Contribution of ship emissions to the concentration and deposition of air pollutants in Europe

Abstract

Emissions from the marine transport sector are one of the least-regulated anthropogenic emission sources and contribute significantly to air pollution. Although strict limits were introduced recently for the maximum sulfur content in marine fuels in the SECAs (sulfur emission control areas) and in EU ports, sulfur emissions outside the SECAs and emissions of other components in all European maritime areas have continued to increase in the last two decades. We have used the air quality model CAMx (Comprehensive Air Quality Model with Extensions) with and without ship emissions for the year 2006 to determine the effects of international shipping on the annual as well as seasonal concentrations of ozone, primary and secondary components of PM${}_{2.5}$, and the dry and wet deposition of nitrogen and sulfur compounds in Europe. The largest changes in pollutant concentrations due to ship emissions were predicted for summer. Concentrations of particulate sulfate increased due to ship emissions in the Mediterranean (up to 60 %), the English Channel and the North Sea (30–35 %), while increases in particulate nitrate levels were found especially in the north, around the Benelux area (20 %), where there were high NH${}_{\mathrm{3}}$ land-based emissions. Our model results showed that not only are the atmospheric concentrations of pollutants affected by ship emissions, but also depositions of nitrogen and sulfur compounds increase significantly along the shipping routes. NO${}_x$ emissions from the ships, especially in the English Channel and the North Sea, cause a decrease in the dry deposition of reduced nitrogen at source regions by moving it from the gas phase to the particle phase which then contributes to an increase in the wet deposition at coastal areas with higher precipitation. In the western Mediterranean region, on the other hand, model results show an increase in the deposition of oxidized nitrogen (mostly HNO${}_{\mathrm{3}})$ due to the ship traffic. Dry deposition of SO${}_{\mathrm{2}}$ seems to be significant along the shipping routes, whereas sulfate wet deposition occurs mainly along the Scandinavian and Adriatic coasts. The results presented in this paper suggest that evolution of NO${}_x$ emissions from ships and land-based NH${}_{\mathrm{3}}$ emissions will play a significant role in future European air quality.

Introduction

There have been many studies on the effects of air and road traffic emissions and projections of their future levels (Cuvelier et al., 2007; Schurmann et al., 2007; Westerdahl et al., 2008; Koffi et al., 2010; Uherek et al., 2010; Wilkerson et al., 2010; Hodnebrog et al., 2011). Relatively few studies, on the other hand, have dealt with the impacts of ship emissions in detail (Corbett et al., 2007; Eyring et al., 2010; Huszar et al., 2010; Jonson et al., 2015).

The marine transport sector, which is one of the least-regulated anthropogenic emission sources, contributes significantly to air pollution, particularly in coastal areas (Marmer and Langmann, 2005; Gonzalez et al., 2011). Emissions from maritime transport in European waters constitute a significant share of worldwide ship emissions of air pollutants and greenhouse gases (EEA, 2013). Shipping is one of the fastest-growing sources of greenhouse gas emissions due to transport and is also a major source of air pollution, which causes health problems, acid rain and eutrophication (Brandt et al., 2013).

Legislation on air pollutants and greenhouse gases from the maritime sector is a major challenge because of the characteristics of the shipping sector, which include global trade operations based in different countries. The efforts of the European Union (EU) and the International Maritime Organization (IMO) to tackle emissions from international shipping are different and, to date, there is no integrated legislation. Globally, the International Maritime Organization (IMO) regulates emissions through the International Convention for the Prevention of Pollution from Ships (MARPOL) and its Annex VI (http://www.imo.org/OurWork/Environment/PollutionPrevention/Pages/Default.aspx). The latest fuel sulfur limits in emission control areas (ECAs) were set at 0.1 % as of 1 January 2015. Reductions of NO${}_x$ emissions from marine diesel engines are also regulated but these focus only on new ships, where limits are defined as a function of speed and installation year.

In Europe, the maximum sulfur content of the marine fuel used by ships operating in the sulfur emission control areas (SECAs) – the Baltic Sea, the English Channel and the North Sea – was restricted to 1.0 % in July 2010 and further reduced to 0.1 % in January 2015. The EU sulfur directive has limited the sulfur content to 0.1 % in harbor areas since January 2010. Although more stringent NO${}_x$ emission limits legislated by the IMO have forced marine diesel engine manufacturers to consider a variety of different emission reduction technologies, there are no NECAs (NO${}_x$ emission control areas) in Europe yet. Since the IMO NO${}_x$ emissions regulations refer only to new ships, the impact of these regulations is minimal at present and probably will likely remain so in the near future (EEA, 2013).

The highest level of detail on ship movements can be obtained with the AIS (Automatic Identification System) data. The AIS was developed and made compulsory by the IMO for all ships over 300 gross tonnage to minimize the probability of groundings and collisions of ships. These signals allow very accurate positioning of vessels and their emissions. When combined with knowledge on each ship's engine and possible abatement techniques, a realistic estimation of fuel consumption and emissions can be made. Jalkanen et al. (2009) presented an automated system that is based on AIS signals, to evaluate exhaust emissions from marine traffic in the Baltic Sea area. A pilot project using the AIS data to estimate shipping emissions in the port of Rotterdam allowed for calculation of emissions on a much finer geographical grid than could be done previously (Denier van der Gon and Hulskotte, 2010). In the near future, AIS data are expected to be used to improve accuracy of emission estimates in a larger area in Europe.

Johansson et al. (2013) reported that the emission limitations from 2009 to 2011 have had a significant effect on reducing the emissions of SO${}_x$ in the northern ECA in Europe. On the other hand, sulfur emissions in sea areas outside the SECAs and emissions of other species – especially NO${}_x$ – in all sea areas around Europe have been increasing over the past few decades, while land-based emissions have been gradually decreasing. The revised Gothenburg Protocol specifies national emission reduction commitments in Europe to be achieved by 2020 (http://www.unece.org/env/lrtap/multi_h1.html). These commitments, however, are only for land-based sources and do not cover emissions from international shipping. According to the European Environment Agency, emissions of nitrogen oxides from international maritime transport in European waters are projected to increase and could be equal to land-based sources by 2020 (EEA, 2013). It is therefore important to understand the impacts of shipping emissions on both concentrations and deposition of specific air pollutants. Most of the previous studies have dealt with the impacts of ship emissions on a global and continental scale, while there have only been a few studies available that quantify the impact of ship emissions on smaller scales, using models with high resolution. In this modeling study, we investigated the impacts of ship emissions on European air quality in detail by analyzing the seasonal and spatial variations of the contributions from the shipping sector to the concentrations of ozone and PM${}_{2.5}$ components, as well as to the deposition of nitrogen and sulfur compounds.

Method

The models used in this study are the Comprehensive Air Quality Model with Extensions (CAMx), version 5.40 (http://www.camx.com) and the Weather Research and Forecasting Model (WRF-ARW), version 3.2.1 (http://wrf-model.org/index.php). The model domain covered all of Europe with a horizontal resolution of 0.250${}^{\circ }$ $\times$ 0.125${}^{\circ }$, which corresponds approximately to 19 km $\times$ 13 km around the central latitudes of the model domain. We used 6 h ECMWF (European Centre for Medium-Range Weather Forecasts, http://www.ecmwf.int/) data to provide initial and boundary conditions for the WRF model. WRF uses 31 vertical layers up to 100 hPa, of which 14 were used in CAMx, with the lowest layer being 20 m thick. The initial and boundary concentrations were obtained from the MOZART (Model of Ozone and Related Chemical Tracers) global model for the studied period (Horowitz et al., 2003). MOZART uses geographic latitude–longitude coordinates and has a resolution of 1.895${}^{\circ }$ $\times$ 1.875${}^{\circ }$. Data were extracted for the area covered by our model domain and adapted to our horizontal grid cells and vertical layers using our preprocessors (Oderbolz et al., 2012). Photolysis rates were calculated using the TUV photolysis preprocessor (http://cprm.acd.ucar.edu/Models/TUV/). Ozone column densities were extracted from TOMS data (http://ozoneaq.gsfc.nasa.gov/OMIOzone.md). Dry deposition of gases in CAMx is calculated using a state-of-the-science leaf area index (LAI)-based resistance model (Zhang et al., 2003). For surface deposition of particles, CAMx includes diffusion, impaction and/or gravitational settling. CAMx uses separate scavenging models for gases and aerosols to calculate wet deposition. The gas-phase mechanism used in this study was CB05 (Carbon Bond mechanism 5; Yarwood et al., 2005). Concentrations of particles with a diameter smaller than 2.5 $\mathrm{\mu }$m were calculated using the fine/coarse option of CAMx. Calculation of secondary organic aerosols (SOAs) was based on the semi-volatile equilibrium scheme SOAP (Strader et al., 1999) that partitions condensable organic gases to seven types of SOAs. This is the traditional two-product approach which treats the primary organic aerosols as non-volatile.

The gridded TNO-MACC data for 2006 were used as the basic anthropogenic emission inventory (Denier van der Gon et al., 2010). The annual emission data for 10 SNAP (Selected Nomenclature for sources of Air Pollution) categories per grid cell in geographic the latitude–longitude coordinate system (with a grid resolution of 0.125${}^{\circ }$ $\times$ 0.0625${}^{\circ }$, which corresponds approximately to 9 km $\times$ 7 km around the central latitudes of the model domain) were converted to hourly gridded data using the monthly, weekly and diurnal profiles provided by TNO. Wildfire, sea salt and mineral dust emissions were not included in the inventory. There are some estimates of fires using the fire radiative power (FRP) from satellites (Sofiev et al., 2013). However, occurrence and intensity of such emissions, as well as vertical distributions, vary significantly spatially and temporally, making their parameterization difficult. Sea salt is mainly found on coarse particles and sea salt modeling would mainly improve formation of coarse nitrate (Sellegri et al., 2001). Similarly, mineral dust is more relevant for coarse particles (Athanasopoulou et al., 2010). Since our focus in this work was only on the fine fraction of particles (PM${}_{2.5}$), we believe that lack of such emissions did not affect our results significantly.

The biogenic emissions (isoprene, monoterpenes, sesquiterpenes) were calculated as described in Andreani-Aksoyoglu and Keller (1995) using the temperature and shortwave irradiance from the WRF output, the global USGS land-use data, and the GlobCover 2006 inventory. All emissions were treated as area emissions in the first model layer. We performed CAMx simulations for 2006 with (base case) and without (no ship) ship emissions. Figure S1 shows the annual emissions from ships. Temporal profiles for ship emissions show a small increase ($\sim$ 10 %) in summer with respect to winter (Denier van der Gon et al., 2011). Concentrations, as well as dry and wet deposition of pollutants, were calculated over the entire year.

Model performance and uncertainties

The model performance for simulations reported in this paper has been thoroughly evaluated and the results can be found in Aksoyoglu et al. (2014). It is, however, necessary to give some information about the model performance here. Accuracy of the state-of-the-art air quality models such as CAMx depends largely on the quality of the input data, such as meteorological fields and emissions. It is well known that reproducing meteorological parameters like wind fields under difficult conditions – especially in wintertime – is challenging. Uncertainty in emissions varies depending on pollutant and source. In general, some emission sources are difficult to estimate regionally, such as agricultural activities. For example, ammonia emissions and their daily and diurnal variations are related to actual climate conditions in a particular year. According to Kuonen et al. (2014), uncertainty estimates for emissions vary between 10 and 300 %, depending on pollutant and source.

Biogenic emission models require a detailed vegetation inventory, emission factors (based on very few data) for each specific species, and temperature and radiation data (Guenther et al., 2012; Oderbolz et al., 2013). In spite of extensive efforts in this field, biogenic emission models still have high uncertainty, mostly due to lack of sufficient measurements of these species. Evaluation of deposition is another challenge, since measurement techniques are available only for wet deposition. Dry deposition can only be estimated using gas-phase concentrations and dry deposition velocities.

By keeping these uncertainties in mind, the general performance of both WRF and CAMx models was reasonably good for the modeled period. The model evaluation of the CAMx model suggested a relatively good model performance with a mean bias of 4 ppb and $-$1.9 $\mathrm{\mu }$g m${}^{-\mathrm{3}}$ for annual ozone and PM${}_{2.5}$ concentrations, respectively. There was some underestimation of PM${}_{2.5}$ in January–February, when unusually high concentrations were reported in Europe due to severe meteorological conditions. The agreement between measurements and meteorological model results was good, with high correlation coefficients (0.76–0.98) and low mean bias error, MBE ($-$1.13 for air temperature, 0.57 for wind speed). These values fulfill the desired accuracy suggested by Cox et al. (1998). Details of the model performance of the base run including ship emissions can be found in Aksoyoglu et al. (2014).

Results and discussion

Annual impacts

The annual mean surface ozone was predicted to be about 40 ppb over the sea and coastal areas when emissions from the marine transport sector were excluded (Fig. S2 in the Supplement). Ship emissions cause an increase in the mean surface ozone by 4–5 ppb (5–10 %) in the Mediterranean Sea (Fig. 1). On the other hand, ozone levels decrease by about 5–6 ppb (10–20 %) around the English Channel and the North Sea due to enhanced titration caused by NO${}_x$ emissions from ships. It was shown in an earlier sensitivity study for the same year that ozone formation in that area was VOC-limited because of high NO${}_x$ $/$ VOC ratios, whereas a NO${}_x$-sensitive regime was predicted for the Mediterranean region (Aksoyoglu et al., 2012).

Contribution of ship emissions to mean surface O${}_{\mathrm{3}}$ in 2006: left in ppb (base case $-$ no ship), right in % ((base case $-$ no ship)$\times$100/(base case)).

[Figure omitted. See PDF]

The modeled mean annual concentration of PM${}_{2.5}$ varied between 5 and 40 $\mathrm{\mu }$g m${}^{-\mathrm{3}}$ for the year 2006 without ship emissions in Europe (Fig. S3). The highest concentrations were predicted around the Benelux area, northern Italy and eastern Europe. The concentration of PM${}_{2.5}$ increased along the shipping routes as well as the coastal areas when emissions from the ship traffic were included (Fig. 2). These changes were caused not only by primary PM (elemental carbon (EC) and primary organic aerosol (POA)) emissions from ships, but also by an increase in the concentration of precursor species leading to the formation of secondary PM (particulate nitrate (NO${}_{\mathrm{3}})$, sulfate (SO${}_{\mathrm{4}})$, ammonium (NH${}_{\mathrm{4}}$) and secondary organic aerosol (SOA)). The largest contribution was predicted in the western Mediterranean (up to 45 %) as well as along the north European coast (10–15 %). Studies with other models using the 2005 inventory, at a relatively coarse resolution of about 50 km, showed a similar spatial distribution but predicted a lower contribution (15–25 %) in the Mediterranean (EEA, 2013). The difference is likely due to the use of different emission inventories, in addition to the different domain resolutions. The finer resolution used in this study was able to capture the local effects more clearly.

Contribution of ship emissions to the mean PM${}_{2.5}$ concentration in 2006: left in $\mathrm{\mu }$g m${}^{-\mathrm{3}}$ (base case $-$ no ship), right in % ((base case $-$ no ship) $\times$ 100/(base case)).

[Figure omitted. See PDF]

Seasonal impacts

Ozone

We analyzed the changes in the surface ozone mixing ratios caused by the ship emissions in each season separately (Fig. 3). The effects were stronger in summer and there was a difference in the seasonal variation between north and south. Ship emissions were predicted to cause a decrease in ozone in the north, including the area of the English Channel, the North Sea and the Baltic Sea, in all seasons except summer. Ozone decreased in summer due to ship traffic only around the English Channel by $-$20 %, while it increased by about 5 % in the eastern part of the North Sea and the Baltic Sea (Fig. 3b). These results are in the same range as those found by Huszar et al. (2010) for 2004. The area around the English Channel is a high-NO${}_x$ region, leading to a reduction of the surface ozone concentration as a result of the contribution from ship emissions, as discussed in Sect. 3.1.

Contribution of ship emissions to mean surface O${}_{\mathrm{3}}$ (%) in (a) spring, (b) summer, (c) fall and (d) winter ((base case $-$ no ship) $\times$ 100/(base case)).

[Figure omitted. See PDF]

On the other hand, an opposite effect was predicted for the southern part of the model domain. Emissions from shipping led to increased surface ozone in all seasons except winter. No increase, but instead a small decrease, in winter ozone was predicted along the shipping routes (Fig. 3d). In summer, the contribution of the ship emissions to the mean surface ozone varied between $+$10 and $+$20 % in the Mediterranean, with a negative change of about $-$5 % over a very small area at the Strait of Gibraltar (Fig. 3b). Marmer et al. (2009) reported the maximum contribution of ships to surface ozone in summer 2006 as 12 % over the Strait of Gibraltar, using a global model with a horizontal resolution of 1${}^{\circ }$ $\times$ 1${}^{\circ }$. The finer horizontal resolution used in our study (0.250${}^{\circ }$ $\times$ 0.125${}^{\circ }$) enabled us to distinguish the change in contribution of ship emissions to ozone from $+$20 % over the northwest African coast to $-$5 % at the Strait of Gibraltar.

PM${}_{2.5}$

The model results suggested that emissions from the international shipping increase PM${}_{2.5}$ concentrations in all seasons (Fig. 4). The largest contribution of ship traffic was predicted in summer, when concentrations increased not only around the shipping routes, but also over the coastal areas. The change in PM${}_{2.5}$ concentrations caused by shipping emissions in summer was about 20–25 % in the north around the English Channel and the North Sea, whereas a much larger contribution was predicted in the western Mediterranean (40–50 %). In winter, the contribution decreased to 5–10 % in the north and 15–20 % in the south.

Contribution of ship emissions to PM${}_{2.5}$ (%) in (a) spring, (b) summer, (c) fall, and (d) winter ((base case $-$ no ship)$\times$100/(base case)).

[Figure omitted. See PDF]

Impacts on aerosol components in summer

In this section, the contribution of ship emissions in summer to the individual components of PM${}_{2.5}$ is investigated, because the effects are stronger in that season (see Fig. 4b). In order to understand which components are affected more by ship emissions, we first analyzed the effects on primary and secondary species. The contribution of ship emissions to the concentrations of primary and secondary PM${}_{2.5}$ is shown in Fig. 5. Elevated concentrations of the primary carbonaceous components EC and POA were predicted only along the shipping routes in the Mediterranean and in the north around the English Channel and the North Sea (Fig. 5a), whereas the concentrations of secondary aerosols (SAs) containing secondary inorganic aerosols (SIAs) and SOAs increased over a larger area (Fig. 5b). These results suggest that the effects on the concentrations of secondary particles (via formation by oxidation of gaseous precursors) are more significant than the effects on primary particles (by direct emissions). As seen in Fig. 5b, the concentrations of secondary aerosols increased not only over the sea areas but also over the continent, due to emissions from international shipping.

Contribution of ship emissions ($\mathrm{\mu }$g m${}^{-\mathrm{3}})$ to (a) the primary aerosol (PA) and (b) the secondary aerosol (SA) concentration in summer 2006 (base case $-$ no ship).

[Figure omitted. See PDF]

Detailed analysis of model results revealed that the change in the secondary aerosol concentration due to ship emissions occurs mainly in the inorganic fraction (Fig. 6a–c). The concentrations of particulate nitrate and ammonium increased by about 10–20 % around the Benelux area and northern Italy, where there are high land-based ammonia emissions (Fig. 6a and b). These results indicate that NO${}_x$ emissions from the ships and ammonia emissions from the land lead to the formation of ammonium nitrate. On the other hand, particulate sulfate increased along the shipping routes and coastal areas, with the largest effects (50–60 %) in the western Mediterranean and the North African coast (Fig. 6c). The contribution to the SOA concentration was relatively small (< 10 %) and was mainly found in the north (Fig. 6d). We note that the results for SOA might look different if a VBS (volatility basis set) scheme were used to calculate the organic aerosol (OA) concentrations (Donahue et al., 2006), but this could not be done because the volatility distribution of ship emissions is not well known yet (Pirjola et al., 2014).

Contribution of ship emissions ($\mathrm{\mu }$g m${}^{-\mathrm{3}})$ to the secondary aerosol concentration: (a) NO${}_{\mathrm{3}}$, (b) NH${}_{\mathrm{4}}$, (c) SO${}_{\mathrm{4}}$ and (d) SOA in summer 2006 (base case $-$ no ship). Note that the scale in (d) is 10 times smaller than the others.

[Figure omitted. See PDF]

Contribution to deposition

Nitrogen deposition

The atmospheric deposition of pollutants raises serious concerns for ecosystems. In general, the main nitrogen sources are emissions of nitrogen oxides from combustion processes and ammonia from agricultural activities. The deposition of atmospheric nitrogen species constitutes a major nutrient input to the biosphere, which enhances forest growth. Despite this, increased nitrogen input into terrestrial ecosystems represents a potential threat to forests. Enhanced nitrogen deposition can cause soil acidification, eutrophication and nutrient imbalances, causing a reduction in biodiversity. The deposition of atmospheric nitrogen compounds occurs via dry and wet processes. NO${}_{\mathrm{2}}$, NH${}_{\mathrm{3}}$, nitric acid (HNO${}_{\mathrm{3}}$), and nitrous acid (HONO) are the most important contributors to nitrogen dry deposition. Nitrogen wet deposition results from the scavenging of atmospheric N constituents.

The predicted annual deposition of total nitrogen in Europe, based only on the land emissions, varied between 5 and 45 kg N ha${}^{-\mathrm{1}}$ yr${}^{-\mathrm{1}}$ in 2006 (Fig. 7, left panel) and was mainly dominated by dry deposition (Fig. S4). The largest dry deposition was generally over the regions with high ambient NH${}_{\mathrm{3}}$ concentrations (the Benelux area and northern Italy) as also reported previously (Flechard et al., 2011). In the rest of the area, dry deposition of oxidized nitrogen was dominant.

Annual nitrogen deposition only due to land-based emissions (left) and contribution of ship emissions to N deposition (right) (base case $-$ no ship).

[Figure omitted. See PDF]

As seen in the right panel of Fig. 7, ship emissions caused an increase in N deposition along the shipping routes, except for a few high-NH${}_{\mathrm{3}}$ locations where a small decrease in deposition was predicted. Analysis of the changes in the dry and wet deposition showed that the main contribution of ship emissions was to dry N deposition, while wet deposition increased slightly (10 %) in the North Sea (Figs. 8 and S5).

Further investigation of the changes in the dry deposition showed that ship emissions caused an increase in the dry deposition of HNO${}_{\mathrm{3}}$ in the Mediterranean, whereas there was a small decrease ($-$2 %) in the NH${}_{\mathrm{3}}$ deposition in ammonia-rich areas (Fig. 9). Dry deposition of ammonia occurred close to the source areas. Our results suggest that NO${}_x$ emissions from ships were responsible for transformation of some gaseous ammonia to particulate ammonium (see Fig. S6), which has a lower dry deposition velocity than gaseous NH${}_{\mathrm{3}}$ but contributes to an increased wet deposition especially over the North Sea (Fig. 8, right panel). The largest contribution of the ship traffic emissions to deposition of oxidized nitrogen (in the form of HNO${}_{\mathrm{3}})$ was in the Mediterranean Sea (see Fig. 9, right panel).

Contribution of ship emissions to the annual dry N deposition (left) and wet N deposition (right) (base case $-$ no ship).

[Figure omitted. See PDF]

Contribution of ship emissions to the annual dry NH${}_{\mathrm{3}}$ deposition (left) and dry HNO${}_{\mathrm{3}}$ deposition (right) (base case $-$ no ship).

[Figure omitted. See PDF]

Sulfur deposition

After emission, sulfur dioxide is further oxidized in the atmosphere, with sulfuric acid and sulfate as the final products. Sulfate is mostly removed by wet deposition, with various effects on ecosystems, including acidification of marine ecosystems and soil, vegetation damage, and corrosion. Excluding the ship emissions, the largest total sulfur deposition was predicted to occur in the eastern part of Europe (with high land-based SO${}_{\mathrm{2}}$ emissions) (Fig. 10, left panel) and was dominated by dry deposition (Fig. S7, left panel). Wet deposition was predicted to be relatively higher in areas with high precipitation (Fig. S7, right panel). Generally, the importance of dry deposition of sulfur decreased and the importance of wet deposition increased with distance from the source, along with the decrease in the SO${}_{\mathrm{2}}$ $/$ sulfate ratio.

Sulfur deposition only due to land-based emissions (left) (no ship) and due to ship emissions (right) (base case $-$ no ship).

[Figure omitted. See PDF]

Our simulations showed that ship emissions contributed substantially to the sulfur deposition along the shipping routes and the coastal areas (Fig. 10, right panel; see Fig. S8 for relative contribution). The western Mediterranean and the North African coast were especially affected by the sulfur deposition from ship traffic. As shown in Fig. 11, the contribution to the dry SO${}_{\mathrm{2}}$ deposition dominated along the shipping routes, while the effect on wet SO${}_{\mathrm{4}}$ deposition was smaller and was mostly in areas with higher precipitation. Comparison of the right panel of Fig. 10 with the left panel of Fig. 11 clearly shows that the contribution of ship emissions to sulfur deposition is mainly in the form of SO${}_{\mathrm{2}}$ dry deposition.

Contribution of ship emissions to dry SO${}_{\mathrm{2}}$ deposition (left) and wet SO${}_{\mathrm{4}}$ deposition (right) (base case $-$ no ship).

[Figure omitted. See PDF]

Conclusions

Although regulations for emissions from the maritime traffic sector – especially for sulfur – have been tightened over the last few years, the impacts are limited at present in Europe, since there are no NECAs yet and the IMO emission limits refer only to new ships. The European Environment Agency estimated that emissions of nitrogen oxides from international maritime transport in European waters could be equal to land-based sources by 2020. The model results presented in this study give an overview of the effects of ship emissions on the concentrations and depositions of air pollutants in Europe, based on the 2006 emission inventory.

Our results suggest that emissions from marine engines cause a decrease of 10–20 % in annual surface ozone in the area of the English Channel and the North Sea, but they lead to an increase (5–10 %) in the Mediterranean Sea. There was a difference in the seasonal variation between north and south. Ship emissions were predicted to cause a decrease in ozone in the north, covering the area of the English Channel, the North Sea and the Baltic Sea in all seasons except summer. Ozone decreased in summer due to ship traffic only around the English Channel, while it increased by about 5 % in the North and the Baltic seas. On the other hand, an opposite effect was predicted for the southern part of the model domain. Emissions from shipping led to an increase in the surface ozone in all seasons except winter. In contrast, a small decrease in winter ozone was predicted along the shipping routes, especially in the western Mediterranean. Based on these results, we conclude that ship emissions cause an increase in ozone in seasons with active photochemistry (i.e., summer in the north and spring to fall in the south).

The PM${}_{2.5}$ concentrations increased by up to 45 % in the Mediterranean Sea, and 10–15 % in the North Sea, Baltic Sea and along the coastal areas, due to ship traffic. The impacts predicted for the Mediterranean region are larger than those reported in other studies. The finer resolution used in this work captured the local effects more accurately. Significant effects of ship emissions on the air quality were predicted not only along the shipping routes but also over a large part of the European continent. Although increased concentrations of primary organic aerosols and elemental carbon were predicted only along the shipping routes, secondary pollutants were affected over a larger area. The effects of ship emissions were larger in summer, predominantly on secondary inorganic aerosols, whereas SOA concentrations increased by less than 10 %. One should keep in mind, however, that the results for SOA might look different if a VBS scheme is used to calculate the OA concentrations, but this could not be done in this study due to lack of information about the volatility distribution of ship emissions. Ship emissions increased the particulate sulfate concentrations in the Mediterranean as well as in the North Sea. On the other hand, particulate nitrate concentrations increased due to the NO${}_x$ emissions from shipping, especially around the Benelux area, where there are high land-based NH${}_{\mathrm{3}}$ emissions.

Consumption of gaseous NH${}_{\mathrm{3}}$ for particulate nitrate formation resulted in a small decrease in the dry deposition of reduced nitrogen in its source regions and an increase in wet deposition along the shorelines with high precipitation rates. Deposition of nitrogen was predicted to increase in the Mediterranean mainly due to an increase in the deposition of oxidized nitrogen compounds (mainly HNO${}_{\mathrm{3}})$. On the other hand, the increase in dry deposition of SO${}_{\mathrm{2}}$ along the shipping routes was larger than the increase in wet deposition of SO${}_{\mathrm{4}}$ along the Scandinavian and Adriatic coast.

The model results achieved in this study suggest that emissions from ship traffic have significant impacts on air quality, not only along the shipping routes but also over a large part of the European continent. While SO${}_{\mathrm{2}}$ emissions in European waters will continue to decrease due to regulation of the sulfur content in marine fuels, NO${}_x$ emissions are expected to increase further in the future and could be equal to or even larger than the land-based emissions from 2020 onwards. Impacts of regulations for NO${}_x$ emissions from marine diesel engines are expected to be limited in the near future.

In an earlier study, we predicted that there would be a significant reduction in PM${}_{2.5}$ ($\sim$ 30 %) and in oxidized nitrogen deposition ($\sim$ 40 %) in Europe by 2020 compared to 2005, assuming a baseline scenario where land-based emissions were reduced according to the Gothenburg Protocol scenarios (Aksoyoglu et al., 2014). Increasing emissions from marine transport, however, might partly outweigh the benefit of reductions of land-based emissions.

As a final remark, we have to consider the following issues for future European air quality:

• In general, there is a clear need to improve the emission inventories to reduce the uncertainties. Since ammonia is a very important precursor for the secondary inorganic aerosol formation, more accurate estimates of its emissions are needed for future simulations.

• With significant future reductions of NO${}_x$ emissions from ship traffic, changing chemical regimes around the northern coast would affect the impacts on ozone as well as the formation of secondary inorganic aerosols. Decreasing NO${}_x$ $/$ VOC ratios would affect ozone formation, whereas decreasing NO${}_x$ $/$ NH${}_{\mathrm{3}}$ ratios might change the formation of secondary inorganic aerosols as well as nitrogen deposition, since ammonia land emissions are not expected to decrease significantly in the near future compared to sulfur and nitrogen emissions in Europe.

The Supplement related to this article is available online at doi:10.5194/acp-16-1895-2016-supplement.

Acknowledgements

This work was carried out within the framework of the NADIP project (NO${}_x$ Abatement in Diesels: Process analysis, optimization and impact) and partially funded by the Competence Center Energy and Mobility (CCEM) in Switzerland. We would like to thank NADIP project leader P. Dimopoulos Eggenschwiler for his support. We thank TNO for providing us with the anthropogenic emission inventory, as well as the European Centre for Medium-Range Weather Forecasts (ECMWF) for the meteorological and the global air quality model data. We gratefully acknowledge J. Keller for his contribution to input generation. Our thanks extend to ENVIRON for their continuous support of the CAMx model. Finally, we would like to thank the editor and the referees for their useful comments and suggestions that helped to improve the clarity and readability of the paper. Edited by: N. Mihalopoulos