Spatial distribution and seasonal variations of atmospheric sulfur deposition over Northern China

Introduction Conclusions References

precipitation patterns in Northern China.However, the spatial and inter-annual differences in the wet deposition were not significant, which were influenced by the precipitation amount, scavenging ratio and the concentrations of atmospheric S compounds.In contrast, the relatively large dry deposition of SO 2 and sulfate during cold season, especially at industrial areas, was reasonably related to the local emissions from home heating.Although seasonal fluctuations were constant, clear spatial differences were observed in the total S deposition flux and higher values were also found in industrial areas with huge emissions of SO 2 .These findings indicate that human activity has dramatically altered the atmospheric S deposition and thus regional S cycles.To systematically illustrate the potential effects of acidifying deposition on the receiving environment, we calculated the deposition of "potential acidity" that takes into account the microbial transformation of ammonium to nitrate in the ecosystems, resulting in the release of hydrogen ions.The estimated total "acid equivalents" deposition of S and

Introduction
Acid deposition, originating largely from man-made emissions of three gaseous pollutants, sulfur dioxide (SO 2 ), nitrogen oxides (NO x ) and ammonia (NH 3 ), is a global environmental issue because of its transboundary impacts of the biogeochemical cycles (Rodhe et al., 2002;Flechard et al., 2011).Of concern is the acid deposition that exceeds the critical loads of an ecological system will cause long-term harmful effects, with several deleterious consequences ranging from eutrophication of coastal waters to acidification of lakes, streams, and soils with attendant reductions in species diversity (Reuss et al., 1987;Sickles et al., 2009).With acid deposition behaving as one of the most important environmental issues, decision-makers need to determine which emission controls are needed and to what extent emission reductions will promote ecosystem recovery from acid deposition (Driscoll et al., 2001).To address these concerns, it is essential to gain a quantitative knowledge of acid deposition through monitoring, including SO 2 , NO x and NH 3 as well as particulate and aqueous forms such as sulfate (SO 2− 4 ), nitrate (NO − 3 ) and ammonium (NH + 4 ).Figures

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Full In recent decades, the increasing emissions of acidic compounds of sulfur (S) and nitrogen (N) make East Asia one of the most serious acid deposition regions following Europe and North America (Endo et al., 2010;Ayers et al., 2000;Kuribayashi et al., 2012).As one of the most rapidly developing nations in East Asia, China is exposed to acid deposition levels that are amongst the highest around the world (Larssen et al., 2006), with a major contribution from reduced N (NH x ) relative to oxidized N (NO y ) (Larssen et al., 2011) .Under strong environmental pressure, the Chinese government has implemented a variety of control programs including the adjustment of energy structure (Zhao et al., 2009).As a consequence, the increasing trend of annual SO 2 emissions has leveled off since 2006, after which a declining trend has been seen (Lu et al., 2010).In contrast, reactive N emissions by intensive farming and livestock (NH 3 ) and by traffic and industry (NO x ) are still rising (Liu et al., 2010).These changes in emissions may enhance the N deposition, and thus change the acid deposition pattern in the future.This argument is partially supported by the fact that the ratio of equivalent concentrations of SO 2− 4 to NO − 3 in precipitation gradually decreased from 5.4 to 1.7 in Beijing (Wang et al., 2012) and from 4.6 to 1.5 in Guangzhou (Fang et al., 2011) during the last few decades.If this trend continues, HNO 3 is expected to play a more important role than H 2 SO 4 in determining precipitation acidity in the coming decades.However, the above ratios of larger than unity indicate that precipitation acidity in China was still mainly caused by H 2 SO 4 , formed via the oxidation of SO 2 (Wang et al., 2012).
Although national SO 2 emissions will reduce approximately by 2 % yr −1 from 2010 to 2020 if current policy is fully implemented, the exceedance of the critical loads for soil acidification caused by S deposition is still higher than 20 % of the country's territory in 2020 (Zhao et al., 2009).Monitoring information on the S deposition is thus necessary to check whether emission reduction measures are working efficiently (Larssen et al., 2011).Furthermore, enhanced spatial resolution of acid deposition pattern is a key requirement in the application of the critical loads approach to emission control policies (Aherne and Farrell, 2002).Nevertheless, long-term measurement data of acid Introduction

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Full deposition are not yet available on a regional scale in China, although the importance of such data has been emphasized previously (Larssen et al., 2006).Acid deposition monitoring networks have been established in many regions around the world, such as Europe (EMEP), US (CASTNET), Canada (CAPMoN), East Asia (EANET) and Africa (IDAF).In China, acid deposition monitoring conducted routinely across the whole country is lacking.Monitoring of acid rain started in the 1980s, but at few stations only, with a less suitable device setup and incomplete data series (Zhao et al., 1988).Under current national monitoring networks, only general urban air quality parameters and rainwater pH are measured, whereas little has been done involving chemistry determination (Tang et al., 2010).A rapid advance in recent years has been made to investigate the chemical composition responsible for the variations of precipitation pH (Wai et al., 2005;Huang et al., 2008;Sun et al., 2010;Yang et al., 2012;Tsai et al., 2011), but most are limited to a single site/year.Acid deposition can occur by dry deposition of gases and particles and by wet deposition from rain, snow, clouds and fog (Flechard et al., 2011).Thus, data presented only for precipitation chemistry were not enough to accurately assess effects of acid deposition; both the wet and dry deposition flux of S and N must be known.Acid deposition is also spatially and temporally variable (Larssen and Carmichael, 2000), highlighting an urgent requirement to organize an up-to-date national deposition network across the country.Although some sites in Southern China have participated in the EANET, however, the EANET coverage is very sparse in Northern China.To date, fewer measurement results have been focused on N deposition in Northern China (Shen et al., 2009), and field-based data regarding S deposition, in particular dry deposition, are even more scarce.
Northern China is a densely populated and industrialized region with intensive agricultural activities.Associated with the increasing energy use, economic boom and population growth during the past several decades, the substantial emissions of acidic precursors (SO 2 , NO x and NH 3 ) have intensified the acid deposition problem over the target areas.For example, the precipitation acidity in China remained steady or showed a decreasing trend during the 1990s.After 2000, however, an increasing trend Introduction

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Full of precipitation acidity was observed at many sites in Northern China (Tang et al., 2010).Besides nutrients, anthropogenic emissions of toxic trace elements and organic pollutants are of concern regarding the related impacts on receiving terrestrial and aquatic ecosystems (Sakata et al., 2006).However, measuring wet and dry deposition of such matters has so far received little attention.To advance our current understanding of their transport and fate from local to regional and global scales, quantitative assessment of atmospheric deposition is required to update with detailed spatial and seasonal descriptions.Therefore, a new monitoring network, including ten sites well distributed within the target area, was set up in late 2007.The focus of the program is to study the wet and dry deposition of important trace species, including carbon, N, S, phosphorus, heavy metals, and polycyclic aromatic hydrocarbons.The network and the resulting scientific research identified that the acidity of precipitation in Northern China generated by H 2 SO 4 and HNO 3 was primarily buffered by CaCO 3 and NH 3 prior to deposition (Wang et al., 2012).The neutralization of precipitation acidity by alkaline species in the air, however, does not necessarily reduce the potential for acidification of aquatic and terrestrial ecosystems (Galloway et al., 1984).Evidently, a solution of CaSO 4 can cause short-term acidification of surface waters as a result of cationexchange reactions in acid soils (Galloway et al., 1987).Neutralization by NH 3 is also illusory, because the resulting NH + 4 contributes to acidification of soil and surface water through chemical processes in the soil (Larssen et al., 2006;Driscoll et al., 2001).Therefore, NH x must be taken into account when quantifying the potential amount of total acid deposition.Additionally, how and to what extent the S and N are contributing to the total acid deposition is largely unknown, though indications of high deposition flux of N have been reported in a companion paper (Pan et al., 2012).
Here we complement the previous study by additionally estimating atmospheric deposition flux of S via precipitation (wet deposition) and as gases and particles (dry deposition) on the basis of a multi-year monitoring at ten selected sites.The objective of this study was to explore the spatial and temporal variations of atmospheric S deposition and its contribution to total acidifying deposition in Northern China.Combined Introduction

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Full with N deposition, the unique dataset provides an insight into the total deposition of potential acidity on a regional scale.The knowledge gained from the study provides the up-to-date scientific basis for the validation of emission inventories and atmospheric chemistry models and the crafting of control strategies to reduce their emissions and ecological impacts in the target area.
2 Materials and methods

Site descriptions
Continuous measurements were carried out during the period from December 2007 to November 2010 at ten sites using the uniform monitoring protocol.Based on the urban geographies, energy structures and ecosystem types, the ten sites were classified into five categories: the two megacity sites located in the Beijing (BJ, 39.96In order to interpret the geographical distributions of the estimated S deposition, emission patterns of the precursor SO 2 have to be known.Recently, an Asian emission inventory for the year 2006 was developed that includes a database for the study area (Fig. 1a), based on the INTEX-B experiment (Zhang et al., 2009).Hereafter, we refer to this emission inventory as the "INTEX-B".Anthropogenic SO 2 released to the Introduction

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Full atmosphere in the INTEX-B was estimated to be 248, 336 and 2281 Gg in Beijing, Tianjin and Hebei province, respectively, making Northern China one of the main "hot spots" of SO 2 emissions in Asia.Power plants and industry are the main sources of SO 2 , contributing more than 70 % of the total emissions.Emissions of SO 2 in the target region that are aggregated into grids at a resolution of 30 min × 30 min are shown in Fig. 1b, combined with the 3-yr mean total S deposition that is categorized into five grades and plotted on the maps.Comparisons between depositions and emissions are discussed in Sect.4.2.

Wet deposition measurements
Daily rainwater and event-based snow samples were collected using an automatic wetdry sampler and a clean plastic bucket, respectively.During the 3-yr observation period, a total of 1306 precipitation samples (including rainwater and snow) were collected.All of the precipitation samples were stored, transported, filtered, and analyzed following the prescribed procedure (Pan et al., 2010a).Of the total number of precipitation samples, 16 % were discarded from final analysis due to low volumes (< 20 ml) that did not permit a complete chemical analysis.The concentrations of SO 2− 4 in the filtrates of the remaining 1107 samples were determined using an ion chromatography system (Model ICS-90, Dionex Corporation, Sunnyvale, CA, USA) and the standard laboratory procedure of the State Key Laboratory of Atmospheric Boundary Layer Physics and Atmospheric Chemistry (Wang et al., 2012)

Particulate dry deposition measurements
Dry-deposited particles were collected monthly using a surrogate surface made of polyurethane foam (PUF) filter, which was placed in a glass bucket.The sampling and chemical analysis procedures have been detailed previously (Pan et al., 2012) and are only described briefly here.During a rain event, the glass bucket was covered with a lid of the automatic wet-dry collector.After the rainfall ceased, the lid was lifted and rotated to cover the aperture that collected the rainwater.This feature of the automatic sampler is highly useful for simultaneously collecting rainwater and deposited particles with minor mixing of the two.After collection, the content of water-extractable SO Measuring dry deposition is known to be much more challenging (Wesely and Hicks, 2000) than that of wet deposition, in terms of measuring the amount and composition of precipitation.Due to gravitational settling, large particles will get trapped easily by PUFbased method used in the study.However, it is difficult to capture fine particles, because impaction and interception with vegetative canopies are the main processes determining their fallout from the air, and such a phenomenon is not reproduced in the design of any standardized artificial collection device (Wesely and Hicks, 2000).Therefore, collection of the dry-deposited particles using surrogate surface method has potential underestimation and the p SO 2− 4 measured in the study represents a rather conservative estimation.Despite these precautions, the method can avoid particle bounce and is a cost-effective and convenient way to delineate spatial and temporal information of dry-deposited particles at a variety of locations and over varying time intervals (Pan et al., 2010b).Introduction

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Gaseous dry deposition measurements
The inferential method, which combines the measured concentration and a modeled dry deposition velocity (V d ), was used to estimate the gaseous dry deposition flux (Schwede et al., 2011).The method has been used successfully to estimate the dry deposition of reactive N species including NO, NO 2 and NH 3 .Gaseous measurements of SO 2 are monthly integrated samples using passive diffusing techniques (Analysts, CNR -Institute of Atmospheric Pollution, Rome, Italy).The ambient SO 2 concentrations were calculated from the SO 2− 4 levels in the extracts of diffusive samplers corrected by the local temperature and humidity conditions (Costabile et al., 2006).To test the suitability of this sampling technique, the concentration of SO 2 was parallel measured by using a pulsed UV fluorescence analyzer (Model 43i, Thermo-Fisher Scientific, Co., Ltd., Franklin, MA, USA).The results from the passive sampler exposed for one month compare well with the average SO 2 values measured by the in situ continuous active analyzer during the corresponding period, indicating that the passive technique used is reliable for such a study (Wu et al., 2010).
In this study, the hourly V d of SO 2 and other gases (O 3 , NO, NO 2 and NH 3 ) was simulated using the Models-3/Community Multiscale Air Quality (CMAQ v4.6) system (Byun and Ching, 1999).In its second dry deposition scheme (M3Dry) of CMAQ v4.6, the well-known resistance parameterization for V d was introduced (Wesely, 1989).Three resistances are considered in this theory: aerodynamic resistance (R a ), quasi-laminar resistance (R b ), and surface or canopy resistance (R c ), among which the R c is the most difficult one to simulate (Zhang et al., 2008).The V d is defined as (1) This model has proven to be suitable for regional air pollution and acid deposition simulations in China (Zhao et al., 2009).The driving meteorological inputs were provided by the fifth-generation NCAR/Penn State Mesoscale Model (MM5 v3.7).The ambient concentrations of SO 2 measured by passive samplers were multiplied with the average Introduction

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Full V d during the corresponding period to estimate the monthly gaseous dry deposition flux of SO 2 ( g SO 2 , kg S ha −1 month −1 ).
We are however aware that the inferential method is not the best one to clarify gaseous dry deposition flux but an alternative operational tool to compensate for the absence of measured data at regional scales (Flechard et al., 2011).While uncertainties existed in the V d simulations (Pan et al., 2012), the SO 2 flux is evaluated with the inferential method to give an idea of its contribution to the S budget and thus acid deposition, keeping in mind that it has to be validated with micrometeorological measurements in future field campaigns.

Statistics
To evaluate the total atmospheric S deposition flux ( t SO x ), we compiled the flux data in terms of S from dry-deposited particles, gas and precipitation (including rain and snow) chemistry measurements.Annual w SO2− 4 , p SO 2− 4 , g SO 2 , and t SO x (generally expressed in the units of kg S ha −1 yr −1 ) are integrated from monthly flux datasets.
Total potential acidifying deposition flux (N plus S, hereafter as total acid deposition) is given in mole of protons (H + ) ha −1 yr −1 (converted to keq ha  A one-way analysis of variance (ANOVA) and nonparametric tests were performed to examine the significance of differences in the annual flux of S for all ten sites over the three years of the study.A linear regression analysis was used to investigate relationships between precipitation and the wet deposition of SO Introduction

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Wet deposition of sulfate
The mean annual w SO 2− 4 at the ten sites was relatively high and varied from 14.6 to 24.7 kg S ha −1 yr −1 , with an average of 19.6 kg S ha −1 yr −1 during the 3-yr period (Fig. 2).
The largest values were observed at the TG site (24.7 kg S ha −1 yr −1 ) followed by the CZ site (22.9 kg S ha −1 yr −1 ).The w SO The relatively low flux occurred at the TJ site (15.5 kg S ha −1 yr −1 ), and, as expected, the 3-yr mean values were lowest at the XL site (14.6 kg S ha −1 yr −1 ).However, the differences were not significant (p > 0.05) between the sites or the different years, thus indicating the absence of a geographic pattern in the spatial distribution of w SO 2− 4 .The mean monthly w SO 2− 4 monitored for the three years ranged from 0 to 6.1 kg S ha −1 month −1 (Fig. 3a).Seasonal variations of w SO 2− 4 showed similar trends at each site, with a maximum in summer and minimum in winter.The w SO 2− 4 was high in June, July, August and September which is in agreement with the rainy season in Northern China.Indeed, the ten-site 3-yr average total w SO 2− 4 during the summer (June to August) was 11.5 kg S ha −1 , which accounted for 58 % of the annual mean w SO 2− 4 .Accordingly, the rainfall amount received was 364 mm during the summer in the study area, which contributed 66 % of the annual precipitation.In contrast, the minimum w SO 2− 4 level obtained in the winter months was attributable to a low precipitation.In general, a significant linear correlation between the monthly w SO 2− 4 and precipitation was observed at each site (0.52 < r 2 < 0.93, p < 0.001); therefore, the precipitation amount is likely to be very important in explaining the seasonal trend of w SO the precipitation amount between the BJ and YF sites has no significant difference, we suggest that the notably high flux observed in these two months at the YF site is due to local emissions.Additionally, a large standard deviation of w SO 2− 4 was found both in March and May at the YF site.However, such fluctuations can be well explained by the variable depths of precipitation in different years.

Dry deposition of particulate sulfate
The highest mean annual p SO 2− 4 during the 3-yr period was found at the TS site (24.8 kg S ha −1 yr −1 ) (Fig. 2), with monthly mean values at this site significantly higher than other sites with the exception of TJ and BD (p > 0.05).The second highest result was observed at the TJ site (19.5 kg S ha −1 yr −1 ), followed by the BD, LC and TG sites, with values of 15.2, 13.4 and 13.1 kg S ha −1 yr −1 , respectively.However, the monthly mean p SO 2− 4 at these sites was not significantly different (p > 0.05).The p SO 2− 4 was comparable for the YC, CZ, BJ and YF sites with values of 10.3, 9.6, 9.3 and 8.4 kg S ha −1 yr −1 , respectively; differences in the monthly mean values at these sites were also not significant (p > 0.05).Annual mean p SO 2− 4 was lowest at the XL site (4.4 kg S ha −1 yr −1 ), mostly reflecting less emissions surrounding the rural site.
The monthly mean flux at this site differed significantly from those at the other sites (p < 0.05).In general, annual mean p SO 2− 4 was higher at industrial and urban sites (e.g.TS and TJ) than at suburban and rural sites (e.g.CZ and XL).This observed spatial variation indicates that the annual p SO 2− 4 strongly depends on emissions, increasing with large fossil fuel combustion.Although the LC site is located in an agricultural area, the influence of SO x plume emitted from industry sources of Shijiazhuang city (the capital of Hebei province) might be the reason why the relatively high annual During the 3-yr observation campaign, the mean monthly p SO 2− 4 ranged from 0.1 to 5.1 kg S ha −1 month −1 (Fig. 3b).At most of the sites, the p SO 2− 4 was higher in the winter and the early spring than in the other seasons, which coincides with the period of home heating in the target areas.This is more pronounced in industrial and urban areas with high population density and thus huge coal consummation even on a relatively small spatial scale (e.g.TJ, TS and BD).In contrast, the lower atmospheric SO x emissions and efficient wet deposition result in lower p SO 2− 4 during summertime.Different from most sites, the p SO 2− 4 was elevated in June, August and October at the TG, LC and CZ sites, respectively, which is similar to that of w SO 2− 4 .These notably high values might be attributable to local sources.In general, the mean monthly p SO 2− 4 was not significantly different (p > 0.05) between different years at most of the sites with the exception of BD, BJ, CZ, TG and TS, thus indicating the inter-annual variations of p SO 2− 4 .At these five sites, the p SO 2− 4 was found to be lower in 2009 than 2008 and 2010.

Dry deposition of sulfur dioxide
The mean annual g SO 2 during the 3-yr period is presented in Fig. 2. The results demonstrated a relatively high g SO 2 at the industrial TS and urban TJ sites (55.2 and 48.8 kg S ha −1 yr −1 , respectively); regardless, the difference of monthly means between them was not significant (p > 0.05).The mean annual g SO 2 was high and comparable at the LC, BD, TG, CZ and YC sites with values of 38.9, 37.8, 33.6, 28.9 and 27.3 kg S ha −1 yr −1 , respectively, and no significantly different monthly mean values were found among the former four sites (p > 0.05).The spatial variations of the g SO 2 were similar to those of p SO 2− 4 , both of which illustrated higher flux at certain industrial and urban sites than that of agricultural and rural sites.In general, the monthly mean g SO 2 was significantly higher at the TJ and TS sites than that of other site with the exception of BD and LC (p > 0.05).In contrast, the monthly mean g SO 2 at the BJ, Introduction

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Full The monthly mean g SO 2 during the 3-yr period ranged from 0.4 to 9.0 kg S ha −1 month −1 (Fig. 3c).Between March and September, the seasonal variation of g SO 2 was not distinct at most of the sites, probably because they mainly come from industrial emissions that have less of a seasonal cycle.In contrast, the higher g SO 2 values were found during winter and late autumn (November) at the sites of TS, TJ, LC, TG and BD.The notably high g SO 2 in cold months suggests an important source, which is related to the fact that extra coal is consumed for home heating.
The year-to-year variations in the annual average g SO 2 were comparatively small at each site with the exception of BD, CZ, TG, TJ and TS, which showed significantly lower values in 2010 than in 2009 or 2008 (p < 0.05).This is inconsistent with the decreasing trend of SO 2 emissions after the year 2006 due to sources control measures implemented in China.Since the gas dry deposition flux is determined by both the concentrations and V d of the target species, the flux estimates can be significantly different because of differences in the simulated V d (Pan et al., 2012).We believe the concentration of SO 2 may serve as a better index than the flux for tracking the inter-annual trends of emissions.As expected, it was observed that the SO 2 concentrations (data not shown) declined year by year in the target area.In general, the ten-site average concentrations of SO

Atmospheric total deposition of sulfur in Northern China
Several measurements have been conducted in China in an effort to evaluate the wet and dry deposition of SO x separately, but few have been performed to determine them simultaneously, especially including dry deposition.To our knowledge, this has not been reported in Northern China before, though previous modeling studies illustrated that S deposition will be enhanced.The mean annual t SO x estimated in this study ranged from 35.0 to 100.7 kg S ha −1 yr −1 over the 3-yr sampling period at the ten sites (Fig. 2).With the exception of BD, LC and TJ, the year-to-year variations in the annual average t SO x at most of sites were not significant (p > 0.05).Nevertheless, a declining tendency was found for the sites of BD, BJ, LC and TS, while the inter-annual variations at the other sites were more ambiguous.The discrepancy of inter-annual t SO x trends between sites is reasonable, because the year-to-year variations of w SO 2− 4 and p SO 2− 4 do not show any marked trend.As presented in Fig. 3d, the monthly mean t SO x during the 3-yr period ranged from 1.4 to 11.7 kg S ha −1 month −1 ; however, the seasonal trends at most sites did not significantly differ, indicating that the t SO x was constant throughout the entire year as a result of opposite seasonal trends of wet vs. dry deposition of S at a given site.The spatial variation of t SO x was in agreement with that of gaseous and particulate dry deposition; the values at the LC, BD, TG, CZ and YC sites were higher than those at the BJ, YF and XL sites and lower than those at the TS and TJ sites.The overall mean t SO x in Northern China was 64.8 kg S ha −1 yr −1 during the 3-yr period, which was approximately 4 to 10 times that observed by CASTNET, EMEP and EANET with values of 6.2, 6.2 and 17.6 kg S ha −1 yr −1 , respectively (Endo et al., 2010).Since most sites in these abroad networks were located in less polluted areas, perhaps using the rural XL site as the critical baseline of the heavily S deposition would be more reflective of the effects of human-induced emissions in the target areas.Even then, the t SO x at Introduction

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Full the site XL (35.0 kg S ha −1 yr −1 ) was approximately 2 to 5 times that observed by the above monitoring networks, indicating a high level of S deposition in Northern China as a result of huge SO 2 emissions.In addition, the total S deposition at the XL site was likely underestimated because fog water is an important atmospheric deposition process in mountainous forest areas but not measured in this study; however, the rate of S deposition through fog, wet and dry deposition to high-elevation forests in the US (16.3 kg S ha −1 yr −1 ) (Miller et al., 1993) and Japan (16.7-27.7 kg S ha −1 yr −1 ) (Shimadera et al., 2012) was still significantly lower than the present study.In the future, north-central and eastern China is expected to receive the most S depositions by the CMAQ simulations, attributed mainly to high densities of energy consumption and emissions (Zhao et al., 2009).Besides the present field-evidence data, the predicted spatial pattern of S deposition in China can be further confirmed by the fact that the total airborne SO x input into the farmland in Jiangxi, China is estimated to be 103 kg S ha −1 yr −1 (Wang et al., 2003), which is greater than the results found for the agricultural sites of LC and YC (70.2 and 58.0 kg S ha −1 yr −1 , respectively) in this study.Therefore, the sufficiently high S deposition in vast areas in China suggests the need for integrated assessment of the effects of excess S deposition on different ecosystems.

Anthropogenic imprints on sulfur deposition in Northern China
In order to address the above concerns and thus help policy-makers mitigate S deposition, the sources of SO 2 have to be known.Evidently, several points may support the significant influences of human activity on the S deposition in the target areas.First, the relatively high t SO x was found at the sites of TS, TJ, BD and LC, which is surrounded by large SO 2 emissions sources (Fig. 1b).Although the discrepancy between emissions and depositions observed at the sites of BJ and YC may be attributable to the overestimations in current gridded inventories of INTEX-B, the relatively low t SO x at the XL and YF sites indicated fewer anthropogenic emissions.Overall, the t SO x tends to Introduction

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Full increase in relation to human-induced SO 2 emissions and a spatial pattern is identified here.
Second, it has been noted elsewhere that dry deposition of trace species is dominated by local emissions (Balestrini et al., 2000); the idea was verified here.The sites that had significantly higher dry deposition of S were mostly located in the industrial areas such as TS, TJ and BD, or influenced by the industry plume, in particular LC.Although the wet deposition, when considered in terms of annual flux, showed no significant differences between sites, the high local emissions near the sites had apparent effect on the precipitation flux.With the aim of identifying potential factors influencing the spatial pattern of wet depositions, we performed a statistical analysis of data from the 3-yr period following previous studies (Sakata et al., 2006).The results illustrated a moderate linear relationship between the annual w SO 2− 4 and the corresponding precipitation amounts (p < 0.05).However, only 23 % of the variance of the w SO 2− 4 (r 2 = 0.23) was explained by the annual precipitation (Fig. 4), implying marked differences in the scavenging ratio and the atmospheric concentrations of SO x across Northern China, which is similar to that of NH + 4 (Pan et al., 2012).Specifically, the annual w SO 2− 4 at the sites of BD, LC and TG tends to be much higher than that expected from the precipitation amount, which is indicative of large contribution from man-made emissions.Such a contribution also resulted in peak values of w SO 2− 4 in September and October at the CZ site (Fig. 3a), and in March and May at the YF site, as discussed in Sect.3.1.
Third, in dry deposition the seasonal trend of g SO 2 and p SO 2− 4 was ostensibly linked to local anthropogenic emissions.The relatively high values of dry deposition in cold months clearly suggested the importance of emissions from coal combustion sources for the total S deposition.In winter, the SO 2 concentration from passive samples was about 3 times that in summer, explained as due to an additionally higher contribution from coal burning during home heating periods.Furthermore, significant fluctuations of g SO 2 between most of industrial sites in late autumn and winter, in association with SO 2 concentrations, reflected the differences in the emission strength.Introduction

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Contribution of different pathways to total S deposition flux
The g SO 2 was the major contributor to t SO x at most sites, and the contribution ranged from 39 % (BJ) to 58 % (TJ), averaging 49 %.Only at the megacity BJ and its suburban site YF was the contribution of g SO 2 less than that of w SO 2− 4 .In contrast, the contribution of S deposition through g SO 2 was larger than 50 % at the sites of TJ, BD, TS and LC, which is due in part to the proximity of the sites to high emission areas.The mean annual g SO 2 in this study reached 32.4 kg S ha −1 yr −1 .This result is much higher than that modeled in South Korea (3.5 kg S ha −1 yr −1 ) (Park and Lee, 2003) and Japan (approximately 3.1 kg S ha −1 yr −1 ) (Endo et al., 2010), and values determined over a coniferous canopy in The Netherlands (4.3 kg S ha −1 yr −1 ) using the aerodynamic gradient technique (Wyers and Duyzer, 1997).Given the influence of local SO 2 emissions from intensive industrialization, it is not surprising that the g SO 2 at the TJ and TS sites reached 48.8 and 55.2 kg S ha −1 yr −1 , respectively.However, these values are much lower than the result from the farmland in Jiangxi, China (79.6 kg S ha −1 yr −1 ), which could be attributable to the influence of SO 2 emissions from anthropogenic sources (Wang et al., 2003).The influence of SO 2 emissions could also explain why the sites BD and LC had a higher g SO 2 value than the other agricultural, urban and rural locations, including YC, BJ, YF and XL (p < 0.05).The w SO 2− 4 contributed 19-43 % to t SO x with a ten-site, 3-yr mean of 32 %, which was 17 % lower than that of g SO 2 .Of note is the fact that the contribution of w SO 2− 4 to t SO x exceeds that of g SO 2 at the BJ and YF sites, implying that precipitation was the largest contributor to the total S deposition.The ten-site average w SO 2− 4 in Northern China was 19.6 kg S ha −1 yr −1 from 2008 to 2010.This result is comparable to the values from Jiangxi (17.2 kg S ha −1 yr −1 ) (Wang et al., 2003) and the other Chinese subtropical forest catchment (14.4-30.4kg S ha −1 yr −1 ) (Larssen et al., 2011), but much higher than the modeled w SO (6 %) (Wang et al., 2003).The relatively high SO 2 dry deposition and its role played in the total S deposition probably reflected heavy SO 2 pollution in China emitted from anthropogenic sources.

Acid deposition in Northern China and its implications
Acid deposition of SO 2− 4 has been widely documented due to concerns about the impacts of this deposited compound on the acidity of the environment.Besides SO 2− 4 , however, NO − 3 , NH + 4 , NO x , NH 3 , and SO 2 also contribute to the acidification of ecosystems.Thus, there is a necessity for quantification of total acid deposition to inform the control strategies for atmospheric pollution.Specifically, the total S deposition in this study ranged from 2.2 to 6.3 keq ha −1 yr −1 with a ten-site mean of 4.1 keq ha −1 yr −1 (Fig. 5), which is much higher than the critical loads (< 2.0 keq ha −1 yr −1 ) for acidification in most areas of Northern China (Zhao et al., 2009).Since the critical loads are defined as the indicator of ecosystem sensitivity, the total S deposition exceeding the critical loads indicated that harmful ecological effects will happen or may have already occurred (Larssen et al., 2011).
With the aim of determining the contribution of acidifying species into the environment, the total deposition of "acid equivalents" of S and N was estimated by summing the wet and the particulate dry deposition flux of SO  (Pan et al., 2012), the total acid deposition ranged from 4.2 to 11.6 keq ha −1 yr −1 with a ten-site mean of 8.4 keq ha −1 yr −1 (Fig. 5) falling in the range of 1.9-13 keq ha −1 yr −1 in forested catchments in Southern China (Larssen et al., 2011).However, acid deposition in the study is roughly 10 times the results in Malaysia (0.3-0.5 keq ha −1 yr −1 ) (Ayers et al., 2000), and is also much higher than previous estimates from high-elevation forests in the US (2.2 keq ha −1 yr −1 ) (Miller et al., 1993) and the values (0.8-1.9 keq ha −1 yr −1 ) reported by CASTNET, EMEP and EANET (Endo et al., 2010).The acid deposition value in Northern China is sufficiently Introduction

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Full high so as to suggest the need for integrated assessment of acidic deposition and potential environmental consequences in the region (Zhao et al., 2009;Larssen et al., 2011;Hao et al., 2001).
The atmospheric deposition of S has been of interest for many years; however, the importance of N on acidification risks is clearly demonstrated in our study if the acidification effect of N is combined with that of S. As illustrated in Fig. 5, the contribution of N to acid deposition was higher than that of S at most of sites with the exception of TS, TJ, BD and XL.The relatively high contribution of S to total acid deposition at these four sites highlights that further SO 2 abatement in industrial regions is still needed, further considering the large S deposition in rural areas of Northern China.It should be noted that the contribution of S to total acid deposition was almost in the same range as that of N at most of sites (Fig. 5), with the exception of the BJ, TG, YC and LC sites where NH 3 deposition is large (Pan et al., 2012).Thus, to mitigate acid deposition and its potential impacts on receiving ecosystems, a multi-pollutant perspective and joint control strategies to abate SO 2 and NH 3 simultaneously in the target areas are recommended (Zhao et al., 2009).

Conclusions
The proportion of cities in Northern China experiencing acid rain (pH < 5.6) has increased during recent years, arousing great concerns about the effects of the deposited material on the acidity of the environment.Compared with precipitation, atmospheric dry deposition of acidifying species is suggested to be an important process determining their lifetime in air and their input to various ecosystems.Furthermore, the acid deposition is likely to be much greater in Northern China because of large emissions of reactive N and S species.However, few studies have been carried out over the target area to simultaneously quantify both the wet and the dry deposition.Additionally, the contribution of S to acid deposition remains unclear, though N deposition has been reported in our companion paper (Pan et al., 2012).These are, to our knowledge, the Introduction

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Full ecosystems, thus prompting concerns regarding ecological impacts.The contribution of S to total acid deposition is on the same level as that of N at most of sites; however, the acidification risks of S were more pronounced in the industrial areas.To mitigate acid deposition in the target areas, priority should be given to strategies for abating SO 2 and NH 3 simultaneously.Introduction

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Full Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | . Monthly wet deposition flux of SO 2Discussion Paper | Discussion Paper | Discussion Paper | PUF filter was determined by the ion chromatography system.Monthly dry depothe PUF filter by the surrogate surface area and the corresponding sampling period.
Screen / Esc Printer-friendly Version Interactive Discussion Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Screen / Esc Printer-friendly Version Interactive Discussion Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | release of H + .N deposition data are presented in Pan et al. (2012).
at a given site.Compared with other sites, the w SO 2− 4 at the YF site showed a different seasonal pattern with relatively high values in March and May.Discussion Paper | Discussion Paper | Discussion Paper | at the LC site.Compared with TJ and TG, another megacity of BJ and its suburban site YF had relatively low p SO 2− 4 , probably reflecting the restriction of SO x emissions from industry and the relocation of steel industry from BJ to TS. Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | of 19.5, 18.2 and 16.0 kg S ha −1 yr −1 , respectively, which is indicative of fewer sources of emissions.
2 in 2009 reduced approximately by 18 % compared to 2008 and further declined 36 % from 2009 to 2010, which corresponds well to the decreasing tendencies of SO 2 emissions in recent years since 2006.Although differences between the three years are not significant at the 0.05 level at some sites, it is important to point out that g SO 2 also tended to decline between 2008/2009 and 2010.Overall, the tensite average g SO 2 in 2010 was 38 % and 45 % lower than in 2008 and 2009.To check whether control measures are working as anticipated, in future work it is intended to investigate the long-term trends of S deposition in the context of evaluating the impacts on ecosystems.
Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | the estimated dry deposition flux of SO 2 , NO 2 , NO and NH 3 .Taking N deposition data from our companion paper into account Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper |

Fig. 1 .Fig. 2 .
Fig. 1.Locations of the study area (a) and the sampling sites (b) in Northern China 767 with SO 2 emission distributions.The total sulfur deposition data are means of 768 three-year observations from December 2007 to November 2010.The emission data 769 of SO 2 are from 2006 (Zhang et al., 2009), with a resolution of 30 min×30 min.The 770 definition of the site codes is found in Sect.2.1.771

Fig. 3 .Fig. 4 .Fig. 5 .
Fig. 3. Seasonal variations of the atmospheric deposition flux of sulfur at the ten selected sites in Northern China.The data shown are the monthly mean ± standard deviations of 3yr observations (from December 2007 to November 2010).The definition of the site codes is found in Sect.2.1.

east of Tianjin city; two agricultural sites 6 km southwest of Yucheng city (YC, 36
.85 • N, 116.55 • E) and at Luancheng (LC, 37.89 • N, 114.69 • E), which is 4 km southeast of Shijiazhuang city; and one rural site at Xinglong (XL, 40.38 • N, 117.57• E) on Mt.Yan, with an elevation of 960 m a.s.l.Comprehensive descriptions of the stations can be found in Pan et al. (2012).