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
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.
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
Locations of investigated islands for soil sampling and trajectory of the vessel in the Russian Arctic.
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 m
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
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
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
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-
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
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.
Dry deposition fluxes (
The wet deposition fluxes (
The air–water diffusive fluxes (
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
(
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:
Nine heavy metals were measured, and the average concentration for each
metal in each sea can be found in Table S3. The highest
Occurrence of heavy metals. Results show the concentrations of
heavy metals in the
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
Measured atmosphere–ocean exchange of heavy metals.
The dry deposition that involves aerosols binding to heavy metals (
Wet deposition of
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
(
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
Occurrence of PAHs in the Russian Arctic Ocean. Concentrations of
PAHs in the
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
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 (
Measured atmosphere–ocean exchange of PAHs.
The estimated
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
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.
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
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:
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.
The authors declare that they have no conflict of interest.
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.
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).
This paper was edited by Ralf Ebinghaus and reviewed by two anonymous referees.