Biochar and ash derived from silicon-rich rice husk decrease inorganic arsenic species in rice grain

Biochar and ash derived from silicon-rich rice husk decrease inorganic arsenic species in rice grain

Science of the Total Environment 684 (2019) 360–370 Contents lists available at ScienceDirect Science of the Total Environment journal homepage: www...

3MB Sizes 0 Downloads 2 Views

Science of the Total Environment 684 (2019) 360–370

Contents lists available at ScienceDirect

Science of the Total Environment journal homepage:

Biochar and ash derived from silicon-rich rice husk decrease inorganic arsenic species in rice grain Parapond Leksungnoen a, Worachart Wisawapipat a,⁎, Daojarus Ketrot a, Surachet Aramrak a, Sumontha Nookabkaew b, Nuchanart Rangkadilok b,c, Jutamaad Satayavivad b,c,d a

Department of Soil Science, Faculty of Agriculture, Kasetsart University, Bangkok 10900, Thailand Laboratory of Pharmacology, Chulabhorn Research Institute (CRI), Kamphaeng Phet 6, Laksi, Bangkok 10210, Thailand Center of Excellence on Environmental Health and Toxicology (EHT), Bangkok 10400, Thailand d Environmental Toxicology Program, Chulabhorn Graduate Institute (CGI), Kamphaeng Phet 6, Laksi, Bangkok 10210, Thailand b c




• Rice husk biochar and ash significantly decreased As(III) accumulation in grain. • Rice husk ash significantly decreased As (V) uptake in grain. • Biochar and ash promoted residual As phase formation and demoted rice grain As. • Oxalate-extractable As of flooded soil could explain inorganic grain As contents. • Mn oxides were the major sources and sinks for As released to soil solution.

a r t i c l e

i n f o

Article history: Received 21 February 2019 Received in revised form 10 May 2019 Accepted 17 May 2019 Available online 18 May 2019 Editor: Xinbin Feng Keywords: Sequential extraction Porewater Paddy soil Agricultural wastes Iron sulfides Biogeochemistry

a b s t r a c t Exposure to arsenic (As) through rice consumption potentially threatens millions of people worldwide. Understanding is still lacking the recycling impacts of rice residues on As phytoavailability in paddy soils and is of indisputable importance in providing a sustainable and effective measure to decrease As accumulation in rice grain. Herein, we examined the effects of rice husk biochar (RHB) and rice husk ash (RHA) on As grain speciation, and As dynamics in the soil porewater and solid-phase fractions. The results corroborated that both the RHB and RHA (0.64% w/w) treatments significantly (p b 0.05) decreased inorganic As accumulation in rice grain to 0.27–0.29 mg kg−1, which was below the maximum inorganic As level in husked rice (0.35 mg kg−1) established by the Codex. The residual phase (F6 = 90% of total soil As) as quantified by the sequential extraction was the dominant As pool; the fractions were subsequently transformed into several As pools associated with soluble and exchangeable (F1), organically bound (F2), Mn oxides (F3), poorly crystalline (F4) and crystalline (F5) Fe oxides during the rice growing periods. The Si-rich amendments enhanced the residual phase formation upon soil flooding, which decreased the As availability to rice plant. The inorganic grain-As concentrations were well explained by the soil-extractable As concentrations in the F2, F3, F5, and F6 fractions. The pore-water analysis indicated that Mn oxides were important sources and sinks for As released to the soil solution. Our findings shed light on the beneficial role of RHB and RHA in alleviating inorganic As uptake in paddy rice. © 2019 Elsevier B.V. All rights reserved.

⁎ Corresponding author. E-mail address: [email protected] (W. Wisawapipat). 0048-9697/© 2019 Elsevier B.V. All rights reserved.

P. Leksungnoen et al. / Science of the Total Environment 684 (2019) 360–370

1. Introduction Elevated arsenic (As) levels found in rice grain are of global concern because As is a potentially carcinogenic metalloid that poses an imperative health threat to millions of people worldwide through its intake from the consumption of As-rich rice (Ma et al., 2016; Williams et al., 2005). Rice is highly capable of assimilating As from soils resulting in As-laden rice grain (Abedin et al., 2002; Ma et al., 2008; Su et al., 2010). Moreover, paddy rice cultivation substantially enhances the As uptake into rice grain during soil flooding under anaerobic rice cultiva2− tion, which promotes the reduction of As(V) (H2AsO− 4 and HAsO4 ) to the more toxic As(III) (H3AsO03), and thus increases the mobility, availability, and bioaccessibility of As in terrestrial environments (Marin et al., 1993). Furthermore, the reductive dissolution of Mn(III/IV)- and Fe(III)-(oxyhydr)oxides, referred to as Mn and Fe oxides hereafter, under anoxic conditions considerably intensifies the release of As adsorbed to or substituted into the structure of these minerals (Weber et al., 2009a). Conversely, aerobic rice production such as intermittently flooded rice paddies has been successfully introduced into rice production to mitigate As uptake in rice (Arao et al., 2009; Li et al., 2009; Xu et al., 2008) by mechanisms of the oxidation of highly mobile As(III) to less mobile As(V) along with the adsorption of As(V) onto Mn(III/ IV)- and Fe(III)-oxides (Ehlert et al., 2016; Parsons et al., 2013). In addition to water management approaches, uses of silicon (Si) fertilizer—as Si is a beneficial element for rice—have been efficiently utilized to diminish As assimilation in rice (Li et al., 2009), because As(III) uses the similar Si-uptake transporters (Lsi1 and Lsi2) in rice (Ma et al., 2008). Soil incorporations of Si compounds such as silica gel and diatomaceous earth as well as Si-rich agricultural residues such as fresh rice husk (FH), rice husk ash (RHA) and rice husk biochar (RHB) can alleviate As accumulation in rice (Liu et al., 2014; Seyfferth and Fendorf, 2012). Recently, it has been reported that Si-rich materials can improve Si bioavailability and enhance the proportion of ferrihydrite on root plaque and decreased the As uptake (Limmer et al., 2018). The use of biochars derived from diverse biomass including RHB has been demonstrated to be an effective soil amendment to enhance soil carbon sequestration (Lal, 2004; Lehmann et al., 2006) and fertility (Van Zwieten et al., 2010; Zhang et al., 2012) and to decrease trace metal(loid) bioavailability (Beesley et al., 2011; Namgay et al., 2010; Park et al., 2011) and greenhouse gas emissions (Khan et al., 2013) in soils. However, recent studies have shown that biochar addition into paddy soils increased the dissolved As in soils through enhancing the abundance of Fe-reducing bacteria, and promoting the reductive dissolution rate of Fe oxides (Wang et al., 2017). To date, information is poorly understood on the effects of RHB and RHA on the transformation of solid-phase As partitioning in soils during rice growth and its role in As accumulation in rice. Therefore, this study examined the growth and As grain speciation in rice (Oryza sativa L.) as affected by RHB and RHA addition, and investigated the effects of Si-rich residues on As changes in porewater and the solid-phase fractions of As-rich paddy soil. The relationships of grain-As concentrations with As extractability in soils using sequential extraction were also investigated to unravel the soil chemical factors controlling As uptake by rice grain. 2. Materials and methods 2.1. Soil collection and characterization For the pot experiment, large quantities of surface soil samples (0–25 cm depth) were used from an As-enriched paddy soil collected in a rice paddy field in Saraburi province, Thailand. This soil was classified as Calcic Endoaquert (Soil Survey Staff, 2014). The sample was airdried, gently pulverized and sieved to a particle size b2000 μm using a stainless steel sieve. The general physicochemical properties of the studied soil—texture, pH by 1:1 H2O extraction (pHH2O), organic carbon content, and cation exchange capacity (CEC)—were analyzed using


standard procedures (Gee and Bauder, 1986; Sparks et al., 1996). Total As and Si concentrations in pressed powder samples were measured using X-ray fluorescence spectrometry (XRF; S8Tiger; Bruker; Billerica, MA, USA). The accuracy of the As and Si concentration in the soil was measured against a standard reference specimen (STSD-3; CCRMP, CANMET Mining, and Mineral Sciences Laboratories; Hamilton, Ontario, Canada). The As and Si concentrations in STSD-3 were 25.7 ± 1.9 mg kg−1 and 21.2 ± 0.047%, respectively, (mean ± standard deviation of triplicates), which were in agreement with the certified values of 28 mg kg−1 and 23%, respectively. The studied paddy soil was calcareous (pHH2O = 7.7) with a clayey texture (Table 1). This soil had medium amounts of organic carbon (1.2%), high nitrogen (0.15%), a meager amount of Bray-II extractable P (4.7 mg kg−1), and a very high CEC value (84 cmolc kg−1). The total As concentration in the soil material was 83 mg kg−1, which exceeded the value of the background soil As concentrations of 10 mg kg−1 (Kabata-Pendias, 2011) and the average concentration of uncontaminated Thai paddy soils of 1.7 mg kg−1 (Prakongkep et al., 2008). The source of As contamination in the studied soil was considered to be from natural origins as the area has been used for paddy rice production with no As-related industry in the surrounding area. 2.2. Biochar preparation and characterization Fresh rice husk material was obtained from a rice mill processing non-glutinous rice. For rice husk biochar (RHB) production, the husk material was placed into a 200 L traditional kiln and charred for 4 h (slow pyrolysis) at an operating temperature of approximately 270–350 °C. Each biochar sample was ground to a particle size of b2000 μm before chemical analysis and the pot experiment. Smallholder farmers can afford to own a traditional kiln and to produce biochar on site; however, it should be noted that O2 may partially diffuse into the kiln. Furthermore, biochar produced using a low operating temperature (b550 °C) potentially has a high sorption capacity (Manyà, 2012) that is suitable for lowering the solubility and mobility of potentially toxic trace elements such as As. Rice husk ash (RHA) was prepared by the combustion of the fresh husk under the ambient atmosphere. Chemical characterization of the RHB and RHA was measured using standard procedures (Sparks et al., 1996). For acid washed biochar (AWB) preparation, ground RHB was washed with 0.1 M HCl at a solid-to-solution ratio of 1:200, then agitated using a horizontal mechanical shaker for 1 h. The acid washing procedure was repeated 12 times and targets the removal of silicate minerals and other minerals from the biochar. The entrained acid solution (H+ and Cl− ions) was washed out with deionized (DI) water. An AgNO3 solution was used to check the presence of Cl− ions in the solutions. The AWB was then oven dried at 60 °C for 24 h and kept for the pot experiment. The pH and electrical conductivity (EC) of the RHB and RHA were determined using a 1:5 (w/v) solid-to-solution ratio. The ash content was analyzed by combusting the RHB in a muffle furnace at 750 °C for 4 h. The total As and Mn concentrations in the RHB and RHA were analyzed using aqua regia digestion (3:1 (v/v) of HCl and HNO3) (Chen and Ma, 2001). The total Si concentrations in the RHB, AWB, and RHA were determined using microwave-NaOH/H2O2 digestion (Seyfferth and Fendorf, 2012). The As and Si concentrations in digests were determined using atomic absorption spectroscopy (AAS, Agilent Technologies, Tokyo, Japan) after filtering through a filter paper (Whatman Grade 42, Dassel, Germany). The hydride generation technique (HGAAS) was used for As determination. The RHB material was neutral (pHH2O = 7.3) with a low soluble salt content (EC = 1.1 dS m−1) and a relatively high cation exchangeable capacity (CECNH4OAc = 26 cmolc kg−1). The RHB contained a large amount of total Si (180 g kg−1), which was consistent with a high mineral ash content of 39%. The AWB material had a slightly lower concentration of total Si (154 g kg−1) compared to the RHB because diluted HCl is

pH by 1:1 H2O extraction for soil sample and 1:5 H2O extraction for RHB, RHA, and AWB sample. EC using 1:5 H2O extraction. Total As, Cu, Fe, Mn, and Zn concentrations in the soil sample were determined using X-ray fluorescence spectrometry, whereas those in the RHB, RHA, and AWB were obtained from aqua regia digestion (Chen and Ma, 2001). The total Si concentrations in RHB, RHA, and AWB were obtained from microwave-NaOH/H2O2 digestion (Seyfferth and Fendorf, 2012).

52 – – – 38 – – – 84 ± 3.4 26 ± 3.0 29 ± 0.63 – 4.7 ± 0.07 – – – 0.15 ± 0.03 – – – 1.6 ± 0.03 2.2 ± 0.09 1.8 ± 0.05 1.8 ± 0.22 0.71 ± 0.01 1.11 ± 0.06 6.6 ± 0.03 0.33 ± 0.01 7.7 ± 0.01 7.3 ± 0.01 9.3 ± 0.01 3.2 ± 0.01 Soil RHB RHA AWB


83 0.33 ± 0.01 2.3 ± 0.02 0.71 ± 0.22 10 – – –

Silt Sand CEC (cmolc kg−1) Avail. P (mg kg−1) Total N (%) OC (%) EC (dS m−1) pH (H2O) Materials


218 180 ± 6.3 224 ± 127 154 ± 11 113 31 ± 0.71 130 ± 2.3 16 ± 3.3 (mg kg−1) (%)




117 3.7 ± 0.49 14 ± 0.63 6.2 ± 2.0


79734 197 ± 43 3484 ± 181 116 ± 13


3377 2774 ± 74 9544 ± 213 199 ± 7.3

Si (g kg−1) Zn

P. Leksungnoen et al. / Science of the Total Environment 684 (2019) 360–370

Table 1 Average values (mean ± SD) of physicochemical propertiesa and total element concentrationsb of the paddy soil, rice husk biochar (RHB), rice husk ash (RHA), acid washed biochar (AWB) used in this study.


less effective in extracting Si, resulting in large amounts of Si remaining in the AWB sample even after the 12 successive acid-washing steps. The RHA material was highly alkaline (pHH2O = 9.3) with a relatively high soluble salt content (EC = 6.6 dS m−1). The RHA sample had the highest concentration of total Si (224 g kg−1) compared to the other materials. In addition, the samples of RHB, RHA, and AWB contained very low As concentrations of 0.33, 2.3, and 0.71 mg kg−1, respectively (Table 1). 2.3. Pot experiment Pot experiments were conducted in July 2015 in a greenhouse at ambient temperature with natural sunlight. The experiment was performed with three replications to elucidate the impact of RHB and RHA on As speciation in rice (Oryza sativa L. var. RD31) grain. A sample of 8 kg of airdried soil was placed into each polyethylene container. The RHB and RHA materials were homogeneously incorporated into the soil at rates of 0.16, 0.32, and 0.64% w/w, which were referred to as RHB0.16, RHA0.16, RHB0.32, RHA0.32, RHB0.64, and RHA0.64, respectively. These rates corresponded to the field application rates of 3.12, 6.25 and 12.50 t ha−1, respectively, based on a soil depth of 15 cm and a bulk density of 1.3 g cm−3. A control treatment (C) without soil amendments and an AWB treatment at the rate of 0.64% w/w (AWB0.64) were included. The basal fertilizers were applied and mixed in all treatments at the NPK rates of 178, 53, and 106 mg nutrient kg−1 soil, respectively. The containers were flooded to ~5 cm above the soil surface using DI water. Rhizon samplers (Rhizon MOM; 10 cm length, 2.5 mm outside diameter; Rhizosphere Research Products, Wageningen, the Netherlands) were deployed into the soil at a downward 45° angle for collecting soil porewater from rhizospheres. Five rice seedlings were transplanted into each pot after 2 days of flooding. 2.4. Plant growth, harvesting, and chemical analysis Rice was grown under continuous flooding conditions and irrigated with DI water throughout the entire rice growth period. The number of tillers (~77 days depending on treatment), heading date, and plant height (at 120 days) were measured. After grain maturity, the straw was cut 4 cm above the soil surface. Roots were separated from the soil, washed with tap water, and rinsed with DI water. The weights of straw and roots were measured after drying in an oven at 65 °C for 48 h. The numbers of filled grains and unfilled grains and the weight of filled grains were determined. Unpolished rice grains were separated from the filled grains for the analysis of the total concentration and speciation of As. The amounts of total As in samples of straw, roots, and husk were separately digested using an HNO3/H2O2 procedure. The concentration of As in root Fe plaque was also measured following the procedure of Hu et al. (2007). Briefly, 1 g of fresh root sample was extracted with 30 mL of 1 M HCl for 30 min, and the dissolved As concentration in the extracted sample was measured using the HGAAS technique. 2.5. Total and speciation of arsenic in rice grain For analysis of total As in rice grain, 0.2 g of finely ground samples were weighed into Teflon digestion vessels and mixed with 6 mL of trace-metal-grade concentrated HNO3. After allowing 30 min of settling time at room temperature, the vessels were closed and digested in a MAR5 microwave digestion system (CEM Corporation, Mathews, NC, USA). The temperature was ramped for 15 min to reach 190 °C, and subsequently, digestion was allowed for an additional 30 min at 190 °C. The clear aliquot was diluted to 50 mL, and the As concentration in the digests was measured using inductively coupled plasma mass spectrometry (8800 Triple Quadrupole ICP-MS, Agilent Technologies, Tokyo, Japan). The duplicate analysis of NIST-certified reference material (CRM) 1568b rice flour was performed to ascertain the analysis, which yielded an excellent recovery for As of 97 ± 0.84%.

P. Leksungnoen et al. / Science of the Total Environment 684 (2019) 360–370

For analysis of As speciation in rice grain, 0.25 g of finely ground samples of unpolished rice grain was weighed into polypropylene centrifuge tubes and mixed with 10 mL of 0.15 M HNO3. The closed screwcapped tubes were transferred into a shaking water bath and continuously agitated at 95 °C for 3 h. The suspension was centrifuged at 2205 ×g for 5 min and the supernatant (1 mL) was filtered through a 0.2 μm polyvinylidene fluoride membrane filter, and then analyzed for grain As species using high-performance liquid chromatography (HPLC)-ICP-MS. Separation of As species was achieved on an X-Select Charged Surface Hybrid (CSH) C18 column (4.6 mm × 150 mm, and 5 μm internal diameter; Water Corporation, Milford, MA, USA) with an optimized mobile phase (1.0 mL min−1) of 7.5 mM (C4H9)4NOH and 10 mM (NH4)H2PO4 at pH 8.25 (95%) and methanol (5%). To validate inorganic As species (Asi = AsIII + AsV) and organic As (dimethylarsinic acid: DMA and monomethylarsonic acid: MMA) species, repeated extractions and analyses (n = 5) of NIST CRM (1568b) were performed. The analysis had recoveries for Asi, DMA, and MMA of 101 ± 5, 89 ± 2, and 124 ± 3%, respectively.


suspensions were centrifuged at 2205 ×g for 15 min and the As concentrations in the extracts were measured using HGAAS. In this study, the SE results were referred to as the (F1) soluble and exchangeable As, (F2) organically bound As, (F3) As bound to Mn oxides, (F4) As bound to poorly crystalline Fe oxides, (F5) As bound to crystalline Fe oxides, and (F6) residual phases. 2.8. Data analysis The effects of RHB and RHA on the elemental concentrations in plant parts and growth parameters (filled grain weight, % filled grain, heading date, tiller number, height, straw biomass, root biomass, and root length) were analyzed using ANOVA and Duncan's multiple range test (DMRT) with statistical software. The normality and homogeneity of data were tested using the Shapiro-Wilks and Levene's tests, respectively before ANOVA analysis. Most data except for As(V) and DMA had a normal distribution and equal variance. The Games-Howell test was used instead of the DMRT for data that deviated from normality and homogeneity. Significance was tested at the p b 0.5 level.

2.6. Porewater collection and analysis 3. Results and discussion To investigate the dynamics of As, Si, and major dissolved ions (Mn, Fe, S, and Cu) in the soil porewater treated with RHB and RHA, porewater was collected at different defined periods (0, 1, 3, 5, 7, 10, 14, 21, 28, 42, 56, 70, 84, 98, and 112 days) for chemical analysis. The pH and redox potential (Eh) of the samples of porewater and wet soils were measured at the time of solution sampling. The pH and Eh of soil porewater were measured under ambient conditions, with the values being simultaneously recorded after each porewater sample had been taken to minimize changes in the redox-sensitive parameter. The Eh of the soil and porewater was measured using a combined platinum Ag/ AgCl electrode (51343200 InLab Redox ORP, Mettler Toledo, Columbus, Ohio, USA), and was calibrated against a redox-buffer solution (Eh = Emeasured + 207 mV). Acidified aliquots (% v/v of conc. HCl) of porewater samples were frozen for measurements of major cations using an AAS and for sulfate using a spectrophotometer. 2.7. Solid-phase fractions of arsenic A six-step sequential extraction (SE) procedure was conducted on wet soil pastes of selected samples (the control, RHB0.64, RHA0.64, and AWB0.64) during rice growth (1, 14, 42, 56, and 98 days) to investigate the dynamics in soil As partitioning during rice growth (Table 2). This chemical extraction technique quantitatively identified As pools of different solubilities in the studied soil samples (Huang and Kretzschmar, 2010; Voegelin et al., 2008). In brief, 1 g samples of wet soil solids were weighed into separate 50 mL centrifuge tubes and equilibrated sequentially with: (F1) 0.05 M NH4H2PO4, (F2) 0.1 M pyrophosphate at pH 7, (F3) 0.1 M NH2OH-HCl + 1 M NH4OAc at pH 6.0, (F4) 0.2 M NH4-oxalate at pH 3.25, and (F5) 0.1 M ascorbic acid + 0.2 M NH4oxalate at pH 3.25. The F6 fraction, the residual fraction, was calculated from the difference between the total As determined using XRF analysis and the sum of fractions F1–F5. After equilibration in each step, the soil

3.1. Impacts on arsenic speciation in grain and yield The total grain As concentration (sum of all As species) in the present study varied from 0.29 to 0.38 mg kg−1, depending on the treatment (Fig. 1a). The analysis of As speciation in the grain showed As(III) and DMA as the two dominant species in the rice grain. In all treatments, the results suggested that 84–93% of the total As in the grain was present as As(III), 5.9–14% was present as DMA, and 0.27–3.1% was present as As(V). On average, inorganic As was estimated to host 90% of the total grain As. Our results were consistent with several reports indicating the dominance of As(III), DMA, and As(V) in rice grain around the globe (Limmer et al., 2018; Seyfferth et al., 2016; Williams et al., 2005). The RHA0.64 and RHB0.64 treatments significantly decreased grain-As (III) concentrations (Fig. 1b). In addition, all RHA rates significantly decreased grain-As(V) concentrations (Fig. 1c). This demonstrated that these amendments were highly efficient and capable of alleviating human health concerns from the intake of As-rich rice. The RHA treatments were more effective at decreasing inorganic As than the RHB treatments when applied at the same rate, as they reduced both As (III) and As(V). The highest RHA and RHB rates (RHA0.64 and RHB0.64) significantly lowered the inorganic As concentrations in rice grain to 0.27–0.29 mg kg−1 (20–24% reduction), which is below the maximum guideline for inorganic As in husked rice (0.35 mg kg−1) as proposed by the Codex Committee on Contaminants in Foods (CCCF, 2018). The decrease in the As(III) accumulation in rice from RHB and RHA application could be primarily attributed to improving the Si phytoavailability in the soil solution (see Section 3.2) and the Si uptake in the rice plant (Figs. 2, A.1), which can suppress the As(III) uptake by rice roots through the competitive absorption of Si with As(III) in the Si uptake pathway (Ma et al., 2008). This coincided well with several recent research studies demonstrating that applications of Si-rice residues can

Table 2 Summary of six-step sequential extraction of As and hypothetical interpretation of the six fractions and solid to solution ratio (SSR). As fraction a

F1 F2a F3b F4b F5b F6 a

Extracting solution

Target pool

Extraction conditions

SSR (g mL−1)

0.05 M (NH4)H2PO4 0.1 M Na2H2P2O7 (pH 7) 0.1 M NH2OH-HCl + 1 M NH4OAc (pH 6) 0.2 M NH4-oxalate (pH 3.25) 0.1 M ascorbic acid + 0.2 M NH4-oxalate (pH 3.25) Total As determined by XRF minus the sum of As in fractions F1–F5

Soluble and exchangeable Organically bound Bound to Mn oxides Bound to weakly crystalline Fe-oxides Bound to crystalline Fe-oxides Residual fractions

2h 1h 0.5 h 2 h in dark 2 h in water bath at 96 °C

1:25 1:20 1:25 1:25 1:25

Huang and Kretzschmar (2010). Voegelin et al. (2008). Fraction F3, soil residuum was washed with 1 M NH4OAc at pH 6 for 10 min (SSR = 1:12.5). Soil residuum from fractions F4 and F5 was washed with 0.2 M NH4oxalate at pH 3.25 for 10 min in the dark (SSR = 1:12.5). b


P. Leksungnoen et al. / Science of the Total Environment 684 (2019) 360–370

Fig. 1. Effects of rice husk biochar (RHB), acid-wash biochar (AWB) and rice husk ash (RHA) on (a) total As concentrations (sum of As species), (b) As(III) concentrations, (c) As (V) concentrations, and (d) organic Aso (DMA + MMA) concentrations in unpolished grain. Data indicate mean ± SD values (n = 3). Different lowercase letters above columns in the same subfigure are significantly different at p b 0.05. Columns with no letters indicate not significantly different (p N 0.05).

reduce As uptake in rice grain (Amaral et al., 2017; Liu et al., 2014; Seyfferth et al., 2016; Wang et al., 2016). Moreover, a recent study showed that Si-rich materials could promote the formation of a large fraction of poorly ordered ferrihydrite in Fe root plaque, which could be attributed to dissolved Si that can delay the formation of crystalline Fe minerals (Limmer et al., 2018). Therefore, considerable amounts of As particularly As(V) could be retained more in the Fe plaque and thus decrease grain As accumulation because ferrihydrite is more chemically reactive than the more crystalline Fe(III) oxides such as goethite and

lepidocrocite (Ona-Nguema et al., 2005). This was in accordance with our results showing a higher As concentration in the Fe plaque and lower concentrations of root-As and grain-As(V) in the RHA treatments relative to the control or the RHB treatments (Figs. 1 and A.2). However, elevated concentrations of dissolved Si in aqueous soil solution may lead to the polymerization of silicate on the surfaces of Fe(III) oxides, which could suppress the long-term availability of chemically reactive surfaces of Fe(III) oxides to adsorb inorganic As including As(III) and As(V) (Christl et al., 2012).

Fig. 2. Effects of rice husk biochar (RHB), acid-wash biochar (AWB), and rice husk ash (RHA) on Si concentrations in (a) unpolished grain, (b) husk, (c) straw, (d) roots, and (e) iron plaque. Data indicate mean ± SD (n = 3). Different lowercase letters on columns indicate significantly different at p ≤ 0.05. Columns with no letters indicate not significantly different (p N 0.05).

P. Leksungnoen et al. / Science of the Total Environment 684 (2019) 360–370


Fig. 3. Effects of rice husk biochar (RHB), acid-wash biochar (AWB), and rice husk ash (RHA) on (a) filled grain weight, (b) % filled grain, (c) heading date, (d) tiller number, (e) height120d, (f) straw biomass, (g) root biomass, and (h) root length. Data are mean ± SD (n = 3). Different lowercase letters above columns in the same subfigure are significantly different at p b 0.05. Columns with no letters indicate not significantly different (p N 0.05).

Neither RHA nor RHB amendments adversely affected the rice yield and the number of filled grains (Fig. 3a and b). The RHA treatments, especially at the highest rate, significantly accelerated the heading date and increased the numbers of tillers, the height, straw biomass, root biomass, and root length (Fig. 3c–h), presumably due to surplus mineral nutrients from the ash material. 3.2. Dynamics of soil pH, Eh, and major dissolved species The dynamics of soil pH, Eh, and aqueous (dissolved) Mn, Fe, S, Cu, As, and Si demonstrated similar trends in all treatments, but the magnitudes of the dissolved metal(loids) greatly varied between the treatments (Fig. 4a). The soil pH slightly decreased from ~7.5 to 6.7 during the rice growth stage of soil flooding, but gradually increased to circumneutral conditions at the beginning of soil drainage (ripening state started at 96 days), which resulted from the HCO− 3 /CO2 equilibrium. The decrease in the soil pH resulted from the production of H+ from the dissociation of HCO− 3 and H2CO3, whereas the increase in the soil pH was due to the consumption of H+ by the Fe(III) reduction (Kögel-Knabner et al., 2010). Soil Eh gradually decreased from about +400 mV to +100 mV at 112 days, indicating reducing conditions during rice growth. Dissolved Mn clearly increased after 7 days, whereas Fe started to dissolve after 28 days, reflecting the onset of microbial reduction of Mn(III/IV)- and Fe(III)-oxides. A simultaneous increase of dissolved As with increasing dissolved Mn and Fe illustrated the release of adsorbed

As to the soil solution through the mechanisms of the reductive dissolution of Mn(III/IV)- and Fe(III)-oxides. The mobilization of As was observed after 30 days when the Eh dropped to about +350 mV (Fig. 4b). The relationship between soil solution Eh and aqueous As was adequately described by an exponential function (Fig. 5a, R2 = 0.60, p b 0.001). Soil porewater As concentrations increased rapidly when the pH was lower than 7.5 (Fig. 5b), mainly due to the competitive adsorption of bicarbonate with As(III) (Stachowicz et al., 2007) and the decrease in the soil pH during soil flooding. The positive linear correlations of dissolved As with Mn (R2 = 0.73, Fig. 5c, p b 0.001) and Fe (R2 = 0.55, Fig. 5d, p b 0.001) elaborated the crucial contribution of Mn(III/IV)- and Fe(III)-oxides to the soil porewater As. The higher relationship of dissolved As with dissolved Mn than with dissolved Fe delineated that Mn oxides could be the primary source of As released into the soil solution rather than the sink for As under continuous flooding rice cultivation. Furthermore, birnessite—a MnO2 mineral—has been documented to be highly capable of oxidizing As(III) to As(V) and Fe (II) to Fe(III), and thereby birnessite decreases As mobility through As (V) sorption to Fe(III) oxides as previously observed in a column experiment (Ehlert et al., 2016). This was consistent with recent research that supplied Mn oxide (hausmannite) into paddy soils and found that it enhanced As(III) oxidation, decreased dissolved As in the porewater, and decreased As concentrations into rice grain (Xu et al., 2017). However, the chemically reactive surfaces of Mn oxides for the As(III) oxidation could be passivated by dissolved Fe(II) and Mn (II) during the initial period of the flooded conditions (Ehlert et al., 2016). However, the long


P. Leksungnoen et al. / Science of the Total Environment 684 (2019) 360–370

Fig. 4. Dynamics of soil (a) pH, (b) Eh and (c, d, e, f, g, and h) dissolved elements in porewater during rice growth in a naturally As-rich paddy soil incorporated with rice husk biochar (RHB), acid-wash biochar (AWB), and rice husk ash (RHA). Values are mean ± SD values (n = 3) of selected sampling periods (0, 7, 28, 56, 70, and 112 days).

Fig. 5. Relationships between As and (a) Eh, (b) pH, (c) dissolved Mn and (d) dissolved Fe in porewater of the studied paddy soil incorporated with rice husk biochar (RHB), acid-wash biochar (AWB), and rice husk ash (RHA). Each data point represents the average values (n = 3) of individual data points collected at different sampling times (0, 1, 3, 5, 7, 10, 14, 21, 28, 42, 56, 70, 84, 98, and 112 days). *** indicates significant at p b 0.001.

P. Leksungnoen et al. / Science of the Total Environment 684 (2019) 360–370

period of soil flooding of 4 months in this study could have successively promote the release of inorganic As adsorbed to freshly precipitated Mn (III/IV) and Fe(III) oxides that are commonly observed during the microbial respiration of electron acceptors according to the typical sequence of redox reactions in paddy soils (Borch et al., 2010; Fulda et al., 2013). The elevated concentrations of As by this reductive dissolution of the Mn(III/IV) and Fe(III) oxides vitally increased the As bioavailability and thereby increased As-grain accumulation (see Section 3.4). In the high-ash treatment (RHA0.64), aqueous Mn concentrations were higher than for the other treatments (Fig. 4c), due to the contribution of Mn from the ash material (total Mn = 9544 mg kg−1). Conversely, in the high-biochar treatment (RHB0.64), the concentration of aqueous Fe was higher than in the other treatments (Fig. 4d), which could have been due to the redox properties of biochars that behaved as an electron shuttle between Fe(III)-reducing bacteria and Fe(III) minerals facilitating the reduction of Fe(II) (Wang et al., 2017). Moreover, the adsorption of Fe(II) onto biochar instead of Fe(III) oxides effectively increased the extent of Fe(III) reduction by preventing the surface passivation of Fe(III) oxides (Kappler et al., 2014), which resulted in an increase in the dissolved As concentrations in the RHB0.64 treatment. However, other studies have shown that the dissolved organic carbon that may occur in the biochars potentially displace some metal(loids) including As adsorbed to soil constituents and thereby increases the amounts of dissolved As (Williams et al., 2011; Winkel et al., 2008). In fact, less As(III) was observed in the rice grain with the high RHB0.64 treatment (Fig. 1b), but it was high in the porewater indicating that there were several mechanisms for As exclusion from the rice plant. This was attributed to the RHB providing available Si pools for rice plants as was observed by the higher Si uptake in rice grain as discussed above (Fig. 2). The RHB addition also decreased As accumulation in the Fe plaques on rice roots (Fig. A.1). Moreover, the use of RHB at the high rate enhanced the transformation of As fractions to residual fractions during rice growing periods (see below). The adsorption of As(III) by the RHB could be another As exclusion mechanism (Fig. A.3). The dissolved sulfate concentrations gradually decreased with soil reduction and were almost entirely dissolved after 112 days of flooding (Fig. 4e). This assumed the reduction of sulfate into sulfides. The dissolved Cu was consumed at 112 days. The decrease in the dissolved sulfate with Cu may have been due to the colloid-sized precipitates of copper sulfides such as covellite (CuS) and chalcocite (Cu2S), which were observed to be the dominant sulfide minerals in flooded paddy soil contaminated with trace metals (Fig. 4f) (Fulda et al., 2013; Weber et al., 2009b). The current study found slightly elevated concentrations of As in the porewater in the RHB0.64 treatment (deviating from the control treatment of ~0.17 μmol L−1) that could be of global environmental concern (Fig. 4g). However, the As porewater concentrations only slightly exceeded the World Health Organization guideline for As in drinking water (0.13 μmol L−1). Moreover, paddy rice cultivation is considered as a relatively closed system, because irrigation water is typically flooded during the rice growing stages, and then allowed to evaporate during the ripening stages. Afterward, the paddy fields are subjected to oxic conditions that cause the oxidative precipitation of Fe(III)- and Mn(IV)-oxides along with oxidizing As(III) to As(V), which decreases As mobility and solubility. The dissolved Si concentrations were high, with the values being dependent on sampling time and treatment (Fig. 4h). Elevated dissolved Si (N2000 μM) could be attributable to rice residues containing abundant poorly crystalline Si phytolith (Neethirajan et al., 2009), whose solubility is 17 times higher than well crystalline quartz (Fraysse et al., 2006). This was consistent with the high values of RHB and RHA dissolved Si in the soil solution. The low operating temperatures for the RHB and RHA production in this study (b350 °C) promoted the dominance of the poorly crystalline Si phase in both materials. The dissolved Si in the RHB0.64 treatment was higher in the control series after 30 days, which was due to dissolved Si in the soil solution slowly diffusing into


the pores of the biochar and thereby decreasing the amount of pristine Si in the soil solution during the first 30 days of rice growth. However, the soil solution absorbed into the biochar matrix could subsequently dissolve Si from silicate minerals occurring inside the pores and elevated the concentration of soil porewater Si after 30 days of soil reduction. 3.3. Dynamics of solid-phase fractions of arsenic during rice growth The solubility of As in the studied soil incorporated with RHB and RHA was observed using a six-step sequential extraction (Fig. 6). The extracted As fractions are shown as a relative percentage of the total amount of As measured by XRF in Table A1. On the first day of the pot experiment, in the control series, 90% of total soil As was identified in fraction F6, suggesting the dominance of As pools in the soils of some stable As minerals such as calcium arsenate minerals (orthoarsenate: Ca3[AsO4]2 or yukonite: Ca2Fe3[AsO4]3[OH]4·4H2O), or As sulfide minerals (arsenopyrite: FeAsS, orpiment: As2S3, or realgar As2S4) as previously observed in diverse soil types (Mandaliev et al., 2014; Meunier et al., 2010; Sadiq, 1986). The relative fractions of As associated with crystalline Fe oxides (F5) and poorly crystalline Fe oxides (F4) were the second largest As pools in the soil (F4 + F5 = 6.7% of total soil As). The fractions attributed to soluble and exchangeable As (F1 = 1.4%), organically bound As (F2 = 0.87%), and As bound to Mn oxides (F3 = 0.88%) contributed minor amounts to the total As in the soil. During the 98 days of rice cultivation, a substantial transformation of As partitioning in the soil was observed. The residual fraction (F6) in the control treatment decreased from 90% (on the first day) to 69% (at 98 days) of the total soil As due to the slow reduction kinetics of stable minerals that were subsequently transformed to other As fractions. The increasing amounts of As in the F4 and F5 fractions in all treatments may indicate As(III) sorption to newly formed mackinawite (FeS) under reducing conditions (Bostick and Fendorf, 2003; Boye et al., 2017). This result was also consistent with a recent study showing that the use of zero-valent iron in contaminated paddy soils could promote the formation of amorphous Fe oxides which provide a new reactive surface for As adsorption, resulting in decreased As solubility and bioavailability (Qiao et al., 2018). The increased As concentration in the F2 fraction suggested the contribution of natural organic matter (NOM) to stabilize As under strong reducing conditions through the covalent binding of As(III) to the sulfhydryl groups of NOM as previously observed in a naturally As-enriched peatland (Langner et al., 2012). This was in contrast to several studies which have documented that organic amendment and dissolved organic matter played a disputable role in promoting As mobility (Beesley et al., 2014; Williams et al., 2011; Winkel et al., 2008) through displacing adsorbed As from surfaces of Fe(III) oxides (Redman et al., 2002), and thus increased the rice grain As uptake (Norton et al., 2013). Soluble and exchangeable As in the control treatment increased by 5% during paddy rice cultivation, and this could be attributed to As (V) reduction to As(III) that has a higher mobility than As(V), and to the release of As through dissimilatory Fe(III) mineral dissolution during soil drainage at 98 days. The RHA and RHB treatments slightly decreased the soluble and exchangeable As in the studied soil (at 98 days) presumably due to adsorption processes of As with the ash and biochar materials. Our As (III) adsorption on the RHB demonstrated an S-curve isotherm, suggesting a low affinity of the RHB surfaces for As(III) sorption at low surface coverage (Fig. S3). It suggested that the biochars reactive sites were in competition with dissolved silicic acids (H4SiO4) for As(III) (H3AsO3). This was in accordance with the greater As(III) sorption capacity for the AWB with the lower Si concentration. However, the biochar surfaces could acquire preference as As binding sites when the As concentration exceeded the competition capacity of the H4SiO4. Recent research has shown that pristine and hematite/magnetiteengineered biochar can suppress As mobility through several mechanisms including precipitation, complexation, and electrostatic


P. Leksungnoen et al. / Science of the Total Environment 684 (2019) 360–370

Fig. 6. Relative distribution of solid-phase As fractions of the studied paddy soil incorporated with rice husk biochar (RHB), acid-wash biochar (AWB), and rice husk ash (RHA) during rice growth in a naturally As-rich paddy soil.

interactions (Li et al., 2017). In addition, there was evidence showing that oak wood-derived biochar could remove both As(III) and As(V) through the adsorption on the\\OH,\\COOH,\\CO, and\\CH3 surface functional groups of the biochar (Wang et al., 2015). Moreover, both the RHA and RHB treatments increased the F6 fraction during days 42 to 56, which was consistent with the decreased As concentration in the F5 fraction. This suggested the formation pathway of sparingly soluble As sulfides under anaerobic conditions from the newly formed crystalline phases of iron sulfide, and thereby a decrease in the As solubility under reducing conditions. This was consistent with the result of Burton et al. (2014) showing that As adsorption by mackinawite under microbial sulfate reduction could be an important mechanism to mitigate As mobility in sulfidic flooded soils. The As in the F3 fraction gradually increased from day 1 (0.69–0.88%) to day 56 (0.70–1.86%) and then decreased at day 98 during soil drainage (0.51–0.72%). This suggested that Mn(II) minerals such as rhodochrosite (MnCO3) which could be formed during soil reduction, contributed to As sorption under anoxic conditions. 3.4. Associations of solid-phase fractions of arsenic with arsenic speciation in rice grain Time-averaged As extractability in various fractions of the studied soil during rice cultivation (1, 14, 42, 56, and 98 days) was calculated and related to As grain accumulation to deliberate how solid-phase fractions of As in soils affect the accumulation of As in rice grains (Fig. A.4). The results demonstrated positive linear relationships of inorganic grain As with extractable As in the fractions F2 (R2 = 0.87), F3 (R2 = 0.93), and F5 (R2 = 0.96). The As concentration in fraction F6 was highly negatively correlated to the inorganic grain As (R2 = 0.90). However, extractable As in fraction F1, interpreted as a soluble and exchangeable form that is the most important to As bioavailability in terrestrial environments, poorly described the inorganic grains As (R2 = 0.06). These results suggested that the extractants targeting exchangeable As pools in soils may be inappropriate for explaining As phytoavailability to paddy rice plants. Selective extractions associated with organically bound (F2), Mn oxides (F3), and crystalline Fe oxides (F5) that could be dissolvable under soil reduction are more effective for describing As than the soluble and exchangeable forms such as (NH4)H2PO4 extraction or dissolved As in soil solution. In addition, because the As extractability in these fractions was subjected to change upon flooding conditions (Fig. 6). We propose that the soil samples must be equilibrated to reach reducing conditions before the soil extraction process. This allows the adsorbed- or coprecipitated-As to be released through the reductive dissolution of Fe and Mn oxides, which are accessible by

paddy rice, and precisely delineate As availability in soils to paddy rice. In other words, soil amendment incorporations or agricultural practices that potentially affect the transformation of the As pools associated with organic matter, Mn oxides, and crystalline Fe oxide fractions, and increase the residual As pools could be promising procedures to mitigate As accumulation in rice grain. 4. Conclusions Recycling rice residues enriched in Si into paddy soils could be a practical way to mitigate excessive As accumulation as part of maintaining paddy rice sustainably. This study highlighted that the highest rate (0.64% w/w) of incorporation of rice husk ash (RHA) and rice husk biochar (RHB) significantly decreased inorganic grain As by 20–24% while maintaining the required rice yield and grain quality. The effect of the RHA on decreasing As(V) uptake in rice grain was also more pronounced than that of the RHB, presumably because of the enhancement of As accumulation in the Fe plaque. The improving available Si in the soil solution that competitively inhibited As(III) uptake and transportations toward grain could also be an important mechanism in decreasing As(III) accumulation in rice. Dissolved Si from the rice residues such as rice husk and straw would typically be much more soluble than Si from silicate minerals. However, entrained silicate minerals from both biochar and ash formed by the respective pyrolysis and carbonization processes could slowly dissolve the available Si into soil porewater and suppress As uptake in the rice grain over several cycles compared to fresh rice residues. The highest application rate of RHA and RHB at 0.64% w/w used in the present study corresponded to a field application rate of 12.50 t ha−1, which is about 10-times higher than the available quantity of rice husk in each crop cycle (~1.25 t ha−1). Thus, it is necessary for local farmers in As-contaminated areas to obtain additional fresh rice husk or RHA from nearby rice mill industries or to gradually recycle obtainable Si-rich rice residues on sites, with at least 10 consecutive cycles being necessary for the sustainably produce of rice with safe As levels. Nonetheless, the biochar yield from fresh rice husk is about 35%, suggesting a challenge to obtain sufficient biochar quantity as a considerable amount of the fresh husk (ca. 37.5 t ha−1) and time are needed for the biochar production. In addition to RHB and RHA applications, soil incorporation of fresh straw after the harvest should be a more practical procedure to improve Si availability in soils and to reduce the number of cycles of Si-rich residues required to decrease As bioavailability. Further research should examine the effects of RHA and RHB on the speciation of As uptake in rice under paddy field conditions. Investigation

P. Leksungnoen et al. / Science of the Total Environment 684 (2019) 360–370

into biochar stability and its long-term impacts on the mobility, bioavailability, and speciation of As in rice cropping systems should also be conducted and compared to ash materials having less stability. Declaration of Competing Interest The authors declare no competing financial interests. Acknowledgments We gratefully acknowledge the Kasetsart University Research and Development Institute (KURDI), Bangkok, Thailand for financial support (Grant No. 18.58). Anonymous reviewers are appreciated for their constructive suggestions to improve the clarity of the manuscript. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2019.05.247. References Abedin, M.J., Cresser, M.S., Meharg, A.A., Feldmann, J., Cotter-Howells, J., 2002. Arsenic accumulation and metabolism in rice (Oryza sativa L.). Environ. Sci. Technol. 36, 962–968. Amaral, D., Lopes, G., Guilherme, L.R.G., Seyfferth, A.L., 2017. A new approach to sampling intact Fe plaque reveals Si-induced changes in Fe mineral composition and shoot As in rice. Environ. Sci. Technol. 51, 38–45. Arao, T., Kawasaki, A., Baba, K., Mori, S., Matsumoto, S., 2009. Effects of water management on cadmium and arsenic accumulation and dimethylarsinic acid concentrations in Japanese rice. Environ. Sci. Technol. 43, 9361–9367. es9022738. Beesley, L., Moreno-Jiménez, E., Gomez-Eyles, J.L., Harris, E., Robinson, B., Sizmur, T., 2011. A review of biochars' potential role in the remediation, revegetation and restoration of contaminated soils. Environ. Pollut. 159, 3269–3282. envpol.2011.07.023. Beesley, L., Inneh, O.S., Norton, G.J., Moreno-Jimenez, E., Pardo, T., Clemente, R., Dawson, J.J., 2014. Assessing the influence of compost and biochar amendments on the mobility and toxicity of metals and arsenic in a naturally contaminated mine soil. Environ. Pollut. 186, 195–202. Borch, T., Kretzschmar, R., Kappler, A., van Cappellen, P., Ginder-Vogel, M., Voegelin, A., Campbell, K., 2010. Biogeochemical redox processes and their impact on contaminant dynamics. Environ. Sci. Technol. 44, 15–23. Bostick, B.C., Fendorf, S., 2003. Arsenite sorption on troilite (FeS) and pyrite (FeS2). Geochim. Cosmochim. Acta 67, 909–921. envpol.2011.07.023. Boye, K., Lezama-Pacheco, J., Fendorf, S., 2017. Relevance of reactive Fe:S ratios for sulfur impacts on arsenic uptake by rice. Soil Syst 1, 1–13. soils1010001. Burton, E.D., Johnston, S.G., Kocar, B.D., 2014. Arsenic mobility during flooding of contaminated soil: the effect of microbial sulfate reduction. Environ. Sci. Technol. 48, 13660–13667. CCCF, 2018. Joint FAO/WHO Food Standards Programme Codex Committee on Contaminants. Foods. Working Document for Information and Use in Discussions Related to Contaminants and Toxins in the GSCTFF, 12th Session Utrecht, The Netherlands, 12–16 March 2018. Food and Agriculture Organization of the United Nations, Rome, Italy. Chen, M., Ma, L.Q., 2001. Comparison of three aqua regia digestion methods for twenty Florida soils. Soil Sci. Soc. Am. J. 65, 491–499. sssaj2001.652491x. Christl, I., Brechbühl, Y., Graf, M., Kretzschmar, R., 2012. Polymerization of silicate on hematite surfaces and its influence on arsenic sorption. Environ. Sci. Technol. 46, 13235–13243. Ehlert, K., Mikutta, C., Kretzschmar, R., 2016. Effects of manganese oxide on arsenic reduction and leaching from contaminated floodplain soil. Environ. Sci. Technol. 50, 9251–9261. Fraysse, F., Pokrovsky, O.S., Schott, J., Meunier, J.-D., 2006. Surface properties, solubility and dissolution kinetics of bamboo phytoliths. Geochim. Cosmochim. Acta 70, 1939–1951. Fulda, B., Voegelin, A., Kretzschmar, R., 2013. Redox-controlled changes in cadmium solubility and solid-phase speciation in a paddy soil as affected by reducible sulfate and copper. Environ. Sci. Technol. 47, 12775–12783. Gee, G.W., Bauder, J.W., 1986. Particle-size analysis. In: Klute, A. (Ed.), Methods of Soil Analysis. Part 1. Physical and Mineralogical Methods. American Society of Agronomy, Inc., Madison, Wisconsin USA, pp. 383–411. Hu, Z.Y., Zhu, Y.G., Li, M., Zhang, L.G., Cao, Z.H., Smith, F.A., 2007. Sulfur (S)-induced enhancement of iron plaque formation in the rhizosphere reduces arsenic accumulation in rice (Oryza sativa L.) seedlings. Environ. Pollut. envpol.2006.06.014.


Huang, J.H., Kretzschmar, R., 2010. Sequential extraction method for speciation of arsenate and arsenite in mineral soils. Anal. Chem. 82, 5534–5540. 10.1021/ac100415b. Kabata-Pendias, A., 2011. Trace Elements in Soils and Plants. 4th edition. CRC press, Boca Raton, FL. Kappler, A., Wuestner, M.L., Ruecker, A., Harter, J., Halama, M., Behrens, S., 2014. Biochar as an electron shuttle between bacteria and Fe (III) minerals. Environ. Sci. Technol. Lett. 1, 339–344. Khan, S., Chao, C., Waqas, M., Arp, H.P.H., Zhu, Y.G., 2013. Sewage sludge biochar influence upon rice (Oryza sativa L) yield, metal bioaccumulation and greenhouse gas emissions from acidic paddy soil. Environ. Sci. Technol. 47, 8624–8632. 10.1021/es400554x. Kögel-Knabner, I., Amelung, W., Cao, Z., Fiedler, S., Frenzel, P., Jahn, R., Kalbitz, K., Kölbl, A., Schloter, M., 2010. Biogeochemistry of paddy soils. Geoderma 157, 1–14. https://doi. org/10.1016/j.geoderma.2010.03.009. Lal, R., 2004. Soil carbon sequestration impacts on global climate change and food security. Science 304, 1623–1627. Langner, P., Mikutta, C., Kretzschmar, R., 2012. Arsenic sequestration by organic sulphur in peat. Nat. Geosci. 5, 66–73. Lehmann, J., Gaunt, J., Rondon, M., 2006. Bio-char sequestration in terrestrial ecosystems– a review. Mitig. Adapt. Strat. GL. 11, 395–419. Li, R.Y., Stroud, J.L., Ma, J.F., McGrath, S.P., Zhao, F.-J., 2009. Mitigation of arsenic accumulation in rice with water management and silicon fertilization. Environ. Sci. Technol. 43, 3778–3783. Li, H., Dong, X., da Silva, E.B., de Oliveira, L.M., Chen, Y., Ma, L.Q., 2017. Mechanisms of metal sorption by biochars: biochar characteristics and modifications. Chemosphere 178, 466–478. Limmer, M.A., Mann, J., Amaral, D.C., Vargas, R., Seyfferth, A.L., 2018. Silicon-rich amendments in rice paddies: effects on arsenic uptake and biogeochemistry. Sci. Total Environ. 624, 1360–1368. Liu, W.-J., McGrath, S.P., Zhao, F.-J., 2014. Silicon has opposite effects on the accumulation of inorganic and methylated arsenic species in rice. Plant Soil 376, 423–431. https:// Ma, J.F., Yamaji, N., Mitani, N., Xu, X.-Y., Su, Y.-H., McGrath, S.P., Zhao, F.-J., 2008. Transporters of arsenite in rice and their role in arsenic accumulation in rice grain. Proc. Natl. Acad. Sci. 105, 9931–9935. Ma, L., Wang, L., Jia, Y., Yang, Z., 2016. Arsenic speciation in locally grown rice grains from Hunan Province, China: spatial distribution and potential health risk. Sci. Total Environ. 557-558, 438–444. Mandaliev, P.N., Mikutta, C., Barmettler, K., Kotsev, T., Kretzschmar, R., 2014. Arsenic species formed from arsenopyrite weathering along a contamination gradient in circumneutral river floodplain soils. Environ. Sci. Technol. 48, 208–217. https://doi. org/10.1021/es403210y. Manyà, J.J., 2012. Pyrolysis for biochar purposes: a review to establish current knowledge gaps and research needs. Environ. Sci. Technol. 46, 7939–7954. 10.1021/es301029g. Marin, A.R., Masscheleyn, P.H., Patrick, W.H., 1993. Soil redox-pH stability of arsenic species and its influence on arsenic uptake by rice. Plant Soil 152, 245–253. https://doi. org/10.1007/BF00029094. Meunier, L., Walker, S.R., Wragg, J., Parsons, M.B., Koch, I., Jamieson, H.E., Reimer, K.J., 2010. Effects of soil composition and mineralogy on the bioaccessibility of arsenic from tailings and soil in gold mine districts of nova scotia. Environ. Sci. Technol. 44, 2667–2674. Namgay, T., Singh, B., Singh, B.P., 2010. Influence of biochar application to soil on the availability of As, Cd, Cu, Pb, and Zn to maize (Zea mays L.). Aust. J. Soil Res. 48, 638–647. Neethirajan, S., Gordon, R., Wang, L., 2009. Potential of silica bodies (phytoliths) for nanotechnology. Trends Biotechnol 27, 461–467. tibtech.2009.05.002. Norton, G.J., Adomako, E.E., Deacon, C.M., Carey, A.-M., Price, A.H., Meharg, A.A., 2013. Effect of organic matter amendment, arsenic amendment and water management regime on rice grain arsenic species. Environ. Pollut. 177, 38–47. 10.1016/j.envpol.2013.01.049. Ona-Nguema, G., Morin, G., Juillot, F., Calas, G., Brown, G.E., 2005. EXAFS analysis of arsenite adsorption onto two-line ferrihydrite, hematite, goethite, and lepidocrocite. Environ. Sci. Technol. 39, 9147–9155. Park, J.H., Choppala, G.K., Bolan, N.S., Chung, J.W., Chuasavathi, T., 2011. Biochar reduces the bioavailability and phytotoxicity of heavy metals. Plant Soil 348, 439–451. Parsons, C.T., Couture, R.M., Omoregie, E.O., Bardelli, F., Greneche, J.M., Roman-Ross, G., Charlet, L., 2013. The impact of oscillating redox conditions: arsenic immobilisation in contaminated calcareous floodplain soils. Environ. Pollut. j.envpol.2013.02.028. Prakongkep, N., Suddhiprakarn, A., Kheoruenromne, I., Smirk, M., Gilkes, R.J., 2008. The geochemistry of Thai paddy soils. Geoderma 144, 310–324. j.geoderma.2007.11.025. Qiao, J., Liu, T., Wang, X., Li, F., Lv, Y., Cui, J., Zeng, X., Yuan, Y., Liu, C., 2018. Simultaneous alleviation of cadmium and arsenic accumulation in rice by applying zero-valent iron and biochar to contaminated paddy soils. Chemosphere 195, 260–271. https://doi. org/10.1016/j.chemosphere.2017.12.081. Redman, A.D., Macalady, D.L., Ahmann, D., 2002. Natural organic matter affects arsenic speciation and sorption onto hematite. Environ. Sci. Technol. 36, 2889–2896. Sadiq, M., 1986. Solubility relationships of arsenic in calcareous soils and its uptake by corn. Plant Soil 91, 241–248.


P. Leksungnoen et al. / Science of the Total Environment 684 (2019) 360–370

Seyfferth, A.L., Fendorf, S., 2012. Silicate mineral impacts on the uptake and storage of arsenic and plant nutrients in rice (Oryza sativa L.). Environ. Sci. Technol. 46, 13176–13183. Seyfferth, A.L., Morris, A.H., Gill, R., Kearns, K.A., Mann, J.N., Paukett, M., Leskanic, C., 2016. Soil incorporation of silica-rich rice husk decreases inorganic arsenic in rice grain. J. Agric. Food Chem. 64, 3760–3766. Soil Survey Staff, 2014. Keys to Soil Taxonomy. 12th edition. United States Department of Agriculture, Natural Resources Conservation Service, Washington DC. USA. Sparks, D.L., Page, A., Helmke, P., Loeppert, R., Soltanpour, P., Tabatabai, M., Johnston, C., Sumner, M., 1996. Methods of Soil Analysis. Part 3-Chemical Methods. Soil Science Society of America Inc., American Society of Agronomy, Inc., Madison, Wisconsin, USA. Stachowicz, M., Hiemstra, T., Van Riemsdijk, W.H., 2007. Arsenic-bicarbonate interaction on goethite particles. Environ. Sci. Technol. 41, 5620–5625. es063087i. Su, Y.H., McGrath, S.P., Zhao, F.J., 2010. Rice is more efficient in arsenite uptake and translocation than wheat and barley. Plant Soil 328, 27–34. s11104-009-0074-2. Van Zwieten, L., Kimber, S., Morris, S., Chan, K., Downie, A., Rust, J., Joseph, S., Cowie, A., 2010. Effects of biochar from slow pyrolysis of papermill waste on agronomic performance and soil fertility. Plant Soil 327, 235–246. Voegelin, A., Tokpa, G., Jacquat, O., Barmettler, K., Kretzschmar, R., 2008. Zinc fractionation in contaminated soils by sequential and single extractions: influence of soil properties and zinc content. J. Environ. Qual. 37, 1190–1200. jeq2007.0326. Wang, S., Gao, B., Zimmerman, A.R., Li, Y., Ma, L., Harris, W.G., Migliaccio, K.W., 2015. Removal of arsenic by magnetic biochar prepared from pinewood and natural hematite. Bioresour. Technol. 175, 391–395. Wang, H.-Y., Wen, S.-L., Chen, P., Zhang, L., Cen, K., Sun, G.-X., 2016. Mitigation of cadmium and arsenic in rice grain by applying different silicon fertilizers in contaminated fields. Environ. Sci. Pollut. Res. 23, 3781–3788. s11356-015-5638-5.

Wang, N., Xue, X.M., Juhasz, A.L., Chang, Z.Z., Li, H.B., 2017. Biochar increases arsenic release from an anaerobic paddy soil due to enhanced microbial reduction of iron and arsenic. Environ. Pollut. 220, 514–522. envpol.2016.09.095. Weber, F.-A., Hofacker, A.F., Voegelin, A., Kretzschmar, R., 2009a. Temperature dependence and coupling of iron and arsenic reduction and release during flooding of a contaminated soil. Environ. Sci. Technol. 44, 116–122. es902100h. Weber, F.-A., Voegelin, A., Kretzschmar, R., 2009b. Multi-metal contaminant dynamics in temporarily flooded soil under sulfate limitation. Geochim. Cosmochim. Acta 73, 5513–5527. Williams, P.N., Price, A., Raab, A., Hossain, S., Feldmann, J., Meharg, A., 2005. Variation in arsenic speciation and concentration in paddy rice related to dietary exposure. Environ. Sci. Technol. 39, 5531–5540. Williams, P.N., Zhang, H., Davison, W., Meharg, A.A., Hossain, M., Norton, G.J., Brammer, H., Islam, M.R., 2011. Organic matter-solid phase interactions are critical for predicting arsenic release and plant uptake in Bangladesh paddy soils. Environ. Sci. Technol. 45, 6080–6087. Winkel, L., Berg, M., Amini, M., Hug, S.J., Johnson, C.A., 2008. Predicting groundwater arsenic contamination in Southeast Asia from surface parameters. Nat. Geosci. 1, 536–542. Xu, X.Y., McGrath, S.P., Meharg, A.A., Zhao, F.-J., 2008. Growing rice aerobically markedly decreases arsenic accumulation. Environ. Sci. Technol. 42, 5574–5579. https://doi. org/10.1016/j.fcr.2011.11.020. Xu, X., Chen, C., Wang, P., Kretzschmar, R., Zhao, F.-J., 2017. Control of arsenic mobilization in paddy soils by manganese and iron oxides. Environ. Pollut. 231, 37–47. https://doi. org/10.1016/j.envpol.2017.07.084. Zhang, A., Bian, R., Pan, G., Cui, L., Hussain, Q., Li, L., Zheng, J., Zheng, J., Zhang, X., Han, X., 2012. Effects of biochar amendment on soil quality, crop yield and greenhouse gas emission in a Chinese rice paddy: a field study of 2 consecutive rice growing cycles. Field Crops Res 127, 153–160.