases in NO3- and
SO42- wet deposition; however, this recovery has been outpaced
by the rate of decline in acidic wet deposition (Strock et al., 2014;
Lawrence et al., 2015). The increasing temporal trends in pH can be
partially attributed to aerosol acidity. A correlation of 0.29 between daily
pH and c / a was found in this study. Based on the annual trend between the
1994 and 2010 period, the annual average pH and c / a (data combined at 13 of
15 sites) had very similar trends (Fig. 5b) and the correlation coefficient
improved to 0.86.
Scavenging ratios
General statistics and comparisons with the literature
A summary of the monthly average scavenging ratio (W) (on a mass basis)
statistics for the inorganic ions and trace gases are provided in Table S3.
Monthly WCa ranged from 120 to 14 338. The minimum and maximum
values were found at Egbert and Algoma, respectively. WMg ranged
from 131 to 11 243. The lowest W was recorded at Chapais, while the
highest W was recorded at Kejimkujik. WNa ranged from 76 to
12 165. The lowest W was recorded at Chapais, while the highest W was
recorded at Kejimkujik similar to Mg2+. The WK ranged from
69 to 5565 and had lower values compared to Ca2+, Mg2+, and
Na+ because K+ is predominantly in fine particulate matter except
for an additional coarse mode found at coastal locations (Zhang et al.,
2008). For this reason, it is assumed that the W of fine PM
(WfPM) was equivalent to the WK for inland sites,
while WK/2 was assumed for coastal sites. The minimum and maximum
WK were found at Montmorency and Kejimkujik, respectively. Since
Ca2+, Mg2+, and Na+ are mainly associated with coarse
particulate matter, the average W of these ions was used as an estimate of
the W of coarse particles (WcPM). The range of WcPM
was 83–12 165. The minimum and maximum values were found at Chapais and
Kejimkujik, respectively. The WCl ranged from 210 to 35 521. The
minimum and maximum values were found at Chapais and Algoma, respectively.
Compared to other ions, WCl
was larger and had greater variability. Overall, the range in the average W
for these ions among the 13 sites was within the average W from previous
studies (Table S4).
Monthly scavenging ratios were determined for particulate NO3-
(pNO3-) and HNO3 separately because both gaseous and
particulate forms can contribute to NO3- wet deposition. Monthly
W ranged from 135 to 4272 for pNO3- and 7–16 658 for HNO3.
Based on the average scavenging ratio, WHNO3 was greater than
WpNO3. The average WpNO3 from 13 sites was within the range of
literature values in Table S4; however, the majority of the WpNO3 in
the literature are determined from total nitrate in precipitation and
pNO3- in air. Thus, most of the WpNO3 are overestimated.
Scavenging ratios of pNO3-based on total nitrate in precipitation
are higher by a factor of 1.4–18 depending on the site (average: factor of
6).
In this study, the average WHNO3 at some sites was higher
than those in Table S4; however, they were most comparable to the average
determined by Cadle et al. (1990) likely because of the similarity in the
methods of determining WHNO3. The method first calculates
the pNO3- scavenged and then the difference between the total
NO3- and the pNO3- scavenged is assumed to be contributed by
HNO3 scavenging. One major difference in the approach was that Cadle et
al. (1990) used WK as a surrogate for WpNO3 to
estimate the pNO3- scavenged, whereas in this study
WpNO3 considered the seasonal particle size distribution of
pNO3-. The average WHNO3 at some sites in this study
were different than those determined by Hicks (2005) (Table S4), who assumed
only HNO3 contributed to NO3- wet deposition. The values were
also different from that of Kasper-Giebl et al. (1999), who used a multiple
linear regression (MLR) approach. For comparison purposes,
WpNO3 and WHNO3 derived from MLR are also
shown in Tables S4 and S5. The MLR results show
WHNO3 > WpNO3. However, this
empirical method generated higher WpNO3 and lower
WHNO3 at some locations compared to the method used in this
study. Table S5 indicates that the MLR model fit was considered weak to
moderate (R2=0.15–0.34) depending on the site.
Monthly WpNH4 ranged from 63 to 4356. The lowest W was
recorded at Egbert, while the highest W was recorded at Algoma. Scavenging
ratios of NH3 were undetermined because NH3 air concentrations were
not available. Average WpNH4 was also lower than some of
the values in the literature (Table S4), which are likely overestimated
because the values were based on the total NH4+ precipitation
concentrations, instead of pNH4+ scavenged by precipitation.
Comparison of these two methods of calculation indicates that the use of
total NH4+ precipitation concentration overestimated the scavenging
ratios by 4–48 % (average: 22 %) depending on the location. Despite
the coexistence of NH4+ and SO42- in the atmosphere, the
difference between the average scavenging ratios of these ions can vary by
4–98 % (average: 32 %). The range of monthly scavenging ratios for
pSO42- and SO2 were 75–3146 and 0.3–12 068, respectively. The
average WpSO4 among the sites in this study were lower than
some of the literature values in Table S4. Most of the studies excluded the
wet scavenging of SO2 because pSO42- was assumed to be the
dominant contributor to SO42- wet deposition. This method
overestimates the scavenging ratio of pSO42- by 18–85 %
(average: 44 %) compared to the method used in this study. According to
the limited number of WSO2 estimates (Table S4), the
precipitation scavenging of SO2 is less important compared to
pSO42- because of the lower WSO2. In this study,
SO2 and pSO42- can be equally important at times in terms of
scavenging ratios. The average WSO2 at more than half of the
sites in this study was greater than literature averages. This is
potentially due to the different methodologies for calculating
WSO2 and precipitation type. The MLR method yielded higher
WpSO4 and lower WSO2 at some locations
compared to the method used in this study (Tables S4 and S5). The approach
used in this study was similar to Cadle et al. (1990); however, that study
used WNH4 as a surrogate for WpSO4 based on
the assumption these ions are typically found in the same aerosols. There
are,
however, large uncertainties with WNH4 because of the
scavenging by both pNH4+ and NH3. As mentioned earlier, most of
the WNH4 in literature are overestimated because of the
exclusion of NH3. Use of these values could lead to a high bias in
WpSO4 and subsequently lower WSO2. Aside
from the methodology, the scavenging ratios determined by Cadle et al. (1990)
were based on snowfall events, which favor the scavenging of particles over
gases (Hicks, 2005; Zhang et al., 2013, 2015).
The variability in scavenging ratios among different ions and trace gases
within the same month is shown in Fig. S4. The range in scavenging ratios
can be quite large. This is expected because the different physical and
chemical properties of the pollutants affect their wet scavenging
efficiencies. For particulate matter, coarse particles (e.g., Ca2+,
Mg2+, and Na+) are scavenged more efficiently than fine particles
(e.g. K+) (Galloway et al., 1993; Guerzoni et al., 1995; Tuncel and
Ungör, 1996). The different solubilities of gaseous pollutants
(HNO3 and SO2) can also explain the differences in scavenging
ratios for different pollutants. WHNO3 values were 1.4 to 6 times higher than
those of WSO2 depending on the location (Table S3), which is consistent
with the higher solubility of HNO3 compared to SO2 (H=2.1×105 M atm-1 vs. 1.2 M atm-1; Zhang et al., 2006; Sander,
2015).
Variations in scavenging ratios
Although the monthly scavenging ratios of most ions span a wide range, the
average scavenging ratios of particulate ions were within a factor of
1.6–4.5 among the 13 sites (Table S3). The average scavenging ratios of
Na+, Cl-, and Ca2+ have larger spatial variability than other
particulate ions. This may reflect the different precipitation scavenging
efficiencies of particles in the marine boundary layer and continental
atmospheres as hypothesized by Galloway et al. (1993). One related theory is
that the higher relative humidity in marine environments is conducive to the
hygroscopic growth of sea-salt aerosols which increases its scavenging
efficiency (Hennigan et al., 2008). The average scavenging ratios ranged
from 497 to 996 for fine particles (factor of 2 spatial variability) and
666 to 2077 for coarse particles (factor of 3 spatial variability). Larger
scavenging ratios of fine particles were found at inland sites, where soil
and biomass emissions are sources of K+. For coarse particles, the
larger scavenging ratios at coastal sites (Table S3) were likely attributed
to oceanic source of Na+ and Mg2+. The greatest spatial
variability was in the scavenging ratios of gases including HNO3 and
SO2. The average scavenging ratios varied by a factor 7.4 and 10.7,
respectively. The large variability in scavenging ratios between sites is
expected because the air and precipitation concentrations and precipitation
type and amounts among other factors can also vary with location.
A pronounced seasonal variation can be seen in the monthly average scavenging
ratios. The average scavenging ratio of most of the ions and HNO3 were
lowest during July or August (Fig. S5), which resembles the monthly
NO3-, SO42-, and NH4+ variations in a previous study
(Kasper-Giebl et al., 1999). There were two periods when the average
scavenging ratios peaked: one peak during April–May and a second peak during
September–October were observed for most of the ions and HNO3. A
similar pattern was obtained for WpNO3 when MLR was used to
generate monthly scavenging ratios. However, different patterns were obtained
for WHNO3 and WpSO4 derived from MLR, as
shown by the higher values during winter and lower values in the warm seasons
(Fig. S5 and Table S6).
Monthly variations in WSO2 were different from those of
other ions and HNO3. The average WSO2 was lower during
winter and peaked in the summer, which is supported by the MLR results as
well (Fig. S5 and Table S6). This result is also consistent with the seasonal
WSO2 patterns from multiple US sites (Hicks, 2005) and other
studies suggesting the inefficient scavenging of SO2 by snow
(Kasper-Giebl et al., 1999, and references therein). However, limited field
measurements of dissolved SO2 (measured as sulfite) in precipitation
samples indicate that the highest precipitation concentrations were found in
the colder months, which is consistent with solubility theory (Hales and
Dana, 1979; Dana, 1980). This finding does not necessarily contradict the
results from this study because WSO2 also depends on the air
concentration. The higher ambient SO2 likely due to higher combustion
emissions associated with winter heating (Fig. S6) resulted in lower
scavenging ratios, and vice versa during warmer months. Besides temperature
effects on solubility, precipitation pH and the presence of NH3 and
H2O2 could also affect SO2 wet scavenging (Zhang et al.,
2006).
Most of the long-term scavenging ratio trends were not statistically
significant according to the seasonal Kendall test, but some statistically
significant trends were found at a few locations and for some nitrogen and
sulfur species. At Longwoods, there was a statistically significant
declining trend in the scavenging ratio of pNO3-; however, the
magnitude of the trend was only -6.3 (< 1 %) per year, which is
small compared to the scavenging ratio (values in the hundreds to
thousands). At Algoma, a significant increasing trend in the scavenging
ratio of pSO42- was found with a slope of +11.4 (1.4 %) per
year. At many of the sites, the lack of long-term trends in the scavenging
ratios of sulfur and nitrogen species reflect the decreasing trends in both
wet deposition and air concentrations (Tables 2 and 4). There are also many
factors that can affect the precipitation concentrations, such as particle
sizes, air concentrations, rainfall intensity, and precipitation and cloud
types, which vary geographically and could change over time.
Average percent contributions of gas- and particulate-phase species
to nitrate (a), ammonium (b), and sulfate (c) wet
deposition.
Monthly variation in the contributions of gas- and particulate-phase
species to nitrate (a), ammonium (b), and
sulfate (c) wet deposition. Error bars represent the standard
deviation of the percent contributions by gas- and particulate-phase species
between sites.
Relative contributions of particulates and gases to nitrate,
ammonium,
and sulfate wet deposition
In the previous section, scavenging ratios were determined for particulate
(pNO3-, pNH4+, pSO42-) and gaseous (HNO3
and SO2) species. In this section, the relative percent contributions
of these particulate and gaseous species to nitrate, ammonium, and sulfate
wet deposition are determined. The average ±1σ of
pNO3- contributions to nitrate wet scavenging
(%pNO3-) was 28 ± 23 % for all years of data at the 13 locations. Percent HNO3 contributions to nitrate wet scavenging
(%HNO3) was 72 ± 23 %. Based on the average, %HNO3
dominated %pNO3- at most of the sites (Fig. 9a). Geographical
variations were observed in the %pNO3- and %HNO3.
Average %pNO3- were higher at the two lowest latitude sites and
coastal sites (Fig. 9a). One reason is because of the higher pNO3-
air concentrations at the lower-latitude locations (Longwoods and Egbert)
discussed in Sect. 3.1.1. %pNO3- were also higher at the
coastal locations (Saturna and Kejimkujik) likely because of the
partitioning of HNO3 to sea-salt aerosols (Pryor and Sørensen, 2000;
Fischer et al., 2006), which are typically coarse particles and hygroscopic
and hence more efficiently removed by precipitation. In contrast,
%HNO3 were greater at the higher latitude continental locations
(Fig. 9a). Relative contributions of pNH4+ and NH3 were
70 ± 19 % and 30 ± 19 %, respectively. Precipitation
scavenging of pNH4+ was greater than NH3 at all sites
(Fig. 9b). The percent pSO42- contributions to sulfate wet
scavenging (%pSO42-) was 63 ± 20 %, while percent
SO2 contributions to sulfate wet scavenging (%SO2) was
37 ± 20 %. Average %pSO42- were greater than that of
%SO2 at most of the sites (Fig. 9c). No pronounced geographical
patterns were observed in the relative contributions of gases and
particulates to ammonium or sulfate wet deposition. Knowledge of the
relative scavenging contributions of gases and particles may improve the wet
deposition modeling of nitrate, ammonium, and sulfate, which continues to
show discrepancies between model and observations (Appel et al., 2011; Zhang
et al., 2012; Kajino and Aikawa, 2015; Qiao et al., 2015). Furthermore, gas-
and particulate-phase pollutants may not come from the same source, since
particulate matter can be re-emitted from natural sources and human
activity. Aerosol formation is modeled separately in chemical transport
models and the wet deposition of particles and gases use different
parameterizations in these models. The scavenging coefficient of gases in
models depends on Henry's law constant or gas diffusivity and reactivity,
whereas the scavenging coefficient of particles is a function of particle
size distribution and collection efficiency among other factors (Gong et
al., 2011). Studies also suggest different efficiencies between rainout and
washout scavenging mechanisms for gases and aerosols (Gong et al., 2011;
Kajino and Aikawa, 2015). These parameterizations can differ between
different chemical transport models as well leading to large uncertainties
in the wet deposition estimates (Tost et al., 2007). Gas/aerosol wet
scavenging observations can also be used to evaluate those from wet
deposition simulations (Kajino and Aikawa, 2015).
A seasonal variation was observed in the relative contributions of gases and
particulates to nitrate, ammonium, and sulfate wet deposition. The
contributions by particulates to nitrate, ammonium, and sulfate wet
scavenging were greater during cold months and lower during summer (Fig. 10). This pattern is consistent with studies suggesting that particle
scavenging by snow is more efficient than the scavenging by rain for an
equivalent amount of precipitation (Zhang et al., 2013, 2015). Based on
scavenging ratios, Zhang et al. (2015) found that snow scavenging can be 10 times more efficient than the rain scavenging of polycyclic aromatic
compounds (PAC), and that snow scavenging of particulate-phase PAC can
exceed that of gas-phase PAC by a similar magnitude. In contrast to particle
scavenging, greater precipitation scavenging of gases was observed in the
warm seasons (Fig. 10). This inverse relationship between particle and gas
wet scavenging resulted because of Eq. (3), which assumes that the
precipitation scavenging in excess of particle wet scavenging was due to gas
scavenging. This assumption needs to be validated by independently deriving
the gas scavenging contributions; however, the results based on the
assumption are consistent with precipitation scavenging theories.
In terms of nitrate scavenging, the largest difference between gas and
particulate wet scavenging were observed during warm months (factor of 4
higher for HNO3), whereas smaller differences were seen during cold
months (Fig. 10a). The large contribution by HNO3 is expected because
it is one of the most soluble gases and is effectively scavenged by rainout
(Chang, 1984; Garrett et al., 2006). %HNO3 were also fairly high
during the cold months because of higher solubility at lower temperatures
and the high absorption and retention of HNO3 and other strong acids
on ice crystals (Diehl et al., 1995; Clegg and Abbatt, 2001). Snow
scavenging of HNO3 can also exceed below-cloud rain scavenging of
HNO3 for an equivalent precipitation rate (Chang, 1984).
Particle wet scavenging exceeded gas scavenging contributions to ammonium
wet deposition in most months by a factor of 2.6 except during May–June
(Fig. 10b). For sulfate scavenging, larger differences between gas and
particulate wet scavenging were found during cold months (factor of 2 higher
for pSO4-), while small differences were observed during warm
months (Fig. 10c). The greater scavenging of pSO4- during colder
months is likely attributed to the effectiveness of particle scavenging by
snow as discussed earlier. Another explanation for the large disparity
between particle and gas scavenging in the cold months could be the low
absorption of SO2 by ice crystals, especially on low pH ice surfaces
(Clegg and Abbatt, 2001). As for the smaller difference between particle and
gas scavenging during warm months, experiments suggest that snow scavenging
of SO2 can be increased by the presence of H2O2 and
relatively higher temperatures of ∼ 0 ∘C (Mitra et al.,
1990). The latter conditions are conducive to the formation of a
quasi-liquid layer on the ice surface, which may increase the dissolution of
SO2 (Clegg and Abbatt, 2001).
Conclusions
Long-term air concentrations, wet deposition, and scavenging ratios of
inorganic ions were analyzed using CAPMoN data. Geographical variability in
the air concentrations of inorganic ions can be attributed to proximity of
the sites to anthropogenic sources, oceanic sea-salt emissions, and
agricultural emissions. Annual wet deposition geographical patterns for
Ca2+, Na+, Cl-, NH4+, NO3-, and
SO42- were similar to those in air. Widespread declines were
observed for NH4+ (1994–2010) and SO42- (1983–2010) in
air, which was attributed to decreases in SO2, NOx, and local
NH3 emissions. NO3- air concentrations increased from
1991 to 2001 and then decreased from 2001 to 2010, consistent with the trends in
NOx emissions in Canada over these two decades and in the USA over the
last decade. However, widespread declines in annual wet deposition were only
found for NO3- and SO42- from 1984 to 2011. SO42- air concentrations and annual wet deposition
declined ∼ 2 times faster in southern Ontario and southern Québec than other locations
because of the proximity of the sites to industrial emission sources.
Aerosol acidity and acid rain had greater impacts on southern and eastern
Canada than western Canada. Temporal trends show aerosol acidity and acid
rain have been decreasing simultaneously from 1994 to 2010, consistent with
large declines in nitrate and sulfur species and slight declines or lack of
change in base cations.
Scavenging ratios of particulate NH4+, SO42-, and
NO3- in the literature may be overestimated on average by 22 %,
44 %, and a factor of 6, respectively, because the wet scavenging of
gases was excluded. The wet scavenging of HNO3 dominated particulate
NO3- at most locations, while the wet scavenging of particulate
NH4+ and SO42- was more efficient than NH3 and SO2,
respectively. The wet scavenging of particles was more efficient in the cold
months likely because of the scavenging by snow. Greater gas scavenging was found in the warm months compared to the cold months, which was opposite in trend to particulate wet scavenging.
Long-term trends in inorganic ions provide greater insight into which
Canadian regions are still susceptible to or likely recovering from acid
rain impacts, and the effectiveness of environmental policies at mitigating
acid rain. Particulate inorganic ions and trace gas scavenging ratios
provide a measure of the wet scavenging efficiencies and are potentially
useful surrogates for the wet scavenging of other pollutants provided that
they have similar physicochemical properties (e.g. solubility, particle
sizes). These results can be considered in future wet deposition
modeling to improve the prediction of nitrate, ammonium, and sulfate wet
deposition. Scavenging ratios can potentially be used to obtain a rough
first-order estimate of the wet deposition at other locations considering
the uncertainties in both the scavenging ratios and in wet deposition
modeling.