Soil organic matter in rainfed cropping systems of the Australian cereal belt(*).
The Australian cereal belt stretches as an arc from north-eastern Australia to south-western Australia (24 [degrees] S-40 [degrees] S and 125 [degrees] E-147 [degrees] E; Fig. 1), with mean annual temperatures from 14 [degrees] C (temperate) to 26 [degrees] C (subtropical), and annual rainfall ranging from 250 mm to 1500 mm. With high potential evapotranspiration, often exceeding 2000 mm annually, most of the cereal belt lies in semi-arid to arid regions.
The predominant soil types are Chromosols, Sodosols, Vertosols, and Kandosols with significant areas of Ferrosols, Kurosols, Podosols, and Dermosols (Isbell 1996), covering approximately 20 Mha of grain cropping and 21.1 Mha of improved pastures (NGGIC 1999c). According to McGarity (1975), Vertosols are mainly found in the southern Queensland and northern New South Wales cereal belt, Chromosols and Sodosols in the southern cereal belt, and Kandosols in the western cereal belt (Western Australia).
Major crops include wheat (Triticum aestivum), barley (Hordeum vulgate), lupin (Lupinus angustifolius), chickpea (Cicer arietinum), canola (Brassica napus), field pea (Pisum sativum), sorghum (Sorghum bicolor), maize (Zea mays), soybean (Glycine max), fababean (Viciafaba), and dryland cotton (Gossypium hirsutum). The dominant cropping systems in south-western and southern Australia are wheat-lupins, wheat-ley pasture, wheat-chickpea, wheat-canola, barley-field peas-wheat, and a combination of these cropping systems for longer rotations. In the north-east, sorghum forms a significant component of the grain cropping systems, such as sorghum-fallow-wheat. Also, ley pastures in the rotation are less common in the north-east region.
Soil organic matter (SOM) is the organic fraction of the soil but excludes undecayed plant and animal residues (Soil Science Society of America 1997). Its functions in soil are summarised in Fig. 2.
Land clearing and development for cropping often lead to reduced organic matter levels in soil (Russell and Williams 1982; Dalal and Mayer 1986a, 1986b). This is primarily due to changes in temperature, moisture fluxes, and soil aeration; to exposure of new soil surfaces resulting from aggregate disruption; to reduced input of organic materials and increased export of nutrients; and frequently to increased soil erosion. Large soil carbon losses contribute to increased [CO.sub.2] emissions to the atmosphere (Gifford et al. 1992) and frequently lead to land degradation and agricultural unsustainability (Lal 1998). Conversely, restoration of SOM improves soil quality, maintains/enhances agricultural sustainability, and increases the soil C sink to mitigate elevated [CO.sub.2] concentration in the atmosphere (AGO 2000). It is, therefore, appropriate to review soil organic matter in the semi-arid and arid Australian cereal belt, since agricultural sustainability and enhancement of the C sink in soil are often interrelated.
The objectives of this review are to summarise current understanding of organic matter in terms of its components and functions in soil and biomass productivity, and soil C sinks and sources for greenhouse gases; to examine management practices that affect SOM levels in Australian cropping soils and management practices that enhance sinks for C in terms of sustainability and greenhouse gases, and soil quality; and to identify areas of future SOM research.
Components of soil organic matter
Temperature and moisture (net effect of rainfall and potential evapotranspiration) are the major determinants of plant biomass production and, along with soil type, determine the SOM content as well as the rate of change of SOM under different land uses. Since the SOM content depends upon the relative rates at which organic materials are added to the soil and lost from it through decomposition, it can mathematically be expressed as follows (Bartholomew and Kirkham 1960):
(1) d SOM/dt = A - k SOM
(2) or [SOM.sub.t] = [SOM.sub.0] exp (-kt) + A/k [1 - exp(-kt)]
where [SOM.sub.0] and [SOM.sub.t] are the SOM contents initially (t = 0), and at a given time, t, respectively; A (mass of SOM per unit area) is the rate at which organic matter is returned to the soil; and k (reciprocal of time) is the rate of loss of SOM. Since SOM is heterogeneous and its various components decompose at different rates, Eqn 2 is modified to:
(3) [SOM.sub.t] = [SOM.sub.1] exp ([-k.sub.1]t) + [A.sub.1]/[k.sub.1] [1 - exp ([-k.sub.1]t)] + ...
+ [SOM.sub.n] exp ([-k.sub.n]t) + [A.sub.n]/[k.sub.n] [1 -exp ([-k.sub.n]t)]
where [SOM.sub.1] and [SOM.sub.n] refer to different components of organic matter, such as microbial biomass and light fraction (active C pool), heavy or clay-size C fraction (passive, inert SOM), and the remaining slow C pool (Jenkinson and Rayner 1977; Dalal and Mayer 1986c, 1986d, 1987a, 1987b; Parton et al. 1987, 1996; Jenkinson et al. 1992) and [k.sub.1] and [k.sub.n] refer to the rate of loss of individual components. SOM dynamic models consider these components as well as the factors affecting these components such as climate, organic matter addition, soil matrix, cultural practices, and some soil degradation processes such as soil erosion (Parton et al. 1996).
(4) When t = 0, Eqn 3 is reduced to [SOM.sub.t] =[ SOM.sub.1] + ... +[SOM.sub.n] =[SOM.sub.0], and at time t = [infinity] [SOM.sub.t] = [A.sub.1]/[k.sub.1] + ... + [A.sub.n]/[k.sub.n] = [SOM.sub.e], which may be simplified to give:
[SOM.sub.e] = A/k
where [SOM.sub.e] is organic matter content at steady state after a long period of consistently similar soil management (Bartholomew and Kirkham 1960) or under native vegetation. The simplified form of Eqns 2 and 3 is:
(5) [SOM.sub.t] = [SOM.sub.e] + ([SOM.sub.o] - [SOM.sub.e]) exp (-kt)
where k is the overall rate of loss (decomposition) of various organic matter components. The turnover period of SOM is then reciprocal of the k value. Organic matter turnover in soil at a steady state can be estimated from Eqn 4, especially in stable natural ecosystems where it can reasonably be assumed that organic matter additions balance organic matter losses. Eqn 5 is used in this review to discern trends in soil C changes with time, since there is limited information available on various components of SOM in the Australian cropping soils.
In Eqn 3 above, it is assumed that various components of soil organic matter decompose at different rates. However, it is difficult to define, identify, fractionate, and assign distinct functions to its various components (Christensen 1992; Baldock and Skjemstad 1999). The operational scheme presented in Fig. 3 is adapted from Baldock and Skjemstad (1999) to fractionate various components of SOM and is summarised below into 3 major categories: active C pool, slow C pool, and passive C pool.
The active C pool is characterised by its rapid rate of turnover in soil, mostly in [is less than] 10 years in mesic subhumid to semi-arid environments. In the operational scheme suggested in Fig. 3, it is described as carbon content of particulate organic matter (POM-C) or macroorganic matter and can be approximated to organic C in the sand-size fraction [is greater than] 20 [micro]m (Dalal and Mayer 1986c) or [is greater than] 53 [micro]m (Cambardella and Elliott 1993). The so-called light fraction SOM of density [is less than] 1.6 Mg/[m.sup.3] (Greenland and Ford 1964) or [is less than] 2 Mg/[m.sup.3] (Dalal and Mayer 1986d) has also been used as an active C pool. Parton et al. (1996) suggested 2-3 times the microbial biomass is equivalent to the active C pool in soil. However, it appears from Table I that the active pool could be 10 times the microbial biomass pool.
The active C pool organic matter consists mostly of plant debris (Turchenek and Oades 1979), and mesofauna and fungal hyphae (Oades and Ladd 1977), and small but significant amounts of microbial biomass and its products in the silt-size and clay-size fractions (Dalai and Mayer 1987b). The amounts of active pool of organic C vary widely in Australian soils (Table 1). In undisturbed soils, POM-C varies from 1% to 85% (Greenland and Ford 1964; Christensen 1992; Gregorich and Janzen 1996), depending on the nature of ecosystem, vegetation, climate, and soil characteristics. Data collated from a limited number of soils show the mean values of 34% for POM-C, 23% for the light-fraction C, and 23% for 10 times the microbial biomass C (Table 1).
Table 1. Amounts of active C pool in 0-0.1 m depth estimated as particulate organic matter (POM >20 [micro]m or >53 gm, light fraction (LF <2 Mg/[m.sup.3] or <1.6 Mg/[m.sup.3]) organic matter, or microbial biomass (MB) in relatively undisturbed soils in the Australian cereal belt Soil POM-C (t/ha) (% total organic C) Waco (Vertosol), Qld 2.7(A) 19.9 Langlands-Logie (Vertosol), Qld 7.2)A) 32.8 Riverview (Kandosol), Qld 5.0(A) 31.9 Namoi (Vertosol), NSW 6.9(D) 51.5 Talbarear (Vertosol), NSW 2.5(D) 35.1 Meadows (Sodosol), SA Glen Osmond (Chromosol), SA Mean values 4.9 34.2 [+ or -] [+ or -] 2.2 11.3 Soil LF-C (t/ha) (% total organic C) Waco (Vertosol), Qld 2.9(B) 20.8 Langlands-Logie (Vertosol), Qld 5.9(B) 26.6 Riverview (Kandosol), Qld 3.8(B) 23.9 Namoi (Vertosol), NSW Talbarear (Vertosol), NSW Meadows (Sodosol), SA 10.1(E) 28.5 Glen Osmond (Chromosol), SA 4.9(E) 17.4 Mean values 5.5 23.40 [+ or -] [+ or -] 2.8 4.5 Soil 10 x MB-C (t/ha) (% total organic C) Waco (Vertosol), Qld 3.8(C) 27.8 Langlands-Logie (Vertosol), Qld 5.0(C) 22.8 Riverview (Kandosol), Qld 2.6(C) 16.8 Namoi (Vertosol), NSW Talbarear (Vertosol), NSW Meadows (Sodosol), SA Glen Osmond (Chromosol), SA Mean values 3.80 22.50 [+ or -] [+ or -] 1.2 5.1 (A) >20 [micro]m, Dalal and Mayer (1986c). (B) <2 Mg/[m.sup.3], Dalai and Mayer (1986d). (C) Dalal and Mayer (1987b). (D) >53 [micro]m, Chan (1997). (E) <1.6 Mg/[m.sup.3], Golchin et al. (1994).
The active (or labile) C pool, estimated from chemical extractions such as KMn[O.sup.4] (Blair et al. 1995), is not considered in this report since the nature of soil C extracted is undefined.
The passive pool is taken as the amount of charcoal (char C) in the soil (Skjemstad et al. 1996), and/or physically protected organic matter, such as in the clay fraction (Parton et al. 1987). Values of char C and the clay-size C for a limited number of undisturbed soils are given in Table 2. Almost 25% of the total C may occur as char C and the clay-size C could be as high as 50%. However, the latter may be an overestimate since it also contains small amounts of active C such as microbial biomass and lysed microbial cells in undisturbed soils, as demonstrated by clay-size C loss when these soils are cultivated (Dalal and Mayer 1986c). The clay-size C as a proportion of the initial total organic C remaining after 45-70 years of cultivation appears to be about 25%, a figure similar to that of char C, about 22% in these soils (Table 2; Skjemstad et al. 2001).
Table 2. Amounts of passive C pool in 0-0.1 m depth estimated as charcoal organic matter (Char C) or clay fraction (<2 gm) organic C in relatively undisturbed soils in the Australian cereal belt The values in parentheses are the estimated equilibrium values using Eqn 5, after 45-70 years of cultivation Soil Char C (t/ha) (% total organic C) Waco (Vertosol), Qld 5.5(A) 30.2 Langlands-Logie (Vertosol), Qld 4.6(A) 17.0 Riverview (Kandosol), Qld Gumercha (Chromosol), SA 5.2(C) 20.0 Mean values 5.1 [+ or -] 22.4 [+ or -] 0.5 6.9 Soil Clay-size C (t/ha) (% total organic C) Waco (Vertosol), Qld 7.2 (5.4)(B) 52.4 (39.4) Langlands-Logie (Vertosol), Qld 9.1 (4.3)(B) 41.2 (19.5) Riverview (Kandosol), Qld 7.37(B) 47.0 Gumercha (Chromosol), SA 10.9(C) 42.0 Mean values 8.6 [+ or -] 45.7 [+ or -] 1.7 5.2 (4.9) (29.5) (A) Skjemstad et al. (2001); (B) Dalal and Mayer (1986c); (C) Skemstad et al. (1996).
The slow pool is then calculated by subtracting the sum of the active pool and the passive pool from the total organic C in soil. It is estimated to be about 50% of the total organic C in soil. Total organic C contents in various Australian soils are listed by Baldock and Skjemstad (1999) and have been discussed in detail by Spain et al. (1983).
Climate and management effects on soil organic matter
As mentioned earlier, rainfall, temperature, and potential evapotranspiration determine the plant biomass production under adequate nutrition conditions in rainfed agriculture. Therefore, in a steady state, generally plant residue C inputs determine the organic C levels in cultivated soils (Eqn 1).
Dalal and Mayer (1986b) observed that, in the contiguous region (similar rainfall and temperature), initial organic C contents (virgin soils) as well as organic C contents after 20 years of cultivation for cereal cropping were closely correlated with the mean annual precipitation. However, in soil under native vegetation the rate of increase in organic C values for each mm of precipitation was almost twice that in cultivated soils (48 kg C/mm v. 29 kg C/mm) in the top 0.1 m depth. This was primarily due to the lesser amount of organic residues produced and returned to the soil, and/or accelerated rate of organic matter decomposition in cultivated soils. Increased incidence of erosion may also be a contributing factor (Cook et al. 1992). In wider regions, however, the ratio precipitation:potential evapotranspiration determines the plant biomass production (Parton et al. 1987, 1996).
Temperature affects both the rate of SOM decomposition and C inputs from plant biomass. Generally, SOM loss is faster in the tropics than in subtropical or temperate regions. Bridge and Bell (1994) observed 59% loss in total organic C (0-0.1 m) from a Ferrosol after more than 50 years of cultivation in semi-arid subtropical Queensland, whereas Cogle et al. (1995) measured a 56% loss in organic C after about 10 years of cropping in subhumid tropical Queensland, thus substantiating the rapid decomposition of organic C in a moist tropical environment (Dalal and Probert 1997). Ladd et al (1981) showed that plant residue decomposed twice as fast in Nigeria (26 [degrees] C) as South Australia (16 [degrees] C), with a temperature-dependent rate quotient of close to 2 for every 10 [degrees] C change in temperature, as expected from temperature-dependent biological decomposition.
In a geographically contiguous region, Dalal and Mayer (1987b) found that microbial biomass C increased with increasing mean annual precipitation and decreased with increasing mean annual temperature; similar climate effects are expected on other measures of active C pools in soils.
Short-term seasonal effects on soil organic C, mainly the active C pool, are primarily reflected through the carbon inputs from the growing plant biomass and plant residues (Dalal 1998). For example, Dalal et al. (1994) observed that the soil microbial biomass in a Vertosol was significantly higher immediately after crop harvest than following the 6-month's clean fallow period. This effect was further accentuated when the 18-month's fallow was practised.
Cultivation of a soil that previously supported native vegetation or pasture generally leads to a reduced level of soil organic C (Dalal and Mayer 1986b; Loch et al. 1987). The organic matter level of a cultivated soil eventually attains a steady state, where rate of formation of new SOC from organic residues (plant and crop residues, roots and root exudates, organic wastes, manures and green manures) equals rate of SOM decomposition (Eqn 4), provided that soil and crop management practices remain essentially similar over a long period.
Eqns 3-5 have been used to describe SOM dynamics in soils under cultivation (Jenkinson and Rayner 1977; Dalal and Mayer 1986b, 1986c, 1986d, 1986e; Parton et al. 1987, 1996; Jenkinson et al. 1992). The rates of loss and turnover period of SOC in different soils vary considerably. Organic C from Kandosols turns over much faster than from Vertosols in a similar climate (Dalal and Mayer 1986b, 1986c, 1986d). At least part of the effect is due to the clay content, the decomposition being more rapid in a coarse soil than in a fine-textured soil. This is illustrated in the results presented in Table 3 on total soil organic C loss when Vertosols and Kandosols were subjected to long periods of cultivation. The rate of net organic C loss declined exponentially as the clay content in soil increased, irrespective of clay mineralogical composition (k value = 0.05+8.1 exp(-0.11 clay), [r.sup.2] = 0.99). Using spectroscopic techniques, including [sup.13]C nuclear magnetic resonance spectroscopy, Skjemstad et al. (1986) showed that, in some Vertosols (high in mixed clay minerals), the chemical nature of soil organic C remained essentially similar in spite of 60% C lost upon cultivation and cereal cropping. Alkyl and O-alkyl groups remained dominant throughout the C loss to 60% (Skjemstad et al. 2001). However, in other Vertosols (high in smectitic clay), Skjemstad and Dalal (1987) observed that besides the association of SOM with clay, SOM stability against decomposition in these soils may be due to increased aromaticity (aryl-C), possibly charcoal (Skjemstad et al. 2001), as well as shorter, more highly branched alkyl chains.
Table 3. Net organic C loss from 0-0.1 m layers of Vertosois and Kandosols subjected to long-term cultivation and cereal cropping in southern Queensland From Dalal and Mayer (1986b) Soil Clay content Initial organic C (%) (t/ha) Waco clay (Vertosol) 72 8.26 Thallon clay (Vertosol) 59 4.41 Langlands-Logie clay (Vertosol) 49 7.76 Cecilvale clay (Vertosol) 40 10.18 Billa Billa loamy clay (Vertosol) 34 8.27 Riverview sandy loam (Kandosol) 18 9.38 Soil Rate of loss, k (1/year) Waco clay (Vertosol) 0.065 Thallon clay (Vertosol) 0.069 Langlands-Logie clay (Vertosol) 0.080 Cecilvale clay (Vertosol) 0.180 Billa Billa loamy clay (Vertosol) 0.259 Riverview sandy loam (Kandosol) 1.224
The changes in soil oranic C with different periods of cultivation (tillage) and cereal cropping in a Haplustert (Vertosol, black earth), a Chromustert (Vertosol, grey cracking clay), and a Paleustalf (Kandosol, red earth) are shown in Fig. 4 (Dalal and Mayer 1986b).
Russell and Williams (1982) summarised the percentage loss of organic C from a number of Australian soils; these varied from 10% to 60% over 10-80 years of cultivation. However, time-series SOM decline measurements were not made. Similar observations have since been made by Standley et al. (1990) and Chan et al. (1997).
In a number of these studies cited above, changes in bulk density with the loss in SOC and, consequently, on depth of sampling were not considered. Hence equivalent-sampling depths may differ from actual sampling depths (Dalal and Mayer 1986b), so these estimates of rates of soil organic C loss should be considered approximate.
Mechanisms for the loss of SOM under tillage
Tillage associated with cropping has often been considered responsible for the decline in soil organic carbon (Hallsworth et al. 1954; Russell and Williams 1982; Dalal and Mayer 1986a; Grace et al. 1995). Mechanical actions of tillage, which result in fracturing, pulverisation, and mixing of soil, can directly and indirectly result in the reduction of SOC concentration (Rovira and Greacen 1957; Roberts and Chan 1990). There is a popular view that disruption of the soil aggregates during tillage leads to exposure of physically protected soil organic C to microbial breakdown and therefore it is the cause of reduction in soil organic C under cropping (Davidson 1986; Chartres et al. 1992). Whereas increases in soil respiration (measured as increases in carbon dioxide) have been demonstrated as a result of mechanical disturbance of soil, the equivalent amount of organic carbon loss is generally rather small compared with the total loss in soil organic C (Rovira and Greacen 1957; Roberts and Chan 1990). Roberts and Chan (1990) measured the amount of soil organic carbon loss of a sandy loam as a result of simulated tillage in the laboratory and found that the loss per operation, depending on the severity of soil disturbance due to the simulated tillage operations, ranged from 0.005 to 0.037 g C/kg soil. This is in good agreement with results of other researchers (Table 4).
Table 4. Enhanced organic carbon losses (g C/kg soil) as a result of C respiration due to tillage operations Carbon losses Tillage operations References 0.019 Simulated tillage under Rovira and Greacen laboratory conditions (1957) 0.005-0.037 Simulated tillage under Roberts and Chan laboratory conditions (1990) 0.036-0.16 Disk harrow to 7.6 cm and Reicosky and Lindstrom chisel ploughing to 7.6 cm (1993)
Nevertheless, for the same soil in the field under different tillage practices, the additional loss in soil organic carbon in the top 0.05 m after 2 years of conventional tillage when compared with no-till was 4.8 g C/kg soil (Chan and Mead 1988) or 6 t C/ha (bulk density, 1.25 Mg/[m.sup.3]). According to the simulated tillage experiment (Table 4), this is equivalent to 130-960 tillage operations, depending on the intensity of the soil disturbance. Therefore, the loss due to this mechanism alone cannot account for the observed soil organic C loss as recorded in the field under conventional tillage. Other SOM loss processes such as water erosion and wind erosion may also have been involved.
Apart from the direct effects of tillage, the SOM also declines under continuous cropping due to lower annual C inputs under a cropping system than under a grassland or forest ecosystem. The content of soil SOM that can be maintained in a given region ultimately depends on the production and input of plant biomass and/or addition of off-site C source such as farmyard manure and organic wastes/materials.
SOC pools in cultivated soils
As mentioned above, a more complete understanding of the dynamics of SOM requires knowledge of different C pools, their amounts, and their rates of turnover (Eqn 3). These C pools include an active pool (sand-size or POM-C), a slow pool, and a passive pool (char C or a proportion of clay-size C).
The experimental values corresponding to the conceptual soil C pools are given in Table 5 for the Langlands-Logie soil series subjected to 45 years of cultivation and cereal cropping. In general, the active C pool decreases the most, and much faster than the slow C pool (total C less sand-size C and clay-size C) and passive C pools, whereas the slow pool is intermediate between the active and passive pools. The significance of these C pools, including microbial biomass, is considered below.
Table 5. Rates of loss of organic C from the whole soil and active, slow, and passive pools in 0-0.1 m layer of the Langlands-Logie soil series (Vertosol) subjected to 45 years of cultivation and cereal cropping Soil C pool Initial C Steady state C [C.sub.o] [C.sub.e] (kg/ha) (kg/ha) Whole soil(A) 22070 7760 Active pool Light fraction, <2 Mgm(3C) 5880 1200 Sand-size fraction, >20 7230 60 [micro]m(B) Slow pool Silt-size fraction, 2-20 5210 1910 [micro]m(B)? Passive pool Clay-size fraction, <2 9110 4300 [micro]m(B) Char C(D) 5600 3300 Soil C pool Percent C los Rate of 100 x ([C.sub. loss o] - [C.sub. (1/year) e])/[C.sub.o] (%) Whole soil(A) 65 0.080 Active pool Light fraction, <2 Mgm(3C) 79 0.194 Sand-size fraction, >20 99 0.109 [micro]m(B) Slow pool Silt-size fraction, 2-20 63 0.078 [micro]m(B)? Passive pool Clay-size fraction, <2 53 0.039 [micro]m(B) Char C(D) 41 -- (A) Dalal and Mayer (1986b). (B) Dalal and Mayer (1986c). (C) Dalal and Mayer (1986d). (D) Skjemstad et al. (2001).
Active C pool
Soil microbial biomass
Microbial biomass C is considered a sensitive indicator of the effects of management practices over relatively short periods compared with total SOC (Powlson et al. 1987; Sparling 1992; Dalai 1998). For example, in Rutherglen, northern Victoria, soil organic C was 12% higher but microbial biomass was 43% higher after 9 years of no-till and stubble retention than under conventional till and stubble burned on a Chromosol (Haines and Uren 1990). Returns of reduced amounts of organic residues to cultivated soils would rapidly decrease the amount of microbial biomass C. Since microbial biomass is closely related to a suite of biological and biochemical activities in soil (Dalai and Mayer 1987b), the reduced amount of microbial biomass upon cultivation of previously fertile soils leads to soil fertility depletion due to lower labile organic N, P, and S, and to reduced retentive capacity (immobilisation) for mineral N in most soils (Dalai 1998). Furthermore, continuous cultivation for cereal cropping, especially monoculture, may reduce not only the total amount but also may alter the composition of microbial biomass, which could lead to biological degradation (Pankhurst 1997).
Light fraction, [is less than] 2 Mg/[m.sup.3] and sand-size organic matter (macroorganic matter)
The proportion of light fraction (or sand-size fraction) to total SOM provides an earlier and a sensitive indication of the consequences of different soil management practices than the total SOM (Christensen 1992; Gregorich and Janzen 1996; Chan 1997). Moreover, since crop residues and other organic materials in soil maintain the active C pool, non-return of these materials (e.g. stubble removal and/or burning) deprives soil of its active fraction of organic C (Chan 1997).
The amount and rate of loss of light fraction and sand-size organic C in cultivated soils is very rapid (Table 5). Upon cultivation, [is greater than] 80% of the light fraction SOM can be lost from soil within 45-70 years (Dalal and Mayer 1986d; Skjemstad and Dalal 1987).
Slow pool, silt-size organic C
The rate of SOC loss from the silt-size fraction is much slower than that from the active pool (sand-size and the light fraction) (Table 5). Dalal and Mayer (1986c) observed that the proportional loss of SOC from cultivation for cereal cropping for 20-70 years was similar to that of SOC in the silt-size fraction of 5 Vertosols and 1 Chromosol. Thus, silt-size C remained essentially similar (26-27%) upon cultivation (20-70 years) of virgin soils.
Passive pool, clay-size C and char C
The rate of loss of organic C from the clay size fraction is generally lower than that from the other fractions and the whole soil (Table 5) (Dalal and Mayer 1986b, 1986c, 1986d; Parton et al. 1987, 1996; Christensen 1992), resulting in an increasing proportion of total SOM (often [is greater than] 60%) remaining in the clay fraction after long-term cultivation. Skjemstad et al. (1996) demonstrated that the passive C might indeed mostly occur as char C in many Australian soils. This was substantiated by the large proportion of char C remaining in two Vertosols cultivated for 45-50 years (Skjemstad et al. 2001) (Table 5). Other studies have confirmed the differential loss of organic C from different soil C fractions (Chan 1997). Therefore, the magnitude of loss not only of SOM but also of various fractions varies widely in Australian soils and thus the consequences of SOM loss on soil fertility depletion may vary in their type and magnitude considerably.
Modelling SOM turnover
Most SOM models assume that different components of organic matter differ in their rates of decomposition and turnover periods (Jenkinson and Rayner 1977; Van Veen and Paul 1981; Parton et al. 1987, 1996; Jenkinson et al. 1992), although most of these are only conceptual components, and therefore, currently, not all of the conceptual fractions of SOM have been experimentally verified.
We have attempted to simulate using the `CENTURY' model (Parton et al. 1996) organic C dynamics from the experimentally determined values (given in Table 5) for the Langlands-Logie soil series. The results are presented in Fig. 5.
Dalal and King (1994) found that the magnitudes of decline in total soil organic C and slow component were reasonably close to the experimentally determined values, although the dynamics of the active component of SOM (Dalal and Mayer 1986c, 1986d) were poorly simulated (Dalal and King 1994).
The SOM models were generally developed and validated in temperate regions. Parton et al. (1989) pointed out at least 3 deficient areas of understanding, which could make applicability of these models more appropriate. These include: (1) functional relationships of soil texture, clay mineralogy, and especially Fe and Al on SOM dynamics; (2) clay mineralogy and parent material effects on the formation of `passive' SOM fractions; and (3) low pH and Al toxicity in acidic soils effects on microbial transformations.
Probert et al. (1995) applied the CENTURY model to simulate changes in total N (as a measure of organic C) in a Vertosol used for winter cereal cropping for 20 years. It simulated total N levels satisfactorily for no-till and stubble-retained systems but poorly for conventional-tilled and stubble-burned systems. Obviously, considerable improvements in initialisation and parameterisation of models are required to simulate organic matter dynamics in the Australian cereal belt.
Also, there is a further need to experimentally verify the various `conceptual' pools of SOM used in these models. Most SOM models assume a `protected' or `recalcitrant' fraction of SOM, which usually correlates with the clay (and silt) fraction, although the nature and magnitude of aggregate-size distribution in soil would be more relevant (Dalai and Mayer 1986b; Elliott 1986) to SOM dynamics than the total clay (and silt) contents, especially in conjunction with the density fractions of SOM. Unfortunately, methods used to fractionate SOM in different aggregate-size and density fractions are poorly described, and thus, comparative evaluation of the importance of aggregate-size and density distribution of SOM on SOM dynamics is made extremely difficult. It is imperative, therefore, to standardise the methods of determining the fractions of SOM based upon soil aggregate size and density and then to ascertain whether aggregate size and density fractions of SOM from long-term cultivated soils are biologically meaningful in SOM turnover in Australian soils.
Moreover, char C may be a dominant passive C pool in the Australian soils (Skjemstad et al. 1996), since fire has been a salient feature of the Australian landscape for thousands of years.
Consequences of soil organic matter loss
Nutrient depletion from cultivated soils
Loss of SOM on cultivation of virgin or pasture lands is also accompanied by the loss of its constituent nutrients, especially if they are not readily retained by the soil, such as N and S. Besides the loss of nutrients associated with loss of SOM, available nutrient depletion can occur by one or more processes of depletion, such as nutrient removal in product and plant residue, soil erosion, burning, leaching, and denitrification of N (Dalal and Probert 1997).
Loss of nitrogen from cultivated soils
The estimates of the rate of soil total N losses from cultivated Vertosols vary widely. Martin and Cox (1956) measured 0.8% and Hallsworth et al. (1954) 2% annual decrease in soil N from the top 0.15 m layer, whereas Waring and Teakle (1960) estimated that 3% of total N was removed by wheat crops annually. Russell (1981) measured N losses of 5% annually from an initially highly fertile Vertosol cropped to sorghum continuously for 10 years.
The mean annual rates of N loss from the profiles of 5 Vertosols and 1 Kandosol subjected to 20-70 years of cereal cropping were 31.3-67.1 kg/ha.year of N (Dalal and Mayer 1987a), which were in most soils similar to that accounted for in crop removal. Some of the N loss was compensated for by the application of fertiliser N. Similar studies done elsewhere on soils subjected to long periods of cultivation have shown that in most cases, soil N losses were primarily accounted for by N removal by crops (Haas et al. 1957; Williams and Lipsett 1961), especially where no organic manures or fertilisers have been added. However, in soils of initially high fertility, N losses by other factors, such as by deep leaching and possibly by denitrification, especially from applied fertiliser N, do occur (Dalal and Probert 1997).
Potentially mineralisable N
The active C pool loss is more closely related to the depletion of mineralisable nutrients than to the total amounts when a soil under a natural ecosystem is brought under cultivation and cropping (Ayanaba et al. 1976; Dalal and Mayer 1987b, 1990; Chan 1997).
Dalal and Mayer (1986a) found that total N and anaerobic mineralisable N declined in a Vertosol (Waco soil) by 35% and 55%, respectively, after a mean cultivation period of 26 years. In a Kandosol (Riverview soil), after a mean cultivation period of only 7 years, total N and anaerobic mineralisable N declined by 28% and 51%, respectively. Thus, the decreases in anaerobic mineralisable N due to cropping exceeded those in total N, microbial biomass N, and even in N mineralisation potentials ([N.sub.o]) (Dalal and Mayer 1987b).
Loss of P from cultivated soils
Since organic P is a constituent of organic matter, the decline in the latter with cultivation also leads to a decline in the soil organic P content (Dalal 1997). Also, crop removal and other P-depleting factors, such as erosion and runoff, would deplete the plant-available P pool in soil.
Chan et al. (1988) measured substantially lower (39-71%) available P, using Bray No. 1 extractant, in a number of Vertosols (north-western New South Wales) cultivated for 8-50 years than in those under permanent pasture. They ascribed the decrease in available P to the increase in soil pH upon cultivation. Standley et al. (1990) also found that the available P (bicarbonate-extractable) in a Chromustert from central Queensland decreased from 31 mg P/kg soil initially to [is less than] 20 mg P/kg soil after 7 years of cropping for sorghum, but they found no significant change in soil pH. Dalal(1997) found that the bicarbonate-extractable P during 45 years of cultivation for cereal cropping declined at the rate of 1.2 mg P/kg.year. The bicarbonate-extractable P is derived from both organic P and inorganic P in soil (Dalal and Mayer 1986a; Dalal 1997). In general, reduction in P in soil due to the loss of `labile' P has been closely associated with SOM losses (Tiessen and Stewart 1983). However, the role of SOM in contributing towards and maintaining plant-available P in soils has not been well understood.
Bulk density changes
Increases in bulk density can be brought about by compaction (heavy machinery and tillage) and loss in SOM. Harte (1984) recorded an increase in bulk density (0-0.2 m soil depth) from 1.30 Mg/[m.sup.3] to 1.62 Mg/[m.sup.3] (25% increase) in a loamy Chromosol, from 1.42 Mg/[m.sup.3] to 1.62 Mg/[m.sup.3] (14% increase) in a sandy Chromosol, and from 1.13 to 1.22 Mg/[m.sup.3] (8% increase) in a Ferrosol after 15 years of continuous cultivation and cropping in northern New South Wales, Australia. Harte (1984) also found significant negative correlations between soil organic matter content and bulk density in these soils, and thus substantial increases in bulk density were associated with declines in soil organic matter.
However, changes in SOC levels may not always be associated with changes in bulk density, especially due to changes in tillage practices. For a Chromosol in Wagga Wagga, Chart et al. (1992) and Chan and Heenan (1993)observed similar bulk density (1.31 Mg/[m.sup.3]) despite a 27% difference in SOC in the 0-0.05 m layer between the conventional tilled and no-till treatments.
The increase in bulk density following the loss in organic matter is much greater in soils where organic matter is prominent in aggregate formation and stabilisation (Chromosols, Kandosols, and Kurosols) than where it is less so (Vertosols, where clay matrix and shrinks--well characteristics predominate, and Ferrosols, where oxides of Fe and Al predominate) (Dalal and Bridge 1996).
Aggregation and SOC are usually closely correlated (Table 6). The striking results from Table 6 are that aggregation in Kandosols and Chromosols (21.5-32.1) is 2-7 times more responsive to the change in SOM than that in Vertosols (4.3-10.5). Dalal et al. (1991) also found that aggregation (silt and clay aggregated) declined 5 times faster in Kandosols than in Vertosols when these soils were cultivated for 20-45 years for cereal cropping, and a rate of loss in soil aggregation was similar to that of loss of SOM from these soils (Dalal and Mayer 1986b). Also, the loss of organic matter and subsequent decrease in aggregate stability occurs at a more rapid rate immediately following the cultivation of virgin or grassland soils (Tisdall and Oades 1980). This is because the POM-C or light fraction C, which is present as transient binding organic matter (polysaccharides, polyuronides, mucilages, microbial debris) and temporary binding organic matter (roots, fungal hyphae), disappears at a much faster rate than the whole SOM (Greenland and Ford 1964; Dalal and Mayer 1986d; Chan 1997). Conversely, aggregates of [is less than] 0.125 mm size decrease as the organic matter increases in soil (Bell et al. 1998). Based on the hierarchical model of soil structure (Tisdall and Oades 1982), the transient and temporary forms of SOM are responsible for the stabilisation of the macro-aggregates ([is greater than] 0.25 mm). This explains why the stability of the macroaggregates is so dependent on soil and agronomic management practices.
Table 6. Linear relationship between aggregation (y = % mean weight diameter in mm) and soil organic C (x = % organic C) Soil order Regression Regression R value constant coefficient [Chromosols(A)] -20.3 21.5 0.96 [Kandosols(B)] -5.8 32.1 0.67 [Vertosols(C)] 72 10.5 0.80 [Vertosols(D)] 70 4.3 0.76 (A) Tisdall and Oades (1980). (B) Chan et al. (1992). (C) Calculated from Chan et al. (1988) for aggregates > 0.05 mm diameter. (D) R. C. Dalal (unpubl. data).
Cook et al. (1992) found that the decrease in SOC was closely related to the decrease in mean weight diameter of water-stable aggregates of a Vertosol under native grass, which decreased from 0.54 to 0.26 mm after 64 years of cultivation. Consequently, the saturated hydraulic conductivity decreased from 14.4 mm/h to only 2.5 mm/h. Connolly et al. (1997) also observed a close relationship between hydraulic conductivity and SOC in the top layer of a number of cultivated soils in southern Queensland.
Cook et al. (1992) found that macroaggregates [is greater than] 0.25 mm in a Vertosol decreased rapidly, from 55% under natural vegetation to 30% after 64 years of conventional cultivation, as did the loss of SOC. Similarly, Chan et al. (1988) found that a Vertosol that under native pasture had 60% macroaggregates [is greater than] 0.5 mm contained only 3% of this size aggregates after 45 years of mechanical cultivation for cereal cropping, apparently due to loss of active C pool (Chan 1997).
The severity of mechanical cultivation (also animal treading, mechanical shearing, and earth works) on aggregate breakdown is felt most in non-swelling clay soils, especially at low soil organic matter content (Dalal and Bridge 1996). In swelling clay soils, such as Vertosols, other aggregate stabilising agents such as clay content and electrolyte concentration, and aggregate dispersing agents such as exchangeable sodium (Dalal 1989; Dalal and Bridge 1996), primarily affect aggregate stability. The role of organic matter in aggregate stability in these soils is ambiguous since it is interrelated with, as well as affecting, other aggregate stabilising agents (Prebble 1987; Dalal 1989).
Cation exchange capacity (CEC)
In many soils, especially with low activity clays, organic matter contributes significantly to the total CEC. Moody et al. (1997) observed that the effective CEC of Ferrosols was significantly correlated with both clay content and total organic C content. Loss of SOM from low activity clay soils, especially Ferrosols and Kandosols, seriously reduced their CEC and increased their zero-point charge ([pH.sub.0]). For example, Gillman (1984) showed that decline of SOC from 4% to 1% resulted in an increase in [pH.sub.0] from 3.9 to 4.2. If the pH of that soil has also declined from pH 6 to pH 5, the loss in organic C from 4% to 1% would greatly reduce (10-40%) its pH-dependent CEC (Moody 1994).
The quantity and quality of organic matter is associated with the activity and abundance of different organisms in soil (Pankhurst 1997). The microbial activity and abundance is affected by (1) reducing the amount and diversity of litter input; (2) decreasing population size, species composition, and diversity of saprophytes and detritivores; and (3) reducing the time interval between litter inputs, and changing the site of decomposition from the soil surface to within the soil. Of these decomposers, bacteria account for almost 70% of organic matter turnover and N cycling in soil since they contain diverse metabolic capabilities and utilise many sources of energy and C in soil (Swift et al. 1980). The soil macrofauna--earthworms, termites, and ants--may account for a further 10-15% in forest and grassland soils but much less in cultivated soils (Curry and Good 1992), presumably due to reduced quantity and quality of SOM in the latter.
Currently, there is only circumstantial evidence of direct association between the quantity and quality of organic matter and soil biodiversity. For example, Yeates and Bird (1994) found that the pasture soil and the native shrub soil contained 30% and 80%, respectively, more nematode taxa than the adjacent cultivated wheat soil. The latter soil contained less SOM than the pasture and the native shrub soils. However, direct tillage and monoculture practices also affect nematode taxa distribution in soil (Gupta and Yeates 1997). There is a need, therefore, to quantify the relationship between soil biodiversity, organic matter quantity and quality, and the processes of organic matter decomposition and accumulation in soil.
Carbon emissions to the atmosphere
Loss of soil organic C in rainfed agricultural soils inevitably leads to predominantly [CO.sub.2] emission to the atmosphere (Gifford et al. 1992). Under anaerobic conditions, losses of soil organic C as [CH.sub.4] and volatile acids also occur (NGGIC 1999a).
Carbon dioxide-equivalent ([CO.sub.2]-e) emission estimates due to land use changes have 3 major components: area of change, age of activity, and rate of change in C per unit area. Each of these components currently has considerable uncertainty (Carter et al. 1998; Dalal and Carter 2000). The age and area of C changes could potentially be mapped at high resolution from LANDSAT satellite imagery, which dates from the early 1970s. To date, most mapping has been used to determine the current (after 1990) land use or potential use rather than providing spatially explicit time-series data.
At least 20 years of cropping history is needed to quantify the area and timing of land use change so the majority of soil organic C changes can be accounted for (Carter et al. 1998). For many land use changes, delayed release of soil organic C continues at a rate that declines exponentially with time; so a large proportion of emissions occur within a 20-30-year time frame. However, the initial soil organic C amounts and rates of decline are poorly known for most land use changes, especially for the whole soil profile or at least for the top 1 m depth, in most of the Australian cereal belt. This leads to great uncertainty of 75-80% around a potentially large amount of carbon in the national inventory of [CO.sub.2]-e emissions (NGGIC 1999a, 1999b).
Soil organic C decline under cultivation is better estimated with rundown curves established for a number of soil types within one zone of similar climate (Dalal and Mayer 1986b). However, to estimate C losses from all cultivated lands it is necessary to generalise these relationships across climate and clay content and then apply these relationships to areas of cultivation of known age. The lack of suitably detailed soil attribute maps, and cultivation age maps makes this process difficult.
Table 7 summarises the soil organic C loss from the 0-1.2 m depths from 6 soil series, which varied in clay content from 18% to 72%, and period of cereal cropping from 20 to 70 years. Soil C loss varied from 194 to 1015 kg/ha.year. It was inversely related to the clay content of these soils. Heenan et al. (1995) reported a soil C loss of 400 kg/ha.year for a Red Chromosol with 29% clay content, from the 0-0.1 m layer, which would account for [is greater than] 66% from the whole soil profile, according to the soil C loss and clay content relationship derived from Table 7. Unfortunately, there is hardly any information available for the soil C loss from deeper soil depths in cultivated Australian soils.
Table 7. Estimates of soil carbon loss and [CO.sub.2]-equivalent emissions from the whole soil profile (0-1.2 m depth) in the Australian cultivated soils used for cereal cropping R. C. Dalai (unpubl. data). Except for Langlands-Logie, the relationship between the organic C loss rate (y) and clay content (x) was described by the equation: y = -55.4 [+ or -] 40.2 [+ or -] 191.2 [+ or -] 12.7 (100/x), r = 0.98; since the intercept is not significantly different from zero the equation, was reduced to: y = 175.6 [+ or -] 5.9 (100/x) Soil Clay Total organic C in Organic C loss content uncultivated soils (kg/ha.l.2 m.year) (%) (kg/ha. 1.2 m) Waco clay 72 97715 [+ or -] 4144 194 [+ or -] 59 Thallon clay 59 71553 [+ or -] 2425 322 [+ or -] 66 Langlands-Logie 49 104065 [+ or -] 8602 882 [+ or -] 143 Cecilvale 40 94558 [+ or -] 3635 377 [+ or -] 135 Billa Billa 34 80223 [+ or -] 6915 508 [+ or -] 177 Riverview 18 76507 [+ or -] 5746 1015 [+ or -] 229 Soil [CO.sub.2]-e emission (kg/ha.year) Waco clay 711 [+ or -] 216 Thallon clay 1181 [+ or -] 242 Langlands-Logie 3234 [+ or -] 524 Cecilvale 1382 [+ or -] 495 Billa Billa 1863 [+ or -] 649 Riverview 3722 [+ or -] 840
From the soil C loss and clay content relationship given in Table 7, an estimate of [CO.sub.2]-e emissions from the Australian cereal belt can be made. Assuming that the weighted mean clay content (0-0.1 m depth) of 20 Mha of the cropping soils is 24%, comprising 3 Mha of fine-textured soils (45% clay), 12 Mha of medium-textured soils (25% clay), and 5 Mha of coarse-textured soils (15% clay), with an estimated annual soil C loss of 730 [+ or -] 23 kg/ha, then 2.6 [+ or -] 0.08 t [CO.sub.2]-e/ha.year or 52 Mt [CO.sub.2]-e would be emitted annually. Almost half of this loss occurs from the top 0-0.1 m depth (Dalal and Mayer 1986b). This can be compared with a net emission of 496.1 Mt [CO.sub.2]-e in 1997 for all sectors, which comprised 94.2 Mt [CO.sub.2]-e from the agricultural sector and 64.8Mt [CO.sub.2]-e from the forest and grassland conversion (NGGIC 1999a, 1999b). Therefore, more than half of the [CO.sub.2]-e from the agricultural sector is due to soil organic C loss from the Australian cereal belt.
The size of these organic C emissions and their effects on soil organic C might be determined on a statistical basis by detailed sampling of an adequate number of paired sites of known age. To produce maps of emissions, calibrated models that account for plant growth, climate, land management, and effect of soil properties on organic C loss rates would need to be combined with maps describing the temporal and spatial nature of land use change and maps of soil properties (Dalal and Carter 2000).
It may be difficult to separate the results of historic land use change from climate change (changes in temperature and precipitation), and feedback effects from enhanced biomass production from increased [CO.sub.2] concentration in the atmosphere ([CO.sub.2] fertilisation). A combination of measurement and process modelling will be needed to differentiate the effects of these processes.
Organic matter restorative practices
Management practices that lead to improvement in plant biomass production would likely lead to increased C inputs and hence to increased organic matter in soil. Sustaining or enhancing soil productivity depends at least partly on soil and crop management practices that maintain or increase soil organic matter. Management practices such as pasture leys (Hossain et al. 1996), crop rotation and fertilisers (Cogle et al. 1995), no till and crop residue retention (Dalal 1989), and manure applications may influence organic matter levels in soil. These management practices include use of improved pasture species in leys in ley-crop rotations, modification of stocking rates in ley pasture grazing, plant residue retention (stubble retention), and reduced tillage (minimum or no-till with controlled traffic practices), soil fertility management (legumes and fertiliser application), sustainable land management practices to reduce land degradation, C input from external sources such as manures, and changes in fire frequency.
In the temperate cereal belt, the beneficial effects of a pasture phase in increasing soil organic carbon level and maintaining soil structure are well documented (e.g. Greenland 1971; Tisdall and Oades 1980; Russell and Williams 1982). Using the data from the permanent rotation experiment at the Waite Institute (established 1926), Russell and Williams (1982) and Grace et al. (1995) demonstrated that soil organic C tended to increase with the increasing frequency of pasture in the wheat-pasture rotation (WP). Whereas the wheat-fallow rotation reduced soil organic C by [is greater than] 60% after 68 years (1925-1993), permanent pasture since 1950 has almost restored the original level (30 t C/ha in 1993 v. 34 t C/ha in 1925 in 0-0.1 m depth, assumed bulk density of 1.25 Mg/[m.sup.3]). Grace et al. (1995) concluded that a 2 wheatcrops: 4 years of pasture phase is required to maintain the long-term steady-state soil organic C level in the Urrbrae loam (Red Chromosol). Additional benefits due to increases in soil organic C were: an overall increase in porosity; a decrease in bulk density; and an increase in aggregate stability (Greenland 1971; Tisdall and Oades 1982; Chan et al. 1992; Tisdall 1996; Bell et al. 1999).
In semi-arid north-eastern Australia, Chan et al. (1997) found that organic C in degraded Vertosols (Typic Chromustert) increased from 6.5 t C/ha to 8 t C/ha (0-0.05 m depth, with assumed bulk density of 1 Mg/[m.sup.3]) after 4 years by restoration of pasture with barrel medic (Medicago truncatula) and Mitchell grass (Astrebla lappacea).
Increase in organic C concentrations due to the inclusion of legumes in grass pastures is mainly from increased biomass C input due to [N.sub.2] fixed by the legume, which is subsequently utilised by the associated grass species, with increased C input to the soil (Bruce 1965; Dalal et al. 1995).
The native high N status of soils cleared from Acacia harpophylla and other leguminous native vegetation has been used to establish improved grass pastures (Graham et al. 1981), e.g. buffel grass (Cenchrus cilliaris), Chloris gayana, Paspalum spp., and Panicum spp. For example, a Vertosol cleared of Acacia harpophylla maintained original organic C concentrations under Panicum maximum for at least 11 years. In fact, there was a general trend for increase in organic C with the period of pasture growth (Skjemstad et al. 1994). However, buffel grass or green panic pasture productivity declines after a certain period when N supply becomes limited due to most of the N being tied up in roots and standing biomass (Robertson et al. 1997). Nitrogen application or inclusion of a legume restores pasture productivity (Graham 1987) and, therefore, may further increase soil organic C concentration.
On a Vertosol, Whitehouse and Littler (1984) observed substantial increases in organic C after 2-4 years of lucerne + prairie grass pasture; organic C increased from 18 t C/ha to 20.5 t C/ha (assumed bulk density, 1 Mg/[m.sup.3]) after 4 years of pasture growth (0-0.15 m depth). Similarly, Dalal et al. (1995) measured a rate of increase in organic C of 650 kg C/ ha.year in a Vertosol under grass + legume pasture (purple pigeon grass, Setaria incrassata; Rhodes grass, Chloris gayana; lucerne, Medicago sativa; and annual medics, M. scutellata and M. truncatula) for 4 years; most of the increases in organic C occurred in the top 0.05 m layer. Skjemstad et al. (1994) reported an increase of 550 kg C/ha.year in a Vertosol under Rhodes grass in a similar environment. Increase in organic C under pasture was attributed to the much higher input of C through the grass root biomass (10 t root dry matter/ha.year) compared with continuous wheat cropping (2 t root dry matter/ha.year); above-ground residue inputs were essentially similar (2.5 t residues/ha.year) (Dalal et al. 1995). Chan (1997) observed similar effects of pasture on organic C concentrations in a Vertosol in northern New South Wales. He also found that almost 70% of the organic C increases could be attributed to the increase in particulate organic C.
Wheat cropping after the pasture phase resulted in organic C decline, but even after 4 years it remained above the organic C concentration in the continuous wheat cropping treatment (Dalal et al. 1995). Also, the decreasing organic C trends during the cropping phase were slower than the relatively faster increase in organic C in soil under the grass + legume phase earlier, presumably due to the increased plant C input from higher crop yields compared with continuous wheat cropping.
Whereas a grass + legume pasture phase had a positive effect on soil organic C, 2-year rotations of lucerne--wheat and medic--wheat after 8 years under conventional tillage had a negligible effect on organic C concentrations (Dalal et al. 1995). The plant C inputs, especially from root biomass, in these treatments were [is less than or equal to] 50% of those from the grass + legume pasture. Therefore, a relatively small amount of C input and the rapid rate of turnover of legume C in these short legume rotations do not increase organic matter in these Vertosols. This is confirmed by Holford (1990) who observed no increase in organic C after 4 years of lucerne pasture on a Vertosol in northern New South Wales.
In a subtropical Ferrosol, Bell et al. (1997) measured similar amounts of organic C in the top 0-0.3 m depth after 4 years of 2 grass pastures (Pennisetum clandestinum and Chloris gayana) and in the adjoining soil under continuous cultivation and cropping, although the bulk density was significantly lower under the grass pastures. However, Cogle et al. (1995) found that 8-10 years of grass pasture established after clearing a tropical Ferrosol from woodland reduced organic C loss from 22 t C/ha under cultivation to 13 t C/ha under grass pasture. Organic C loss was further reduced to only 8 t C/ha when a legume (Stylosanthes hamata) was also grown in the grass pasture for 4 years.
In summary, significant increases in organic C or retarding organic C loss from Australian cereal belt soils can be achieved by crop rotations that include a grass + legume pasture phase. However, the optimal duration for the pasture phase in relation to the cropping phase is less well known for the tropical and subtropical than for the semi-arid areas in the temperate region (Grace et al. 1995).
Cereal-grain legume and other crop rotations
Little information is available on the effect of different crops on long-term soil organic C under continuous cropping and crop rotations on different soils. Dalal et al. (1995) found that the 2-year rotation of chickpea (Cicer arietinum)--wheat after 8 years under conventional tillage had a negligible effect on organic C concentrations. Skjemstad et al. (1994) observed that organic C loss from a fertile Vertosol was faster under black gram (Vigna mungo) than under sorghum cropped for 11 years, a consequence of lower plant C input from the former (0.2 t/ha.year of above-ground dry matter) than from the latter (0.5-0.75 t/ha.year of dry matter), and possibly similar C input from the below-ground plant biomass. Similarly, Grace et al. (1995) reported that a wheat-pea rotation contained less organic C than continuous wheat (16 t C/ha v. 19 t C/ha, assuming a bulk density of 1.25 Mg/[m.sup.3]) in the 0-0.1 m layer after 68 years of cropping.
Heenan et al. (1995) found that a lupin-wheat rotation on a Red Chromosol after 11 years lost a much smaller amount of soil organic C (250 kg C/ha.year) than continuous wheat (400 kg C/ha.year) when both were managed under conventional tillage and stubble burning practice.
Crop rotations involving cereals and crops other than grain legumes such as canola have not been in practice in long-term experiments. Grace et al. (1995) found that a wheat--oats--fallow rotation in the long-term trial at Waite Agricultural Research Institute had soil organic C lower than continuous wheat. It can be assumed that disease-breaking crops such as canola, which may enhance cereal production, could result in higher plant C inputs and hence a likely increase in soil organic C. Unfortunately, no long-term studies exist to validate this assumption.
No-till and crop residue retention
No-till (NT) and crop residue retention reduce the loss of organic matter in soil. For example, Dalal (1989) measured higher organic C contents in a Vertosol under NT and crop residue retention than under conventional till (CT); a positive interaction of tillage practice x crop residue x N fertiliser application was observed after 13 years. These increases in organic C occur in the top 0-0.025 m or 0-0.05 m layers in Vertosols, after 18 years of NT practice (Dalal et al. 1991, 1995). Thus, NT practice enhances SOM stratification even in a Vertosol, although total amounts of SOM may be similar to that in the CT practice.
Similarly, in Ferrosols, NT practice enhanced or reduced the decline in organic C concentrations in the top layers (Table 8). Besides providing C inputs, surface residue cover also reduces raindrop impact and enhances water infiltration, which may increase plant biomass production, and reduce soil erosion loss from the surface with higher organic C content. On the other hand, Fettell and Gill (1995) reported no effect of different tillage practices including NT practice and/or stubble management practice on a Red Chromosol after 14 years of cereal cropping, although fertiliser N application significantly increased soil organic C. This has been confirmed on other soils by Haines and Uren (1990) and Carter and Mele (1992). These results were attributed to the fact that cereal grain yields and hence plant C input, especially below ground, were essentially similar under NT and stubble retention compared with conventional tillage practice.
Table 8. Effect of no-till (NT) for 4-5 years and conventional till (CT) on organic C contents of a tropical Ferrosol Soil depth Organic C (t/ha) (m) CT NT 0-0.05 6.2 9.5(*) 0.05-0.10 5.4 8.1(*) 0.10-0.20 8.5 12.3 (*) Significant difference in organic C between NT and CT.
A collation of soil organic C contents from field trials under NT (3-19 years duration) on the coarse-textured soils (Red Chromosols and Kandosols) around the Australian cereal belt revealed that significantly higher soil organic C levels than for conventional tillage were found only in the wetter temperate areas ([is greater than] 500 mm) and the differences were restricted to the top 0.025 m or 0-0.1 m depth (Chan et al. 1998). In Wagga Wagga, New South Wales, with an annual rainfall of 550 mm, NT combined with stubble retention was required to slow the decline in soil organic C level under continuous cropping on a Red Chromosol (Heenan et al. 1995). The lack of a positive increase in soil organic C due to NT in other drier areas, [is less than] 500 mm rainfall in temperate regions (Fig. 6) and [is less than] 700 mm rainfall in the subtropical regions (Dalal et al. 1995; Bell et al. 1997), was probably a reflection of a number of factors, namely low crop yield (due to low rainfall), partial removal of stubble by grazing, and the high decomposition rate (due to the high temperature). There is evidence suggesting that under continuous cropping in the drier areas, the soil organic C level continues to decline even under NT (Dalal et al. 1995; Chan et al. 1998).
However, Kern and Johnson (1993) estimated that NT would lead to 28% more soil organic C retained than conventional tillage in the soils of the USA (Fig. 6); most of the gains are likely to occur in the cool and moist regions.
Fertilisers and manures
A direct consequence of organic matter decline is the concomitant decreases in nutrient supply from soils, especially that of N (Dalal and Mayer 1986a; Moody 1994). This N loss can be remedied by application of N (and other nutrients) from fertilisers, manures, and other sources.
At Taree, South Australia, soil organic C declined by about 10% over a 10-year period from a Red Chromosol under continuous wheat cropping and with no fertiliser application, whereas N fertiliser application at the rate of 80 kg N/ha was found to prevent such decline (Carter et al. 1993). On the other hand, N fertiliser applied at a rate of 120 kg N/ha did not increase soil organic C under continuous wheat cropping on a Red Chromosol in Wagga Wagga, New South Wales, probably due to enhanced acidification of the soil (Chan etal 1992).
Dalal (1992) observed that N fertiliser application of 69 kg N/ha.year for 22 years to a Vertosol reduced the soil N loss compared with nil fertiliser application. Even the N rates exceeding the crop requirements could not reverse soil N decline, although the high N rates increased nitrate-N accumulation in the soil profile. For the first 13 years, however, Dalal (1989) observed a positive effect on organic C from the combined NT practice, residue retained, and N application treatment.
Similarly, Russell (1981) and Skjemstad et al. (1994) found that N applications (100 kg N/ha.year) to sorghum (Sorghum bicolor) grown on a fertile Vertosol for 11 years did not arrest soil organic C decline. Even on a fertility-depleted Vertosol, Dalal et al. (1995) found no significant increase in organic C from a 75 kg N/ha.year application for 8 years compared with nil N fertiliser application. This was primarily because C inputs to soil through plant biomass, especially root biomass, were essentially similar in both treatments. Again, the N application resulted in nitrate-N accumulation in the soil profile.
On Ferrosols, Chromosols, and Kandosols, fertiliser application may result in increased organic matter in soil, although Cogle et al. (1995) did not report any differences in organic C concentrations between the fertilised treatments (80 kg N + 23 kg P/ha.year applied to sorghum for 4 years) and the unfertilised treatments on a Red Kandosol. Moody (1994) reported 60% decline in organic C in cultivated and cropped Ferrosols compared with the virgin Ferrosols, in spite of the fact that many of the cultivated Ferrosols had been fertilised.
It appears that in the semi-arid environment, C inputs from annual crops are limited by the availability of water (rainfall); thus, insufficient plant biomass is produced to meet C lost from soil even though crop yields are significantly increased with fertiliser applications.
In pastures where legume production is limited by P and S supply (Russell and Williams 1982), application of fertilisers such as superphosphate has resulted in increased biomass production and increased soil organic N (Russell 1960; Watson 1969) and, hence, increased organic C in soil. For example, Russell and Williams (1982) estimated an increase of organic N of 50 kg N/ha.year from a mean superphosphate application of 15 kg P/ha.year. With an estimated C :N ratio of 12 for the resultant SOC, an increase of 600 kg C/ha.year is expected. This value is similar to that measured by Dalai et al. (1995), 650 kg C/ha.year, for the legume+grass pasture at Warra, southern Queensland.
Manure applications increase organic C in Vertosols (E. Powell, pers. comm.), Ferrosols (M. Bell, unpubl, data), and other soils, although long-term benefits of manure applications in arable cropping in the Australian soils have not been measured. High-lignin amendments such as farmyard manure, which are more recalcitrant to decomposition than plant residues, usually result in higher soil organic C concentrations (Poulton 1995).
Benefits of increased organic matter in soil
Increased SOM reverses the adverse consequences of organic matter loss from soil. Lal (1998) has provided detailed literature on benefits of SOM for agricultural sustainability, including improved soil structure, nutrient retention, reduced soil erosion, and improved soil fertility. Williams and Lipsett (1961) and Russell and Williams (1982) summarised the Australian research on the benefits of increased SOM on soil properties and agricultural productivity.
However, there is limited information available on the increase in different C pools of SOM and their beneficial effects on soil properties and agricultural sustainability in the Australian cereal belt. Moreover, their effects on soil microbial biodiversity (Pankhurst et al. 1995) and herbicide and pesticide retention and their degradation (Briggs 1981; Kookana et al. 1998; Ahmad 1999) in soil are almost unknown.
Greenhouse C sink
Assuming that the rates of [CO.sub.2]-e losses from cultivated soils, calculated from Table 7, can be reversed by using restorative practices, the potential sinks available in these soils would be 1040 Mt [CO.sub.2]-e (52 Mt [CO.sub.2]-e/year from the soil profile over a 20-year period), that is, more than 2 times the current total annual [CO.sub.2]-e emissions from Australia. Therefore, SOM provides an attractive greenhouse C sink option because it is a large sink and provides long-term C storage (AGO 2000). Moreover, increased C sequestration in soil, hence increased SOM, also provides associated environmental and production benefits (Lal 1998).
Management practices mentioned above, that is, ley pasture--crop rotation, cereal crop rotation with grain legumes or other disease-break crops such as canola, no-till, stubble retention, and fertiliser management, and organic residues and manure application, may contribute to an increased C sink in soil (Table 9). A combination of the above management practices may provide a synergistic effect in increasing C sequestration in soil. For example, Dalal (1989) reported a significant increase in SOM in a Vertosol after 13 years only when no-till practice, stubble retention, and fertiliser N were practised in combination. It is estimated that only 23% of the potential sink of 50 Mt [CO.sub.2]-e/year in the cultivated soils of the Australian cereal belt, about 11.8 Mt [CO.sub.2]-e/year over 20-year period, can be realised with the application of improved agricultural management practices (Table 9). The remaining 77% potential soil carbon sink can be utilised by increasing the cropping intensity, high plant and animal biomass inputs, and if increasing cropping areas are used for growing perennial crops, agroforestry, and forestry crops.
Table 9. Potential greenhouse C sink clue to management practice in the rainfed cultivated Australian cereal belt, excluding methane and nitrous oxide emissions/absorption Management practice Area Plant input (Mha) (t C/ha.year) No-till, stubble retained(A) 2.5 1.6 Improved ley pasture(B) 21 4.2 Sugarcane trash retention(B) 0.4 4.2 Cotton(B) 0.4 1.7 Others (manure application)(C) 0.1 0.5 Total 24.4 Management practice [CO.sub.2] sequestration Mt [CO.sub.2]/year No-till, stubble retained(A) 0.2 Improved ley pasture(B) 10.8 Sugarcane trash retention(B) 0.2 Cotton(B) 0.1 Others (manure application)(C) 0.1 Total 11.4 Management practice Mt [CO.sub.2] per 20-year period No-till, stubble retained(A) 4.0 Improved ley pasture(B) 216 Sugarcane trash retention(B) 4.0 Cotton(B) 2.0 Others (manure application)(C) 2.0 Total 228.0 (A) In regions with annual rainfall >500 mm in temperate areas (Chan et al. 1998) and >700 mm in subtropical and tropical areas (Dalal 1989; Dalal and Carter 2000). (B) Adapted from AGO (2000), NGGI (1999c), and Dalal and Carter (2000). (C) Calculated from AGWISE (1999) assuming annual 4 t/ha manure application.
Current agricultural practices for food and fibre production in the Australian cereal belt have inevitably led to the loss of both the quantity and quality of organic matter from the soils. Land degradation (decline in soil fertility, soil structure, and soil microbial diversity, and increase in soil erosion and salinity and sodicity) at least partly due to organic matter loss is widespread. Contribution to the greenhouse [CO.sub.2]-e emissions from the loss in organic matter from these soils is estimated to be in excess of 50 Mt [CO.sub.2]-e/year. Apparently, the rate of organic C loss is inversely related to the soil's clay content.
Sustainable cropping and land management practices are, therefore, essential for maintaining crop productivity while maintaining or enhancing organic matter in soil (Cogle et al. 1995; Dalal et al. 1995; Bell et al. 1997; Chan et al. 1997), and increasing the soil's C sink to mitigate greenhouse gas emissions. Except for the positive effects of grass + legume pastures or fertilised grass pastures in both temperate and subtropical regions, very little information is available about other land and crop management practices to enhance organic matter in soil. For example, the positive effect of no-till and stubble retention practices is apparently limited by the amount of C inputs, which is limited by rainfall, and rarely exceeds SOM losses from decomposition and other soil degradation processes.
Since soil organic C is heterogeneous in nature, the significance and functions of its various components are ambiguous. Therefore no attempt was made to recommend a soil organic C level or its components that will be sufficient to sustain and enhance the biophysical and chemical roles ascribed to organic matter in soil. Some commercial soil testing services and States' analytical laboratories provide a broad classification of low, medium, and high levels of soil organic C (Bruce and Rayment 1982; Peverill et al. 1991; Purdie 1995). It is essential that the relationship between levels of total soil organic C or its components and the most affected soil property(ies) be established and then quantified before the concentrations or amounts of soil organic C and/or its components can be used with confidence for the purpose that its use is intended. Furthermore, there is scant information on the interactions between pesticides and herbicides, soil microbial diversity, soil-borne diseases, and soil organic matter. Thus, the environmental role of organic matter in soil requires intensive investigation.
The challenge for sustainable land management is to balance the needs for the production of food and fibre and other land uses with the necessity to maintain sufficient land cover to minimise the adverse impact of soil degradation processes while maintaining or enhancing the organic C sink in soil to mitigate greenhouse gases emissions. Until recently, both food and fibre production systems involved intensive cultivation, which resulted in lack of plant residue soil cover, often utilising the native soil fertility, which led to declining yields, and mainly using monoculture systems, which again have resulted in decreasing yields and hence lower C inputs to the soil. Studies should be undertaken to quantify the size of the above-ground biomass returned to the soil, the size of the below-ground C including roots and charcoal, and the rates of turnover (C fluxes) of both below-ground and above-ground C pools. Since C components differ in their turnover rates, understanding of the function and nature of these components through improved techniques is required. This approach should also lead to improved simulation of soil organic C dynamics and better prediction of the short-and long-term consequences of land and vegetation management practices and environmental pollutants on SOM management while enhancing soil C sinks in the Australian cereal belt.
We thank Cecelia McDowall and Michelle Perry for assistance with literature search. We also thank the reviewers and Weijin Wang, Phil Moody, and John Carter for their comments and suggestions.
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Manuscript received 5 May 1999, accepted 10 October 2000
(*) This review is one of a series commissioned by the journal's Editorial Advisory Committee.
R.C. Dalal(A) and K. Y. Chan(B)
(A) CRC for Greenhouse Accounting, Department of Natural Resources, Resource Sciences Centre, 80 Meiers Road, Indiooropilly, Qld 4068, Australia; email: Ram.Dala@ dnr.qld.gov.au
(B) Wagga Wagga Agricultural Institute, NSW Agriculture, PMB, Wagga Wagga, MSW 2650, Australia.
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|Author:||Dalal, R. C.; Chan, K. Y.|
|Publication:||Australian Journal of Soil Research|
|Article Type:||Statistical Data Included|
|Date:||May 1, 2001|
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