Effects of nitrogen-fixing and non-nitrogen-fixing tree species on soil properties and nitrogen transformation during forest restoration in southern China

Abstract

The role of different plantation tree species in soil nutrient cycling is of great importance for the restoration of degraded lands. The objective of the present study was to evaluate the potential of N-fixing and non-N-fixing tree species to recuperate degraded land in southern China. The soil properties and N transformations in six forest types (two N-fixing plantations, three non-N-fixing plantations and a secondary shrubland) established in 1984 were compared. The N-fixing forests had 40–50% higher soil organic matter and 20–50% higher total nitrogen concentration in the 0–5 cm soils than the non-N-fixing forests. Soil inorganic N was highest under the secondary shrubland. The N-fixing Acacia auriculiformis plantation had the highest soil available P. There were no significant differences in soil N mineralization and nitrification among the forest types, but seasonal variation in these N processes was highly significant. In the rainy season, the rates of N mineralization (7.41–11.3 kg N ha⁻¹ month⁻¹) were similar to values found in regional climax forests, indicating that soil N availability has been well recovered in these forest types. These results suggest that N-fixing species, particularly Acacia mangium, are more efficient in re-establishing the C and N cycling processes in degraded land in southern China. Moreover, the N-fixing species A. auriculiformis performed better than other species in improving soil P availability.

Key words:

• afforestation
• nitrogen mineralization
• nitrogen-fixing species
• soil chemical properties
• South China

Introduction

Subtropical and tropical forests are among the most productive terrestrial ecosystems on earth. However, millions of hectares of subtropical and tropical forests are being deforested or degraded as a result of human activities (Lamb et al. 2005). Thus, restoration at both regional and global scales is critical for the sustainability of the earth’s ecosystem (Gardiner et al. 2003).

The roles that different tree species (e.g. N-fixing versus non-N-fixing species, coniferous versus deciduous) play in restoring soil processes have been extensively studied (Alvarez-Aquino et al. 2004; Franco and De Faria 1997; Lamb 1998; Macedo et al. 2008; Malcolm et al. 2008; Russell et al. 2007). Nitrogen-fixing species, owing to their ability to fix N₂ through microbial symbiosis, can increase soil C and N and have been widely used as pioneer plants in the recovery of degraded land in the tropics and subtropics (Fisher 1995; Johnson and Curtis 2001). In Brazil, for example, N-fixing tree species have been successful in revegetating degraded land, mostly as a result of their contribution of 12 Mg ha⁻¹ year⁻¹ dry litter and 0.19 Mg ha⁻¹ year⁻¹ N (Franco and De Faria 1997). Macedo et al. (2008) also found that using N-fixing species (Acacia auriculiformis and Acacia mangium among others) for the recovery of tropical forests increased soil C and N stocks by 1.73 and 0.13 Mg ha⁻¹ year⁻¹, respectively. In Australian degraded subtropical land, intentionally elevated densities of N-fixing species tended to increase the litter and soil N content when re-establishing a self-sustaining eucalyptus forest (Grant et al. 2007). In these studies, both the soil organic matter (SOM) and soil N content were increased by N-fixing trees.

Nitrogen mineralization, a major process supplying mineral N to plants in terrestrial ecosystems, is a soil microbial process that is regulated by many abiotic and biotic factors (Hayatsu et al. 2008; Matsuoka et al. 2006; Sano et al. 2006). The use of N-fixing tree species in forest restoration also greatly affects the rates of N mineralization and other N transformations (Knoepp and Swank 1998; Knops et al. 2002). Rhoades et al. (1998) found that nitrification rates were fivefold faster and NO₃ pools fourfold greater under N-fixing trees than under pasture grasses. Scowcroft et al. (2004) also reported significantly increased N availability in surface soils as a result of reforestation with koa, an N-fixing tree native to Hawaii. In Brazil, researchers found that rapid N accumulation by N-fixing tree species caused higher soil nitrification rates over immobilization (Siddique et al. 2008).

Southern China, located mostly in a subtropical region, has 25 million hectares of forest plantations, most of which have been established on degraded land (Peng et al. 2009; Ren et al. 2007). Nitrogen-fixing Acacia spp. were introduced to this region in the 1960s as afforestation trees to conserve water and soil and to improve soil fertility in degraded areas (Yang et al. 2009). Although the role of different tree species in restoring soil processes has been investigated in numerous studies, only a few studies have been conducted in southern China, and most of these have focused on the effects of forest management and conversion (Mo et al. 2003; Xiang et al. 2009; Yan et al. 2008). Specific knowledge on N-fixing species and non-N-fixing species in forest restoration in this region is still scarce.

In the present study, we investigated soil properties and N transformations in two N-fixing species plantations, one eucalyptus plantation, two native species plantations and one secondary shrub, all of which were established in 1984 on a degraded grassland site. The objective of the present study was to compare soil properties and N transformations among the six forest types to evaluate the potential of N-fixing and non-N-fixing species to recuperate degraded land in southern China.

Materials and methods

Study area

The experimental site is located at the Heshan National Field Research Station of Forest Ecosystems (60.7 m a.s.l., 22°34′N, 112°50′E), which is in the subtropical region of southern China (Fig. 1). The region has a subtropical monsoon climate. The mean annual temperature is 21.7°C, the mean annual rainfall is 1,700 mm and the annual potential evaporation is approximately 1,600 mm (Fig. 2). The annual cycle includes a hot and rainy (growing) season (from April to September) and a cool and dry (dormant) season (from October to March). Climax vegetation in this region is subtropical evergreen broad-leaved forest. As a result of long-term disturbances, the soil in this area has eroded and the original vegetation has almost disappeared, leading to vast areas of degraded land (Yu and Peng 1996).

Figure 1 Map of study site (Heshan station).

Figure 2 Monthly mean soil temperature (measured at a soil depth of 5 cm) and rainfall amount at Heshan station in 2007.

Experimental design

The experimental area is typical of the region, with low hills (<30 m vertical distance between the lowest and highest points) and small catchments (each having an area of approximately 5–8 ha). The slope of the area is between 20 and 30° and the soil type is classified as an acrisol developed from sandstone, with a pH of approximately 4.0. In 1984, six adjacent catchments vegetated only with grass were chosen for a scientific study on the basis of their similarity. Five experimental plantations (Eucalyptus citriodora monoculture, Acacia mangium monoculture [N-fixing], Acacia auriculiformis monoculture [N-fixing], Schima superba monoculture and a native species mixed plantation [mainly Schima wallichii and Castanopsis hystrix]) were randomly allocated to each catchment and trees were planted on a 2.5 m × 3 m grid. There was also an unplanted area (approximately 2 ha) left to observe the effects of natural revegetation. These forests have been protected for scientific research since their establishment (Li et al. 2001). Unfortunately this design has inherent limitations because site and forest-type effects are confounded. However, in this particular situation, the initial soil properties were similar among the plantations, for example, in 1986, 2 years after initial planting, the SOM (0–15 cm) ranged from 14.33 ± 1.08 g kg⁻¹ in the eucalyptus plantation to 16.39 ± 0.21 g kg⁻¹ in the A. mangium plantation; this difference was not significant (Tan 2008). Thus, any differences found later among plantations can be considered to result mainly from the forest type because of the homogenous nature of the experimental area prior to afforestation (Li et al. 2001). In 2007, three 20 m × 20 m replicate plots were randomly selected in each forest type (including the secondary shrubland). Detailed vegetation information of the six forest types in 2007 is shown in Table 1.

Soil sampling and measurement of nitrogen transformations

In 2007, an in situ soil-core technique (Raison et al. 1987) was used to estimate soil net N mineralization and N leaching rates. Nine sample points were randomly located in each 20 m × 20 m replicate plot. At each of these points, two polyvinyl chloride tubes (4.6 cm in diameter and 15 cm in height) were hammered into the soil to a depth of 10 cm. Forest floor litter was removed before sampling. One of the two tubes from each subplot was retrieved and sent to the laboratory (S0). The other tube, with a lid on the top and holes on the upper 5 cm sidewall for aeration, was incubated in situ for 1 month (30 days) before being retrieved (S1). Initial soil sampling was done in June and December, representing the typical rainy and dry seasons in this region.

All soil cores were transported to the laboratory immediately, stored at 4°C and extracted for mineral N within 48 h of sampling. Before extraction, each of the nine cores from the same plot was manually divided into two layers (0–5 cm and 5–10 cm) and soils from the same section (from the same plot) were pooled and mixed thoroughly. Visible roots and stones were removed manually. Twenty grams of fresh soil from each layer was extracted with 100 mL of 2 mol L^–1 KCl solution (1:5) and filtered (202# filter paper, Shuangquan Corp., Shanghai, China). The concentrations of ammonium and nitrate in the extraction solution were determined using a flow injection autoanalyzer (FIA) (Lachat Instruments, Loveland, CO, USA); ammonium was determined using the salicylate–nitroprusside method and nitrate by sulfanilamide colorimetry after the Cd-core reduction to nitrite. Soil moisture was determined by weight loss after oven-drying at 105°C for 24 h. Bulk density was calculated based on the weight of the dried soils in all tubes.

Table 1 Vegetation description of the six forest types (including a shrubland) at Heshan Station

Forest type	Main species	Height (m)	Coverage (%)	d.b.h. (cm)
AM	Acacia mangium	20.3	90	23.2
AA	Acacia auriculiformis	19.4	80	22.5
EU	Eucalyptus citriodora	25.4	90	18.5
SS	Schima superba	12.1	80	14.3
MX	Schima wallichii	12.3	80	13.6
	Castanopsis hystrix	11.2		12.7
SL	Litsea cubeba	3.3	60	–
	Evodia lepta	5.6		–
Data were collected in October 2007, 23 years after the initial planting. AA, Acacia auriculiformis; AM, Acacia mangium; d.b.h., diameter at breast height; EU, Eucalyptus citriodora; MX, native species mixture; SL, secondary shrubland; SS, Schima superba.

Net N mineralization was calculated as the increase in ammonium plus nitrate N between the initial soil sample (S0) and the incubated sample (S1), and net nitrification was the increase in nitrate. Soil net N mineralization and nitrification in the 0–10 cm soils were computed as the summation of changes in the 0–5 cm soil layer and the 5–10 cm layer. Although the method used here may contain some “artificial” effects (e.g. root severing and the exclusion of a root effect on soil moisture) on soil N processes (Jussy et al. 2004), other studies have suggested that such effects are minor and a short incubation period (<4 weeks) was recommended to overcome this problem (Adams et al. 1989; Stenger et al. 1996). In this case, we incubated for 1 month (30 days), a time period practiced by many researchers (Uri et al. 2008; Yan et al. 2008).

Soil chemical properties

Soil chemical properties (i.e. soil pH, soil exchangeable cations, organic matter, total N and C/N) were determined using soil samples from the S0 cores collected in June 2007. All soil samples were air-dried and passed through a 2 mm sieve. Soil pH was measured in a 1:5 mixture of soil : deionized water. Soil exchangeable K⁺, Na⁺, Ca²⁺ and Mg²⁺ were extracted with 1 mol L^–1 NH₄Ac (Liu et al. 1996) and measured by inductively coupled plasma (ICP) (Perkin Elmer, Waltham, MA, USA). Soil available P was extracted with Bray-2 solution (Bray and Kurtz 1945) and determined using the molybdate blue colorimetric method. Soils for the analyses of total N (TN) and organic matter were ground to pass through a 0.25-mm sieve. The TN concentration was determined by micro-Kjeldahl digestion followed by colorimetric determination on the Lachat FIA. Soil organic carbon (SOC) was determined using the wet combustion method, and SOM was calculated as SOM = 1.73 × SOC (Liu et al. 1996). The soil C/N ratio was calculated as the ratio of SOC to soil TN.

Statistical analyses

In the present study, all soil variables in the six forest types were measured in 0–5 cm soils and 5–10 cm soils separately. A two-way anova, with forest type (including secondary shrubland, n = 6) and soil depth (n = 2) as the main factors, was used to analyze the respective effects of forest type and soil depth (as well as their interactions) on soil NH₄ ⁺, NO₃ ⁻, available P and other soil properties (SOM, TN, C/N, bulk density, soil pH and exchangeable cations). In each soil layer, if a significant effect of forest type was found, a least significant difference post-hoc test was carried out after a one-way anova to test the differences of the above variables between the two specific forest types. Area-based (0–10 cm soils) measurements of N transformations in each season, including N mineralization rates and nitrification rates, were analyzed using a one-way anova. The equality of variance in the data was tested by Levene’s test. As the variances of exchangeable K⁺, Na⁺, Ca²⁺ and Mg²⁺ were not homogeneous, a non-parametric rank analysis (Scheirer–Ray–Hare test) was used for these variables (Scheirer et al. 1976). Pearson correlation was used to detect any relationships among soil pH, SOM, TN and exchangeable K, Na, Ca and Mg. All analyses and computations were carried out in SPSS 15.0 (SPSS Inc., Chicago, IL, USA) and Excel 2003 (Microsoft Corp., Redmond, WA, USA) software.

Results

Soil carbon and nitrogen pools and pH

Soil organic matter and TN concentration all varied significantly among forest types and between soil depths (Table 2). There was also a significant positive correlation between SOM and TN (P < 0.01; Table 3). In the 0–5 cm soils, plantations of the N-fixing species A. mangium and A. auriculiformis had significantly higher SOM and TN concentrations than the E. citriodora, S. superba and mixed native species plantations (Table 2). In the 5–10 cm soils, however, there was no difference among the forest types (Table 2). Both SOM and TN concentrations decreased with soil depth. Soil C/N ratios did not differ among forest types (Table 2), but were significantly higher in the 0–5 cm soils than in the 5–10 cm soils (P < 0.001).

Table 2 General soil properties in the 0–5 cm and 5–10 cm soils of the six forest types at Heshan station in 2007

Variables	Soil layer	AM	AA	EU	SS	MX	SL	Ft	Sd	Ft × Sd
SOM (g kg^−1)	0–5 cm	78.9^c ± 3.78	73.4^b ± 7.68	50.2^a ± 4.42	49.4^a ± 6.55	46.3^a ± 9.89	52.0^ab ± 2.69	<0.01	<0.001	ns
	5–10 cm	32.6 ± 4.03	28.5 ± 4.55	30.2 ± 9.17	23.5 ± 3.80	17.6 ± 3.28	29.1 ± 2.17
TN (g kg^−1)	0–5 cm	1.39^c ± 0.08	1.13^b ± 0.10	0.91^a ± 0.01	0.84^a ± 0.01	0.88^a ± 0.07	0.98^ab ± 0.05	<0.01	<0.001	<0.01
	5–10 cm	0.68 ± 0.03	0.71 ± 0.07	0.73 ± 0.11	0.64 ± 0.10	0.73 ± 0.06	0.66 ± 0.05
C/N	0–5 cm	33.2 ± 2.33	37.4 ± 4.97	32.1 ± 2.72	33.8 ± 4.27	30.5 ± 4.26	30.7 ± 0.09	ns	<0.001	ns
	5–10 cm	27.7 ± 2.38	23.1 ± 2.02	26.5 ± 10.3	21.4 ± 5.05	15.4 ± 1.71	26.0 ± 2.00
BD (g cm^−3)	0–5 cm	0.88 ± 0.03	1.04 ± 0.07	0.91 ± 0.04	1.01 ± 0.03	0.98 ± 0.07	0.99 ± 0.05	<0.05	<0.001	ns
	5–10 cm	1.16^ab ± 0.02	1.31^c ± 0.02	1.27^bc ± 0.02	1.17^abc ± 0.06	1.04^a ± 0.09	1.25^bc ± 0.04
Soil pH	0–5 cm	3.81 ± 0.02	3.85 ± 0.03	3.88 ± 0.04	3.87 ± 0.04	3.88 ± 0.03	3.85 ± 0.02	<0.05	<0.05	ns
	5–10 cm	3.81 ± 0.06	3.92 ± 0.02	3.94 ± 0.02	3.96 ± 0.04	3.97 ± 0.03	3.85 ± 0.03
AA, Acacia auriculiformis; AM, Acacia mangium; BD, bulk density; EU, Eucalyptus citriodora; Ft, forest types effects; MX, native species mixture; Sd, soil depth effects; SL, secondary shrubland; SOM, soil organic matter; SS, Schima superba; TN, total nitrogen. Data are mean ± standard error (n = 3). In each layer, a one-way anova was used to detect differences among forest types; means sharing the same superscript were not significantly different at P = 0.05 (least significant difference).

Table 3 Pearson’s correlation analysis of soil pH, soil organic matter, total nitrogen and exchangeable K, Na, Ca and Mg

	pH	K	Na	Ca	Mg	SOM
K	−0.016	1
Na	−0.046	−0.166	1
Ca	−0.107	−0.324	0.550*	1
Mg	−0.535*	0.005	0.267	0.478*	1
SOM	−0.720**	0.050	0.507*	0.199	0.660**
TN	−0.511*	−0.097	0.567*	0.447	0.674**	0.732**
Significant correlations are indicated as follows: *P < 0.05, **P < 0.01. Sample sizes are 18 for all variables. This analysis was based on the average values of these variables in the 0–5 cm and 5–10 cm soils. SOM, soil organic matter; TN, total nitrogen.

Soil pH was similar among the six forest types (Table 2). For example, in the 0–5 cm soil, the highest pH value (3.88) was found in the mixed plantation and in the Eucalyptus plantation, but this value was only 0.07 pH unit higher than the lowest value (3.81), which was recorded in the A. mangium plantation. The pH in the 0–5 cm soils was significantly lower than that in the 5–10 cm soils (P < 0.05), in accordance with higher SOM concentration in the 0–5 cm soils. In the correlation analysis, soil pH was mainly negatively related to SOM and TN (P < 0.05; Table 3).

Soil exchangeable cations

Soil exchangeable cations varied significantly among forest types, with the exception of Mg²⁺, which varied greatly among replicates and had no pronounced variation among treatments (Fig. 3). The soil exchangeable K⁺ concentration under the S. superba plantation was the highest in both soil layers, and significantly greater than the values recorded in the other plantations (Fig. 3a,b). The three exotic-species plantations (E. citriodora, A. mangium and A. auriculiformis) had approximately 20% higher exchangeable Na⁺ concentration than the S. superba plantation and the shrubland in both soil layers (Fig. 3c,d). For Ca²⁺, the secondary shrubland had the lowest concentration among the six forest types in the 0–5 cm and 5–10 cm soils. In particular, almost no exchangeable Ca²⁺ was detected in the 5–10 cm soil of the secondary shrubland.

Soil exchangeable cation concentrations all declined significantly with an increase in soil depth (Fig. 3). Soil exchangeable K⁺ and Na⁺ in the 5–10 cm soils were over 70% of those in the corresponding upper soils, whereas they were only 50% and even less for exchangeable Ca²⁺ and Mg²⁺.

Soil inorganic nitrogen and available phosphorus

Soil extractable ammonium varied significantly among the forest types in both the rainy and dry seasons (Fig. 4a,b; P < 0.05 for both). In June, the N-fixing A. mangium plantation and the shrubland had the highest soil ammonium concentration, and ammonium concentrations were 2–6 mg kg⁻¹ lower in the 5–10 cm soil than in the 0–5 cm soil. The highest ammonium concentration in the 5–10 cm soil was found under shrubland (8.0 mg kg⁻¹). In December, the shrubland and the A. mangium plantation still had the highest concentration of ammonium, and the pattern among forests was similar to that in the rainy season (Fig. 4b).

Unlike soil ammonium, soil nitrate concentration differed significantly among forest types, but only in the rainy season (P = 0.001; Fig. 4c). In June, soil nitrate was highest under shrubland, followed by the N-fixing A. mangium in both soil layers (Fig. 4c). In December, soil nitrate did not differ significantly among forests owing to large variance within treatments (Fig. 4d).

Soil available P varied significantly among forest types in the rainy season, but not in the dry season (Fig. 4c,d). In June, the N-fixing A. auriculiformis plantation had the highest available P concentration in both soil layers, and the concentration was significantly higher than that under Eucalyptus citriodora, A. mangium and the shrubland (Fig. 4e). In December, the available P concentration greatly declined, and was only approximately 1.0 mg kg⁻¹ in both soil layers (Fig. 4f).

Nitrogen mineralization and nitrification

Differences in soil N mineralization and nitrification among forest types were not significant in either season (Table. 4). Net N mineralization rates were high in the rainy season, ranging from 7.4 to 11.3 kg N ha⁻¹ month⁻¹, but negligible or even negative (net immobilization) in the dry season. Nitrification dominated the process of N mineralization. In the rainy season, nearly 100% of mineralized N was nitrified. Soil nitrification rates were much lower in the dry season than in the rainy season, which was similar to the trend for N mineralization (Table. 4).

Figure 3 Soil exchangeable (a) K+, (b) Na+, (c) Ca2+ and (d) Mg2+ in the 0–5 cm and 5–10 cm soils under the six forest types at Heshan station (mean ± standard error, n = 3). AA, Acacia auriculiformis; AM, Acacia mangium; EU, Eucalyptus citriodora; MX, native species mixture; SL, secondary shrubland; SS, Schima superba. Bars sharing the same superscript letter are not significantly different at P = 0.05 (least significant difference).

Figure 4 (a, b) Soil extractable ammonium, (c, d) nitrate and (e, f) available P in the rainy season (a,c,e) and (b,d,f) dry season of 2007 in the six forest types at Heshan station (mean ± standard error, n = 3). AA, Acacia auriculiformis; AM, Acacia mangium; EU, Eucalyptus citriodora; MX, native species mixture; SL, secondary shrubland; SS, Schima superba. Bars sharing the same superscript letter are not significantly different at P = 0.05 (least significant difference).

Discussion

Effects on soil properties

In the restoration of degraded areas, C input and an increase in soil N content are of great importance because they can enhance the capacity of the system to support a more complex community (Franco and De Faria 1997; Macedo et al. 2008). Nitrogen-fixing species have been used as an N source in the recovering of tropical and subtropical systems, including degraded mining land (Franco and De Faria 1997), deforested land (Siddique et al. 2008) and agroforestry (Handayanto et al. 1995). In these studies, not only soil N, but also SOM has been increased by N-fixing species (Deans et al. 1999; Franco and De Faria 1997; Macedo et al. 2008). The SOM and TN results in the present study also showed that the two N-fixing species (A. mangium and A. auriculiformis) were able to restore C and N cycling better than the non-N-fixing species and natural revegetation (shrubland) in southern China; these results are consistent with observations from other tropical and subtropical regions (Macedo et al. 2008; Siddique et al. 2008; Stock et al. 1995).

The higher SOM and TN concentration in the 0–5 cm soils, relative to the non-N-fixing plantations (Table 2), may result from higher litter production and lower litter decomposition of N-fixing species. It is known that litter production and the rate of litter decomposition are the most important factors by which tree species regulate the size and distribution of soil C and N pools (Aerts and de Caluwe 1997; Finzi et al. 1998a; Stump and Binkley 1993). Li et al. (2000) reported litter-fall mass at this site in the order of A. mangium (11.1 t ha⁻¹) > S. superba (6.5 t ha⁻¹) > A. auriculiformis (4.8 t ha⁻¹) > Eucalyptus citriodora (2.6 t ha⁻¹). In addition, the litter of N-fixing A. mangium and A. auriculiformis had a much lower decomposition rate compared with other species (Eucalyptus citriodora and S. superba) (Li et al. 2000, 2001). Thus, the large differences in litter production and the rate of litter decomposition between N-fixing and non-N-fixing species may have contributed to the higher soil C and N pools under the two N-fixing species.

Table 4 Net nitrogen mineralization and nitrification rates (data were combined in the 0–10 cm soils) in the six forest types at Heshan station in 2007

Forest type	Nitrogen mineralization (kg N ha^−1 month^−1)		Nitrification (kg N ha^−1 month^−1)
	Rainy	Dry	Rainy	Dry
AM	11.0 ± 1.62	−1.46 ± 1.93	14.51 ± 1.85	0.37 ± 0.54
AA	10.9 ± 2.08	0.93 ± 0.92	14.67 ± 1.40	4.09 ± 0.35
EU	8.56 ± 3.24	−2.46 ± 0.86	11.04 ± 2.95	1.27 ± 0.94
SS	11.3 ± 2.96	−2.01 ± 0.81	14.25 ± 2.24	1.44 ± 1.62
MX	7.41 ± 1.09	−0.78 ± 0.90	10.36 ± 1.12	1.79 ± 0.2
SL	8.03 ± 1.99	1.68 ± 3.14	10.30 ± 2.10	1.82 ± 0.41
P-value	ns	ns	ns	ns
AA, Acacia auriculiformis; AM, Acacia mangium; EU, Eucalyptus citriodora; MX, native species mixture; SL, secondary shrubland; SS, Schima superba. Data are mean ± standard error (n = 3).

In the present study, an acidification effect by N-fixing tree species was not significant. However, previous studies have shown that N-fixing species acidified their rooting soil more than non-N-fixing species (Haynes 1983). For example, in Hawaii, Rhoades and Binkley (1996) observed that soil pH declined more dramatically (from 5.9 to 4.5) in an N-fixing (Albizia) plantation than in a Eucalyptus plantation (from 5.9 to 5.0). Yamashita et al. (2008) also found that the surface soil pH in an 8-year-old N-fixing A. mangium plantations was 1 pH unit lower than that in a non-N-fixing Imperata cylindrica grassland. In the present study, however, there was very little difference (0.1 unit) in soil pH between the N-fixing and non-N-fixing species in either the 0–5 cm or the 5–10 cm soils (Table 2). Although nitrogen fixation and nitrification processes could lead to soil acidification, other factors, such as initial soil condition, atmospheric acid deposition, the export of cations by leaching or plant uptake and litter decomposition, could also acidify soil (Finzi et al. 1998b; Yamashita et al. 2008). At the present site, soil pH in the initial soil condition was strongly acidic, below pH 4.2 (Yu and Peng 1995). As soil has a strong buffering capacity in acidic conditions owing to dissolving aluminum from clay minerals (van Breemen et al. 1984), these buffering processes may contribute to the insignificant difference between the N-fixing and non-N-fixing sites. Another possibility is that there is no significant difference in nitrification rates among forest types. Nitrification could reduce soil pH by producing protons in its chemical processes, thus resulting in a decline in soil pH under N-fixing tree sites.

The soil in the N-fixing A. auriculiformis plantation had the highest available P concentration in both soil layers. We also observed a higher soil available P concentration under A. auriculiformis seedlings than under other species, such as A. mangium and S. superba, in a pot experiment (Wang F, 2009, unpubl. data). In southern China, P deficiency is a general problem for most plantations (Ren et al. 2007). Thus, Acacia auriculiformis may be a better choice to improve soil P availability in comparison to other species. As tree species can regulate soil P availability through their effect on soil pH and phosphatase activities (Zou et al. 1995), further studies should be done to show that A. auriculiformis can affect soil P availability.

Effects on soil nitrogen transformations

There were no significant differences in N mineralization or nitrification between the N-fixing species and the non-N-fixing species in the present study. This result differed from previous studies, which have demonstrated that soils under N-fixing species had higher N transformation rates than those under non-N-fixing species (Bernhard-Reversat 1988; Siddique et al. 2008). Our result was also inconsistent with observations made in these plantations after 13 years of growth, when Li et al. (2001) found that nitrification rates in the two N-fixing species plantations were much greater than those in the non-N-fixing E. citriodora and S. superba stands. One possibility is that over 20 years other sources of N input, such as regional atmospheric deposition, have greatly increased N availability in all forests. Fang et al. (2008) measured high inorganic N deposition in this region at a rate of over 30 kg N ha⁻¹ year⁻¹ and another 10–20 kg N ha⁻¹ year⁻¹ deposition in the form of dissolved organic N.

In the present study, net N mineralization rates in the rainy season ranged from 7.41 to 11.3 kg N ha⁻¹ month⁻¹ (Table 4). The rates in these forests are similar to values found in climax forest in this region. In the Dinghushan climax evergreen broad-leaved forest, 65 km from our sites, Fang (2006) detected a 6.7 kg N ha⁻¹ month⁻¹ soil N mineralization rate. Li et al. (2006) worked in an evergreen broad-leaved forest in Yunnan Province and found that annual N mineralization in the 0–15 cm soils was 159.12 kg N ha⁻¹ year⁻¹, equivalent to 12–18 kg N ha⁻¹ month⁻¹. We estimated that the annual N mineralization rate of the 0–10 cm soils in the six forest types ranged from 60 to 100 kg N ha⁻¹ year⁻¹. Thus, the values were similar to those found in natural climax forests of subtropical China. Based on the study in Dinghushan, Mo et al. (2003) suggested that over a period of approximately 50 years, successful rehabilitation of soil N availability on severely degraded lands is possible. However, our results suggested that soil N availability could be mostly recovered after 23 years of forest restoration.

Contrary to our expectations, soil inorganic N, particularly nitrate, was highest not under the two N-fixing plantations, but under the naturally regenerated shrubland. This result differs from the patterns of soil total N pools (Table. 2). In general, variations in soil inorganic N can be attributed to three factors: plant uptake, N mineralization and nitrification, and losses through leaching. As there were no differences in soil N mineralization and N leaching loss (data not shown) among forest types in the present study, plant uptake is probably the main factor that led to the variation in soil inorganic N. In a 2007 survey, the height of the overstory vegetation was less than 6 m in the shrubland and over 10 m in the plantations (Table 1). In the A. mangium and E. citriodora stands, the tallest trees were over 25 m. Therefore, it is likely that the low plant uptake in the shrubland explains the high inorganic N concentration in that system.

Conclusions

Soil properties and N transformations in six forest types were compared in the present study to evaluate the effects of N-fixing and non-N-fixing species on forest restoration in southern China. The N-fixing forests had 40–50% higher SOM and 20–50% higher TN concentration in the 0–5 cm soils than the non-N-fixing forests, suggesting that N-fixing species, particularly A. mangium, are more efficient in re-establishing C and N cycling processes in the degraded land of southern China. There were no significant differences in soil N mineralization and nitrification among the forest types. In the rainy season, the rates of N mineralization (7.41–11.3 kg N ha⁻¹ month⁻¹) were similar to values in regional climax forests, indicating that soil N transformations have been well recovered in these 23-year-old plantations. Soil inorganic N was highest under the secondary shrubland and second highest under the N-fixing Acacia mangium. Lower plant uptake could be responsible for the higher inorganic N in the shrubland. The present results show that after 23 years’ growth, N-fixing species performed better in restoring soil C and N pools and their cycling.

Acknowledgments

The project was supported by National Basic Research Program of China (973 Program, 2009CB421101) and the National Natural Science Foundation of China (30630015 and 30870442). We also thank Professor Murray McBride from Cornell University and Dylan Horvath and Tim Scott from Binghamton University-SUNY for their critical comments on this manuscript.