Monitoring Effect of SO2 Emission Abatement on Recovery of Acidified

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Monitoring effect of SO2 emission abatement on recovery of acidified soil and stream water in southwest China Qian Yu, Ting Zhang, Xiaoxiao Ma, Ronghua Kang, Jan Mulder, Thorjørn Larssen, and Lei Duan Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b01147 • Publication Date (Web): 03 Aug 2017 Downloaded from http://pubs.acs.org on August 4, 2017

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Environmental Science & Technology

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Monitoring effect of SO2 emission abatement on recovery of acidified

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soil and stream water in southwest China

3 4 5

Qian Yu †, Ting Zhang†, Xiaoxiao Ma‡, Ronghua Kang§, Jan Mulder§, Thorjørn Larssen#, Lei Duan†,∇,*

6 7

†

Environment, Tsinghua University, Beijing 100084, China;

8 9 10

State Key laboratory of Environmental Simulation and Pollution Control, School of

‡

State Grid Xingyuan Company Limited, Beijing 100761, China;

§

Department of Environmental Sciences, Norwegian University of Life Sciences, Box 5003, NO-1432 Ås, Norway.

11 12

#

Norwegian Institute for Water Research, Gaustadalleen 21, 0349 Oslo, Norway;

13

∇

Collaborative Innovation Centre for Regional Environmental Quality, Tsinghua

14

University, Beijing 100084, China;

15 16

ABSTRACT: Following Europe and North America, East Asia has become a global

17

hotspot of acid deposition, with very high deposition of both sulfur (S) and nitrogen (N)

18

occurring in large areas in southwest and southeast China. Great efforts have been made

19

in reducing national emission of sulfur dioxide (SO2) since 2006 in China. However, the

20

total emission of nitrogen oxides (NOx) continued to increase until 2011. In order to

21

evaluate the effects of SO2 and NOx emission abatement on acid deposition and

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acidification of soil and water, we monitored the chemical composition of throughfall,

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soil water, and stream water from 2001 to 2013 in a small, forested catchment near

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Chongqing city in Southwestern China. The deposition of S decreased significantly,

25

whereas N deposition increased in the recent years. This clearly showed the effect of

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SO2 abatement but not of NOx. Overall the rate of acid deposition was reduced.

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However, there was delay in the recovery of soil and surface water from acidification,

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probably due to desorption of previously stored sulfate (SO42-) and increase in nitrate

29

(NO3-) leaching from soil. The average acid input by N transformations has greatly

30

exceeded the H+ input directly by atmospheric deposition. The reversal of acidification

31

with an increase in pH of soil water, requires additional abatement of emissions of both

32

SO2 and NOx.

33

KEYWORDS: sulfur deposition, nitrogen deposition; acidification; sulfur adsorption;

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nitrogen saturation; Chongqing

35 36

INTRODUCTION

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Atmospheric acid deposition has been shown to affect the chemistry of soils and

38

surface waters and accelerate their acidification. Trends of soil water in forested

39

catchments and other ecosystems sensitive to acid deposition in the late 1980s and in the

40

1990s, were analyzed in many countries in Europe and North America.1~4 After the

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Convention on Long-Range Transboundary Air Pollution to reduce emissions of

42

acidifying pollutants such as sulfur dioxide (SO2) and nitrogen oxides (NOx) was

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implemented, significant recovery of acidified soil water and surface water in forest

44

ecosystems was reported.5~8

45

Since emission control in Europe and North America has taken effect, East Asia has

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become the global hotspot of acid deposition. Elevated deposition of both sulfur (S) and

47

nitrogen (N) occurs in southwest and southeast China.9~11 Acid deposition was first


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reported in China in the late 1970s, with the lowest annual average pH of rainwater and

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highest S deposition in southwest China, especially around Chongqing city.12~14 After a

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30-year continuous increase, the emissions of SO2 and NOx have been reduced at the

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national scale in recent years, following the implementation of the National SO2 Total

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Emission Control in 2006 and the National NOx Total Emission Control in 2011,

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respectively.15 However, emissions of NH3 have continued to increase because of

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technical and economic difficulties in controlling these emissions. So far, there have

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been few long-term monitoring reports of the effects of acid deposition on soils and

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waters in China.16

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Much of our understanding of the effect of acid deposition on soils and waters is

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based on work in temperate regions of Europe and North America, where mineral

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weathering is the main acid-buffering process in soils.17 Further extension of acid

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deposition to Eastern and Southern Asia, and South America in recent years,13,18,19

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warrants the collection of long-term monitoring data on soil and water acidification in

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these regions. Although acid forest soils in tropical and subtropical regions,

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characterized by slow mineral weathering and thus limited acid-buffering capacity, are

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potentially sensitive to acidification,20,21 both atmospheric base cation deposition,22,23

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denitrification,16 and S accumulation and adsorption24 may enhance the acid neutralizing

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capacity of these soils.

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Here, we use long-term monitoring data from a forested catchment near Chongqing

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city in southwest China during 2001-2013 to examine the trends in soil and water



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acidification, and to assess the effect of recent reductions in national S and N emissions.

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Accompanying the increasing importance of nitrogen oxides (NOx) and ammonia (NH3)

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emissions, the acidifying effect of nitrogen (N) deposition has become a research and

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management focus in recent years.25~27 The main objectives of this study were: (1) to

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assess the effect of SO2 emission abatement on acid deposition, (2) to investigate the

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response of soil water and stream water chemistry to changes in acid deposition, and (3)

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to evaluate the contribution of NOx and NH3 emissions to soil and water acidification.

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METHODOLOGY

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Study site. Tieshanping (29°38′N, 106°43′E), one of the five long-term monitoring

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catchments involved in the IMPACTS (Integrated Monitoring Program on Acidification

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of Chinese Terrestrial Systems) project,9 is a 16-ha subtropical headwater catchment

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about 25 km northeast from the center of Chongqing City in southwestern China (Fig.1).

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As one of the biggest cities in southwest China, Chongqing has high emissions of both

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SO2 and NOx. The SO2 emissions of Chongqing province (with an area of 8.24 Mha;

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Chongqing as the capital city) reached the highest value in 2006 (0.86 Mt yr-1) and

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decreased afterwards in response to the national SO2 emission abatement implemented

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in 2006 (Fig. 2). The NOx emission increased until 2011, when the national NOx

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abatement was implemented. Due to the relatively constant N fertilizer consumption in

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Chongqing province since 2001, it is unlikely that the emission of NH3 changed much

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during 2001-2013.


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Annual mean temperature and precipitation (1971 to 2000) at the nearest

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meteorological station in Chongqing was 18.2°C and 1105 mm, respectively. The

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bedrock of the catchment is sandstone. The forest soil is locally called Yellow Earth

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(corresponding to a Haplic Acrisol in FAO, and representative for this part of China),

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with a thin (< 1 cm) O horizon and low pH value (3.7~4.1 from the O horizon to the

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lower B horizon).24 Based on earlier studies, the Acrisol at Tieshanping has a dense

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argic B horizon at about 30 cm soil depth which reduced hydraulic conductivity, thus

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creating interflow along the hill slope.28 The major biogeochemical processes

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determining water transport and S and N transformations in soil are thus limited to the

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upper soil horizons (0-30cm). The vegetation is Masson pine (Pinus massoniana Lamb.)

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dominated forest, which is widely distributed in the humid subtropical areas of south and

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southwest China.29 The annual average pH of precipitation at the site ranged from 4.0 to

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4.2 during 2001–2004, with elevated deposition of S (10 keq·ha−1·yr−1), N (3

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keq·ha−1·yr−1, about two-thirds as NH4+ and the rest as NO3-) and calcium (Ca) (6

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keq·ha−1·yr−1).17 The high acid deposition and ambient air SO2 concentration caused

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acidification of soil and stream water, and impacted forest health.30 In general, the study

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site could present sensitive forested catchment to acidification in this region.

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Sampling and chemical analysis. In the catchment of Tieshanping we studied two 10

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m by 10 m sampling plots named C, and L (Fig. 1), each of which had four throughfall

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collectors placed in the corners and a set of ceramic suction lysimeters placed close to

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the center of the plots, at two different depths (about 3 and 30 cm) for soil water



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collection. The self-made throughfall collectors consisted of a 10-cm-diameter funnel to

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collect throughfall and an opaque bottle to store water. Plot C is located at the foot of the

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hillslope with an elevation of 560 m above sea level, near the stream, whereas plot L

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(548 m in elevation) is located near the summit of the hillslope. Continuous sampling of

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throughfall and soil water was performed weekly at both plots from 2001 to 2013 (Note

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that in 2007, only throughfall was sampled and analyzed). Stream water was also

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collected by fixed volume (500 mL) weekly (2001-2004, 2009-2013) at a dam (The dam

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was moved from downstream of a pool to upstream in 2008), where a flow-measuring

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flume was installed to record the stream water flux.

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Every four weeks, the samples from all throughfall collectors, each lysimeter, and

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stream were pooled into a bulk sample (monthly sample with volume proportion)

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respectively for chemical analysis. All water samples were filtered (before filtration a

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sub-sample was used for pH measurement) and analysed for major cations (Ca2+, Mg2+,

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K+, Na+, and NH4+) and anions (NO3−, SO42−, Cl−, and F−) by ion chromatography.

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Monomeric (mainly labile) aluminium (Ala) fraction of soil water and surface water was

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determined by measuring the amount of 8-hydroxy-quinoline aluminium complex that

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formed within 20 s.31 The organic fraction (Alo) of Ala was determined by passing the

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sample through a pH-adjusted cation exchange column before 8-hydroxy-quinoline

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complex formation. Inorganic monomeric Al (Ali) was calculated as the difference

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between Ala and Alo.32 Al speciation, especially Al3+ concentration, was calculated by

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the ALCHEMI program.33


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Data analysis. The flux of each ion was calculated from ion concentration and water

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flux in each compartment. Since the water flux in soil was not directly measured in this

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study, it was estimated on the basis of the Na+ balance assumption (The ratio of soil

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water flux to throughfall water flux is assumed to equal the ratio of Na+ concentration in

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throughfall to that in soil water). 24 It should be noted that Cl- is generally assumed to be

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conservative, and thus can be used as a tracer to estimate soil water flux.34 However,

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long-term monitoring in this catchment showed significant sink of Cl- in soil, as

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indicated by similar or even lower Cl- concentration in soil water and stream water than

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in throughfall water (Fig. S1 in Supplementary Information). In comparison with Cl-,

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Na+ seemed more conservative in this catchment (Fig. S2). Considering the existence of

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potential sources/sinks of Na+ in soil such as mineral weathering and cation exchange,

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even extremely low in this catchment,24,35 the estimation of soil water flux based on the

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real ratios of Na+ concentrations between throughfall and soil water might cause

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uncertainty. Therefore, a fixed ratio of soil water flux to throughfall water flux was

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assumed, 0.79 for S1 and 0.43 for S2, according to a fertilizing experiment by NaNO3

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addition in the same catchment,24 where Na+ was more conservative than the untreated

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soil because soil Na+ weathering or exchange could be ignored due to huge amount of

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Na+ added.

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The acid neutralizing capacity (ANC) was calculated as the sum of base cations (Ca2+

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+ Mg2+ + K+ + Na+ + NH4+, in equivalent) minus the sum of acid anions (SO42- + NO3- +

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Cl- + F-, in equivalent) based on the charge balance approach.36 The molar ratio of base



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cation to Al3+, Bc/Al, was calculated as the sum of base cation (Ca2+ + Mg2+ + K+, in

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molar concentration) divided by the molar concentration of Al3+. In addition to H+

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deposition, N transformations may contribute to soil acidification (H+N):37

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H+N = (NH4+i –NH4+o) + (NO3-o – NO3-i)

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where NH4+i and NO3-i were throughfall fluxes of NH4+ and NO3- respectively, and

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NH4+o and NO3-o were the leaching fluxed of NH4+ and NO3- in soil water.

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Statistically significant temporal changes in the chemical composition of throughfall

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and soil water was analyzed by the flow-adjusted, seasonal Kendall test (SKT) and the

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Sen slope.38-40 The analysis was done for two periods, 2001-2006 and 2008-2013,

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respect to the SO2 emission control since 2006. Monthly data of pH, SO42-, NO3-, Ca2+,

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and Al3+ concentration were used for analysis. In addition, post hoc analysis using

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one-way ANOVA was applied to analyze differences in water chemistry in different

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periods (2001-2006 and 2008-2013). Difference were considered significant at the p ≤

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0.05 level.

167 168

RESULTS

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Throughfall. The monthly concentration of major ions, such as SO42-, NO3-, NH4+,

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Ca2+, and Al3+ in throughfall water, as well as the water amount and pH of throughfall,

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show clear seasonal variation (Fig. 3). The pH (annual average ranging from 3.35 to

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4.53) was high in summer but low in winter, which coincided with the seasonal variation

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of precipitation (shown by throughfall water amount). Throughfall pH was only greater


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than 5.6 during a few months in summer. The concentrations of SO42-, NO3-, and Ca2+

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were all higher in winter than in summer. In general, lower concentrations of SO42- and

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NO3- due to the diluting effect of abundant rain in summer might lead to higher pH. The

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NH4+ concentration showed not only peaks in winter, but also some peaks in summer,

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which reflected high NH3 emission from agricultural field in the growing season.

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The pH of throughfall showed a significant trend of decrease during 2001-2006

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(negative Sen slope by SKT; at a rate of -0.0953 yr-1), but a significant increase during

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2008-2013 (at a rate of 0.1792 yr-1) (Table 1). Several ions had significant

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increasing/decreasing trends in concentration values. For example, until 2006, the SO42-

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concentration in throughfall increased significantly (at 97.25 µeq⋅L-1⋅yr-1). However,

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after 2008 the concentration of SO42- decreased (at a rate of -110.5 µeq⋅L-1⋅yr-1). In

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contrast, the NO3- concentration showed an increasing trend during the whole period (at

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12.59 µeq⋅L-1⋅yr-1); This rate became even higher in recent years. The concentration of

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NH4+ and Ca2+ also showed increasing trends during the whole period. The faster rate of

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SO42- concentration decreasing than the increasing rate of NO3- concentration after 2008

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(while no significant trends in NH4+ and Ca2+) indicated decreasing trend of throughfall

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water acidity, coincided with the increasing trend of pH.

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Soil water and surface water. Temporal variations of monthly concentration of

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SO42- and Ca2+ in soil water and stream water coincided with those in throughfall (with

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some delay), but with smaller amplitude (especially in S2 and stream water) (Fig. 3).

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Due to nitrification or uptake, little NH4+ occurred in soil water (especially in S2). As a



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result of NO3- formation by nitrification, more frequent variation of NO3- concentration

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was found in S1 than in throughfall, according to the peaks of DIN (NH4+ + NO3-) in

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throughfall. The pH of soil water showed opposite trend to throughfall, i.e., lower in

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summer but higher in winter. The stream water chemistry was even more stable than that

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of soil water in the bottom layer (S2).

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Unlike throughfall, soil water and stream water showed a decreasing trend of pH in

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the whole period (at -0.0171, -0.0369 and -0.0877 yr-1 in S1, S2 and SW respectively)

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(Table 1). The annual average pH of stream water was in the range of 3.96-4.85, which

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was greater than that of soil water (S1, 3.39-4.03；S2, 3.59-4.30). Unlike the decreasing

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trend of SO42- concentration in throughfall after 2008, the SO42- concentration of soil

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water and stream water showed no significant trend during the same period, even

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increasing trend during the whole study period (at 23.53 and 40.64 µeq⋅L-1⋅yr-1 in S2 and

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SW respectively). The concentration of NO3- and NH4+ in soil water and stream water

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showed also an increasing trend, similar as in throughfall. The increase in Ca2+

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concentration in stream water might neutralize the increase in SO42- concentration (i.e.,

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the changing rates of Ca2+ and SO22- were similar), while decreasing pH coincided with

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increasing NO3-. It should be noted that more variable stream water chemistry during

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2009-2013 occurred than during 2001-2004 period (Fig. 3), because the dam for stream

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water monitoring was moved from the downstream to upstream of a small pond.

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Accompanying the decrease in pH, the Al3+ concentration in soil water showed an

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increasing trend during the whole period (at 20.80 and 32.26 µeq⋅L-1⋅yr-1 in S1 and S2


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respectively) (Table 1). However, there was significant decline in the saturation of the

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soil solution with respect to Al(OH)3, as can be seen from the significant decline in log

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Q (defined as log Al + 3 pH)41 (Fig. 3).

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Fluxes. Yearly fluxes of S, N, Ca + Mg, and H+ in throughfall, soil water, and stream

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water showed also variation during 2001-2013 (Fig.4). The fluxes of S and Ca + Mg

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showed generally a decreasing trend along the hydrological continuums (i.e., from

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throughfall to soil water, then to stream water). In contrast, the fluxes of total inorganic

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nitrogen (TIN) in top soil water (S1) were larger than that in throughfall. The average H+

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depostion based on throughfall was significantly lower during 2008-2013 than during

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2001-2006 (one-way ANOVA; Table S1), which coincided with the significantly higher

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average pH of throughfall during 2008-2013 than during 2001-2006. However, the

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average H+ flux in soil water (S2) was significantly higher during 2008-2013 than

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during 2001-2006, while there is no significant difference in H+ flux in stream water.

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There were no significant differences in fluxes of SO42-, NO3- and Ca2+ between the two

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periods in stream water, with only exception of NO3- depostion, which was significantly

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higher during 2008-2013 than during 2001-2006 (one-way ANOVA; Table S1).

232 233

DISCUSSION

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Change of acid deposition. The significant declining trend of S depositon after 2008

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found in this study is among the first to document a successful SO2 emission abatement

236

in China. During the monitoring period, the annual S deposition (based on throughfall)



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increased to maximum values in 2004 and 2005 (>14 keq·ha-1·yr-1; Fig. 4). After 2008,

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the S deposition showed a decreasing trend (significant negative Sen slope; Table 1),

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reaching 7.7 keq·ha-1·yr-1 in 2013, which is smaller than in 2001. This temporal pattern

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in throughfall deposition of S, with its decline after 2008, coincides with a decrease in

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SO2 emission in Chongqing province (Fig. 2). In contrast, the annual N deposition

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continued to increase, reaching 5.3 keq·ha-1·yr-1 in 2010 - 2012, which is twice as much

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as in 2001 (2.5 keq·ha-1·yr-1) (Fig. 4). This trend also coincides with the fast increasing

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NOx emission before 2011 (Fig. 2), which resulted in larger increase in NO3- than NH4+

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in N deposition, and decrease in NH4+/NO3- ratio (Fig. 4). The net effect of the two N

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species on throughfall was thus decrease in pH. The effect of decreasing NOx emission

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after 2011 has not been seen yet.

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Although S and N deposition at the site were large compared to values in Europe and

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North America in 1980s,9 the precipitation was not exceptionally acidic, with annual

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average pH of throughfall of 3.79, which is within the range of values reported for

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typically acidified regions in North America and Europe (3.5~4.5).1~3,42 The annual

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average pH of deposition would have been around 3.0 without the high deposition of

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Ca2+ (and other base cations), which reflected the importance of the alkaline dust,

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mainly from anthropogenic sources such as cement production and iron steel industry in

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China.23 The annual Ca2+ deposition showed an increasing trend (due to increasing

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anthropogenic emission of Ca-content dust),43 reaching a value as high as 8.2

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keq·ha-1·yr-1. The annual Mg2+ deposition fluxes (see Fig. S3 for monthly variation in


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concentration) amounted to about 23% of Ca2+ deposition (Fig. 4), and also contributed

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greatly to ANC. As a result of decreasing S deposition and increasing Ca deposition

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since 2008, the H+ deposition decreased after reaching maximum values of about 4

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keq·ha-1·yr-1 (Fig. 4).

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S and N budet. The overall budget (i.e., the difference between throughfall

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deposition flux and stream water output) indicated considerable retention of S, N and Ca

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in the catchment (Table 2). The S output with soil water in the top layer (S1) was lower

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than input by deposition, especially in the earlier years. This highlights a significant

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SO42- sink in the upper soil. In a recent study we found that SO42- adsorption in the acid

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soils was the most likely mechanism removing SO42- from soil water.24 The sulfate

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adsorbed pool was 16.3-28.5 keq ha-1 in the upper 30 cm soil layer, and there might be

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larger adsorption capacity in deeper soil layer.16 In addition, incorporation to organic

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matter might be another important sulfate rentention mechanism, especially in the topper

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soil with higer organic carbon content. The organic S pool was estimated to be 11 keq

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ha-1 according to the ratio of organic S to organic C

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soil.24 Since the S deposition was averaged to 11.5 keq ha-1 yr-1 during 2001-2006 (Table

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S1), this SO42- sink in top soil (3.33 keq ha-1 yr-1 in Table2) amounted to about 29% of

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the S deposition. As SO42- adsorption is a process of ligand exchange with OH-,44 it may

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account for an increase in the ANC with about 3 keq ha-1 yr-1, which is larger than the

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annual H+ depostion (2.60 keq ha-1 yr-1). Adsorption and desorption may also be

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important to control short-term fluctuations of the SO42- flux, such as the observed

48

and the organic C content in



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year-to-year variation. When S deposition decreased after 2008, the S sink in S1

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decreased to 1.30 keq ha-1 yr-1, which accounted for only 12% of the S deposition

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(averagely 11.0 keq ha-1 yr-1) during 2008-2013 (Table S1). In addition, much more S

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retention occurred below the upper 30 cm soil layer (Table 2). Together with SO42-

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adsorption in the deeper soil, SO42- reduction to sulfides in the wet soils of the

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groundwater discharge zone (GDZ) may contribute significantly to S sink in the

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catchment, which need be tested in the future.

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In soil water, the flux of NH4+-N was very low, while that of NO3--N was large and

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even exceeded the sum of the input of NH4+ and NO3- in throughfall (Fig. 4). This

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indicates that most NH4+ derived from throughfall deposition and mineralization of soil

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organic matter was nitrified. The large leaching of NO3--N from the top soil layer (S1)

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indicates large N internal cycle, as illustrated by the rapid mineralization of annual

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litterfall.45

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decreased due to immobilization, nutrient uptake, and denitrification16. As a result, the

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upper 30cm soil layer showed a slight N sink (averaged 1.38 keq ha-1 yr-1; according to

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Table 2), which indicated little less N flux in soil solution (S2) than in throughfall

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deposition. In addition, there were larger variations of N fluxes from year to year in the

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soil solution; the fluxes were larger in the wet years, which can be attributed to

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mineralized N from litterfall and soil organic matter at high moisture conditions (driven

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by rainfall). It may be concluded that N saturation46 is extreme in hillslope soils at

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Tieshanping with all N deposition leached. This is also shown in the studies of soil N

From the top layer (S1) to the lower layer (S2), the N flux in soil water


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status,45 N mass balances,24 natural 15N abundances47,48 and 15N tracer49 in the catchment.

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Similar to S, a high N retention occurs in the groundwater discharge zone, that is, after

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percolation water has left the upper 30 cm of soil but before the water leaves the

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catchment as stream water at the weir (Table 2). Denitrification, especially in GDZ, may

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be the key reason of N loss.16 Since loss of nitrate by denitrification provides ANC,

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stream water was much less acidified than soil water, by nitrogen deposition.

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Like S, Ca2+ sink mainly occurred in the upper 30 cm of the soil (Table 2). In China,

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Ca2+ deposition is a major source of Ca2+ in soil water.16,23 The soil Ca2+ sink was too

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large to be explained by plant uptake, since the Ca2+ uptake by forest trees at the site was

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estimated to only 0.25-1.10 keq ha-1 yr-1.50 A possible mechanism might be the building

310

up of cation exchange capacity (CEC),16 possibly due to the increasing content of soil

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organic matter caused by N and Ca deposition.27,51

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Acidification and recovery. The concentration of SO42- and NO3- in soil water in the

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catchment was very high, even one order higher than that seen in Europe in 1980s and

314

1990s.1~4 Although the concentration of base cations (especially Ca2+) was also very

315

high, it was not high enough to compensate for the negative charge associated with

316

SO42- and NO3-. Thus, the pH of soil water and stream water was relatively low.

317

The pH and the concentration (and deposition) of SO42- in throughfall showed

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acidification of rainwater during 2001-2006, and recovery (not only alkalinisation

319

because of significant decrease in S deposition) after 2008. However, pH of soil water

320

did not show a significant increase in recent years, even with significant increase in pH



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of stream water (Table 1 and Table S1). The delayed recovery of forest soil and surface

322

water from acidification was also widely found in Europe and North America, and

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attributed to the release of previously stored SO42- from soil, coupled to the leaching of

324

Al3+ and base cations.3,4,52~54 The large adsorption of SO42- by soil on oxide surfaces

325

buffered soil acidity thus limiting the concentration of H+ in soil water, especially in the

326

earlier years (2001-2006). This occurred even with much NH4+ nitrified to produce

327

acidity. In recent years, accompanying the decrease in SO42- concentration in soil water

328

due to the decrease in S deposition, SO42- adsorption largely decreased and the

329

previously adsorbed sulfate outflowed from the upper 30 cm soil layer, which led to an

330

increase in ANC sink and thus a H+ source in this layer (Table 2). However, the total

331

SO42- sink remained unchanged because of the increase in SO42- adsorption in deeper

332

soil layer (Table 2).

333

Associated with the increasing N deposition, N transformations played an

334

increasingly important role in the acidification of soils and waters. The average acid

335

input by N transformation was 3.03 keq ha-1 yr-1, exceeding the H+ input directly by

336

atmospheric deposition (1.91 keq ha-1 yr-1) (Fig. 5). The exceedance was larger than the

337

previous findings.25 Since the S flux in soil water and stream water did not show

338

significantly increasing trend in recent years, the decrease in pH in stream water (Table

339

1) might be caused by increase in NO3- leaching. Actually, since NOx emission

340

abatement only began in 2011, its effect on acidification was not expected to see.


Page 16 of 37

Page 17 of 37


341

The Al3+ concentration of soil water did not change with the decrease of pH,

342

providing further evidence that gibbsite equilibrium does not present a good model to

343

describe the Al solubility.24,55,56 There was significant decline in log Q (defined as log

344

Al + 3 pH), while Al3+ concentration increasing (Table 1). The results may support the

345

earlier hypothesis that in strongly acidified forest soils complexion by solid phase

346

organics controls the solubility of Al, even in mineral soil layers, relatively low in

347

organic C.41,57

348

The Bc/Al in soil water is commonly used as a risk indicator for forest decline due to

349

acidification,58 and 1.0 is widely used as a critical limit in critical load assessments in

350

Europe.59 In this study, the molar ratios of Bc/Al in soil water of S2 were averaged to

351

2.42 at Tieshanping, much higher than the critical limit used. However, possible damage

352

to forest by acidification had occurred at this site due to high Al concentration. 32 This

353

implies that a higher critical limit should be applied in subtropical China.60 In addition,

354

no significant trend of increasing Bc/Al was found after 2008 than before 2006.

355 356

POLICY IMPLICATION

357

The long-term monitoring on throughfall, soil water and stream water during 2001 to

358

2013 in Tieshanping catchment, representing catchments sensitive to acidification in

359

southwest China, showed clear effect of emission abatement on recovery of acidified

360

soil and stream water. Coincided with the decrease in SO2 emission in Chongqing

361

province after 2006, a significantly decreasing trend of S deposition was detected during



362

2008-2013. In contrast, N deposition continued increasing in accordance with the

363

increase in NOx (until 2011) and NH3 emission. As a net effect, acid deposition had

364

been reduced, and the acidification of soil water and surface water had leveled off. The

365

trends of deposition in Tieshanping from 2001-2013 could represent the regional or even

366

national trends. It should be noted that high SO42- adsorption in the acid soil, together

367

very high deposition of base cations, especially Ca2+, provided large ANC to the

368

ecosystem. Those might be the main reasons that the soil water acidification in China

369

was not very serious in comparison with that in Europe and North America, even with

370

much higher S and N deposition. However, the recovery of acidified soil water and

371

surface water was delayed, due probably to desorption of previously stored sulfate

372

(SO42-) and enhanced leaching of nitrate (NO3-) from soil. Since soil acidification is still

373

very serious, shown by quite low pH in soil water, future emission abatement of both

374

SO2 and NOx should be strengthened.

375 376

ASSOCIATED CONTENT

377

Supporting Information

378

Additional information on the water chemistry during the periods and the calculation of

379

soil water flux are presented in a supplementary document.

380

AUTHOR INFORMATION

381

Corresponding Author

382

* Phone: +86 10 6278 3758; fax: +86 10 6277 3597; e-mail: [email protected]

383

Notes


Page 18 of 37

Page 19 of 37


384

The authors declare no competing financial interest.

385 386

ACKNOWLEDGEMENT

387

The authors are grateful for the financial support of the National Natural Science

388

Foundation of China (21221004), the State Environmental Protection Public Welfare

389

Project of China (201209001) and the Research Council of Norway (209696/E10) to

390

carry out this study.

391 392

REFERENCE

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552

Page 28 of 37

Tables

553 554

Table 1 Sen slopes of throughfall (CTF), soil water (S1 and S2), and stream water (SW)

555

concentrationa Component

Period

pH

SO42-

NO3-

NH4+

Ca2+

Al3+

logQ

CTF

2001~2013

0.0073**

n.s.

12.59**

6.232*

16.69*

-

-

2001~2006

-0.0953**

97.25**

9.756**

n.s.

n.s.

-

-

2008~2013

0.1792**

-110.5**

12.78**

n.s.

n.s.

-

-

2001~2013

-0.0171**

n.s.

30.94**

2.681**

n.s.

20.80**

-0.0798**

2001~2006

-0.0073**

n.s

29.70*

n.s

n.s

43.80**

n.s.

2008~2013

0.0042*

n.s

n.s

n.s

n.s

n.s.

n.s.

2001~2013

-0.0369**

23.53**

26.18**

n.s

n.s.

32.26**

-0.0963**

2001~2006

-0.0263**

44.70**

35.54**

n.s.

n.s

32.57**

n.s.

2008~2013

n.s.

n.s.

n.s.

n.s.

n.s.

n.s.

n.s.

2001~2013

-0.0877**

40.64**

9.452**

2.580**

40.82**

-

-

2001~2006

-

-

-

-

-

-

-

2008~2013

-0.0713*

n.s.

n.s.

n.s.

n.s.

-

-

S1

S2

SW

556

a

Analyzed by seasonal Kendall test (SKT), with change of concentration in µeq⋅L-1⋅yr-1, and change of pH in yr-1；

557

* Trend significant at P

Monitoring Effect of SO2 Emission Abatement on Recovery of Acidified

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