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Soil Acidification in Response to Acid Deposition in Three Subtropical Forests of Subtropical China
Pedosphere 20(3): 399–408, 2010 ISSN 1002-0160/CN 32-1315/P c 2010 Soil Science Society of China Published by Elsevier Limited and Science Press
Soil Acidification in Response to Acid Deposition in Three Subtropical Forests of Subtropical China∗1 LIU Ke-Hui1,2 , FANG Yun-Ting3 , YU Fang-Ming4 , LIU Qiang5 , LI Fu-Rong1 and PENG Shao-Lin1,∗2 1 State
Key Laboratory of Biocontrol, School of Life Science, Sun Yat-Sen University, Guangzhou 510275 (China) of Applied Science and Technology, Guilin University of Electronic Technology, Guilin 541004 (China) 3 Dinghushan Forest Ecosystem Research Station, South China Botanical Garden, Chinese Academy of Sciences, Dinghu, Zhaoqing, 526070 (China) 4 College of Resource and Environment, Guangxi Normal University, Guilin 541004 (China) 5 Department of Biology, Hainan Normal University, Haikou 571158 (China) 2 College
(Received August 11, 2009; revised February 21, 2010)
ABSTRACT Long-term changes in soil pH, the current status of soil acidification, and the response of bulk soil and soil water pH to experimental nitrogen addition under three subtropical forests were investigated in Dinghushan Biosphere Reserve of subtropical China. The results showed that the mineral soil pH at 0–20 cm depth declined significantly from 4.60–4.75 in 1980s to 3.84–4.02 in 2005. Nitrogen addition resulted in the decrease of pH in both bulk soil and soil water collected at 20-cm depth. The rapid decline of soil pH was attributed to long-term high atmospheric acid deposition (nitrogen and sulphur) therein. The forest at earlier succession stage with originally higher soil pH appeared to be more vulnerable to acid deposition than that at later succession stage with originally low soil pH. Key Words:
forest succession, nitrogen addition, soil pH
Citation: Liu, K. H., Fang, Y. T., Yu, F. M., Liu, Q., Li, F. R. and Peng, S. L. 2010. Soil acidification in response to acid deposition in three subtropical forests of subtropical China. Pedosphere. 20(3): 399–408.
Soil acidification can be a natural process of proton-generating reactions, including the protolysis of carboxylic acid and natural organic acids and the uptake of cations. However, this process can be accelerated, particularly by deposition of sulphur (S) and nitrogen (N) compounds (Richter, 1986; Kazda and Zvacek, 1989). Soil acidification likely results in loss of base cations and mobilization of aluminium (Al) and other potentially toxic elements (Reuss et al., 1987; Aastrup et al., 1995; Das et al., 1997; Pichtel et al., 1997; Gilliam et al., 2005). Previous studies showed that forest soil acidification caused by acid deposition was one of major environmental issues in developed countries and was likely to become worse in some developing countries (Pan, 1992; Larssen et al., 1999; Tomlinson, 2003). In a German forest, soil pH has decreased by approximately 1.0 pH unit down to a depth of 2 m since 1950 (Grenzius, 1984). In southern Sweden, pH in topsoil has decreased considerably during the last 15–35 years, by about 1.0 pH unit in less acid soils and 0.5 in acid soils (Falkengren-Grerup, 1987). The decrease of soil pH occurred throughout the soil profiles and was accompanied by considerable losses of exchangeable of base cations, which decreased by about 50% of the pool in 1949/1954 (Falkengren-Grerup, 1987; Falkengren-Grerup et al., 1987). Similarly, the calcium pools in forests in Norway and the northern US have been reported to decline about 50% in the last 45 years (Likens et al., 1996). Soil pH has reached as low as 3.2–3.6 in areas of southern Germany (Z¨ottl et al., 1989). ∗1 Supported
by the National Natural Science Foundation of China (Nos. 30270282 and 40703030), the Key Project of the Chinese Ministry of Education (No. 704037), the Doctoral Scientific Research Foundation of Guilin University of Electronic Technology, China (No. Z20718), and the Guangxi Provincial Department of Education, China (No. 200707MS048). ∗2 Corresponding author. E-mail: [email protected].
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Southern China has become the third-largest producer of acid rain in the world, after the northeastern US and central Europe (Larssen et al., 1999). The total annual SO2 emissions in China increased constantly from 1982 to 1997 and slightly decreased in 1998–2004, but then increased and reached the highest point of 25.5 million tons in 2005 (MEPC, 1996–2006). The emissions of NOx have also increased rapidly in the past several years (MEPC, 1996–2006). Dai et al. (1998) reported that in some forest soils in southern China the soil pH declined by 0.1–1.0 pH units, cation exchange capacity (CEC) by 53%–76%, and base saturation by 30%–59% over 35 years, which might be the results of increased emissions of SO2 and NOx (Zhao and Seip, 1991; Pan, 1992; Dai et al., 1998). The Guangdong Province in subtropical China has been experiencing a long-term severe acid deposition, with low mean annual pH (4.13–5.58) and high frequency of acid rain (rainfall of pH < 5.6 being 33.5%–55.0%) (Environmental Protection Bureau of Guangdong Province, 1996–2006). In some areas of this province, high deposition of N and S has been frequently reported. In Dinghushan Biosphere Reserve located in central Guangdong Province, throughfall N input was measured at 21–50 kg N ha−1 year−1 during 1994–2008 (Huang et al., 1994; Fang et al., 2008), and throughfall S input was reported to be 20–40 kg S ha−1 year−1 in the 1990s (Seip et al., 1999). Rainfall pH at Dinghushan Biosphere Reserve ranged from 4.35 to 5.56 in 2000 (Liu et al., 2001) and from 4.08 to 5.22 in 2003–2004 (Fang et al., 2005). In this study, the long-term changes in soil pH in last two decades were presented from three representative forest types of Dinghushan Biosphere Reserve, comprising a forest succession sequence typical of subtropical China. Current status of soil acidification in these forests was reported as well as the responses of pH in bulk soil and soil water to experimental N addition. MATERIALS AND METHODS Study site The Dinghushan Biosphere Reserve (23◦ 09 21 –23◦ 11 30 N and 112◦ 30 39 –112◦ 33 41 E) is located in subtropical China and covers approximately 1 200 ha. This reserve is 20 km east to small Zhaoqing City (330 000 inhabitants) and about 90 km west to the metropolitan Guangzhou (10 million inhabitants). This reserve was established in 1956 by the Chinese government and the Forest Ecosystem Research Station was founded in 1978. One year later, the reserve was included into United Nations Educational Scientific and Cultural Organization’s Man and the Biosphere Network for the humid tropics. This site is ideal for studying the effects of air pollutants on forest ecosystems because typical forests therein represent a forest succession sequence prevailing in subtropical China (Peng, 1996) and have experienced heavy acid deposition over the past several decades (Huang et al., 1994; Seip et al., 1999; Fang et al., 2008). The reserve has a subtropical monsoon humid climate with a mean annual rainfall of 1 900 mm and mean annual evaporation of 1 115 mm (Kong et al., 1993). The rainfall is distributed unevenly over the year, with 75% in March–August and only 6% in December–February. The mean temperature is 21.4 ◦ C, with an average temperature of the coldest (January) and warmest (July) month of 12.6 and 28.0 ◦ C, respectively, and the mean relative humidity is 80.8% (Kong et al., 1993). In this study, three forests were chosen: a Pinus massoniana forest (PMF), a pine and broadleaf mixed forest (PBMF) and a monsoon evergreen broadleaf forest (MEBF), representing the early, medium and late succession stages, respectively. MEBF has been well protected for 400 years for religion reasons. It is a regional climax community with typical monsoon evergreen broadleaf species. The other two young forests were originated from the 1930’s clear-cut and subsequent pine plantation (Fang et al., 2006). The colonization from natural dispersal of regional broadleaf species has changed plant composition in the mixed forest, where the major species are P. massoniana, Schima superba, and Castanopsis chinensis, whereas the pine forest is dominated by P. massoniana under continuous human disturbances before 1990, primarily harvesting of understory and litter (Mo et al., 2003). The soils at the three study sites are lateritic red soil (Oxisol of U.S. Soil Taxonomy) formed from sandstone, with soil depths varying at each site (He et al., 1982; Mo et al., 2003). The soil is generally
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less than 30 cm to bedrock in the PMF, 30–60 cm in the PBMF, and > 60 cm in the MEBF (Mo et al., 2003). The MEBF soil has higher organic matter content, total N, exchangeable bases, Fe2 O3 and Al2 O3 , but lower bulk density than the soils of the two young forests, PMF and PBMF (Table I). TABLE I Physicochemical properties of the mineral soils (0–20 cm depth) in three forests with different succession stages, Pinus massoniana forest (PMF), pine and broadleaf mixed forest (PBMF) and monsoon evergreen broadleaf forest (MEBF), in the Dinghushan Biosphere Reserve Forest
Organic mattera)
PMF PBMF MEBF
27.3 34.5 53.5
Total Na)
Total exchangeable basesb)
0.9 1.0 1.9
18.39 14.04 19.50
g kg−1
a) Cited
from Mo et al. (2003);
b) Cited
Cation exchange capacityb)
mmol kg−1 84.91 86.88 100.55
from Liu (2003);
c) Cited
Fe2 O3 c)
Al2 O3 c)
mg per 100 g soil 346.8 461.7 438.3 320.5 744.6 590.7
Bulk densitya) g cm−3 1.41 1.30 1.21
from Zhang (1990).
Long-term change in soil pH and current status of soil acidification Mineral soil pH at 0–20 cm depth of the study forests has been monitored since 1980 (He et al., 1982; Zhang, 1990; Fu et al., 1995; Xia et al., 1997; Liu et al., 2001; Mo et al., 2003). In May 2005, 10–12 soil cores at 0–20 cm depth were sampled again in the permanent plots of each forest, after removing the organic layer, to measure pH and concentrations of exchangeable elements. In laboratory, soil samples were air-dried, crushed with a crabstick and passed through a 2 mm sieve prior to later chemical analysis. Soil pH in deionized water suspension was measured using a pH meter (pHS-25, Leici, Shanghai, China) at a ratio of 1:2.5 (soil:water), after shaking for 0.5 h and then equilibrating for 0.5 h. The detection limit was 0.01 pH. Sieved soil samples (5 g) were extracted with 50 mL 1 mol L−1 CH3 COONH4 , shaken for 2 h, and then filtered through Whatman filter paper No. 42 (Liu et al., 1996). The extract was used to determine exchangeable Ca, Mg, K, Na, Al, Fe, Mn, and Cd with inductively-coupled plasma emission spectroscopy. In May 2005, one sampling plot (5 m × 6 m) was selected in each forest and 30 sampling sites in each plot were selected with the space between each sampling site being 100 cm. After removing the organic layer, the mineral soil at each sampling site was taken with a soil auger (2.5 cm diameter, 120 cm long) and separated into three horizons (0–10, 10–20, and 20–40 cm). Soil pH was measured by the same method described above. Responses of pH in bulk soil and soil water to experimentally increased N input In the Dinghushan Biosphere Reserve, an N addition experiment was begun in 2003 in all three study forests to quantify the risks and consequences of increased nitrogen deposition in subtropical forest ecosystems (Fang et al., 2006). Four N treatments were established within the MEBF (control, low N, medium N, and high N) and three N treatments (control, low N, and medium N) were established in the PMF and PBMF. The total applications of nitrogen were 0, 50, 100, and 150 kg ha−1 year−1 for the control, low N, medium N, and high N treatments, respectively. Three replicate plots were randomly selected for each treatment in each forest. Each plot measured 10 m × 20 m, with buffer strips of about 10 m around each plot. The nitrogen in the form of NH4 NO3 solution was sprayed monthly on the forest floor. During each N application, the fertilizer NH4 NO3 was weighed, mixed with 20 L of water, and applied to the plots below the canopy. Two passes were made across each plot to ensure an even distribution of fertilizer. The control plots received the same amount of water with no additional N (Fang et al., 2006). Mineral soil was collected to a 10 cm depth five times over the first 3 years of the experiment (April, August and November 2004, June 2005, and August 2006); pH was determined to evaluate the effects
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of increased N deposition. For each sampling, 12–15 soil cores (2.5 cm in diameter) in each plot were collected randomly and combined into a composite sample. Soil pH was measured with the same method described above. Soil water was collected at the 20 cm depth (below the major rooting zone) from all plots except in one of the medium N plots in the MEBF due to shallow soils (Fang et al., 2006). In each plot, two zero-tension tray lysimeters (755.4 cm2 per tray) were installed in April/May 2003 and soil water was sampled twice a month (one prior to the N addition and the other 15 days after the N addition) in 2004–2006. Soil water from the two lysimeters within each plot was combined on the date of collection, and pH in each soil water sample was measured with the same pH meter above. Statistical analysis To quantify the change rate of soil pH over the past two decades, a linear regression was performed for each forest. Two-way analysis of variance (ANOVA) with Tukey’s HSD (honestly significant differences) test was performed to test the differences of soil pH between soil horizons and between forest types. Two-way ANOVA with sampling date and N treatments as main factors was performed to identify the effects of N addition on soil pH for each forest. For soil water pH, since samples were collected and analysed frequently, a repeated measures ANOVA was used to examine the overall N treatment effects for each forest and in each year, respectively. All statistical analyses were conducted using SPSS version 10.0 for Windows. Statistically significant differences were identified at P < 0.05. RESULTS Soil pH changes during the last two decades Soil pH in all three forests in the Dinghushan Biosphere Reserve declined rapidly over time (Fig. 1), significantly and linearly (r = 0.94–0.98, P < 0.001) from 4.60–4.75 in 1980s to 3.84–4.02 in 2005. However, the magnitude of the declines varied slightly with the stage of forest succession. Soil pH declined the fastest in the PMF, and the slowest in the MEBF (Fig. 1). The annual decrease rates were about 0.047, 0.045, and 0.044 pH units in the PMF, PBMF, and MEBF, respectively.
Fig. 1 pH changes with time in the 0–20 cm mineral soil in the Pinus massoniana forest (PMF), pine and broadleaf mixed forest (PBMF) and monsoon evergreen broadleaf forest (MEBF) of the Dinghushan Biosphere Reserve.
pH throughout soil profiles and soil exchangeable cations Soil pH values for each soil horizon and each forest measured in 2005 are shown in Table II. The pH throughout the soil profile averaged 4.00±0.10 in the PMF, 3.98±0.08 in the PBMF, and 3.86±0.08 in the MEBF. Soil pH increased with soil depth, with values being significantly higher in the deepest horizon (20–40 cm) than at the surface (0–10 cm) in the study forests (F = 12.028, P < 0.001; Table II). At the same soil horizons, soil pH in both the PMF and the PBMF was significantly higher than that in the MEBF (F = 12.164, P < 0.001; Table II).
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TABLE II Soil pH in each horizon of the Pinus massoniana forest (PMF), pine and broadleaf mixed forest (PBMF) and monsoon evergreen broadleaf forest (MEBF) of the Dinghushan Biosphere Reserve Forest
Soil depth
PMF PBMF MEBF
0–10 cm
10–20 cm
20–40 cm
3.97±0.02a) bAb)
4.02±0.10aAB 4.04±0.07abA 3.84±0.07abB
4.04±0.08aA 4.07±0.07aA 3.90±0.08aB
3.87±0.10bA 3.78±0.05bB
a) Means±standard
errors (n = 30). followed by the same lowercase and uppercase letters are not significantly different among forests within the same soil horizon and among soil horizons within one forest at P < 0.05 by Tukey’s honestly significant difference test. b) Means
In general, concentrations of CH3 COONH4 -exchangeable elements in mineral soil (0–20 cm) increased with forest succession (Table III). The mean soil concentrations of the exchangeable elements Ca, K, Mg, Al, Fe, Mn, and Cd in the MEBF were about 1.2–3.1 and 1.3–2.5 times higher than those in the PMF and PBMF. The concentrations of Na were similar in the three forests. Base saturation was 7.4%, 8.0%, and 12.2% in the PMF, PBMF, and MEBF, respectively. TABLE III Base saturation and concentrations of CH3 COONH4 -exchangeable elements in the 0–20 cm mineral soils of the Pinus massoniana forest (PMF), pine and broadleaf mixed forest (PBMF) and monsoon evergreen broadleaf forest (MEBF) of the Dinghushan Biosphere Reserve Forest PMF PBMF MEBF
Base saturation % 7.4ba) (1.1b) ) 8.0b (1.5) 12.2a (0.9)
K+ 33.2b (5.8) 33.8b (7.6) 84.2a (15.2)
Na+ 21.3a (3.2) 19.8a (4.9) 23.5a (6.6)
Ca2+ 55.0c (6.3) 92.9b (11.3) 146.6a (25.2)
Mg2+ 73.8b (15.6) 56.7b (11.3) 136.2a (18.3)
Al3+ mg kg−1 312.2b (11.3) 337.6b (10.2) 423.5a (15.3)
Fe3+
Mn2+
Cu2+
Cd2+
45.1c (5.6) 105.2b (11.6) 134.5a (10.3)
1.8ab (0.3) 1.2b (0.6) 2.2a (0.4)
1.0b (0.0) 1.0b (0.0) 1.5a (0.1)
35.0c (4.3) 49.4b (4.5) 109.3a (11.5)
a) Means
followed by the same letter within each column are not significantly different at P < 0.05 by Tukey’s honestly significant difference test (n = 10–12). b) Standard error.
Effect of N addition on pH values in bulk soil and soil water Soil pH generally declined after N addition in all three forests during the study period, but the values differed among forest types (Fig. 2). In the PMF, two-way ANOVA showed that the effect of low N treatment was significant (P = 0.002) on pH, and mean pH in the low N plots was 0.10 lower than that in the controls (Fig. 2). In the PBMF and MEBF, N additions resulted in slight decreases in soil pH (0.02–0.05), but it was not significant. Seasonal differences in soil pH values were significantly (P < 0.001) in all three forests, with the highest in November 2004 and the lowest in August 2006 (Fig. 2). The differences between seasonal maximum and minimum of pH values were 0.16, 0.18 and 0.21 in the PMF, PBMF, and MEBF, respectively. Repeated measures ANOVA on pH values in soil water for each forest and in each year showed a clear decline following N addition in all three forests, and the decrease was less pronounced in the third year especially in the MEBF (Fig. 3). In the PMF, N treatment effect on soil water pH was significant in the first year (P = 0.047) and marginally significant in the second and third years (P = 0.070 and 0.091, Fig. 3). In this forest, annual mean pH in soil water was lower in N addition plots than that in control plots by 0.08–0.19 pH units. In the PBMF, a marginally significant effect of N treatment on soil water pH values was only observed in the first year (P = 0.07). In the MEBF, significant N
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Fig. 2 Changes of pH in the 0–10 cm mineral soils of different N treatments in the Pinus massoniana forest (PMF), pine and broadleaf mixed forest (PBMF) and monsoon evergreen broadleaf forest (MEBF) of the Dinghushan Biosphere Reserve during the period of 2004–2006. Vertical bars indicate standard errors of the means (n = 3).
Fig. 3 Changes of soil water pH of different N treatments in the Pinus massoniana forest (PMF), pine and broadleaf mixed forest (PBMF) and monsoon evergreen broadleaf forest (MEBF) of the Dinghushan Biosphere Reserve during the period of 2004–2006. Vertical bars indicate standard errors of the means (n = 3).
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treatment effect on soil water pH was found in the first two years (P = 0.008 and 0.009, respectively), but disappeared in the third year (P = 0.400). In the PBMF and MEBF, the annual means of pH value in soil water were 0.01–0.15 pH units lower in N addition plots than in their corresponding controls. DISCUSSION Our results demonstrated that the forest soils were highly acidified in the Dinghushan Biosphere Reserve in subtropical China, with low exchangeable Ca (55.0–146.6 mg kg−1 ), low pH (3.8–4.0) and high exchangeable Al (312.2–423.5 mg kg−1 ) in the top 20 cm mineral soils. Mineral soil pH at 0–20 cm depth decreased with forest succession; pH in the MEBF (at the later successional stage) was 0.18 and 0.14 pH units lower than those in PBMF and PMF (at the earlier successional stage). These differences could be attributed to the different accumulation of bases in biomass associated with forest succession because they share the same parent soil materials and weathering conditions. Soil pH increased with soil depths, but all were lower than 4.1 (Table II), suggesting the study soils are well acidified throughout the profiles. In this study, soil pH declined rapidly from 1980s to 2005 (Fig. 1). Although soil pH during this period might be monitored in different seasons, seasonal variation in soil acidity has been shown to be minimal (Xia et al., 1997; Fig. 2). Furthermore, this seasonal variation is much smaller than the observed decline. Thus, we can conclude that the study forests were indeed experiencing an accelerated soil acidification. This decline was similar to that in other forests suffering from acid deposition in Guangxi and Hunan provinces in southern China (Dai et al., 1998) and in Mount Lushan in Jiangxi Province over 35 years (Pan, 1992). This decline was also within the range reported for 22 Swedish forests, where pH in the topsoil decreased by about 1.0 pH unit in less acid soils and by 0.5 in acid soils (Falkengren-Grerup, 1987). Although soil pH decreased, the concentrations of base cations were higher in this study than those in previous reports (He et al., 1982; Liu, 2003), suggesting that availability of soil base cations may initially increase with acid deposition. This observation was consistent with the results of Lu (2009), where the contents of exchangeable Ca2+ increased by 13%–52% in PMF and MEBF after 26 months of N application, and exchangeable Mg2+ were increased by 9%–23%. No increase in the soil exchangeable cations was observed in the MEBF, likely because available cations have been largely depleted in this old-growth forest due to long-term biomass accumulation and to great leaching companying with nitrate loss (Fang et al., 2008). At the same time, the concentrations of some heavy metals (Mn and Cu) were increased as compared with previous study (Liu, 2003). The same phenomena happened in the field N application (Lu et al., 2009). This result was also consistent with the results of experimental simulation in Hunan Province of South China, where acid deposition led to greater releases of heavy metals in the soil with pH < 4.2 (Liao et al., 2005). Several other studies in temperate and boreal forest ecosystems have also indicated that the levels of soil heavy elements are enhanced by acid deposition (Aastrup et al., 1995; Das et al., 1997; Pichtel et al., 1997; De Schrijver et al., 2006). In this study, we observed a significant and linear decline in soil pH during the last two decades. One may think what will happen to soil pH in the future. Will theses soils maintain constant declines in pH as observed over past two decades? Several processes occur in soils to buffer pH; these depend largely on soil pH (Ulrich, 1983, 1989). Generally, when soil pH is 5.0–4.2, the soils primarily release cations to neutralize acid inputs (cation exchange buffering). When soil pH is below 4.2, aluminium is mobilized (aluminium buffering); below 3.8, iron is mobilized (iron buffering) (Ulrich, 1983). The range of soil pH (3.7–5.2) in the study forests in the last two decades (Fig. 1) suggests that the main process by which the soil interacts with acid deposition is cation exchange buffering. Once mobilization of aluminium and iron takes place in soil, pH decline will be at slower rates than that in the cation exchange buffering stage. Thus, we predict that pH will probably become constant for a while, but finally decline rapidly. This prediction was in part supported by our observation in the N addition experiment. The absolute decrease in pH in bulk soil was quite small following the N addition, although the effect of N addition
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on soil pH was statistically significant (Fig. 2). Furthermore, the effect of N addition on soil water pH collected below the upper 20 cm soil layer was less pronounced in the third year. On the other hand, our results indicate that cation nutrients (Ca, Mg, and K) in the study forest soils will finally become depleted, becoming limiting factors to ecosystem productivity. Subsequently, the soil will release more toxic heavy metals if acid deposition keeps constant or increases further in the future. Soil acidification can be caused by both natural processes and acid deposition. Studies in unpolluted areas of Europe indicated that natural processes were the most important influences on soil acidification (Moreno Marcos and Gallardo Lancho, 2002). A study from one 30-year permanent plot indicated that 62% of soil acidification was attributed to natural process, and 38% to acid deposition (Markewitz et al., 1998). This was also found by Warfvinge et al. (1995) who studied soil acidification by mitigation strategies in northern forests in Sweden. In areas heavily polluted by acid deposition, soil acidification was mainly caused by anthropogenic factors, such as deposition of SO2 and NOx . The importance of acid deposition in the acidification of forest soils has been confirmed by Zhao and Seip (1991), Das et al. (1997), Augusto et al. (1998) and De Schrijver et al. (2006). In this study, the MEBF was over 400 years old, but the pH in this forest was only 0.14 and 0.12 pH units lower than that in the PMF and PBMF. However, during the last 20 years, soil pH values decreased by about 0.72 pH units in the MEBF, which was largely more than the result by natural succession. This result, therefore, suggests that the role of forest succession in soil acidification was minor in our study forests. Our observation was consistent with the study of Falkengren-Grerup et al. (1987) in Sweden, where biological acidification was almost negligible. Also, based on the net primary productivity and elemental composition, Duan et al. (2002) calculated N uptake rate and base cations for each vegetation type in China, and found that the direct vegetation uptake contributed very little to soil acidification in subtropical China. Results from Li et al. (2003) further illustrated that litter only from a few species can be expected to acidify soils. The Dinghushan Biosphere Reserve forests have been experiencing a high N deposition of 21–50 kg N ha−1 year−1 since 1990 (Huang et al., 1994; Fang et al., 2008), which is comparable to the highest deposition levels observed in Europe and the two sites in southern China, and higher than that in any forest of North America and Japan (Fang et al., 2008). This large N deposition must be contributing to the rapid soil acidification in the Dinghushan Biosphere Reserve. Reuss and Johnson (1986) demonstrated that enhanced N deposition could cause soil acidification via several pathways, with the overall change depending on the properties of the ecosystem, the form of N addition, and the anion or cation associated with the added N. Several studies in natural and agricultural systems of the temperate zone have shown that both atmospheric and fertilizer inputs of N can lead to acidification, depletion of base cations, and mobilization of potentially toxic aluminium ions and further soil acidification (Hallbacken and Tamm, 1986; Falkengren-Grerup, 1987; Lawrence et al., 1995; Likens et al., 1996; Bergholm et al., 2003). One field N addition experiment also proved that soils and soil water became further acidified during the first three years of treatments. Decreased exchangeable cations and increased Al3+ and Mn2+ also occurred after 26 months of N applications (Lu et al., 2009). At our study site, high S deposition (20–67 kg S ha−1 year−1 in throughfall) was also observed in the last two decades (Seip et al., 1999), which combined with N deposition can result in more acidic condition in soils than deposition of S or N alone (Sogn and Abrahamsen, 1998). Therefore, we conclude that the rapid acidification observed in the study forest soils could be caused by high acid deposition more than by natural processes. The results from this N addition experiment showed that the soils in all three forests could be further acidified by enhanced N deposition, and that it was most pronounced in forests in the early succession stage (the PMF) than in late stages (the PBMF and MEBF) (Fig. 3). This result supports the long-term observation of soil acidity during the past two decades with the fastest decline in the PMF (Fig. 1). Low base cation concentrations, CEC and organic matter pool in our forests, especially in the PMF, may lead to poor buffering capacity to acids, as Tao et al. (2002) reported. This result indicates that forest soils with originally higher pH values seem to be more sensitive to acid deposition than those
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with originally lower pH values. This speculation is in agreement with the report for Swedish forests (Falkengren-Grerup, 1987). CONCLUSIONS The forest soils at the study site were strongly acidified and soil pH values had declined rapidly. High acid deposition might be the major driving force to the rapid pH decline. Forests early in succession may be more susceptible to acid inputs than those late in succession in subtropical China. ACKNOWLEDGEMENTS We would like to thank the Dinghushan Forest Ecosystem Station, Chinese Academy of Sciences for providing the experimental condition. Thanks are also given to Mrs. XIAO Yan-Xiu, Mr. MO DingShen, and Mr. FANG Xiao-Ming at the Dinghushan Forest Ecosystem Station, Chinese Academy of Sciences for their kindly help in field and laboratory experiments, and to the two anonymous reviewers and Dr. F. S. GILLIAM at Marshall University, USA for their valuable comments in improving the manuscript. REFERENCES Aastrup, M., Iverfeldt, A., Bringmark, L., Kvarn¨ as, H., Thunholm, B. and Hultberg, H. 1995. Monitoring of heavy metals in protected forest catchments in Sweden. Water Air Soil Poll. 85: 755–760. Augusto, L., Bonnaud, P. and Ranger, J. 1998. Impact of tree species on forest soil acidification. Forest Ecol. Manag. 105(1–3): 67–78. Bergholm, J., Berggren, D. and Alavi, G. 2003. Soil acidification induced by ammonium sulphate addition in a Norway spruce forest in Southwest Sweden. Water Air Soil Poll. 148(1–4): 87–109. Dai, Z. H., Liu, Y. X., Wang, X. J. and Zhao, D. W. 1998. Changes in pH, CEC and exchangeable acidity of some forest soils in southern China during the last 32–35 years. Water Air Soil Poll. 108: 377–390. Das, P., Samantaray, S. and Rout, G. R. 1997. Studies on cadmium toxicity in plants: a review. Environ. Pollut. 98: 29–36. De Schrijver, A., Mertens, J., Geudens, G., Staelens, J., Campforts, E., Luyssaert, S., De Temmerman, L., De Keersmaeker, L., De Neve, S. and Verheyen, K. 2006. Acidification of forested podzols in North Belgium during the period 1950– 2000. Sci. Total Environ. 361(1–3): 189–195. Duan, L., Huang, Y. M., Hao, J. M. and Zhong, Z. P. 2002. Vegetation uptake of nitrogen and base cation in China and its role in soil acidification. Chin. J. Environ. Sci. (in Chinese). 23(3): 68–74. Environmental Protection Bureau of Guangdong Province. 1982–2006. Envirommental reports of Guangdong Province. Available online at http://www.gdepb.gov.cn/ (verified on May 15, 2006). Falkengren-Grerup, U. 1987. Long-term changes in pH of forest soils in southern Sweden. Environ. Pollut. 43(2): 79–90. Falkengren-Grerup, U., Linnermark, N. and Tyler, G. 1987. Changes in acidity and cation pools of south Swedish soils between 1949 and 1985. Chemoshphere. 16(10–12): 2239–2248. Fang, Y. T., Gundersen, P., Mo, J. M. and Zhu, W. X. 2008. Input and output of dissolved organic and inorganic nitrogen in subtropical forests of South China under high air pollution. Biogeosciences. 5: 339–352. Fang, Y. T., Mo, J. M., Jiang, Y. Q., Li, D. J. and Gundersen, P. 2005. Acidity and inorganic nitrogen concentrations in soil solution in short-term response to N addition in subtropical forests. J. Trop. Subtrop. Bot. (in Chinese). 13: 123–129. Fang, Y. T., Zhu, W. X., Mo, J. M., Zhou, G. Y. and Gundersen, P. 2006. Dynamics of soil inorganic nitrogen and their responses to nitrogen additions in three subtropical forests, South China. J. Environ. Sci. 18: 752–759. Fu, S. L., Yi, W. M. and Ding, M. M. 1995. Mineralization of soil microbial C, N, P, and K in different vegetations types at Dinghushan Biosphere Reserve. Acta Phytoecol. Sin. (in Chinese). 19: 217–224. Gilliam, F. S., Lyttle, N. L., Thomas, A. and Adams, M. B. 2005. Soil variability along a nitrogen mineralization and nitrification gradient in a nitrogen-saturated hardwood forest. Soil Sci. Soc. Am. J. 69: 247–256. Grenzius, R. 1984. Strong acidification of Berlin’s forest soils. Eur. J. Forest Res. 103: 131–139. Hallbacken, L. and Tamm, C. O. 1986. Changes in soil acidity from 1927 to 1982–1984 in a forest area of south-west Sweden. Scand. J. Forest Res. 1: 219–232. He, J. H., Chen, Z. Q. and Liang, Y. A. 1982. The soil of Dinghushan Biosphere Reserve. J. Trop. Subtrop. Forest Ecosyst (in Chinese). 1: 25–38. Huang, Z. L., Ding, M. M. and Zhang, Z. P. 1994. The hydrological processes and nitrogen dynamics in a monsoon evergreen broad-leafed forest of Dinghu Shan. Acta Phytoecol. Sin. (in Chinese). 18: 194–199.
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Effects of N and S deposition on leaching from an acid forest soil and growth of Scots pine (Pinus sylvestris L.) after 5 years of treatment. Forest Ecol. Manag. 103: 177–190. Tao, F., Hayash, Y. and Lin, E. 2002. Soil vulnerability and sensitivity to acid deposition in China. Water Air Soil Poll. 140: 247–260. Tomlinson, G. H. 2003. Acidic deposition, nutrient leaching and forest growth. Biogeochemistry. 65: 51–81. Ulrich, B. 1983. Soil acidity and its relation to acid deposition. In Ulrich, B. and Pankrath, J. (eds.) Effects of Accumulation of Air Pollutants in Forest Ecosystems. Reidel Publishing, Boston. pp. 127–146. Ulrich, B. 1989. Effects of acidic precipitation on forest ecosystem in Europe. In Adriano, D. C. and Johnson, A. H. (eds.) Acidic Precipitation. Spring-Verlay, Berlin. 2: 189–272. Warfvinge, P., L¨ ofgren, S. and Lundstrm, U. 1995. Implications of natural acidification for mitigation strategies in northern Sweden. Water Air Soil Poll. 85: 499–504. Xia, H. P., Yu, Q. 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Soil Acidification in Response to Acid Deposition in Three Subtropical Forests of Subtropical China∗1 LIU Ke-Hui1,2 , FANG Yun-Ting3 , YU Fang-Ming4 , LIU Qiang5 , LI Fu-Rong1 and PENG Shao-Lin1,∗2 1 State
Key Laboratory of Biocontrol, School of Life Science, Sun Yat-Sen University, Guangzhou 510275 (China) of Applied Science and Technology, Guilin University of Electronic Technology, Guilin 541004 (China) 3 Dinghushan Forest Ecosystem Research Station, South China Botanical Garden, Chinese Academy of Sciences, Dinghu, Zhaoqing, 526070 (China) 4 College of Resource and Environment, Guangxi Normal University, Guilin 541004 (China) 5 Department of Biology, Hainan Normal University, Haikou 571158 (China) 2 College
(Received August 11, 2009; revised February 21, 2010)
ABSTRACT Long-term changes in soil pH, the current status of soil acidification, and the response of bulk soil and soil water pH to experimental nitrogen addition under three subtropical forests were investigated in Dinghushan Biosphere Reserve of subtropical China. The results showed that the mineral soil pH at 0–20 cm depth declined significantly from 4.60–4.75 in 1980s to 3.84–4.02 in 2005. Nitrogen addition resulted in the decrease of pH in both bulk soil and soil water collected at 20-cm depth. The rapid decline of soil pH was attributed to long-term high atmospheric acid deposition (nitrogen and sulphur) therein. The forest at earlier succession stage with originally higher soil pH appeared to be more vulnerable to acid deposition than that at later succession stage with originally low soil pH. Key Words:
forest succession, nitrogen addition, soil pH
Citation: Liu, K. H., Fang, Y. T., Yu, F. M., Liu, Q., Li, F. R. and Peng, S. L. 2010. Soil acidification in response to acid deposition in three subtropical forests of subtropical China. Pedosphere. 20(3): 399–408.
Soil acidification can be a natural process of proton-generating reactions, including the protolysis of carboxylic acid and natural organic acids and the uptake of cations. However, this process can be accelerated, particularly by deposition of sulphur (S) and nitrogen (N) compounds (Richter, 1986; Kazda and Zvacek, 1989). Soil acidification likely results in loss of base cations and mobilization of aluminium (Al) and other potentially toxic elements (Reuss et al., 1987; Aastrup et al., 1995; Das et al., 1997; Pichtel et al., 1997; Gilliam et al., 2005). Previous studies showed that forest soil acidification caused by acid deposition was one of major environmental issues in developed countries and was likely to become worse in some developing countries (Pan, 1992; Larssen et al., 1999; Tomlinson, 2003). In a German forest, soil pH has decreased by approximately 1.0 pH unit down to a depth of 2 m since 1950 (Grenzius, 1984). In southern Sweden, pH in topsoil has decreased considerably during the last 15–35 years, by about 1.0 pH unit in less acid soils and 0.5 in acid soils (Falkengren-Grerup, 1987). The decrease of soil pH occurred throughout the soil profiles and was accompanied by considerable losses of exchangeable of base cations, which decreased by about 50% of the pool in 1949/1954 (Falkengren-Grerup, 1987; Falkengren-Grerup et al., 1987). Similarly, the calcium pools in forests in Norway and the northern US have been reported to decline about 50% in the last 45 years (Likens et al., 1996). Soil pH has reached as low as 3.2–3.6 in areas of southern Germany (Z¨ottl et al., 1989). ∗1 Supported
by the National Natural Science Foundation of China (Nos. 30270282 and 40703030), the Key Project of the Chinese Ministry of Education (No. 704037), the Doctoral Scientific Research Foundation of Guilin University of Electronic Technology, China (No. Z20718), and the Guangxi Provincial Department of Education, China (No. 200707MS048). ∗2 Corresponding author. E-mail: [email protected].
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Southern China has become the third-largest producer of acid rain in the world, after the northeastern US and central Europe (Larssen et al., 1999). The total annual SO2 emissions in China increased constantly from 1982 to 1997 and slightly decreased in 1998–2004, but then increased and reached the highest point of 25.5 million tons in 2005 (MEPC, 1996–2006). The emissions of NOx have also increased rapidly in the past several years (MEPC, 1996–2006). Dai et al. (1998) reported that in some forest soils in southern China the soil pH declined by 0.1–1.0 pH units, cation exchange capacity (CEC) by 53%–76%, and base saturation by 30%–59% over 35 years, which might be the results of increased emissions of SO2 and NOx (Zhao and Seip, 1991; Pan, 1992; Dai et al., 1998). The Guangdong Province in subtropical China has been experiencing a long-term severe acid deposition, with low mean annual pH (4.13–5.58) and high frequency of acid rain (rainfall of pH < 5.6 being 33.5%–55.0%) (Environmental Protection Bureau of Guangdong Province, 1996–2006). In some areas of this province, high deposition of N and S has been frequently reported. In Dinghushan Biosphere Reserve located in central Guangdong Province, throughfall N input was measured at 21–50 kg N ha−1 year−1 during 1994–2008 (Huang et al., 1994; Fang et al., 2008), and throughfall S input was reported to be 20–40 kg S ha−1 year−1 in the 1990s (Seip et al., 1999). Rainfall pH at Dinghushan Biosphere Reserve ranged from 4.35 to 5.56 in 2000 (Liu et al., 2001) and from 4.08 to 5.22 in 2003–2004 (Fang et al., 2005). In this study, the long-term changes in soil pH in last two decades were presented from three representative forest types of Dinghushan Biosphere Reserve, comprising a forest succession sequence typical of subtropical China. Current status of soil acidification in these forests was reported as well as the responses of pH in bulk soil and soil water to experimental N addition. MATERIALS AND METHODS Study site The Dinghushan Biosphere Reserve (23◦ 09 21 –23◦ 11 30 N and 112◦ 30 39 –112◦ 33 41 E) is located in subtropical China and covers approximately 1 200 ha. This reserve is 20 km east to small Zhaoqing City (330 000 inhabitants) and about 90 km west to the metropolitan Guangzhou (10 million inhabitants). This reserve was established in 1956 by the Chinese government and the Forest Ecosystem Research Station was founded in 1978. One year later, the reserve was included into United Nations Educational Scientific and Cultural Organization’s Man and the Biosphere Network for the humid tropics. This site is ideal for studying the effects of air pollutants on forest ecosystems because typical forests therein represent a forest succession sequence prevailing in subtropical China (Peng, 1996) and have experienced heavy acid deposition over the past several decades (Huang et al., 1994; Seip et al., 1999; Fang et al., 2008). The reserve has a subtropical monsoon humid climate with a mean annual rainfall of 1 900 mm and mean annual evaporation of 1 115 mm (Kong et al., 1993). The rainfall is distributed unevenly over the year, with 75% in March–August and only 6% in December–February. The mean temperature is 21.4 ◦ C, with an average temperature of the coldest (January) and warmest (July) month of 12.6 and 28.0 ◦ C, respectively, and the mean relative humidity is 80.8% (Kong et al., 1993). In this study, three forests were chosen: a Pinus massoniana forest (PMF), a pine and broadleaf mixed forest (PBMF) and a monsoon evergreen broadleaf forest (MEBF), representing the early, medium and late succession stages, respectively. MEBF has been well protected for 400 years for religion reasons. It is a regional climax community with typical monsoon evergreen broadleaf species. The other two young forests were originated from the 1930’s clear-cut and subsequent pine plantation (Fang et al., 2006). The colonization from natural dispersal of regional broadleaf species has changed plant composition in the mixed forest, where the major species are P. massoniana, Schima superba, and Castanopsis chinensis, whereas the pine forest is dominated by P. massoniana under continuous human disturbances before 1990, primarily harvesting of understory and litter (Mo et al., 2003). The soils at the three study sites are lateritic red soil (Oxisol of U.S. Soil Taxonomy) formed from sandstone, with soil depths varying at each site (He et al., 1982; Mo et al., 2003). The soil is generally
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less than 30 cm to bedrock in the PMF, 30–60 cm in the PBMF, and > 60 cm in the MEBF (Mo et al., 2003). The MEBF soil has higher organic matter content, total N, exchangeable bases, Fe2 O3 and Al2 O3 , but lower bulk density than the soils of the two young forests, PMF and PBMF (Table I). TABLE I Physicochemical properties of the mineral soils (0–20 cm depth) in three forests with different succession stages, Pinus massoniana forest (PMF), pine and broadleaf mixed forest (PBMF) and monsoon evergreen broadleaf forest (MEBF), in the Dinghushan Biosphere Reserve Forest
Organic mattera)
PMF PBMF MEBF
27.3 34.5 53.5
Total Na)
Total exchangeable basesb)
0.9 1.0 1.9
18.39 14.04 19.50
g kg−1
a) Cited
from Mo et al. (2003);
b) Cited
Cation exchange capacityb)
mmol kg−1 84.91 86.88 100.55
from Liu (2003);
c) Cited
Fe2 O3 c)
Al2 O3 c)
mg per 100 g soil 346.8 461.7 438.3 320.5 744.6 590.7
Bulk densitya) g cm−3 1.41 1.30 1.21
from Zhang (1990).
Long-term change in soil pH and current status of soil acidification Mineral soil pH at 0–20 cm depth of the study forests has been monitored since 1980 (He et al., 1982; Zhang, 1990; Fu et al., 1995; Xia et al., 1997; Liu et al., 2001; Mo et al., 2003). In May 2005, 10–12 soil cores at 0–20 cm depth were sampled again in the permanent plots of each forest, after removing the organic layer, to measure pH and concentrations of exchangeable elements. In laboratory, soil samples were air-dried, crushed with a crabstick and passed through a 2 mm sieve prior to later chemical analysis. Soil pH in deionized water suspension was measured using a pH meter (pHS-25, Leici, Shanghai, China) at a ratio of 1:2.5 (soil:water), after shaking for 0.5 h and then equilibrating for 0.5 h. The detection limit was 0.01 pH. Sieved soil samples (5 g) were extracted with 50 mL 1 mol L−1 CH3 COONH4 , shaken for 2 h, and then filtered through Whatman filter paper No. 42 (Liu et al., 1996). The extract was used to determine exchangeable Ca, Mg, K, Na, Al, Fe, Mn, and Cd with inductively-coupled plasma emission spectroscopy. In May 2005, one sampling plot (5 m × 6 m) was selected in each forest and 30 sampling sites in each plot were selected with the space between each sampling site being 100 cm. After removing the organic layer, the mineral soil at each sampling site was taken with a soil auger (2.5 cm diameter, 120 cm long) and separated into three horizons (0–10, 10–20, and 20–40 cm). Soil pH was measured by the same method described above. Responses of pH in bulk soil and soil water to experimentally increased N input In the Dinghushan Biosphere Reserve, an N addition experiment was begun in 2003 in all three study forests to quantify the risks and consequences of increased nitrogen deposition in subtropical forest ecosystems (Fang et al., 2006). Four N treatments were established within the MEBF (control, low N, medium N, and high N) and three N treatments (control, low N, and medium N) were established in the PMF and PBMF. The total applications of nitrogen were 0, 50, 100, and 150 kg ha−1 year−1 for the control, low N, medium N, and high N treatments, respectively. Three replicate plots were randomly selected for each treatment in each forest. Each plot measured 10 m × 20 m, with buffer strips of about 10 m around each plot. The nitrogen in the form of NH4 NO3 solution was sprayed monthly on the forest floor. During each N application, the fertilizer NH4 NO3 was weighed, mixed with 20 L of water, and applied to the plots below the canopy. Two passes were made across each plot to ensure an even distribution of fertilizer. The control plots received the same amount of water with no additional N (Fang et al., 2006). Mineral soil was collected to a 10 cm depth five times over the first 3 years of the experiment (April, August and November 2004, June 2005, and August 2006); pH was determined to evaluate the effects
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of increased N deposition. For each sampling, 12–15 soil cores (2.5 cm in diameter) in each plot were collected randomly and combined into a composite sample. Soil pH was measured with the same method described above. Soil water was collected at the 20 cm depth (below the major rooting zone) from all plots except in one of the medium N plots in the MEBF due to shallow soils (Fang et al., 2006). In each plot, two zero-tension tray lysimeters (755.4 cm2 per tray) were installed in April/May 2003 and soil water was sampled twice a month (one prior to the N addition and the other 15 days after the N addition) in 2004–2006. Soil water from the two lysimeters within each plot was combined on the date of collection, and pH in each soil water sample was measured with the same pH meter above. Statistical analysis To quantify the change rate of soil pH over the past two decades, a linear regression was performed for each forest. Two-way analysis of variance (ANOVA) with Tukey’s HSD (honestly significant differences) test was performed to test the differences of soil pH between soil horizons and between forest types. Two-way ANOVA with sampling date and N treatments as main factors was performed to identify the effects of N addition on soil pH for each forest. For soil water pH, since samples were collected and analysed frequently, a repeated measures ANOVA was used to examine the overall N treatment effects for each forest and in each year, respectively. All statistical analyses were conducted using SPSS version 10.0 for Windows. Statistically significant differences were identified at P < 0.05. RESULTS Soil pH changes during the last two decades Soil pH in all three forests in the Dinghushan Biosphere Reserve declined rapidly over time (Fig. 1), significantly and linearly (r = 0.94–0.98, P < 0.001) from 4.60–4.75 in 1980s to 3.84–4.02 in 2005. However, the magnitude of the declines varied slightly with the stage of forest succession. Soil pH declined the fastest in the PMF, and the slowest in the MEBF (Fig. 1). The annual decrease rates were about 0.047, 0.045, and 0.044 pH units in the PMF, PBMF, and MEBF, respectively.
Fig. 1 pH changes with time in the 0–20 cm mineral soil in the Pinus massoniana forest (PMF), pine and broadleaf mixed forest (PBMF) and monsoon evergreen broadleaf forest (MEBF) of the Dinghushan Biosphere Reserve.
pH throughout soil profiles and soil exchangeable cations Soil pH values for each soil horizon and each forest measured in 2005 are shown in Table II. The pH throughout the soil profile averaged 4.00±0.10 in the PMF, 3.98±0.08 in the PBMF, and 3.86±0.08 in the MEBF. Soil pH increased with soil depth, with values being significantly higher in the deepest horizon (20–40 cm) than at the surface (0–10 cm) in the study forests (F = 12.028, P < 0.001; Table II). At the same soil horizons, soil pH in both the PMF and the PBMF was significantly higher than that in the MEBF (F = 12.164, P < 0.001; Table II).
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TABLE II Soil pH in each horizon of the Pinus massoniana forest (PMF), pine and broadleaf mixed forest (PBMF) and monsoon evergreen broadleaf forest (MEBF) of the Dinghushan Biosphere Reserve Forest
Soil depth
PMF PBMF MEBF
0–10 cm
10–20 cm
20–40 cm
3.97±0.02a) bAb)
4.02±0.10aAB 4.04±0.07abA 3.84±0.07abB
4.04±0.08aA 4.07±0.07aA 3.90±0.08aB
3.87±0.10bA 3.78±0.05bB
a) Means±standard
errors (n = 30). followed by the same lowercase and uppercase letters are not significantly different among forests within the same soil horizon and among soil horizons within one forest at P < 0.05 by Tukey’s honestly significant difference test. b) Means
In general, concentrations of CH3 COONH4 -exchangeable elements in mineral soil (0–20 cm) increased with forest succession (Table III). The mean soil concentrations of the exchangeable elements Ca, K, Mg, Al, Fe, Mn, and Cd in the MEBF were about 1.2–3.1 and 1.3–2.5 times higher than those in the PMF and PBMF. The concentrations of Na were similar in the three forests. Base saturation was 7.4%, 8.0%, and 12.2% in the PMF, PBMF, and MEBF, respectively. TABLE III Base saturation and concentrations of CH3 COONH4 -exchangeable elements in the 0–20 cm mineral soils of the Pinus massoniana forest (PMF), pine and broadleaf mixed forest (PBMF) and monsoon evergreen broadleaf forest (MEBF) of the Dinghushan Biosphere Reserve Forest PMF PBMF MEBF
Base saturation % 7.4ba) (1.1b) ) 8.0b (1.5) 12.2a (0.9)
K+ 33.2b (5.8) 33.8b (7.6) 84.2a (15.2)
Na+ 21.3a (3.2) 19.8a (4.9) 23.5a (6.6)
Ca2+ 55.0c (6.3) 92.9b (11.3) 146.6a (25.2)
Mg2+ 73.8b (15.6) 56.7b (11.3) 136.2a (18.3)
Al3+ mg kg−1 312.2b (11.3) 337.6b (10.2) 423.5a (15.3)
Fe3+
Mn2+
Cu2+
Cd2+
45.1c (5.6) 105.2b (11.6) 134.5a (10.3)
1.8ab (0.3) 1.2b (0.6) 2.2a (0.4)
1.0b (0.0) 1.0b (0.0) 1.5a (0.1)
35.0c (4.3) 49.4b (4.5) 109.3a (11.5)
a) Means
followed by the same letter within each column are not significantly different at P < 0.05 by Tukey’s honestly significant difference test (n = 10–12). b) Standard error.
Effect of N addition on pH values in bulk soil and soil water Soil pH generally declined after N addition in all three forests during the study period, but the values differed among forest types (Fig. 2). In the PMF, two-way ANOVA showed that the effect of low N treatment was significant (P = 0.002) on pH, and mean pH in the low N plots was 0.10 lower than that in the controls (Fig. 2). In the PBMF and MEBF, N additions resulted in slight decreases in soil pH (0.02–0.05), but it was not significant. Seasonal differences in soil pH values were significantly (P < 0.001) in all three forests, with the highest in November 2004 and the lowest in August 2006 (Fig. 2). The differences between seasonal maximum and minimum of pH values were 0.16, 0.18 and 0.21 in the PMF, PBMF, and MEBF, respectively. Repeated measures ANOVA on pH values in soil water for each forest and in each year showed a clear decline following N addition in all three forests, and the decrease was less pronounced in the third year especially in the MEBF (Fig. 3). In the PMF, N treatment effect on soil water pH was significant in the first year (P = 0.047) and marginally significant in the second and third years (P = 0.070 and 0.091, Fig. 3). In this forest, annual mean pH in soil water was lower in N addition plots than that in control plots by 0.08–0.19 pH units. In the PBMF, a marginally significant effect of N treatment on soil water pH values was only observed in the first year (P = 0.07). In the MEBF, significant N
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Fig. 2 Changes of pH in the 0–10 cm mineral soils of different N treatments in the Pinus massoniana forest (PMF), pine and broadleaf mixed forest (PBMF) and monsoon evergreen broadleaf forest (MEBF) of the Dinghushan Biosphere Reserve during the period of 2004–2006. Vertical bars indicate standard errors of the means (n = 3).
Fig. 3 Changes of soil water pH of different N treatments in the Pinus massoniana forest (PMF), pine and broadleaf mixed forest (PBMF) and monsoon evergreen broadleaf forest (MEBF) of the Dinghushan Biosphere Reserve during the period of 2004–2006. Vertical bars indicate standard errors of the means (n = 3).
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treatment effect on soil water pH was found in the first two years (P = 0.008 and 0.009, respectively), but disappeared in the third year (P = 0.400). In the PBMF and MEBF, the annual means of pH value in soil water were 0.01–0.15 pH units lower in N addition plots than in their corresponding controls. DISCUSSION Our results demonstrated that the forest soils were highly acidified in the Dinghushan Biosphere Reserve in subtropical China, with low exchangeable Ca (55.0–146.6 mg kg−1 ), low pH (3.8–4.0) and high exchangeable Al (312.2–423.5 mg kg−1 ) in the top 20 cm mineral soils. Mineral soil pH at 0–20 cm depth decreased with forest succession; pH in the MEBF (at the later successional stage) was 0.18 and 0.14 pH units lower than those in PBMF and PMF (at the earlier successional stage). These differences could be attributed to the different accumulation of bases in biomass associated with forest succession because they share the same parent soil materials and weathering conditions. Soil pH increased with soil depths, but all were lower than 4.1 (Table II), suggesting the study soils are well acidified throughout the profiles. In this study, soil pH declined rapidly from 1980s to 2005 (Fig. 1). Although soil pH during this period might be monitored in different seasons, seasonal variation in soil acidity has been shown to be minimal (Xia et al., 1997; Fig. 2). Furthermore, this seasonal variation is much smaller than the observed decline. Thus, we can conclude that the study forests were indeed experiencing an accelerated soil acidification. This decline was similar to that in other forests suffering from acid deposition in Guangxi and Hunan provinces in southern China (Dai et al., 1998) and in Mount Lushan in Jiangxi Province over 35 years (Pan, 1992). This decline was also within the range reported for 22 Swedish forests, where pH in the topsoil decreased by about 1.0 pH unit in less acid soils and by 0.5 in acid soils (Falkengren-Grerup, 1987). Although soil pH decreased, the concentrations of base cations were higher in this study than those in previous reports (He et al., 1982; Liu, 2003), suggesting that availability of soil base cations may initially increase with acid deposition. This observation was consistent with the results of Lu (2009), where the contents of exchangeable Ca2+ increased by 13%–52% in PMF and MEBF after 26 months of N application, and exchangeable Mg2+ were increased by 9%–23%. No increase in the soil exchangeable cations was observed in the MEBF, likely because available cations have been largely depleted in this old-growth forest due to long-term biomass accumulation and to great leaching companying with nitrate loss (Fang et al., 2008). At the same time, the concentrations of some heavy metals (Mn and Cu) were increased as compared with previous study (Liu, 2003). The same phenomena happened in the field N application (Lu et al., 2009). This result was also consistent with the results of experimental simulation in Hunan Province of South China, where acid deposition led to greater releases of heavy metals in the soil with pH < 4.2 (Liao et al., 2005). Several other studies in temperate and boreal forest ecosystems have also indicated that the levels of soil heavy elements are enhanced by acid deposition (Aastrup et al., 1995; Das et al., 1997; Pichtel et al., 1997; De Schrijver et al., 2006). In this study, we observed a significant and linear decline in soil pH during the last two decades. One may think what will happen to soil pH in the future. Will theses soils maintain constant declines in pH as observed over past two decades? Several processes occur in soils to buffer pH; these depend largely on soil pH (Ulrich, 1983, 1989). Generally, when soil pH is 5.0–4.2, the soils primarily release cations to neutralize acid inputs (cation exchange buffering). When soil pH is below 4.2, aluminium is mobilized (aluminium buffering); below 3.8, iron is mobilized (iron buffering) (Ulrich, 1983). The range of soil pH (3.7–5.2) in the study forests in the last two decades (Fig. 1) suggests that the main process by which the soil interacts with acid deposition is cation exchange buffering. Once mobilization of aluminium and iron takes place in soil, pH decline will be at slower rates than that in the cation exchange buffering stage. Thus, we predict that pH will probably become constant for a while, but finally decline rapidly. This prediction was in part supported by our observation in the N addition experiment. The absolute decrease in pH in bulk soil was quite small following the N addition, although the effect of N addition
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on soil pH was statistically significant (Fig. 2). Furthermore, the effect of N addition on soil water pH collected below the upper 20 cm soil layer was less pronounced in the third year. On the other hand, our results indicate that cation nutrients (Ca, Mg, and K) in the study forest soils will finally become depleted, becoming limiting factors to ecosystem productivity. Subsequently, the soil will release more toxic heavy metals if acid deposition keeps constant or increases further in the future. Soil acidification can be caused by both natural processes and acid deposition. Studies in unpolluted areas of Europe indicated that natural processes were the most important influences on soil acidification (Moreno Marcos and Gallardo Lancho, 2002). A study from one 30-year permanent plot indicated that 62% of soil acidification was attributed to natural process, and 38% to acid deposition (Markewitz et al., 1998). This was also found by Warfvinge et al. (1995) who studied soil acidification by mitigation strategies in northern forests in Sweden. In areas heavily polluted by acid deposition, soil acidification was mainly caused by anthropogenic factors, such as deposition of SO2 and NOx . The importance of acid deposition in the acidification of forest soils has been confirmed by Zhao and Seip (1991), Das et al. (1997), Augusto et al. (1998) and De Schrijver et al. (2006). In this study, the MEBF was over 400 years old, but the pH in this forest was only 0.14 and 0.12 pH units lower than that in the PMF and PBMF. However, during the last 20 years, soil pH values decreased by about 0.72 pH units in the MEBF, which was largely more than the result by natural succession. This result, therefore, suggests that the role of forest succession in soil acidification was minor in our study forests. Our observation was consistent with the study of Falkengren-Grerup et al. (1987) in Sweden, where biological acidification was almost negligible. Also, based on the net primary productivity and elemental composition, Duan et al. (2002) calculated N uptake rate and base cations for each vegetation type in China, and found that the direct vegetation uptake contributed very little to soil acidification in subtropical China. Results from Li et al. (2003) further illustrated that litter only from a few species can be expected to acidify soils. The Dinghushan Biosphere Reserve forests have been experiencing a high N deposition of 21–50 kg N ha−1 year−1 since 1990 (Huang et al., 1994; Fang et al., 2008), which is comparable to the highest deposition levels observed in Europe and the two sites in southern China, and higher than that in any forest of North America and Japan (Fang et al., 2008). This large N deposition must be contributing to the rapid soil acidification in the Dinghushan Biosphere Reserve. Reuss and Johnson (1986) demonstrated that enhanced N deposition could cause soil acidification via several pathways, with the overall change depending on the properties of the ecosystem, the form of N addition, and the anion or cation associated with the added N. Several studies in natural and agricultural systems of the temperate zone have shown that both atmospheric and fertilizer inputs of N can lead to acidification, depletion of base cations, and mobilization of potentially toxic aluminium ions and further soil acidification (Hallbacken and Tamm, 1986; Falkengren-Grerup, 1987; Lawrence et al., 1995; Likens et al., 1996; Bergholm et al., 2003). One field N addition experiment also proved that soils and soil water became further acidified during the first three years of treatments. Decreased exchangeable cations and increased Al3+ and Mn2+ also occurred after 26 months of N applications (Lu et al., 2009). At our study site, high S deposition (20–67 kg S ha−1 year−1 in throughfall) was also observed in the last two decades (Seip et al., 1999), which combined with N deposition can result in more acidic condition in soils than deposition of S or N alone (Sogn and Abrahamsen, 1998). Therefore, we conclude that the rapid acidification observed in the study forest soils could be caused by high acid deposition more than by natural processes. The results from this N addition experiment showed that the soils in all three forests could be further acidified by enhanced N deposition, and that it was most pronounced in forests in the early succession stage (the PMF) than in late stages (the PBMF and MEBF) (Fig. 3). This result supports the long-term observation of soil acidity during the past two decades with the fastest decline in the PMF (Fig. 1). Low base cation concentrations, CEC and organic matter pool in our forests, especially in the PMF, may lead to poor buffering capacity to acids, as Tao et al. (2002) reported. This result indicates that forest soils with originally higher pH values seem to be more sensitive to acid deposition than those
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with originally lower pH values. This speculation is in agreement with the report for Swedish forests (Falkengren-Grerup, 1987). CONCLUSIONS The forest soils at the study site were strongly acidified and soil pH values had declined rapidly. High acid deposition might be the major driving force to the rapid pH decline. Forests early in succession may be more susceptible to acid inputs than those late in succession in subtropical China. ACKNOWLEDGEMENTS We would like to thank the Dinghushan Forest Ecosystem Station, Chinese Academy of Sciences for providing the experimental condition. Thanks are also given to Mrs. XIAO Yan-Xiu, Mr. MO DingShen, and Mr. FANG Xiao-Ming at the Dinghushan Forest Ecosystem Station, Chinese Academy of Sciences for their kindly help in field and laboratory experiments, and to the two anonymous reviewers and Dr. F. S. GILLIAM at Marshall University, USA for their valuable comments in improving the manuscript. REFERENCES Aastrup, M., Iverfeldt, A., Bringmark, L., Kvarn¨ as, H., Thunholm, B. and Hultberg, H. 1995. Monitoring of heavy metals in protected forest catchments in Sweden. Water Air Soil Poll. 85: 755–760. Augusto, L., Bonnaud, P. and Ranger, J. 1998. Impact of tree species on forest soil acidification. Forest Ecol. Manag. 105(1–3): 67–78. Bergholm, J., Berggren, D. and Alavi, G. 2003. Soil acidification induced by ammonium sulphate addition in a Norway spruce forest in Southwest Sweden. Water Air Soil Poll. 148(1–4): 87–109. Dai, Z. H., Liu, Y. X., Wang, X. J. and Zhao, D. W. 1998. Changes in pH, CEC and exchangeable acidity of some forest soils in southern China during the last 32–35 years. Water Air Soil Poll. 108: 377–390. Das, P., Samantaray, S. and Rout, G. R. 1997. Studies on cadmium toxicity in plants: a review. Environ. Pollut. 98: 29–36. De Schrijver, A., Mertens, J., Geudens, G., Staelens, J., Campforts, E., Luyssaert, S., De Temmerman, L., De Keersmaeker, L., De Neve, S. and Verheyen, K. 2006. Acidification of forested podzols in North Belgium during the period 1950– 2000. Sci. Total Environ. 361(1–3): 189–195. Duan, L., Huang, Y. M., Hao, J. M. and Zhong, Z. P. 2002. Vegetation uptake of nitrogen and base cation in China and its role in soil acidification. Chin. J. Environ. Sci. (in Chinese). 23(3): 68–74. Environmental Protection Bureau of Guangdong Province. 1982–2006. Envirommental reports of Guangdong Province. Available online at http://www.gdepb.gov.cn/ (verified on May 15, 2006). Falkengren-Grerup, U. 1987. Long-term changes in pH of forest soils in southern Sweden. Environ. Pollut. 43(2): 79–90. Falkengren-Grerup, U., Linnermark, N. and Tyler, G. 1987. Changes in acidity and cation pools of south Swedish soils between 1949 and 1985. Chemoshphere. 16(10–12): 2239–2248. Fang, Y. T., Gundersen, P., Mo, J. M. and Zhu, W. X. 2008. Input and output of dissolved organic and inorganic nitrogen in subtropical forests of South China under high air pollution. Biogeosciences. 5: 339–352. Fang, Y. T., Mo, J. M., Jiang, Y. Q., Li, D. J. and Gundersen, P. 2005. Acidity and inorganic nitrogen concentrations in soil solution in short-term response to N addition in subtropical forests. J. Trop. Subtrop. Bot. (in Chinese). 13: 123–129. Fang, Y. T., Zhu, W. X., Mo, J. M., Zhou, G. Y. and Gundersen, P. 2006. Dynamics of soil inorganic nitrogen and their responses to nitrogen additions in three subtropical forests, South China. J. Environ. Sci. 18: 752–759. Fu, S. L., Yi, W. M. and Ding, M. M. 1995. Mineralization of soil microbial C, N, P, and K in different vegetations types at Dinghushan Biosphere Reserve. Acta Phytoecol. Sin. (in Chinese). 19: 217–224. Gilliam, F. S., Lyttle, N. L., Thomas, A. and Adams, M. B. 2005. Soil variability along a nitrogen mineralization and nitrification gradient in a nitrogen-saturated hardwood forest. Soil Sci. Soc. Am. J. 69: 247–256. Grenzius, R. 1984. Strong acidification of Berlin’s forest soils. Eur. J. Forest Res. 103: 131–139. Hallbacken, L. and Tamm, C. O. 1986. Changes in soil acidity from 1927 to 1982–1984 in a forest area of south-west Sweden. Scand. J. Forest Res. 1: 219–232. He, J. H., Chen, Z. Q. and Liang, Y. A. 1982. The soil of Dinghushan Biosphere Reserve. J. Trop. Subtrop. Forest Ecosyst (in Chinese). 1: 25–38. Huang, Z. L., Ding, M. M. and Zhang, Z. P. 1994. The hydrological processes and nitrogen dynamics in a monsoon evergreen broad-leafed forest of Dinghu Shan. Acta Phytoecol. Sin. (in Chinese). 18: 194–199.
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