Dealumination as a mechanism for increased acid recoverable aluminium in Waikato mineral soils.
The Waikato Region (Fig. 1) covers much of the Central North Island of New Zealand, extending south of Lake Taupo and Turangi to the Bombay Hills, and with a land area of 25 000 [km.sup.2] The Waikato region can be divided into distinct topographical areas (Table 1).
About 70% of this area is production land. Of this, ~81% is in pastoral agriculture (mainly dairy, beef and sheep farming), 18% is in plantation forestry, and <1% is in various types of horticulture (Table 2).
Soil quality monitoring started in 1995 in the Waikato Region, initially as part of the national 500 Soils project (Hill et al. 2003). The monitoring network now includes 128 soil quality sites. Sampling sites were chosen to cover a representative range of Soil Orders and land uses including background and different farming systems. Testing for trace elements in these samples was initiated in 2003 as the deficiency or excess of trace elements in soils can have a major bearing on soil health even though they are present at low concentrations. Some trace elements are essential micronutrients for plants and animals, and others are not. However, both essential and nonessential elements can become toxic at higher concentrations. Three sub-regional transects, complementary to the soil quality monitoring, were also sampled.
The aim of the soil quality monitoring program is to assess the regions soils for production and environmental protection. It is part of maintaining appropriate productivity, while reducing adverse environmental effects, and contributing to human and animal health. This paper focuses on a potentially interesting soil process, dealumination, as the mechanism for the increase observed in strong acid recoverable Al and associated elements in farmed soils compared with background soils in the soil quality monitoring program.
Soil quality monitoring sites were chosen to cover a representative range of Soil Orders and land uses. Sampling consisted of 25 soil cores (0-100mm) over a 50-m transect, which are combined to form composites for analysis (Sparling et al. 2002).
[FIGURE 1 OMITTED]
Sub-regional transect sampling was based on a 2-km grid, independent of landuse or Soil Order. Samples were taken at 2 depths, 0 100 and 100-200mm, from the sides of rectangular test-pits. Data from 3 large sub-regional transects were combined with data from the soil quality monitoring to provide a sufficiently large dataset (251 samples) for valid statistical analysis. The breakdown of sites by land use was 36 background forest, 37 cropping and horticultural, and 178 pastoral. Only data for mineral soils is presented in this paper as organic soils have low concentrations of Al.
Background sites listed in Tables 3 and 4 were identified on the basis of their current land use and what is known of their land use history. There are far fewer of these than the farmed land sites. They were all long-term forest, uninfluenced by anthropogenic activities for the life of the trees. Some of these sites may have been logged or cleared by early generations, but atmospheric inputs in New Zealand soils are relatively low, and for the most part these sites are regarded as being close enough to background to serve as a useful point of comparison.
Samples are analysed for an established set of soil quality chemical and physical parameters following Spading et al. (2002) and for 32 elements following EPA 200.2 (total recoverable metals hydrochloric/nitric acid digestion). Measurements were made at IANZ-accredited laboratories (soil quality chemistry at Landcare Research, Palmerston North, soil quality physical parameters at Landcare Research, Hamilton, and elemental analysis by ICP-MS at Hill Laboratories, Hamilton). Samples were also assayed for total fluorine at Hill Laboratories using the alkali-fusion/ion-selective electrode method developed by Massey University ('Methods of Soil Analysis' 2nd edn, Pt 2, 26-4.3.3). A subset of samples was analysed twice, and/or at different laboratories, as a check on precision and accuracy of the results. In addition, in-house quality control standards and a standard reference fiver sediment (AGAL-10) were also analysed. Obtained values were in range with the following exceptions:
* Mercury (4x) and Sn (1 x) were reanalysed due to quality control problems.
* The AGAL-10 values obtained for Cr & Ni are lower than the certified range values due to the certified results being obtained using a total digestion, which releases more interstitially bound elements than EPA 200.2. Recovery of non-interstitially-bound elements would not be affected, and are available to either digestion method.
Relative enrichment (or depletion) of each analyte in farmed soils was determined by calculating the ratio of the mean results from farmed soils (0-100 mm) with the mean from background sites (also 0-100mm). Where necessary, data were log-transformed to form a normal distribution. Pooled Student's t-tests were used to assess significance of the difference between each pair of means. For the subset of samples from sub-regional soil transects, where soil was collected at 2 depths, relative surface enrichment was assessed by comparing the 0-100 with the 100-200 mm results. In this case, significance was assessed using paired Student's t-tests after data were transformed to a normal distribution.
Results and discussion
Measurement showed significant enrichment and depletion associated with soil depth and land use (Tables 3 and 4). Some of these relate to well-known sources of trace elements in farming, whereas other results were previously unsuspected. These are discussed in turn below. A ranking of relative enrichments in farmed soils is shown in Fig. 2. The discussion is on a gravimetric basis. As soil bulk density has significantly increased (by 17%) in farmed soils compared with background ones (Table 4), consistent with surface compaction, the magnitude of the changes in soil chemistry and biochemistry are likely to be greater on a volumetric or per ha basis.
Enrichments associated with known sources
For several of the elements showing higher concentrations in farmed soils with accompanying evidence of surface enrichment, sources are well known, and are only discussed here synoptically. Total recoverable P is enriched 6-fold in farmed soils, almost all of which can be attributed to use of superphosphate fertiliser. Cd, F, and U, the 3 contaminant elements commonly linked to this source, show the same behaviour as P, in both having significantly higher concentrations in farmed soils and being surface-enriched (P< 0.0001 in all cases) (Table 3). The similar behaviour of Ca can be attributed to the combined influence of superphosphate fertiliser (which is 40% gypsum, or CaS[O.sub.4]) and lime. In the case of Zn, the significant (P<0.000I) increase in fanned soils and evidence of surface enrichment (P = 0.002) is attributed to the widespread use of zinc sulfate as a protective agent against sporodesmin, the fungal ryegrass toxin that causes liver damage which results in facial eczema in grazing stock. Use of zinc sulfate to prevent facial eczema results in annual zinc loadings to soil of 5-7 kg/ha.year and an annual average accumulation rate in the order of 1 mg/kg.year. This source alone is sufficient to account for the observed doubling of average zinc concentrations from 30mg/kg in background soils to 60mg/kg at the time of this work. In reality the doubling of the soil Zn concentration on the fanned soils is likely to be much higher given that a significant proportion of the farmed soils samples was not from pastoral landuse activities.
[FIGURE 2 OMITTED]
Behaviour of the major nutrients N and K is similar to P and its associated contaminants (Cd, F, and U). Mg and Na are slightly different. Like NPK, these elements are likely to be added to fanned soils either intentionally or inadvertently as part of other treatments, and both show evidence of surface enrichment suggesting the presence of an external source. However, average concentrations of Na farmed soils are no higher than those at background sites, and for Mg evidence for enrichment in farmed soils is comparatively marginal (enrichment factor 1.16, P = 0.017). For these 2 elements, these results suggest that external sources do exist (as surface enrichment is observed), but that exports from farm systems are roughly balancing imports.
In the context of these known sources, the results for Al, and a cluster of elements associated with aluminosilicates, are unusual.
The relative behaviours of Fe and Al illustrate this, due to their similarities. First, at background sites, average concentrations of acid extractable Fe (25 600mg/kg) and AI (24 700 mg/kg) are the highest of any of the metals assayed, and close to each other. Second, both Fe and Al show relative surface depletion, i.e. enrichment at 100-200mm compared with 0-100mm (Table 3). This is unremarkable, and probably reflects the higher organic (and lower mineral) content of the 0-100 mm layer.
Given their high natural concentrations (>2%), we would not expect any common external source to be able to cause a measurable increase in Fe or Al in farmed soils. In keeping with this, Fe shows no such evidence of a concentration increase in farmed soils (Table 3).
By contrast, farmed soils show a significant (P<0.0001) concentration increase in acid-recoverable Al. This increase is remarkable not so much for the size of the enrichment factor (1.5), but for the element involved and the amount of additional Al represented--an additional 13 700 mg/kg of acid-extractable Al in the farmed soils. Assuming an average of 70 years of farming, this would imply an annual average increase in acid-extractable Al in farmed Waikato soils of 175 mg/kg.year.
Acid-recoverable concentrations of eight other elements follow the same general pattern as Al. Ag, Bi, La, Li, Mn, Mo, Sn, and T1 are also enriched in farmed soils compared with background soils but do not show significant surface enrichment (Table 3). Lack of surface enrichment suggests that acid-recoverable concentrations of these elements have increased in farmed soils as a result of soil processes rather than through external additions to the soil surface, as with Al.
Clear differences between recoverable Al in farmed and background soils within the same Soil Orders can be seen when the data is further broken down, although caution has to be exercised due to the low numbers of background sites for each soil order (Table 5). Only the 5 Soil Orders with the largest land area (and number of sites) are presented. Increases are seen in Allophanic and Brown soils (1.5 times), Granular soils (1.2 times), and Pumice soils (1.1 times), while no increase is seen in Gley Soils. The Pumice Soils have lower acid recoverable Al concentrations than the other Soil Orders, probably due to the low clay and high glass dominated sand content of these soils. Differences in the Soil Orders are attributed to differing mineralogies and explained in detail below.
It is central to interpretation that (with the exception of F which is total F) the trace element results describe the total acid-extractable fractions of each element. For some elements, this measurement is likely to under-estimate total concentrations. Standard acid digestions do not readily dissolve the larger crystalline aluminosilicates, but would be expected to readily dissolve metastable and stable secondary minerals, the ultra-fine hydrated oxide films, and other amorphous forms of these elements. Several of the soils in this study are derived from volcanic ash and do not have significant amounts of crystalline aluminosilicates, but instead consist of allophane, which are clay-sized, short-range ordered aluminosilicates associated with the weathering of volcanic ashes and glasses. Nevertheless, standard acid digestions of these short-range ordered aluminosilicate soils leave an undissolved residue, which can be dissolved in HF (Taylor 1997). The apparent increase in Al in farmed soils with no evidence of surface enrichment is more likely to reflect a shift towards more acid-recoverable forms of Al, than an increase in total concentration.
Only a fraction of the total Al content of soils is mobile and plays a role in soil fertility (Kabata Pendias and Pendias 2001). However, the large inert aluminosilicate mineral pool is subject to gradual weathering and break down, with soluble aluminium silicate complexes being important intermediates in the weathering reactions of primary AI-Si minerals and the formation of metastable and stable secondary minerals in soils (Browne and Driscoll 1992). Secondary minerals form by 1 of 2 processes: (1) transformation of primary minerals, e.g. mica [right arrow] vermiculite [right arrow] smectite or chlorite; (2) neoformation, which is the precipitation of the breakdown products of primary and secondary minerals out of soil solution, e.g. smectite, iron oxides (Churchman 2006).
Significant clay mineral alterations may occur within relatively short periods of time. Delmelle et al. (2007) have shown rapid acid dissolution of volcanic ash occurs within eruption plumes and this process is favoured by the presence of fluoride. Egli et al. (2001) showed Al removal from soil mineral interlayers seemed to be enhanced by dissolved F and organic complexing agents in soil near an aluminium smelter. After the reduction of high F deposition rates from the smelter, the accumulated soil organic matter was decomposed over 24 years. The combined decrease in F and humus led to chain reactions with losses of major elements and a dealumination of clay minerals, i.e. removal of interlayered Al of 2:1 minerals and consequent formation of smectites (Egli et al. 2004). In the topsoil, vermiculite and chlorite were transformed to smectite. In the subsoil, mica and chlorite were apparently weathered to smectites.
The F loading rate for a typical Waikato farm is about a tenth of that beside an aluminium smelter. A typical annual fertiliser application is 400 kg/ha.year of single superphosphate containing 1.5% F, equivalent to 6 kg/ha.year, or to ~3% of the average background in Table 3. In comparison, deposition rates for F near aluminium smelters of ~80 kg F/ha.year have been reported by Wilcke et al. (2000) and Polomski et al. (1982), although levels as high as 1000 kg F/ha.year can be calculated from information in Vike (2005).
Accelerated dissolution of Al may be explained by dependence of the rate on the concentration of adsorbed protons and adsorbed fluorides. Both fluoride and adsorbed protons enhance the dissolution rates via a series of coupled pathways (Nordin et al. 1999), e.g. slow conversion of fluoroaluminium complexes to insoluble species, such as gibbsite [Al[(OH).sub.3]], and possible precipitation of fluorite (Ca[F.sub.2]) or fluoroapatite [[Ca.sub.5](PO4)3F], has been suggested in soils amended with Al or Ca (Rai et al. 2007). These results suggested the formation of mineral phases from organo-metal complexes and transformation of amorphous phases into more crystalline ones, such as geothite (Egli et al. 2001).
This is similar to the process seen in podzolisation where weathering of micas into smectites can occur (Righi et al. 1999). Interestingly, Spodosols are reported to retain F more strongly than Aqueptic soils due to their much larger Al-oxide contents (Simard and Lafrance 1996).
These secondary minerals are more soluble in acid; for example, montmorillonite and smectite were substantially degraded in HCl (Palkova et al. 2003). Acid attack of the smectite structure occurred at both interlayer surfaces and edges, while development of non-swelling interlayers in heated montmorillonites substantially reduced but did not stop dissolution of these samples. It is therefore plausible that increased biochemical or chemical activity in farmed soils should result in increased weathering or attack of primary crystalline aluminosilicates, increasing the relative content of more easily extractable secondary aluminosilicates.
Two specific mechanisms that could favour Al mobilisation from clay surfaces include partial dissolution by local areas of high acidity associated with fertiliser granules, and surface complexation and extraction by the fluoride and residual hydrofluoric acid present in phosphate fertiliser. When superphosphate fertilisers are added to soils, monocalcium phosphate dissolves, resulting in a release of phosphoric acid. The pH in the immediate vicinity of a superphosphate fertiliser granule is <2 (McLaren and Cameron 1990). As [H.sub.2]P[O.sub.4.sup.-] ions adsorb onto soil colloids, part of the acid produced is neutralised by the release of O[H.sup.-] ions. Acidification processes result in increasing solubility of aluminium, releasing [Al.sup.3+] ions into solution (Berdren et al. 1997; Habs 1997); for example, increasing nitric acid concentration from pH 5 to 3 in forest soils increased aluminium concentration in leachate by 21% (James and Riha 1989). A low pH induces weathering of soil minerals with a high release of [Al.sup.3+] (Beyer et al. 1993) and mobilisation increases drastically when soil solution pH decreases to <4.5 (Falkengrengrerup and Bergkvist 1995). Dissolved [Al.sup.3+] ions will exist in a distribution of forms depending on pH; these include [Al[([H.sub.2]O).sub.6]].sup.3+.sub.aq], [Al[([H.sub.2]O).sub.5]OH].sup.2+.sub.aq], [Al[([H.sub.2]O).sub.4][(OH).sub.2]].sup.+.sub.aq], [Al[([H.sub.2]O).sub.3][(OH).sub.3]].sup.0.sub.aq], [Al[([H.sub.2]O).sub.2][[(OH).sub.4]].sup.-.sub.aq], and more complex species such as [[Al.sub.13][[(OH).sub.32].sup.7+.sub.aq]. Progressive conversion of attached water ligands ([H.sub.2]O) to hydroxide (O[H.sup.-]) (hydrolysis) occurs as the pH increases. This results in release of protons. As dissolved [Al.sup.3+] ions diffuse through porewater to more neutral pH areas, they will hydrolyse, releasing secondary acid back to solution. Diffusion and hydrolysis of [Al.sup.3+] ions that have been mobilised by acid will therefore have the net effect of recycling (but spreading) the original acidity. The neutralising process of adsorption of [H.sub.2]P[O.sub.4.sup.-] ions results in the opposite effect and the likely precipitation of secondary soil minerals. However, the release and reprecipitation of aluminium from soil is also dependent on the supply of complexation agents and the capacity of the soil to provide aluminium hydroxides (Gundersen and Rasmussen 1990).
Some nitrogen fertilisers can also increase soil acidity but the extent depends on the amounts of N[H.sub.4.sup.+], urea, or biologically fixed N that are lost from the soil, e.g. in drainage waters or farm products (Bolan et al. 1991). Such acidity remains in the soil system unless active application of liming materials is made to the soil, so there is likely less opportunity for the formation of secondary soil minerals. Nitrogen fertilisers also generally lack a strong electro-negative ion such as fluoride, as a complexation agent. Nitrogen fertilisers are likely to have a lesser influence on the dealumination process than phosphate fertilisers.
Continuing the discussion on phosphate fertilisers, New Zealand superphosphate contains ~1.5% F, while reactive phosphate rock (RPR) contains ~3% F, but concentrations vary due to source (O'Hara et al. 1982; Ballance 2006). Addition of F to soil increases pH and concentrations of total organic C, Al, Fe, Ca, Mg, K, Mn, and P in solution (Elrashidi and Lindsay 1987; Arnesen 1998; Arnesen and Krogstad 1998; Meeussen et al. 1999).
Additional F is known to sorb onto Al and Fe oxides and hydroxides in acid soils (Elrashidi and Lindsay 1986; Wilcke et al. 2000) and in Ultic Soils (Harrington et al. 2003). Sorption onto soils is a precursor to Al and Fe dissolution. The [F.sup.-] ion replaces -OH/[H.sub.2]O groups bound to surface Al atoms by ligand exchange, loosens other Al-OH bonds, and facilitates the dissolution of Al from the surface (Nordin et al. 1999). Even at low pH, high concentrations of F are able to form complexes with [Al.sup.3+] that would otherwise exist as the free ion (Sigfusson et al. 2006). A significant proportion of soil F can also exist as fluoro-aluminium complexes at near-neutral pH and these species may play a crucial role in controlling diffusive mobility of F in soil (Rai et al. 2007). Due to their relatively large surface area, dissolution of amorphous Al oxides is more prominent than that of more crystalline phases (Farah et al. 1987). Above all, any allophane present is going to be strongly attacked by fluoride ions. This mechanism is the basis for the quick test for allophane developed by Fieldes and Perrott (1966). The test involves adding NaF to the soil and measuring the resulting pH rise. The reaction of the [F.sup.-] ion with the highly accessible and reactive Al in allophane (allophane is normally 2 : 1 Al : Si and the Al is on the surface of the structure) releases OH ions, which cause a rise in pH. Other, more crystalline, clay minerals such as halloysite do react with F ions but to much lesser extent. While we are proposing that F ions are reacting with Al in clay minerals, the mechanism will happen more readily and to a greater extent in allophane than in halloysite, which has the Al groups more enclosed within the structure (less on the surface). Furthermore, Al is also less abundant in halloysite (Al : Si = 1 : 1) than allophane (Churchman 2006).
At high dissolved F concentrations (>7 mM), likely mechanisms of F retention include adsorption of A1F solution complexes, entrapment in the interparticle pore fluid, and precipitation into solution and/or onto the soil surface (Harrington et al. 2003). Manoharan et al. (1996) have calculated that in soil porewater, fluoride species should primarily exist as aluminium complexes. HF is also the only common acid capable of dissolving silicates. More recently, Manoharan et al. (2007) have shown that F added in phosphate fertiliser significantly increases soil solution concentrations of Al irrespective of solution pH. It is therefore conceivable that the high (15 000 mg/kg) F and residual HF content of superphosphate fertiliser has been working to mobilise A1 from clay mineral surfaces (and presumably also Si, which was not measured in this work). A portion of this would then be free for neoformation as metastable and stable secondary minerals. If the significant enrichment of total recoverable Al in farmed soils does reflect a shift to more acid-soluble forms (through whatever mechanism), we might expect to see an associated increase in recoverable concentrations of those trace elements that are usually locked up inside aluminosilicates. Of these, Li and U could be good markers. Li is primarily associated with aluminosilicates (Kabata Pendias and Pendias 2001), and a significant proportion of natural U is also associated with residual phases. As will be discussed later, such parallel enrichment of recoverable Li and U is observed.
An increase in acid-extractable Al does not necessarily imply an increase in porewater soluble Al, because recovery in an acid digest is not the same thing as water solubility. It is possible that an increase in metastable and stable secondary mineral forms of Al may be associated with more Al in soil porewater. The solubility of aluminium in equilibrium with solid phase aluminium hydroxide is highly dependent on pH and on complexing agents such as fluoride, silicate, phosphate, and organic matter, but not on the clay mineralogy (Bache and Ross 1991; Browne and Driscoll 1992; Dahlgren and Walker 1993; Sjostrom 1994; Masaya et al. 2006). At pH >5.6, naturally occurring aluminium compounds exist predominantly in an undissolved hydroxide form such as gibbsite [Al[(OH).sub.3]] or as aluminosilicates. Under acidic conditions, the most soluble form of aluminium is organically bound aluminium, while the amorphous aluminium hydroxy forms are more soluble than the crystalline forms (Sjrstrrm 1994); for example, aluminium hydroxides, at pH >5.6, and soil organic matter, at pH <5.0, were found to control aluminium equilibria in New Zealand soils (Adams et al. 2000).
Complexation with organic matter (Gurung et al. 1996) or F (Manoharan et al. 1996) of porewater Al is likely to ameliorate potential Al toxicity; for example, most labile Al in the vicinity of an aluminium smelter was bound to F and the activity of [Al.sup.3+] was extremely low (Gago et al. 2002). Large additions of phosphate fertiliser allowed crops to withstand higher concentrations of solution aluminium (Bache and Ross 1991), probably due to enhanced plant growth masking aluminium toxicity. However, high concentrations of F in soil appear to inhibit biological processes resulting in the accumulation of soil organic matter (Egli et al. 2001). Fluoroaluminate complexes may mobilise Al (Totsche et al. 2000; Wilcke et al. 2000) and be effective analogues of inorganic phosphate. Possibly, fluoroaluminate complexes and other phosphate analogues inhibit plant phospholipase D by competing with a substrate phosphate group for substratebinding sites, thereby preventing the formation of an enzymephosphatidyl intermediate (Li 2003). The accumulation of P-mimicking fluoroaluminates in the soil has been shown to restrict barley root growth (Manoharan et al. 2007) and may affect the phosphate absorption of other plant species (Facanha and Okorokova-Facanha 2002). Conversely, an increase in AI solubility could result in [Al.sup.3+] ions competing with other cations for exchange sites; for example, Al in solution played an important role in the mobilisation of cadmium (Berggren 1992).
It appears that 2 competing effects should be considered in this case: on one hand the presence of high F should result in less toxic Al fluoride species being formed in porewater, but on the other hand an increase in amorphous forms of Al may result in increased availability of both Al and the range of trace elements usually associated with crystalline aluminosilicates.
It is therefore possible that these results represent an increase in more easily dissolved forms of Al in farmed soils with time. This is consistent with a study utilising analysis of total elements using HF digestions to study clay mineral changes in a soil near an aluminium smelter over 24 years, which found no significant Al enrichment (Egli et al. 2001).
Several elements behave similarly to and are correlated with Al (Table 6). Like Al, acid-recoverable concentrations of these elements are statistically higher in farmed than background soils, which enrichment factors following the order: Li (2.5)>La (2.1)>Al, Mn (1.5)>Ag, Bi, Mo, Sn, Tl (1.4). Also like Al, these elements show no evidence of surface enrichment (Table 3). Whatever caused the apparent increase in acid-recoverable Al may also be responsible for the observed increases in these other elements.
This idea is supported by extremely high correlations (P<0.0001) between Al and each of these elements. For Bi, Li, La, and Mo, the correlation with Al is also the highest correlation between each element and any other element (Al-La scatter plot shown in Fig. 3). A slightly higher correlation than that with Al is observed between Tl and Mn (r=0.836).
An interesting question is posed with Cr, which often forms part of the residual phase in soils and could be expected to follow the same trends as AI and the other associated elements described above. Instead, Cr was ~20% (P<0.002) lower in farmed soils than background ones, although there was no significant surface enrichment and there is a significant correlation with Al (r=0.476). If this is a real effect and not a statistical false positive, then possible explanations are that farm exports exceed imports, and/or enhanced Cr leaching is occurring on farmed soils. Of the common forms of Cr, the chromate (Cr(VI)) oxyanion is considered to be the most mobile. A possible cause of increased leaching could therefore be through more oxidation of chromic (Cr(III)) to chromate (Cr(VI)) in farmed soils. However, like AI(III), Cr(III) is a small highly charged cation, and another explanation would be though enhanced solubilisation of chromic Cr through fluoride complexation.
As discussed earlier, the idea of an anthropogenic source of new Al is unlikely due to the high natural concentrations of Al (difficulty of an external source making any difference), the lack of evidence for surface enrichment, and the reality that acid digests do not recover the full Al load. This leads to the possibility that some process in farmed soils has caused an increase in acid-soluble forms of Al from aluminosilicate parent materials. If this interpretation is accurate, the correlations in Table 4 would then suggest that a reasonable fraction of total Bi, La, Li, Mo, Sn, and T1 is locked inside aluminosilicate minerals in natural soils. If some gradual degradation of a proportion of the aluminosilicates does occur in farmed soils, this would be expected to result in an increasing portion of total Bi, La, Li, Mo, Sn, and Tl becoming acid-available. A similar grouping was found by Tyler (2004), who placed P, Al, Li, La, and U all in same group following PCA factor analysis of HCl extractions of a Swedish Podzol.
[FIGURE 3 OMITTED]
Uranium concentrations from our study provide further evidence for dealumination, although U chemistry is more complicated due to an external source of U. This external source of U in New Zealand soils is phosphate fertiliser (McBride and Spiers 2001; Taylor and Kim 2007). Concentrations of total recoverable U in Waikato farmed soils appear to be 2.5 times their background mean and also show signs of more marginal but significant surface enrichment (P< 0.0001, Table 3). While correlations exist between U and P (Fig. 4), U correlates even more strongly with total recoverable Al (Fig. 5), suggesting that U is being (or has been) mobilised from the portion of aluminosilicates that have become degraded. Several sources suggest that 60 mg/kg would be a reasonable upper estimate of the average historic U content of New Zealand superphosphate fertilisers (Menzel 1968; McBride and Spiers 2001; Taylor and Kim 2007).
[FIGURE 4 OMITTED]
[FIGURE 5 OMITTED]
Loading estimates using an assumed figure of 60mg/kg for average historic U in superphosphate fertiliser would suggest that phosphate fertilisers might be responsible for about 75% of the total recoverable U increase observed in Waikato soils (Table 7). Note that in top-down estimates, concentrations have been corrected to 0-200 mm equivalents. The mean ratio 0-100/0-200mm for U is 1.17. Under these circumstances, superphosphate might account for 0.012 mg/kg. year against an observed increase of 0.016 mg/kg.year. In order for superphosphate to account for the entire increase, an historic average U concentration of at least 80 mg/kg would be required ('at least' because the simple bottom-up loading estimate assumes no losses).
There is therefore scope for the existence of a second source to account for the rest of the observed increase in of total recoverable U in Waikato soils. The correlation between total recoverable U and Al in Waikato soils (Fig. 5) suggests that this may be tied in with the observed increase in total recoverable Al (Table 3). In sequential extractions, a large proportion of the natural U in soils is known to be associated with resistant residual phases, which are most typically silicate and aluminosilicate minerals (Taylor 1997). The combination of surface enrichment with the correlation with Al suggests that like Li and La, recoverable U may also be increasing as a consequence of an increase in amorphous forms of Al drawn from the soil aluminosilicate and possibly silicate pool.
The relative slopes of Figs 4 and 5 suggest that on a weight-for-weight basis, 1 mg P may have contributed ~14 times more U than 1 mg recoverable Al from a 2 : 1 alumiosilicate. An assumed U concentration in superphosphate of 60mg/kg translates to ~700 mg U/kg P (Table 7). Dividing by 14 gives 50 mg U/kg A1. This would imply a U content of an idealised 2 : 1 layer clay mineral (sheet aluminosilicate) in the region of 7 mg/kg, which is similar to the U content of volcanic ash from recent volcanic eruptions in the region (Briggs et al. 1993).
The overall suggested picture regarding the 3-fold increase in acid-recoverable U in Waikato regional soils would then be that:
* about three-quarters of the increase in recoverable U may be attributable to use ofsuperphosphate fertilisers (comparatively small mass but high U content);
* the other quarter may have come from aluminosilicate mineral weathering resulting in an increase in secondary minerals (low U content, but large available mass).
Another interesting question would be the possible enhanced production of [SiF.sub.4] gas from soil and its contribution to the fluorine cycle and the atmospheric halide load. Significant Si-depletion of surface Si[O.sub.2] has been observed in Hawaiian basalt flows, explained by the thermodynamic relationship between HF and SiF4. At temperatures <375[degrees]C SiF4 dominates the fluoride gases found in volcanic plumes, while HF dominates >375[degrees]C (White and Hochella 1992).
An increase in acid-recoverable Al concentrations in Waikato farmed soils not readily explained by an external source due to lack of enrichment at the soil surface. However, it could be explained as an increase in the concentration of acid-recoverable A1 as a result of accelerated weathering or chemical attack of primary crystalline and short-range order aluminosilicates. Acid-recoverable concentrations of La, Li, and Pb that are normally retained inside aluminosilicates (in residual phases) were also significantly higher in farmed than background soils but are not selectively enriched at the soil surface.
This process may also contribute one-quarter of the observed increase in acid-recoverable U. If it is occurring, accelerated Al weathering may be a normal part of an increase in soil productivity, or may be facilitated by an external agent capable of attacking crystalline aluminosilicates. A candidate in the latter category is the F (and/or possibly free HF) in phosphate fertilisers, because this is now known to substantially increase Al in soil porewater. Two specific mechanisms that could favour Al mobilisation from clay surfaces include partial dissolution by local areas of high acidity associated with fertiliser granules, and surface complexation and extraction by the fluoride and residual hydrofluoric acid present in phosphate fertilisers.
Based on the high reactivity between F and both Al and Si, potential exists for significant production of [SiF.sub.4(g)] as another side-effect of phosphate fertiliser use. It is unknown what annual flux this source may contribute to the atmospheric F load.
Our thanks to 2 anonymous referees for helpful comments; to Jock Churchman, The University of Adelaide, for advice on mineralogy; to Landcare Research and Environmental Waikato field staff for collecting samples.
Manuscript received 23 March 2009, accepted 31 July 2009
Adams M, Hawke D, Nilsson N, Powell K (2000) The relationship between soil solution pH and [Al.sup.3+] concentrations in a range of South Island (New Zealand) soils. Australian Journal of Soil Research 38, 141-153. doi: 10.1071/SR98095
Arnesen A (1998) Effect of fluoride pollution on pH and solubility of Al, Fe, Ca, Mg, K and organic matter in soil from Ardal (Western Norway). Water, Air, and Soil Pollution 103, 375-388. doi: 10.1023/ A: 1004921600022
Arnesen A, Krogstad T (1998) Sorption and desorption of fluoride in soil polluted from the aluminium smelter at Ardal in Western Norway. Water, Air, and Soil Pollution 103, 357-373. doi: 10.1023/ A: 1004900415952
Bache B, Ross J (1991) Effect of phosphorus and aluminum in the response of spring barley to soil acidity. Journal of Agricultaral Science 117, 299 305. doi: 10.1017/S0021859600067022
Ballance (2006) Ballance Fact Sheets: Fluorine. Ballance Agric.-nutrients Ltd, Tauranga, New Zealand. Available at: www.ballance.co.nz/ fafluorine.html (accessed 4 September 2006).
Berden M, Nilsson S, Nyman P (1997) Ion leaching before and after clear-cutting in a Norway spruce stand effects of long-term application of ammonium nitrate and superphosphate. Water, Air, and Soil Pollution 93, 1-26. doi: 10.1007/BF02404745
Berggren D (1992) Speciation and mobilization of aluminum and cadmium in Podzols and Cambisols of Sweden. Water, Air, and Soil Pollution 62, 125 156. doi: 10.1007/BF00478457
Beyer L, Blume H-P, Henss B, Peters M (1993) Soluble aluminium- and iron-organic complexes and carbon cycle in Hapludaffs and Haplorthods under forest and cultivation. The Science of the Total Environment 138, 57-76. doi: 10.1016/0048-9697(93)90405-U
Bolan N, Hedley M, White R (1991) Processes of soil acidification during nitrogen cycling with emphasis on legume based pastures. Plant and Soil 134, 53 63.
Briggs RM, Giffur MG, Moyle AR, Taylor SR, Norman MD, Houghton BF, Wilson CJN (1993) Geochemical zoning and eruptive mixing in ignimbrites from the Mangakino volcano, Taupo Volcanic Zone. Journal of Volcanology and Geothermal Research 56, 175 203. doi: 10.1016/0377-0273(93)90016-K
Browne B, Driscoll C (1992) Soluble aluminum silicates: stoichiometry, stability, and implications for environmental geochemistry. Science 256, 1667-1670. doi: 10.1126/science.256.5064.1667
Churchman GJ (2006) Soil phases: the inorganic solid phase. In 'Soils: basic concepts and future challenges'. (Eds G Certini, R Scalenghe) pp. 23-44. (Cambridge University Press: Cambridge, UK)
Dahlgren R, Walker W (1993) Aluminum release rates from selected Spodosol Bs horizons--effect of pH and solid-phase aluminum pools. Geochimica et Cosmochimica Acta 57, 57-66. doi: 10.1016/0016-7037 (93)90468-C
Delmelle P, Lambert M, Dufrene Y, Gerin P, Oskarsson N (2007) Gas/ aerosol-ash interaction in volcanic plumes: new insights from surface analyses of fine ash particles. Earth and Planetary Science Letters 259, 159-170. doi: 10.1016/j.epsl.2007.04.052
Egli M, Durrenberger S, Fitze P (2004) Spatio-temporal behaviour and mass balance of fluorine in forest soils near an aluminium smelting plant: short- and long-term aspects. Environmental Pollution 129, 195-207. doi: 10.1016/j.envpol.2003.10.005
Egli M, Mirabella A, Fitze P (2001) Clay mineral transformations in soils affected by fluorine and depletion of organic matter within a time span of 24 years. Geoderma 103, 307-334. doi: 10.1016/S0016-7061(01) 00046-5
Elrashidi M, Lindsay W (1986) Chemical equilibria of fluorine in soils: a theoretical development. Soil Science 141, 274-280. doi: 10.1097/ 00010694-198604000-00004
Elrashidi M, Lindsay W (1987) Effect of fluoride on pH, organic matter and solubility of elements in soils. Environmental Pollution 47, 123-133. doi: 10.1016/0269-7491(87)90042-X
Facanha A, Okorokova-Facanha A (2002) Inhibition of phosphate uptake in corn roots by aluminum-fluoride complexes. Plant Physiology 129, 1763-1772. doi: 10.1104/pp.001651
Falkengrengrerup U, Bergkvist B (1995) Effects of acidifying air-pollutants on soil/soil solution chemistry of forest ecosystems. Annali di Chimica 85, 317-327.
Fieldes M, Perrott K (1966) The nature of allophane in soils: Part 3-Rapid field and laboratory test for allophane. New Zealand Journal of Science 9, 623-629.
Gago C, Marcos M, Alvarez E (2002) Aqueous aluminium species in forest soils affected by fluoride emissions from an aluminium smelter in NW Spain. Fluoride 35, 110-121.
Gundersen P, Rasmussen L (1990) Nitrification in forest soils--effects from nitrogen deposition on soil acidification and aluminum release. Reviews of Environmental Contamination and Toxicology 113, 1-45.
Gurung S, Stewart R, Loganathan P, Greg P (1996) Aluminium--organic matter fluoride interactions during soil development in oxidised mine waste. Soil Technology 9, 273-279. doi: 10.1016/0933-3630(95) 00042-9
Habs H (1997) 'Aluminum. Environmental health criteria 194.' (World Health Organization: Geneva)
Harrington L, Cooper E, Vasudevan D (2003) Fluoride sorption and associated aluminum release in variable charge soils. Journal of Colloid and lnterface Science 267, 302-313. doi: 10.1016/S00219797(03)00609-X
Hewitt AE (2003) 'New Zealand Soil Classification.' Landcare Research Science Series No. 1. (Manaaki Whenua Press: Lincoln, New Zealand)
Hill RB, Sparling GP, Framption C, Cuff J (2003) National soil quality review and programme design. Ministry for the Environment Technical Report 74, Ministry for the Environment, Wellington, New Zealand.
James B, Riha S (1989) Aluminum leaching by mineral acids in forest soils. I. Nitric sulfuric-acid differences. Soil Science Society of America Journal 53, 259-264.
Kabata-Pendias A, Pendias H (2001) 'Trace elements in soils and plants.' (CRC Press: Boca Raton, FL)
Li L (2003) Aluminum fluoride inhibition of cabbage phospholipase D by a phosphate-mimicking mechanism. FEBS Letters 461, 1-5. doi: 10.1016/ S0014-5793(99)01414-3
Manoharan V, Loganathan P, Parfitt R, Tillman R (1996) Changes in soil solution composition and aluminium speciation under legume-based pastures in response to long-term phosphate fertiliser applications. Australian Journal of Soil Research 34, 985-498. doi: 10.1071/ SR9960985
Manoharan V, Loganathan P, Tillman R, Parfitt R (2007) Interactive effects of soil acidity and fluoride on soil solution aluminium chemistry and barley (Hordeum vulgare L.) root growth. Environmental Pollution 145, 778-786. doi: 10.1016/j.envpol.2006.05.015
Masaya S, Junko H, Katsuhiro T (2006) The behavior of uranium in forming hydroxyl aluminum silicate ion. Clay Science 12, 270-273.
McBride M, Spiers G (2001) Trace element content of selected fertilizers and dairy manures as determined by ICP-MS. Soil Science & Plant Analysis 32, 13-156. doi: 10.1081/CSS-I00102999
McLaren RG, Cameron KC (1990) 'Soil science.' (Oxford University Press: Auckland)
Meeussen J, Scheidegger A, Hiemstra T, Van Riemsdijk W, Borkovec M (1999) Predicting multicomponent adsorption and transport of fluoride at variable pH in a goethite-silica sand system. Abstracts of Papers of the American Chemical Society 217, U750.
Menzel RG (1968) Uranium, radium, and thorium content in phosphate rocks and their possible radiation hazard. Journal of Agricultural and Food Chemistry 16, 231-234. doi: 10.1021/jf60156a002
Nordin J, Sullivan D, Phillips B, Casey W (1999) Mechanisms for fluoride-promoted dissolution of bayerite [beta-Al(O[H.sub.3])(s)] and boehmite [gamma-AlOOH]: F-19-NMR spectroscopy and aqueous surface chemistry. Geochimica et Cosmochimica Acta 63, 3513-3524. doi: 10.1016/S0016-7037(99)00185-4
O'Hara P, Fraser A, James M (1982) Superphosphate poisoning of sheep: the role of fluoride. New Zealand Veterinary Journal 30, 199-201.
Palkova H, Madejova J, Righi D (2003) Acid dissolution of reduced-charge Li- and Ni-montmorillonites. Clays and Clay Minerals 51, 133-142. doi: 10.1346/CCMN.2003.0510202
Polomski J, Fluhler H, Blaser P (1982) Accumulation of airborne fluoride in soils. Journal of Environmental Quality 11, 457-462.
Rai K, Agarwal M, Dass S, Shrivastav R (2007) Diffusive mobility of fluoride in soil: Some mechanistic aspects. Communications in Soil Science and Plant Analysis 38, 57-68. doi: 10.1080/001036206 01093629
Righi D, Huber K, Keller C (1999) Clay formation and podzol development from postglacial moraines in Switzerland. Clay Minerals 34, 319-332. doi: 10.1180/000985599546253
Sigfusson B, Gislason S, Paton G (2006) The effect of soil solution chemistry on the weathering rate of a Histic Andosol. Journal of Geochemical Exploration 88, 321-324. doi: 10.1016/j.gexplo.2005.08.067
Simard R, Lafrance P (1996) Fluoride sorption and desorption indices in Quebec soils. Communications in Soil Science and Plant Analysis 27, 853-866. doi: 10.1080/00103629609369602
Sparling GP, Rijkse WC, Wilde H, van der Weerden T, Beare M, Francis G (2002) Implementing soil quality indicators for land: Research Report for 2000-01 and Finial Report for MfE Project Number 5089. Landcare Research Contract Report, LC0102/015. Landcare Research, Hamilton.
Taylor MD (1997) The fate of uranium contaminants of phosphate fertilisers in New Zealand. MSc Thesis, Waikato University, Hamilton, New Zealand.
Taylor MD, Kim ND (2007) The fate of uranium contaminants of phosphate fertilizer. In 'Loads and fate of fertilizer-derived uranium'. (Eds LJ De Kok, E Schnug) p. 229. (Backhuys Publishers: Leiden, The Netherlands)
Totsche K, Wilcke W, Korber M, Kobza J, Zech W (2000) Evaluation of fluoride-induced metal mobilization in soil columns. Journal of Environmental Quality 29, 454-459.
Tyler G (2004) Vertical distribution of major, minor, and rare elements in a Haplic Podzol. Geoderma 119, 277-290. doi: 10.1016/j.geoderma. 2003.08.005
Vike E (2005) Uptake, deposition and wash off of fluoride and aluminium in plant foliage in the vicinity of an aluminium smelter in Norway. Water, Air, and Soil Pollution 160, 145-159. doi: 10.1007/s11270-005-3862-1
White A, Hochella M (1992) Surface-chemistry associated with the cooling and subaerial weathering of recent basalt flows. Geochimica et Cosmochimica Acta 56, 3711-3721. doi: 10.1016/0016-7037(92) 90164-E
Wilcke W, Totsche K, Korber M, Kobza J, Zech W (2000) Fluoromobilization of metals in a Slovak forest soil affected by the emissions of an aluminum smelter. Journal of Plant Nutrition and Soil Science-Zeitschift fur Pflanzenernahrung und Bodenkunde 163, 503-508. doi: 10.1002/1522-2624(200010)163:5<503::AID-JPLN503> 3.0.CO;2-R
M. D. Taylor (A,B) and N. D. Kim (A)
(A) Environment Waikato, PO Box 4010, Hamilton East, Hamilton 3247, New Zealand.
(B) Corresponding author. Email: email@example.com
Table 1. Topographical areas of the Waikato Rehion, New Zealand Topographical area Features Eastern Ranges Volcanic hill country. Good quality soils in low lying areas Lowlands and Plains Flat and gently rolling land with highly versatile soils. Large areas of wetlands Western and Central Steep, sedimentary hill country, large tracts of Hill Country native forest, extensive cave and karst systems Taupo Volcanic Zone Pumice lands, high landscape values, geothermal features Topographical area Prominent land uses Dominant Soil Orders (A) Eastern Ranges Dairy farming and horticulture Brown (low lands), beef and sheep farming Lowlands and Plains Urban and rural settlement, Allophanic, dairy farming, horticulture Granular, and cropping Organic Western and Central Beef and sheep farming, exotic Brown, Hill Country forestry and native forest Granular Taupo Volcanic Zone Beef and sheep farming, dairy Pumice farming, exotic forestry and native forest (A) New Zealand Soil Classification (Hewitt 2003). Table 2. Land use in the Waikato Region Land use Land area (ha) Maize 3 364 Potatoes 2 309 Onions 2 102 Kiwifruit 800 Asparagus 560 Field/seed peas 364 Apples 300 Avocadoes 200 Grapes 200 Broccoli 136 Forestry 329 780 Pasture, sheep 333 440 Pasture, dairy 623 010 Pasture, beef 471 350 Table 3. Summarised results for analyses common to both soil quality monitoring and transects All units mg/kg dry weight apart from pH and C : N ratio. n.s., Not significant 0-100/ Paired t-test Background (0-100 mm) 100-200 mm P-value Average Range pH 1.03 <0.02 5.1 4.2-6.0 Total C 1.51 <0.0001 78000 24 000-20 000 Fe 0.92 <0.0001 25600 4700-76 000 Al 0.91 <0.0001 24700 4500-70 000 Total N 1.66 <0.0001 4000 1000-11 000 Ca 1.89 <0.0001 1400 390-3300 Mn 1.34 n.s 780 50-2960 Mg 1.47 <0.0001 760 140-2010 K 1.65 <0.0001 490 170-1300 P 1.96 <0.0001 350 160-730 Total F 1.23 <0.0001 190 70-300 Na 1.23 <0.001 160 90-280 Ba 1.02 0.02 97 15-310 V 0.97 0.02 68 5-300 Zn 1.49 <0.0001 28 11-58 Sr 1.44 <0.0001 19 5-57 Cr 1.13 n.s l8 1-150 Cu 1.28 0.001 16 4-55 Pb 1.03 n.s 11 3-32 La 1.02 n.s 11 2-55 Rb 1.22 <0.0001 7.60 1.1-22 Co 0.94 0.001 5.90 0.90-28 As 0.96 0.004 5.10 1.0-25 Li 1.01 n.s 3.90 0.60-9.4 Ni 1.25 0.005 3.90 0.56-21 B 1.70 <0.0001 2.90 1.0-8.5 Cs 1.08 0.04 1.60 0.30-5.3 Sn 0.99 n.s 1.14 0.38-2.6 U 1.17 <0.0001 0.79 0.19-2.5 Mo 1.07 n.s 0.76 0.23-1.80 Tl 1.13 n.s 0.22 0.057-0.60 Hg 0.93 <0.0001 0.19 0.019-0.50 Bi 0.97 0.008 0.18 0.059-0.40 Cd 1.74 <0.0001 0.11 0.030-0.30 Ag 1.06 n.s 0.11 0.030-0.32 Sb 1.40 n.s 0.076 0.020-0.17 C : N 0.97 n.s 20 11-32 Olsen P 2.05 <0.0001 6 1-17 Farmed (0-100 mm) Farmed/ Pooled background t-test Average Range pH 5.9 4.5-7.5 1.15 <0.0001 Total C 74000 22 000-210 000 0.95 n.s Fe 27000 2000-325 000 1.05 n.s Al 37000 6600-89 000 1.50 <0.0001 Total N 6700 1900-17 300 1.68 <0.0001 Ca 4530 720-14 700 3.24 <0.0001 Mn 1180 75-8080 1.51 0.008 Mg 880 180-3900 1.16 0.017 K 750 160-2200 1.53 <0.0001 P 2130 330-10 000 6.09 <0.0001 Total F 440 120-900 2.32 <0.0001 Na 160 50-340 1.00 n.s Ba 150 25-09 1.55 <0.0001 V 56 5-290 0.82 n.s Zn 62 1-258 2.21 0.002 Sr 20 5-56 1.05 0.036 Cr 14 1-220 0.78 0.002 Cu 24 3-250 1.50 0.079 Pb 16 3-95 1.45 <0.001 La 23 2-179 2.09 <0.001 Rb 10 0.9-31 1.32 0.024 Co 6.80 0.84-25 1.15 0.028 As 8.60 0.70-94 1.69 0.001 Li 9.90 0.80-32 2.54 <0.0001 Ni 6.10 1.0-34 1.56 <0.0001 B 3.70 1.0-10 1.28 0.010 Cs 2.80 0.47-11 1.75 <0.0001 Sn 1.59 0.42-5.4 1.39 0.001 U 2.00 0.50-5.5 2.53 <0.0001 Mo 1.03 0.20-4.8 1.36 0.042 Tl 0.31 0.060-1.4 1.41 0.034 Hg 0.16 0.027-0.50 0.84 n.s Bi 0.25 0.050-2.5 1.39 0.014 Cd 0.71 0.10-2.0 6.45 <0.0001 Ag 0.15 0.048-0.65 1.36 0.010 Sb 0.12 0.020-1.30 1.58 0.002 C : N 11 8-17 0.55 <0.0001 Olsen P 48 3-354 8.00 <0.0001 Table 4. Summarised results for analyses from soil quality monitoring only n.s., Not significant Measurement Background Average Range N[O.sub.3]-N (mg/kg) 10 0.01-56 N[H.sub.4]-N (mg/kg) 15 0.36-46.6 Anaerobically min. N (mg/kg) 109 26-310 Canon exch. capacity (cmol/kg) 27 9-55 Base saturation 36 18-61 Exchang. canons (cmol(+)/kg): Ca 5.8 1.6-14 Mg 2.2 0.18-6.0 K 0.57 0.28-1.3 Na 0.24 0.05-0.64 Bulk density (t/[m.sup.3]) 0.70 0.39-1.16 Particle density (t/[m.sup.3]) 2.34 2.08-2.61 Total porosity (%v/v) 7l 54-82 Macroporosity (-5 kPa) (%v/v) 23 3.3-44 Air capacity (-10 kPa) (%v/v) 27 3.9-51 Agg. stability (mean wt diam., mm) 2.20 1.95-2.87 Total water (%v/v) 31 13-58 Available water (%v/v) 17 4.3-52 Topsoil depth (mm) 120 40-200 Rooting depth (mm) 850 270-1200+ Measurement Farmed Average Range N[O.sub.3]-N (mg/kg) 49 3.9-218 N[H.sub.4]-N (mg/kg) 11 0.01-82 Anaerobically min. N (mg/kg) 157 12-528 Canon exch. capacity (cmol/kg) 29 14-55 Base saturation 56 16-82 Exchang. canons (cmol(+)/kg): Ca 13 4.6-26.3 Mg 1.70 0.48-6.5 K 1.04 0.25-2.6 Na 0.17 0.01-1.1 Bulk density (t/[m.sup.3]) 0.82 0.46-1.35 Particle density (t/[m.sup.3]) 2.36 2.09-2.71 Total porosity (%v/v) 66 49-78 Macroporosity (-5 kPa) (%v/v) 11 0.5-32 Air capacity (-10 kPa) (%v/v) 13 0.7-38 Agg. stability (mean wt diam., mm) 2.12 0.71-2.78 Total water (%v/v) 30 8-1960 Available water (%v/v) 13 3.7-57 Topsoil depth (mm) 170 50-360 Rooting depth (mm) 720 200-1200+ Measurement Farmed/ Pooled background t-test N[O.sub.3]-N (mg/kg) 4.90 <0.0001 N[H.sub.4]-N (mg/kg) 0.73 0.04 Anaerobically min. N (mg/kg) 1.44 0.009 Canon exch. capacity (cmol/kg) 1.07 n.s Base saturation 1.56 0.0003 Exchang. canons (cmol(+)/kg): Ca 2.24 <0.0001 Mg 0.77 n.s K 1.82 0.0004 Na 0.71 0.046 Bulk density (t/[m.sup.3]) 1.17 <O.0001 Particle density (t/[m.sup.3]) 1.01 n.s Total porosity (%v/v) 0.93 <0.0001 Macroporosity (-5 kPa) (%v/v) 0.48 <0.0001 Air capacity (-10 kPa) (%v/v) 0.48 <0.0001 Agg. stability (mean wt diam., mm) 0.96 n.s Total water (%v/v) 0.97 n.s Available water (%v/v) 0.76 n.s Topsoil depth (mm) 1.42 <0.0001 Rooting depth (mm) 0.85 0.02 Table 5. Acid-recoverable Al in farmed soils compared to background for the 5 Soil Orders with the largest land area in the Waikato Soil order Clay mineralogy Land use n Av. Al Allophaic Allophane, imogolite, and Background 3 28 000 ferrihydrite Farmed 57 44 000 Brown Micalillite and vermiculite Background 5 22 000 Farmed 24 35 000 Granular Kaolin-group, halloysite Background 3 35 000 Farmed 31 41 000 Pumice Allophane and imogolite; clay Background 2 10 000 content often <10%; Farmed 10 11 000 prominent glass in sand fraction Gley Various, reflects the ungleyed Background 2 35 000 material from which these Farmed 44 33 000 soils are derived Table 6. Correlations (log-transformed) of elements significantly enriched in farmed soils compared with background sources but not significantly surface-enriched Ag Bi La Li Mn Mo Sn Tl Al 0.585 0.673 0.719 0.732 0.566 0.726 0.541 0.676 Ag 1 0.584 0.676 0.401 0.413 0.636 0.616 0.457 Bi 1 0.513 0.579 0.397 0.658 O.bll 0.502 La 1 0.500 0.608 0.523 0.373 0.696 Li 1 0.579 0.524 0.506 0.559 Mn 1 0.540 0.318 0.836 Mo 1 0.555 0.562 Sn 1 0.358 Table 7. Estimates of the accumulation rate of U in Waikato soils using 2 approaches Variable Number Unit Bottom-up estimate U concentration in superphosphate 60 mg/kg Average amount of super added annually 400 kg/ha (dairy & beef weighted average) Amount of U loaded annually 24 000 mg/ha Mass of soil in 1 ha to 200mm depth at 2 000 000 kg/ha BD 1 t/[m.sup.3] Expected annual concentration increase 0.012 mg/kg.year if all U retained in top 200 mm Top-down estimate Estimated natural U concentration 0.79 mg/kg Present day geomean U concentration 1.84 mg/kg (0-200 mm) Approximate amount accumulated over 1.05 mg/kg 65 years Approximate observed annual accumulation 0.016 mg/kg.year rate
|Printer friendly Cite/link Email Feedback|
|Author:||Taylor, M.D.; Kim, N.D.|
|Publication:||Australian Journal of Soil Research|
|Date:||Dec 1, 2009|
|Previous Article:||Scaling analysis of soil water retention parameters and physical properties of a Chinese agricultural soil.|
|Next Article:||Crop performance as affected by three opening configurations for no-till seeder in annual double cropping regions of northern China.|