Environmental risk indicators for soil phosphorus status.
There are several instances in Australia of eutrophication of receiving waters being caused by diffuse-source phosphorus (P) mainly attributed to agricultural land use. Notable examples are the Peel Harvey Inlet (Department of Environment and Heritage 2006), the Barwon-Darling River (Cottingham et al. 1995), and the Great Barrier Reef (Baker et al. 2003), and there is a compelling need to be able to identify soils whose high P status poses a risk to water quality.
In the USA, 'trigger' extractable soil P concentrations have been set, above which impacts on water quality are considered likely (summarised in Gartley and Sims 1994). However, the rationale for deriving these triggers varies, and ranges from expert opinion on 'acceptable levels' [generally 345 times higher than soil P test levels considered necessary for optimum growth (Gartley and Sims 1994)] to defining the 'change-point' in the relationship between extractable soil P and solution P where solution P concentration begins to increase more rapidly as soil test P increases (e.g. Hesketh and Brookes 2000; Kleinman et al. 2000). While this relationship is soil-dependent because of differences between soils in P buffer capacity (Hesketh and Brookes 2000; Penn et al. 2005), attempts have been made to derive a more generalised relationship to define the change-point by using a 'P saturation' index based on a soil P test and an indicator of the maximum P sorption capacity of the soil. In particular, a P saturation index derived from the Mehlich 3 extractant (Jones 1990) and defined as the molar ratio [Mehlich 3-P/ (Mehlich 3-Al + Mehlich 3-Fe)] has been related to solution P (using 0.01 M CaCl2-extractable P as the surrogate measure of soil solution P). The change-point of this relationship has been used as the criterion for identifying soils whose P status is considered to be excessively high and likely to result in significant environmental risk (Sims et al. 2002). However, several instances have been recorded where there is a single linear relationship between the Mehlich P saturation index and solution P concentration in both runoff and drainage water (e.g. Pautler and Sims 2000), with no indication of a change-point. In addition, Koopmans et al. (2002) have demonstrated that the change-point is an artefact of the experimental conditions. They found that the change-point is dependent on the soil : solution ratio, range of solution P concentrations, and range of P buffer capacities of the soils used, and may be absent under particular experimental circumstances. As a result, the change-point is not an acceptable trigger for discriminating high risk from low risk.
Another trigger that has been used to identify high-risk soils is the Mehlich P saturation index, which corresponds to an oxalate P saturation index [molar oxalate-extractable P/molar (oxalate-extractable Al + Fe)] value of 25% (Sims et al. 2002). This value is used in the Netherlands as the trigger value above which soils may not be fertilised with P, and is based on research indicating that this value corresponds to a leachate solution P concentration of 0.1 mg P/L (Van der Zee et al. 1990).
It is apparent from this review that there is no universally accepted environmental risk index for soil P. Furthermore, none of the indices used in the USA or Europe are routinely determined in Australia. This raises two questions: what method(s) should be used to assess risk in Australia, and how should the threshold or trigger value be set? To answer these questions, bioavailable P must first be put in context from an environmental risk viewpoint, and then the commonly used soil P test methods evaluated in this context.
Biologically available P is divided operationally into two sources: dissolved reactive P (DRP) and bioavailable particulate P (BPP). DRP comprises dissolved and colloidal organic and inorganic P that passes through a 0.45-um filter and is reactive to the acid molybdate procedure of Murphy and Riley (1962). This P is considered to be immediately bioavailable to microorganisms (Bostr6m et al. 1988). Phosphorus extracted from soils by dilute (0.005M or 0.01 M) calcium chloride (CaCI2-P) is highly correlated with DRP in runoff (Dougherty et al. 2008; Schindler et al. 2009) and drainage water (Hesketh and Brookes 2000). This extractant can therefore be used as the benchmark method to estimate relative DRP concentrations emanating from soils of different P status. The other source of biologically available P is BPP, which is readily desorbed from sorption sites when DRP in the surrounding solution is reduced. Algal bioassays indicate that P desorbed to iron-oxide impregnated filter paper (FeO-P) is highly correlated with algal-available P (Chaston 2002), and FeO-P is considered to be the benchmark method for assessing BPP (Sharpley 1993).
The appropriate benchmark method to use to assess environmental risk will depend on whether DRP, BPP, or both are considered to be the major source(s) of bioavailable P. This assessment will depend on land use, the main form in which P is being transported off-site (i.e. in solution or sorbed on suspended sediment), the nutrient enrichment ratio, and the effectiveness of the delivery pathway to the receiving water. For example, a review of data on the P form in runoff from several studies of pastures in Australia indicated that DRP comprised from <14% to >90% of total P in runoff depending on surface cover (Nash and Halliwell 1999). Where erosion is a risk to water quality, BPP comprises the primary P source (Sharpley 1993). These results indicate that environmental risk assessment requires two indices for measuring the P status of the source soil: one for assessing DRP and the other, BPP. In this complex situation, a defensible risk index might simply be to determine whether or not a soil's P status is higher than is necessary to meet the productivity goals of a particular land use (as is used in the USA--see above). Because crops vary in their P requirements and, by extension, the critical soil P test value for optimum yield (e.g. Moody and Bolland 1999), it is apparent that setting a single trigger value is not appropriate unless it is set at a value above that commensurate with optimum productivity (e.g. 90% maximum yield) of all possible land uses.
It is desirable from a practical viewpoint that any assessment of environmental risk should be based on the same soil tests that are used to make P fertiliser recommendations for crops/ pastures. One set of soil tests then meets the goals of both productivity and environmental risk assessment. Several diagnostic soil P tests are used in Australia to assess the likelihood of crop and pasture yield responses to applied P. The most widely used soil P tests are Colwell-P, Olsen-P, and acid- (or BSES-) P, and critical soil test levels or ranges for optimum yield have been defined for many crop and pasture species (Moody and Bolland 1999). Recently, the single-point P buffer index (PBI) (Burkitt et al. 2002) has been adopted by many Australian laboratories to allow critical Colwell-P values for particular crops/pastures to be adjusted for soil P buffer capacity (Moody 2007). However, neither the environmentally relevant Ca[Cl.sub.2]-P nor FeO-P is in routine use as a diagnostic soil P test. This raises the question of whether the soil tests which are currently used to guide productivity might also be useful to assess potential environmental risk.
This paper addresses this question. It also investigates the validity of the change-point approach to setting trigger values across a diverse suite of soils, and compares this approach with alternative criteria.
Materials and methods
Thirty-one surface (0-0.10 m) samples were collected from the fertilised rows of horticultural tree plantations in south-eastern Queensland and northern New South Wales (26 soils) and from grain-growing soils in south-western and northern Queensland (5 soils). Soil types comprised the following Australian Soil Orders (Isbell 2002): Ferrosols (10), Kurosols (10), Kandosols (3), Tenosols (2), Dermosols (2), and Vertosols (4). Samples were air-dried at 40[degrees]C and ground <2 mm before analysis.
Soil solution extraction
Soil samples (150 g) were weighed into screw-top plastic bottles with removable bases, and de-ionised water was added to bring the soils to field capacity (-10 kPa matric suction). The bottles were incubated at 25[degrees]C for 4 days and the soil solution then removed by centrifugation (Aitken and Outhwaite 1987). The extracted soil solution was filtered (<0.45 um) and analysed for DRP using the Murphy and Riley (1962) colourimetric method.
Soil P methods
Bicarbonate-extractable P was determined following the procedures of Colwell (1963) (Colwell-P) and Olsen et al.
[FIGURE 1 OMITTED] (1954) (Olsen-P) as described in Methods 9B2 and 9C2 in Rayment and Higginson (1992). Phosphorus was also extracted by the Mehlich 3 method (0.2 M C[H.sub.3]COOH, 0.25 M N[H.sub.4]N[O.sub.3], 0.015M N[H.sub.4]F, 0.0013M HN[O.sub.3], 0.001M EDTA; Jones 1990) and by 0.005 M Ca[Cl.sub.2] (Ca[Cl.sub.2]-P) for 17h at a soil:extractant ratio of 1:5. Phosphorus desorbable to an iron-oxide impregnated filter paper (FeO-P) was determined as described by Guo et al. (1996) by shaking 1.0g (<2mm) soil with a filter paper strip (20 mm by 100 mm) in 40 mL of 0.002 M Ca[Cl.sub.2] for 16h. Sorbed P was then extracted by 0.2 M [H.sub.2]S[O.sub.4]. Except for Mehlich 3-P, which was determined by inductively coupled plasma-atomic emission spectroscopy (ICP-AES), P in all extracts was determined by the colourimetric procedure of Murphy and Riley (1962).
The single-addition PBI was determined as described in Burkitt et al. (2002) and calculated as:
PBI (Colwell) = (Ps + Colwell-P)/[c.sup.0.41] (1)
where Ps is the amount of P sorbed from an initial addition of 1000 mg P/kg, and c is the final solution concentration (mg P/L).
A surrogate index for estimating soil solution P concentration was calculated as (Colwell-P/PBI). The Mehlich P saturation index was calculated as [Mehlich P/Mehlich (Fe + Al)] using molar concentrations.
Additional soil analyses
Mehlich 3 extractable Fe and Al were determined by ICP-AES. Total organic carbon (TOC) was determined by the Heanes (1984) method, and particle size analysis (clay and silt) was determined following soil dispersion with sodium hexametaphosphate and reading by hydrometer. Effective cation exchange capacity was calculated as the sum of exchangeable cations and exchangeable acidity (Method 15J1 in Rayment and Higginson 1992).
SigmaStat[R] 3.1 was used to carry out linear and curvilinear regressions.
The soils exhibited a wide variation in all properties, and the median Colwell-P, Olsen-P, and PBI values were higher, and the range in PBI greater, than those of the 90 surface soils used by Burkitt et al. (2002) to develop the PBI (Table 1).
Relationship between soil solution P and Ca[Cl.sub.2]-P
There was a strong curvilinear relationship ([R.sup.2] = 0.89, P < 0.001) between soil solution P and Ca[Cl.sub.2]-P concentrations, supporting the use of Ca[Cl.sub.2]-P as a convenient measure of soil solution P (Fig. 1).
Correlations between P measures
Linear correlation coefficients between the various soil P tests are presented in Table 2. There was no correlation between Ca[Cl.sub.2]-P and FeO-P, which are the benchmark measures of DRP and BPP, respectively. Of the routine soil P tests, Olsen-P was most highly correlated with FeO-P, while the index (Colwell-P/ PBI) was highly correlated with Ca[Cl.sub.2]-P and soil solution P, as was the Mehlich P saturation index.
Several studies have reported change-points in the relationship between Ca[Cl.sub.2]-P and Mehlich P saturation index at a saturation index of ~0.20 (e.g. Bloesch and Rayment 2006). Although only three soils in the current study had Mehlich P saturations >0.15, there was no indication of a curvilinear trend between Ca[Cl.sub.2]-P and Mehlich P saturation (Fig. 2).
Threshold values of environmental risk indicators Threshold concentrations of various indicators have been used to assess the risk of off-site P movement in drainage or runoff, and indicators can be broadly grouped as assessing DRP [e.g. ANZECC and ARMCANZ (2000) guidelines for surface water quality noting that these guideline concentrations are at catchment scale, not field scale] or BPP (e.g. acceptable soil P test levels). While there is a tendency to think of environmental risk indicators as being different to crop productivity indicators in terms of methods used and interpretation, the ability of soil to supply adequate (productivity viewpoint) or excess (environmental viewpoint) bioavailable P is governed by the principles of the quantity-intensity concept of nutrient availability (Schofield 1955). The quantity factor is the amount of P available for biological uptake over a specified period, whereas the intensity factor is the instantaneous soil solution P concentration. The quantity and intensity factors are linked by the buffer capacity factor (quantity of P released for a unit change in solution P concentration). As demonstrated by the results in Table 2, the quantity and intensity factors need not be correlated. In terms of assessing the environmental risk of a soil's P status, indicators of DRP must be correlated with soil P intensity, whereas indicators of BPP must be correlated with soil P quantity.
[FIGURE 2 OMITTED]
As neither of the benchmark methods used to measure DRP and BPP, i.e. Ca[Cl.sub.2]-P and FeO-P, respectively, is in routine use in Australia, surrogate measures of these analyses are required. Probert and Moody (1998) demonstrated that a (quantity/buffer capacity) index (Colwell-P/single-point P sorption index) was highly correlated with plant uptake because it was correlated with P intensity. This finding is confirmed for the soils used in this current study; the index (Colwell-P/PBI) was highly correlated with both soil solution P and Ca[Cl.sub.2]-P (Table 2). Olsen-P is the best indicator of FeO-P (Table 2), although Colwell-P is also highly correlated with FeO-P. Both Olsen-P and Colwell-P are indicators of P quantity by virtue of the ability of the bicarbonate anion in their common extractant to displace sorbed P and to solubilise calcium phosphates by depressing the solution Ca concentration. In summary, it is suggested that the index (Colwell-P/PBI) can be used as an indicator of DRP and Olsen-P (or Colwell-P) as an indicator of BPP.
Productivity and environmental risk assessment
A range of indicators and threshold values have been used to assess soil P status for environmental risk, and equivalent threshold values for the suite of soils in this study are presented for Olsen-P and (Colwell-P/PBI) (Table 3). For comparison, critical values of Olsen-P and (Colwell-P/PBI) for optimum yield of various crops and pastures are also indicated in Table 3. Critical (Colwell-P/PBI) values were calculated for the specified crops/pastures from equations in Moody (2007) using the median PBI value (162) of the soils in the current study (Table 1).
It is apparent that threshold values of (Colwell-P/PBI) as a surrogate for DRP concentrations in the Australian water quality guidelines at catchment scale are an order of magnitude lower than the critical values needed in soils for productivity. The threshold value of (Colwell-P/PBI) calculated as a surrogate for the DRP value used in The Netherlands for drainage is higher than the critical values for wheat and pasture but not for potato. The DRP environmental risk threshold values are therefore often in conflict with soil solution DRP values required for productivity. However, the inevitable dilution/removal of soil solution DRP as it moves in areal scale from plot ([m.sup.2]) to block and sub-catchment ([km.sup.2]) (as shown by Barlow et al. 2005) simply means that realistic threshold DRP values need to be set at the plot scale taking land use and soil type into account.
The Olsen-P values corresponding to BPP thresholds are an order of magnitude higher than the critical values required for optimum productivity. However, the threshold FeO-P values in Table 3 correspond to maximum algal growth observed in both studies (Chaston 2002 and Sharpley 1993), and because the algal growth response to increasing FeO-P was linear, setting a lower acceptable level of algal growth (based on allowable chlorophyll-a concentrations for instance) would lower the FeO-P threshold. For example, FeO-P at 50% maximum chlorophyll-a concentration (700 [micro]g chlorophyll-a/ L) was 70mgP/kg (Chaston 2002), which corresponds to 67 mgOlsen-P/kg. Nonetheless, the fact that there is a link between algal growth and Olsen-P will allow threshold Olsen-P values to be set for whatever level of algal growth is considered to be environmentally acceptable.
The widely used commercial soil P tests of Olsen-P, Colwell-P, and PBI have been shown to be applicable for assessing the P status of soil for risk of off-site P movement in both dissolved and bioavailable particulate form. The challenge to environmental regulators is to set threshold values at the plot scale that realistically reflect the likely downstream environmental impact of runoff and/or suspended sediment at catchment scale. The levels of these soil P tests that are necessary for optimum productivity of various cropping systems are known, allowing a direct linkage to be made between productivity and environmental risk. Such a linkage has not often been transparent in the past because environmental risk indicators had no credentials for assessing productivity.
My special thanks are due to Siok Ay Yo, Grant Pu, Bruce Compton, and staff of the Chemistry Centre, Environment and Resource Sciences, Queensland Department of Environment and Resource Management, for undertaking the soil analyses reported here.
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Manuscript received 12 July 2010, accepted 23 September 2010
P. W. Moody
Environment and Resource Sciences, Queensland Department of Environment and Resource Management, Ecosciences Precinct, Dutton Park, Qld 4102, Australia. Email: Phil.Moody@derm.qld.gov.au
Table 1. Range and median values for some properties of the 31 surface (0-0.10m) soils ECEC, Effective cation exchange capacity; PBI, single-point P buffer index Median Property Range value pH (1 : 5 water) 4.6-8.7 5.6 EC (1 : 5 water) (dS/m) 0.03-1.0 0.11 Total organic carbon (%) 0.5-6.8 3.8 ECEC ([cmol.sub.c]/kg) 3.5-26.0 9.0 Clay (%) 5-55 32 Colwell-P (mg/kg) 20-837 133 Olsen-P (mg/kg) 8-161 41 PBI 26-1891 162 Table 2. Linear correlation coefficients (r) between soil phosphorus parameters for the 31 surface (0-0.10m) soils PBI, Single-point P buffer index; ** P <0.01; *** P < 0.001 Soil Colwell soln PBI P/PBI P Mehlich P 0.883 *** 0.813 *** 0.579 ** sat. index FeO-P 0.317 0.304 0.426 Colwell-P -0.052 -0.122 0.773 *** Olsen-P 0.228 0.190 0.506 Ca[C1.sub.2]-P 0.925 *** 0.944 *** -0.594 *** PBI -0.602 *** -0.641 *** Soil soln P 0.902 *** Ca[C1.sub.2]-P Olsen-P Colwell-P FeO-P Mehlich P 0.906 *** 0.185 0.085 0.278 sat. index FeO-P 0.357 0.955 *** 0.828 *** Colwell-P -0.032 0.893 *** Olsen-P 0.266 Ca[C1.sub.2]-P PBI Soil soln P Table 3. Threshold values of soil phosphorus parameters for assessing environmental risk and critical productivity threshold values PBI, Single-point P buffer index Bioavailable P Threshold indicator Measure Environmental Measure Value risk indicator BPP Algal growth FeO-P 120 mg/kg (A) 190 mg/kg (A) DRP Water quality DRP 4-20 [micro]g/L (C) 100 [micro]g/L Surrogate threshold indicator Critical Measure Value productivity threshold Olsen-P 114 mg/kg Perennial pasture: 181 mg/kg 9 mg/kg (E) Annual pasture: 13 mg/kg (E) Colwell 0.01-0.07 Potato: 0.51 (G) P/PBI 0.35 Wheat: 0.21 (G) Temperate pasture: 0.22 (G) (A) Chaston (2002); (B) Sharpley (1993); (C) ANZECC and ARMCANZ (2000); (D) Van der Zee et al. (1990); (E) Gourley and James (1997); (E) Gourley (1987); (F) Moody (2007) calculated at PBI=162.
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