Manganese oxidation and reduction in soils: effects of temperature, water potential, pH and their interactions.
Manganese (Mn) is an essential plant micronutrient, and Mn toxicity and deficiency in plants have both been well documented (Hannam and Ohki 1988). Manganese toxicity can limit commercial plant production and necessitate soil liming and drainage and the use of Mn-tolerant plant species and varieties (Zhang et al. 2006; Setter et al. 2009; Hayes et al. 2012; Hemandez-Soriano et al. 2012), yet toxicity thresholds for soil Mn tests are poorly defined because plant-available Mn concentrations can fluctuate widely and rapidly with changes in soil temperature and water content (Uren 1999). A greater understanding of Mn behaviour in soils most likely would provide for more proficient management and containment of Mn toxicity in the field.
The behaviour of Mn in soils is most simply represented by the concept of a balance between divalent Mn ([Mn.sup.2+]) and insoluble Mn oxides (Mn[O.sub.x]). However, the transformations between these two nominal forms are not well understood because of a seemingly misplaced faith in thermodynamics to predict the fluctuations in the concentration of water-soluble plus exchangeable [Mn.sup.2+] (WS+E [Mn.sup.2+]), a key measure of the likelihood of Mn toxicity (Uren 2013). There are two problems here: (i) molecular oxygen ([O.sub.2]) is a poor oxidant of [Mn.sup.2+] even though [O.sub.2] is essential for [Mn.sup.2+] oxidation; and (ii) the failure of many researchers to accept or appreciate that the microbial oxidation of [Mn.sup.2+] is the primary determinant of the concentration of WS+E [Mn.sup.2+] in soils which are neither waterlogged nor air-dry. The incompetence of [O.sub.2] as an oxidant in the abiotic oxidation of [Mn.sup.2+] arises out of the fact that [O.sub.2] has two half-filled, highest occupied molecular orbitals, which makes it difficult for [O.sub.2] to accept a single electron from [Mn.sup.2+] in the first step of the reaction (Luther 2005, 2010). This first step is therefore rate-limiting, but despite this limitation, Mn-oxidising microbes can overcome this barrier and catalyse the oxidation of [Mn.sup.2+] at rates of the order of [10.sup.3] times faster than abiotic oxidation, even at pH 8 (Morgan 2005). Some microbes do so by producing enzymes which directly oxidise [Mn.sup.2+] (e.g. Thompson et al. 2006), while other Mnoxidising microbes produce extracellular superoxide (Learman et al. 2011; Hansel et al. 2012), and providing hydrogen peroxide which forms as a byproduct is destroyed, the shortcomings of [O.sub.2] as an oxidant in the abiotic reaction are overcome. The widespread production of superoxide by bacteria (Diaz et al. 2013) may explain why so many bacteria in a diverse array of situations and environments can oxidise [Mn.sup.2+].
Further, if the oxidation of [Mn.sup.2+] in soils is abiotic then it would be expected to increase with (i) increasing activity of [Mn.sup.2+], (ii) increasing pH, (iii) increasing temperature, and (iv) increasing activity of molecular [O.sub.2], but as pointed out by Uren (2013), there is long-standing evidence that, although [Mn.sup.2+] oxidation may increase initially as each of these four variables increase, the extent of oxidation decreases after a maximum is reached. For example, Thompson et al. (2005) found that 21[degrees]C was the optimum temperature for the oxidation of [Mn.sup.2+] in soil and that [Mn.sup.2+] oxidation was totally inhibited at 3[degrees]C and 52[degrees]C. Similarly, Bromfield and David (1976) found an optimum pH (pH 6.5) for oxidation of [Mn.sup.2+] and that no oxidation occurred at pH 5.4 or at pH 7.9. Also, in soil-agar cultures, high concentrations of Mn:+ inhibited the oxidation of [Mn.sup.2+] (e.g. Leeper and Swaby 1940) and high [O.sub.2] pressure inhibited surface oxidation of [Mn.sup.2+] and forced [Mn.sup.2+] oxidation deeper into soil-agar cultures (Uren and Leeper 1978). Thus, there is ample evidence, some direct and some circumstantial, that the oxidation of [Mn.sup.2+] in soil is carried out predominantly by microorganisms.
The chemical reduction of Mn oxides in soils involves competition for electrons between insoluble Mn oxides of varying reactivity and accessibility, and soluble highly mobile oxidants such as molecular [O.sub.2] and nitrate; the reducing agents must be able to diffuse to the Mn oxides if the latter are to be reduced. The reducing agents may be either soluble mobile organic moieties derived from both chemical and microbial sources, or they may be derived from anaerobic respiration. In aerobic soils, molecular [O.sub.2] because of its high mobility and redox potential will be the preferred oxidant rather than Mn oxides. Lovley (1995) contends that the microbial reduction of both Fe and Mn oxides in soils and sediments does not occur in the presence of [O.sub.2] and that most of the readily biodegradable organic matter in soils is unable to reduce Mn[O.sub.x] nonenzymatically. In addition, the non-uniform distribution of Mn oxides in soils (McKenzie 1975; Uren 1990) and of the sites of microbial degradation of biodegradable organic matter (Paul 2007) decreases the likelihood of direct reduction of Mn[O.sub.x] either by microbes or by their metabolic byproducts. Understandably, the opportunities for both microbial and chemical reduction of Mn oxides are limited in aerobic soils but are increased in waterlogged soils in the absence of [O.sub.2] by virtue of Mn[O.sub.x] becoming the alternative electron acceptor after nitrate has been reduced (Ponnamperuma 1972).
As the soil drains, air enters macropores, and the abiotic oxidation of [Fe.sup.2+] and various organic reducing agents begins. The activity of molecular [O.sub.2] in these macropores will depend on the soil water potential, conversely the air porosity, and the temperature. The continuity of drained pores increases as drainage proceeds, but little is known of the air porosity at which microbial [Mn.sup.2+] oxidation commences. It is likely that the microbial oxidation of [Mn.sup.2+] could begin at low air porosity, say 5%, because microbial oxidation has been found to occur commonly in soil and other environments in 'suboxic' zones at oxic/anoxic interfaces (Uren and Leeper 1978; Thompson et al. 2005; Tebo et al. 2010). Once oxidation commences, the WS+E [Mn.sup.2+] would immediately decrease in volumes of soil where the macropores have drained, and eventually an equilibrium would be reached and the WS+E [Mn.sup.2+] would stabilise at a value determined by the relative rates of microbial oxidation of [Mn.sup.2+] and chemical reduction of reactive Mn oxides.
This paper reports the results of five experiments designed to provide greater understanding of the effects of soil water potential, temperature and pH on Mn redox reactions and hence on changes in concentrations of WS+E [Mn.sup.2+], an indicator of the potential for Mn toxicity. The experiments followed 3 years of research in which 5-10-fold increases in WS+E [Mn.sup.2+], up to maximum concentrations of 250mg/kg, were observed in field soils in response to both waterlogging and hot, dry weather (Sparrow and Uren 19876), and where concentrations of WS+E [Mn.sup.2+] >100mg/kg were associated with toxic concentrations of Mn in wheat. Although the effects of temperature and pH on the reduction of Mn oxides and the microbial oxidation of [Mn.sup.2+] were investigated on some of these soils (Sparrow and Uren 1987a), the effects of waterlogging and subsequent drainage were not reported, nor was the effect of varying the soil water potential from near saturation to wilting point.
Biological inhibitors have been used in soil studies to eliminate microbial processes and allow the rate of chemical reduction of Mn oxides to be determined in the absence of microbial [Mn.sup.2+] oxidation (Leeper and Swaby 1940; Mann and Quastel 1946; Grasmanis and Leeper 1966; Uren and Leeper 1978; Sparrow and Uren 1987a). Compared to other inhibitors, sodium azide has been shown to interfere least with purely chemical Mn reactions (Rosson et al. 1984). Here it is assumed, as argued above, that in aerobic soil in the absence of microbial oxidation the increase in WS+E [Mn.sup.2+] is due to chemical Mn[O.sub.x] reduction. No-one has reported on the use of microbial inhibitors to study Mn reactions in soils that are waterlogged or have been drained after waterlogging, and two of the experiments in this paper address this knowledge gap.
Materials and methods
Bulk samples of surface (0-150 mm) silty loams from northeastern Victoria, Australia, were collected from paddocks with a long history under a cereal-pasture rotation, sieved through a 2-mm sieve and stored in a field-moist condition at ambient temperature until required. The soils (hereafter referred to as soil A and soil B) were acidic, with pH (1 : 2 soil: 0.01 M Ca[Cl.sub.2]) of 4.1 and 4.0, respectively. The soils contained Walkley-Black organic C of 1.4% and 1.3%, and had 810 (soil A) and 560 (soil B) mg/kg of Mn reducible by 0.2% quinol in 1 M ammonium acetate at pH 7 (Leeper 1947), and a further 180 and 100 mg/kg, respectively, of Mn reducible by 0.4% sodium dithionite in 1 m ammonium acetate at pH 7. They had previously been studied because of concerns about Mn toxicity in pastures and crops produced on them (Sparrow and Uren 1987b).
Experiment 1. Effect of a microbial inhibitor (sodium azide) on Mn transformations in soil waterlogged at different temperatures
A sample of 400 g (oven-dry soil equivalent) of soil A was weighed into each of 24 pots (8 cm in diameter and 12 cm high). The samples were waterlogged with distilled water to give a 1 cm layer of water at the surface, and then either sodium azide (Na[N.sub.3], 100mg/kg on an oven-dry soil basis, administered as an aqueous solution) or sodium chloride (also as an aqueous solution with the amount of Na equivalent to that added as Na[N.sub.3]) was added to each pot. Preliminary experiments comparing several Na[N.sub.3] rates up to 800 mg/kg had shown that 100 mg/kg was enough to inhibit microbial [Mn.sup.2+] oxidation while minimising interference of exchange reactions by high rates of [Na.sup.+].
Lids were fixed to each pot and a needle was used to make a small hole in each lid to facilitate gas exchange. Three pots of each treatment were then placed at random in the dark in an oven (30[degrees] [+ or -] 1[degrees]C), or controlled environment cabinets (20[degrees] [+ or -] 1[degrees]C, 10[degrees] [+ or -] 1[degrees]C), or in a refrigerated cold room (4[degrees] [+ or -] 1[degrees]C). Water levels in each pot were checked and made good every 2 days. The soil in each pot was sampled after 2, 4, 8 and 14 days, and WS+E [Mn.sup.2+] was extracted using 1% NaHCaEDTA in 1 m ammonium acetate at pH 8.3 (Beckwith 1955). This extractant has the advantage that its pH and Ca content minimise the reduction of Mn oxides during extraction (Beckwith 1955; Bromfield and David 1978). In this experiment and in the others in this paper, WS+E [Mn.sup.2+] was extracted from soil samples at their prevailing water content, and concentrations were subsequently corrected to a dry-weight basis. The solution concentrations of Mn in this and subsequent experiments were measured by flame atomic absorption spectroscopy.
Using GENSTAT, data were analysed by analysis of variance as a completely randomised factorial (azide treatment x temperature x time) design, and for significant treatments and interactions, means were compared by least significant difference (l.s.d.) at P=0.05.
Experiment 2. Effect of a microbial inhibitor on Mn in soil drained after severe waterlogging at 3[degrees]C
Twelve sintered glass funnels, 42 mm in diameter and 40 mm deep, no. 4 porosity, were assembled in a controlled environment cabinet. Eight funnels were connected with tubing to one reservoir and the remaining four funnels to a second reservoir. The funnels, tubing and reservoirs were filled with freshly boiled but cooled distilled water so that no air was present. Soil A (50 g) was added to each funnel and the soil samples were then waterlogged and incubated at 30[degrees]C in the dark. Small plastic cups containing waterlogged soil were also placed in the cabinet and these were sampled every few days to monitor the rate of increase in WS+E [Mn.sup.2+].
After 14 days, the monitoring indicated that all of the reducible Mn had been converted to WS+E [Mn.sup.2+]. At this time, the soils in the funnels were drained to -1 kPa for 15 min to allow the soil in a random selection of four of the eight funnels connected to one reservoir to be sampled and analysed for WS+E [Mn.sup.2+]. The funnels were briefly waterlogged again and 500 mg/kg (oven-dry soil basis) of Na[N.sub.3] as a 1% w/v solution was added to the soil in each of the four funnels connected to one reservoir, while an equivalent amount of Na as 2% w/v sodium chloride solution was added to the soil in the other eight funnels. After 20 min of equilibration, the soils were all drained to -10kPa. Reservoir levels were checked daily and adjusted as necessary. Each funnel was sampled after 3 and 6 days of drainage, and the concentration of WS+E [Mn.sup.2+] determined.
Data were analysed by analysis of variance using GENSTAT, with drainage time and azide treatment as factors. For significant treatments and interactions, means were compared by l.s.d. at P=0.05.
Experiment 3. Effect of liming acidic soils on WS+E [Mn.sup.2+] during and after waterlogging
Four 3-kg (oven-dry soil equivalent) samples of soils A and B were weighed and the following rates (w/w) of analytical grade CaC[O.sub.3] added: 0, 0.08%, 0.15% and 0.32% CaC[O.sub.3], corresponding to additions of 0, 0.4, 0.8 and 1.61 CaC[O.sub.3]/ha, assuming a bulk density of 1.3 t/[m.sup.3]. Samples of soil and lime were then mixed thoroughly and 500 mL of distilled water was added to aid the reaction of the lime. After 2 days of incubation in plastic bags, the gravimetric water content of each sample was brought to 15%, after which the samples were stored in the dark at room temperature for 10 months. The pH of each soil after 5 and 10 months is shown in Table 1.
After 10 months, for each soil and lime treatment, 150-g samples were weighed into 200-mL plastic cups. There were three replicates. Distilled water was added slowly and gently to each cup until a layer of water 5 mm deep at the surface was formed. Each cup was then covered loosely with a plastic bag to reduce evaporation and placed in a completely randomised design in an incubator, in the dark at 30 [+ or -] 1[degrees]C. This temperature was chosen to hasten soil reactions both during waterlogging and after drainage and so decrease the duration of the experiment. Water levels were checked daily and any losses made good. Soil samples were taken for WS+E [Mn.sup.2+] analysis after 4 and 8 days of waterlogging.
After 8 days of waterlogging, surface water was removed by gentle suction and 10 holes were made in the bottom of each cup, using a sharp needle. The cups were then wedged inside a 5-cm-diametcr, no. 2 porosity, sintered glass funnel placed within a rubber collar in the neck of a 1-L Buchner flask. The tip of the sintered glass funnel emptied into a 30-mL test tube inside the Buchner flask. By attaching a vacuum pump to the Buchner flask, the soils in the cups were drained to a water potential of about-10 kPa, i.e. field capacity. Final gravimetric water contents were 37 [+ or -] 1% for soil A and 29 [+ or -] 1% for soil B. The bulk density for both soils was ~1.3 t/[m.sup.3]. The cups were weighed, loosely covered with plastic bags and returned to the incubator. Cups were reweighed daily for the remainder of the experiment and water losses made good.
The filtrate from each sample was stored overnight at 4[degrees]C, refiltered through Whatman 542 filter paper, and analysed for WS+E [Mn.sup.2+]. Further soil samples were taken from each cup for WS+E [Mn.sup.2+] and soil pH analysis at 4 and 8 days after drainage. Preliminary work had shown that the pH of the EDTA extract of unlimed soils was not less than 7.8, a value for which Beckwith (1955) found minimal dissolution of Mn oxides. Therefore, we were confident that there was no effect on WS+E [Mn.sup.2+] due to differences in extract pH between lime treatments.
Data were assessed by analysis of variance using Genstat, with time and lime rate as factors. For significant treatments and interactions, means were compared by l.s.d. at P=0.05. Data for each soil were analysed separately.
Experiment 4, parts (i) and (ii). Interaction of temperature and soil water potential after waterlogging on transformations of soil Mn
For both parts of this experiment, (i) and (ii), 28 sintered glass funnels of the kind used in Expt 2 were organised into four groups of seven, with a separate reservoir for each group. Soil A (40 g of oven-dry soil equivalent) was added to each funnel and flooded from below with freshly boiled but cool distilled water for 7 days in a controlled environment cabinet at 20 [+ or -] 1[degrees]C in the dark. Then the soil from one funnel in each group was sampled and analysed for WS+E [Mn.sup.2+]. Half of the funnels were then carefully moved to another controlled environment cabinet set a 10 [+ or -] 1[degrees]C.
Next, in part (i), half of the unsampled funnels at each temperature were drained to -10 kPa, and the rest to -1 kPa. In part (ii), the drainage treatments were -5 and -2 kPa. In each part, three randomly selected funnels at each temperature were sampled after 2 and 8 days and the samples analysed for WS+E [Mn.sup.2+]. Immediately before sampling for Mn analysis, a small sample of soil was taken near the wall of the funnel for determination of water content. During incubation, the drained soils were loosely covered with a plastic bag, and reservoir water levels were maintained daily.
Using GENSTAT, the data were analysed by 3-way (drainage x temperature x time) factorial analysis of variance, assuming a completely randomised design. For significant treatments and interactions, means were compared by l.s.d. at P=0.05.
Experiment 5. Effect of different temperatures and soil water potentials on the oxidation and reduction of Mn in aerobic soils
Samples of soils A (WS+E [Mn.sup.2+], 54 mg/kg) and B (WS+E [Mn.sup.2+], 272 mg/kg) equivalent to 5 g of oven-dry soil were weighed into pre-weighed, 100-mL polypropylene centrifuge tubes and wet up to -1, -10, -100 or-1500 kPa water potential according to soil-water release curves for each soil determined previously. The corresponding gravimetric water contents were 42%, 34%, 16% and 8% for soil A, and 33%, 29%, 10% and 3% for soil B. The tube lids were tightly fitted and nine samples of each soil x water potential treatment were randomly arranged on trays and placed in the dark at each of the following temperatures: 30 [+ or -] 1[degrees]C (oven), 20 [+ or -] 1[degrees]C and 10 [+ or -] 1[degrees]C (incubators), and 4 [+ or -] 1[degrees]C (refrigerated cold room).
Three of each group of nine samples had been moistened with a solution of sodium azide to give a final concentration of 100 mg/kg on an oven-dry soil basis, and to the other six samples of soil A, 200 mg/kg of Mn as MnS[O.sub.4] x 4[H.sub.2]O was added in the moistening process to provide a starting concentration of [Mn.sup.2+] similar to that in soil B. The water content of the soils was checked every second day and losses made good with distilled water. All tubes were opened briefly on these occasions to allow gas exchange.
The samples to which sodium azide had been added were removed and extracted for WS+E [Mn.sup.2+] on day 2. Rates of Mn[O.sub.x] reduction for each treatment were estimated from the increase in the concentration of WS+E [Mn.sup.2+] due to addition of azide. This was done for each treatment by subtracting the mean concentration of WS+E [Mn.sup.2+] in azide-treated soil from the mean starting concentration. No allowance was made for any effect on WS+E [Mn.sup.2+] of the sodium added with the azide because preliminary work had shown no such effect at the concentrations used in this experiment. Half of the remaining samples in each treatment were analysed for WS+E [Mn.sup.2+] on day 4, and the rest on day 8.
For each soil, data from days 4 and 8 were analysed by Genstat using a completely randomised temperature x time x water potential factorial analysis of variance, and for significant treatment effects, the treatment means were compared by the l.s.d. at P = 0.05.
Results and discussion
The concentrations of WS+E [Mn.sup.2+] for each treatment at each sampling date are shown in Fig. 1. There was a significant temperature x time interaction, with faster increases in the concentration of WS+E [Mn.sup.2+] at higher temperatures. The importance of temperature in the reduction of Mn during waterlogging was mentioned by Ponnamperuma (1972), and Hofacker et al. (2013) reported a positive effect of increasing temperatures from 5[degrees]C to 23[degrees]C on the release of dissolved [Mn.sup.2+] from submerged, contaminated, floodplain soils, but Expt 1 represents the first evidence of the magnitude of this effect in agricultural soils. Yamane and Sato (1967) showed that increasing the temperature from 10[degrees]C to 35[degrees]C during flooding of soil increased the rate of decomposition of added glucose, starch, cellulose and gelatine, thus providing more substrate for anaerobic respiration and reduction of Mn[O.sub.x] (Lovley 1995).
A significant inhibition by azide of the release of WS+E [Mn.sup.2+] was observed after 2 days at 30[degrees]C and after 8 days at 20[degrees]C, but after this there was less effect of azide on the rate of release of [Mn.sup.2+] at these temperatures (Fig. 1). The rate of increase in the concentration of WS+E [Mn.sup.2+] in azide-treated soil at 30[degrees]C slowed greatly after day 8, possibly because, by this time, almost all of the Mn oxides in the soil had been reduced to [Mn.sup.2+], or the energy supply (organic matter) had mn out.
There was little effect of sodium azide at either 4[degrees]C or 10[degrees]C, which suggests that microbial Mn[O.sub.x] reduction at these temperatures was minimal. Only very slight increases in the concentration of WS+E [Mn.sup.2+] were observed at 4[degrees]C and 10[degrees]C in soil treated with chloride, so chemical Mn[O.sub.x] reduction at these temperatures did not appear to be very great either. Although it is possible that a longer incubation may have revealed an effect of azide at 4[degrees]C and 10[degrees]C, other unpublished work from our laboratory showed little or no acceleration in the release of [Mn.sup.2+] during 4 weeks of waterlogging at these temperatures.
The results (Table 2) show that the waterlogging appeared to reduce all of the Mn oxides in soil A (>900mg/kg, soil A contained 810mg/kg of Mn reducible in quinol), so that the subsequent rate of chemical Mn[O.sub.x] reduction under drainage would probably have been very low. In the samples to which azide was added at the end of the waterlogging, no re-oxidation of [Mn.sup.2+] had occurred after 6 days of drainage, but in the same time, extensive [Mn.sup.2+] re-oxidation took place in the samples to which chloride was added (Table 2). We conclude that, under drainage, no chemical oxidation of [Mn.sup.2+] occurred but that microbial [Mn.sup.2+] oxidation was significant. The possibility exists that chemical [Mn.sup.2+] oxidation in the azide-treated soil was precisely offset by chemical Mn[O.sub.x] reduction but this seems unlikely, and the rates involved would probably have been very small compared with the rate of microbial oxidation in the chloride-treated soil. This experiment represents further evidence that [Mn.sup.2+] oxidation in soil is a microbial process and not a chemical one, and that microbial [Mn.sup.2+] oxidation can take place soon after severe waterlogging; the air porosity at -10 kPa is ~33%. Siman et al. (1974) and Sparrow and Urcn (1987/)) both measured rapid decreases in WS+E [Mn.sup.2+] in the field following drainage of waterlogged soils, but this is the first evidence clearly linking such decreases to microbial [Mn.sup.2+] oxidation.
Four days after drainage, the soil pH (Table 3) in most treatments was 0.1 or 0.2 units higher than at the beginning of the experiment (Table 1), and there was little change after a further 4 days. Presumably, this was a residual effect of waterlogging, which has been shown to increase the pH of acidic soils (Ponnamperuma 1972; Morales et al. 2002).
Concentrations of [Mn.sup.2+] in the leachates at day 8 (Table 3) generally reflected the concentrations of WS+E [Mn.sup.2+] for the corresponding treatment at that time (Fig. 2). The addition of lime at the lowest rate did not appear to alter the rate of increase in WS+E [Mn.sup.2+] under waterlogged conditions for either soil (Fig. 2). The main effect of lime at this rate was to establish a lower initial concentration of WS+E [Mn.sup.2+] compared with the unlimcd soil. With higher lime rates, the initial concentration of WS+E [Mn.sup.2+] was even lower and its rate of increase under waterlogged conditions was less than rates for unlimed soil. The rates of increase in the concentrations of WS+E [Mn.sup.2+] appeared to decrease after day 4 in some treatments, but this behaviour was not consistent.
The failure of the lowest rate of lime to inhibit the rate of increase in the concentration of WS+E [Mn.sup.2+] under waterlogging may be due to a stimulation of microbial activity and hence microbial Mn[O.sub.x] reduction at the raised pH, as suggested by Graven et al. (1965). For the lowest rate of lime used in this experiment, such stimulation may have been sufficient to compensate for the decrease in the rate of chemical Mn[O.sub.x] reduction that would occur with liming (Leeper 1947). However, at the two higher rates of lime, any stimulation of microbial reduction may not have been sufficient to offset the decrease in chemical Mn[O.sub.x] reduction, thus accounting for the net decrease in the rate of increase observed for these treatments.
These results indicate that, over the time span of this experiment, liming at a rate of at least 0.16% was necessary to decrease the rate of reduction of Mn[O.sub.x] in soil waterlogged at 30[degrees]C. Liming at a rate of 0.08% gave rise to a lower initial concentration of WS+E [Mn.sup.2+] but this advantage was diminished the longer the soil was waterlogged, because the WS+E [Mn.sup.2+] in this treatment increased at a rate similar to that of the control. Liming at higher rates proved much more advantageous because, not only did such treatment lower the initial WS+E [Mn.sup.2+] concentration even further, it also decreased the rate of increase in the concentration of WS+E [Mn.sup.2+]. Longer periods of waterlogging would be required for the WS+E [Mn.sup.2+] to reach toxic concentrations in the soils limed at the higher rates. For example, the unlimed soil B needed ~3 days of waterlogging to exceed the WS+E [Mn.sup.2+] toxicity threshold of 100 mg/kg suggested by the field study of Sparrow and Uren (19876), whereas soil B treated with the highest rate of lime did not reach this threshold after 8 days of waterlogging (Fig. 2).
The WS+E [Mn.sup.2+] in both soils at the two highest rates of lime had returned to their respective starting concentrations by day 12, after 4 days of drainage (Fig. 2), and at the lowest lime rate of 0.08%, after 8 days of drainage. The ability of a soil to re-oxidise Mn rapidly may be important if that soil is drained for only short periods between episodes of waterlogging. Pierce et al. (2010) measured higher concentrations of Mn in plants grown in continuously waterlogged soil than in plants grown in intermittently waterlogged soil, but they made no measurements of soil Mn.
The optimum lime rate in Expt 3 was 0.16% because it was the lowest rate able to prevent WS+E [Mn.sup.2+] from exceeding 100 mg/kg, a concentration associated with Mn toxicity in wheat (Sparrow and Uren 19876), and because it was the lowest rate able to allow WS+E [Mn.sup.2+] to return to its starting concentration after 4 days of drainage (Fig. 2). However, depending on the duration of waterlogging and the prevailing temperature, liming alone may not be able to prevent Mn toxicity in the field. Siman et al. (1974) found that lime, despite increasing soil pH by 0.5-0.7 units and decreasing extractable Mn and plant Mn, did not prevent Mn toxicity in waterlogged pasture plants.
Our findings need to be confirmed at temperatures other than 30[degrees]C, because Expt 1 showed that changes in temperature could affect microbial and chemical Mn[O.sub.x] reduction differently.
The changes in the concentration of WS+E [Mn.sup.2+] following drainage are shown in Fig. 3 for each temperature and level of drainage. The means and standard deviations of the gravimetric soil water content for these treatments are given in Table 4, and they show a steady gradient from 55% at saturation to ~30% at -10 kPa water potential. Assuming a bulk density of 1.3t/[m.sup.3], -10 kPa corresponds to an air porosity of ~34%. There was no obvious difference in the volume and hence density of soil in funnels of different drainage treatments. Table 4 also shows consistency at a given water potential between the 10[degrees]C and 20[degrees]C incubation cabinets.
Significant net re-oxidation of Mn occurred at both temperatures but only at -5 and -10kPa (Fig. 3). At these drainage levels, there was presumably sufficient air in the soil for microbial [Mn.sup.2+] oxidation to be the dominant Mn process. No directly comparable studies on Mn are known to the authors, although Shuman (1980) did incubate soils for 2 weeks at 0, -33 and -500 kPa and found little effect of temperature (10[degrees]C, 25[degrees]C, 40[degrees]C) on extractable Mn at the two lower (aerobic) water potentials. The optimum water potentials for other aerobic soil microbial processes such as respiration and nitrification were reviewed by Rodrigo et al. (1997) and were found to be in the range -10 to -50 kPa.
At -2 kPa in Expt 4, there was no significant change in the concentration of WS+E [Mn.sup.2+] over 8 days, and this was also the case at -1 kPa at 10[degrees]C (Fig. 3). However, at 20[degrees]C there was a significant increase in the concentration of WS+E [Mn.sup.2+] after 8 days of drainage at -1 kPa. We conclude that, at these low levels of drainage, sufficient anaerobic sites remained in the soil for the rate of [Mn.sup.2+] oxidation to be slow and either in balance with or less than the rate of Mn[O.sub.x] reduction. Anaerobic microsites have been found within unsaturated soil aggregates (Tiedje et al. 1984; Sexstone et al. 1985), while Helyar and Conyers (1987) consider that partially drained soils, with saturated aggregates but air-filled macropores, could allow Mn[O.sub.x] reduction to continue inside the aggregates. So far, evidence of anaerobic microbial activity in restricted parts of the soil such as microsites appears limited to studies of denitrification (Hojberg et al. 1994; Stolk et al. 2011) and methane production (Kammann et al. 2009), rather than Mn oxide reduction. Whether anaerobic microsites occur inside aggregates depends very much on the porosity and the poresize distribution of the aggregates and on the concentrations and accessibility (to microbes) of readily biodegradable organic matter inside the aggregate. Unfortunately, such attributes are rarely measured.
The interaction between temperature and degree of drainage with respect to soil Mn is important because it may mean that the extent of drainage necessary to decrease Mn availability after waterlogging, or at least slow down further increases, is greater in spring than in winter. Sparrow and Uren (19876) observed monthly peaks in WS+E [Mn.sup.2+] in field soils in winter and in spring, but weekly or more frequent measures would have enabled a comparison of the speed with which concentrations rose to a maximum and then returned to the baseline in these seasons. The temperature x drainage interaction may be due to the effects of [O.sub.2] availability on the relative rates of microbial oxidation of [Mn.sup.2+] and reduction of Mn[O.sub.x]. Especially in poorly drained soil, each temperature may favour a particular suite of microorganisms, with higher temperatures (lower [O.sub.2] concentrations) favouring Mn[O.sub.x]reducing (anaerobic) organisms, and low temperatures (higher 02 concentrations) favouring [Mn.sup.2+]-oxidising (aerobic) organisms. Oxygen availability should influence the rate of chemical Mn[O.sub.x] reduction too, because [O.sub.2] and Mn oxides both accept electrons from organic reducing agents in soil. Therefore, the influence of [O.sub.2] availability on chemical Mn[O.sub.x] reduction should complement its influence on microbial Mn[O.sub.x] reduction, with low [O.sub.2] concentrations also favouring chemical Mn[O.sub.x] reduction.
Water potential had no effect on the change in the concentration of WS+E [Mn.sup.2+] in either soil at 4[degrees]C or 10[degrees]C during the 8-day incubation (Figs 4 and 5). At no water potential was there a statistically significant change in the concentration of WS+E [Mn.sup.2+] in cither soil at 4[degrees]C between day 4 and day 8, but at 10[degrees]C, soil A showed a slight increase at all water potentials (Fig. 4). At 20[degrees]C and 30[degrees]C, the concentration of WS+E [Mn.sup.2+] of the soils held at -100 kPa and -1500 kPa either increased (soil A with added Mn, Fig. 4) or changed little (soil B, Fig. 5). It is not surprising that soil B showed little increase given its already high concentration of WS+E [Mn.sup.2+]. The increases in the concentration of WS+E [Mn.sup.2+] at -100 kPa and -1500 kPa were greater at 30[degrees]C than at 20[degrees]C.
At -10 kPa and -1 kPa, the concentration of WS+E [Mn.sup.2+] in both soils at 20[degrees]C and 30[degrees]C decreased over the incubation period, the decrease being greater at 20[degrees]C in soil A and at 30[degrees]C in soil B (Figs 4 and 5). For both soils the decrease was greatest at -1 kPa.
Increases in the concentration of WS+E [Mn.sup.2+] due to microbial inhibition by azide were greatest in soil A (Table 5), probably due to its lower initial concentration of WS+E [Mn.sup.2+] and hence its higher inherent propensity for Mn[O.sub.x] reduction. In both soils, the increases were greater at 20[degrees]C and 30[degrees]C (Table 5), as expected, and at 4[degrees]C and 10[degrees]C there was little difference in the rates of increase between different water potentials. At 20[degrees]C and especially at 30[degrees]C, rates of increase in the concentration of WS+E [Mn.sup.2+] of azide-treated soil at -1500kPa were markedly slower that at the higher water potentials (Table 5).
Rates of [Mn.sup.2+] oxidation for these soils at each temperature and water potential are shown in Tables 6 and 7, and were calculated from the corresponding rates of chemical reduction of Mn oxides (Table 5), and the rates of change in the concentration of WS+E [Mn.sup.2+] between days 4 and 8 (Figs 4 and 5). A similar approach was used by Sparrow and Uren (1987a) for soils limed to different pH and similar assumptions apply here, namely that the rate of microbial Mn[O.sub.x] reduction was negligible and that the observed increases in the concentration of WS+E [Mn.sup.2+] in azide-treated soils indicate a rate of chemical Mn[O.sub.x] reduction the same as that in non-azide-treated soil. Manganese oxide reduction by facultative anaerobes is a dominant Mn process in waterlogged soils (Ponnamperuma 1972), but while some microbial Mn[O.sub.x] reduction coupled to the oxidation of organic matter has been observed in solution culture (Ghiorse 1988), it is not considered a significant process in aerobic soil (Lovley 1995).
This analysis shows that actual [Mn.sup.2+] oxidation at 20[degrees]C and 30[degrees]C (Tables 6 and 7) was consistently greater than the rate of decrease in the concentration of WS+E [Mn.sup.2+] indicated in Figs 4 and 5, and that some [Mn.sup.2+] oxidation occurred at 10[degrees]C and 4[degrees]C in soil A but that oxidation in soil B at these temperatures was negligible (Tables 6 and 7). The increase in the concentration of WS+E [Mn.sup.2+] at -100kPa and -1500kPa for soil A at 20[degrees]C and 30[degrees]C (Fig. 4) was insufficient to account for the rate of reduction observed (Table 6). Therefore, some oxidation of Mn must have occurred in this soil at these water potentials (Table 6). Zak et al. (1999) also observed a temperature x water potential interaction in a study of microbial respiration and nitrogen mineralisation. Microbial activity generally increased with incubation temperature, but at the highest temperature of 25[degrees]C, activity was restricted at water potentials less than -300 kPa, whereas activity was not affected by water potential at 5[degrees]C and 10[degrees]C. They attributed the interaction to the decreasing ability of the substrates to diffuse in the drier soil, a limitation only exposed by the higher substrate demand at 25[degrees]C.
For both soils at 20[degrees]C and 30[degrees]C, a change in water potential from -10kPa to -100kPa altered their behaviour dramatically. Instead of showing a decrease in the concentration of WS+E [Mn.sup.2+], there was either no change or an increase in WS+E [Mn.sup.2+] (Figs 4 and 5). Because the rate of Mn[O.sub.x] reduction was little affected by such a change in water potential (Tables 6 and 7), it appears that [Mn.sup.2+] oxidation was severely restricted at -100 kPa compared with-10kPa. At-1500kPa, the rate of reduction was slowed even further, and therefore [Mn.sup.2+] oxidation must have been restricted even more at this water potential than at -100 kPa (Tables 6 and 7) to account for the continued increase in the concentration of WS+E [Mn.sup.2+] in soil A, and the absence of a decrease in soil B (Figs 4 and 5). These observations support the suggestion that [Mn.sup.2+] oxidation is favoured by the microaerophilic conditions, which presumably exist near or at oxic-anoxic interfaces (Thompson et al. 2005; Tebo et al. 2010).
For the two highest water potentials, increasing temperature enhanced [Mn.sup.2+] oxidation more than reduction, which is in accord with the findings of Sparrow and Uren (1987a) where soils were incubated only at -10kPa. The result for -1 kPa is, however, contrary to that for Expt 4, where soils drained at about -1 kPa showed greater Mn[O.sub.x] reduction relative to oxidation at 20[degrees]C compared with 10[degrees]C. In the current experiment, the small amount of soil (5 v. 45 g) incubated in a relatively large tube may have allowed greater diffusion of oxygen through the samples, so favouring [Mn.sup.2+] oxidation (Schjonning et al. 2003).
An assumption in the calculations in Tables 6 and 7 is that the rate of reduction was the same over the final 4 days of the experiment as at the start. The work of Sparrow and Uren (1987a) indicated that this might not be the case where a substantial decrease in the concentration of WS+E [Mn.sup.2+] occurred during the incubation. In such cases (e.g. for soil B at 30[degrees]C, at -1 and -10kPa) the rate of Mn[O.sub.x] reduction at the end of the incubation may have been greater than at the start, when it was measured. Such an effect would only serve to increase the difference in [Mn.sup.2+] oxidation rate between the wet soils and the drier ones, for the latter would probably show a more constant rate of Mn[O.sub.x] reduction during the incubation.
When the soils in Expt 5 were at water potentials less (drier) than -10 kPa the rate of [Mn.sup.2+] oxidation was restricted much more than the rate of reduction of Mn oxides (Tables 6 and 7). This result is expected, since the oxidation of [Mn.sup.2+] is microbial and will be restricted increasingly as the soil water retreats to pores too small to accommodate active colonies of microbes. Stark and Firestone (1995) proposed that microbial cell dehydration was rate-limiting for nitrification at water potentials less (drier) than -600 kPa, whereas substrate availability was the limiting factor in wetter soils. In near air-dry soils, the reduction reaction may not be occurring at all, and it may only be revealed when the soil is wet up at the time of extraction. This may account for the lower rates of apparent Mn oxide reduction in both soils at 20[degrees]C and 30[degrees]C, at -1500kPa compared with higher water potentials (Table 5).
While [Mn.sup.2+] oxidation in our study was less in drier soil, some oxidation did take place, even at -1500 kPa. Nitrification and nitrogen mineralisation have both been observed in soils drier than this (Justice and Smith 1962; Myers et al. 1982). In a climate where fluctuations in soil water potential can be extreme (Sparrow and Uren 19876), it is not surprising to find soils with measurable microbial [Mn.sup.2+] oxidation at a water potential as low as wilting point.
The water potentials in this experiment were obtained by wetting up soil that was initially drier than the driest treatment. Different behaviour, particularly the behaviour of microbial populations, may have been observed in soil dried down rather than wetted up to these same water potentials, so the experiment should be repeated with soil water status controlled in this way. Any expression of Mn toxicity will depend not only on the effect of changes in water potential on soil Mn reactions, but also on the effect of such changes on plant-water relations and hence Mn uptake.
The laboratory experiments described in this paper show a delicate balance between [Mn.sup.2+] oxidation and Mn[O.sub.x] reduction in soils, particularly in poorly drained soils. Relatively small changes in water potential, along with changes in the temperature and pH, can cause this balance to tip in favour of oxidation (decreased Mn availability) or reduction (increased Mn availability), but steady-state situations of zero net change are also possible. In several instances, the behaviour of Mn observed in these experiments parallels observations of other microbially driven processes in both aerobic and anaerobic soils. A major challenge for further research is to assess more fully the implications of our findings for practical soil management and control of Mn toxicity.
Received 27 May 2013, accepted 5 December 2013, published online 9 May 2014
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L. A. Sparrow (A,B,c) and N. C. Uren (A)
(A) Department of Agricultural Sciences, La Trobe University, Bundoora, Vic. 3086, Australia.
(B) Present address: Tasmanian Institute of Agriculture, University of Tasmania, Launceston, Tas. 7250, Australia.
Corresponding author. Email: Leigh.Sparrow@utas.edu.au
Table 1. experiment 3. Effect of CaC[O.sub.3] on soil pH (1 : 2 Ca[Cl.sub.2]) after 5 and 10 months of incubation at 15% gravimetric water content at room temperature Rate of CaC[O.sub.3] Soil A Soil B 5 months 10 months 5 months 10 months 0 4.0 4.0 4.0 3.8 0.08% (0.4 t/ha) 4.3 4.2 4.5 4.2 0.16% (0.8 t/ha) 4.9 4.7 5.0 4.9 0.32% (1.6 t/ha) 5.9 5.7 6.4 6.2 Table 2. Experiment 2. Comparative effect of sodium azide and sodium chloride on the concentrations of water-soluble plus exchangeable (WS+E) [Mn.sup.2+] following the draining of severely waterlogged soil A WS+E [Mn.sup.2+] at the end of waterlogging = 968 [+ or -] 7 mg/kg. Soil held at 30[degrees]C throughout the experiment. l.s.d. (P = 0.05) for treatment x time interaction = 174 mg/kg Time of drainage at -10 kPa 3 days 6 days Na azide (abiotic), +500 mg/kg 947 959 Na chloride (biotic), +450 mg/kg 881 521 Table 3. Experiment 3. Concentration of [Mn.sup.2+] in leachate after 8 days of waterlogging and soil pH after 4 and 8 days of drainage (days 12 and 16) at -10 kPa Soils held at 30[degrees]C throughout the expt Rate of lime Soil A Leachate pH at pH at [Mn.sup.2+] (mg/L) day 12 day 16 0 78 4.0 3.9 0.08% (0.4 t/ha) 35 4.4 4.3 0.16% (0.8 t/ha) 5 5.1 5.0 0.32% (1.6 t/ha) 1 6.3 6.3 Rate of lime Soil B Leachate pH at pH at [Mn.sup.2+] (mg/L) day 12 day 16 0 35 4.0 3.9 0.08% (0.4 t/ha) 32 4.3 4.3 0.16% (0.8 t/ha) 8 4.8 4.8 0.32% (1.6 t/ha) 1 5.9 5.9 Table 4. Experiment 4. Gravimetric water contents (% [+ or -] standard deviation) of soil A at each sampling (after 2 and 8 days of drainage) Drainage Temp. 0 -1 kPa 0 2 days 8 days 10[degrees]C 55 53 [+ or -] 0.7 52 [+ or -] 2.0 20[degrees]C 55 52 [+ or -] 0.3 51 [+ or -] 0.3 Drainage Temp. 0 -2 kPa 0 2 days 8 days 10[degrees]C 55 46 [+ or -] 0.1 47 [+ or -] 0.9 20[degrees]C 55 46 [+ or -] 2.0 46 [+ or -] 0.9 Drainage Temp. 0 -5 kPa 0 2 days 8 days 10[degrees]C 55 38 [+ or -] 0.9 38 [+ or -] 1.3 20[degrees]C 55 37 [+ or -] 1.5 37 [+ or -] 1.5 Drainage Temp. 0 -10 kPa 0 2 days 8 days 10[degrees]C 55 29 [+ or -] 1.1 28 [+ or -] 0.6 20[degrees]C 55 31 [+ or -] 1.0 29 [+ or -] 1.1 Table 5. Experiment 5. Effect of temperature and soil water potential on the increase in the concentration of water-soluble plus exchangeable (WS+E) [Mn.sup.2+] (mg/kg.day) of soils A and B treated with 100mg/kg of sodium azide Soil water 4[degrees]C 10[degrees]C 20[degrees]C 30[degrees]C potential Soil A -1 kPa 7 12 25 56 -l0kPa 7 14 26 56 -l00kPa 8 15 28 59 -1500 kPa 8 14 24 47 l.s.d. (P=0.05) water potential x temperature = 0.9 Soil B -1 kPa 0 1 10 24 -l0kPa 1 2 8 24 -l00kPa 0 1 10 21 -1500kPa 0 1 6 13 l.s.d. (P = 0.05) water potential x temperature = 2.5 Table 6. Experiment 5. Calculated rates of oxidation of Mn for soil A at various temperatures and soil water potentials WS+E, Water-soluble plus exchangeable Soil water Rate of Rate of Rate of potential Mn[O.sub.x] decrease in cone, [Mn.sup.2+] reduction of WS+E [Mn.sup.2+] oxidation (mg/kg.day) (A) (mg/kg.day) (B) (mg/kg.day) 4[degrees]C -1 kPa 7 -1 6 -10kPa 8 -2 6 -100kPa 8 0 8 -1500kPa 8 -1 7 10[degrees]C -1 kPa 12 -3 9 -10kPa 14 -3 11 -100kPa 15 -3 12 -1500kPa 13 -3 10 20[degrees]C -1 kPa 25 15 40 -10kPa 26 11 37 -100kPa 28 -2 26 -1500kPa 24 5 19 30[degrees]C -1 kPa 56 8 64 -10kPa 56 6 62 -100kPa 59 -7 52 -1500kPa 47 -12 35 (A) From Table 5. (B) From Fig. 4. Table 7. Experiment 5. Calculated rates of oxidation of Mn for soil B at various temperatures and soil water potentials WS+E, Water-soluble plus exchangeable Rate of Rate of Rate of Mn[O.sub.x] decrease in cone, [Mn.sup.2+] reduction of WS+E [Mn.sup.2+] oxidation (mg/kg.day) (A) (mg/kg.day) (B) (mg/kg.day) 4[degrees]C -1kPa -1 0 -1 -10kPa 1 0 1 -100kPa -1 0 -1 -1500kPa -1 0 -1 10[degrees]C -1kPa 1 1 2 -10kPa 2 1 3 -100kPa 1 0 1 -1500kPa 1 1 2 20[degrees]C -1kPa 9 13 22 -10kPa 8 10 18 -100kPa 10 -1 9 -1500kPa 6 0 6 30[degrees]C -1kPa 24 15 39 -10kPa 24 13 37 -100kPa 21 1 22 -1500kPa 13 -1 12 (A) From Table 5. (B) From Fig. 5.
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|Author:||Sparrow, L.A.; Uren, N.C.|
|Date:||Aug 1, 2014|
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