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Influence of Soil Properties on Zinc Solubility Dynamics Under Different Redox Conditions in Non–Calcareous Soils
Pedosphere 27(1): 96–105, 2017 doi:10.1016/S1002-0160(17)60299-6 ISSN 1002-0160/CN 32-1315/P c 2017 Soil Science Society of China ⃝ Published by Elsevier B.V. and Science Press
Influence of Soil Properties on Zinc Solubility Dynamics Under Different Redox Conditions in Non-Calcareous Soils Michelle Anne Belen BUNQUIN1,∗ , Susan TANDY2 , Sarah Johnson BEEBOUT1 and Rainer SCHULIN2 1 Crop 2 Soil
and Environmental Sciences Division, International Rice Research Institute, College, Los Ba˜ nos, Laguna 4031 (Philippines) Protection, Institute of Terrestrial Ecosystems, ETH Zurich, Z¨ urich CH-8092 (Switzerland)
(Received April 22, 2016; revised November 14, 2016)
ABSTRACT Zinc (Zn) deficiency in paddy soils is often a problem for rice production. Flooding can decrease metal availability in some noncalcareous soils through different mechanisms associated with soil redox status. Laboratory experiments were performed in order to better understand the processes that governed the dynamics of Zn in non-calcareous paddy soils at varying redox potentials (Eh). Airdried non-calcareous soil samples collected from four different paddy field sites in the Philippines were submerged and incubated in a reaction cell with continuous stirring and nitrogen purging for 4 weeks, and then purged with compressed air for another week to reoxidize the system. The Eh of the four soils started at 120 to 300 mV, decreased to −220 to −300 mV after 100 to 250 h of reduction, and was maintained at this low plateau for about 2 weeks before increasing again upon reoxidation. Zinc solubility showed contrasting patterns in the four soils, with two of the soils showing a decrease in soluble Zn as the Eh became low, probably due to zinc sulfide (ZnS) precipitation. In contrast, the other two soils showed that Zn solubility was maintained during the reduced phase which could be due to the competition with iron (Fe) for precipitation with sulfide. Differences in the relative amounts of S, Fe, and manganese (Mn) oxides in the four soils apparently influenced the pattern of Zn solubility after flooding. Key Words:
microcosm, paddy soil, precipitation, rice, soil sulfide, speciation, Zn availability
Citation: Bunquin M A B, Tandy S, Beebout S J, Schulin R. 2017. Influence of soil properties on zinc solubility dynamics under different redox conditions in non-calcareous soils. Pedosphere. 27(1): 96–105.
INTRODUCTION Zinc (Zn) deficiency limits grain yield and nutritional content in rice and is the most common in perennially wet or calcareous soil. It is an essential nutrient for humans for the development of the immune system and gastrointestinal tract, so that an inadequate supply of Zn causes stunted physical growth, susceptibility to infections, and neurobehavioral abnormalities (Brown et al., 2001). Zinc measurements on dry soil samples are ineffective in predicting Zn fertilizer responses in flooded soils (Impa and Johnson-Beebout, 2012), so it would be useful to identify other routinely measurable soil parameters that may increase our ability to predict flooding-induced Zn deficiency in rice fields. Flooding changes many aspects of soil chemistry which are known to affect Zn availability. The primary redox-induced changes are related to the following factors: available sulfur (S) as sulfate and sulfide, amorphous iron (Fe) and manganese (Mn) (hydr)oxides, organic matter (OM), pH, and soil solution bicarbonate ∗ Corresponding
author. E-mail: [email protected].
concentrations (Neue and Mamaril, 1985; Du Laing et al., 2009). Sulfides can decrease the mobility of trace metals such as Zn, Fe, cadmium (Cd) and copper (Cu) in soil by formation of insoluble sulfide precipitates upon flooding (Du Laing et al., 2007; de Livera et al., 2011). Iron and Mn (hydr)oxides provide adsorption sites that can immobilize trace metals, which can then be released under anaerobic conditions by reductive dissolution of the oxides, and adsorbed again during the reoxidation phase to less crystalline (hydr)oxides that are formed by reprecipitation (Du Laing et al., 2009). Organic matter provides a source of carbon (C) for microorganisms which leads to a faster development of reducing conditions upon flooding. It also protects some metal ions (e.g., Zn, Cd, nickel (Ni), and Cu) from adsorption to the solid phase through complexation with soluble humic and fulvic acids, thus increasing metal mobility (Du Laing et al., 2009). Soil pH affects metal solubility by influencing the formation of metal complexes with dissolved organic C (DOC) and the adsorption of Zn, Fe, and Mn to Fe and Mn (hydr)oxides (Kirk, 2004). The formation of Zn carbonates
INFLUENCE OF SOIL PROPERTIES ON ZN SOLUBILITY
may reduce Zn solubility at either high (in a calcareous soil) or low redox potential (Eh) (in any soil) due to the increase in partial pressure of CO2 as OM decomposes (Kirk, 2004). Also, high bicarbonate concentrations in soil solution hinder root growth and cause root leakage of Zn-inefficient rice genotypes (Rose et al., 2011). There has been disagreement in the literature about which of these processes controls Zn solubility in flooded soils. Kittrick (1976) explained that high hydrogen sulfide (H2 S) could depress Zn2+ concentrations to very low levels in flooded soils by forming zinc sulfide (ZnS), while Lindsay (1979) theorized that iron pyrite (FeS2 ) formation should consume S when the Fe redox reactions control the equilibrium Eh at a level not favorable for ZnS formation. Subsequent studies in Zn-contaminated soils have supported both sides, confirming both that Fe can interfere with ZnS formation in sulfate-limited soils (Weber et al., 2009), and ZnS can be formed in microsites (Hesterberg et al., 2011). However, the relative importance of ZnS formation in Zn-limited soils is not well understood. To isolate geochemical reactions from some of the many variables which can cause imprecise data in field experiments, reaction cells have been used to study redox chemistry questions regarding metal availability in soil (Yu et al., 2007). Continuously stirred reaction cells allow precise monitoring of chemical changes in homogenized soil slurries as the Eh varies. However, stirring with continuous nitrogen (N2 ) purging does not allow the normal build-up of dissolved CO2 in solution, and therefore does not allow meaningful measurement of carbonate equilibria (de Livera et al., 2011). Our overall aim was to understand which types of soils are likely to show flooding-induced Zn deficiency in rice. We hypothesized that although Zn does not change oxidation state due to soil Eh, Zn solubility would be influenced by other redox-active soil parameters, including changes in speciation of S, Fe, and Mn. Our objectives were to understand which chemical reactions are the most important in explaining the decrease in Zn availability in some soils after flooding and to identify a set of soil parameters that may be useful for predicting which rice fields are at risk of floodinginduced Zn deficiency. MATERIALS AND METHODS Soil sampling Samples of non-calcareous soils were collected from four different paddy field sites in the Philippines: an Aeric Endoaquept from Central Mindanao University, Bukidnon (BKN) (7◦ N, 125◦ E) (Carating et al.,
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2014); a Vertic Tropaquept from Bay, Laguna (BAY) (14◦ N, 121◦ E) (Moormann and van Breemen, 1978); an Entic Pellustert from PhilRice, Nueva Ecija (NE) (15◦ N, 120◦ E) (Cassman et al., 1996); and an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna (IRRI) (14◦ N, 121◦ E) (Dobermann et al., 2000). Rice of the first three sites shows moderate Zn-deficiency symptoms in most years, while the fourth site is considered as Znsufficient. The top 20 cm of soil was collected from each site, followed by air drying and homogenization during the grinding process to pass through a 2-mm sieve. Initial soil characterizations are shown in Table I. Experimental setup A laboratory experiment was set up using a modified reaction cell system similar to Yu et al. (2007), with two replicates for each soil. The setup (Fig. 1) consisted of a 2-L reaction beaker containing 400 g soil, 1 600 mL deionized water, and 4 g ground rice straw as a source of OM (C, 391.0 g kg−1 ; N, 8.9 g kg−1 ; S, 1.05 g kg−1 ; Zn, 24 mg kg−1 ; and C/N ratio, 43.9:1). Different components were placed in the system to accommodate the measurements, including gas inlets and outlets, soil slurry sampling tubes, Rhizon Flexr soil solution samplers with pore size of 0.15 µm (Rhizosphere Research Products, the Netherlands), and inhouse fabricated platinum (Pt) electrodes with a reference Ag/AgCl electrode connected to an Oaktonr or Omega pHH-830r pH meter. Slurry was stirred for 4 weeks under anaerobic conditions by continuously purging with N2 , and then reoxidized in the fifth week by purging with air. The experiment was carried out in duplicate for each soil with the addition of ZnSO4 ·7H2 O solution to give an added concentration of 0.91 mg Zn L−1 in the slurry, which is approximately equivalent to fertilization at 5 kg Zn ha−1 . After 21 h of stirring, soil solution was collected in evacuated tubes using Rhizon Flexr soil solution samplers by suction. Soil solution sampling was repeated three times a week. The following parameters were measured in the soil solutions: 1) water-soluble Zn, Fe, and Mn (in samples that had been acidified immediately after sampling by adding one drop of 1 mol L−1 HCl through a syringe) by atomic absorption spectroscopy (Analyst 200 Model Perkin Elmer, Singapore); 2) water-soluble sulfide (in fresh samples without acidification) by an ion-selective electrode (ISE) (USEPA, 1996b) using a 9616BNWP Model (Orion Thermo Scientific, USA) after acid-distillation (USEPA, 1996a); 3) water-soluble sulfate (in samples without acidification) by the turbidimetric method (USEPA,
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TABLE I Selected physical and chemical propertiesa) of the soils collected from four sitesb) in the Philippines Property
BAY BKN NE IRRI Method description
pH (H2 O) Available Zn (mg kg−1 ) CEC (cmol kg−1 ) OC (g kg−1 ) Sand (%) Silt (%) Clay (%) Available P (mg kg−1 ) Exchangeable K (cmol kg−1 ) EC (mS cm−1 ) Total Fe oxides (g kg−1 ) Total Mn oxides (g kg−1 ) Amorphous Fe oxides (g kg−1 ) Amorphous Mn oxides (g kg−1 ) Total Zn (mg kg−1 ) Total S (mg kg−1 ) Total Fe (g kg−1 ) Total Mn (g kg−1 )
7.0 0.20 42.8 47 18 47 35 48 0.83 0.72 9.4 1.0 3.4 0.7 94 1110 50 1.1
5.5 1.55 13.5 20 23 29 48 12 0.27 0.08 53.4 1.6 5.5 1.1 155 139 94 1.6
5.2 1.06 28.0 15 25 28 47 9.3 0.29 0.21 22.6 0.7 10.9 0.5 81 231 59 0.9
6.6 1.88 24.6 12 22 48 30 22 1.68 0.19 32.8 1.8 13.1 1.1 108 220 75 1.8
1:1 soil:water ratio (Thomas, 1996) DTPAc) extraction (Lindsay and Norvell, 1978) Ammonium acetate (pH 7) method (Sumner and Miller,1996) Potassium dichromate oxidation method (Walkley and Black, 1934) Hydrometry (Gee and Bauder, 1979)
Ammonium fluoride (pH 2.6) method (Bray and Kurtz, 1945) Ammonium acetate (pH 7) method (Helmke and Sparks, 1996) 1:1 soil:water ratio (Hesse, 1971) EDTAd) -dithionite extraction (Asami and Kumada, 1959) EDTA-dithionite extraction (Asami and Kumada, 1959) Ammonium oxalate extraction in darkness (Leoppert and Inskeep, 1996) Selective dissolution with hydroxylamine HCl (Gambrell, 1996) X-ray fluorescence spectrometry (X-Lab 2000, Spectro, Germany)
a) CEC
= cation exchange capacity; OC = organic carbon; EC = electrical conductivity. = an Aeric Endoaquept from Central Mindanao University, Bukidnon; BAY = a Vertic Tropaquept from Bay, Laguna; NE = an Entic Pellustert from PhilRice, Nueva Ecija; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. c) DTPA = diethylenetriaminepentaacetic acid. d) EDTA = ethylenediaminetetraacetic acid. b) BKN
Fig. 1
Reaction cell apparatus used to conduct the stirred-slurry experiment.
1978) using a DU730 UV-Vis spectrophotometer (Beckman Coulter, USA); and 4) pH using an Orion 4-star Plus bench top pH meter (Thermo Scienti-
fic, USA). For each parameter, the limit of detection (LOD) was calculated as three times the standard deviation of blank measurements and quality control sta-
INFLUENCE OF SOIL PROPERTIES ON ZN SOLUBILITY
ndard solutions included in each analysis. For the BAY and IRRI soils, additional soil slurry samples were taken for determining diethylenetriaminepentaacetic acid (DTPA)-extractable Zn at only one time point for each phase beginning with an initial slurry sample 20 min after the addition of ZnSO4 ·7H2 O solution. The supernatant was removed after centrifugation of the slurry samples and the solid fraction was extracted with DTPA-triethanolamine (TEA) in a N2 filled glove bag to estimate the initial available Zn in the slurry (Johnson-Beebout et al., 2009). Temperature was measured several times a day using a digital thermometer. The Eh of the soil slurry was measured continuously using duplicate Pt electrodes joined together by a Cu wire to obtain an averaged readout. An Ag/Ag Clelectrode was used as a reference. The Eh corrected with reference to the standard hydrogen electrode (SHE) were carried out using the following equation (PCRA, 2007): Eh = Ehactual
readout
+ 200 − 0.65/(T − 25)
(1)
where Ehactual readout is the actual readout of Eh and T is the temperature (◦ C). The reaction duration was divided into the following four phases for plotting: initial phase (21 h for all soils); partially reduced phase (69 to 189 h for BAY and 69 to 237 h for BKN, NE, and IRRI); most reduced phase (237 to 573 h for BAY, 358 to 597 h for BKN and NE, and 358 to 693 h for IRRI); and reoxidation phase (597 to 765 h for BAY, 693 to 741 h for BKN and NE, and 741 to 765 h for IRRI). The numbers of time points used for analyses were shown in Table II. TABLE II Numbers of time points used for analyses during the four different reaction phases in the soils collected from four sitesa) in the Philippines Phase
Initial phase Partially reduced phase Most reduced phase Reoxidation phase
Number of time point BAY
BKN
NE
IRRI
2 6 14 8
2 8 14 6
2 8 16 4
2 10 14 4
a) BKN
= an Aeric Endoaquept from Central Mindanao University, Bukidnon; BAY = a Vertic Tropaquept from Bay, Laguna; NE = an Entic Pellustert from PhilRice, Nueva Ecija; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna.
Speciation modeling Speciation modeling (Visual MINTEQ ver. 3.0) was conducted to provide indications of what species
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might have formed to control Zn solubility (Gustafsson, 2010). Each sampling point was modeled as if it was at equilibrium, although this is probably not the case, because kinetic models are not sufficiently equipped to account for the complex environment (de Livera et al., 2011). Therefore, the equilibrium model results are only indicative of the compounds that might be formed under specific conditions. The measured Eh was not used as a model input because electrodemeasured Eh values are known to differ sufficiently from theoretical Eh values so as to make them unhelpful for quantitative thermodynamic calculations (McBride, 1994). Zinc, Fe, Mn, sulfide, sulfate, exchangeable K, available P, and DOC concentrations in soil solution were used as model inputs. For ionic strength, the model input was calculated from the measured electrical conductivity (EC) according to Ponnamperuma et al. (1966). For temperature, the model input was set to 25 ◦ C. Adsorption onto Fe and Mn (hydr)oxide surfaces was included in the model to simulate the distribution of ions between adsorbed and dissolved phases in a colloidal suspension. Potential solid phases were also selected in the model to investigate which would possibly be formed during the incubation period and thus affect the solubility of Zn. RESULTS The pH did not change during the experiment in any of the soils, with values in the range of 6.2–7.7 in BAY, 6.2–7.1 in BKN, 6.2–7.1 in NE, and 6.0–7.1 in IRRI (data not shown). Since pH changes in flooded soils are driven by carbonate equilibria (Kirk, 2004), it is not surprising that our stirred system with N2 purging did not exhibit these changes. The BAY soil had the fastest reduction rate among all soils, reaching −245 mV within 160 h (Fig. 2). The reduction rates of the BKN and NE soils were slower than that of the BAY soil, as the Eh reached the reduced phase plateau after about 250 h. The decrease in the Eh of the IRRI soil was the slowest compared with other soils, reaching the reduced phase plateau after about 300 h. For all soils, it was found in a pre-experiment that about 70% of the added Zn (0.91 mg L−1 ) immediately became unavailable in the aqueous phase within 20 min of addition (data not shown). In the BAY soil, the water-soluble Zn concentration was 0.10 mg L−1 during the initial phase, then decreased as the soil became more reduced until it reached 0.03 mg L−1 during the most reduced phase, and then it increased again up to 0.05 mg L−1 when the soil was reoxidized (Fig. 3a). In the BKN and NE soils, water-soluble Zn concentrations started at 0.14 and 0.15 mg L−1 , respectively, both de-
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Fig. 2 Redox potential (Eh) during the four different reaction phases in the soils collected from four sites in the Philippines. BKN = an Aeric Endoaquept from Central Mindanao University, Bukidnon; BAY = a Vertic Tropaquept from Bay, Laguna; NE = an Entic Pellustert from PhilRice, Nueva Ecija; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. Vertical bars indicate the standard errors of the means (n = 2).
creased slightly to 0.12 mg L−1 during the most reduced phase, and showed no change during reoxidation. The range of water-soluble Zn concentrations in the IRRI soil was similar to that of BAY, but there was an increase from 0.05 mg L−1 during the initial phase to 0.09 mg L−1 during the partially reduced phase, followed by a decrease during the most reduced conditions to 0.05 mg L−1 which did not increase during the reoxidation phase. In a similar preliminary experiment with IRRI soil, the estimated plant-available Zn using DTPA-extractable Zn showed a decrease as the Eh decreased (Fig. 4), which was also observed in this
M. A. B. BUNQUIN et al.
study for BAY and IRRI soils (Fig. 5). Water-soluble Fe in the BAY soil peaked at 1.5 mg L−1 during the partially reduced phase and slightly decreased throughout the most reduced phase and greatly reduced in the reoxidizing phase (Fig. 3b). Iron solubility in the BKN soil did not reach its peak of 6 mg L−1 until the most reduced phase, and then dropped steeply upon reoxidation. A similar trend was also found for the NE soil, except that Fe became more soluble earlier in the partially reduced phase than for BKN and its peak concentration was 5 mg L−1 during the most reduced phase. The IRRI soil showed a different trend in water-soluble Fe concentrations, with the highest concentration of 6 mg L−1 during the initial phase. It decreased in the reduced phases to about 1 mg L−1 , and then decreased further upon the reoxidation phase. The trends of water-soluble Mn observed between redox conditions and for all soils followed those of water-soluble Fe (Fig. 3). The peak concentrations were higher for Mn (4.1 mg L−1 for BAY, 8.7 mg L−1 for BKN, and 9.6 mg L−1 for IRRI) than Fe (2.2 mg L−1 for BAY, 7.0 mg L−1 for BKN, and 5.8 mg L−1 for IRRI) in all soils except in NE soil (3.4 mg L−1 Mn and 7.8 mg L−1 Fe). There was also a much larger difference in peak Mn concentrations between BKN and NE soils (5.3 mg L−1 ) than peak Fe concentrations (0.8 mg L−1 ). The BAY soil showed the highest concentrations of water-soluble sulfate-S of all the soils (55 mg L−1 ), while NE and IRRI showed similar initial concentrations (about 25 mg L−1 ), and BKN showed the lowest (6 mg L−1 ) (Fig. 6a). The BAY and IRRI soils showed
Fig. 3 Water-soluble Zn (a), Fe (b), and Mn (c) during the four different reaction phases in the soils collected from four sites in the Philippines. BKN = an Aeric Endoaquept from Central Mindanao University, Bukidnon; BAY = a Vertic Tropaquept from Bay, Laguna; NE = an Entic Pellustert from PhilRice, Nueva Ecija; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. Vertical bars indicate the standard errors of the means (n = number of time points for each phase of four sites as shown in Table II).
INFLUENCE OF SOIL PROPERTIES ON ZN SOLUBILITY
Fig. 4 Changes of diethylenetriaminepentaacetic acid (DTPA)available Zn concentration along with redox potential (Eh) in the preliminary experiment of an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. Vertical bars indicate the standard errors of the means (n = 3).
a decrease in water-soluble sulfate-S throughout the reduced phases, and stayed low during reoxidation phase. The BKN soil showed an increase in watersoluble sulfate-S during reduction to 18 mg L−1 in the most reduced phase, followed by a decrease upon reoxidation. The NE soil had the smallest fluctuation in water-soluble sulfate-S, from 20–22 mg L−1 during the first two phases to 18–20 mg L−1 during the most reduced and the reoxidation phases. Water-soluble sulfide-S concentrations in the BAY and IRRI soils increased to their maxima of 10–13 mg L−1 when the soils were not yet fully reduced, and then decreased during the most reduced phase without change during the reoxidation phase (Fig. 6b). In the
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Fig. 5 Diethylenetriaminepentaacetic acid (DTPA)-available Zn concentrations during the four different reaction phases in the BAY and IRRI soils collected from the Philippines. BAY = a Vertic Tropaquept from Bay, Laguna; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. Vertical bars indicate the standard errors of the means (n = 2).
BKN and NE soils, water-soluble sulfide-S concentrations were very low in the beginning, increased only up to 5 mg L−1 during the reduced phases, and then decreased slightly during the reoxidation phase. Speciation modeling suggested that water-soluble Zn and Zn bound to Mn and Fe (hydr)oxides occurred in all soils during the initial phase, while ZnS was only found in the BAY soil (Table III). During the reduced phases, Zn solubility in the BAY and IRRI soils seemed to be controlled by ZnS precipitation, while Zn was fairly evenly distributed between ZnS, Zn-Mn(hydr)oxide, and water-soluble species in the BKN and NE soils.
Fig. 6 Water-soluble sulfate (a) and sulfide (b) during the four different reaction phases in the soils collected from four sites in the Philippines. BKN = an Aeric Endoaquept from Central Mindanao University, Bukidnon; BAY = a Vertic Tropaquept from Bay, Laguna; NE = an Entic Pellustertfrom PhilRice, Nueva Ecija; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. Vertical bars indicate the standard errors of the means (n = number of time points for each phase of four sites as shown in Table II).
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TABLE III Speciation modeling results of Zn, Fe, and S during the four different reaction phasesa) in the soils collected from four sitesb) in the Philippines Ele- Species ment
BAY F1
BKN F2
F3
F4
F1
NE F2
F3
F4
F1
IRRI F2
F3
F4
F1
F2
F3
F4
Zn
ZnS +++c) Water-soluble Zn + Zn bound to Mn (hydr)oxide + Zn bound to Fe (hydr)oxide +
+++ +++ +++ – +++ +++ +++ – ++ ++ + + +++ ++ ++ ++ +++ ++ ++ ++ + ++ ++ +++ ++ ++ ++ + + + – + + + + +
+++ +++ – +++ +++ +++ ++ + + + + – ++ + +++ ++ ++ + + + ++ + + +
Fe
FeS2 +++ Water-soluble Fe + Fe bound to Fe (hydr)oxide –
+++ +++ +++ – ++ ++ + – – – – –
+++ +++ ++ +++ +++ +++ ++ + +++ ++ ++ ++ + – ++ + – –
S
Soluble oxidized S Soluble reduced S S bound to Fe (hydr)oxide FeS2 ZnS
++ ++ ++ ++ +
+++ ++ ++ + +
++ ++ ++ ++ +
+++ +++ +++ – + ++ + – – + – –
+++ +++ ++ ++ ++ ++ ++ ++ ++ + – ++ + – +
++ ++ ++ ++ +
++ ++ ++ ++ +
++ + ++ – –
++ ++ +
++ ++ + ++ +++ ++ ++ ++ + +
++ ++ ++ + +
++ – ++ ++ –
++ ++ ++ ++ +
++ ++ ++ ++ +
++ ++ ++ + +
a) F1,
F2, F3, and F4 = initial, partially reduced, most reduced, and reoxidation phases, respectively. = an Aeric Endoaquept from Central Mindanao University, Bukidnon; BAY = a Vertic Tropaquept from Bay, Laguna; NE = an Entic Pellustert from PhilRice, Nueva Ecija; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. c) +, ++, and +++ represent low to high extents of formation. b) BKN
DISCUSSION In all soils, at least 70% of the added Zn was no longer present in the soil solution within 20 min after Zn addition, while still in an oxidized environment. It seems likely that most of this Zn was adsorbed to Mn and Fe (hydr)oxides in the soil. At the experimental pH, Mn (hydr)oxides would be more negatively charged than ferrihydrite due to their lower point of zero charge. Therefore, it is likely that more Zn was bound to Mn (hydr)oxides than to Fe (hydr)oxides (Stumm and Morgan, 1996), as suggested by the speciation modeling for all four soils (Table III). During the initial phase, Zn solubility was different among soils, remaining higher in the NE and BKN soils than in the BAY or IRRI soil (Fig. 3). For the BAY soil, small amounts of sulfide were measured in solution even during the initial phase (Fig. 6). Soluble sulfide, polysulfide clusters complexed to Fe, Cu, and Zn (< 0.2 µm), and ZnS particles (< 0.45 µm) have all been found in oxic waters (Rozan et al., 2000; Priadi et al., 2012). This soluble sulfide could have precipitated small amounts of Zn as ZnS, even at high redox potential, as suggested by speciation modeling (Table III). As the redox potential decreased, DTPA-extractable Zn (which represents the exchangeable Zn pool) decreased in both the BAY and IRRI soils (Figs. 4 and 5), but water-soluble Zn decreased in the BAY soil but not the IRRI soil (Fig. 3a). In the IRRI soil, which had low sulfide concentrations,
Zn was likely to be bound in the initial phase to Mn (hydr)oxides and some to Fe (hydr)oxides, as suggested by the speciation modeling, whereas in the BAY soil, there was abundant sulfide, making precipitation of ZnS more likely initially and throughout the reduced phases. In the IRRI soil, soluble Zn increased in the partially reduced phase, apparently due to the reductive dissolution of Fe and Mn (hydr)oxides, as indicated by their increasing concentrations during this phase, which would have released (hydr)oxide-bound Zn into solution. During the most reduced phase, there was a decrease in soluble Fe concentration, as FeS formation became favoured (Lindsay, 1979) during the reduction of the IRRI soil, due to its high amorphous Fe (hydr)oxide content. In the reoxidation phase, soluble Zn did not increase in all soils, perhaps due to the persistence of ZnS in the oxic solution during the short reaction time (Hong et al., 2011). The speciation modeling results suggested the possibility of persistent Fe and Zn sulfides during the reoxidation phase. In the BKN and NE soils, soluble Zn decreased only slightly from the initial phase to the reduced phases, apparently because of high soluble Fe and relatively low sulfide. The speciation modeling results suggested that some ZnS formation was possible, but there was not sufficient sulfide to precipitate all of the Fe and Zn in solution. Microbial growth, which induces reduction, is dependent on soil OM as a C source. As expected, the BAY soil, which had the highest OM (organic C = 47
INFLUENCE OF SOIL PROPERTIES ON ZN SOLUBILITY
g kg−1 ), had the fastest Eh reduction (or decrease in Eh) compared to the other soils (Fig. 2), allowing other redox reactions to proceed faster. The rate of reduction for the other soils was observed to have closely followed the soil OM concentration. Organic matter is also related to the concentration of dissolved organic ligands (DOL), such as fulvic acids in the system. These ligands can form soluble metal complexes, which may increase metal mobility by preventing soluble metals from taking part in precipitation reactions such as the formation of metal sulfides (Frohne et al., 2011). In addition to the soil OM, litter decomposition has been shown to affect metal mobility in soils, primarily through its effect on total organic C and through the release of S and minerals (Du Laing et al., 2008). In our study, rice straw was added to all soils at a rate comparable to the incorporation of crop residues in a paddy soil, which could have contributed up to 0.05 mg L−1 Zn to the solution as the straw decomposed. This release of Zn from straw may have contributed to the measured Zn reported in Fig. 3, implying that Zn from the incorporated crop residues may positively influence Zn availability in soil solution, while the additional C source was likely to influence it negatively. In contrast to Zn, the S that may have been released during straw decomposition was small (2.1 mg L−1 S) relative to the sulfate-S measured in solution, indicating that the soil OM played a more important role than straw decomposition for S in this system. In this study, the speciation modeling predicted that both inorganic Zn and DOL-complexed Zn would be present in the soluble phase. During the initial conditions, soluble S was found to be primarily in the form of sulfate in all four soils (Fig. 6). It is known that sulfate can bind to Fe (hydr)oxides (Ishiguro et al., 2006), which may prevent it from later being available for reduction. Based on the initial Fe hydroxide content of the soils, it was expected that the amount of sulfate adsorbed to Fe (hydr)oxides at the start of the experiment would be least for the BAY soil and most for the IRRI soil (Table I). The sulfate concentration in solution would likely be determined by the initial soil sulfate concentration, which was much higher for the BAY soil than for the other soils (probably due to its higher total S content). As the Eh decreased, there was a shift from soluble sulfate to soluble sulfide species in the soils except for BKN, as microbial communities initiated sulfate reduction at Eh starting at −150 mV (Mansfeldt, 2004). In the case of the BKN soil, the decrease in Eh was relatively slower than the other soils, so reduction of sulfate to sulfide did not happen as fast. In the partially reduced
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phase, Fe oxide dissolution occurred first, which could have released some sorbed sulfate. It may have taken time for the Fe oxides to dissolve due to the high total Fe oxide content in this soil before sulfate reduction could take place. Sulfide peaked in the BAY and IRRI soils at the start of the partially reduced phase and the concentrations were higher than those that would have been expected based on the Eh, perhaps indicating the release of occluded sulfide minerals. Once soluble sulfide is formed, it is possible that it could precipitate with Zn, and thus remove both sulfide and Zn from solution (de Livera et al., 2011; Frohne et al., 2011). This ZnS precipitation seems to have happened for the BAY soil, where there were fluctuations in both soluble Zn and sulfide concentrations throughout the incubation period. Iron may also form sulfide minerals but at a lower Eh compared to that of Zn and is more soluble (Frohne et al., 2011). However, large amounts of soluble Fe under reduced conditions may compete with Zn for the available sulfide. In the IRRI soil, Fe competed with Zn for precipitation with sulfide directly, so at the start of the reduction there was an increase in the soluble Zn, but this did not last long, because in the most reduced phase, as sulfide became more available, it was enough to precipitate both Zn and Fe. However, for the BKN and NE soils, Zn sulfide precipitation apparently did not limit Zn solubility as there was more soluble Fe and less sulfide available than in the BAY and IRRI soils and complexation with DOC might have occurred to keep Zn in solution. The soluble Fe content of soils is related both to the initial amorphous Fe (hydr)oxide content and the level of reduction in the soil. Soluble Fe in the BAY, BKN, and NE soils followed the Eh pattern, while the IRRI soil showed a high Fe content initially due to its high amorphous Fe content, and also upon reduction. These contrasting results between soils indicated a need for further research to test more specific hypotheses about the conditions under which FeS competes with ZnS for formation and precipitation. Therefore, ZnS precipitation seems to be the primary cause of decreased Zn solubility under reducing conditions in flooded (non-calcareous, noncontaminated) soils in agreement with Kittrick (1976), but it does not necessarily control overall Zn solubility due to interactions with other factors. In some soils, there were indications suggesting that Fe may prevent ZnS formation, as suggested by Lindsay (1979). The main soil characteristic that resulted in the predicted formation of ZnS in our experiments was the presence of soluble sulfide. The amount of soluble sulfide was influenced by Eh, total soil S content, the interaction
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of soluble sulfate with Fe oxide surfaces, and the precipitation of soluble sulfide with Fe. Redox potential was primarily influenced by the amount of soil OM. CONCLUSIONS Results of this study provided indications identifying that the most important cause of decreasing Zn availability in flooded (non-calcareous, noncontaminated) soils was likely to be the precipitation of ZnS especially in high-S soils like the BAY soil, while sorption of Zn onto Fe and Mn (hydr)oxides may be equally important for sulfate-limited soil under oxidized conditions. In some soils, this ZnS precipitation was sufficient to cause Zn deficiency in rice plants, whereas in other soils, there was still enough Zn left in the soil solution to maintain Zn sufficiency, usually through complexation with dissolved OM. The most important reactions controlling ZnS formation were likely to be the OM-mediated decrease in Eh, the redox-mediated presence of soluble sulfide, Fe2+ competition with Zn2+ for sulfide precipitation, and the presence of Fe oxide to sorb sulfate and prevent it from becoming sulfide. These results emphasized the importance of further research to test a wider range of soils, including multiple combinations of Zn, Fe, and S forms, to determine if Fe or S can be managed in flooded soils to prevent Zn deficiency caused by ZnS precipitation or if these results can be used to more precisely predict which fields need Zn fertilization. It may be possible to find a set of soil parameters that can be measured in routine dry-soil tests, which would be helpful in predicting which soils are likely to demonstrate flooding-induced Zn deficiency. The parameters that were most likely to be of use included OM, amorphous Fe/Mn oxide content, and some form of sulfur, along with pH. It is not clear which sulfur species (total S, organic S, sulfate-S, sulfide-S, or a percentage of total S in one form) would be the most important in predicting whether or not Zn sulfide formation will limit Zn solubility. It may be the ratio of a particular form of sulfur to a species of Fe or Mn that provides the best predictive ability. This hypothesis needs to be further tested on a wide range of soils to enable multiple regression analysis to determine which parameters are the most useful. ACKNOWLEDGEMENTS This research was supported by the Global Rice Science Partnership (GRiSP) Staff Development Fund and the Swiss Agency for Development and Cooperation (SDC) awarded to Dr. S. M. Impa, International
Rice Research Institute, Philippines, through its Research Fellow Partnership Programme. We would like to thank Mr. Jerone Onoya, Mr. Lean Mercado, and Mrs. Ellen Genil from International Rice Research Institute, Philippines for technical and logistical assistance, as well as International Rice Research Institute Analytical Service Laboratory for the soil characterization data. We gratefully acknowledge the input from internal reviewers, Drs. S. M. Impa, M. C. R. Alberto, and Bill Hardy from International Rice Research Institute, Philippines, which helped to improve the manuscript. REFERENCES Asami T, Kumada K. 1959. A new method for determining free iron in paddy soils. Soil Sci Plant Nutr. 5: 141–146. Bray R H, Kurtz L T. 1945. Determination of total, organic, and available forms of phosphorus in soils. Soil Sci. 59: 39–46. Brown K H, Wuehler S E, Peerson J M. 2001. The importance of zinc in human nutrition and estimation of the global prevalence of zinc deficiency. Food Nutr Bull. 22: 113–125. Carating R B, Galanta R G, Bacatio C D. 2014. The soils of the Philippines. In Hartemink A E (ed.) World Soils Book Series. Springer Science+Business Media, Dordrecht, the Netherlands. pp. 51–106. Cassman K G, Dobermann A, Sta. Cruz P C, Gines G C, Samson M I, Descalsota J P, Alcantara J M, Dizon M A, Olk D C. 1996. Soil organic matter and the indigenous nitrogen supply of intensive irrigated rice systems in the tropics. Plant Soil. 182: 267–278. de Livera J, McLaughlin M J, Hettiarachchi G M, Kirby J K, Beak D G. 2011. Cadmium solubility in paddy soils: Effects of soil oxidation, metal sulfides and competitive ions. Sci Total Environ. 409: 1489–1497. Dobermann A, Dawe D, Roetter R P, Cassman K G. 2000. Reversal of rice yield decline in a long-term continuous cropping experiment. Agron J. 92: 633–643. Du Laing G, Vanthuyne D R J, Vandecasteele B, Tack F M G, Verloo M G. 2007. Influence of hydrological regime on pore water metal concentrations in a contaminated sedimentderived soil. Environ Pollut. 147: 615–625. Du Laing G, Bontinck A, Samson R, Vandecasteele B, Vanthuyne D R J, Meers E, Lesage E, Tack F M G, Verloo M G. 2008. Effect of decomposing litter on the mobility and availability of metals in the soil of a recently created floodplain. Geoderma. 147: 34–36. Du Laing G, Rinklebe J, Vandecasteele B, Meers E, Tack F M G. 2009. Trace metal behaviour in estuarine and riverine floodplain soils and sediments: A review. Sci Total Environ. 407: 3972–398 5. Frohne Z, Rinklebe J, Diaz-Bone R A, Du Laing G. 2011. Controlled variation of redox conditions in a floodplain soil: Impact on metal mobilization and biomethylation of arsenic and antimony. Geoderma. 160: 414–424. Gambrell R P. 1996. Manganese. In Sparks D L et al. (eds.) Methods of Soil Analysis. Part 3. Chemical Methods. Soil Science Society of America and American Society of Agronomy, Madison. pp. 665–682. Gee G W, Bauder J W. 1979. Particle size analysis by hydrometer: a simplified method for routine textural analysis and a
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(in French). Available online at www.cefracor.org/html/publications.htm (verified on April 8, 2015). Ponnamperuma F N, Tianco E M, Loy T A. 1966. Ionic strength of the solutions of flooded soils and other natural aqueous solutions from specific conductance. Soil Sci. 102: 408–413. Priadi C, Le Pape P, Morin G, Ayrault S, Maillot F, Juillot F, Hochreutener R, Llohrens I, Testemale D, Proux O, Brown G E. 2012. XAFS evidence for amorphous zinc sulfide as a major zinc species in suspended matter from the Seine River downstream of Paris, Ile-de-France, France. Environ Sci Technol. 46: 3712–3720. Rose M T, Rose T J, Pariasca-Tanaka J, Widodo, Wissuwa M. 2011. Revisiting the role of organic acids in the bicarbonate tolerance of zinc-efficient rice genotypes. Funct Plant Biol. 38: 493–504. Rozan T F, Lassman M E, Ridge D P, Luter G W. 2000. Evidence for iron, copper and zinc complexation as multinuclear sulphide clusters in oxic rivers. Nature. 406: 879–882. Stumm W, Morgan J J. 1996. Aquatic Chemistry. Chemical Equilibrium and Rates in Natural Waters. 3rd Edn. WileyInter science, New York. Sumner M E, Miller W P. 1996. Cation exchange capacity, and exchange coefficients. In Sparks D L et al. (eds.) Methods of Soil Analysis. Part 3. Chemical Methods. Soil Science Society of America and American Society of Agronomy, Madison. pp. 1201–1229. Thomas G W. 1996. Soil pH and soil acidity. In Sparks D L et al. (eds.) Methods of Soil Analysis. Part 3. Chemical Methods. Soil Science Society of America and American Society of Agronomy, Madison. pp. 475–490. US Environmental Protection Agency (USEPA). 1978. Sulfate, Method 375.4 (Turbidimetric). Methods for chemical analysis of water and wastes. Available online at www.caslab.com/EPA-Methods/PDF/EPA-Method-3754.pdf (verified on April 8, 2015). US Environmental Protection Agency (USEPA). 1996a. Method 9030B. Acid soluble and acid insoluble sulfides: Distillation. Available online at www.epa.gov/epawaste/hazard/testmethods/sw846/pdfs/9030b.pdf (verified on April 8, 2015). US Environmental Protection Agency (USEPA). 1996b. Method 9215. Potentiometric determination of sulfide in aqueous samples and distillates with ion-selective electrode. Available online at: www.epa.gov/epawaste/hazard/testmethods/ sw846/pdfs/9215.pdf (verified on April 8, 2015). Walkley A, Black I A. 1934. An examination of the Degtjareff method for determining soil organic matter, and a proposed modification of the chromic acid titration method. Soil Sci. 37: 29–38. Weber F A, Voegelin A, Kretzschmar R. 2009. Multi-metal contaminant dynamics in temporarily flooded soil under sulfate limitation. Geochim Cosmochim Acta. 73: 5513–5527. Yu K, B¨ ohme F, Rinklebe J, Neue H-U, De Laune R D. 2007. Major biogeochemical processes in soils—A microcosm incubation from reducing to oxidizing conditions. Soil Sci Soc Am J. 71: 1406–1417.
Influence of Soil Properties on Zinc Solubility Dynamics Under Different Redox Conditions in Non-Calcareous Soils Michelle Anne Belen BUNQUIN1,∗ , Susan TANDY2 , Sarah Johnson BEEBOUT1 and Rainer SCHULIN2 1 Crop 2 Soil
and Environmental Sciences Division, International Rice Research Institute, College, Los Ba˜ nos, Laguna 4031 (Philippines) Protection, Institute of Terrestrial Ecosystems, ETH Zurich, Z¨ urich CH-8092 (Switzerland)
(Received April 22, 2016; revised November 14, 2016)
ABSTRACT Zinc (Zn) deficiency in paddy soils is often a problem for rice production. Flooding can decrease metal availability in some noncalcareous soils through different mechanisms associated with soil redox status. Laboratory experiments were performed in order to better understand the processes that governed the dynamics of Zn in non-calcareous paddy soils at varying redox potentials (Eh). Airdried non-calcareous soil samples collected from four different paddy field sites in the Philippines were submerged and incubated in a reaction cell with continuous stirring and nitrogen purging for 4 weeks, and then purged with compressed air for another week to reoxidize the system. The Eh of the four soils started at 120 to 300 mV, decreased to −220 to −300 mV after 100 to 250 h of reduction, and was maintained at this low plateau for about 2 weeks before increasing again upon reoxidation. Zinc solubility showed contrasting patterns in the four soils, with two of the soils showing a decrease in soluble Zn as the Eh became low, probably due to zinc sulfide (ZnS) precipitation. In contrast, the other two soils showed that Zn solubility was maintained during the reduced phase which could be due to the competition with iron (Fe) for precipitation with sulfide. Differences in the relative amounts of S, Fe, and manganese (Mn) oxides in the four soils apparently influenced the pattern of Zn solubility after flooding. Key Words:
microcosm, paddy soil, precipitation, rice, soil sulfide, speciation, Zn availability
Citation: Bunquin M A B, Tandy S, Beebout S J, Schulin R. 2017. Influence of soil properties on zinc solubility dynamics under different redox conditions in non-calcareous soils. Pedosphere. 27(1): 96–105.
INTRODUCTION Zinc (Zn) deficiency limits grain yield and nutritional content in rice and is the most common in perennially wet or calcareous soil. It is an essential nutrient for humans for the development of the immune system and gastrointestinal tract, so that an inadequate supply of Zn causes stunted physical growth, susceptibility to infections, and neurobehavioral abnormalities (Brown et al., 2001). Zinc measurements on dry soil samples are ineffective in predicting Zn fertilizer responses in flooded soils (Impa and Johnson-Beebout, 2012), so it would be useful to identify other routinely measurable soil parameters that may increase our ability to predict flooding-induced Zn deficiency in rice fields. Flooding changes many aspects of soil chemistry which are known to affect Zn availability. The primary redox-induced changes are related to the following factors: available sulfur (S) as sulfate and sulfide, amorphous iron (Fe) and manganese (Mn) (hydr)oxides, organic matter (OM), pH, and soil solution bicarbonate ∗ Corresponding
author. E-mail: [email protected].
concentrations (Neue and Mamaril, 1985; Du Laing et al., 2009). Sulfides can decrease the mobility of trace metals such as Zn, Fe, cadmium (Cd) and copper (Cu) in soil by formation of insoluble sulfide precipitates upon flooding (Du Laing et al., 2007; de Livera et al., 2011). Iron and Mn (hydr)oxides provide adsorption sites that can immobilize trace metals, which can then be released under anaerobic conditions by reductive dissolution of the oxides, and adsorbed again during the reoxidation phase to less crystalline (hydr)oxides that are formed by reprecipitation (Du Laing et al., 2009). Organic matter provides a source of carbon (C) for microorganisms which leads to a faster development of reducing conditions upon flooding. It also protects some metal ions (e.g., Zn, Cd, nickel (Ni), and Cu) from adsorption to the solid phase through complexation with soluble humic and fulvic acids, thus increasing metal mobility (Du Laing et al., 2009). Soil pH affects metal solubility by influencing the formation of metal complexes with dissolved organic C (DOC) and the adsorption of Zn, Fe, and Mn to Fe and Mn (hydr)oxides (Kirk, 2004). The formation of Zn carbonates
INFLUENCE OF SOIL PROPERTIES ON ZN SOLUBILITY
may reduce Zn solubility at either high (in a calcareous soil) or low redox potential (Eh) (in any soil) due to the increase in partial pressure of CO2 as OM decomposes (Kirk, 2004). Also, high bicarbonate concentrations in soil solution hinder root growth and cause root leakage of Zn-inefficient rice genotypes (Rose et al., 2011). There has been disagreement in the literature about which of these processes controls Zn solubility in flooded soils. Kittrick (1976) explained that high hydrogen sulfide (H2 S) could depress Zn2+ concentrations to very low levels in flooded soils by forming zinc sulfide (ZnS), while Lindsay (1979) theorized that iron pyrite (FeS2 ) formation should consume S when the Fe redox reactions control the equilibrium Eh at a level not favorable for ZnS formation. Subsequent studies in Zn-contaminated soils have supported both sides, confirming both that Fe can interfere with ZnS formation in sulfate-limited soils (Weber et al., 2009), and ZnS can be formed in microsites (Hesterberg et al., 2011). However, the relative importance of ZnS formation in Zn-limited soils is not well understood. To isolate geochemical reactions from some of the many variables which can cause imprecise data in field experiments, reaction cells have been used to study redox chemistry questions regarding metal availability in soil (Yu et al., 2007). Continuously stirred reaction cells allow precise monitoring of chemical changes in homogenized soil slurries as the Eh varies. However, stirring with continuous nitrogen (N2 ) purging does not allow the normal build-up of dissolved CO2 in solution, and therefore does not allow meaningful measurement of carbonate equilibria (de Livera et al., 2011). Our overall aim was to understand which types of soils are likely to show flooding-induced Zn deficiency in rice. We hypothesized that although Zn does not change oxidation state due to soil Eh, Zn solubility would be influenced by other redox-active soil parameters, including changes in speciation of S, Fe, and Mn. Our objectives were to understand which chemical reactions are the most important in explaining the decrease in Zn availability in some soils after flooding and to identify a set of soil parameters that may be useful for predicting which rice fields are at risk of floodinginduced Zn deficiency. MATERIALS AND METHODS Soil sampling Samples of non-calcareous soils were collected from four different paddy field sites in the Philippines: an Aeric Endoaquept from Central Mindanao University, Bukidnon (BKN) (7◦ N, 125◦ E) (Carating et al.,
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2014); a Vertic Tropaquept from Bay, Laguna (BAY) (14◦ N, 121◦ E) (Moormann and van Breemen, 1978); an Entic Pellustert from PhilRice, Nueva Ecija (NE) (15◦ N, 120◦ E) (Cassman et al., 1996); and an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna (IRRI) (14◦ N, 121◦ E) (Dobermann et al., 2000). Rice of the first three sites shows moderate Zn-deficiency symptoms in most years, while the fourth site is considered as Znsufficient. The top 20 cm of soil was collected from each site, followed by air drying and homogenization during the grinding process to pass through a 2-mm sieve. Initial soil characterizations are shown in Table I. Experimental setup A laboratory experiment was set up using a modified reaction cell system similar to Yu et al. (2007), with two replicates for each soil. The setup (Fig. 1) consisted of a 2-L reaction beaker containing 400 g soil, 1 600 mL deionized water, and 4 g ground rice straw as a source of OM (C, 391.0 g kg−1 ; N, 8.9 g kg−1 ; S, 1.05 g kg−1 ; Zn, 24 mg kg−1 ; and C/N ratio, 43.9:1). Different components were placed in the system to accommodate the measurements, including gas inlets and outlets, soil slurry sampling tubes, Rhizon Flexr soil solution samplers with pore size of 0.15 µm (Rhizosphere Research Products, the Netherlands), and inhouse fabricated platinum (Pt) electrodes with a reference Ag/AgCl electrode connected to an Oaktonr or Omega pHH-830r pH meter. Slurry was stirred for 4 weeks under anaerobic conditions by continuously purging with N2 , and then reoxidized in the fifth week by purging with air. The experiment was carried out in duplicate for each soil with the addition of ZnSO4 ·7H2 O solution to give an added concentration of 0.91 mg Zn L−1 in the slurry, which is approximately equivalent to fertilization at 5 kg Zn ha−1 . After 21 h of stirring, soil solution was collected in evacuated tubes using Rhizon Flexr soil solution samplers by suction. Soil solution sampling was repeated three times a week. The following parameters were measured in the soil solutions: 1) water-soluble Zn, Fe, and Mn (in samples that had been acidified immediately after sampling by adding one drop of 1 mol L−1 HCl through a syringe) by atomic absorption spectroscopy (Analyst 200 Model Perkin Elmer, Singapore); 2) water-soluble sulfide (in fresh samples without acidification) by an ion-selective electrode (ISE) (USEPA, 1996b) using a 9616BNWP Model (Orion Thermo Scientific, USA) after acid-distillation (USEPA, 1996a); 3) water-soluble sulfate (in samples without acidification) by the turbidimetric method (USEPA,
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TABLE I Selected physical and chemical propertiesa) of the soils collected from four sitesb) in the Philippines Property
BAY BKN NE IRRI Method description
pH (H2 O) Available Zn (mg kg−1 ) CEC (cmol kg−1 ) OC (g kg−1 ) Sand (%) Silt (%) Clay (%) Available P (mg kg−1 ) Exchangeable K (cmol kg−1 ) EC (mS cm−1 ) Total Fe oxides (g kg−1 ) Total Mn oxides (g kg−1 ) Amorphous Fe oxides (g kg−1 ) Amorphous Mn oxides (g kg−1 ) Total Zn (mg kg−1 ) Total S (mg kg−1 ) Total Fe (g kg−1 ) Total Mn (g kg−1 )
7.0 0.20 42.8 47 18 47 35 48 0.83 0.72 9.4 1.0 3.4 0.7 94 1110 50 1.1
5.5 1.55 13.5 20 23 29 48 12 0.27 0.08 53.4 1.6 5.5 1.1 155 139 94 1.6
5.2 1.06 28.0 15 25 28 47 9.3 0.29 0.21 22.6 0.7 10.9 0.5 81 231 59 0.9
6.6 1.88 24.6 12 22 48 30 22 1.68 0.19 32.8 1.8 13.1 1.1 108 220 75 1.8
1:1 soil:water ratio (Thomas, 1996) DTPAc) extraction (Lindsay and Norvell, 1978) Ammonium acetate (pH 7) method (Sumner and Miller,1996) Potassium dichromate oxidation method (Walkley and Black, 1934) Hydrometry (Gee and Bauder, 1979)
Ammonium fluoride (pH 2.6) method (Bray and Kurtz, 1945) Ammonium acetate (pH 7) method (Helmke and Sparks, 1996) 1:1 soil:water ratio (Hesse, 1971) EDTAd) -dithionite extraction (Asami and Kumada, 1959) EDTA-dithionite extraction (Asami and Kumada, 1959) Ammonium oxalate extraction in darkness (Leoppert and Inskeep, 1996) Selective dissolution with hydroxylamine HCl (Gambrell, 1996) X-ray fluorescence spectrometry (X-Lab 2000, Spectro, Germany)
a) CEC
= cation exchange capacity; OC = organic carbon; EC = electrical conductivity. = an Aeric Endoaquept from Central Mindanao University, Bukidnon; BAY = a Vertic Tropaquept from Bay, Laguna; NE = an Entic Pellustert from PhilRice, Nueva Ecija; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. c) DTPA = diethylenetriaminepentaacetic acid. d) EDTA = ethylenediaminetetraacetic acid. b) BKN
Fig. 1
Reaction cell apparatus used to conduct the stirred-slurry experiment.
1978) using a DU730 UV-Vis spectrophotometer (Beckman Coulter, USA); and 4) pH using an Orion 4-star Plus bench top pH meter (Thermo Scienti-
fic, USA). For each parameter, the limit of detection (LOD) was calculated as three times the standard deviation of blank measurements and quality control sta-
INFLUENCE OF SOIL PROPERTIES ON ZN SOLUBILITY
ndard solutions included in each analysis. For the BAY and IRRI soils, additional soil slurry samples were taken for determining diethylenetriaminepentaacetic acid (DTPA)-extractable Zn at only one time point for each phase beginning with an initial slurry sample 20 min after the addition of ZnSO4 ·7H2 O solution. The supernatant was removed after centrifugation of the slurry samples and the solid fraction was extracted with DTPA-triethanolamine (TEA) in a N2 filled glove bag to estimate the initial available Zn in the slurry (Johnson-Beebout et al., 2009). Temperature was measured several times a day using a digital thermometer. The Eh of the soil slurry was measured continuously using duplicate Pt electrodes joined together by a Cu wire to obtain an averaged readout. An Ag/Ag Clelectrode was used as a reference. The Eh corrected with reference to the standard hydrogen electrode (SHE) were carried out using the following equation (PCRA, 2007): Eh = Ehactual
readout
+ 200 − 0.65/(T − 25)
(1)
where Ehactual readout is the actual readout of Eh and T is the temperature (◦ C). The reaction duration was divided into the following four phases for plotting: initial phase (21 h for all soils); partially reduced phase (69 to 189 h for BAY and 69 to 237 h for BKN, NE, and IRRI); most reduced phase (237 to 573 h for BAY, 358 to 597 h for BKN and NE, and 358 to 693 h for IRRI); and reoxidation phase (597 to 765 h for BAY, 693 to 741 h for BKN and NE, and 741 to 765 h for IRRI). The numbers of time points used for analyses were shown in Table II. TABLE II Numbers of time points used for analyses during the four different reaction phases in the soils collected from four sitesa) in the Philippines Phase
Initial phase Partially reduced phase Most reduced phase Reoxidation phase
Number of time point BAY
BKN
NE
IRRI
2 6 14 8
2 8 14 6
2 8 16 4
2 10 14 4
a) BKN
= an Aeric Endoaquept from Central Mindanao University, Bukidnon; BAY = a Vertic Tropaquept from Bay, Laguna; NE = an Entic Pellustert from PhilRice, Nueva Ecija; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna.
Speciation modeling Speciation modeling (Visual MINTEQ ver. 3.0) was conducted to provide indications of what species
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might have formed to control Zn solubility (Gustafsson, 2010). Each sampling point was modeled as if it was at equilibrium, although this is probably not the case, because kinetic models are not sufficiently equipped to account for the complex environment (de Livera et al., 2011). Therefore, the equilibrium model results are only indicative of the compounds that might be formed under specific conditions. The measured Eh was not used as a model input because electrodemeasured Eh values are known to differ sufficiently from theoretical Eh values so as to make them unhelpful for quantitative thermodynamic calculations (McBride, 1994). Zinc, Fe, Mn, sulfide, sulfate, exchangeable K, available P, and DOC concentrations in soil solution were used as model inputs. For ionic strength, the model input was calculated from the measured electrical conductivity (EC) according to Ponnamperuma et al. (1966). For temperature, the model input was set to 25 ◦ C. Adsorption onto Fe and Mn (hydr)oxide surfaces was included in the model to simulate the distribution of ions between adsorbed and dissolved phases in a colloidal suspension. Potential solid phases were also selected in the model to investigate which would possibly be formed during the incubation period and thus affect the solubility of Zn. RESULTS The pH did not change during the experiment in any of the soils, with values in the range of 6.2–7.7 in BAY, 6.2–7.1 in BKN, 6.2–7.1 in NE, and 6.0–7.1 in IRRI (data not shown). Since pH changes in flooded soils are driven by carbonate equilibria (Kirk, 2004), it is not surprising that our stirred system with N2 purging did not exhibit these changes. The BAY soil had the fastest reduction rate among all soils, reaching −245 mV within 160 h (Fig. 2). The reduction rates of the BKN and NE soils were slower than that of the BAY soil, as the Eh reached the reduced phase plateau after about 250 h. The decrease in the Eh of the IRRI soil was the slowest compared with other soils, reaching the reduced phase plateau after about 300 h. For all soils, it was found in a pre-experiment that about 70% of the added Zn (0.91 mg L−1 ) immediately became unavailable in the aqueous phase within 20 min of addition (data not shown). In the BAY soil, the water-soluble Zn concentration was 0.10 mg L−1 during the initial phase, then decreased as the soil became more reduced until it reached 0.03 mg L−1 during the most reduced phase, and then it increased again up to 0.05 mg L−1 when the soil was reoxidized (Fig. 3a). In the BKN and NE soils, water-soluble Zn concentrations started at 0.14 and 0.15 mg L−1 , respectively, both de-
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Fig. 2 Redox potential (Eh) during the four different reaction phases in the soils collected from four sites in the Philippines. BKN = an Aeric Endoaquept from Central Mindanao University, Bukidnon; BAY = a Vertic Tropaquept from Bay, Laguna; NE = an Entic Pellustert from PhilRice, Nueva Ecija; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. Vertical bars indicate the standard errors of the means (n = 2).
creased slightly to 0.12 mg L−1 during the most reduced phase, and showed no change during reoxidation. The range of water-soluble Zn concentrations in the IRRI soil was similar to that of BAY, but there was an increase from 0.05 mg L−1 during the initial phase to 0.09 mg L−1 during the partially reduced phase, followed by a decrease during the most reduced conditions to 0.05 mg L−1 which did not increase during the reoxidation phase. In a similar preliminary experiment with IRRI soil, the estimated plant-available Zn using DTPA-extractable Zn showed a decrease as the Eh decreased (Fig. 4), which was also observed in this
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study for BAY and IRRI soils (Fig. 5). Water-soluble Fe in the BAY soil peaked at 1.5 mg L−1 during the partially reduced phase and slightly decreased throughout the most reduced phase and greatly reduced in the reoxidizing phase (Fig. 3b). Iron solubility in the BKN soil did not reach its peak of 6 mg L−1 until the most reduced phase, and then dropped steeply upon reoxidation. A similar trend was also found for the NE soil, except that Fe became more soluble earlier in the partially reduced phase than for BKN and its peak concentration was 5 mg L−1 during the most reduced phase. The IRRI soil showed a different trend in water-soluble Fe concentrations, with the highest concentration of 6 mg L−1 during the initial phase. It decreased in the reduced phases to about 1 mg L−1 , and then decreased further upon the reoxidation phase. The trends of water-soluble Mn observed between redox conditions and for all soils followed those of water-soluble Fe (Fig. 3). The peak concentrations were higher for Mn (4.1 mg L−1 for BAY, 8.7 mg L−1 for BKN, and 9.6 mg L−1 for IRRI) than Fe (2.2 mg L−1 for BAY, 7.0 mg L−1 for BKN, and 5.8 mg L−1 for IRRI) in all soils except in NE soil (3.4 mg L−1 Mn and 7.8 mg L−1 Fe). There was also a much larger difference in peak Mn concentrations between BKN and NE soils (5.3 mg L−1 ) than peak Fe concentrations (0.8 mg L−1 ). The BAY soil showed the highest concentrations of water-soluble sulfate-S of all the soils (55 mg L−1 ), while NE and IRRI showed similar initial concentrations (about 25 mg L−1 ), and BKN showed the lowest (6 mg L−1 ) (Fig. 6a). The BAY and IRRI soils showed
Fig. 3 Water-soluble Zn (a), Fe (b), and Mn (c) during the four different reaction phases in the soils collected from four sites in the Philippines. BKN = an Aeric Endoaquept from Central Mindanao University, Bukidnon; BAY = a Vertic Tropaquept from Bay, Laguna; NE = an Entic Pellustert from PhilRice, Nueva Ecija; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. Vertical bars indicate the standard errors of the means (n = number of time points for each phase of four sites as shown in Table II).
INFLUENCE OF SOIL PROPERTIES ON ZN SOLUBILITY
Fig. 4 Changes of diethylenetriaminepentaacetic acid (DTPA)available Zn concentration along with redox potential (Eh) in the preliminary experiment of an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. Vertical bars indicate the standard errors of the means (n = 3).
a decrease in water-soluble sulfate-S throughout the reduced phases, and stayed low during reoxidation phase. The BKN soil showed an increase in watersoluble sulfate-S during reduction to 18 mg L−1 in the most reduced phase, followed by a decrease upon reoxidation. The NE soil had the smallest fluctuation in water-soluble sulfate-S, from 20–22 mg L−1 during the first two phases to 18–20 mg L−1 during the most reduced and the reoxidation phases. Water-soluble sulfide-S concentrations in the BAY and IRRI soils increased to their maxima of 10–13 mg L−1 when the soils were not yet fully reduced, and then decreased during the most reduced phase without change during the reoxidation phase (Fig. 6b). In the
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Fig. 5 Diethylenetriaminepentaacetic acid (DTPA)-available Zn concentrations during the four different reaction phases in the BAY and IRRI soils collected from the Philippines. BAY = a Vertic Tropaquept from Bay, Laguna; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. Vertical bars indicate the standard errors of the means (n = 2).
BKN and NE soils, water-soluble sulfide-S concentrations were very low in the beginning, increased only up to 5 mg L−1 during the reduced phases, and then decreased slightly during the reoxidation phase. Speciation modeling suggested that water-soluble Zn and Zn bound to Mn and Fe (hydr)oxides occurred in all soils during the initial phase, while ZnS was only found in the BAY soil (Table III). During the reduced phases, Zn solubility in the BAY and IRRI soils seemed to be controlled by ZnS precipitation, while Zn was fairly evenly distributed between ZnS, Zn-Mn(hydr)oxide, and water-soluble species in the BKN and NE soils.
Fig. 6 Water-soluble sulfate (a) and sulfide (b) during the four different reaction phases in the soils collected from four sites in the Philippines. BKN = an Aeric Endoaquept from Central Mindanao University, Bukidnon; BAY = a Vertic Tropaquept from Bay, Laguna; NE = an Entic Pellustertfrom PhilRice, Nueva Ecija; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. Vertical bars indicate the standard errors of the means (n = number of time points for each phase of four sites as shown in Table II).
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TABLE III Speciation modeling results of Zn, Fe, and S during the four different reaction phasesa) in the soils collected from four sitesb) in the Philippines Ele- Species ment
BAY F1
BKN F2
F3
F4
F1
NE F2
F3
F4
F1
IRRI F2
F3
F4
F1
F2
F3
F4
Zn
ZnS +++c) Water-soluble Zn + Zn bound to Mn (hydr)oxide + Zn bound to Fe (hydr)oxide +
+++ +++ +++ – +++ +++ +++ – ++ ++ + + +++ ++ ++ ++ +++ ++ ++ ++ + ++ ++ +++ ++ ++ ++ + + + – + + + + +
+++ +++ – +++ +++ +++ ++ + + + + – ++ + +++ ++ ++ + + + ++ + + +
Fe
FeS2 +++ Water-soluble Fe + Fe bound to Fe (hydr)oxide –
+++ +++ +++ – ++ ++ + – – – – –
+++ +++ ++ +++ +++ +++ ++ + +++ ++ ++ ++ + – ++ + – –
S
Soluble oxidized S Soluble reduced S S bound to Fe (hydr)oxide FeS2 ZnS
++ ++ ++ ++ +
+++ ++ ++ + +
++ ++ ++ ++ +
+++ +++ +++ – + ++ + – – + – –
+++ +++ ++ ++ ++ ++ ++ ++ ++ + – ++ + – +
++ ++ ++ ++ +
++ ++ ++ ++ +
++ + ++ – –
++ ++ +
++ ++ + ++ +++ ++ ++ ++ + +
++ ++ ++ + +
++ – ++ ++ –
++ ++ ++ ++ +
++ ++ ++ ++ +
++ ++ ++ + +
a) F1,
F2, F3, and F4 = initial, partially reduced, most reduced, and reoxidation phases, respectively. = an Aeric Endoaquept from Central Mindanao University, Bukidnon; BAY = a Vertic Tropaquept from Bay, Laguna; NE = an Entic Pellustert from PhilRice, Nueva Ecija; IRRI = an Aquandic Epiaquoll from the experimental farm of International Rice Research Institute, Laguna. c) +, ++, and +++ represent low to high extents of formation. b) BKN
DISCUSSION In all soils, at least 70% of the added Zn was no longer present in the soil solution within 20 min after Zn addition, while still in an oxidized environment. It seems likely that most of this Zn was adsorbed to Mn and Fe (hydr)oxides in the soil. At the experimental pH, Mn (hydr)oxides would be more negatively charged than ferrihydrite due to their lower point of zero charge. Therefore, it is likely that more Zn was bound to Mn (hydr)oxides than to Fe (hydr)oxides (Stumm and Morgan, 1996), as suggested by the speciation modeling for all four soils (Table III). During the initial phase, Zn solubility was different among soils, remaining higher in the NE and BKN soils than in the BAY or IRRI soil (Fig. 3). For the BAY soil, small amounts of sulfide were measured in solution even during the initial phase (Fig. 6). Soluble sulfide, polysulfide clusters complexed to Fe, Cu, and Zn (< 0.2 µm), and ZnS particles (< 0.45 µm) have all been found in oxic waters (Rozan et al., 2000; Priadi et al., 2012). This soluble sulfide could have precipitated small amounts of Zn as ZnS, even at high redox potential, as suggested by speciation modeling (Table III). As the redox potential decreased, DTPA-extractable Zn (which represents the exchangeable Zn pool) decreased in both the BAY and IRRI soils (Figs. 4 and 5), but water-soluble Zn decreased in the BAY soil but not the IRRI soil (Fig. 3a). In the IRRI soil, which had low sulfide concentrations,
Zn was likely to be bound in the initial phase to Mn (hydr)oxides and some to Fe (hydr)oxides, as suggested by the speciation modeling, whereas in the BAY soil, there was abundant sulfide, making precipitation of ZnS more likely initially and throughout the reduced phases. In the IRRI soil, soluble Zn increased in the partially reduced phase, apparently due to the reductive dissolution of Fe and Mn (hydr)oxides, as indicated by their increasing concentrations during this phase, which would have released (hydr)oxide-bound Zn into solution. During the most reduced phase, there was a decrease in soluble Fe concentration, as FeS formation became favoured (Lindsay, 1979) during the reduction of the IRRI soil, due to its high amorphous Fe (hydr)oxide content. In the reoxidation phase, soluble Zn did not increase in all soils, perhaps due to the persistence of ZnS in the oxic solution during the short reaction time (Hong et al., 2011). The speciation modeling results suggested the possibility of persistent Fe and Zn sulfides during the reoxidation phase. In the BKN and NE soils, soluble Zn decreased only slightly from the initial phase to the reduced phases, apparently because of high soluble Fe and relatively low sulfide. The speciation modeling results suggested that some ZnS formation was possible, but there was not sufficient sulfide to precipitate all of the Fe and Zn in solution. Microbial growth, which induces reduction, is dependent on soil OM as a C source. As expected, the BAY soil, which had the highest OM (organic C = 47
INFLUENCE OF SOIL PROPERTIES ON ZN SOLUBILITY
g kg−1 ), had the fastest Eh reduction (or decrease in Eh) compared to the other soils (Fig. 2), allowing other redox reactions to proceed faster. The rate of reduction for the other soils was observed to have closely followed the soil OM concentration. Organic matter is also related to the concentration of dissolved organic ligands (DOL), such as fulvic acids in the system. These ligands can form soluble metal complexes, which may increase metal mobility by preventing soluble metals from taking part in precipitation reactions such as the formation of metal sulfides (Frohne et al., 2011). In addition to the soil OM, litter decomposition has been shown to affect metal mobility in soils, primarily through its effect on total organic C and through the release of S and minerals (Du Laing et al., 2008). In our study, rice straw was added to all soils at a rate comparable to the incorporation of crop residues in a paddy soil, which could have contributed up to 0.05 mg L−1 Zn to the solution as the straw decomposed. This release of Zn from straw may have contributed to the measured Zn reported in Fig. 3, implying that Zn from the incorporated crop residues may positively influence Zn availability in soil solution, while the additional C source was likely to influence it negatively. In contrast to Zn, the S that may have been released during straw decomposition was small (2.1 mg L−1 S) relative to the sulfate-S measured in solution, indicating that the soil OM played a more important role than straw decomposition for S in this system. In this study, the speciation modeling predicted that both inorganic Zn and DOL-complexed Zn would be present in the soluble phase. During the initial conditions, soluble S was found to be primarily in the form of sulfate in all four soils (Fig. 6). It is known that sulfate can bind to Fe (hydr)oxides (Ishiguro et al., 2006), which may prevent it from later being available for reduction. Based on the initial Fe hydroxide content of the soils, it was expected that the amount of sulfate adsorbed to Fe (hydr)oxides at the start of the experiment would be least for the BAY soil and most for the IRRI soil (Table I). The sulfate concentration in solution would likely be determined by the initial soil sulfate concentration, which was much higher for the BAY soil than for the other soils (probably due to its higher total S content). As the Eh decreased, there was a shift from soluble sulfate to soluble sulfide species in the soils except for BKN, as microbial communities initiated sulfate reduction at Eh starting at −150 mV (Mansfeldt, 2004). In the case of the BKN soil, the decrease in Eh was relatively slower than the other soils, so reduction of sulfate to sulfide did not happen as fast. In the partially reduced
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phase, Fe oxide dissolution occurred first, which could have released some sorbed sulfate. It may have taken time for the Fe oxides to dissolve due to the high total Fe oxide content in this soil before sulfate reduction could take place. Sulfide peaked in the BAY and IRRI soils at the start of the partially reduced phase and the concentrations were higher than those that would have been expected based on the Eh, perhaps indicating the release of occluded sulfide minerals. Once soluble sulfide is formed, it is possible that it could precipitate with Zn, and thus remove both sulfide and Zn from solution (de Livera et al., 2011; Frohne et al., 2011). This ZnS precipitation seems to have happened for the BAY soil, where there were fluctuations in both soluble Zn and sulfide concentrations throughout the incubation period. Iron may also form sulfide minerals but at a lower Eh compared to that of Zn and is more soluble (Frohne et al., 2011). However, large amounts of soluble Fe under reduced conditions may compete with Zn for the available sulfide. In the IRRI soil, Fe competed with Zn for precipitation with sulfide directly, so at the start of the reduction there was an increase in the soluble Zn, but this did not last long, because in the most reduced phase, as sulfide became more available, it was enough to precipitate both Zn and Fe. However, for the BKN and NE soils, Zn sulfide precipitation apparently did not limit Zn solubility as there was more soluble Fe and less sulfide available than in the BAY and IRRI soils and complexation with DOC might have occurred to keep Zn in solution. The soluble Fe content of soils is related both to the initial amorphous Fe (hydr)oxide content and the level of reduction in the soil. Soluble Fe in the BAY, BKN, and NE soils followed the Eh pattern, while the IRRI soil showed a high Fe content initially due to its high amorphous Fe content, and also upon reduction. These contrasting results between soils indicated a need for further research to test more specific hypotheses about the conditions under which FeS competes with ZnS for formation and precipitation. Therefore, ZnS precipitation seems to be the primary cause of decreased Zn solubility under reducing conditions in flooded (non-calcareous, noncontaminated) soils in agreement with Kittrick (1976), but it does not necessarily control overall Zn solubility due to interactions with other factors. In some soils, there were indications suggesting that Fe may prevent ZnS formation, as suggested by Lindsay (1979). The main soil characteristic that resulted in the predicted formation of ZnS in our experiments was the presence of soluble sulfide. The amount of soluble sulfide was influenced by Eh, total soil S content, the interaction
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of soluble sulfate with Fe oxide surfaces, and the precipitation of soluble sulfide with Fe. Redox potential was primarily influenced by the amount of soil OM. CONCLUSIONS Results of this study provided indications identifying that the most important cause of decreasing Zn availability in flooded (non-calcareous, noncontaminated) soils was likely to be the precipitation of ZnS especially in high-S soils like the BAY soil, while sorption of Zn onto Fe and Mn (hydr)oxides may be equally important for sulfate-limited soil under oxidized conditions. In some soils, this ZnS precipitation was sufficient to cause Zn deficiency in rice plants, whereas in other soils, there was still enough Zn left in the soil solution to maintain Zn sufficiency, usually through complexation with dissolved OM. The most important reactions controlling ZnS formation were likely to be the OM-mediated decrease in Eh, the redox-mediated presence of soluble sulfide, Fe2+ competition with Zn2+ for sulfide precipitation, and the presence of Fe oxide to sorb sulfate and prevent it from becoming sulfide. These results emphasized the importance of further research to test a wider range of soils, including multiple combinations of Zn, Fe, and S forms, to determine if Fe or S can be managed in flooded soils to prevent Zn deficiency caused by ZnS precipitation or if these results can be used to more precisely predict which fields need Zn fertilization. It may be possible to find a set of soil parameters that can be measured in routine dry-soil tests, which would be helpful in predicting which soils are likely to demonstrate flooding-induced Zn deficiency. The parameters that were most likely to be of use included OM, amorphous Fe/Mn oxide content, and some form of sulfur, along with pH. It is not clear which sulfur species (total S, organic S, sulfate-S, sulfide-S, or a percentage of total S in one form) would be the most important in predicting whether or not Zn sulfide formation will limit Zn solubility. It may be the ratio of a particular form of sulfur to a species of Fe or Mn that provides the best predictive ability. This hypothesis needs to be further tested on a wide range of soils to enable multiple regression analysis to determine which parameters are the most useful. ACKNOWLEDGEMENTS This research was supported by the Global Rice Science Partnership (GRiSP) Staff Development Fund and the Swiss Agency for Development and Cooperation (SDC) awarded to Dr. S. M. Impa, International
Rice Research Institute, Philippines, through its Research Fellow Partnership Programme. We would like to thank Mr. Jerone Onoya, Mr. Lean Mercado, and Mrs. Ellen Genil from International Rice Research Institute, Philippines for technical and logistical assistance, as well as International Rice Research Institute Analytical Service Laboratory for the soil characterization data. We gratefully acknowledge the input from internal reviewers, Drs. S. M. Impa, M. C. R. Alberto, and Bill Hardy from International Rice Research Institute, Philippines, which helped to improve the manuscript. REFERENCES Asami T, Kumada K. 1959. A new method for determining free iron in paddy soils. Soil Sci Plant Nutr. 5: 141–146. Bray R H, Kurtz L T. 1945. Determination of total, organic, and available forms of phosphorus in soils. Soil Sci. 59: 39–46. Brown K H, Wuehler S E, Peerson J M. 2001. The importance of zinc in human nutrition and estimation of the global prevalence of zinc deficiency. Food Nutr Bull. 22: 113–125. Carating R B, Galanta R G, Bacatio C D. 2014. The soils of the Philippines. In Hartemink A E (ed.) World Soils Book Series. Springer Science+Business Media, Dordrecht, the Netherlands. pp. 51–106. Cassman K G, Dobermann A, Sta. Cruz P C, Gines G C, Samson M I, Descalsota J P, Alcantara J M, Dizon M A, Olk D C. 1996. Soil organic matter and the indigenous nitrogen supply of intensive irrigated rice systems in the tropics. Plant Soil. 182: 267–278. de Livera J, McLaughlin M J, Hettiarachchi G M, Kirby J K, Beak D G. 2011. Cadmium solubility in paddy soils: Effects of soil oxidation, metal sulfides and competitive ions. Sci Total Environ. 409: 1489–1497. Dobermann A, Dawe D, Roetter R P, Cassman K G. 2000. Reversal of rice yield decline in a long-term continuous cropping experiment. Agron J. 92: 633–643. Du Laing G, Vanthuyne D R J, Vandecasteele B, Tack F M G, Verloo M G. 2007. Influence of hydrological regime on pore water metal concentrations in a contaminated sedimentderived soil. Environ Pollut. 147: 615–625. Du Laing G, Bontinck A, Samson R, Vandecasteele B, Vanthuyne D R J, Meers E, Lesage E, Tack F M G, Verloo M G. 2008. Effect of decomposing litter on the mobility and availability of metals in the soil of a recently created floodplain. Geoderma. 147: 34–36. Du Laing G, Rinklebe J, Vandecasteele B, Meers E, Tack F M G. 2009. Trace metal behaviour in estuarine and riverine floodplain soils and sediments: A review. Sci Total Environ. 407: 3972–398 5. Frohne Z, Rinklebe J, Diaz-Bone R A, Du Laing G. 2011. Controlled variation of redox conditions in a floodplain soil: Impact on metal mobilization and biomethylation of arsenic and antimony. Geoderma. 160: 414–424. Gambrell R P. 1996. Manganese. In Sparks D L et al. (eds.) Methods of Soil Analysis. Part 3. Chemical Methods. Soil Science Society of America and American Society of Agronomy, Madison. pp. 665–682. Gee G W, Bauder J W. 1979. Particle size analysis by hydrometer: a simplified method for routine textural analysis and a
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