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Phosphorus accumulation and other changes in soil properties as a consequence of vegetable production, Sydney region, Australia.


Vegetable production is common in many peri-urban areas around the world and is important from many agricultural, historical, cultural, social, economic, and environmental perspectives. For example, vegetable production within the Sydney region has a history as long as European settlement (about 200 years) and is a main supplier of fresh vegetables consumed by Sydney's 4 million inhabitants. Total agricultural production within the Sydney region is estimated to be worth more than AU$1 billion annually (Gillespie and Mason 2003). Fresh market vegetables comprise 40% of total agricultural production within the Sydney region (Gillespie and Mason 2003), which highlights its importance to the New South Wales economy.

It has been reported in other parts of the world that soils used for vegetable growing undergo pedological changes and transform into characteristic soil types, namely Fimic Anthrosols (Zhang et al. 2003). Apart from heavy metals (Jinadasa et al. 1997), little information is available on the extent and magnitude of changes in soil properties as a consequence of vegetable production in Sydney. This information is useful for developing indicators for measuring the sustainability of vegetable production in the Sydney region and tools for informing future land use and management in these areas. Changes in the nutrient status, particularly with respect to nitrogen (N) and phosphorus (P), of agricultural soils within metropolitan catchments can influence the water quality and ecology of Sydney's waterways. This is highlighted by the toxic blue-green algal blooms that occurred in the Hawkesbury-Nepean River during the early 1990s, which were partly attributed to non-point pollution from agriculture. Catchments characterised by cropping land uses have high rates of soil loss and are important sources of diffuse sediment in the waterways in the Sydney area (Erskine et al. 2002). However, improved understanding of the link between land use and water quality is necessary before strategies for alleviating negative impacts arising from agriculture within the Sydney region can be developed.

The objectives of this paper are to: (1) report the results of a survey of vegetable growing soils in the Sydney region, (2) compare the physical and chemical properties of vegetable soils with those of adjacent unfarmed 'reference' soils, and (3) discuss the potential environmental implications of the results, particularly with respect to P accumulation and nutrient management.

Materials and methods

Site selection

Soils used for vegetable production in the Sydney region fall into 5 major soils types, namely Rudosols (active), Rudosols (old), Chromosols, Kandosols, and Hydrosols (Table 1) (Isbell 1996). Although the Hydrosols represent some of the oldest continuously used vegetable soils in Australia, they are not widespread and only represent a very small proportion of the soils used for growing vegetables in Sydney.

The soil survey was designed with the assistance of NSW Department of Primary Industries' agricultural extension officers and leading local growers to identify vegetable farms in the Sydney region which: (i) were located on representative soil types (Table 1); (ii) had a history of vegetable production > 10 years (vegetable soil); and (iii) had an unfarmed reference area adjacent to the site used for vegetable production (reference soil).

Many of the farms included in this investigation were also used by Jinadasa et al. (1997) in their studies on heavy metals. In all cases covering the different soil types, a range of common vegetable crops had been grown and there was no evidence suggesting that particular crops were being grown on particular soil types. At each site, the reference soil was located on the same soil type but had not been used for cropping and was ideally under natural vegetation or at least used for extensive grazing. The vegetable farms located on the Hydrosols were exceptional because these sites are now surrounded by highly urbanised areas, and consequently, it was not possible to confidently identify an undisturbed reference site; therefore, they were not included in the present investigation. In total, 34 and 27, 'vegetable' and 'reference' sites were sampled, respectively (Table 1). Some vegetable sites did not have adjacent reference sites.

Soil sampling

Sampling was carried out in 2004 after harvesting of the vegetable crops. At each site, 6 soil cores (50 mm in diameter) were collected (0-0.30 m) from both the vegetable and reference sites. At each vegetable site, a pair of soil cores was taken at random from within each of 3 randomly selected vegetable beds (approx. 2 m wide). Each soil core was stratified into 0-0.10, 0.10-0.20, and 0.20-0.30m depths. Each pair of cores from the same bed and depth was bulked and placed into a plastic bag to form triplicate composite samples for each depth at each vegetable site. At each reference site, 3 sampling points were selected at random and at each point, 2 cores (0-0.30 m) were collected. These were sectioned and bulked in the same way as those from the adjacent vegetable site.

In addition, a spade was used to collect composite, intact, surface soil samples (0-0.10m) from 5 randomly selected locations at each vegetable and reference site. These were dried at 36[degrees]C and stored for subsequent assessment of soil structural stability.

Sample preparation

Each bag containing the composite soil cores was weighed and gently

broken up into fragments before air drying at 36[degrees]C for a week or until constant mass was achieved. The air-dried soils were ground to < 2 mm for subsequent chemical analyses.

Soil chemical analyses

The < 2 mm vegetable and reference soil samples were analysed for pH, electrical conductivity, (EC), total carbon (C), total nitrogen (N), exchangeable cations, bicarbonate-extractable P, Ca[Cl.sub.2]-extractable P, and acid-extractable P. Total carbon and nitrogen were determined by dry combustion using LECO (Nelson and Sommers 1982). Exchangeable cations were determined using the method of Gillman and Sumpter (1986); [MATHEMATICAL EXPRESSION NOT REPRODUCIBLE IN ASCII.], EC, bicarbonate-extractable P (Colwell-P), and 0.005 M Ca[Cl.sub.2]-extractable P (Ca[Cl.sub.2]-P) were determined following Rayment and Higginson (1992). The method used in measuring Ca[Cl.sub.2]-P in our ease involved an 18-h shaking in 0.005 M Ca[Cl.sub.2] (Rayment and Higginson 1992), which was longer than for the methods recommended for assessing environmental impact of P (e.g. Reddy et al. 1980; Jaszbcrenyi and Loch 1997). Environmental impact assessments usually employ a shorter shaking time of 1-2 h in 0.01 M Ca[Cl.sub.2] solution. Acid-extractable P (total P) was determined following USEPA (1996), which involved digestion with a mixture of nitric acid and hydrochloric acid and hydrogen peroxide. Although the procedure was not expected to solubilise those elements bound in silicate structure, it is likely to serve as a useful indicator of the total P pool that could become 'environmentally' available in the soil (USEPA 1996).

Soil structural stability

Soil clods (about 10mm in diameter) were gently separated from the air-dried composite surface samples of the vegetable and reference soils. About 20 g of clods was weighed and wet-sieved for 10 min at 38 mm stroke length and 30 strokes/min using 2 sieves with apertures of 2 mm and 250 [micro]m, respectively, in a 2-L cylindrical container. The proportion of water-stable aggregates > 2 mm and > 250 [micro]m diameter was calculated from the mean of 2 measurements per sample.

Statistical analysis

Preliminary analyses of the results for soil chemical properties indicated that pH, Al, Ca, and C did not need any transformation prior to analysis. However, the data for EC, K, Mg, Na, P, and N required log-transformation to normalise variance. A mixed model was then developed to explore the relationship between land use, soil type, sampling depth, and sites, and their effects on the soil parameters of interest (Eqn 1):

Soil parameter = (land use) + (soil type) + (depth) + (land use * soil) + (land use * depth) + (depth * soil) + (depth. land use, soil) + sites + error (1)

The effects of the italicised terms were assumed to be random and normally distributed and the errors were assumed to have an auto-regressive correlation between depths. A residual maximum likelihood (REML) estimation was used to estimate all parameters. All analyses were undertaken using GENSTAT VI (VSN International Ltd 2003).



Typical of the soils of the Sydney region, the reference soils contained low concentrations of Colwell-P and Ca[Cl.sub.2]-P (Table 2). Mean Colwell-P in the topsoil (0-0.10m) was 15mg/kg and decreased to 7mg/kg at depth (0.20-0.30m). Likewise, mean Ca[Cl.sub.2]-P concentration in the surface (0-0. 10 m) of the reference soils was very low (0.07 mg/kg) and below detection limit at depth.

Total P, Colwell-P, and Ca[Cl.sub.2]-P concentrations in the vegetable soils were all significantly (P < 0.05) higher than the corresponding reference soil at all depths (Table 2). Significant land use x depth interactions were found for all 3 forms of P (Table 2). The ratios of total P, Colwell-P, and Ca[Cl.sub.2]-P in the vegetable soils over the reference soils at 0-0.30 m depth were all > 1, indicating accumulation of the various forms of P in the vegetable soils. Mean total P concentrations in the vegetable soils were 6, 6, and 5 times that of the Reference soils for the 0-0.10, 0.10-0.20, and 0.20-0.30 m layers, respectively (Table 2). Mean Colwell-P concentrations in the vegetable soils were 18, 44, and 19 times that of the reference soils for the 0-0.10, 0.10-0.20, and 0.20-0.30 m layers, respectively. Average Colwell-P concentrations in the 0-0.10 and 0.10-0.20 m layers of the vegetable soils were > 260 mg/kg (Table 2), whereas the mean concentration at depth (0.20-0.30 m) was 134 mg/kg. Mean Ca[Cl.sub.2]-P concentrations in the vegetable soils were 97, 239, and 123 times those of the reference soils for 0-0.10, 0.10-0.20, and 0.20-0.30m layers, respectively (Table 2). The greatest P accumulation was found at 0.104).20 m depth for all 3 forms of P. Moreover, the ratios were in the order total P < Colwell-P < Ca[Cl.sub.2]-P, suggesting accumulation of proportionally more labile forms of P.

There were significant (P < 0.05) differences in Ca[Cl.sub.2]-P concentrations between soil types, but not for total P and Colwell-P. There was no significant land use x soil type interaction for any of the P forms evaluated, indicating that the impact of land use on soil P concentration was similar for all soil types evaluated.

Regression analysis indicated that Colwell-P and total P concentrations in the surface soils (0-0.10m) across all soil types can be described by a linear relationship ([r.sup.2] =0.85, P < 0.001) (Fig. 1a). No significant relationship between Ca[Cl.sub.2]-P and total P across all soil types was determined (Fig. 1b).


Total carbon and total nitrogen

Significant land use x depth interactions were observed for both C and N. Significant differences (P < 0.05) in soil organic carbon concentrations were observed between the surface soils (0-0.10m) of the vegetable and reference soils, but not in the deeper layers (Table 3). Mean soil carbon concentration in the surface (0-0.10 m) of soils used for vegetable production was only 57.3% (16.5 v. 28.8g/kg) of the reference soils (Table 3).

Total N concentrations in the 0-0.10 m layer of the vegetable soils were 50% (P < 0.05) lower than those in the corresponding reference soils (Table 3). However, total N of the vegetable soil did not change with depth (P > 0.05) (Table 3).

Other soil chemical properties

The effect of land use on all other soil chemical properties evaluated was statistically significant (P < 0.05), with the exception of exchangeable Mg (Table 3). Compared with the unfarmed reference soils, the vegetable soils were higher in pH, EC, and exchangeable Ca, Na, and K, but lower in exchangeable Al (Table 3). The effect of soil type was significant (P < 0.05) for all chemical properties, except for pH and EC. The effect of depth was significant for all chemical properties except for pH, Al, and EC (Table 3).

Significant land use x soil type interactions were found for pH and exchangeable Al (Table 4). Mean pH of vegetable soils at all depths (0-0.30m) was significantly (P < 0.05) lower than the reference soils for the Rudosols (active) but significantly higher for the Kandosols. The effect of land use on soil pH was not significant for the other 2 soil types. The concentrations of exchangeable Al in the Kandosols used for vegetable production were significantly lower than those in the unfarmed reference soils.

Significant land use x depth interactions were found for exchangeable Ca and Mg. The concentrations of exchangeable Ca in the vegetable soils were significantly (P < 0.05) higher than those in the reference soils at all depths. The percentage increase was 50, 126, and 131% for the 0-0.10, 0.10-20, and 0.20-0.30 m layers, respectively (Table 3). The concentration of exchangeable Mg in the surface soils (0-0.10 m) was similar for the vegetable and reference soils, but was significantly higher in the vegetable soils in the 2 lower layers, i.e. 0.10-0.20 and 0.20-0.30 m.

Soil structural stability

Percentages of water-stable aggregates > 2 mm and > 250 [micro]m were both significantly (P < 0.05) higher in the reference soils than the vegetable soils (Table 5). The percentage of water-stable aggregates in the vegetable soils was close to zero for both size fractions.


Our results have revealed significant changes in the properties of soils as a result of vegetable production in the Sydney area. Of particular concern is the accumulation of P in the vegetable soils when compared to the unfarmed soils. Soils of the Sydney area are inherently low in P (Walker 1960) and many native plants have adapted to the low P status of the soils (Handreck 1997). The 3 forms of P evaluated in this study, total P, Colwell-P, and Ca[Cl.sub.2]-P, were used to represent the acid-extractable P pool (and so include the labile as well as much of the non-labile P), the concentration of plant-available or labile P, and the concentration of P in the soil solution (highly labile P), respectively. Bicarbonate-extractable P (Colwell-P) also provides a measure of the quantity component (q) of labile P (Dalai and Hallsworth 1976; Helyar and Spencer 1977; Holford 1997; Redding et al. 2002), and Ca[Cl.sub.2]-P measures the intensity component (I) of labile P (Moody et al. 1988, 1990). This information is useful when considering soil buffering capacity, P availability to plants, and the potential for environmental impacts as a consequence of the soil becoming saturated with P. The method used in measuring Ca[Cl.sub.2]-P in our ease involved a longer shaking time (18 h v. 1-2 h, respectively) but lower molarity of Ca[Cl.sub.2] (0.005 v. 0.01 M) than those recommended for assessing environmental impact of P. Although no research has been carried out to assess the difference in the quantity of P extraction under the 2 different sets of conditions, previous work by Moody et al. (1988) has indicated that Ca[Cl.sub.2]-P extracted using 18-h shaking with 0.005 M Ca[Cl.sub.2] is highly correlated with soil solution P. Generally, a Colwell-P concentration of 150mg/kg in the surface soil is considered adequate for vegetable production (NSW Agriculture 1997). Our results revealed mean Colwell-P concentrations nearly twice this threshold to a depth of 0.20 m (Table 2), suggesting that the quantity of available P in the vegetable soils was in excess of plant requirements. The effect of such excessive P concentrations on vegetable productivity is not clear, but excessive soil P, especially in labile forms, has serious implications for growth of native plants in the future (Handreck 1997) and for P transport and consequent negative off-site impacts on water quality.

Of greater concern is the potential for environmental impacts arising from elevated concentrations of P in the soils used for vegetable production. Phillips (2002) reported that soils containing elevated solution P concentrations can act as non-point sources of pollution of surface and shallow ground water. Reddy et al. (1980) and Redding et al. (2002) reported considerable increases in the proportion of Ca[Cl.sub.2]-P as a proportion of total P in the surface 0-0.05 m as a result of repeated applications of effluent. Those workers concluded that the potential for P transport in runoff and subsoil leaching increased with Ca[Cl.sub.2]-P concentrations in the soil. Similarly, Pote et al. (1999) and Dougherty et al. (2006) observed increased concentrations of P in runoff when soil test P levels were high. Therefore, the high concentrations of P we observed in soils used for vegetable production around Sydney suggest they have increased potential to export P and negatively affect water quality.

In the present investigation, the increased ratios of total P, Colwell-P, and Ca[Cl.sub.2]-P in the vegetable soils over that of the reference soils demonstrates that not only has vegetable production increased the size of each of the non-labile, labile, and solution P pools, it has also caused a shift in the relative size of the different pools. A higher proportion of the P in the vegetable soils is in labile forms, suggesting that soil P buffering capacity has been exceeded. Moreover, the highest ratios of vegetable to reference for Ca[Cl.sub.2]-P and Colwell-P (239 : 1 and 44 : 1, respectively) were observed in the 0.10-0.20 m soil layer (Table 2). These were a direct consequence of the different distribution patterns of the 2 forms of P down the profile between the vegetable and reference soils. Whereas the concentration of both Ca[Cl.sub.2]-P and Colwell-P in the reference soils decreased rapidly with depth, the decreases were much smaller in the vegetable soils (e.g. comparing concentrations at 0-0.10 m with those at 0.10-0.20 m) (Table 2). Tillage practices, such as rotary cultivation and bed formation, are likely to have distributed P into the 0.10-0.20 m layer and our data suggest this zone is becoming saturated with P.

The linear relationship between Colwell-P and total P concentrations in the surface soils (0-0.10m) across all soil types evaluated (Fig. 1) indicates the Colwell extraction is useful for measuring the q component of labile P across a range of soil types. In contrast, the different relationships between Ca[Cl.sub.2]-P and total P for each soil type (Fig. 1) highlights how soil properties such as clay content, pH, and buffering capacity can influence the concentration of P in the soil solution (I component).

The accumulation of P in the vegetable soils, as reported in this survey, is a direct consequence of a long history of heavy applications of poultry manure and inorganic P fertilisers. Both are usually applied in the beginning of the season during bed formation and all the soil samplings were performed after the crops were harvested. Therefore, the high P levels found in the soils cannot be attributed to the timing of fertiliser application of the current season. Moreover, we suggest this would correspond to the period when the concentration of available P in the vegetable soils should be the lowest, due to depletion of the available P pool by the preceding crop.

Poultry litter contains 2.6% N and 1.8% P (Griffiths 2004) and is a popular source of nutrients for vegetable growers within the Sydney basin (Dorahy et al. 2005). Applying poultry manure at agronomic rates for N (e.g. 150 kg N/ha) corresponds to approximately 100 kgP/ha, which is in excess of the P requirements of many vegetable crops. As highlighted by the results of our soil survey, this has led to the accumulation of P in the soil over time. Consequently, alternative management practices are required to decrease P accumulation in soils used for vegetable production within the Greater Sydney Region. Given that available P concentrations were well in excess of crop requirements, we suggest production could be sustained in many instances without the need for further P additions. Wells et al. (2000) investigated vegetable production in a Kandosol at Somersby, NSW, under a number of management systems, including conventional practice (urea + high rates of poultry litter) and an organic system, which used a combination of green waste compost and rock phosphate. They found that Bray extractable P concentrations in the topsoil (0-0.10m) of the conventional system increased from 87 to 361 mg/kg, a 3.14-fold increase over a 3.5-year period. In contrast, the organic system resulted in only a 0.93-fold increase (87 to 168mg/kg) (Wells et al. 2000). Regardless of the management system employed or the form of P used, there is a clear need to use soil testing to identify whether P fertiliser is required and match subsequent P application rates with crop P requirements.

The lower C concentrations and soil structural stability of the vegetable soils in comparison to the unfarmed reference soils detected in this survey was of particular relevance to the future management of these soils. Elsewhere, soils used for vegetable growing can undergo distinct pedological changes and become transformed into distinct soil types, namely Fimic Anthrosols (Zhang et al. 2003) or Anthroposols (Isbell 1996). Compared to the original soils, surface horizons of Fimic Anthrosols used for vegetable production are characterised by high concentrations of organic matter, nutrients, particularly phosphorus (P), and well-developed soil structure, and may have elevated heavy metal concentrations (Zhang et al. 2003).

Zhang et al. (2003) reported that full development of Fimic Antrosols from a range of parent soils could occur within 30 years in many parts of China. The fact that many of the vegetable soils investigated in the present survey had grown vegetables for > 30 years, and in some instances > 100 years, highlights how different cultural and management practices can influence soil characteristics. One explanation could be the lower use of organic inputs such as manure, composts, and night soils in vegetable production in Sydney compared with that in China. For example, Huang et al. (2006) reported that intensive vegetable growing only accumulated soil organic matter where heavy applications of organic fertilisers in the form of cow manure were used. Excessive tillage practices, such as rotary hoeing and bed formation, as well as the heavy use of inorganic fertilisers, would exacerbate the loss of carbon from the soils used in the Greater Sydney Region for vegetable production.

Commensurate with the observed decline in soil organic matter in the soils used for vegetable production was a significant decrease in soil structural stability (Table 5). The complete absence of water-stable macro-aggregates (> 250 [micro]m) indicated significant soil physical degradation, which may contribute to the high rates of surface runoff, sediment transport, and nutrient export reported from vegetable farms within the Sydney Basin (Hollinger et al. 2001). Wells et al. (2000) and Sarooshi et al. (2002) both reported significant increases in soil organic carbon concentrations as a result of applying compost to vegetable growing soils. This highlights the potential of using composts as alternative inputs in vegetable growing to reverse the current soil organic carbon decline and structural degradation.

The observed changes in pH, EC, and exchangeable cations are also attributed to management practices associated with conventional vegetable production systems around Sydney. Lime is regularly applied to maintain soil pH, and poultry manure is also rich in basic cations, namely Ca, K, Mg, and Na. Likewise, salts associated with inorganic and organic inputs are likely to increase soil EC over time.

Soil type is also an important factor when considering changes in some soil chemical characteristics and the impact of management practices upon them. For example, Kandosols are inherently acidic (pH 4.7, Table 4) and would be subjected to higher rates of lime application than the younger Alluvial soils, which have higher pH in their natural state (Walker 1960). In the latter soils, pH decreased (Table 4), possibly as a consequence of downward leaching of nitrates in the absence of lime application.

This study clearly demonstrates an urgent need to develop alternative management practices, which preserve the productivity and profitability of vegetable production systems within the Greater Sydney Region, while minimising potential environmental impacts arising from them. Further research is required to integrate options such as decreasing tillage and application of inorganic P fertilisers and poultry manure, with alternative inputs such as low N and P containing composts and cover crops to maintain soil organic carbon and improve soil structure, as well as avoid P accumulation in the soil.


Our survey has indicated that soils used for vegetable production within the Sydney region have considerably different chemical and physical properties to those of adjacent unfarmed 'reference' soils. The most significant changes in these soils relate to the accumulation of P, loss of soil carbon, and decline in soil structural stability, possibly as a consequence of applying heavy rates of inorganic fertilisers and poultry manure, as well as excessive cultivation. These changes have implications with respect to increasing the potential for off-farm movement of runoff, sediments, and nutrients, and subsequent impacts on the water quality of the Hawkesbury-Nepean River and its tributaries. Consequently, we advocate the need to develop strategies for improving nutrient management and tillage practices in vegetable production systems within the Sydney region.


Financial support of the Department of Environment and Conservation (NSW) and NSW Treasury for this project is acknowledged. We thank all the owners of the vegetable farms in participating in this survey. We also thank Leigh James, Ashley Senn, Cedric Hawkins, and Stephen Ng for their assistance in identifying the sites used in this survey.


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K.Y. Chan (A,B,G), C.G. Dorahy (A,C), S. Tyler (A,B,D), A.T. Wells (E), P.P. Milham (F), and I. Barchia (C)

(A) Centre for Recycled Organics in Agriculture.

(B) NSW Department of Primary Industries, Locked Bag 4, Richmond, NSW 2753, Australia.

(C) NSW Department of Primary Industries, PMB 8, Camden, NSW 2570, Australia.

(D) Current address: 432 Zara Rd., Chillingham, NSW 2484, Australia.

(E) NSW Department of Primary Industries, Locked Bag 26, Gosford, NSW 2250, Australia.

(F) NSW Department of Primary Industries, Forest Rd, Orange, NSW 2800, Australia.

(G) Corresponding author. Email:
Table 1. Major soil types used for vegetable growing in the Sydney
region, New South Wales, Australia, and the number of farms included
in the present investigation

Soil types Australian soil Number
 classification of sites

Alluvial (active) Rudosols 5
Alluvial (old) Rudosols 7
Podzolic (shale) Chromosols 13
Earth (sandstone) Kandosols 9

Table 2. Comparison of total P, Colwell-P, and Ca[Cl.sub.2]-P
concentrations (mg/kg) at 3 sampling depths in unfarmed reference and
adjacent vegetable soils from the Sydney region, NSW, Australia

Values in parentheses are the log-transformed means used for
statistical comparison; l.s.d. at P = 0.05 to compare land use within
each of the soil depths and mean soil depth

Soil land use Sampling depth (m)
 0-0.10 0.10-0.20 0.20-0.30


Reference 0.07 (-1.185) 0.02 (-1.792) 0.01 (-1.896)
Vegetable 6.77 (0.830) 4.78 (0.679) 1.23 (0.090)
l.s.d. (P = 0.05) (0.382) (0.382) (0.382)


Reference 15 (1.176) 6 (0.806) 7 (0.823)
Vegetable 273 (2.436) 264 (2.423) 134 (2.126)
l.s.d. (P = 0.05) (0.289) (0.289) (0.289)

Total P

Reference 320 218 180
Vegetable 1436 1344 835
l.s.d. (P = 0.05) 524 524 524

Soil land use Mean


Reference 0.03 (-1.624)
Vegetable 4.26 (0.533)
l.s.d. (P = 0.05) (0.415)


Reference 9 (0.935)
Vegetable 224 (2.328)
l.s.d. (P = 0.05) (0.262)

Total P

Reference 239
Vegetable 1205
l.s.d. (P = 0.05) 476

Table 3. Comparison of basic chemical properties at 3 sampling depths
in unfarmed reference and adjacent vegetable soils from the Sydney
region, NSW, Australia

Values in parentheses are the log-transformed means used for
statistical comparison; l.s.d. at P = 0.05 to compare land use within
each of the 3 soil depths and mean soil depth

Soil land use Sampling depth (m)
 0-0.10 0.10-0.20 0.20-0.30

Carbon (g/kg)

Reference 28.8 15.1 11.3
Vegetable 16.5 14.7 11.2
l.s.d. (P = 0.05) 4.5 n.s. n.s.

Nitrogen (g/kg)

Reference 1.5 (-0.817) 0.7 (-1.159) 0.5 (-1.346)
Vegetable 1.0 (-1.001) 0.9 (-1.065) 0.6 (-1.230)
l.s.d. (P = 0.05) (0.152) n.s. n.s.


Reference 5.31 5.34 5.31
Vegetable 5.87 5.84 5.70
l.s.d. (P = 0.05) n.s. n.s. n.s.

EC (dS/m)

Reference 0.09 (-1.028) 0.08 (-1.101) 0.08 (-1.086)
Vegetable 0.18 (-0.751) 0.14 (-0.842) 0.13 (-0.885)
l.s.d. (P = 0.05) (0.273) n.s. n.s.

Exch. Ca ([cmol.sub.c]/kg)

Reference 5.94 3.89 3.35
Vegetable 8.88 8.77 7.72
l.s.d. (P = 0.05) 2.83 2.83 2.83

Exch. K ([cmol.sub.c]/kg)

Reference 0.40 (-0.398) 0.24 (-0.617) 0.21 (-0.686)
Vegetable 0.64 (-0.197) 0.39 (-0.404) 0.30 (-0.528)
l.s.d. (P = 0.05) (0.17) n.s. n.s.

Exch. Mg ([cmol.sub.c]/kg)

Reference 1.84 (0.265) 1.26 (0.101) 1.41 (0.150)
Vegetable 1.83 (0.262) 1.79 (0.253) 1.99 (0.298)
l.s.d. (P = 0.05) n.s. (0.12) (0.12)

Exch. Na ([cmol.sub.c]/kg)

Reference 0.21 (-0.686) 0.22 (-0.661) 0.27 (-0.569)
Vegetable 0.33 (-0.482) 0.35 (-0.462) 0.37 (-0.433)
l.s.d. (P = 0.05) n.s. n.s. n.s.

Exch. Al ([cmol.sub.c]/kg)

Reference 0.56 0.57 0.51
Vegetable 0.15 0.15 0.16
l.s.d. (P = 0.05) 0.25 0.25 0.25

Soil land use Mean

Carbon (g/kg)

Reference 18.4
Vegetable 14.1
l.s.d. (P = 0.05) 4.3

Nitrogen (g/kg)

Reference 0.9 (-1.107)
Vegetable 0.8 (-1.099)
l.s.d. (P = 0.05) n.s.


Reference 5.32
Vegetable 5.80
l.s.d. (P = 0.05) 0.46

EC (dS/m)

Reference 0.08 (-1.072)
Vegetable 0.15 (-0.826)
l.s.d. (P = 0.05) (0.242)

Exch. Ca ([cmol.sub.c]/kg)

Reference 4.39
Vegetable 8.46
l.s.d. (P = 0.05) 2.75

Exch. K ([cmol.sub.c]/kg)

Reference 0.28 (-0.376)
Vegetable 0.44 (-0.567)
l.s.d. (P = 0.05) (0.16)

Exch. Mg ([cmol.sub.c]/kg)

Reference 1.50
Vegetable 1.87
l.s.d. (P = 0.05) n.s.

Exch. Na ([cmol.sub.c]/kg)

Reference 0.23 (-0.638)
Vegetable 0.35 (-0.459)
l.s.d. (P = 0.05) (0.17)

Exch. Al ([cmol.sub.c]/kg)

Reference 0.55
Vegetable 0.15
l.s.d. (P = 0.05) 0.21

n.s., Not significant at P = 0.05.

Table 4. Comparison of mean pH and exchangeable Al in unfarmed
reference and vegetable soils in relation to the major soil types used
for vegetable production in Sydney region, NSW, Australia

Soil land use Soil type

 Rudosol Rudosol Chromosol Kandosol
 (active) (old)

Reference 6.56 5.84 5.19 4.73
Vegetable 5.30 6.06 5.73 6.12
l.s.d. (P = 0.05) 1.04 1.04 1.04 1.04

Exch. Al ([cmol.sub.c]/kg)

Reference 1.1 2.0 5.7 10.2
Vegetable 1.5 1.0 2.0 1.0
l.s.d. (P = 0.05) 3.9 3.9 3.9 3.9

Table 5. Comparison of the percentage of water-stable aggregates
(>2 mm and >250 [micro]m) in surface soil (0-0.10m) collected from
unfarmed soil (Reference) and soils used for vegetable production in
the Sydney region, NSW, Australia

Soil land use Water-stable aggregates (%)

 >2 mm >250 [micro]m

Reference 29.4 66.0
Vegetable 0.2 0.9
l.s.d. (P = 0.05) 11.1 5.4
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Article Details
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Author:Chan, K.Y.; Dorahy, C.G.; Tyler, S.; Wells, A.T.; Milham, P.P.; Barchia, I.
Publication:Australian Journal of Soil Research
Geographic Code:8AUST
Date:Mar 1, 2007
Previous Article:Indexing soil quality: a new paradigm in soil science research.
Next Article:Assessment of the influence of soil structure on soil strength/soil wetness relationships on Red Ferrosols in north-west Tasmania.

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