An Emerging Global Understanding of Arsenic in Rice (Oryza sativa) and Agronomic Practices Supportive of Reducing Arsenic Accumulation

Abstract

Arsenic uptake in rice (Oryza sativa) is recognized as a global health emergency, requiring the development of agronomic protocols to reduce human exposure to rice having elevated arsenic concentrations. Recent rice-arsenic investigations have centered around numerous agronomic approaches, including: (i) rice breeding and cultivar selection, (ii) altering irrigation water applications to reduce arsenic soil availability, (iii) application of soil amendments which either support arsenic adsorption on iron-plaque or provide antagonistic competition for root uptake, and (iv) phytoremediation. Given that rice cultivars vary in their arsenic accumulation capacity, this manuscript review concentrates on the influences of water management, soil amendments, and phytoremediation approaches on arsenic accumulation. Water management, whether alternating wetting and drying or furrow irrigation, provides the greatest potential to alleviate arsenic uptake in rice. Phytoremediation has great promise in the extraction of soil arsenic; however, the likelihood of multiple years of cultivating hyperaccumulating plants and their proper disposal is a serious limitation. Soil amendments have been soil applied to alter the soil chemistry to sequester arsenic or provide competitive antagonism towards arsenic root uptake; however, existing research efforts must be further field-evaluated and documented as producer-friendly protocols. The usage of soil amendments will require the development of agribusiness supply chains and educated extension personnel before farm-gate acceptance.

1. Introduction

The objectives for this manuscript are two-fold: (i) to specify the arsenic chemistry in soil with a special reference to rice (Oryza sativa), and (ii) to discern agronomic practices that either accentuate or diminish arsenic accumulation in rice.

2. Arsenic as a health issue and its presence in soil

The World Health Organization has established the inorganic arsenic maximum tolerable daily intake at 2 μg kg⁻¹ body weight [1]. Inorganic arsenic intake may lead to gastrointestinal, cardiovascular, central nervous system diseases, as well as bone marrow depression and selective cancers (kidney, lung, bladder) [1, 2]. The World Health Organization and the United States Environmental Protection Agency have established drinking water standards at 10 μg As L⁻¹. Compounding the arsenic water and food threshold levels is problematic because arsenic speciation influences arsenic toxicity, with arsenite (As(III)) being perceived as appreciably more toxic than arsenate (As(V)) [3].

Arsenic soil surface horizon concentrations vary from 0.1 to 67 mg kg⁻¹, with a geometric mean of 5.8 mg kg⁻¹ [4]. Among sedimentary deposits, argillaceous sediments generally have greater arsenic concentrations (trace to 13 mg kg⁻¹) [4]. In Missouri, Aide et al. [5] measured soil arsenic concentrations in 22 pristine soil profiles and reported that the epipedons exhibited arsenic concentrations from 2 to 12 mg kg⁻¹, whereas the argillic and cambic horizons exhibited greater arsenic concentrations, ranging from 10 to 30 mg kg⁻¹. The source of the observed arsenic was speculated to be simply inherited in the parent material. Naturally occurring As-bearing minerals include: arsenopyrite (FeAsS), cobaltite ((Co,Fe)AsS), enargite (Cu₃AsS₄), erythrite (Co₃(AsO₄)₂ 8H₂O), orpiment (As₂S₃), proustite (Ag₃AsS₃), realgar (AsS), and tennantite (Cu₁₂As₄S₁₃) [3, 4].

3. Soil chemistry of arsenic and arsenic speciation

Arsenite (As(III)) exists as the hydroxyl species (H₃AsO₃ - H₂AsO₃¹⁻)), whereas arsenate (As(V)) exists as an oxyanion (H₂AsO₄¹⁻ or HAsO₄²⁻). Arsenite and arsenate may: (i) form complexes with soil organic matter, (ii) become adsorbed onto Mn- Al- and Fe-oxyhydroxides, (iii) become adsorbed onto phyllosilicates, (iv) leach or percolate to deeper soil horizons, or (v) undergo plant uptake [6, 7, 8, 9, 10, 11, 12, 13, 14, 15, 16]. Aide et al. [6] in a soil chemistry review of arsenic in the soil environment discussed (i) arsenic acid–base chemistry, (ii) commonly occurring As-bearing minerals, (iii) thermodynamics of arsenic oxidation – reduction, (iv) arsenic adsorption onto phyllosilicates and Fe-oxyhydroxides, and (v) competitive adsorption. Plant physiology and arsenic have recently been reviewed [7, 8].

4. Arsenite and arsenate as acids

Given the arsenite’s high pK_a1 of 9.2, the dominant arsenite species will be H₃AsO₃ [14]. H₂AsO₃⁻¹ is a weak acid with a pK_a2 = 12.7, thus HAsO₃⁻² has a small activity within the normal alkaline pH range of most soil environments. Conversely, arsenate (H₃AsO₄) readily deprotonates to form H₂AsO₄⁻¹ (pK_a1 = 2.3). Additionally, H₂AsO₄⁻¹ will deprotonate to form HAsO₄⁻² (pK_a2 = 6.8), thus H₂AsO₄⁻¹ and HAsO₄⁻² are the dominant arsenate species in most soils. HAsO₄⁻² will deprotonate to form AsO₄⁻³ (pK_a3 = 11.6); however, this species will only exist in the most extreme alkaline soil environments. Monomethylarsonic acid (MMA or CH₃AsO(OH)₂ with pK₁ = 3.6 and pK₂ = 8.2) and dimethylarsenic acid (DMA or (CH₃)₂AsO(OH) with pK₁ = 6.2) may also readily exist in soil environments [17].

5. Arsenite and arsenate as oxidation and reduction species

Wagman et al. [18] used standard free energies of formation to determine half-cell reactions for arsenate to arsenite reduction:

$$\left(0.5\right){\mathrm{H}}_2\mathrm{A}\mathrm{s}\mathrm{O}{4}^{-}+\mathrm{e}-+1.5{\mathrm{H}}^{+}=0.5As{\left(OH\right)}_3+0.5{\mathrm{H}}_2\mathrm{O}pK=-10.84$$

$$\left(0.5\right){{\mathrm{HAsO}}_4}^{2-}+\mathrm{e}-+2{\mathrm{H}}^{+}=0.5As{\left(OH\right)}_3+{\mathrm{H}}_2\mathrm{O}pK=-14.22$$

Using protocols from Essington [19], the predominance diagram (Figure 1) illustrates the relative stability regions for arsenite and arsenate species expected in the soil environment, ranging from pH 3 to pH 9. The Pourbaix diagram (predominance diagram) shows the transitional nature of As(V) as a proton donor and the reduction of As(V) to As(III). The demarcation of oxic, suboxic, and anoxic regimes was discussed in Essington and we note that arsenate largely exists in oxic to suboxic regimes [19]. Arsenite formation in anoxic soil environments is thermodynamically favored in increasingly acidic soil environments.

Figure 1.

Predominance diagram showing arsenic species predominance zones for given pe and pH as master variables (created by authors of this manuscript).

Arsenic reduction is mediated by the soil’s microbial population, effectively supporting electron donation from suitable organic substrates. Dissimilatory arsenate-reducing bacteria can effectively reduce arsenate to arsenite by using arsenate as a terminal electron acceptor [20, 21, 22]. Xu et al. [23] demonstrated that reduction of arsenate to arsenite post root uptake, coupled with efflux from the root to the rhizosphere, also contributes to arsenate reduction. Qiao et al. [24] employed anaerobic microcosms to demonstrate that humic substances facilitate arsenic reduction. Fulvic acid was more effective in reducing arsenic than humic acid, and humic acid was more effective in reducing arsenic than humin. As a carbon source, fulvic acid supported microbial activity and reduced fulvic acid acted as an electron shuttle to reduce Fe(III)-oxyhydroxides and As(V). Arsenic may be co-precipitated with Fe-oxyhydroxides, and the reductive dissolution of these Fe-oxyhydroxides may promote the release of arsenate, which may then be subsequently reduced to arsenite. Mn-oxyhydroxides have been implicated in the oxidation of arsenite to arsenate [25, 26, 27, 28, 29].

6. Groundwater irrigation as an arsenic source for Rice accumulation

In the Indo-Gangetic Plain, Vicky-Singh et al. [30] documented arsenic concentrations of soil surface horizons and surface and groundwater resources. They reported that tube-well water ranged from 5.3 to 17.3 μg L⁻¹ and soil horizons ranged 1.09 to 2.48 mg kg⁻¹, with data showing that tube well water irrigation was contributing arsenic to soil. In the Mekong Delta (Vietnam), Huang et al. [31] documented multiple groundwater samples having arsenic concentration greater than 50 μg L⁻¹ and demonstrated that As(III) was the more abundant valance state. The historical applications of As-bearing groundwater correlated with arsenic soil accumulation. Radu et al. [28] performed a batch experiment involving pyrolusite (MnO₂) to show that second order kinetics, which incorporated MnO₂ concentrations, described arsenite oxidation. Subsequent arsenate adsorption was appropriate described using the Langmuir equation.

Farooq et al. [32] investigated arsenic accumulation associated with irrigation and agronomic practices in the Bengal Delta. Two different fields were irrigated with different arsenic concentrations in the groundwater, with one field planted to wheat and the other field planted to rice. These authors indicated that the more concentrated As-bearing groundwater in the rice field did not increase the arsenic soil concentrations as significantly as the wheat field, which was irrigated with less concentrated As-bearing groundwater. The authors proposed and provided evidence that greater quantities of rice plant residue, with its production of organic acids, supported arsenic diffusion to deeper soil horizons. Arsenic concentrations exceeding 10 μg L⁻¹ appear to be more frequent in the western United States, demonstrating that the local geology is important in influencing water quality [33]. In a recent review, Mohanty [2] documented the efficacy and deficiencies of technologies involving treatment of arsenic-bearing groundwater, which may be employed to improve irrigation water quality.

7. Adsorption of Arsenite and arsenate species

Arsenite and arsenate species experience pH-dependent adsorption and co-precipitation with Fe-oxyhydroxides, most notably ferrihydrate (β-FeOOH), lepidocrocite (γ-FeOOH), goethite (α-FeOOH), and hematite (Fe₂O₃) [3, 17]. Surface protonation of goethite (pK = −9.6) permits the interface to acquire amphoteric positive charge densities sufficient to promote monodentate or bidentate arsenite adsorption [6, 29]. Arsenite and arsenate adsorption may result in both monodentate and bidentate bonding structures [3, 10, 14, 17, 22, 34, 35, 36].

The optimal pH for arsenic adsorption depends on (i) experimental protocols, and (ii) the presence of phosphate, silicic acid, naturally occurring organic acids, and other competing anions. The optimal pH for the adsorption of arsenite on Al- and Fe-oxyhydroxides ranges from pH 7 to 10, [14, 34, 35, 36, 37, 38, 39, 40, 41, 42, 43], whereas the optimal pH for the adsorption of arsenate on Al- and Fe-oxyhydroxides varies across the pH range of 4 to 7 [17, 37, 44, 45, 46]. Cornu et al. [13] observed an arsenate adsorption pH dependency with both kaolinite and humic acid treated kaolinite. Interestingly, Cornu et al. [13] observed that arsenate adsorption onto humic acid treated kaolinite was greater than for untreated kaolinite when the electrolyte solution was a Ca(NO₃)₂ media, whereas arsenate adsorption was substantial decreased on humic acid treated kaolinite in NaNO₃ media. Goldberg [14] investigated arsenic adsorption on Al-oxides, Fe-oxides and reference phyllosilicates (kaolinite, illite and montmorillonite). Arsenate adsorption was pH-dependent with arsenate adsorption less evident on transition to alkaline media. Arsenate adsorption decreased above pH 9 for Al-oxides, above pH 7 for Fe-oxides and pH 5 for the reference clays. Arsenite adsorption showed a maximum adsorption near pH 8 for non-crystalline aluminum oxides and exhibited little pH dependence on non-crystalline Fe-oxides [14].

Jackson and Miller [17] evaluated various concentrations of phosphate (pH 3 and 7) to extract arsenite, arsenate, dimethylarsinic acid, and monoethylarsonic acid adsorbed onto goethite and non-crystalline Fe-oxyhydroxides. Phosphate was demonstrated to displace arsenite and arsenate. Khaodhiar et al. [47] prepared iron oxide coated sand (Fe₂O₃) to show that arsenate adsorption was strongly adsorbed at acidic to slightly acidic pH values and adsorption decreased with increasing pH. Grafe et al. [35] investigated arsenite and arsenate adsorption on goethite and observed that arsenate adsorption decreased gradually and continuously from pH 3 to pH 11. Arsenite adsoption was shown to have a maximum adsorption at pH 9. The influence of either fulvic acid or humic acid addition resulted in a reduction in adsorption for both arsenite and arsenate.

Sulfate, carbonate, and dissolved organic matter have been shown to be relatively less effective than phosphate in displacing arsenic [37]. Using goethite as the bonding surface, Luxton et al. [34] showed that silicic acid (H₄SiO₄) was able to effectively displace arsenic. Swedlund and Webster [48] demonstrated that H₄SiO₄ may displace arsenite from ferrihydrate. Zhang and Selim [39] observed arsenic desorption by phosphate, whereas Xu et al. [43] documented arsenic phosphorus-induced desorption from crystalline and non-crystalline aluminum oxides. Smith and Naidu [46] provided data on the kinetics of arsenic desorption, illustrating the importance of studies to understand the equilibrium is rarely achieved in natural systems.

Yamamura et al. [20] amended soils with As(V) laden Fe oxyhydroxides with solution supplemented with either lactate or acetate. After 40 days, there was a greater arsenic release rate in lactate amended systems, suggesting that lactate is a suitable carbon source and both dissimilatory metal(loid) reducers and anaerobic fermenters support arsenic extraction. Razzak et al. [49] documented oxidation–reduction processes in groundwater support simultaneous release of iron and arsenic, thus demonstrating that groundwater irrigation may be an effective arsenic source.

8. Influence of iron plaque on Rice roots and its effect on arsenic uptake

Aquatic plants frequently show accumulations of iron and manganese coatings (Fe-plaque) on root systems, commonly attributed to more oxidized soil conditions in the root rhizosphere leading to ferrous to ferric ion production and subsequent hydrolysis to a Fe-Mn oxyhydroxide status. Many researchers have investigated whether Fe-plaque on rice root systems act as a preferential adsorption site for arsenic, thus limiting the potential for arsenic accumulation in plant organs [12, 43, 50, 51, 52]. Most authors acknowledge that the degree of arsenic adsorption by Fe-plaque and the protection afforded towards limiting arsenic accumulation in plant tissue is dependent on (i) soil pH, (ii) the soils oxidation oxidation–reduction status of the bulk soil and the rhizosphere, (iii) the microbial composition, (iv) the quantity of Fe-plaque present on the rice roots, (v) the stage of growth of the rice plant, (vi) the arsenic flux towards the root system and (vii) the presence of competing anionic species in the adsorption processes.

Dong et al. [50] observed that Fe-Mn plaque formation on rice roots was increased because of inoculation with Fe/Mn-oxidizing bacterial strains. The activity of bacterial strains, in combination with exogenous ferrous iron, significantly decreased As and Cd uptake in rice. Interestingly the untreated check showed the following rice plant arsenic concentrations: 354 mg kg⁻¹ for roots, 14.2 mg kg⁻¹ for stem (culm), 24.4 mg kg⁻¹ for leaf, and 0.81 mg kg⁻¹ for brown rice. Conversely, the bacterial strains plus exogenous Fe(II) showed the following rice plant organ arsenic concentrations: 259 mg kg⁻¹ for roots, 13.0 mg kg⁻¹ for stem (culm), 19.2 mg kg⁻¹ for leaf, and 0.72 mg kg⁻¹ for brown rice.

9. Arsenic and plant physiology

Arsenic may limit rice growth and development [1]. Finnegan and Chen [7] and Sharma et al. [8] each reviewed the plant physiology of arsenic on plant growth and development. These authors discussed evidence that arsenite and arsenate are taken up by root cells, but arsenate is rapidly reduced to arsenite. Cellular disruption may be caused by both arsenite and arsenate; however, the mechanisms are distinctly different.

Arsenite is dithiol reactive and readily binds and potentially inactivates selective cysteine containing enzymes and dithiol co-factors. As(III) enters root cells via aquaporin (nodulin26-like intrinsic proteins) and xylem export to stems may occur. Arsenite may bind from one to three sulfhydryl groups, influencing the physiologic behavior of transcription factors, signal transduction proteins, proteolytic proteins, metabolic enzymes, redox regulatory enzymes, and structural proteins. The binding of As(III) to thiols may constitute the main detoxification pathways [7, 8]. Arsenate may replace Pi in critical biochemical reactions: (i) glycolysis, (ii) oxidative phosphorylation, (iii) phospholipid metabolism, (iv) DNA and RNA metabolism, and (v) cellular signaling [7, 8]. Both arsenite and arsenate may increase oxidative stress by inducing the production of reactive oxygen species; that is, the production of superoxide (O₂•⁻), hydroxyl radical (•OH), and peroxide (H₂O₂). Glutathione (a tripeptide with linkage between the carboxyl group of the glutamate side-chain and cysteine) is an antioxidant that assists in preventing reactive oxygen species from disrupting cellular function. Ascorbate may also limit reactive oxygen species damage [7, 8].

Other metabolic consequences of arsenic include: (i) chloroplast shape irregularities and reduction of chlorophyll content, (ii) altered carbohydrate metabolism involving sucrose and starch, (iii) reduced micronutrient uptake, (iv) altered ATP synthesis, (v) altered stomatal conductance, (vi) altered lipid metabolism and the integrity of cellular membranes [8]. Belefant-Miller and Beaty [53] observed the plant distribution of arsenic in rice plants might influence “straighthead”. Yan et al. [54] identified soil arsenic bioavailability is associated with “straighthead” disorder in rice. Lim et al. [55] reviewed the effect of arsenic compounds on plant growth. In a subsequent review, Kofronova et al. [56] focused on arsenic physiology in hyperaccumulating plants and documented the following research outcomes: (i) arsenic interfered with basic cellular metabolism, including carbohydrate metabolism in photosynthesis, (ii) arsenite and arsenate were xylem transported, (iii) arsenate reduction was associated with arsenate reductase and arsenite interacted with glutathione for passage into the cell’s vacuole, (iv) arsenate interfered with cell wall physiology, decreased ribulose-1,5 biphosphate carboxylase/oxygenase, and competed with phosphorus in oxidative phosphorylation. Arsenite interfered with hormonal physiology and restricted pigment system II and chlorophyll functioning.

In a greenhouse project, Jung et al. [57] amended soil at arsenic rates of 0 (untreated check), 25, 50 and 75 mg As kg⁻¹. The 50 mg As kg⁻¹ amended level inhibited shoot growth. Chauhan et al. [58] observed that the presence of increased heavy metal activity and arsenic availability reduced the activity of key soil enzymes, suggesting that bacterial diversity and microbial functioning were impaired.

10. Phytoremediation of arsenic impacted soils

Considerable research has focused on the subtropical fern Pteris vittata as an effective phytoremediation species for arsenic removal from impacted soils [59, 60]. In a two-year study, Lei et al. [59] observed that P. vittata effectively removed soil accumulated arsenic; however, the study highlighted the need to assess arsenic atmosphere deposition to ascertain the proper arsenic removal capacity. Rahman et al. [61] performed solution culture and soil container greenhouse trials involving Pteris multifida to assess this species efficacy to hyperaccumulate arsenic. P. multifida was able to accumulate arsenate. P. multifida also was more suitable for non-tropical climate than Pteris vittate. Yang et al. [60] performed a greenhouse trial to assess the influence of monoammonium phosphate and citric acid amendments to improve the efficacy of Pteris vittate to hyperaccumulate arsenic. Both monoammonium phosphate and citric acid augmented the phytoremediation efficacy of P. vittata.

11. Soil amendments and their efficacy in reducing Rice arsenic accumulation

Numerous rice researchers have documented the effectiveness of soil amendments to mitigate rice arsenic accumulation. Toor and Haggard [44] and Wu et al. [62] investigated the effectiveness of phosphate, whereas Li et al. [63], Wei et al. [64] and Swedlund and Webster [48] each investigated the effectiveness of silicon (Si). Wei et al. [64] evaluated several Si-bearing products to evaluate their efficacy to increase rice yield and reduce rice uptake of arsenic, lead (Pb) and cadmium (Cd). The Si-bearing materials increased rice yield and reduced root to shoot transfer of As, Pb and Cd. Zou et al. [65] and Gemeinhardt et al. [66] investigated the effectiveness of ferrous sulfate, demonstrating that Fe²⁺ oxidation and Fe-oxyhydroxide synthesis in the rhizosphere may provide a substrate for arsenic adsorption. Wu et al. [62] investigated biochar modified with Fe compounds as soil amendments to reduce arsenic bioavailability, with Fe-oxyhydroxide-sulfate showing promise as an effective amendment by reducing arsenic extraction with NaHCO₃.

The application of phosphorus amendments in greenhouse pot culture experiments with wheat in dry cultures and rice in flood cultures revealed that phosphorus applications increased arsenic concentrations in both the wheat and rice experiments [67]. Thin film diffusive gradient technology showed that arsenic release from the soil’s solid phase was augmented by phosphorus competition. Kaur et al. [68] documented that selenium was effective in reducing arsenic uptake. Arsenic concentrations were lowered in the roots, straw, and seed because of the selenium amendments. Future research is desired to explore selenium as an effective soil amendment to reduce arsenic rice accumulation. Wang et al. [69] showed promise that microalgae in paddy fields could sequester arsenic prior to rice root uptake, thus limiting arsenic accumulation in rice.

12. Irrigation management to limit Rice arsenic accumulation

Irrigation management of rice has been extensively studied to determine if restricted water application may result in reduced arsenic uptake [70, 71, 72, 73, 74, 75]. In Missouri, Aide et al. [71], in a two-year rice study, investigated two irrigation regimes involving delayed flood and furrow irrigation on silt loam and clayey soils to assess arsenic uptake. Across both years and soil-types, rice total arsenic uptake was substantially reduced in rough rice seed for the furrow irrigated regime. Aide and Goldschmidt [72] in a two-year project involving 20 rice varieties similarly demonstrated that furrow irrigated rice had dramatically reduced arsenic concentrations in paddy (rough) rice compared to delayed flood irrigated rice. All furrow irrigated rice had rough rice total arsenic concentration below 0.1 mg kg⁻¹, with 17 of the 20 rice varieties having less than 0.05 mg As kg⁻¹. The mean arsenic concentrations for the delayed flood rice regime were approximately 0.28 mg As kg⁻¹.

Aide demonstrated that furrow irrigation, involving three rice varieties in 2018 [75] and six varieties in 2019 [73], resulted in substantially smaller arsenic concentrations in rice straw and rough rice seed than delayed flood. Aide [74] in a review of water availability and research involving water-restricting irrigation regimes in Egypt, India and Eastern Asia demonstrated that alternate wetting and drying irrigation frequently was water conserving and limited arsenic uptake. Many of the research citations documented rice yields that were comparable to traditional irrigation regimes; however, additional research remains to be performed to provide consistency in yield attainment. An additional benefit of reduced irrigation of rice was a reduction in methane emission, a potent greenhouse gas [70].

Carrijo et al. [76] performed a compelling rice meta-study involving 56 studies comparing continuous flood with alternate wetting and drying (introduction of unsaturated soil water conditions), with most of the studies derived from Asia. They defined and partitioned alternate wetting and drying irrigation regimes into “safe” or “mild” (where the soil water matric potential was equal to or smaller than −20 kPa) and “severe” (where the soil water matric potential was below −20 kPa). The meta study documented the following: (i) the presence of unsaturated soil water conditions imposed during the entire growing season depressed rice yields, whereas unsaturated soil water conditions prior to either heading only (vegetative) or post heading (reproductive) only demonstrated little to zero yield loss, (ii) in most cases mild or safe alternate wetting and drying do not depress rice yields, whereas severe alternate wetting and drying showed yield reductions, (iii) yield losses were more significant in low organic matter soils or soils having alkaline pH levels, (iv) compared to the continuous flood system the alternate wetting and drying systems exhibited smaller water use rates and where mild alternate wetting and drying was practiced the water use efficiency was greater.

In China, He et al. [77] compared rice growth characteristics and yields in flood and non-flood systems and documented that rice root length density, leaf dry weight, shoot dry weight, and root activity were greater in the non-flood irrigation system at mid-tillering. Yields were typically greater in the flood system across all treatments. In California, Li et al. [78] investigated several alternate wetting-drying irrigation systems with respect to continuous flood. The alternate wetting and drying were imposed at panicle initiation or 50% heading, with various degrees of drying established for each crop growth stage. At crop maturity total arsenic concentrations were greatest for the root system (14.8 mg As kg⁻¹), whereas straw arsenic concentrations were 0.64 mg As kg⁻¹. The arsenic concentrations in the root systems were primarily associated with Fe-plaque. Grain arsenic concentrations, when compared to continuous flood, were 57% redcued for brown rice and 63% reduced for polished rice. As the driest alternate wetting-drying episode, rice grain exhibited 78% less DMA and 40% less inorganic arsenic when compared to continuous flood. In California, Carrijo et al. [79] observed rice irrigation involving continuous flood and three alternate wetting and drying irrigation regimes with differences in drying severity (low, medium (−71 kPa) and high (−154 kPa) and three timings of the drying episodes (panicle initiation, booting and heading. Imposition of the medium and high drying episodes decreased arsenic uptake by 41 to 61%. The booting and heading drying episodes showed better arsenic mitigation responses.

13. Prospects and research needs

Arsenic accumulation in rice is a substantial global concern [1, 4, 6]. The soil chemistry of arsenic accumulation in rice is rapidly being elucidated; however, studies have yet to develop consistent desirable outcomes with respect to irrigation technology, soil amendments, phytoremediation, and yield maintenance. Alternate wetting and drying and furrow irrigation are competing irrigation regimes, with research showing substantial reductions in arsenic accumulation. However, rice yield maintenance, implementing reliable nitrogen fertilization practices, and providing effective weed management programs remain problematic, especially when food security and traditions may be compounding realities. Water scarcity and climate change provide both opportunities and setbacks to altering irrigation methods [80].

The understanding of rice physiology and arsenic is beginning to be formulated. Das et al. [81] illustrated the importance of biochemical relationships involving ascorbate-glutathione cycle and thiol metabolism to support reducing yield suppression in arsenic impacted rice. Wu et al. [82] showed the promise of arsenic-phosphate interactions involving phosphate transporter expression in rice. Thus, understanding arsenic root uptake at the cellular membrane level and its subsequent movement within the plant, combined with rice breeding and cultivar selection, remain clear avenues of research to reduce the human daily uptake of arsenic.

The prospect of reducing arsenic uptake rests with a global effort to: (i) produce cultivars that restrict arsenic uptake to root cells and exude arsenic to the rhizosphere, and (ii) alter irrigation practices to provide sufficient intervals of oxic soil environment to mitigate arsenic bioavailability. These approaches will also provide other environmental advantages, including water conservation and reduced methane emission [83].