Heavy metals and polycyclic aromatic hydrocarbons (PAHs) can
greatly influence biotic activities and organic sources in the ocean.
However, fluxes of these compounds as well as their fate, transport, and net
input to the Arctic Ocean have not been thoroughly assessed. During
April–November of the 2016 “Russian High-Latitude Expedition”, 51 air
(gases, aerosols, and wet deposition) and water samples were collected from the
Russian Arctic within the Barents Sea, the Kara Sea, the Laptev Sea, and the East
Siberian Sea. Here, we report on the Russian Arctic assessment of the
occurrence of 35 PAHs and 9 metals (Pb, Cd, Cu,
Co, Zn, Fe, Mn, Ni, and Hg) in dry and wet deposition as well as the atmosphere–ocean fluxes of 35
PAHs and Hg0. We observed that Hg was mainly in the gas phase and that
Pb was most abundant in the gas phase compared with the aerosol and
dissolved water phases. Mn, Fe, Pb, and Zn showed higher levels than the
other metals in the three phases. The concentrations of PAHs in aerosols and
the dissolved water phase were approximately 1 order of magnitude higher
than those in the gas phase. The abundances of higher molecular weight PAHs
were highest in the aerosols. Higher levels of both heavy metals and PAHs
were observed in the Barents Sea, the Kara Sea, and the East Siberian Sea, which
were close to areas with urban and industrial sites. Diagnostic ratios of
phenanthrene/anthracene to fluoranthene/pyrene showed a pyrogenic source for
the aerosols and gases, whereas the patterns for the dissolved water phase
were indicative of both petrogenic and pyrogenic sources; pyrogenic sources
were most prevalent in the Kara Sea and the Laptev Sea. These differences
between air and seawater reflect the different sources of PAHs through
atmospheric transport, which included anthropogenic sources for gases and
aerosols and mixtures of anthropogenic and biogenic sources along the
continent in the Russian Arctic. The average dry deposition of ∑9 metals and ∑35 PAHs was 1749 and 1108 ng m-2 d-1, respectively. The average wet deposition of ∑9 metals and ∑35 PAHs was 33.29 and
221.31 µg m-2 d-1, respectively. For the atmosphere–sea
exchange, the monthly atmospheric input of ∑35 PAHs was estimated
at 1040 t. The monthly atmospheric Hg input was approximately 530 t.
These additional inputs of hazardous compounds may be disturbing the
biochemical cycles in the Arctic Ocean.
Introduction
The increasing anthropogenic activities associated with growing industries
within boundary areas of the Arctic for economic reasons, including
hydrocarbon exploration sites and mines in the Russian Arctic, represent
potential pollution sources to Arctic ecosystems (Walker et al., 2003;
Dahle et al., 2009; Ji et al., 2019). Additionally, the Arctic has long been
contaminated by pollutants transported to polar areas from distant locations
outside of this region (Hung et al., 2016). For example,
anthropogenic sources of pollutants in the Arctic have been found to come
from the Norilsk industrial area on the Taymyr Peninsula (Reimann et al.,
1997; Zhulidov et al., 2011) and from the copper–nickel mining industry on
the Kola Peninsula (Boyd et al., 2009; Jaffe et al., 1995). For
pollutants transported from outside of the Arctic, reducing global emissions
would be an ideal strategy to lessen the impacts of pollutants on Arctic
ecosystems. For example, worldwide emissions of mercury will have increased
by 25 % in 2020 over 2005 levels according to previous estimations (Pacyna et al., 2010). Mercury is a key problematic
pollutant in the Arctic because it is a neurotoxic pollutant significantly
influencing northern latitudes via human exposure from eating seafood
and marine mammals (Stow et al., 2015). Thus, global emission
reductions could help to alleviate problems associated with long-range
mercury transport and contamination in the Arctic. In regard to sources
close to the Arctic, these may inevitably cause localized ecological risks
or risks over a wider regional range. For instance, Fernandes
and Sicre (1999) showed that atmospheric transport of anthropogenic
polycyclic aromatic hydrocarbons (PAHs) to the Eurasian Arctic mainly
originated from eastern Europe and Russia. PAHs in aerosols from lower
latitudes were deposited on soils and ice in winter and transported by rivers
to the ocean by the occurrence of freshet (Fernandes and Sicre,
1999). The previous study also showed a strong net deposition in the marine
transect from East Asia to the Arctic, and the controlling sources both
contained East Asia as a potential continental source region and forest fires in the Arctic as a seasonal and regional source (Ma
et al., 2013). In addition, high concentrations of heavy metals (Mn, Zn, Ni,
Fe, and Cd) were observed in the west Arctic Ocean (Chukchi Sea); this
enrichment was not only from Pacific-origin inflow water from the Bering
Strait but also from additional sources such as melting sea ice and river
water discharge (Kondo et al., 2016). Also of concern is the fact
that the melting of contaminated
ice may lead to more pollutant emission into the Arctic Ocean with rapid warming of the global climate, which could
harm its fragile ecosystems.
Pollutants can be transported to the Arctic through both seawater and
atmospheric pathways; the atmospheric pathway is the quickest and most
direct way for long-range pollutant transportation, e.g., pollutants can be
transported from distant sources to the Arctic within several days or weeks
(Shevchenko et al., 2003). Reports have revealed that some
pollutants such as heavy metals and polycyclic aromatic hydrocarbons (PAHs)
can be transported with aerosols over thousands of kilometers to Arctic
regions (Rahn and Lowenthal, 1984; Maenhaut et al., 1989; Shaw, 1991;
Cheng et al., 1993). Approximately 100 t of airborne mercury originating
from industrial sources is deposited in the Arctic Ocean annually
(Valenti, 2006). While there is evidence that atmospheric inputs make
large contributions to the chemical budgets in marine areas, the exact role
of these inputs in the Arctic Ocean remains uncertain and may have been
previously underestimated (Duce
et al., 1991). Numerous studies have shown that aerosol transport is
essential to transfer atmospheric compounds from air to ocean, and that this
process is susceptible to changes in the climate of Arctic regions (Leck
et al., 1996; Sirois and Barrie, 1999; Bigg and Leck, 2001). The compounds
in aerosols over the Russian Arctic have been reported to show maximal
concentrations during the winter/spring season; in addition, 50 % of the
air pollutants were found to have originated from Russian Arctic pollution
(Shevchenko et al., 2003). It has also been reported that the
natural biodegradation rates of exogenous compounds in the Arctic Ocean
could be lower than those in more temperate oceans such as the Atlantic and
Pacific (Bagi et al., 2014). In addition, Vieira
et al. (2019) found that Fe, Mn, and Co were predominantly controlled by
reductive benthic inputs, and that their levels were affected by the
biological processes of uptake and release in the Arctic Ocean. Due to
their toxicity and persistence, high concentrations of heavy metals or other
persistent pollutants such as PAHs may disturb the benthic fluxes in
cross-shelf mixing in Arctic regions, which could result in adverse effects
on marine life and, with the eventual biomagnification in the food web, on
humans as well. However, the long-term influence of heavy metals and PAHs on
biogeochemical cycles in the Arctic Ocean remains poorly understood.
Atmosphere–seawater exchange is the main process that controls the
residence time and levels of chemical compounds in the Arctic Ocean. In
particular, atmospheric deposition is a significant source for pollutants in
seawater, and dry deposition in the ocean has been widely studied
(Jickells and Baker, 2019; Wang et al., 2019; Park et al., 2019).
Although wet deposition (precipitation scavenging) is regarded as playing a
predominant role in eliminating pollutants in both gas and particulate
phases, current reports on the spatial distribution of pollutants from
wet (snow) deposition in high-latitude oceans are scarce
(Custódio et al., 2014). Moreover, for volatile or
semivolatile compounds, the volatilization process is an important pathway
for atmosphere–seawater exchanges. Therefore, the atmosphere–water
exchange of volatile or semivolatile compounds can be estimated by the net
flux of pollutants either volatilizing from seawater to air or depositing
from air to seawater (Rasiq et al., 2019; Cheng et al., 2013; Totten et
al., 2001). Gonzalez-Gaya et al. (2016) reported on a global
assessment of atmosphere–ocean fluxes of 64 PAHs; the net atmospheric PAH
input to global ocean was 0.09 Tg per month. The atmosphere–seawater
exchange rate is greatly influenced by atmospheric temperature variations,
and the direction and magnitude of fluxes of compounds between air and
seawater vary seasonally (Bamford et al., 1999; Hornbuckle et al., 1994).
Additionally, inorganic salt ions can decrease the aqueous solubility of
organic compounds such as PAHs (Rasiq et al., 2019). During the
melting of sea ice in the Arctic Ocean, the magnitude and direction of
atmosphere–seawater fluxes may be different from those in tropical and
subtropical oceans (Gonzalez-Gaya et al., 2016; Rasiq et al., 2019). The
Arctic Ocean is considered as a sink that receives global airborne
pollutants (Environment Canada, Fisheries and Oceans Canada and Indian and Northern Affairs Canada Arctic, 2008); however, the fate of
atmosphere–ocean exchange of trace metals and organic compounds remains
unclear.
In this study, two categories of pollutants (i.e., 9 heavy metals and 35
PAHs) were measured in the Arctic Ocean, in aerosols, gas, and seawater, and
atmosphere–ocean exchanges of Hg and PAHs were studied. We hypothesized
about the relative equilibrium of chemical exchanges between seawater and
air and calculated the net diffusion of atmosphere–ocean exchange of Hg and
PAHs in the Arctic Ocean for an evaluation of the double-directional
exchange. Meanwhile, the dry and wet deposition of heavy metals and PAHs in
the Russian Arctic Ocean were determined. The distributions of heavy metals
and PAHs in each sea of the Arctic Ocean and in various phases were also
characterized to identify possible sources from the continents.
Materials and methodsStudy area and sample collection
All samples were collected during the period from 9 April to 10 November 2016
as part of the “Russian High-Latitude Expedition” carried out on the
Mikhail Somov vessel (this vessel traveled from the city of Arkhangelsk to Wrangel
Island). A total of 51 air and water samples, and 8 wet deposition samples
were gathered from locations ranging from the southern inlet of the Barents
Sea (from west sites to Vaygach Island) to across the Kara Sea (to Gerkules
Island), Laptev Sea (to Bennett Island), and East Siberian Sea (to Wrangel
Island) (Fig. 1).
Locations of investigated islands for soil sampling and trajectory
of the vessel in the Russian Arctic.
Aerosol and gas phase
Air samples, including aerosols and concurrent gases as described elsewhere
(Reddy et al., 2012; Shoeib and Harner, 2002; Galarneau et al., 2017;
Grosjean, 1983; Wu, 2014), were collected by a high-volume sampler set up at
the top of a main rod. A wind vane was connected to the high-volume sampler
so that samples could be collected only if the wind was derived from the bow
to prevent contamination from ship emissions. The average sampled air volume
was 632 m3 (412–963 m3) per sample. The aerosols were sampled on
Teflon filters (P0325-100EA, Fluoropore, Darmstadt, Germany), and the
compounds in the gas phase were then collected over pre-cleaned polyurethane foams
(PUFs). After sampling, the filters and PUFs were tightly covered with
aluminum foil for air-tightness, then immediately placed in polyethylene
bags, and frozen at -20∘C prior to chemical analyses.
Wet deposition and water
Wet deposition samples were collected through a cleaned stainless steel
funnel connected to a glass bottle during eight snow events. Snowfall
samples were melted thoroughly at room temperature. Water samples were
gathered continuously from surface seawater (at a depth of 5 m) along the vessel,
and these samples were immediately filtered onto borosilicate microfiber
glass filters (AP1504700, MilliporeSigma, Darmstadt, Germany). Then, the
compounds in the dissolved phase were retained on XAD sorbent tubes
subjected to controlled flows. The mean filtered water volume was 1239 mL
(135–2876 mL). The XAD tubes were stored at 5 ∘C before their
extraction in the laboratory.
Heavy metal extraction and analysis
For metal determinations in the aerosol, gas phase, wet deposition, and
water samples, Teflon filters, PUFs, and dissolved phases were first
Soxhlet-extracted for 8 h using HNO3. The samples were then diluted
with deionized water to 23 mL and subjected to inductively coupled plasma
mass spectrometry (ICP-MS) analysis. Specifically, the contents of Pb, Cd,
Cu, Co, Zn, Fe, Mn, Ni, and Hg were analyzed on an ICP-MS instrument (Thermo
Scientific ICE 3500, Waltham, MA, USA) while making use of rhodium (Rh) as
an internal standard. High-resolution (10 000) data were collected to avoid
any mass interference problems.
PAH extraction and analysis
For PAH determinations in the gas, aerosol, and dissolved phase samples,
published procedures were used (Berrojalbiz et al., 2011; Castro-Jimenez
et al., 2012; Gonzalez-Gaya et al., 2014). Snow-melt water was extracted by
using solid phase Oasis HLB (3 cc/60 mg) cartridges on board. Briefly,
cartridges were preconditioned with 5 mL methanol, 10 mL of a mixture of
methanol:dichloromethane (1 : 2), and 10 mL deionized water. Afterward, each
sample was combined with a recovery standard and concentrated by N2
until near dryness. Then, it was eluted with 5 mL hexane, 5 mL of a mixture
of hexane:dichloromethane (1 : 2), and 10 mL deionized water.
A total of 35 PAH species were quantified, including naphthalene,
methylnaphthalene (sum of two isomers), 1,4,5-trimethylnaphthalene,
1,2,5,6-tetramethylnaphthalene, acenaphthylene, acenaphthene, fluorene,
dibenzothiophene, anthracene, 9-methylfluorene, 1,7-dimethylfluorene,
9-n-propylfluorene, 2-methyldibenzothiophene, 2,4-dimethyldibenzothiophene,
2,4,7-trimethyldibenzothiophene, 3-methylphenanthrene,
1,6-dimethylphenanthrene, 1,2,9-trimethylphenanthrene,
1,2,6,9-tetramethylphenanthrene, fluoranthene, pyrene, benzo[a]anthracene,
chrysene, 3-methylchrysene, 6-ethylchrysene, 1,3,6-trimethylchrysene,
benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[a]pyrene, perylene,
dibenzo[a,h]anthracene, indeno[1,2,3-cd]pyrene, dibenzo[a,h]anthracene, and
benzo[g,h,i]perylene. PAH quantification was performed by gas chromatography–mass
spectrometry (GC-MS). Specifically, we used a gas chromatograph coupled with
a triple quadrupole mass selective detector (GS-MS, ITQ 1100, Thermo
Scientific, USA) equipped with a DB-5ms chromatographic capillary column (30 m × 0.25 mm i.d. and 0.25 µm film, Agilent Technologies, Santa
Clara, CA, USA) operating in electron impact (EI) mode and with selected ion
monitoring (SIM) as reported previously (Gonzalez-Gaya et al., 2014).
Internal standards (anthracene-d10, p-terphenyl-d14,
pyrene-d10, and benzo[b]fluoranthene-d12) were added before operating
the GC-MS instrument for the quantification of PAHs, and the recovery of
perdeuterated standards (acenaphthene-d10, chrysene-d12,
phenanthrene-d10, and perylene-d12) was determined by addition
prior to the procedures of extraction; these values were then used for the
correction of measured concentrations.
Quality assurance and quality control
Analyses of every sample and phase were conducted in the laboratory with
field blanks to determine the analytical limits and recoveries.
Breakthroughs of aerosols and gas phases were checked for the Teflon filter
and PUF samples. Approximately 90 % of the metals and PAHs were obtained
during the first half of the sample analysis, while the remaining 10 %
were obtained during the second half; for the PAHs, these mostly consisted
of compounds with two to three rings. Six blanks (field and laboratory) were
collected for the gas phase, while seven field banks and eight laboratory
blanks were used for the dissolved phase, all of which were extracted along
with the rest of the samples during the analytical procedure. For the gas
phase, average ∑ metal values were approximately 0.049 and 0.052 ng per
sample in the field and laboratory blanks, respectively, and average ∑ PAH values were approximately 2.44 and 2.06 ng per sample in the field and
laboratory blanks, respectively (Tables S1 and S2 in the Supplement).
For the aerosols, average ∑ metal values were 0.046 and 0.065 ng per
sample in the field and laboratory blanks, respectively, and average ∑ PAH values were 2.95 and 2.96 ng per sample in the field and laboratory
blanks, respectively. Similarly, for the dissolved phase, values of 0.053 and
0.052 ng per sample were obtained for the ∑ metals and values of 2 and
1.73 ng per sample were obtained for the ∑ PAHs. All measured PAHs from
field samples exceeded the field and laboratory blank concentrations;
therefore, the quantified compounds did not subtract the blank values. Mean
recoveries of perdeuterated standards used as surrogates in dissolved
samples were as follows: 63 % for acenaphthene-d10, 54 % for
chrysene-d12, 73 % for phenanthrene-d10, and 82 % for
perylene-d12.
All concentrations in each medium were corrected by the surrogate recovery
for individual samples. The detection limit was used for the lowest limit of
the calibration curve. The quantification limit was equivalent to the
average blank concentration for each phase.
Data processing
Dry deposition fluxes (FDD, ng m-2 d-1) were calculated from
field measurements of trace metals and PAHs collected during the expedition
over eight time periods. Aerosol deposition fluxes for the metals were
calculated as follows:
FDD(metal)=CdVd,
where Cd is the concentration of atmospheric aerosols and Vd is
the velocity of deposition (m s-1). Vd was calculated as shown in Eq. (2), and the details have been described elsewhere (Zhang et
al., 2001):
Vd=ugrav+1Ra+Rs,
where ugrav is the gravitational settling velocity, and Ra and
Rs are the aerodynamic resistance for gaseous species and the surface
resistance, respectively. Rs can be calculated as follows:
Rs=1ε0u∗EBEIMR1,
where ε0 is an empirical constant (ε0=3)
and u∗ is the friction velocity calculated for gases. EB is
the collection efficiency of Brownian diffusion as a function of the Schmidt
number Sc:
EB=(Sc)-γ,
where γ is an empirical constant (γ=0.5). EIM is
the collection efficiency from impaction based on the following formulas
(Peters and Eiden, 1992):
5EIM=st0.8+st2,6st=VgravV∗2Va,
where Va is the kinematic viscosity for air (m2 s-1). The
correction factor (R1) is the fraction of particles close to the
surface:
R1=exp-st1/2.
For PAHs, the specific compound deposition velocity (Vd, cm s-1)
was derived from an empirical parameterization (Gonzalez-Gaya et al.,
2014):
logVd=-0.261logPL+0.387U10Chls-3.082,
where PL is the subcooled liquid vapor pressure of each PAH, U10
is the 10 m height wind speed, and Chls is the concentration of surface
chlorophyll. With Eq. (8), one can estimate the Vd for each PAH and
sampling period by taking PL from references and using the
field-measured U10 and Chls. In this study, FDD(PAH) values
were estimated from the measured concentrations in the aerosol phase
(CA, ng m-3) as follows:
FDD(PAHs)=864VdCA,
where 864 is the unit conversion factor.
The wet deposition fluxes (FW, ng m-3 d-1) of metals/PAHs
were estimated using the quantified concentrations of metals/PAHs from
the collected snow and the precipitated volume of snow-melt water per
surface and time period for each of the eight snow events during the
expedition.
The air–water diffusive fluxes (FAW, ng m-2 d-1) for Hg/PAHs
were calculated according to Fick's law:
FAW=KAWCGH′-1000CTW,
where CG and CTW represent the concentration measured in the gas
phase (ng m-3) and dissolved phase (ng L-1), respectively. H′
is the temperature dependence of Henry's law constant, and H′ values
for PAHs were taken from Bamford et al. (1999); H′ for Hg was
calculated from Eq. (11) for seawater (Andersson et al., 2008):
H′=exp-2404.3T+6.92,
where T is the temperature of the surface water (K). H′ was corrected
by the field-measured salinity. KAW represents the air–water mass
transfer rate (m d-1) calculated by a two-film model (Singh and Xu, 1997) and while considering the nonlinear
wind-speed effect. CTW for Hg was directly measured concentrations, and
CTW values for PAHs were calculated using the measured
concentrations in the dissolved phase (CW) as follows:
CT=CW1+kDOCDOC,
where KDOC was taken as the value of 10 % of the octanol–water
partitioning coefficient (KOW) (Burkhard, 2000), and DOC
represents the dissolved organic carbon (mg L-1).
KAW was calculated by the two-film model:
1KAW=1KW1KAH′,
where KW and KA are the mass transfer coefficients (m d-1) of
Hg and PAHs in the water and air films, respectively. The mass transfer
coefficient of CO2 in the water phase (KW,CO2, m d-1)
can be used to calculate KW (Gonzalez-Gaya et al., 2016), which
is a wind-speed quadratic function at a height of 10 m (U10, m-1) (Nightingale et al., 2000). A Weibull distribution of wind speed
was assumed to parameterize KW,CO2 because average wind speeds
were used during the sampling period as the gas and dissolved phase
concentrations were averaged values for the sampling transects;
KW,CO2 was calculated using a previously reported method (Livingstone and Imboden, 1993):
KW,CO2=0.24[0.24η2Γ1+2ξ+0.061ηΓ1+1/ξ],
where η and ξ are the constants of scale and shape in the Weibull
distribution, respectively, and Γ represents a gamma
function. ξ=2 (Rayleigh distribution) was used as recommended (Gonzalez-Gaya et al., 2016). η is related to wind speed and was
calculated with U10=ηΓ(1+1/ξ) (Livingstone and Imboden, 1993).
KW can be calculated as follows:
KW=KW,CO21SCPAH600,
where SCPAH is the Hg / PAH Schmidt number. The same applies for
KA, which was also calculated from wind speed and the H2O mass
transfer coefficient for the air phase (KA,H2O, cm s-1):
16KA,H2O=0.2U10+0.3,17KA=864KA,H2ODi,aDH2O,a,
where Di,a and DH2O,a represent the Hg / PAH and H2O
diffusive coefficients in air, respectively.
The uncertainty was lower than a factor of 1–2 in these estimates for
metals/PAHs. Most of the increasing uncertainty was associated with the
Henry's law constants. The effect of uncertainty on the air–water exchange
net direction was assessed by the ratios of air–water fugacity
(fG/fW) (Figs. S1 and S2 in the Supplement); moreover, the findings revealed that most metals and PAHs were not close to the
equilibrium of air–water. Among the PAHs, net volatilization was detected
only for dibenzothiophene, alkylated phenanthrenes, and fluoranthene. The
details of the uncertainty analysis are shown in Sect. S1 (Supplement).
Gross fluxes of volatilization and absorption depend on the first and second
terms of Eq. (10), respectively. The total accumulated fluxes for the Barents Sea, the Kara
Sea, the Laptev Sea, and the East Siberian Sea were acquired by multiplying the mean
basin flux with its standard deviation by the surface area of each basin.
The estimations of degradation fluxes of PAHs in the atmospheric ocean
boundary were calculated as follows:
Datm=CGf-CGiABLt,
where CGf and CGi are the last concentration after a fixed time in
a closed system (ng m-3) and the concentration in the gas phase at the
initial time (ng m-3), respectively. t is the time period (average 5 h daytime per day), and ABL represents the average height of the atmospheric
boundary layer (380 m). CGf can be calculated as follows:
LnCGiCGf=kOH[OH]t,
where kOH is the rate constant for a PAH reaction with OH radicals
(Keyte et al., 2013), and [OH] is the hydroxyl radical
concentration in the mixed layer (1000–500 hPa) based on the monthly mean
OH radical concentration (Spivakovsky et al.,
2000). The mean concentrations of OH were calculated by Eq. (13). The OH
concentrations ranged between 5.23 and 17.26×105 mol cm-3. In addition, only the PAHs in the gas phase were considered while the
potential degradation of PAHs bound in aerosols was ignored. Considering the
uncertainty of those sources, a relevant error factor of 2–3 was given for
the degradative fluxes based on the individual PAHs. Because of the large
uncertainties in kOH values, the degradation fluxes of PAHs in the
atmosphere could not be calculated.
Results and discussionHeavy metals in the atmosphere and ocean
Nine heavy metals were measured, and the average concentration for each
metal in each sea can be found in Table S3. The highest ∑9 metal
concentrations in the Barents Sea were found in the gas phase (CG, ng m-3), where the average concentration was 0.418 ng m-3 (Fig. S3).
The average values of CG showed no obvious differences among the four
seas, whereas the oceanic area adjacent to the Chukchi Peninsula,
Taymyr–Gydan peninsulas, and the Arkhangelsk region showed higher combined
concentrations of the nine metals (Fig. 2a). High ∑9 metal
concentrations in the aerosol phase (CA, ng m-3) were observed in
the Barents Sea (p<0.05), where the average ∑9 metal
concentration was 2.713 ng m-3 (Fig. 2b). These high levels may have
been associated with the trajectories of air from Russian inland areas. The
distributions of heavy metals in the Russian Arctic Ocean revealed that the
concentrations of the ∑9 metals in seawater were lower than those
in air. The concentrations of each metal in aerosols were comparable to
those previously reported in the Russian Arctic, i.e., the Laptev Sea, the Kara Sea,
the Barents Sea, Severnaya Zemlya, and Wrangel Island (Shevchenko et al.,
2003; Vinogradova and Ivanova, 2017). For other parts of the Arctic Ocean,
the average mass concentrations of each metal from Svalbard, the Fram
Strait, the central Arctic, and Greenland (Ferrero et al., 2019; Maenhaut et
al., 1979, 2002, 1989), were higher than
those found in aerosols in our study. Metals' concentrations in aerosols in
our study were lower than those in the Red Sea and the Mediterranean area (Chen et al., 2008). The average ∑9 metal
concentrations in dissolved water (CW) ranged from 0.526 to 0.896 µg L-1 (the Laptev Sea to the Barents Sea). This is relatively lower
than the concentrations of dissolved trace metals (Mn, Fe, Ni, Zn, and Cd)
previously reported for the western Arctic Ocean (the Chukchi Sea and the Canada
Basin, depth: 5–20 cm) (Kondo et al., 2016). Higher values
of CW were observed in the Barents Sea–Kara Sea region (the Yamal
Peninsula) and in the East Siberian Sea (close to the Chukchi Peninsula) in
comparison with the CW values in other areas (Fig. 2c).
Occurrence of heavy metals. Results show the concentrations of
heavy metals in the (a) gas phase, (b) aerosol phase, and (c) dissolved
water phase. Colored bars show the sum of nine quantified metals. The number
at the bottom of the legend bars in the panels represent the concentration scale (and is the same
for Figs. 3–7).
The abundance of each metal in gases, aerosols, and dissolved water is
dependent on the emission sources. In this study, Fe and Zn were the most
abundant metals detected in aerosols and dissolved water from the Russian
Arctic Ocean, where the average ∑9 metal concentrations in
aerosols and dissolved water were 0.64 ng m-3 and 0.91 ng L-1,
respectively. Pb was the most abundant metal in the gas phase (the average
concentration in the Russian Arctic Ocean =0.14 ng m-3). In
comparison with aerosols and dissolved water, the gas phase contained higher
levels of Hg, which is a finding consistent with the usual form of Hg in the
atmosphere (>98 %) and the tendency for the remaining types of
Hg to adsorb to particles during atmospheric transport (Poissant et al., 2008). In all phases, the proportions of Mn,
Fe, Pb, and Zn were significantly higher than those of other heavy metals.
Additionally, the metal distributions in the Barents Sea and the Kara Sea showed
the highest proportions, followed by the metal distributions in the East
Siberian Sea. On the Taymyr Peninsula (adjacent to the Kara Sea and the Laptev
Sea), there is a mining and metallurgical factory operated by the Norilsk company that processes copper and nickel and is one of the biggest
metallurgical factories in the world. This may be a likely source of metals
in the Kara Sea region (Shevchenko et al., 2003). Due to
the significant differences in the concentrations of metals in the marine
boundary layer both temporally and spatially throughout the Russian Arctic
Ocean (Vinogradova and Polissar, 1995; Shevchenko et al., 1999), as well
as the scarcity of reported data on heavy metals in the atmosphere in this
region, it was difficult to compare our data with historical findings.
However, our data are similar to those reported for September 1993 in the
Kara Sea (Rovinsky et al., 1995).
Measured atmosphere–ocean exchange of heavy metals. (a) Fluxes of
dry deposition for nine heavy metals; (b) fluxes of net diffusive air–water
exchange for Hg. In panel (a), the colored bars represent the sum of nine heavy
metals. In panel (b), the downward bars represent the net deposition into the
ocean, and the upward bars represent the net volatilization of Hg.
The dry deposition that involves aerosols binding to heavy metals (FDD,
ng m-2 d-1) is a major process for heavy metal deposition (Shevchenko et al., 2003). In the Russian Arctic Ocean, the
average FDD of the ∑9 metals ranged from 392 to 8067 ng m-2 d-1 (mean of 1792 ng m-2 d-1). The largest FDD
value was found close to the coast of the East Siberian Sea, where FDD
values of 305 and 224 ng m-2 d-1 were observed for Hg and Pb,
respectively, and were dominant (Fig. 3a). Our results seem to be 1 order of
magnitude higher than those in the Red Sea (mean of 615 ng m-2 d-1) (Chen et al., 2008) and Mediterranean Sea (mean of 264 ng m-2 d-1; Chen et al., 2008). However, this
comparison may not reflect the strength of the emission sources because dry
deposition is highly dependent on the deposition velocity, which is affected
by meteorological conditions such as the humidity, wind speed, and the stability
of the air column (Mariraj Mohan, 2016). The relative humidity in the
Arctic Ocean tends to be higher in coastal areas and, notably, we sampled
during spring–winter when water vapor evaporates from the relatively warmer
surfaces of seawater (Vihma et al., 2008). In addition, the wind over
sampling sites in the Arctic Ocean was ∼7 m s-1 on
average (the highest average wind speed was ∼9 m s-1 in
the Barents Sea), which is significantly higher than the wind in the Red Sea
and the Mediterranean Sea (0.36–1 m s-1) (Chen et al., 2008; Chester et
al., 1999). During the eight snow events encountered in the course of the expedition,
the wet deposition flux of the ∑9 metals (FWD, µg m-2 d-1) ranged from 23 to 32 µg m-2 d-1 (mean of
26 µg m-2 d-1) (Fig. 4a). Data relevant to the wet deposition
flux of heavy metals in the Arctic region include results for Hg, which were
estimated on land in Alaska; the highest deposition was detected along the
southern and southeastern coasts (>0.05µg m-2 d-1) (Pearson et al., 2019). The values were quite
similar to the FWD for Hg in our study (0.05 to 0.09 µg m-2 d-1). Using analysis of variance tests, we did not find any
significant difference in the FWD at different locations (p>0.05) for all heavy metals, whereas a relatively higher FWD for Hg was
observed in coastal areas adjacent to the Taymyr Peninsula with industrial
factories. Pearson et al. (2019) pointed out that there
are larger contributions from Hg wet deposition in the Bering Sea and the Gulf
of Alaska, which are influenced by the western Pacific winds from East
Asia, where high Hg emissions from industrial activities and coal burning
occur (Wong et al., 2006). The Russian Arctic Ocean is
also affected by Pacific winds from East Asia, which could lead to a combination of
heavy metal deposition from both local anthropogenic sources and long-range
transport from Asia. Wet deposition is an important process for the transfer
of heavy metals from gas and aerosol phases to ocean water. Snowfall in the
Arctic is an important fraction of precipitation, but variations in
measurements ranging from 20 % to 50 % can occur under windy conditions
even with sampling equipment designed with wind protection (Rasmussen
et al., 2012). However, snow events are quite sporadic in the Russian Arctic
Ocean during spring–summer compared with the other deposition processes.
Nevertheless, wet deposition in our study was under regional influences and
had a relatively high uncertainty.
Wet deposition of (a) heavy metals and (b) PAH fluxes. The measured wet deposition of heavy metals and PAHs occurred during the eight
snow events encountered during the vessel's expedition.
For many heavy metals that form volatile species, there is additional
evidence that their existence in water is strongly related to releases from
terrestrial environments rather than internal cycling in aquatic systems
(Robert, 2013). For example, following the deposition of
atmospheric Fe, a nonvolatile species, the concentrations in water are
influenced mainly by the particulate phase and its dissolution, whereas for
Hg, a volatile species that predominantly exists in the atmosphere as a gas
(Hg0), the concentrations of volatile Hg species in water are largely
influenced by volatilization and deposition processes at the air–water
interface; portions of the Hg in aquatic systems end up being converted to
methylmercury (Mason and Sheu, 2002; Sunderland and Mason, 2007; Selin et
al., 2007; Strode et al., 2007). Hg concentrations in the gas phase in the
present study were significantly lower than those measured in 1996 in the
Northern Hemisphere (1.5–1.7 ng m-3) and Southern Hemisphere (1.2–1.3 ng m-3) (Steffen et al., 2005; Slemr et al., 2003; Wängberg et
al., 2007; Kim et al., 2005). Steffen et al. (2002) indicated
that there has been increasing retention of Hg in the Arctic region based on
analyses of long-term measurements of atmospheric Hg concentrations.
Diffusive air–water exchange is the dominant process driving the exchange
of Hg in the ocean. The net diffusive air–water exchange (FAW, ng m-2 d-1) was estimated by a two-film resistance model
(Robert, 2013). The net input of Hg was calculated as shown in
Fig. 3b and revealed that there was a net input from the atmosphere to the
ocean at all stations, especially for the stations close to industrial/urban
areas. The integrated monthly FAW fluxes (in tons per month) for Hg were
of the same order of magnitude as the FDD fluxes for Hg and other heavy
metals in the Russian Arctic Ocean (Fig. S4). For Hg, the gross
volatilization and gross absorption in the Russian Arctic Ocean were 250 and
530 t per month, respectively. In consideration of previous studies of
atmospheric mercury depletion events (AMDEs), during which the net input of
Hg in the Arctic was evaluated (Brooks et al., 2006; Lindberg et al.,
2001), we adjusted our sampling times to avoid sampling during sunrise: during this period the autocatalytic release of sea salt aerosols changes the oxidative
photochemistry in the stratified planetary boundary layer where elemental
and reactive Hg in the gas phase is oxidized by reactive halogens. It was
estimated that ∼99–496 t of Hg are deposited annually in
the Arctic during AMDEs (Skov et al., 2006; Ariya et al., 2004). The net
input of Hg in the present study was 1 order of magnitude higher than that
caused by the AMDEs; this discrepancy may result from the fact that the
previous studies considered the terrestrial boundary between air masses and the
snowpack, and that the different locations and seasons were affected by
different meteorological conditions. In northern regions, it has been shown
that Hg undergoes long-range transport from Eurasia, especially during the
winter season (Poissant et al., 2008). These net amounts of Hg
entering into the Arctic Ocean pose potential risks to marine biota because
Hg is poorly mobile and can be retained by aquatic biota that are exposed to
it during the deposition process (Harris et
al., 2007).
PAHs in the atmosphere and ocean
A total of 35 individual PAHs, which included isomer groups such as alkylated
PAHs, were measured. The average concentrations of PAHs in each sea of the
Russian Arctic Ocean are shown in Table S4. The average values of CG
showed no obvious differences in the Kara Sea, the Laptev Sea, and the East Siberian
Sea (p>0.05), and no particularly high levels of CG were
detected at any of the sampling sites (Fig. 5a). The range of ∑35 PAH CG is 19.87–22.14 ng m-3 for the Barents Sea, 19.01–22.34 ng m-3 for the Kara Sea, 19.23–21.70 ng m-3 for the Laptev Sea, and
19.28–22.61 ng m-3 for the East Siberian Sea. The highest CG of
∑35 PAH is observed in the Barents Sea, with a value of 22.61 ng m-3. ∑35 PAH CA in the Barents Sea (0.25–2.95 ng m-3) and the East Siberian Sea (0.24–3.32), with average values of 1.38
and 2.07 ng m-3 , respectively, were higher than those in the Laptev
Sea (0.23–0.89 ng m-3) and the Kara Sea (0.23–0.27 ng m-3), with
average values of 0.30 and 0.25 ng m-3, respectively (Fig. 5b). The
average CA of ∑35 PAH in the present study is higher than
the average CA of ∑64 PAH measured in the South Atlantic
Ocean (0.93 ng m-3) and North Pacific Ocean (0.56 ng m-3), while it is
much lower than the average CA of ∑64 PAH in the Indian
Ocean (10 ng m-3) (Gonzalez-Gaya et al., 2016). The average ∑35 PAH CA of 1.02 ng m-3 in the Russian Arctic Ocean is
comparable to the average ∑64 PAH CA observed in the South
Atlantic Ocean and South Pacific Ocean (1.1 ng m-3)
(Gonzalez-Gaya et al., 2016). ∑18 PAH CA was measured
from the North Pacific towards the Arctic Ocean, ranging from 0.0002 to 0.36 ng m-3, with the highest concentration found in the coastal areas in
East Asia (Ma et al., 2013). These concentrations were
significantly lower than the average levels found in our study. In addition,
Ma et al. (2013) observed relatively higher ∑18 PAH
CA values in the most northern latitudes of the Arctic Ocean, which is
associated with back trajectories of air masses from southern Asia. The
higher levels of CA in our study could be attributed to the proximity of the coastline to larger areas of burning taiga forest and more industrial
sources in the boreal regions of Russian continent. Similar to the pattern
for heavy metals mentioned above, high levels of these chemicals may have
been derived from atmospheric transport from the industrial areas of the
Russian continent. Due to the various sampling methods and the
differences in PAHs measured, and because not all studies separated gas and
particles concentrations, it is quite difficult to compare PAH levels in
aerosols. The average ∑35 PAH CW ranged from 13.07 ng L-1 (in the Laptev Sea) to 69.90 ng L-1 (in the Barents Sea), and its spatial
variability was similar to that of PAH concentrations in aerosols. The ranges
of ∑35 PAH CW for the Barents Sea, the Kara Sea, the Laptev Sea, and
the East Siberian Sea are 12.36–162.05, 12.18–14.04,
11.21–15.82, and 11.40–129.60 ng L-1, respectively.
Higher levels of CW were also found along the coast of the Yamal–Gydan
peninsulas, where active petrol and natural gas industries sites are located (Fig. 5c).
Occurrence of PAHs in the Russian Arctic Ocean. Concentrations of
PAHs in the (a) gas phase, (b) aerosol phase, and (c) dissolved water phase.
Color bars indicate the sum of 35 PAHs, where each PAH corresponds to the
bottom legend (colors range from red for the heaviest molecular weight PAHs
to green for the lightest molecular weight PAHs).
The contribution of each PAH in the gas, aerosol, and dissolved water phases
is determined by its source, volatility, and hydrophobicity (Lima
et al., 2005). The low-molecular-weight PAHs were dominant in gas and
dissolved water (Fig. S5). In the gas phase, low-molecular-weight PAHs
occupied more than 75 % of the ∑35 PAHs, which mainly contained
methylated phenanthrenes, e.g., methylphenanthrene (mean of 1.31 ng m-3), dimethylphenanthrene (mean of 1.27 ng m-3), and
trimethylphenanthrene (mean of 1.32 ng m-3), and methylated
dibenzothiophenes, e.g., methyldibenzothiophene (mean of 1.29 ng m-3),
dimethyldibenzothiophene (mean of 1.27 ng m-3), and
trimethyldibenzothiophene (mean of 1.32 ng m-3). In dissolved water,
methylnapthalene and tetramethylnaphthalene were the most abundant PAHs with
average concentrations of 1.12 and 1.45 ng L-1, respectively. Measured
values of CG, CA, and CW are known to vary with the changes
in each PAH concentration in the marine environment (Berrojalbiz et al.,
2011; Castro-Jimenez et al., 2012; Cabrerizo et al., 2014). However, there
were no previous reports regarding the occurrence of PAHs in the Russian Arctic
atmosphere and ocean.
Dry deposition fluxes for the 35 measured PAHs. Color bars
indicate the sum of the 35 quantified compounds, and each color represents
the individual PAHs in the bottom legend (colors range from red for the
heaviest molecular weight PAHs to green for the lightest molecular weight
PAHs).
The average dry deposition flux (FDD) of the ∑35 PAHs was
1108 ng m-2 d-1. The increasing values of Vd may influence
FDD in the marine environment due to the higher hydrophobicity of
organic compounds, surface microlayer with reduced surface tension, and
lipid floating (Del Vento and Dachs, 2007b). The higher average
Vd was observed for 9-methylfluorene (1.01–10.02 cm s-1),
followed by 1,7-dimethylfluorene (1.06–10.63 cm s-1) (Fig. S6). On a
global scale, higher Vd values were found for heavier PAHs such as
methylchrysene (0.17–13.30 cm s-1) and dibenzo[a,h]anthracene
(0.29–1.38 cm s-1) (Gonzalez-Gaya et al., 2014). The Vd
values reported previously ranged from 0.08 to 0.3 cm s-1 in the
Atlantic Ocean (Del Vento and Dachs, 2007a) and from 0.01 to 0.8 cm s-1 in coastal areas (Holsen and Noll, 1992; Bozlaker et al., 2008;
Esen et al., 2008; Eng et al., 2014); the higher values were observed in
concentrated industrial and urban areas (Bozlaker et al.,
2008). In our study, the highest Vd values were observed in the Barents
Sea; the other three seas had similar Vd values (p>0.05)
that were lower than in the Barents Sea except for 9-methylfluorene and
1,7-dimethylfluorene. The East Siberian Sea exhibited the lowest value of
Vd, whereas the relatively higher Vd values were found for heavier
PAHs (dibenzo[a,h]anthracene, indeno[1,2,3-cd]pyrene, dibenzo[a,h]anthracene
and benzo[g,h,i]perylene) in all seas (Fig. S7). This may be due to heavier
PAHs being principally deposited via heavier aerosols with a higher Vd
because they are bound to hydrophobic aerosols or because of gravity, e.g.,
soot carbon (Gonzalez-Gaya et al., 2014). Dry deposition is a major
process for high-molecular-weight PAHs bound to aerosols (Figs. 6, S8). The deposition values varied mainly in accordance with the PAH
concentrations of aerosols in suspension and the factors influencing the
deposition velocities (wind speed, compound vapor pressure, etc.). The
FWD of the ∑35 PAHs ranged from 14 to 19 µg m-2 d-1. Gonzalez-Gaya et al. (2014) found the highest FWD of
∑64 PAHs in the North Atlantic Ocean (24 µg m-2 d-1) with an average FWD value of ∼8µg m-2 d-1 on the global scale based on rain samples. The higher
FWD values of PAHs were found in urban areas in China (62.6 µg m-2 d-1) (Wang et al., 2016) from rain samples and
the lower FWD values of PAHs (0.02–0.28 µg m-2 d-1)
from both rain and snow samples were observed in high mountain areas in Europe (Arellano et al., 2018). Our FWD
values were within the range of the previously reported global scale and the
difference in wet deposition was mainly dependent on source distance and
precipitation intensity. Wet deposition is an important purging process for
semivolatile organic compounds such as PAHs in the gas and aerosol phase.
Snow events are quite sporadic in the Arctic Ocean and, thus, these have
lower relevance for wet deposition of PAHs in this region.
Measured atmosphere–ocean exchange of PAHs. (a) Net diffusive
air–water exchange fluxes (all net deposition into the ocean) and (b) volatilization air–water exchange fluxes. Color bars indicate the sum of
the 35 quantified compounds, and each color represents the individual PAHs
in the bottom legend (colors range from red for the heaviest molecular
weight PAHs to green for the lightest molecular weight PAHs).
The estimated FAW values revealed that most PAHs had a net input from
the atmosphere to the ocean except for the more volatile PAHs, such as two to three ring PAHs (Fig. 7). The lighter PAHs (two to three rings) appeared more volatile
(978–4892 ng m-2 d-1), whereas heavier PAHs (four to six rings) showed net
deposition (1561–7808 ng m-2 d-1), except dibenzo[a,h]anthracene
(1322 ng m-2 d-1), indeno[1,2,3-cd]pyrene (1238 ng m-2 d-1), trimethylphenanthrene (1901 ng m-2 d-1), and
benzo[g,h,i]perylene (2708 ng m-2 d-1). A net deposition that was 3 orders of magnitude
higher was observed for methylphenanthrene,
dimethylphenanthrene, trimethylphenanthrene, and tetramethylphenanthrene
(Fig. S9). Our results were similar to those observed in other PAH-affected
areas such as the southeast Mediterranean (Castro-Jimenez et
al., 2012), Narragansett Bay (Lohmann et al., 2011), and the North
Atlantic Ocean (Lohmann et al., 2009). Ma et al. (2013)
suggested that slight volatilization of lighter PAHs may exist from
additional sources such as ship ballast and riverine runoff, which is
consistent with our findings that volatilization occurred mainly in the East
Siberian Sea and Barents Sea, where more industrial factories and urban
areas are situated. Our study is also consistent with previous reports
showing that diffusion during air–water exchange is the main process for the
transfer of relatively light volatile organic compounds to the marine
environment (Castro-Jimenez et al., 2012; Jurado et al., 2005). The
integrated monthly FAW values (in tons per month) of five to six ring PAHs were
comparable to FDD values in the Russian Arctic Ocean, whereas only the
East Siberian Sea showed high levels of dry deposition (Fig. S9). In all
four seas, FAW values of three to four ring PAHs were of the same magnitude as
FDD values. The total volatilization and total adsorption of the ∑35 PAHs in the Russian Arctic Ocean amounted to 2600
and 3640 t per month, respectively. Therefore, there was a net input of
the ∑35 PAHs from the atmosphere to the marine environment that
reached 3276 t, which was 100 times higher than for the aerosol-bound
∑35 PAHs that underwent dry deposition (estimated at
∼30 t per month). In other reports, Gonzalez-Gaya et
al. (2016) estimated the global input of PAHs from the atmosphere to the
ocean to be of the order of 90 000 t per month, and Reddy
et al. (2012) reported that the input of PAHs to the ocean in the Gulf of
Mexico in 2010 after the Deepwater Horizon oil spill was 20 000 t. Such
comparisons suggest that the diffusive fluxes in the Russian Arctic Ocean
play an important role in the atmosphere–ocean exchange of PAHs, whereas
there is a lower input of PAHs to the Russian Arctic Ocean on the global
scale.
Atmospheric degradation of PAHs. Estimated fluxes of degraded PAHs
in the gas phase following reaction with OH radicals.
In addition to the transfers of PAHs to the ocean, PAHs can also be degraded
during transport through the atmosphere due to reactions with OH
radicals (Keyte et al., 2013). The degradation flux Datm
of PAHs in the gas phase of the oceanic atmosphere (Fig. 8) was estimated at
3000 t per month for the ∑35 PAHs, which represents an
additional PAH sink (see Sect. 2). In general, the large amounts of PAHs
undergoing net deposition to the ocean and degradation during atmospheric
transport are indicative of large source areas in the Russian Arctic.
Notably, PAH concentrations have been increasing in the atmosphere owing to
wildfires and fossil fuel use over the past century (Zhang and
Tao, 2009). The high-molecular-weight PAHs were dominant in the aerosols
(Fig. 5) originating from pyrolytic sources (Lima et al., 2005).
In addition, high abundances of alkylated PAHs were observed in the gas and
dissolved phases; along with the evaluations of the diagnostic ratios
(Fig. S10), these results were suggestive of pyrogenic sources for PAHs in
gases and aerosols, and mixtures of pyrogenic and petrogenic sources for
PAHs in dissolved water (mostly for the Laptev Sea and the East Siberian Sea). Other sources contributing to the occurrence of PAHs may have involved both
anthropogenic and biogenic sources on land (Cabrerizo et al., 2011).
In the Russian Arctic Ocean, it can be assumed that PAHs in the atmosphere
(gas and aerosol) originated from anthropogenic sources including industrial
and urban activities, while PAHs in seawater, at sites with relatively fewer
anthropogenic sources, i.e., the Laptev Sea and the East Siberian Sea,
originated from a mixture of anthropogenic and biogenic sources. This
indicates that the atmospheric transport of PAHs derived from anthropogenic
activities occurs in all sectors of the Russian Arctic Ocean, while only the
East Siberian Sea and the Laptev Sea have more anthropogenic PAHs in the water
phase.
Because PAHs are toxic, these chemicals can have an adverse influence on
food webs in marine ecosystems (Hylland, 2006). In particular, even
though PAHs are present at natural background levels in the marine
environment, the massive usage of fossil fuels has led to increases in PAH
emissions and excessive PAH concentrations in many marine environments. The
present study indicates that there are high contributions of diffusive
atmospheric PAHs to the Arctic Ocean, and these chemicals are potentially
perturbing the carbon cycle in the ocean and posing risks to the fragile
Arctic marine food webs. Thus, further studies of the impacts of such
chemicals are warranted.
Conclusion
This study presents the occurrence and atmosphere–ocean fluxes of 35 PAHs
and 9 heavy metals in the Arctic Ocean. Dry deposition and wet deposition
fluxes of nine heavy metals in aerosols were estimated at 2205 and 10.95 µg m-2 d-1, respectively. The net gross
absorption of Hg in the Arctic Ocean was estimated at 280 t per month. A
net input of PAHs from the atmosphere to the Arctic Ocean was observed for
most of the PAHs, especially for the low-molecular-weight PAHs. The net
atmospheric input of the 35 PAHs was estimated at 3276 t per month. The
current occurrences of semivolatile aromatic hydrocarbons could have been
derived from biogenic sources and anthropogenic sources from continental
land masses, especially for the locations close to industrial areas. These
inputs of Hg and PAHs may be causing adverse effects on the fragile Arctic
marine ecosystems; this issue warrants further research. In addition, our
study suggests that both PAHs and metals are affected by local depositional
effects and changes of source emission; therefore, the spatial distribution
of these compounds and source identification need to be further
investigated.
Data availability
All original data regarding the concentrations of PAHs and heavy metals in the gas, aerosol, and dissolved phases, as well as dry deposition velocity are shown in the Supplement. All other data utilized for calculations can be accessed by contacting the corresponding author.
The supplement related to this article is available online at: https://doi.org/10.5194/acp-19-13789-2019-supplement.
Author contributions
XJ and EA set up the
sampling equipment and analyzed the samples and the data. XX
also helped to collect and analyze the data. XJ and XX wrote the paper.
Competing interests
The authors declare that they have no conflict of interest.
Acknowledgements
We would like to thank Yu Su from the School of Visual Arts at BFA Computer Art for helping with data visualization, and Kuznetsova Ekaterina for helping with the Russian translation.
Financial support
This research has been supported by the National Key R&D Program of China (grant no. 2016YFE011230), the Russian Foundation for Basic Research (grant nos. 18-44-890003 and 16-34-60010), the Jiangsu Nature Science Fund (grant no. BK20151378), and the Fundamental Research Funds for the Central Universities (grant no. 090514380001).
Review statement
This paper was edited by Ralf Ebinghaus and reviewed by two anonymous referees.
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