Potential N2O emissions from leguminous tree plantation soils in the humid tropics



[1] We compared nitrous oxide (N2O) emissions over 1 year from soils of plantations growing acacia, which is a leguminous plant capable of symbiotic nitrogen fixation in root nodules, and secondary forests in Sumatra, Indonesia. N2O emissions from acacia plantation soils fluctuated seasonally, from high in the wetter season to low in the drier season, whereas N2O emissions from secondary forest soils were low throughout the year. Water-filled-pore-space data showed that denitrification contributed substantially to N2O emissions from soils at acacia sites. The average annual N2O flux in acacia plantations was 2.56 kg N ha−1 a−1, which was eight times higher than that from secondary forest soils (0.33 kg N ha−1 a−1). In secondary forests, NH4+ was the dominant form of inorganic nitrogen. However, in acacia plantations, the NH4+: NO3 ratio was relatively lower than that in secondary forests. These results suggest that secondary forests were nitrogen limited, but acacia plantations were less nitrogen limited. Leguminous tree plantations may increase nitrogen cycling, resulting in greater N2O emissions from the soil. However, on a global warming potential basis, N2O emissions from acacia plantation soils accounted for less than 10% of the carbon uptake by plants. Nevertheless, because of the spread of leguminous tree plantations in Asia, the importance of N2O emissions from leguminous tree stands will increase in the coming decades.

1. Introduction

[2] Nitrous oxide (N2O) is a strong greenhouse gas, with a global warming potential (GWP) 296 times higher than that of CO2. The atmospheric concentration of N2O has increased rapidly in recent decades, with N2O emission rates from natural ecosystems amounting to 4.6–15.9 Tg N a−1, which is equal to the amount from human activities [Prather et al., 2001]. Because the increment in atmospheric N2O represents the difference between N2O sources and sinks [Kroeze et al., 1999], annual emission rates from each source and sink are constrained by the increment in atmospheric N2O. However, for each source, estimates of N2O source emissions are unreliable. For example, N2O emissions from wet tropical forests are the largest natural ecosystem source of N2O, but the total estimate of 3.0 Tg N a−1 has a range of 2.2 to 3.7 Tg N a−1 [Prather et al., 2001]. Emission rate estimates from other sources also range widely, indicating the need for more data to provide better estimates.

[3] N2O is emitted from soils by nitrification and denitrification processes [Davidson et al., 2000]. The nitrification rate is a primary factor regulating N2O emissions, as it both directly controls N2O emissions from nitrification and the rate of nitrate supply for the denitrification process. Therefore N2O flux is correlated with the nitrification rate in many types of forest ecosystems [Ishizuka et al., 2002, 2005a]. The soil water condition, including water-filled pore spaces (WFPS), is another important factor regulating N2O emission from soils [Davidson et al., 2000]. Theoretically, 60% WFPS is optimal for N2O emissions; at <60% WFPS, nitric oxide is dominant, whereas at >60% WFPS, nitrogen gas is dominant. Because of this trade-off between water content and the ratio between emission of N2O and other nitrogenous gases, N2O emissions usually exhibit a complex response to water content fluctuations.

[4] Excess nitrogen input to the soil could promote nitrogenous gas emissions by raising nitrogen cycling in soils. For example, nitrogen fertilizer input stimulates nitrification, resulting in increased N2O emission rates [Garcia-Montiel and Binkley, 1998; Davidson et al., 2000; Erickson et al., 2001]. Erickson et al. [2001] found a high correlation between the amount of nitrogen input from leaf litter and annual nitrogen oxide (N2O + NO) fluxes among six forests of different ages. These results support the hole-in-the-pipe model [Davidson et al., 2000], which posits that an increase in nitrogen availability will result in an increase in nitrogenous gas emissions. Residues of leguminous plant species have a high nitrogen concentration because of symbiotic nitrogen fixation by nitrogen-fixing bacteria (Rhizobium) in legume root nodules. Thus plant residues of leguminous species with high nitrogen concentrations would increase nitrogen availability in soils, resulting in increased N2O emission rates. For example, N2O flux at a site where wheat was cultivated with a leguminous crop was greater than that at a wheat monoculture site [Nishimura et al., 2005; Wagner-Riddle et al., 1997]. Because large amounts of nitrogen are continuously supplied to soils through plant litterfall, leguminous plantations raise nitrogen availability in the soil, making these plantations a possible N2O source [Intergovernmental Panel on Climate Change (IPCC), 2003].

[5] Acacia is among the most important leguminous trees for industrial plantations because of its fast growth and tolerance of acidic and nutrient-poor environments. By 2000, 8,317,000 ha had been planted with acacia worldwide, and this area has likely increased since then [Food and Agriculture Organization of the United Nations (FAO), 2001]. Of the total global acacia plantation area, 96% is located in Asia [FAO, 2001]. To evaluate the impact of leguminous plantations on N2O emissions, we measured soil surface flux at acacia plantation sites and secondary forests for 1 year in Sumatra, Indonesia.

2. Materials and Methods

2.1. Site Description

[6] We carried out field measurements in the humid tropical forest at Muara Enim, South Sumatra Province, Indonesia (3°30′–4°05′S, 103°50′–104°10′E; Figure 1). The land history of this area is as follows [Hardjono et al., 2005]: In South Sumatra Province, large-scale logging began in the late 1960s. In the mid-1970s, a large land rehabilitation program was initiated, but with little success. In 1990, an industrial and community forest plantation program was started to provide woody raw material for pulp and wood industries. Acacia mangium and Albizia falcataria were the main species planted. Apart from the plantations, wetlands were used to grow paddy rice, and uplands were used for farming, estate crops, and community forests. Estate crops, managed by companies or communities, consisted mainly of rubber, oil palm, and coffee.

Figure 1.

Sampling site locations.

[7] In Muara Enim district, one tree plantation company manages 96,840 ha of acacia (A.mangium) plantations supplying the pulp industry, as well as a riparian buffer zone, conservation forest, multispecies plantation, and local species plantation. The plantations were established beginning in 1990 by the conversion of alang-alang (Imperata cylindrica) grassland, scrubland and secondary forest [Siregar et al., 1999; Hardjono et al., 2005]. Small patches of secondary forest, resulting from slash-and-burn agriculture, are distributed in and around the area, which consists mostly of South Sumatra Province flat to undulating terrain. In Subanjeriji and Benakat, the annual average rainfall and temperature from 1991 to 2002 were 2750 mm and 27.3°C [Hardjono et al., 2005]. Although there is no marked dry season, rainfall is lower from May to September. During our experimental period, rainfall was highest in April 2004 (504 mm) and lowest in September 2004 (45.4 mm; Figure 2). The soils are Acrisols [International Society of Soil Science (ISSS) Working Group RB, 1998] with a parent material of Tertiary sedimentary rocks.

Figure 2.

Monthly rainfall during the 1-year study period.

[8] Three 7-year-old acacia forests (AM1, 2, and 3) and three secondary forests (SF1, 2, and 3) were randomly selected for flux measurements. All of the acacia forests were first-rotation plantations, possibly established by conversion of alang-alang grasslands. When the plantations were established, the land was plowed, harrowed, and sprayed with herbicides. Trees were planted in 1997, spaced 2 × 4 m apart. Before planting, 85 g per tree of phosphate fertilizer (SP-36, PT Petrokimia Gresik) were added to each planting hole. No fertilizer or herbicide was applied subsequently. In 2003, when the trees were 7.7 years old, the sum of above- and below-ground biomasses, mean height, mean diameter at breast height (DBH) and basal area were 136.4 Mg ha−1, 23 m, 20.7 cm, and 25.5 m2 ha−1 in AM1; 154.2 Mg ha−1, 23 m, 20.8 cm, and 22.3 m2 ha−1 in AM2; and 131.9 Mg ha−1, 18 m, 16.5 cm, and 21.3 m2 ha−1 in AM3, respectively (Kaneko et al., personal communication, 2007). The soil texture at AM1 and AM2 was clayey, which is the dominant regional soil texture, whereas that at AM3 was clay loam. The three secondary forests were aggrading and well-stocked. According to the land owners, two of the secondary forests (SF1 and SF2) had regenerated after slash-and-burn activity in the 1970s, while the other (SF3) was established in the 1980s. The dominant tree species were Pithecellobium jiringa at SF1 and Schima wallichii at SF2 and SF3. The average growth in tree size at SF3 was less than that at SF1 and SF2. Estimated aboveground biomass derived from an allometric equation, maximum tree height, mean DBH and basal area were 248.4 Mg ha−1, 55 m, 11.6 cm, and 24.1 m2 ha−1 in SF1; 246.7 Mg ha−1, 64 m, 13.0 cm, and 23.7 m2 ha−1 in SF2; and 116.5 Mg ha−1, 24 m, 11.6 cm, and 14.3 m2 ha−1 in SF3, respectively [M. Hara, personal communication, 2006]. The soil texture at SF1 and SF2 was clayey; at SF3 it was sandy clay loam.

2.2. Flux Measurements

[9] We used a static chamber method to measure N2O flux [Ishizuka et al., 2002] and measured the flux every month from October 2003 to October 2004. Six replicate chambers made of 0.32-m diameter PVC tubes 0.15 m long were inserted into the soil 1 d before sampling, spaced 7–8 m apart. The chamber lid had sampling ports and air bags to equilibrate the inside pressure to atmospheric pressure. After sealing the chambers with lids, we used syringes to take 40-mL gas samples at 0, 10, 20, and 40 min. We injected the gas samples into 30-mL glass vacuum vials with butyl rubber stoppers that had been evacuated beforehand in the laboratory. The temperatures inside the chamber and in the soil at 5 cm depth were measured using a temperature sensor (SK-1250MCIII, Sato Keiryoki Mfg. Co. Ltd., Japan). N2O concentrations were analyzed in the laboratory using a gas chromatograph (GC-14B, Shimadzu, Kyoto, Japan) with an electron capture detector. We calculated fluxes using a linear regression slope, using the data at 0, 20, and 40 min elapsed time.

2.3. Soil Sampling and Analysis

[10] Soil samples for chemical and biochemical analysis were taken from two depths, 0–5 and 5–10 cm, in January, April, July, and October 2004. In each plot, a composite soil sample was obtained by mixing six 300-mL samples taken approximately 2 m from each chamber using three 100-mL cylinders (5.1 cm diameter, 5 cm tall). The soils were passed through a 2-mm mesh sieve and stored at 4°C. Soil pH (H2O) was measured using a glass electrode (HM-30G, Toa DDK, Tokyo, Japan) and a well-mixed solution of 10 g soil and 25 mL deionized water. Total carbon and nitrogen contents were determined using an NC analyzer (NC-900, Sumitomo Chemical Co., Osaka, Japan) for air-dried samples. Microbial carbon and nitrogen biomass was determined using the chloroform fumigation extraction method [Vance et al., 1987] with a total organic carbon (TOC) analyzer equipped with a total nitrogen analyzing attachment (Shimadzu TOC-V CSH and TNM-1). Inorganic ammonium (NH4+-N) and nitrate (NO3N) were extracted by shaking a mixture of 5 g fresh soil and 50 mL 2M KCl for 1 h within 3 d of sampling. After shaking, the solution was filtered and refrigerated. Ammonium and nitrate concentrations in the extracted solution were determined using a flow-injection analyzer (Auto Analyzer 3, Bran + Luebbe, Hamburg, Germany). The net nitrogen mineralization rate and net nitrification rate of the soils at AM2 and SF3 were determined according to the 15N dilution method [Sørensen and Jensen, 1991; Wessel and Tietema, 1992; Stark and Hart, 1996]. We added 5 μg N of ammonium sulfate to 5 g of fresh soil. Three of six replicate samples were immediately extracted using 50 mL 2M KCl solution following the extraction method described above, and the other three samples were incubated for 24 h at 25°C. The ammonium and nitrate concentrations in the extract were determined using the flow-injection analyzer. Bulk density and soil texture were determined using samples collected in January 2004. Soil texture was determined by the pipette method [Gee and Bauder, 1986] with a composite sample of six replicates. We also took soil samples every month at depths of 0–5 and 5–10 cm to measure soil water content gravimetrically. WFPS was calculated as follows:

equation image

where ω (kg kg−1) is soil water content, Ws (Mg m−3) is bulk density, and Vp (m3) is total pore volume in 1 m3 of soil.

2.4. Data Analysis

[11] We calculated the minimum significant flux (α = 0.05 [Hutchinson and Livingston, 1993; Verchot et al., 1999; Ishizuka et al., 2005b]) from verification of each chamber flux. The minimum flux considered significantly different from zero was 37 μg N m−2 d−1. We defined fluxes below this minimum as trace. The trace values were replaced with one-half the minimum flux (18.7 μg N m−2 d−1) for further calculation [Gilbert, 1987].

[12] All statistical analyses were performed using JMP 6.0.3 (SAS, Cary, NC). The significance of differences was tested by a parametric t-test (Student's t-test) or a post-hoc Tukey HSD test, followed by analysis of variance (ANOVA).

[13] Because the WFPS calculated for AM3 in February 2004 far exceeded 100% (135%), we excluded this data point from further WFPS analysis.

3. Results

3.1. Soil Properties

[14] Table 1 shows general soil properties. The soil texture differed among sites; the proportion of sand was larger and the proportion of clay was lower at AM3 and SF3 than at the other sites (Table 1). The bulk density at AM3 and SF3 was greater than that at the other sites. The mean total nitrogen content at 0–5 cm soil depth in acacia and secondary forests did not differ (0.109 kg m−2 and 0.106 kg m−2, respectively). The mean nitrate content at 0–5 cm soil depth at acacia sites (0.55 g N m−2, Table 2) was significantly greater than that at secondary forest sites (0.12 g N m−2; p < 0.05, Student's t-test). Other soil properties at 0–5 cm depth did not differ significantly between acacia and secondary forests (p > 0.05, Student's t-test). Many soil properties differed significantly between sites with coarser soil texture (AM3 and SF3) and those with finer texture (AM1, AM2, SF1, and SF2). The mean water content, total nitrogen, and clay content at 0–5 cm soil depth at the coarser texture sites were significantly lower than at the finer texture sites, whereas the mean C:N ratio, ammonium content, and bulk density at the coarser texture sites were significantly greater than at the finer texture sites (p < 0.05, Student's t-test). Microbial biomass N at AM3 was significantly lower than that at the other plots (p < 0.05, Tukey HSD).

Table 1. General Soil Properties at 0–5 cm and 5–10 cm Depth
PlotDepth, cmpH (H2O)aTotal Ca kg m−2Total Na kg m−2CN ratioa1SEBulk Density Mg m−31SESoil Textureb
Avr.1SEAvr.1SEAvr.1SEAvr.Avr.Clay %Silt %Sand %
  • a

    Average of four measurements in Jan., Apr., Jul., and Oct. 04.

  • b

    Data represent one composite sample from six samples taken in Jan. 04 (N. Yamashita, personal communication, 2006).

Table 2. Seasonal Fluctuation of Soil Values for Soil Water Content, Microbial Biomass Carbon and Nitrogen, and Inorganic Nitrogen at 0–5 cm Depth
PlotMonthWater Contenta kg kg−1Microbial BiomassbNH4+–Nc g N m−2NO3–Nc g N m−2
C g C m−2N g N m−2
  • a

    Values represent the mean and standard error of six measurements comprising one composite sample.

  • b

    Values represent the mean and standard error of three measurements comprising one composite sample.

  • c

    Values represent the mean of two measurements comprising one composite sample.


[15] WFPS increased from October to March and decreased from May to August (Figure 3). This trend corresponded with the rainfall trend (Figure 2); precipitation was higher between October and April. WFPS at AM3 was higher than at the other sites due to lower total pore space.

Figure 3.

Seasonal fluctuations in water-filled pore space (WFPS) in each forest. The error bar indicates 1 SD of six replications.

[16] The net nitrification rate at AM2 was greater than that at SF3 in each month (Table 3), but the annual mean was not significantly different (p > 0.05). The nitrogen mineralization rate was low in July and high in October at both sites (Table 3).

Table 3. Nitrogen Mineralization and Nitrification Rate
  Net Mineralization mg N kg−1 d−1Net Nitrification mg N kg−1 d−1
Average7.50 0.99 
Average12.76 0.12 

3.2. Gas Fluxes

[17] N2O flux from acacia forest soils showed a pronounced seasonal fluctuation, being relatively high in the wetter season; this trend was not apparent in the secondary forests (Figure 4). The means and standard deviations of N2O fluxes from November 2003 to October 2004 were 0.48 ± 0.35 (AM1), 0.63 ± 0.54 (AM2), 1.01 ± 0.78 (AM3), 0.10 ± 0.05 (SF1), 0.06 ± 0.04 (SF2), and 0.11 ± 0.07 (SF3) mg N m−2 d−1. Although the annual mean flux at AM3 was not significantly different from that at AM1 or AM2, the mean flux at AM3 during the high flux period between January 2004 and May 2004 was significantly higher than that at AM1 and AM2. By extrapolating these flux measurements to each month from November 2003 to October 2004, the annual N2O flux from the soils at AM1, AM2, AM3, SF1, SF2, and SF3 was estimated to be 1.75, 2.27, 3.67, 0.36, 0.22, and 0.40 kg N ha−1, respectively. The average N2O flux of acacia plantations and secondary forests was 2.56 kg N ha−1 (1 SD was 0.99 kg N ha−1) and 0.33 kg N ha−1 (1 SD was 0.09 kg N ha−1), respectively. N2O flux and WFPS in acacia forests were significantly correlated (R2 = 0.500, p < 0.01; Figure 5).

Figure 4.

Seasonal fluctuations in N2O flux in acacia and secondary forests. The error bar indicates 1 SD of six chambers.

Figure 5.

Relationship between water-filled pore space (WFPS) and N2O flux at six sites (n = 13 at each site).

[18] In acacia plantations, N2O fluxes correlated positively with NH4+ and negatively with NO3 content in the soil (Figure 6).

Figure 6.

Relationship between N2O flux and inorganic nitrogen (NH4+ and NO3). The number on the upper right of the symbol indicates the soil sampling month at acacia sites.

4. Discussion

4.1. Conversion to Acacia Plantations Might Boost N2O Flux From Soils

[19] The conversion to acacia plantations in this region could increase N2O emissions from soils. Annual N2O emissions from acacia plantation soils were 2.56 kg N ha−1, eight times greater than those from secondary forest soils, indicating that conversion of secondary forests into acacia plantations will enhance soil N2O emissions. Because we did not measure the N2O flux in a grassland control, we cannot directly address the effect of grassland conversion on N2O flux. On Sumatra Island, N2O emissions from grassland soils have not been found to exceed those of forest sites [Ishizuka et al., 2002, 2005a; Verchot et al., 2006]. Thus N2O flux from grasslands is likely lower than that from acacia plantations. Further studies are needed to clarify the effect of acacia plantations on soil N2O flux.

[20] Annual N2O emissions from secondary forest soils (0.33 kg N ha−1) are low relative to other tropical forests; e.g., tropical forest emissions are estimated at 0.01–7.68 kg N ha−1 [Breuer et al., 2000] and Amazon secondary forests at 0.94 kg N ha−1 [Verchot et al., 1999] or 0.80 kg N ha−1 [Palm et al., 2002]. These low secondary forest emissions are comparable to those observed at primary forests in central Sumatra (0.13 kg N ha−1 and 0.39 kg N ha−1 [Ishizuka et al., 2002]), but relatively low compared to emissions from southern Sumatra forests (1.47 and 1.80 kg N ha−1 at sites showing high WFPS (90–100% [Verchot et al., 2006]) and montane forests in central Sulawesi (0.29, 1.01, and 1.11 kg N ha−1 [Purbopuspito et al., 2006]). Ishizuka et al. [2005a] suggested that even within a limited geographical region, N2O flux in soils having a udic moisture regime is higher than that in drier soils. Thus the relatively higher N2O flux observed by Verchot et al. [2006] may be at least partly site specific and reflect exceptionally wet soil moisture conditions. To sum up existing data, N2O flux from Indonesian forest soils ranges from 0.13 to 1.80 kg N ha−1.

[21] The mechanism boosting N2O emissions from acacia plantations may be an enhanced soil nitrogen cycling rate. The net nitrification rate and nitrate content at acacia plots were greater than those at secondary forest sites (Table 3). These rapid rates of nitrification and high nitrate supply for the denitrification process may promote soil N2O emissions in acacia plantations. The relationship between N2O flux and the ratio of NH4+ to NO3 (a nitrogen limitation index [Verchot et al., 2006]) clearly differed at acacia plantations and secondary forests (Figure 7): acacia plantations were distributed along the Y axis, whereas secondary forests lay along the X axis. Because excess nitrate should be immediately absorbed by plants and microorganisms in the nitrogen limited sites, the ratio of NH4+ to NO3 should be higher. Therefore the low N2O flux in secondary forests might be due to nitrogen limitation, and the high N2O flux in acacia plantations might be partly due to the mitigation of nitrogen limitation. In the drier season, the NO3 content in AM1 and AM2 was high (Table 2) and nitrogen limitation was less severe, possibly due to the lesser demands of plants and microorganisms for nitrate. The net nitrification rate (Table 3) shows that the nitrate supply continued in April at AM2, although soil nitrate content at that time was relatively low (Table 2). Thus the decrease in NO3 content in the wetter season at AM1 and AM2 (Figure 6) may be due to plant uptake, corresponding to high growth in the wetter season and/or the loss by denitrification.

Figure 7.

Relationship between N2O flux and the ratio of NH4+ to NO3.

[22] The significant difference in wetter-season N2O emissions among the acacia plantations (AM3 > AM1 and AM2) indicates that emissions may vary considerably, depending on soil type. Relatively low-clay-content Acrisols have less structure development and high bulk density because soil shrinkage in drier periods is weaker than that of more clayey Acrisols, and particles of sand, silt, and clay are packed closely [Ohta and Effendi, 1992]. Thus in relatively low-clay-content Acrisols, soil porosity declines and drainage is consequently constrained. This may be why bulk density was higher at AM3, regardless of its higher sand content relative to the other acacia soils. Because of this soil physical property, WFPS at AM3 was high and provided ideal conditions for denitrification. This difference in pore space might also affect the proportion of gaseous species; for example, N2O is dominant at relatively high WFPS (70–90% [Davidson et al., 2000]). The low microbial biomass nitrogen and low nitrate content in AM3 soil (Table 2) indicates that nitrogen cycling at AM3 may be lower than that at the other acacia sites, but the proportion of nitrogen coming out as N2O might be larger at AM3 due to the high WFPS.

4.2. How Significant are N2O Emissions From Fast-Growing Leguminous Tree Plantations to Global Warming?

[23] In terms of global warming potential (GWP), carbon sequestration by acacia trees is far greater than the loss by N2O emission from soils. According to our biomass estimations using an allometric equation formulated by felling trees and excavating their roots and expressed as a function of DBH, the sum of above- and below-ground biomass in 7.7-year-old acacia stands was 136.4 Mg ha−1 for AM1, 154.2 Mg ha−1 for AM2, and 131.9 Mg ha−1 for AM3, respectively, and the mean annual growth in AM1, AM2, and AM3 was 17.7, 20.0, and 17.1 Mg ha−1, respectively (Kaneko et al., personal communication, 2007). The annual carbon uptake at AM1, AM2, and AM3 corresponds to 32.5, 36.7, and 31.4 Mg CO2 ha−1 a−1, respectively. The N2O flux at AM1, AM2, and AM3 (1.75, 2.27, and 3.67 kg N ha−1 a−1, respectively) was equivalent to 0.81, 1.06, and 1.71 Mg CO2 ha−1 a−1, respectively, using 296 as the GWP for N2O in 100 years [Prather et al., 2001]. Thus on a GWP basis, N2O emissions reduced the carbon uptake by plants by 2.5% at AM1, 6.2% at AM2, and 11.7% at AM3. Because the coarser textured soil of AM3 is not common in this plantation area, the N2O emission reduction appears to be less than 10% of the carbon uptake by plants.

[24] To our knowledge, this study is the first to attempt to clarify the effect of leguminous plantations on atmospheric N2O. In future studies, we plan to consider the following issues. Among the most important is determining how tree stand age affects N2O emissions. This study investigated 7-year-old acacia stands. The growth rate of plantation acacias tends to decrease after 6 years [Yonekawa and Miyawaki, 1988], and nitrogen uptake by trees may decrease at sites older than 7 years. Thus nitrogen limitation is less pronounced in older than in younger acacia stands, indicating that N2O emission rates in younger stands may be lower. Nitrogen input through litterfall also may change with increasing age. We are now monitoring N2O emissions at acacia stands of various ages and rotation stages. We are also assessing the influence of various silvicultural practices, such as weeding, harvesting, and harvest-slash management, on N2O flux fluctuations in acacia plantations.

[25] Because soil nitrogen content and N2O emissions from soil are affected by previous land use [Erickson et al., 2002], the effect of continuous land use on leguminous tree plantations requires clarification. For example, acacia plantations in our study region are usually rotated at 6- to 8-year intervals, and some plantation areas are now in their second or third rotation. Continuous nitrogen input to the soil from leguminous tree leaf litter maintains high nitrogen cycling, and the nitrogen level in second or third rotations might be greater than that in the first rotation. Thus investigating how N2O emissions and nitrogen cycling from leguminous tree plantations change with each rotation, and during harvests between rotations, is another important issue.

[26] Other common plantation leguminous trees include Paraserianthes, Leucaena, and Sesbania species, and evaluating the impact of different tree species on N2O emission rates is also important.

[27] Another critical issue is determining how leguminous plantations stimulate N2O emissions in phosphorus-limited conditions, which are common in the Asian humid tropics. Soil N cycling rates are high in many tropical forest ecosystems, relative to higher latitude forests [Vitousek, 1984], in part due to the presence of leguminous trees and other N-fixing plants, resulting in high nitrogen inputs to forest soils [Vitousek, 1984; Brown and Lugo, 1990]. In the neotropics, where leguminous tree species dominate forests [Primack and Corlett, 2005], the ecosystem is considered to be limited by phosphorus rather than nitrogen [Vitousek, 1984]. Thus inorganic nitrogen in soils dominated by NO3 [Verchot et al., 1999; Reiners et al., 1994] contributes to high N2O emissions from natural forests [Keller and Reiners, 1994; Verchot et al., 1999]. However, in the Asian tropics, natural forests at lower elevations are dominated by non-N-fixing dipterocarp trees [Primack and Corlett, 2005], and secondary forest dominants are not usually legumes either, as in this study. We found NH4+ to be the primary inorganic nitrogen in secondary forest soils, as in other Sumatran forests [Verchot et al., 2006; Ishizuka et al., 2002]. This suggests that mature nonlegume forests in this area may be nitrogen limited, with low N2O flux in these nitrogen-limited systems dependent on the low NO3 production rate. Consequently, an increase in total nitrogen input to such soils, for example by planting leguminous trees, increases nitrogen cycling and results in higher N2O emissions. Because excess nitrogen may easily promote N2O emissions from phosphorus-limited soils [Hall and Matson, 1999], leguminous tree plantations at phosphorus-limited sites may result in higher soil N2O emissions. We have initiated experiments to clarify the impact of leguminous plantations at phosphorus-limited sites.

[28] Currently, N2O emissions from leguminous plantations do not contribute seriously to global warming. Assuming that acacia plantations were converted from nonleguminous forests or grasslands and that the change in N2O emissions is identical to the difference between acacia plantations and secondary forest in this study, N2O emissions by acacia plantations account for 18.5 Gg N (=2.23 kg ha−1 a−1 × 8,317,000 ha), which corresponds to only 0.6% of N2O emissions from wet tropical forests (3 Tg N [Prather et al., 2001]). However, because of the spread of acacia and other leguminous tree plantations in Asia [FAO, 2001], the importance of N2O emissions from leguminous tree stands will increase in coming decades.

5. Conclusion

[29] We compared N2O emissions from acacia plantation stands and secondary forests for 1 year in Sumatra, Indonesia. Average annual N2O emissions from acacia plantation soils were 2.56 kg N ha−1 a−1, equivalent to less than 10% of carbon dioxide fixation by the acacias on a GWP basis. However, due to the increase in the area of acacia and other leguminous tree plantations in Asia, the importance of N2O emissions from leguminous tree stands will increase over the next several decades.


[30] This study was supported by the Japan Society for the Promotion of Science. We thank Maya Liony Lioe, Jiyana, Yudhi, Alkho, Helim, and Roji for help with field work at the MHP Company site. We also thank Takayuki Kaneko for providing biomass data, Masashi Hara for vegetation surveys at the research plots, and Mamoru Kanzaki for valuable suggestions.