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Change in soil carbon after land clearing or afforestation in highly weathered lateritic and sandy soils of south-western Australia
Agriculture, Ecosystems and Environment 95 (2003) 143–156
Change in soil carbon after land clearing or afforestation in highly weathered lateritic and sandy soils of south-western Australia D.S. Mendham∗ , A.M. O’Connell, T.S. Grove CSIRO Forestry and Forest Products, Private Bag No. 5, Wembley, WA 6913, Australia Received 29 November 2001; received in revised form 15 May 2002; accepted 7 June 2002
Abstract Soil organic carbon (SOC) is a significant component of the worlds terrestrial carbon stocks, and changes in land-use have potential to change pools of soil C. Land-use change is occurring over large areas of the world, so soil has potential to be a large source or sink for atmospheric C. Changes in amounts of soil C (0–1 m) were examined after clearing of native vegetation for pasture (20–71 years prior to study), and after establishment of Eucalyptus globulus plantations on ex-pasture sites (7–10 years prior to study). A suite of 10 sites in south western Australia were used in a three-way comparison between native vegetation, pasture and plantation. Soil types across the 10 sites included Acrisols, Arenosols, and a Ferrosol. Soil C was measured in the fine earth (<2 mm), the 2–5 mm soil fraction, in charcoal and roots retained on a 5 mm screen, and in surface litter. Differences in soil bulk density between land-uses were accounted for by calculating depths for equivalent weights of soil. Despite large increases in soil fertility with conversion to pasture, amounts of soil C changed little. The amount of C in the 5–20 cm increment was significantly greater under pasture than native vegetation (mean of 8.4 Mg ha−1 , <5 mm), but in surface soil (0–5 cm) and <20 cm, there were no significant differences in soil C content between land-uses. Less C in soil at 5–20 cm under native vegetation was offset by significantly more C in coarse roots (average of 5 Mg ha−1 higher to 1 m depth), surface soil (2–5 mm fraction, 2 Mg ha−1 ), and in standing pools of surface litter (9.8 Mg ha−1 ). Amounts of soil C under plantation were not significantly different from pasture 7–10 years after plantation establishment. However, plantation soils had an average of 3.1 Mg ha−1 more C in coarse roots than pasture, and significant quantities of C in surface litter (average of 7.9 Mg ha−1 ). Overall, soil C content in the sites of this study was relatively stable and the effect of land-use change was limited. © 2002 Elsevier Science B.V. All rights reserved. Keywords: Soil organic carbon; Root C; Litter C; Land-use change; Native vegetation; Pasture; Plantations; Eucalyptus globulus; Western Australia
1. Introduction The amount of soil organic carbon (SOC) globally has been estimated to contain more than three times the atmospheric pool, and more than four times the biotic pool of C (Lal, 2001). Changes in amount of ∗ Corresponding author. Tel: +61-8-9333-6663; fax: +61-8-9387-8991. E-mail address: [email protected] (D.S. Mendham).
SOC with change in land-use have been the subject of significant attention recently because of widespread recognition of the need to limit emissions of greenhouse gases. Change in land-use has potential to either release or sequester soil C. Clearing of native vegetation for agricultural production often leads to a reduction in SOC (Mann, 1986; Davidson and Ackerman, 1993), whilst potential exists for significant increase in SOC through changed management practices (Lal, 2001) or reafforestation (Post and Kwon, 2000).
0167-8809/02/$ – see front matter © 2002 Elsevier Science B.V. All rights reserved. PII: S 0 1 6 7 - 8 8 0 9 ( 0 2 ) 0 0 1 0 5 - 6
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The rate of accumulation or loss of SOC after change in land-use is governed by the balance between addition and decomposition of organic material, which in turn is influenced by changes in net primary productivity and allocation (Van Cleve and Powers, 1995), quality of organic inputs (Berg, 2000), tillage practices (Balesdent et al., 2000), rooting patterns (Jobbágy and Jackson, 2000) and climate. Other site factors also influence SOC dynamics, including soil texture (Golchin et al., 1994) and climate (Van Cleve and Powers, 1995). Much of the arable land in Australia has been cleared of native vegetation in the last 20–100 years, resulting in increased soil fertility through additions of nutrients from chemical fertilizers (Weaver and Reed, 1998), fixation of atmospheric N by pasture legumes (Underwood and Gladstones, 1979), and changes in the quantity and quality of organic matter inputs. Cycling of SOC under managed pasture in south western Australia is potentially quite different from under native vegetation due to higher fertility, shallower rooting patterns of pasture species, input of residues with higher chemical quality (e.g. lower C:N ratio), and annual pasture species versus perennial forest species. Another significant change in land-use recently in south-western Australia is conversion of pasture land to eucalypt plantation, predominantly Eucalyptus globulus, which can provide economic and environmental benefits to farmers. The area under E. globulus in south-western Australia has been expanding by more than 25,000 ha per year (National Forest Inventory, 2000). Cycling of SOC under plantation is also potentially different to that under pasture, with potential differences again in net primary productivity and allocation, rates of root and litter turnover, residue quality, and microclimatic conditions. Detailed studies of C inputs have not been conducted in these ecosystems, but productivity (in terms of aboveground biomass accumulation and litterfall) is up to 3–4 Mg ha−1 per year in native Jarrah forest (O’Connell and Grove, 1996), 4.5–8 Mg ha−1 per year in annual pastures (Rossiter and Ozanne, 1970; Thomson et al., 1998), and 11–43 Mg ha−1 per year in E. globulus plantations (Hingston and Galbraith, 1998). Soil bulk density can also be modified by change in land-use, which affects the calculations of SOC stocks (Fearnside and Barbosa, 1998). Differences in soil bulk density between land-uses can be caused by
compaction from livestock and machinery, differences in volume and turnover of root material, tillage, and biological activity. Soils under pasture typically have higher bulk density, due to compaction caused by agricultural management practices, and lower volumes of coarse roots. Any effect of changed land-use on soil bulk density will influence the amount of soil that is sampled from within a fixed depth. The purpose of this study was to assess impacts of land-use on changes in total soil C to 1 m depth at 10 sites representing the climatic and soil gradient in south-western Australia. Land-uses studied were native vegetation, pasture managed for 20–71 years after clearing, and E. globulus plantations established on pasture land for 7–10 years prior to the study. The 10 sites were selected such that the soils would have been closely matched prior to change in land-use. 2. Materials and methods 2.1. Site selection and soil sampling Soil was sampled from 10 sites representing the range in soils and climates where E. globulus is grown in south-western Australia (Fig. 1). Selected soil characteristics are shown in Table 1. Total C in the surface soil at the sites ranged from 16.7 to 90.0 g kg−1 . Adjacent pasture (20–71 year old) and plantation (7–10 year old) plots were selected at each site, and a plot on a nearby native vegetation remnant on the same landform element was also selected. The native vegetation at each site was predominantly Jarrah (E. marginata) and Marri (E. calophylla) forest, which ranges in productivity, depending on rainfall. The Anning site was originally occupied by Karri (E. diversicolor) forest. The pasture consisted of a mix of annual species, including subterranean clover (Trifolium subterraneum), capeweed (Arctotheca calendula), and a range of grasses (including Lolium rigidum, Hordeum leporinum, and Bromus diandrus). Tillage operations have been relatively rare in these systems, occurring only with winter cropping approximately every 5–10 years. Pasture soil is enriched in nutrients compared to the native soil (see Table 1) because of inputs of N (from clover) and P (from annual applications of superphosphate). Potassium has been applied to some pastures, and trace elements have also
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Fig. 1. Map of south-western Australia showing locations of the 10 study sites (diamonds). The towns of Albany, Manjimup and Perth (circles) are also shown for reference.
been applied at irregular intervals. For more detail on pasture production in this region, see Rovira (1992). Plantations of E. globulus are being established on pasture sites. Seedlings are planted into rip-lines, and generally fertilized once during establishment with N and P with a spot application close to the base of the seedling. The large majority of E. globulus plantations in south-western Australia (including those in this study) are in their first rotation, with average rotation lengths expected to be about 10 years. Plantations in this study were approaching the end of their first rotation. The validity of comparisons between land-uses was qualitatively assessed by comparing pedological features (including horizon depths, color, texture, gravel type and content) of soil profiles in each of the three land-use types at each site. Soil was sampled within each land-use from a 40 m transect located at least 20 m from the boundary of the other land-uses to minimize boundary effects such as shading, tree roots entering the pasture or fertilizer spread into the native vegetation. Each 40 m transect was divided into 4 × 10 m plots, from which soil was collected using thin-walled stainless steel corers of different internal diameters (47, 59, and 70.5 mm i.d., depending on depth). Three soil cores per plot were
collected and bulked for the 0–5 and 5–10 cm depth increments, whilst single cores per plot were collected at depths greater than 10 cm. For depths greater than 10 cm, the samples were collected sequentially from the same hole within each plot, in increments of 10–20, 20–30, 30–50, 50–70, and 70–100 cm. A flange-type plastic sheath was inserted at the top to prevent surface soil and litter from falling into the hole and contaminating deeper soil samples. Additionally, smaller diameter cores were used with depth to prevent upper layers of soil falling into, and contaminating, the deeper horizon samples. All sampling in the plantation was done away from the rip-lines. Litter samples were collected in a circular sampling area (250 mm diameter). Between 1 and 4 litter samples were collected per plot, depending on the spatial heterogeneity of the litter layer, which was qualitatively assessed at each site. Litter was not collected in the pasture because there were negligible amounts due to grazing. 2.2. Soil preparation and analysis Soil and litter samples were sieved through a 5 mm sieve to separate the coarse and fine fractions. Coarse
146
Site name
Field texture
Soil classification (FAO, 1990)
Annual rainfall (mm)
Pasture age (years)
Plantation age (years)
Silt + clay proportion (%)a
pH
Total C (g kg−1 )b
Total N (g kg−1 )b
Total P (g kg−1 )b
Andrews Anning Ayers Gibbs Hall B Jefferies B Moltoni Patmore B Robinson Walker
Loamy fine sand Gravelly clayey sand Gravelly sandy clay loam Grey sand Coarse sand Sand Gravelly sandy loam Gravelly sandy loam Gravelly sandy loam Loamy fine sand
Ferric Acrisol Ferric Acrisol Xanthic Ferralsol Haplic Arenosol Haplic Arenosol Haplic Arenosol Ferric Acrisol Haplic Arenosol Ferric Acrisol Haplic Arenosol
865 1291 983 721 661 655 1358 833 1226 574
20 71 29 31 43 31 28 20 42 23
10 8 9 9 8 8 10 7 10 8
8.41 12.7 23.9 4.97 7.10 3.94 18.8 11.3 28.9 8.41
5.1 6.0 5.9 5.7 5.6 6.0 5.9 6.3 5.8 4.9
42.6 48.4 90.0 18.0 30.1 16.7 55.5 54.6 77.2 19.1
1.62 1.90 3.59 0.37 1.34 0.35 1.77 1.78 2.60 0.40
0.043 0.095 0.239 0.009 0.036 0.007 0.112 0.149 0.125 0.030
a b
(−3.7) (+4.8) (−12.5) (+78.6) (+12.8) (+18.2) (−12.6) (+0.4) (−3.4) (+15.9)
Depth: 0–10 cm. Chemical data shown for 0–10 cm soil under native vegetation with mean percentage difference in managed soils shown in parentheses.
(+55) (+73) (+16) (+344) (+54) (+137) (+54) (+37) (+68) (+150)
(+285) (+511) (+97) (+965) (+289) (+1601) (+450) (+87) (+396) (+133)
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Table 1 Characteristics of each of the sites examined in south-western Australia
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roots, charcoal and gravel fractions (i.e. those retained on the 5 mm sieve) were separated by hand. Large lateral or tap roots were not sampled. The fine soil was further separated by sieving through 2 mm, resulting in size fractions of 2–5 mm and <2 mm. The 2–5 mm soil fraction was measured separately because the native vegetation and plantation species generally have coarser roots than the pasture species, so the standard <2 mm fraction can potentially bias measurements towards finer rooted species. All fractions were oven dried, weighed, ground and analysed for total C by combustion at 1200 ◦ C and measurement of the evolved CO2 on a LECO TOC analyser (CR-412). Corrections were not made for carbonate as all soils were acidic (pH <7), and short combustion times indicated that CO2 was generated from organic, rather than inorganic sources. Soil (<2 mm) was bulked across plots for each depth, and assessed for particle size distribution and pH. The proportion of sand, silt and clay in the samples was analysed using the pipette method of Coventry and Fett (1979) after removal of the organic material by reacting with H2 O2 and dispersion by ultrasonic treatment. Soil pH was assessed after shaking for 1 h in a 1:5 soil:water suspension. Total N and P were measured on surface soil samples (0–10 cm) using a flow injection analyser to assess concentration of N and P in solution after a Kjeldahl digest (Rayment and Higginson, 1992).
land-uses in a stepwise manner. Different parameters were used between land-uses where the regressions were significantly improved by doing so. The distribution of C with depth in each profile generally fitted an exponential regression of the form
2.3. Calculations of soil carbon
2.4. Statistical analysis
Curve-fitting was used for (a) calculating the soil depth intervals for each of the managed land-uses (because changed bulk density influences the depth of soil for a constant weight), and (b) predicting the carbon content within the depth intervals calculated during the first step. Nonlinear regressions of the following form were fitted to depth trends in soil bulk density
Regression analysis was used to relate soil C to profile depth, soil texture and climatic variables using the Genstat statistical package (Lawes Agricultural Trust). Analysis of variance (ANOVA; Genstat, Lawes Agricultural Trust) was used to assess the effect of change in land-use on soil C in the different profile depths. Sites were used as replicates in the ANOVA. There were some differences in soil texture between land-uses at each site, despite efforts to ensure that sites were as closely matched as possible. Hence ANOVA analyses were conducted on total C data only after ensuring that there was no systematic bias in soil texture between the land-uses. The clay + silt content was used as a covariate in the ANOVA of soil C to help account for minor differences caused by soil texture variation between land-uses at each site.
y = a + brx + cx,
(1)
where y was the soil bulk density (<5 mm fraction), x the depth, and a, b, c and r were fitted parameters. Initially a single model was fitted to the data from each site, but the effect of land-use on the soil bulk density trend with depth was evaluated by assessing significance of applying separate constant (a), linear (b, c), and nonlinear (r) parameters to each of the
z = d + esx ,
(2)
where z was the quantity of C calculated per cm of depth (Mg ha−1 cm−1 ), x the depth (cm), and d, e and s were fitted parameters. Effects of land-use on soil C distribution were evaluated in a similar manner to that for soil bulk density, i.e. by assessing the significance of changing the constant (d), linear (e), and nonlinear (s) parameters in a stepwise manner. Different parameters were used between land-uses to calculate amounts of soil C under the different land-uses where regressions were significantly improved by doing so. Weights of soil for fixed depth increments (0–5, 5–10, 10–20, 20–30, 30–50, 50–70, and 70–100 cm) under native vegetation at each site were calculated according to Eq. (1), and depths under pasture and plantation were calculated for an equivalent weight of soil under native vegetation. The calculated depths for equivalent weights of soil in the pasture and plantation ecosystems are shown in Table 2. The total amount of C within each of the corrected depths was calculated by integrating Eq. (2).
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Table 2 Calculated depths (cm) for equivalent weight of soil under native vegetation Site name
Original depth Pasture
Andrews Anning Ayers Gibbs Hall Ba Jefferies B Moltoni Patmore B Robinson Walker a
Plantation
5
10
20
30
50
70
100
5
10
20
30
50
70
100
3.5 2.7 5 6.6 4.5 4.5 2.5 5 2.1 5.2
8 7.3 10 11.7 9.1 9.2 7.4 10 6.8 9.1
17.2 16.6 20 21.8 18.3 18.5 15.9 20 17 17.9
26.5 25.9 30 32 27.4 27.8 24.3 30 27.4 27.3
45.5 44.3 50 52.2 45.8 46.4 40.8 50 48.2 46.7
65 62.8 70 72.5
95.2 90.5 100 102.9
18.3 18 20 19.9 17.5 18.3 17.3 20 17.2 17.7
27.7 27.7 30 29.9 26.2 27.5 25.7 30 26.6 27
47 47 50 50.1 43.8 45.9 42.5 50 45.3 46
97.9 95.2 100 100.5
92.4 82.2 100 100.2 95.6
9.1 8.4 10 9.8 8.7 9.1 8.9 10 7.8 8.8
66.9 66.3 70 70.3
64.8 57.4 70 69 66.2
4.4 3.5 5 4.8 4.3 4.4 4.7 5 3.1 4.4
64.1 59.3 70 63.9 65.1
91.3 84.5 100 91.9 93.8
Soil not sampled depths greater than 50 cm at Hall B site due to presence of a hard pan.
Multiple linear regression analysis (Genstat, Lawes Agricultural Trust) was used to identify the variables influencing profile C stores across the 10 sites. Variables examined included soil texture and environmental attributes such as temperature (mean annual and mean seasonal), summer evaporation, and annual precipitation.
3. Results 3.1. Influence of land-use change on total soil C Land-use had no significant effect on mean amounts of total C in the <2 mm fraction in the surface horizon (0–5 cm) and depths greater than 20 cm (Fig. 2).
Fig. 2. Amounts of soil C in the <2 mm soil fraction. Significant differences between land-uses within each depth range are denoted by ∗: P < 0.05.
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Fig. 3. Amounts of soil C in the 2–5 mm soil fraction. Significant differences between land-uses within each depth range are denoted by ∗: P < 0.05.
However, the soil under native vegetation had significantly less total soil C in the <2 mm fraction in the 5–10 cm increment (P = 0.044), and also in the 10–20 cm increment when tested at the less stringent 0.1 level (P = 0.071). Total C under native vegetation in the 5–20 cm depth range was 6.3 and 5.6 Mg ha−1 lower than that under pasture and plantation, respectively. The average amount of total C in the 2–5 mm fraction ranged from 0.4 to 4.5 Mg ha−1 for the different depths (Fig. 3), with significantly more C under native vegetation than under pasture or plantation only at the 0–5 cm depth. Compared to the 0–100 cm increment, the mean proportion of total soil C in the 0–10 cm depth was 39.1–39.7%, and was not significantly different between land-uses (Table 3). Amounts of C in the 10–30 cm increment as a proportion of the total (0–100 cm) was less under the native vegetation than in the managed treatments, but was balanced by a higher proportion (although not significant) in the 30–100 cm depth range.
C. Coarse roots were more prevalent under the pasture than other land-uses (although not significantly) in surface soil (0–5 cm depth, Fig. 4), but C in pasture roots depths greater than 5 cm depth was significantly less than under native vegetation. There was more root C under plantations than pasture at most depths, significantly between 5 and 50 cm depth. 3.3. Total C in litter There was no surface litter under pasture because of grazing. A relatively small proportion of the plant material is returned in animal waste, but this is spatially heterogeneous and was not assessed Table 3 Proportion of total soil C (%, <5 mm fraction) to 1 m depth in different depth (cm) ranges Depth range
vegetationa
0–10
10–30
30–100
39.3 a 39.7 a 39.1 a
24.5 a 30.5 b 32.0 b
36.2 a 29.8 a 29.0 a
3.2. Total C in roots
Native Pasturea Plantationa
Differences in amounts of root C under different land-uses were more pronounced than with total
a Values with the same letters within each column were not significantly different (tested by L.S.D., α = 5%).
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Fig. 4. Amounts of soil C in the coarse roots (retained on a 5 mm screen). Significant differences between land-uses within each depth range were denoted by ∗: P < 0.1 and ∗∗∗: P < 0.001.
in this study. There were significant amounts of litter under native vegetation (mean of 9.8 Mg ha−1 , Fig. 5a) and plantation (7.18 Mg ha−1 ). The environmental variable most closely associated with the quantity of litter C was long-term annual rainfall (Fig. 5b, y = 0.0172x − 6.95, R 2 = 0.38, S.E. = 6.29, n = 20). Other variables (summer evaporation, soil texture, monthly temperatures) and combinations of these variables (in multiple linear regressions) across both land-uses did not improve on that relationship.
3.4. Total pools of C to 1 m Amounts of total soil C to 1 m ranged from 50 to 163 Mg ha−1 (Table 4) across the 10 sites, with most of the C being in the fine earth (<2 mm) fraction (30–144 Mg ha−1 ). There was less C in the fine earth under native vegetation than under pastures and plantations, but this was offset by higher soil C in the 2–5 mm fraction, and in the coarse roots. There were similar amounts of litter under native vegetation and plantation. Overall, land-use change had no effect
Table 4 Amounts and range of total C (Mg ha−1 ) in the soil, litter, and coarse roots across the 10 sites Native vegetation
Pasture
Plantation
Mean
Range
Mean
Range
Soil Soil 2–5 mma Roots >5 mma Surface littera
61.4 a 16.4 a 6.35 c 9.8 a
30.0–117.0 0–40.9 3.80–13.01 0.7–34.5
78.0 b 11.9 a 1.38 a –
41.1–143.7 0–18.3 0.48–2.96 –
Totala
94.0 a
53.7–143.1
91.3 a
49.6–162.5
<2 mma
a
Means with the same letter within each row were not significantly different.
Mean 78.4 b 12.1 a 4.53 b 7.9 a 102.9 a
Range 41.3–116.5 2.4–25.4 1.63–8.97 1.8–13.4 59.1–144.1
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Fig. 5. Amounts of litter C under the plantation and native vegetation (a) at each of the sites, and (b) relationship with average annual rainfall across the sites.
on pools of total C to 1 m depth. Charcoal that was hand-separated from the >5 mm soil fraction was not included in the above balance, but this fraction added a mean of 3.98 Mg C ha−1 to the total pool across the sites. Quantities of charcoal C were not significantly different between land-uses. Amounts of C were related to the quotient of mean temperature and precipitation (Fig. 6a, R 2 = 0.29, S.E. = 23.0) used by Brown and Lugo (1982) to predict soil C stocks across a climatic gradient. However, the site factor most highly correlated with total C across sites was soil texture (represented by the clay + silt content, Fig. 6b, R 2 = 0.32, S.E. = 22.6). Incorporation of soil texture with summer evaporation and summer mean daily minimum temperature in a multiple linear regression gave a more significant relationship than with the clay + silt content alone (Fig. 6c, R 2 = 0.50, S.E. = 19.3). There were no significant differences between land-use types for any of these relationships.
4. Discussion Amounts of soil C to 1 m in the soils of this study (50–163 Mg ha−1 ) were less than those published previously for Jarrah (184 Mg ha−1 , Hingston et al., 1981) and Karri forest (177–178 Mg ha−1 , Hingston et al., 1979) in south-western Australia, but the three sites in the studies of Hingston et al. (1979, 1981) had relatively high rainfall and productivity. Sites with lower rainfall (which also generally had lower clay content and higher evaporation) generally corresponded to lower profile C contents (Fig. 6c), although there was a relatively high degree of scatter in that relationship. The range found in this study was similar to those published for Brazilian Amazonia (94–113 Mg ha−1 , Fearnside and Barbosa, 1998; Silver et al., 2000), despite a significantly drier climatic regime in south-western Australia. Jobbágy and Jackson (2000) also found a similar range of soil C stored in a range of ecosystems from
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Fig. 6. Relationship between total soil C (<5 mm fraction) and (a) the quotient of mean annual temperature and annual precipitation, (b) the clay + silt content to 1 m, and (c) a function of the clay + silt content 0–100 cm (c), mean summer evaporation (e), and summer mean daily minimum temperature (T).
deserts (62 Mg ha−1 ) to tropical evergreen forests (186 Mg ha−1 ). However, amounts were much less than those under the E. grandis plantations of eastern Australia (200–400 Mg ha−1 in the top 50 cm) studied by Turner and Lambert (2000). The significant relationship between soil C and the quotient of temperature and precipitation (Fig. 6a) followed a similar trend
to that of Brown and Lugo (1982) and Tate (1992), but the amount of soil C at any given climatic ratio was generally higher than found by Tate (1992). The temperature/precipitation ratio was mainly affected by rainfall across the sites in this study because mean annual temperatures across the sites varied little (between 14.8 and 15.5 ◦ C). Soil texture was more closely
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related to soil C stocks than climatic factors (Fig. 6b), but it was not possible to separate the confounding effects of climate and soil texture across the sites, because rainfall was correlated with soil clay content (R 2 = 0.47). Whilst mean annual temperatures were similar across the sites, the daily minima and maxima varied, depending on distance from the sea. Incorporating mean daily summer minimum temperature as a climatic variable improved the regression in Fig. 6c, such that 50.4% of the variation in the amount of soil C (0–100 cm) was described by those three factors. A significant weakness of many studies investigating the effect of changed land-use on soil properties is the reliance on too few sites in the paired approach. This can lead to treatment effects being attributed to the fine-scale differences in soil properties that occur naturally (Fearnside and Barbosa, 1998). Hence the effect of land-use on soil properties is potentially confounded by differences in the soils. To effectively test the effect of land-use change with the paired site approach is only possible if enough sites are sampled to be used as replicates, or that regression analyses can be conducted using independent site properties. The use of 10 well-matched sites in this study was a sound basis for generalizing the effects of land-use change on soil C stores in south western Australia. The three-way comparison of this study allowed comparison between two different land-use changes—from native vegetation to pasture (20–71 years prior to study), and from pasture to plantation (7–10 years prior to study). 4.1. Effect of land clearing on soil C Many studies have shown that clearing of native vegetation, followed by intensive management with regular cultivation often causes large reductions in soil C (Mann, 1986; Davidson and Ackerman, 1993), but conversion to less intensive management with infrequent tillage (such as pasture) has variable effects, ranging from a reduction in total C (Veldkamp, 1994; Guggenberger and Zech, 1999; Glaser et al., 2000; Rhoades et al., 2000), through unchanged amounts (Reiners et al., 1994; Veldkamp, 1994; Bell et al., 1995), to increased amounts of total C (Neill et al., 1997; Fearnside and Barbosa, 1998). Amounts of soil C in this study were significantly higher in 5–20 cm soil under managed land-uses, but some of this C was offset by lower amounts in the roots, litter and
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2–5 mm fraction under pasture (0–5 cm depth). The amount of C in the roots and litter may be more variable on a seasonal basis than soil C (Attiwill et al., 1978; Fahey and Hughes, 1994), but the values found at a single sampling time were useful for comparing likely magnitudes across different land-uses. Soils were sampled in late spring, after the main growing season, so the quantities of roots and litter were probably about average on an annual basis. If the soils had been sampled during the main growing season, the quantity of carbon in fine roots under pasture would probably have been greater, although the roots would still have been highly stratified, and mainly concentrated in the surface soil (0–5 cm). Changes in amounts of soil C content with land-use are caused by a disruption of the balance between inputs from plant litter (above- and below-ground) and losses from decomposition, leaching and erosion (Schlesinger, 1977). Conversion to pasture in these systems has led to marked increases in soil fertility (Mendham et al., 2002) and higher soil moisture content for much of the year (Farrington et al., 1992), so there is potential for higher decomposition rates under pasture. However, the observation that soil C increased under pasture (in the 5–20 cm depth range) suggested that C additions to that depth must also be significantly greater than those under native vegetation. Addition of C to the lower depths may be attributed to mixing of surface layers during cultivation, and/or through higher net primary productivity. Root dynamics may also play an important role in regulating soil C stores. Pasture roots probably have a higher turnover rate as many pasture species are annual and die off during the summer, whereas, many of the native species are perennial, and have woody roots that are longer lasting. Soil fertility is a major limitation to productivity of the native forests in Australia (Specht, 1996), and improvement of nutrient status through P fertilizer addition and N fixation by pasture legumes has markedly increased the productive capacity of these soils. Increasing agricultural productivity through fertilizer application has been found to increase C stores in many agricultural soils (Lal, 2001), but rates of accumulation were relatively modest in soils of the current study— the average increase in <5 mm soil C since pasture establishment was 8.4 Mg ha−1 in the 5–20 cm depth range, whilst the C:N ratios of the surface (0–10 cm)
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soil have decreased from an average from 32.3 to 17.6 (Mendham et al., 2002).
the increase may be relatively small (O’Connell et al., 2000).
4.2. Effect of afforestation on SOC stores 5. Conclusions Afforestation of crop land can reverse the decline in soil C that often occurs under intensive management (Brown and Lugo, 1990; Ihori et al., 1995), but establishment on pasture can lead to little change (Jug et al., 1999) or a marked decrease in soil C stocks (Turner and Lambert, 2000; Thuille et al., 2000). The plantations of this study had only been established for 7–10 years, but potentially large changes in the quantity and quality of C inputs and soil environment had occurred. Conditions for decomposition would be less favourable under plantation than under pasture, as eucalypt litter generally has a high C:N ratio, lignin and tannin content, and soils under eucalypt plantations generally have much lower moisture content (Grove et al., 2001). Despite large potential changes in C dynamics, plantation establishment had no detectable effect on soil C content at any depth range across the sites in this study. In a review of the literature on afforestation, Polglase et al. (2000) also found only a minor effect of afforestation on soil C stores across a wide range of environments. However, these findings contrasted with those of Turner and Lambert (2000) who found large losses of soil C under E. grandis plantation stands on ex-pasture in eastern Australia. Results from that study need to be treated with caution because they were based on an unreplicated chronosequence with doubtful matching of the sites (Polglase et al., 2000). There is potential for higher soil C to accumulate under plantations in the longer-term, as amounts of surface litter C stores were significant, and root C was significantly higher at most depths between 5 and 50 cm under plantation than pasture (Fig. 4). Root C under plantations may have higher potential contribution to soil C stores than under native vegetation because of the higher soil fertility induced by a period of pasture management. Harvesting operations would induce senescence of many of the roots associated with the plantation crop, possibly hastening their addition to the soil C pool. More research is required to examine these effects. Retention of residues following harvesting of plantations may also potentially increase surface soil C in the longer-term, but the magnitude of
Conversion of the native vegetation to pasture led to a significant mean increase of 8.4 Mg ha−1 in total soil (<5 mm) C stores in the 5–20 cm depth range, but this was offset by a mean reduction of 2 Mg ha−1 in the surface soil C (2–5 mm fraction), lower quantities of root C (mean of 5 Mg ha−1 ), and loss of surface litter layer C (mean of 9.8 Mg ha−1 ). Quantities of C in plantation soils after 7 years were not significantly different from those under pasture, but coarse root C was an average of 3.1 Mg ha−1 higher under plantation (consisting of both live and dead biomass), and the surface litter pool was also significant at 7.9 Mg ha−1 . Overall, change in land-use from native vegetation to pasture, or from pasture to plantation in south-western Australia had no significant effect on total soil C stocks, but rather resulted in a redistribution of C within the different fractions found in soil.
Acknowledgements This work was supported financially by the Australian Centre for International Agricultural Research, the RIRDC/LWRRDC/FWPRDC Joint Venture Agroforestry Program, the Western Australian Department of Resource Development, and WA Plantation Resources, Pty. Ltd. The authors wish to thank Stan Rance, Tuyen Pham, Shirley Snelling, George Wan and Paul Damon for their very able technical assistance, and also Geoff Dimmock for classifying and comparing the soil profiles.
References Attiwill, P.M., Guthrie, H.B., Leuning, R., 1978. Nutrient cycling in a Eucalyptus obliqua forest. I. Litter production and nutrient return. Aust. J. Bot. 26, 79–91. Balesdent, J., Chenu, C., Balabane, M., 2000. Relationship of soil organic matter dynamics to physical protection and tillage. Soil Till. Res. 53, 215–230. Bell, M.J., Harch, G.R., Bridge, B.J., 1995. Effects of continous cultivation on ferrosols in subtropical southeast Queensland. I.
D.S. Mendham et al. / Agriculture, Ecosystems and Environment 95 (2003) 143–156 Site characterization, crop yields and soil chemical status. Aust. J. Agric. Res. 46, 237–253. Berg, B., 2000. Litter decomposition and organic matter turnover in northern forest soils. For. Ecol. Manage. 133, 13–22. Brown, S., Lugo, A.E., 1982. The storage and production of organic matter in tropical forests and their role in the global carbon cycle. Biotropica 14, 161–187. Brown, S., Lugo, A.E., 1990. Effects of forest clearing and succession on the carbon and nitrogen content of soil in Puerto Rico and US Virgin Islands. Plant Soil 124, 53–64. Coventry, R.J., Fett, D.E.R., 1979. A Pipette and Sieve Method of Particle-Size Analysis and Some Observations on its Efficacy, CSIRO Division of Soils Divisional Report No. 38. CSIRO, Canberra. Davidson, E.A., Ackerman, I.L., 1993. Changes in soil carbon inventories following cultivation of previously untilled soils. Biogeochemistry 20, 161–193. Fahey, T.J., Hughes, J.W., 1994. Fine root dynamics in a northern hardwood forest ecosystem, Hubbard Brook Experimental Forest NH. J. Ecol. 82, 533–548. Farrington, P., Salama, R.B., Watson, G.D., Bartle, G.A., 1992. Water use of agricultural and native plants in a western Australian wheatbelt catchment. Agric. Water Manage. 22, 357– 367. Fearnside, P.M., Barbosa, R.I., 1998. Soil carbon changes from conversion of forest to pasture in Brazilian Amazonia. For. Ecol. Manage. 108, 147–166. Glaser, B., Turrion, M.B., Solomon, D., Ni, A., Zech, W., 2000. Soil organic matter quantity and quality in mountain soils of the Alay Range, Kyrgyzia, affected by land-use change. Biol. Fertil. Soils 31, 407–413. Golchin, A., Oades, J.M., Skjemstad, J.O., Clarke, P., 1994. Soil structure and carbon cycling. Aust. J. Soil Res. 32, 1043–1068. Grove, T.S., O’Connell, A.M., Mendham, D.S., Barrow, N.J., Rance, S.J., 2001. Sustaining the Productivity of Tree Crops on Agricultural Land in South-Western Australia, Publication No. 01/09. Rural Industries Research and Development Corporation (RIRDC), Canberra. Guggenberger, G., Zech, W., 1999. Soil organic matter composition under primary forest, pasture and secondary forest sucession, Region Huetar Norte, Costa Rica. For. Ecol. Manage. 124, 93– 104. Hingston, F.J., Turton, A.G., Dimmock, G.M., 1979. Nutrient distribution in Karri (Eucalyptus diversicolor F. Muell.) ecosystems in south-west western Australia. For. Ecol. Manage 2, 133–158. Hingston, F.J., Dimmock, G.M., Turton, A.G., 1981. Nutrient distribution in a Jarrah (Eucalyptus marginata Donn ex Sm.) ecosystem in south-west western Australia. For. Ecol. Manage 3, 183–207. Hingston, F.J., Galbraith, J.H., 1998. Application of the process-based model BIOMASS to Eucalyptus globulus ssp. globulus plantations on ex-farmland in south western Australia. II. Stemwood production and seasonal growth. For. Ecol. Manage. 106, 157–168. Ihori, T., Burke, I.C., Lauenroth, W.K., Coffin, D.P., 1995. Effects of cultivation and abandonment on soil organic matter in northeastern Colorado. Soil Sci. Soc. Am. J. 59, 1112–1119.
155
Jobbágy, E.G., Jackson, R.B., 2000. The vertical distribution of soil organic carbon and its relation to climate and vegetation. Ecol. Appl. 10, 423–436. Lal, R., 2001. World cropland soils as a source or sink for atmospheric carbon. Adv. Agron. 71, 145–191. Mann, L.K., 1986. Changes in soil carbon storage after cultivation. Soil Sci. 142, 279–288. Mendham, D.S., O’Connell, A.M., Grove, T.S., 2002. Organic matter characteristics under native forest, long-term pasture, and recent conversion to Eucalptyus plantations in Western Australia: microbial biomass, soil respiration, and permanganate oxidation. Aust. J. Soil Res. 40, 859–872. National Forest Inventory, 2000. National Plantation Inventory Tabular Report, March 2000. Bureau of Resource Sciences, Canberra. Neill, C., Melillo, J.J., Steudler, P.A., Cerri, C.C., de Moraes, J.F.L., Piccolo, M.C., Brito, M., 1997. Soil carbon and nitrogen stocks following forest clearing for pasture in the south-western Brazilian Amazon. Ecol. Appl. 7, 1216–1225. O’Connell, A.M., Grove, T.S., 1996. Biomass Production, Nutrient Uptake and Nutrient Cycling in the Jarrah (Eucalyptus marginata) and Karri (Eucalyptus diversicolor) forests of south-western Australia. In: Attiwill, P.M., Adams, M.A. (Eds.), Nutrition of Eucalypts. CSIRO Publishing, Melbourne, pp. 155–189. O’Connell, A.M., Grove, T.S., Mendham, D., Rance, S.J., 2000. Effects of site management in eucalypt plantations in south-western Australia. In: Nambiar, E.K.S., Tiarks, A., Cossalter, C., Ranger, J. (Eds.), Site Management and Productivity in Tropical Plantation Forests: A Progress Report. Center for International Forestry Research, Bogor, Indonesia, pp. 61–71. Polglase, P.J., Paul, K.I., Khanna, P.K., Nyakuengama, J.G., O’Connell, A.M., Grove, T.S., Battaglia, M., 2000. Change in Soil Carbon Following Afforestation or Reforestation, Technical Report No. 20. National Carbon Accounting System, Australian Greenhouse Office, Canberra. Post, W.M., Kwon, K.C., 2000. Soil carbon sequestration and land-use change: processes and potential. Global Change Biol. 6, 317–327. Rayment, G.E., Higginson, F.R., 1992. Australian Laboratory Handbook of Soil and Water Chemical Methods. Inkata Press, Melbourne. Reiners, W.A., Bouwman, A.F., Parsons, W.F.J., Keller, M., 1994. Tropical rain forest conversion to pasture: changes in vegetation and soil properties. Ecol. Appl. 4, 363–377. Rhoades, C.C., Eckert, G.E., Coleman, D.C., 2000. Soil carbon differences among forest, agriculture, and secondary vegetation in lower montane Ecuador. Ecol. Appl. 10, 497–505. Rossiter, R.C., Ozanne, P.G., 1970. South-western temperate forests, woodlands, and heaths. In: Moore, R.M. (Ed.), Australian Grasslands. Australian National University Press, Canberra, pp. 199–218. Rovira, A.D., 1992. Dryland mediterranean farming systems in Australia. Aust. J. Exp. Agric. 32, 801–809. Schlesinger, W.H., 1977. Carbon balance in terrestrial detritus. Annu. Rev. Ecol. Syst. 8, 51–81.
156
D.S. Mendham et al. / Agriculture, Ecosystems and Environment 95 (2003) 143–156
Silver, W.L., Neff, J., Mcgroddy, M., Veldkamp, E., Keller, M., Cosme, R., 2000. Effects of soil texture on below-ground carbon and nutrient storage in a lowland Amazonian forest ecosystem. Ecosystem 3, 193–209. Specht, R.L., 1996. The influence of soils on the evolution of the eucalypts. In: Attiwill, P.M., Adams, M.A. (Eds.), Nutrition of Eucalypts. CSIRO Publishing, Melbourne, pp. 31–60. Tate, K.R., 1992. Assessment, based on a climosequence of soils in tussock grasslands, of soil carbon storage and release in response to global warming. J. Soil Sci. 43, 697–707. Thomson, C.J., Ewing, M.A., Turner, N.C., Revell, C.K., Le Coultre, I.F., 1998. Influence of rotation and time of germinating rains on the productivity and composition of annual pastures in western Australia. Aust. J. Agric. Res. 49, 225–232. Thuille, A., Buchmann, N., Schulze, E.D., 2000. Carbon stocks and soil respiration rates during deforestation, grassland-use and subsequent Norway spruce afforestation in the southern Alps, Italy. Tree Physiol. 20 (13), 849–857.
Turner, J., Lambert, M., 2000. Change in organic carbon in forest plantation soils in eastern Australia. For. Ecol. Manage. 133, 231–247. Underwood E.J., Gladstones J.S., 1979. Subterranean clover and other legumes. In: Burvill, G.H. (Ed.), Agriculture in Western Australia 1829–1979. University of Western Australia Press, Nedlands, Western Australia, pp. 139–156. Van Cleve K., Powers R.F., 1995. Soil carbon, soil formation, and ecosystem development. In: McFee, W.W., Kelly, J.M. (Eds.), Carbon Forms and Functions in Forest Soils. Soil Science Society of America Inc., MA, USA. Veldkamp, E., 1994. Organic carbon turnover in three tropical soils under pasture after deforestation. Soil Sci. Soc. Am. J. 58, 175–180. Weaver, D.M., Reed, A.E.G., 1998. Patterns of nutrient status and fertiliser practice on soils of the south coast of western Australia. Agric. Ecosyst. Environ. 67, 37–53.
Change in soil carbon after land clearing or afforestation in highly weathered lateritic and sandy soils of south-western Australia D.S. Mendham∗ , A.M. O’Connell, T.S. Grove CSIRO Forestry and Forest Products, Private Bag No. 5, Wembley, WA 6913, Australia Received 29 November 2001; received in revised form 15 May 2002; accepted 7 June 2002
Abstract Soil organic carbon (SOC) is a significant component of the worlds terrestrial carbon stocks, and changes in land-use have potential to change pools of soil C. Land-use change is occurring over large areas of the world, so soil has potential to be a large source or sink for atmospheric C. Changes in amounts of soil C (0–1 m) were examined after clearing of native vegetation for pasture (20–71 years prior to study), and after establishment of Eucalyptus globulus plantations on ex-pasture sites (7–10 years prior to study). A suite of 10 sites in south western Australia were used in a three-way comparison between native vegetation, pasture and plantation. Soil types across the 10 sites included Acrisols, Arenosols, and a Ferrosol. Soil C was measured in the fine earth (<2 mm), the 2–5 mm soil fraction, in charcoal and roots retained on a 5 mm screen, and in surface litter. Differences in soil bulk density between land-uses were accounted for by calculating depths for equivalent weights of soil. Despite large increases in soil fertility with conversion to pasture, amounts of soil C changed little. The amount of C in the 5–20 cm increment was significantly greater under pasture than native vegetation (mean of 8.4 Mg ha−1 , <5 mm), but in surface soil (0–5 cm) and <20 cm, there were no significant differences in soil C content between land-uses. Less C in soil at 5–20 cm under native vegetation was offset by significantly more C in coarse roots (average of 5 Mg ha−1 higher to 1 m depth), surface soil (2–5 mm fraction, 2 Mg ha−1 ), and in standing pools of surface litter (9.8 Mg ha−1 ). Amounts of soil C under plantation were not significantly different from pasture 7–10 years after plantation establishment. However, plantation soils had an average of 3.1 Mg ha−1 more C in coarse roots than pasture, and significant quantities of C in surface litter (average of 7.9 Mg ha−1 ). Overall, soil C content in the sites of this study was relatively stable and the effect of land-use change was limited. © 2002 Elsevier Science B.V. All rights reserved. Keywords: Soil organic carbon; Root C; Litter C; Land-use change; Native vegetation; Pasture; Plantations; Eucalyptus globulus; Western Australia
1. Introduction The amount of soil organic carbon (SOC) globally has been estimated to contain more than three times the atmospheric pool, and more than four times the biotic pool of C (Lal, 2001). Changes in amount of ∗ Corresponding author. Tel: +61-8-9333-6663; fax: +61-8-9387-8991. E-mail address: [email protected] (D.S. Mendham).
SOC with change in land-use have been the subject of significant attention recently because of widespread recognition of the need to limit emissions of greenhouse gases. Change in land-use has potential to either release or sequester soil C. Clearing of native vegetation for agricultural production often leads to a reduction in SOC (Mann, 1986; Davidson and Ackerman, 1993), whilst potential exists for significant increase in SOC through changed management practices (Lal, 2001) or reafforestation (Post and Kwon, 2000).
0167-8809/02/$ – see front matter © 2002 Elsevier Science B.V. All rights reserved. PII: S 0 1 6 7 - 8 8 0 9 ( 0 2 ) 0 0 1 0 5 - 6
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The rate of accumulation or loss of SOC after change in land-use is governed by the balance between addition and decomposition of organic material, which in turn is influenced by changes in net primary productivity and allocation (Van Cleve and Powers, 1995), quality of organic inputs (Berg, 2000), tillage practices (Balesdent et al., 2000), rooting patterns (Jobbágy and Jackson, 2000) and climate. Other site factors also influence SOC dynamics, including soil texture (Golchin et al., 1994) and climate (Van Cleve and Powers, 1995). Much of the arable land in Australia has been cleared of native vegetation in the last 20–100 years, resulting in increased soil fertility through additions of nutrients from chemical fertilizers (Weaver and Reed, 1998), fixation of atmospheric N by pasture legumes (Underwood and Gladstones, 1979), and changes in the quantity and quality of organic matter inputs. Cycling of SOC under managed pasture in south western Australia is potentially quite different from under native vegetation due to higher fertility, shallower rooting patterns of pasture species, input of residues with higher chemical quality (e.g. lower C:N ratio), and annual pasture species versus perennial forest species. Another significant change in land-use recently in south-western Australia is conversion of pasture land to eucalypt plantation, predominantly Eucalyptus globulus, which can provide economic and environmental benefits to farmers. The area under E. globulus in south-western Australia has been expanding by more than 25,000 ha per year (National Forest Inventory, 2000). Cycling of SOC under plantation is also potentially different to that under pasture, with potential differences again in net primary productivity and allocation, rates of root and litter turnover, residue quality, and microclimatic conditions. Detailed studies of C inputs have not been conducted in these ecosystems, but productivity (in terms of aboveground biomass accumulation and litterfall) is up to 3–4 Mg ha−1 per year in native Jarrah forest (O’Connell and Grove, 1996), 4.5–8 Mg ha−1 per year in annual pastures (Rossiter and Ozanne, 1970; Thomson et al., 1998), and 11–43 Mg ha−1 per year in E. globulus plantations (Hingston and Galbraith, 1998). Soil bulk density can also be modified by change in land-use, which affects the calculations of SOC stocks (Fearnside and Barbosa, 1998). Differences in soil bulk density between land-uses can be caused by
compaction from livestock and machinery, differences in volume and turnover of root material, tillage, and biological activity. Soils under pasture typically have higher bulk density, due to compaction caused by agricultural management practices, and lower volumes of coarse roots. Any effect of changed land-use on soil bulk density will influence the amount of soil that is sampled from within a fixed depth. The purpose of this study was to assess impacts of land-use on changes in total soil C to 1 m depth at 10 sites representing the climatic and soil gradient in south-western Australia. Land-uses studied were native vegetation, pasture managed for 20–71 years after clearing, and E. globulus plantations established on pasture land for 7–10 years prior to the study. The 10 sites were selected such that the soils would have been closely matched prior to change in land-use. 2. Materials and methods 2.1. Site selection and soil sampling Soil was sampled from 10 sites representing the range in soils and climates where E. globulus is grown in south-western Australia (Fig. 1). Selected soil characteristics are shown in Table 1. Total C in the surface soil at the sites ranged from 16.7 to 90.0 g kg−1 . Adjacent pasture (20–71 year old) and plantation (7–10 year old) plots were selected at each site, and a plot on a nearby native vegetation remnant on the same landform element was also selected. The native vegetation at each site was predominantly Jarrah (E. marginata) and Marri (E. calophylla) forest, which ranges in productivity, depending on rainfall. The Anning site was originally occupied by Karri (E. diversicolor) forest. The pasture consisted of a mix of annual species, including subterranean clover (Trifolium subterraneum), capeweed (Arctotheca calendula), and a range of grasses (including Lolium rigidum, Hordeum leporinum, and Bromus diandrus). Tillage operations have been relatively rare in these systems, occurring only with winter cropping approximately every 5–10 years. Pasture soil is enriched in nutrients compared to the native soil (see Table 1) because of inputs of N (from clover) and P (from annual applications of superphosphate). Potassium has been applied to some pastures, and trace elements have also
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Fig. 1. Map of south-western Australia showing locations of the 10 study sites (diamonds). The towns of Albany, Manjimup and Perth (circles) are also shown for reference.
been applied at irregular intervals. For more detail on pasture production in this region, see Rovira (1992). Plantations of E. globulus are being established on pasture sites. Seedlings are planted into rip-lines, and generally fertilized once during establishment with N and P with a spot application close to the base of the seedling. The large majority of E. globulus plantations in south-western Australia (including those in this study) are in their first rotation, with average rotation lengths expected to be about 10 years. Plantations in this study were approaching the end of their first rotation. The validity of comparisons between land-uses was qualitatively assessed by comparing pedological features (including horizon depths, color, texture, gravel type and content) of soil profiles in each of the three land-use types at each site. Soil was sampled within each land-use from a 40 m transect located at least 20 m from the boundary of the other land-uses to minimize boundary effects such as shading, tree roots entering the pasture or fertilizer spread into the native vegetation. Each 40 m transect was divided into 4 × 10 m plots, from which soil was collected using thin-walled stainless steel corers of different internal diameters (47, 59, and 70.5 mm i.d., depending on depth). Three soil cores per plot were
collected and bulked for the 0–5 and 5–10 cm depth increments, whilst single cores per plot were collected at depths greater than 10 cm. For depths greater than 10 cm, the samples were collected sequentially from the same hole within each plot, in increments of 10–20, 20–30, 30–50, 50–70, and 70–100 cm. A flange-type plastic sheath was inserted at the top to prevent surface soil and litter from falling into the hole and contaminating deeper soil samples. Additionally, smaller diameter cores were used with depth to prevent upper layers of soil falling into, and contaminating, the deeper horizon samples. All sampling in the plantation was done away from the rip-lines. Litter samples were collected in a circular sampling area (250 mm diameter). Between 1 and 4 litter samples were collected per plot, depending on the spatial heterogeneity of the litter layer, which was qualitatively assessed at each site. Litter was not collected in the pasture because there were negligible amounts due to grazing. 2.2. Soil preparation and analysis Soil and litter samples were sieved through a 5 mm sieve to separate the coarse and fine fractions. Coarse
146
Site name
Field texture
Soil classification (FAO, 1990)
Annual rainfall (mm)
Pasture age (years)
Plantation age (years)
Silt + clay proportion (%)a
pH
Total C (g kg−1 )b
Total N (g kg−1 )b
Total P (g kg−1 )b
Andrews Anning Ayers Gibbs Hall B Jefferies B Moltoni Patmore B Robinson Walker
Loamy fine sand Gravelly clayey sand Gravelly sandy clay loam Grey sand Coarse sand Sand Gravelly sandy loam Gravelly sandy loam Gravelly sandy loam Loamy fine sand
Ferric Acrisol Ferric Acrisol Xanthic Ferralsol Haplic Arenosol Haplic Arenosol Haplic Arenosol Ferric Acrisol Haplic Arenosol Ferric Acrisol Haplic Arenosol
865 1291 983 721 661 655 1358 833 1226 574
20 71 29 31 43 31 28 20 42 23
10 8 9 9 8 8 10 7 10 8
8.41 12.7 23.9 4.97 7.10 3.94 18.8 11.3 28.9 8.41
5.1 6.0 5.9 5.7 5.6 6.0 5.9 6.3 5.8 4.9
42.6 48.4 90.0 18.0 30.1 16.7 55.5 54.6 77.2 19.1
1.62 1.90 3.59 0.37 1.34 0.35 1.77 1.78 2.60 0.40
0.043 0.095 0.239 0.009 0.036 0.007 0.112 0.149 0.125 0.030
a b
(−3.7) (+4.8) (−12.5) (+78.6) (+12.8) (+18.2) (−12.6) (+0.4) (−3.4) (+15.9)
Depth: 0–10 cm. Chemical data shown for 0–10 cm soil under native vegetation with mean percentage difference in managed soils shown in parentheses.
(+55) (+73) (+16) (+344) (+54) (+137) (+54) (+37) (+68) (+150)
(+285) (+511) (+97) (+965) (+289) (+1601) (+450) (+87) (+396) (+133)
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Table 1 Characteristics of each of the sites examined in south-western Australia
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roots, charcoal and gravel fractions (i.e. those retained on the 5 mm sieve) were separated by hand. Large lateral or tap roots were not sampled. The fine soil was further separated by sieving through 2 mm, resulting in size fractions of 2–5 mm and <2 mm. The 2–5 mm soil fraction was measured separately because the native vegetation and plantation species generally have coarser roots than the pasture species, so the standard <2 mm fraction can potentially bias measurements towards finer rooted species. All fractions were oven dried, weighed, ground and analysed for total C by combustion at 1200 ◦ C and measurement of the evolved CO2 on a LECO TOC analyser (CR-412). Corrections were not made for carbonate as all soils were acidic (pH <7), and short combustion times indicated that CO2 was generated from organic, rather than inorganic sources. Soil (<2 mm) was bulked across plots for each depth, and assessed for particle size distribution and pH. The proportion of sand, silt and clay in the samples was analysed using the pipette method of Coventry and Fett (1979) after removal of the organic material by reacting with H2 O2 and dispersion by ultrasonic treatment. Soil pH was assessed after shaking for 1 h in a 1:5 soil:water suspension. Total N and P were measured on surface soil samples (0–10 cm) using a flow injection analyser to assess concentration of N and P in solution after a Kjeldahl digest (Rayment and Higginson, 1992).
land-uses in a stepwise manner. Different parameters were used between land-uses where the regressions were significantly improved by doing so. The distribution of C with depth in each profile generally fitted an exponential regression of the form
2.3. Calculations of soil carbon
2.4. Statistical analysis
Curve-fitting was used for (a) calculating the soil depth intervals for each of the managed land-uses (because changed bulk density influences the depth of soil for a constant weight), and (b) predicting the carbon content within the depth intervals calculated during the first step. Nonlinear regressions of the following form were fitted to depth trends in soil bulk density
Regression analysis was used to relate soil C to profile depth, soil texture and climatic variables using the Genstat statistical package (Lawes Agricultural Trust). Analysis of variance (ANOVA; Genstat, Lawes Agricultural Trust) was used to assess the effect of change in land-use on soil C in the different profile depths. Sites were used as replicates in the ANOVA. There were some differences in soil texture between land-uses at each site, despite efforts to ensure that sites were as closely matched as possible. Hence ANOVA analyses were conducted on total C data only after ensuring that there was no systematic bias in soil texture between the land-uses. The clay + silt content was used as a covariate in the ANOVA of soil C to help account for minor differences caused by soil texture variation between land-uses at each site.
y = a + brx + cx,
(1)
where y was the soil bulk density (<5 mm fraction), x the depth, and a, b, c and r were fitted parameters. Initially a single model was fitted to the data from each site, but the effect of land-use on the soil bulk density trend with depth was evaluated by assessing significance of applying separate constant (a), linear (b, c), and nonlinear (r) parameters to each of the
z = d + esx ,
(2)
where z was the quantity of C calculated per cm of depth (Mg ha−1 cm−1 ), x the depth (cm), and d, e and s were fitted parameters. Effects of land-use on soil C distribution were evaluated in a similar manner to that for soil bulk density, i.e. by assessing the significance of changing the constant (d), linear (e), and nonlinear (s) parameters in a stepwise manner. Different parameters were used between land-uses to calculate amounts of soil C under the different land-uses where regressions were significantly improved by doing so. Weights of soil for fixed depth increments (0–5, 5–10, 10–20, 20–30, 30–50, 50–70, and 70–100 cm) under native vegetation at each site were calculated according to Eq. (1), and depths under pasture and plantation were calculated for an equivalent weight of soil under native vegetation. The calculated depths for equivalent weights of soil in the pasture and plantation ecosystems are shown in Table 2. The total amount of C within each of the corrected depths was calculated by integrating Eq. (2).
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Table 2 Calculated depths (cm) for equivalent weight of soil under native vegetation Site name
Original depth Pasture
Andrews Anning Ayers Gibbs Hall Ba Jefferies B Moltoni Patmore B Robinson Walker a
Plantation
5
10
20
30
50
70
100
5
10
20
30
50
70
100
3.5 2.7 5 6.6 4.5 4.5 2.5 5 2.1 5.2
8 7.3 10 11.7 9.1 9.2 7.4 10 6.8 9.1
17.2 16.6 20 21.8 18.3 18.5 15.9 20 17 17.9
26.5 25.9 30 32 27.4 27.8 24.3 30 27.4 27.3
45.5 44.3 50 52.2 45.8 46.4 40.8 50 48.2 46.7
65 62.8 70 72.5
95.2 90.5 100 102.9
18.3 18 20 19.9 17.5 18.3 17.3 20 17.2 17.7
27.7 27.7 30 29.9 26.2 27.5 25.7 30 26.6 27
47 47 50 50.1 43.8 45.9 42.5 50 45.3 46
97.9 95.2 100 100.5
92.4 82.2 100 100.2 95.6
9.1 8.4 10 9.8 8.7 9.1 8.9 10 7.8 8.8
66.9 66.3 70 70.3
64.8 57.4 70 69 66.2
4.4 3.5 5 4.8 4.3 4.4 4.7 5 3.1 4.4
64.1 59.3 70 63.9 65.1
91.3 84.5 100 91.9 93.8
Soil not sampled depths greater than 50 cm at Hall B site due to presence of a hard pan.
Multiple linear regression analysis (Genstat, Lawes Agricultural Trust) was used to identify the variables influencing profile C stores across the 10 sites. Variables examined included soil texture and environmental attributes such as temperature (mean annual and mean seasonal), summer evaporation, and annual precipitation.
3. Results 3.1. Influence of land-use change on total soil C Land-use had no significant effect on mean amounts of total C in the <2 mm fraction in the surface horizon (0–5 cm) and depths greater than 20 cm (Fig. 2).
Fig. 2. Amounts of soil C in the <2 mm soil fraction. Significant differences between land-uses within each depth range are denoted by ∗: P < 0.05.
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Fig. 3. Amounts of soil C in the 2–5 mm soil fraction. Significant differences between land-uses within each depth range are denoted by ∗: P < 0.05.
However, the soil under native vegetation had significantly less total soil C in the <2 mm fraction in the 5–10 cm increment (P = 0.044), and also in the 10–20 cm increment when tested at the less stringent 0.1 level (P = 0.071). Total C under native vegetation in the 5–20 cm depth range was 6.3 and 5.6 Mg ha−1 lower than that under pasture and plantation, respectively. The average amount of total C in the 2–5 mm fraction ranged from 0.4 to 4.5 Mg ha−1 for the different depths (Fig. 3), with significantly more C under native vegetation than under pasture or plantation only at the 0–5 cm depth. Compared to the 0–100 cm increment, the mean proportion of total soil C in the 0–10 cm depth was 39.1–39.7%, and was not significantly different between land-uses (Table 3). Amounts of C in the 10–30 cm increment as a proportion of the total (0–100 cm) was less under the native vegetation than in the managed treatments, but was balanced by a higher proportion (although not significant) in the 30–100 cm depth range.
C. Coarse roots were more prevalent under the pasture than other land-uses (although not significantly) in surface soil (0–5 cm depth, Fig. 4), but C in pasture roots depths greater than 5 cm depth was significantly less than under native vegetation. There was more root C under plantations than pasture at most depths, significantly between 5 and 50 cm depth. 3.3. Total C in litter There was no surface litter under pasture because of grazing. A relatively small proportion of the plant material is returned in animal waste, but this is spatially heterogeneous and was not assessed Table 3 Proportion of total soil C (%, <5 mm fraction) to 1 m depth in different depth (cm) ranges Depth range
vegetationa
0–10
10–30
30–100
39.3 a 39.7 a 39.1 a
24.5 a 30.5 b 32.0 b
36.2 a 29.8 a 29.0 a
3.2. Total C in roots
Native Pasturea Plantationa
Differences in amounts of root C under different land-uses were more pronounced than with total
a Values with the same letters within each column were not significantly different (tested by L.S.D., α = 5%).
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Fig. 4. Amounts of soil C in the coarse roots (retained on a 5 mm screen). Significant differences between land-uses within each depth range were denoted by ∗: P < 0.1 and ∗∗∗: P < 0.001.
in this study. There were significant amounts of litter under native vegetation (mean of 9.8 Mg ha−1 , Fig. 5a) and plantation (7.18 Mg ha−1 ). The environmental variable most closely associated with the quantity of litter C was long-term annual rainfall (Fig. 5b, y = 0.0172x − 6.95, R 2 = 0.38, S.E. = 6.29, n = 20). Other variables (summer evaporation, soil texture, monthly temperatures) and combinations of these variables (in multiple linear regressions) across both land-uses did not improve on that relationship.
3.4. Total pools of C to 1 m Amounts of total soil C to 1 m ranged from 50 to 163 Mg ha−1 (Table 4) across the 10 sites, with most of the C being in the fine earth (<2 mm) fraction (30–144 Mg ha−1 ). There was less C in the fine earth under native vegetation than under pastures and plantations, but this was offset by higher soil C in the 2–5 mm fraction, and in the coarse roots. There were similar amounts of litter under native vegetation and plantation. Overall, land-use change had no effect
Table 4 Amounts and range of total C (Mg ha−1 ) in the soil, litter, and coarse roots across the 10 sites Native vegetation
Pasture
Plantation
Mean
Range
Mean
Range
Soil Soil 2–5 mma Roots >5 mma Surface littera
61.4 a 16.4 a 6.35 c 9.8 a
30.0–117.0 0–40.9 3.80–13.01 0.7–34.5
78.0 b 11.9 a 1.38 a –
41.1–143.7 0–18.3 0.48–2.96 –
Totala
94.0 a
53.7–143.1
91.3 a
49.6–162.5
<2 mma
a
Means with the same letter within each row were not significantly different.
Mean 78.4 b 12.1 a 4.53 b 7.9 a 102.9 a
Range 41.3–116.5 2.4–25.4 1.63–8.97 1.8–13.4 59.1–144.1
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Fig. 5. Amounts of litter C under the plantation and native vegetation (a) at each of the sites, and (b) relationship with average annual rainfall across the sites.
on pools of total C to 1 m depth. Charcoal that was hand-separated from the >5 mm soil fraction was not included in the above balance, but this fraction added a mean of 3.98 Mg C ha−1 to the total pool across the sites. Quantities of charcoal C were not significantly different between land-uses. Amounts of C were related to the quotient of mean temperature and precipitation (Fig. 6a, R 2 = 0.29, S.E. = 23.0) used by Brown and Lugo (1982) to predict soil C stocks across a climatic gradient. However, the site factor most highly correlated with total C across sites was soil texture (represented by the clay + silt content, Fig. 6b, R 2 = 0.32, S.E. = 22.6). Incorporation of soil texture with summer evaporation and summer mean daily minimum temperature in a multiple linear regression gave a more significant relationship than with the clay + silt content alone (Fig. 6c, R 2 = 0.50, S.E. = 19.3). There were no significant differences between land-use types for any of these relationships.
4. Discussion Amounts of soil C to 1 m in the soils of this study (50–163 Mg ha−1 ) were less than those published previously for Jarrah (184 Mg ha−1 , Hingston et al., 1981) and Karri forest (177–178 Mg ha−1 , Hingston et al., 1979) in south-western Australia, but the three sites in the studies of Hingston et al. (1979, 1981) had relatively high rainfall and productivity. Sites with lower rainfall (which also generally had lower clay content and higher evaporation) generally corresponded to lower profile C contents (Fig. 6c), although there was a relatively high degree of scatter in that relationship. The range found in this study was similar to those published for Brazilian Amazonia (94–113 Mg ha−1 , Fearnside and Barbosa, 1998; Silver et al., 2000), despite a significantly drier climatic regime in south-western Australia. Jobbágy and Jackson (2000) also found a similar range of soil C stored in a range of ecosystems from
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Fig. 6. Relationship between total soil C (<5 mm fraction) and (a) the quotient of mean annual temperature and annual precipitation, (b) the clay + silt content to 1 m, and (c) a function of the clay + silt content 0–100 cm (c), mean summer evaporation (e), and summer mean daily minimum temperature (T).
deserts (62 Mg ha−1 ) to tropical evergreen forests (186 Mg ha−1 ). However, amounts were much less than those under the E. grandis plantations of eastern Australia (200–400 Mg ha−1 in the top 50 cm) studied by Turner and Lambert (2000). The significant relationship between soil C and the quotient of temperature and precipitation (Fig. 6a) followed a similar trend
to that of Brown and Lugo (1982) and Tate (1992), but the amount of soil C at any given climatic ratio was generally higher than found by Tate (1992). The temperature/precipitation ratio was mainly affected by rainfall across the sites in this study because mean annual temperatures across the sites varied little (between 14.8 and 15.5 ◦ C). Soil texture was more closely
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related to soil C stocks than climatic factors (Fig. 6b), but it was not possible to separate the confounding effects of climate and soil texture across the sites, because rainfall was correlated with soil clay content (R 2 = 0.47). Whilst mean annual temperatures were similar across the sites, the daily minima and maxima varied, depending on distance from the sea. Incorporating mean daily summer minimum temperature as a climatic variable improved the regression in Fig. 6c, such that 50.4% of the variation in the amount of soil C (0–100 cm) was described by those three factors. A significant weakness of many studies investigating the effect of changed land-use on soil properties is the reliance on too few sites in the paired approach. This can lead to treatment effects being attributed to the fine-scale differences in soil properties that occur naturally (Fearnside and Barbosa, 1998). Hence the effect of land-use on soil properties is potentially confounded by differences in the soils. To effectively test the effect of land-use change with the paired site approach is only possible if enough sites are sampled to be used as replicates, or that regression analyses can be conducted using independent site properties. The use of 10 well-matched sites in this study was a sound basis for generalizing the effects of land-use change on soil C stores in south western Australia. The three-way comparison of this study allowed comparison between two different land-use changes—from native vegetation to pasture (20–71 years prior to study), and from pasture to plantation (7–10 years prior to study). 4.1. Effect of land clearing on soil C Many studies have shown that clearing of native vegetation, followed by intensive management with regular cultivation often causes large reductions in soil C (Mann, 1986; Davidson and Ackerman, 1993), but conversion to less intensive management with infrequent tillage (such as pasture) has variable effects, ranging from a reduction in total C (Veldkamp, 1994; Guggenberger and Zech, 1999; Glaser et al., 2000; Rhoades et al., 2000), through unchanged amounts (Reiners et al., 1994; Veldkamp, 1994; Bell et al., 1995), to increased amounts of total C (Neill et al., 1997; Fearnside and Barbosa, 1998). Amounts of soil C in this study were significantly higher in 5–20 cm soil under managed land-uses, but some of this C was offset by lower amounts in the roots, litter and
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2–5 mm fraction under pasture (0–5 cm depth). The amount of C in the roots and litter may be more variable on a seasonal basis than soil C (Attiwill et al., 1978; Fahey and Hughes, 1994), but the values found at a single sampling time were useful for comparing likely magnitudes across different land-uses. Soils were sampled in late spring, after the main growing season, so the quantities of roots and litter were probably about average on an annual basis. If the soils had been sampled during the main growing season, the quantity of carbon in fine roots under pasture would probably have been greater, although the roots would still have been highly stratified, and mainly concentrated in the surface soil (0–5 cm). Changes in amounts of soil C content with land-use are caused by a disruption of the balance between inputs from plant litter (above- and below-ground) and losses from decomposition, leaching and erosion (Schlesinger, 1977). Conversion to pasture in these systems has led to marked increases in soil fertility (Mendham et al., 2002) and higher soil moisture content for much of the year (Farrington et al., 1992), so there is potential for higher decomposition rates under pasture. However, the observation that soil C increased under pasture (in the 5–20 cm depth range) suggested that C additions to that depth must also be significantly greater than those under native vegetation. Addition of C to the lower depths may be attributed to mixing of surface layers during cultivation, and/or through higher net primary productivity. Root dynamics may also play an important role in regulating soil C stores. Pasture roots probably have a higher turnover rate as many pasture species are annual and die off during the summer, whereas, many of the native species are perennial, and have woody roots that are longer lasting. Soil fertility is a major limitation to productivity of the native forests in Australia (Specht, 1996), and improvement of nutrient status through P fertilizer addition and N fixation by pasture legumes has markedly increased the productive capacity of these soils. Increasing agricultural productivity through fertilizer application has been found to increase C stores in many agricultural soils (Lal, 2001), but rates of accumulation were relatively modest in soils of the current study— the average increase in <5 mm soil C since pasture establishment was 8.4 Mg ha−1 in the 5–20 cm depth range, whilst the C:N ratios of the surface (0–10 cm)
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soil have decreased from an average from 32.3 to 17.6 (Mendham et al., 2002).
the increase may be relatively small (O’Connell et al., 2000).
4.2. Effect of afforestation on SOC stores 5. Conclusions Afforestation of crop land can reverse the decline in soil C that often occurs under intensive management (Brown and Lugo, 1990; Ihori et al., 1995), but establishment on pasture can lead to little change (Jug et al., 1999) or a marked decrease in soil C stocks (Turner and Lambert, 2000; Thuille et al., 2000). The plantations of this study had only been established for 7–10 years, but potentially large changes in the quantity and quality of C inputs and soil environment had occurred. Conditions for decomposition would be less favourable under plantation than under pasture, as eucalypt litter generally has a high C:N ratio, lignin and tannin content, and soils under eucalypt plantations generally have much lower moisture content (Grove et al., 2001). Despite large potential changes in C dynamics, plantation establishment had no detectable effect on soil C content at any depth range across the sites in this study. In a review of the literature on afforestation, Polglase et al. (2000) also found only a minor effect of afforestation on soil C stores across a wide range of environments. However, these findings contrasted with those of Turner and Lambert (2000) who found large losses of soil C under E. grandis plantation stands on ex-pasture in eastern Australia. Results from that study need to be treated with caution because they were based on an unreplicated chronosequence with doubtful matching of the sites (Polglase et al., 2000). There is potential for higher soil C to accumulate under plantations in the longer-term, as amounts of surface litter C stores were significant, and root C was significantly higher at most depths between 5 and 50 cm under plantation than pasture (Fig. 4). Root C under plantations may have higher potential contribution to soil C stores than under native vegetation because of the higher soil fertility induced by a period of pasture management. Harvesting operations would induce senescence of many of the roots associated with the plantation crop, possibly hastening their addition to the soil C pool. More research is required to examine these effects. Retention of residues following harvesting of plantations may also potentially increase surface soil C in the longer-term, but the magnitude of
Conversion of the native vegetation to pasture led to a significant mean increase of 8.4 Mg ha−1 in total soil (<5 mm) C stores in the 5–20 cm depth range, but this was offset by a mean reduction of 2 Mg ha−1 in the surface soil C (2–5 mm fraction), lower quantities of root C (mean of 5 Mg ha−1 ), and loss of surface litter layer C (mean of 9.8 Mg ha−1 ). Quantities of C in plantation soils after 7 years were not significantly different from those under pasture, but coarse root C was an average of 3.1 Mg ha−1 higher under plantation (consisting of both live and dead biomass), and the surface litter pool was also significant at 7.9 Mg ha−1 . Overall, change in land-use from native vegetation to pasture, or from pasture to plantation in south-western Australia had no significant effect on total soil C stocks, but rather resulted in a redistribution of C within the different fractions found in soil.
Acknowledgements This work was supported financially by the Australian Centre for International Agricultural Research, the RIRDC/LWRRDC/FWPRDC Joint Venture Agroforestry Program, the Western Australian Department of Resource Development, and WA Plantation Resources, Pty. Ltd. The authors wish to thank Stan Rance, Tuyen Pham, Shirley Snelling, George Wan and Paul Damon for their very able technical assistance, and also Geoff Dimmock for classifying and comparing the soil profiles.
References Attiwill, P.M., Guthrie, H.B., Leuning, R., 1978. Nutrient cycling in a Eucalyptus obliqua forest. I. Litter production and nutrient return. Aust. J. Bot. 26, 79–91. Balesdent, J., Chenu, C., Balabane, M., 2000. Relationship of soil organic matter dynamics to physical protection and tillage. Soil Till. Res. 53, 215–230. Bell, M.J., Harch, G.R., Bridge, B.J., 1995. Effects of continous cultivation on ferrosols in subtropical southeast Queensland. I.
D.S. Mendham et al. / Agriculture, Ecosystems and Environment 95 (2003) 143–156 Site characterization, crop yields and soil chemical status. Aust. J. Agric. Res. 46, 237–253. Berg, B., 2000. Litter decomposition and organic matter turnover in northern forest soils. For. Ecol. Manage. 133, 13–22. Brown, S., Lugo, A.E., 1982. The storage and production of organic matter in tropical forests and their role in the global carbon cycle. Biotropica 14, 161–187. Brown, S., Lugo, A.E., 1990. Effects of forest clearing and succession on the carbon and nitrogen content of soil in Puerto Rico and US Virgin Islands. Plant Soil 124, 53–64. Coventry, R.J., Fett, D.E.R., 1979. A Pipette and Sieve Method of Particle-Size Analysis and Some Observations on its Efficacy, CSIRO Division of Soils Divisional Report No. 38. CSIRO, Canberra. Davidson, E.A., Ackerman, I.L., 1993. Changes in soil carbon inventories following cultivation of previously untilled soils. Biogeochemistry 20, 161–193. Fahey, T.J., Hughes, J.W., 1994. Fine root dynamics in a northern hardwood forest ecosystem, Hubbard Brook Experimental Forest NH. J. Ecol. 82, 533–548. Farrington, P., Salama, R.B., Watson, G.D., Bartle, G.A., 1992. Water use of agricultural and native plants in a western Australian wheatbelt catchment. Agric. Water Manage. 22, 357– 367. Fearnside, P.M., Barbosa, R.I., 1998. Soil carbon changes from conversion of forest to pasture in Brazilian Amazonia. For. Ecol. Manage. 108, 147–166. Glaser, B., Turrion, M.B., Solomon, D., Ni, A., Zech, W., 2000. Soil organic matter quantity and quality in mountain soils of the Alay Range, Kyrgyzia, affected by land-use change. Biol. Fertil. Soils 31, 407–413. Golchin, A., Oades, J.M., Skjemstad, J.O., Clarke, P., 1994. Soil structure and carbon cycling. Aust. J. Soil Res. 32, 1043–1068. Grove, T.S., O’Connell, A.M., Mendham, D.S., Barrow, N.J., Rance, S.J., 2001. Sustaining the Productivity of Tree Crops on Agricultural Land in South-Western Australia, Publication No. 01/09. Rural Industries Research and Development Corporation (RIRDC), Canberra. Guggenberger, G., Zech, W., 1999. Soil organic matter composition under primary forest, pasture and secondary forest sucession, Region Huetar Norte, Costa Rica. For. Ecol. Manage. 124, 93– 104. Hingston, F.J., Turton, A.G., Dimmock, G.M., 1979. Nutrient distribution in Karri (Eucalyptus diversicolor F. Muell.) ecosystems in south-west western Australia. For. Ecol. Manage 2, 133–158. Hingston, F.J., Dimmock, G.M., Turton, A.G., 1981. Nutrient distribution in a Jarrah (Eucalyptus marginata Donn ex Sm.) ecosystem in south-west western Australia. For. Ecol. Manage 3, 183–207. Hingston, F.J., Galbraith, J.H., 1998. Application of the process-based model BIOMASS to Eucalyptus globulus ssp. globulus plantations on ex-farmland in south western Australia. II. Stemwood production and seasonal growth. For. Ecol. Manage. 106, 157–168. Ihori, T., Burke, I.C., Lauenroth, W.K., Coffin, D.P., 1995. Effects of cultivation and abandonment on soil organic matter in northeastern Colorado. Soil Sci. Soc. Am. J. 59, 1112–1119.
155
Jobbágy, E.G., Jackson, R.B., 2000. The vertical distribution of soil organic carbon and its relation to climate and vegetation. Ecol. Appl. 10, 423–436. Lal, R., 2001. World cropland soils as a source or sink for atmospheric carbon. Adv. Agron. 71, 145–191. Mann, L.K., 1986. Changes in soil carbon storage after cultivation. Soil Sci. 142, 279–288. Mendham, D.S., O’Connell, A.M., Grove, T.S., 2002. Organic matter characteristics under native forest, long-term pasture, and recent conversion to Eucalptyus plantations in Western Australia: microbial biomass, soil respiration, and permanganate oxidation. Aust. J. Soil Res. 40, 859–872. National Forest Inventory, 2000. National Plantation Inventory Tabular Report, March 2000. Bureau of Resource Sciences, Canberra. Neill, C., Melillo, J.J., Steudler, P.A., Cerri, C.C., de Moraes, J.F.L., Piccolo, M.C., Brito, M., 1997. Soil carbon and nitrogen stocks following forest clearing for pasture in the south-western Brazilian Amazon. Ecol. Appl. 7, 1216–1225. O’Connell, A.M., Grove, T.S., 1996. Biomass Production, Nutrient Uptake and Nutrient Cycling in the Jarrah (Eucalyptus marginata) and Karri (Eucalyptus diversicolor) forests of south-western Australia. In: Attiwill, P.M., Adams, M.A. (Eds.), Nutrition of Eucalypts. CSIRO Publishing, Melbourne, pp. 155–189. O’Connell, A.M., Grove, T.S., Mendham, D., Rance, S.J., 2000. Effects of site management in eucalypt plantations in south-western Australia. In: Nambiar, E.K.S., Tiarks, A., Cossalter, C., Ranger, J. (Eds.), Site Management and Productivity in Tropical Plantation Forests: A Progress Report. Center for International Forestry Research, Bogor, Indonesia, pp. 61–71. Polglase, P.J., Paul, K.I., Khanna, P.K., Nyakuengama, J.G., O’Connell, A.M., Grove, T.S., Battaglia, M., 2000. Change in Soil Carbon Following Afforestation or Reforestation, Technical Report No. 20. National Carbon Accounting System, Australian Greenhouse Office, Canberra. Post, W.M., Kwon, K.C., 2000. Soil carbon sequestration and land-use change: processes and potential. Global Change Biol. 6, 317–327. Rayment, G.E., Higginson, F.R., 1992. Australian Laboratory Handbook of Soil and Water Chemical Methods. Inkata Press, Melbourne. Reiners, W.A., Bouwman, A.F., Parsons, W.F.J., Keller, M., 1994. Tropical rain forest conversion to pasture: changes in vegetation and soil properties. Ecol. Appl. 4, 363–377. Rhoades, C.C., Eckert, G.E., Coleman, D.C., 2000. Soil carbon differences among forest, agriculture, and secondary vegetation in lower montane Ecuador. Ecol. Appl. 10, 497–505. Rossiter, R.C., Ozanne, P.G., 1970. South-western temperate forests, woodlands, and heaths. In: Moore, R.M. (Ed.), Australian Grasslands. Australian National University Press, Canberra, pp. 199–218. Rovira, A.D., 1992. Dryland mediterranean farming systems in Australia. Aust. J. Exp. Agric. 32, 801–809. Schlesinger, W.H., 1977. Carbon balance in terrestrial detritus. Annu. Rev. Ecol. Syst. 8, 51–81.
156
D.S. Mendham et al. / Agriculture, Ecosystems and Environment 95 (2003) 143–156
Silver, W.L., Neff, J., Mcgroddy, M., Veldkamp, E., Keller, M., Cosme, R., 2000. Effects of soil texture on below-ground carbon and nutrient storage in a lowland Amazonian forest ecosystem. Ecosystem 3, 193–209. Specht, R.L., 1996. The influence of soils on the evolution of the eucalypts. In: Attiwill, P.M., Adams, M.A. (Eds.), Nutrition of Eucalypts. CSIRO Publishing, Melbourne, pp. 31–60. Tate, K.R., 1992. Assessment, based on a climosequence of soils in tussock grasslands, of soil carbon storage and release in response to global warming. J. Soil Sci. 43, 697–707. Thomson, C.J., Ewing, M.A., Turner, N.C., Revell, C.K., Le Coultre, I.F., 1998. Influence of rotation and time of germinating rains on the productivity and composition of annual pastures in western Australia. Aust. J. Agric. Res. 49, 225–232. Thuille, A., Buchmann, N., Schulze, E.D., 2000. Carbon stocks and soil respiration rates during deforestation, grassland-use and subsequent Norway spruce afforestation in the southern Alps, Italy. Tree Physiol. 20 (13), 849–857.
Turner, J., Lambert, M., 2000. Change in organic carbon in forest plantation soils in eastern Australia. For. Ecol. Manage. 133, 231–247. Underwood E.J., Gladstones J.S., 1979. Subterranean clover and other legumes. In: Burvill, G.H. (Ed.), Agriculture in Western Australia 1829–1979. University of Western Australia Press, Nedlands, Western Australia, pp. 139–156. Van Cleve K., Powers R.F., 1995. Soil carbon, soil formation, and ecosystem development. In: McFee, W.W., Kelly, J.M. (Eds.), Carbon Forms and Functions in Forest Soils. Soil Science Society of America Inc., MA, USA. Veldkamp, E., 1994. Organic carbon turnover in three tropical soils under pasture after deforestation. Soil Sci. Soc. Am. J. 58, 175–180. Weaver, D.M., Reed, A.E.G., 1998. Patterns of nutrient status and fertiliser practice on soils of the south coast of western Australia. Agric. Ecosyst. Environ. 67, 37–53.