International Journal of Applied Agricultural Research
ISSN 0973-2683 Volume 11, Number 1 (2016) pp. 1-28
© Research India Publications
http://www.ripublication.com
Arsenic In The Soil Environment: A Soil Chemistry
Review
Michael Aide and Donn Beighley* and David Dunn**
Southeast Missouri State University
E-mail: [email protected]
Fisher Delta Research Center
University Missouri
Abstract
Arsenic in the soil environment has gained renewed interest because of the
emerging acknowledgement that arsenic accumulation in rice is a global
concern. This review reflects the current state of research being provided to
the understanding of arsenic in the soil environment with an emphasis on
arsenic uptake in rice. Arsenic speciation and the chemical reactions
associated with arsenite, arsenate and methylated-arsenic species is of prime
importance. The chemistry of soil arsenic is both abiotic and biotic and its
chemistry is complicated by (i) oxidation-reduction processes, (ii) acid-base
reactions, (iii) adsorption-precipitation reactions, and (iv) plant uptake and
accumulation. Ultimately plant genetics and emerging irrigation regimes,
predicated on our collective understanding of the role soil chemistry, provide
the opportunity to alter agriculture production to safeguard the global food
supply.
Keywords: arsenate, arsenite, methylarsenic acid, rice, water quality
Arsenic In Rice Production Soils: A Soil Chemistry Review :
Introduction
Arsenic (As), has atomic number 33 in Group V of the Periodic Table and it is a
metalloid elementhaving an [Ar] 3d10
4s2p
3 electronic configuration. Electron removal
readily produces two stable valence states: (i) As(III) or arsenite having an electronic
configuration [Ar] 3d10
4s2and (ii) As(V) or arsenatehaving an electronic configuration
[Ar] 3d10
. Arsenic and phosphorus are both Group V elements and share similarities
in their soil chemistry. As an example, the covalent atomic radius of phosphorus is
106 picometers and arsenic is 119 picometers (Greenwood and Earnshaw, 1984).
2 Michael Aide and Donn Beighley and David Dunn
The World Health Organization has established a provisional maximum tolerable
daily intake of inorganic arsenic at 2 µg As/ kg body weight (Tuli et al., 2010).
Inorganic arsenic intake may lead to gastrointestinal, cardiovascular and central
nervous symptoms, bone marrow depression, haemolysis, hepatomegaly, melanosis,
polyneuropathy, and encephalopathy. Inorganic As is a non-threshold class 1
carcinogen (Tuli et al., 2010). The key sources of dietary arsenic intake include
drinking water and selected plant materials derived from their cultivation in soils
having enhanced arsenic uptake. World Health Organization (WHO) and the United
States Environmental Protection Agency (USEPA) drinking water limits for arsenic
are identical at 10 µg As/liter.
Arsenic in the Pristine Environment
Naturally occurring soil arsenic concentrations may be attributed to mineral
weathering of As-bearing minerals and the subsequent migration of arsenicusing
geologic and pedogenic pathways. Soils are considered open thermodynamic systems
and are available to exchange mass and energy with their surroundings. Thus, earth
and soil processes involving atmospheric deposition, the hydrologic cycle, plant and
soil organism biocycling, and soil chemical processes all function to disperse and also
intensifyarsenic concentrations. Most streams and lakes in non-impacted regions of
the USA have arsenic concentrations less than 1 µg dissolved As / liter. Similarly,
precipitation in non-impacted areas is commonly less than 1 µg dissolved As / liter,
thus precipitation in non-impacted areas in not a major arsenic source.
Commonly occurring As-bearing minerals include: arsenopyrite (FeAsS), cobalite
((Co,Fe)AsS), enargite (Cu3AsS4), erythrite(Co3(AsO4)2 ●8H2O), orpiment (As2S3),
proustite (Ag3AsS3), realgar (AsS), and tennantite (Cu12As4S13). Minor arsenate
minerals include: annabergite (Ni3(AsO4)2●8H2O), austinite (CaZn(AsO4)(OH),
clinoclase (Cu3(AsO4)(OH)3, conichalcite (CaCu(AsO4)(OH), cornubite
(Cu5(AsO4)2(OH)4, cornwallite (Cu5(AsO4)2(OH)2, and mimetite (Pb6(AsO4)3Cl
(Kabata-Pendias, 2001; Miretzky and Cirelli, 2010). Arsenide minerals include:
loellingite (FeAs2), safforlite (CoAs), niccolite (NiAs), and rammelsbergite (NiAs2).
Mineral weathering is a complex set of geologic and pedogenic processes influenced
by precipitation, temperature, soil drainage class, and biotic processes involving
microbial and plant populations,
Arsenopyrite is commonly accepted as the most abundant arsenic mineral. Pyrite
(FeS2) , galena (PbS), sphalerite (Zn,FeS), marcasite (FeS2) and chalcopyrite
(Cu,FeS2) are commonly known to contain arsenic as an impurity. Pyrite oxidation
has received significant research attention because of its occurrence in mine tailings
and subsequent creation of acid mine drainage. The overall stoichiometric oxidation
of pyrite overall may be written as:
FeS2 + 3.75 O2 + 4 H2O = Fe(OH)3 + H3AsO4 + H2SO4.
Pyrite oxidation may be an entirely abiotic process; however the reaction rates are
substantially greater if mediated by Thiobacillus ferrooxidans, Thiobacillus
thiooxidans, and Leptosprillum ferrooxidans (Dorofeev et al., 1990Harrington et al.,
1998; Zeman et al., 1995).
Arsenic In The Soil Environment: A Soil Chemistry Review 3
The oxidation of pyrite may also be accelerated by nitrates and associated prokaryotic
organisms (Appelo and Postma, 1993):
10FeS2 + 30 NO3- + 20 H2O = Fe(OH)3 + H3AsO4 + H2SO4.
The stoichiometric oxidation of arsenopyrite may be expressed as:
FeAsS + 3.5 O2 + 4H2O = Fe(OH)3 + H3AsO4 + H2AsO4.
It is also important that arsenic sulfite (As2O3) may abiotically form in non-thermal
waters.
Argillaceous sediments generally have greater arsenic concentrations (trace to 13 mg
As/kg) than other sediment types(Kabata-Pendias, 2001). Arsenic surface soil horizon
concentrationsvary from 0.1 to 67 mg As / kg, with a geometric mean of 5.8 mg
As/kg, with much of the variation attributed to soil order or parent material
inheritance (Heitkemper et al., 2009; Kabata-Pendias, 2011). Pettry and Switzer
(2001) surveyed parent materials and more than 200 soils in the State of Mississippi,
observing parent material arsenic concentrations spanning from 1.8 mg As/kg in the
Wilcox Formation to 33.8 mg As/kg in the Winona Formation. Soil arsenic levels
ranged from 0.26 to 24.4 mg As/kg. Chen et al. (2001) investigated 448 Florida soils
for baseline assessment of arsenic. They observed for undisturbed soils a baseline of
6.2 mg As/kg and a baseline of 7.3 mg As/kg for disturbed soils. Chen et al. (2008)
observed arsenic placement in California soils as a result of micronutrient fertilizer
placement. They documented that arsenic was correlated with P and Zn fertilization in
one subregion and correlation of arsenic with Zn in another subregion. In Missouri,
Aide et al. (2013) discussed soil arsenic concentrations in 22 soil profiles having no
history of arsenic contamination and reported that the surface horizons exhibited
arsenic concentrations ranging from 2 to 12 mg As/kg. Subsurface horizons,
particularly argillic horizons, possess substantially greater arsenic concentrations,
ranging from 10 to 30 mg As/kg. Most of the examined soils exhibited significant As-
Fe concentrationcorrelations. As an example, the Alred series (Loamy-skeletal over
clayey, siliceous, semiactive, mesicTypicPaleudalfs: very deep, well-drained soil
formed in cherty hillslope sediments and the underlying clayey residuum) shows
approximately 1 mg As / kg in the A horizon and appreciably greater arsenic
abundances in the deeper argillic horizon. In an earlier study, Aide et al. (2005)
demonstrated that arsenic was strongly associated with pedogenic Fe and Mn-bearing
nodules in poorly-drained Missouri Alfisols developed on silty terraces.
Anthropogenic Activities and Arsenic as a Contaminant
Staed et al. (2009) surveyed soil lead (Pb) and arsenic concentrations in the Arkansas
Ozark Highlands having a history of Pb-arsenate applications to control agricultural
pests. These authors observed significant soil Pb-As concentration correlations with
arsenic concentrations ranging from trace to approximately 20 mg As/kg. Menjoulet
et al. (2009) investigated the arsenic runoff potential from Arkansas soils impacted by
As-bearing poultry litter. During some rainfall events, flow weighted mean arsenic
concentrations of the runoff waters exceeded maximum contaminant levels for
drinking water. Lead arsenate (PbHAsO4) usage in Washington orchards substantially
impacted soil arsenic concentrations, with appreciable arsenic migration below the
near-surface soil horizons (Peryea, 1991). Phosphorus fertilization of these
4 Michael Aide and Donn Beighley and David Dunn
Washington soils promoted arsenic availability (Davenport and Peryea, 1991). Cullen
and Reimer (1989) noted the biotransformations of arsenic in sediments, documenting
the production of organic arsenic and its potential threat to ground water.
Anthropogenic arsenic activities that impact soil include: (i) deposition of coal-fired
power plants particulates and aerosols, (ii) mining and smelting operations, (iii)
application of As-bearing agricultural pesticides, (iv) poultry feed additives, and (v)
irrigation with As-bearing water (Belefant-Miller and Beaty, 2007; Biswas et al.,
2014; Chen, 2008b; Roussel et al., 2000; Staed et al., 2009; Tuli et al., 2010; Vicky-
Singh 2010., Yan et al., 2008; Welch et al., 2000; Garcia-Manyes et al., 2002).
Arsenite is considered to be 25 to 60 times more toxic than arsenate (Miretzky and
Cirelli, 2010).In the Mid-South region of the United States, soil arsenic concentrations
because of As-bearing herbicides previously applied to cotton (Gossypiumhirsutum
L.) is a concern. Huang and Kretzchmar (2010) developed a sequential selective
extraction protocol for arsenic species in soil.
Toor and Haggard (2009) observed that arsenic in poultry litter and granulates did not
enhance arsenic availability when litter application rates were based on phosphorus
soil test recommendations. Fox and Doner (2003) observed arsenic, molybdenum and
vanadium in constructed wetlands. Unique to this study is that the surface of the
constructed wetland was anoxic; whereas the deeper layers were suboxic to oxic.
Total arsenic concentrations decreased with sediment depth in the constructed
wetland, a feature more attributed to arsenic loss at the deeper depths than arsenic
surface accumulation. In the deeper layers moderate reducing conditions favored Fe-
oxyhydroide dissolution, whereas in the more intensely reducing surface layers As-
sulfides may be more stable. Radu et al. (2005) investigated water flow in Fe-oxide
coated sand, noting that arsenite was more mobile at pH 4.5 than pH 9. These authors
also observed that increasing pore water velocities resulted in greater arsenite
mobility’s. Zhang and Selim (2007) investigated the likelihood that colloidal
dispersion and the subsequent transport of arsenic adsorbed to a colloidal fraction in
leachate waters was a substantial factor influencing arsenic mobility in soils.
Mobilization of colloidal material and facilitated As(III)-colloidal Fe-oxide material
was indicated.
Roussel et al. (2000) investigated drainage waters from As-bearing mine tailings in
France. Arsenic transport was 220 times greater as suspended particulate material
(SPM) than dissolved arsenic. Selective sequential extractions revealed that
approximately 78% of the arsenic in the SPM was associated with hydrous oxides,
notably lepidocrocite. The authors cautioned that oxidation-reduction and pH changes
during water transport from the mine tailings to surface water has the potential to
strongly influence arsenic speciation. Ackermann et al. (2010) studied the
bioavailability of German floodplain soils impacted by long term mining activities.
Incubated sediments from the Mulde River were subjected to selective arsenic
extractions, demonstrating that arsenic was primarily bound to ammonium oxalate Fe-
oxyhydroxides (poorly crystalline Fe-oxyhydroxides) and that these Fe-
oxyhydroxides were subject to change during cycling oxidation and reduction
regimes. In point-of-fact, the frequency of the cyclic oxidation-reduction regimes was
important in influencing arsenic and Fe equilibrium attainment. The authors did
Arsenic In The Soil Environment: A Soil Chemistry Review 5
confirm that frequent episodes promoted increased soluble arsenite concentrations.
Variations in the arsenite and arsenate concentration ratios implied little equilibrium
attainment because of both biotic and abiotic processes.
Devesa-Ray et al. (2008) fractionated soils from Spain demonstrating that arsenic was
strongly associated with Al-Fe oxides and the residual fractions. In the
Freiberg/Saxony region, Schug et al. (1999) selected 159 soils for aqua-regia
digestion and reported a mean recovery of 23.6 mg As/kg and having a range of 0.74
to 6690 mg As/kg.
Jones et al. (1999) irrigation influences on soils in the Madison River Basin, Montana.
Arsenic in the Madison River averaged 0.24 µM; however, the headwaters, near
Yellowstone National Park, possessed 5 µM arsenic. Ground water near Yellowstone
National Park averaged 0.67 µM arsenic. Soil receiving irrigation water; however, did
not demonstrate enhanced arsenic solubility and selective sequential extractions did
not demonstrate any differences in their arsenic fractionation. In the Indo-Gangetic
Plains of northwestern India, Vicky-Singh et al. (2010) documented surface horizons
of selected soils and surface and ground water for arsenic. They observed that tube-
well water ranged from 5.33 to 17.27 µg/liter and soils ranges from 1.09 to 2.48 mg
As/kg. Statistical analysis revealed that tube well water was correlated with soil
arsenic concentrations, suggesting that irrigation impacted these soils.
Soil Chemistry of Arsenic
In the natural environment arsenic exists as two distinct chemically species: (i)
arsenite as a hydroxyl species(H3AsO3 –H2AsO31-
) and (ii) arsenate as an oxyanion
(H2AsO41-
or HAsO42-
).In soils, arsenite and arsenate (i) form complexes with soil
organic matter, (ii) become adsorbed onto Al- and Fe-oxyhydroxides, (iii) become
adsorbed onto phyllosilicates, (iv) leach or percolate to deeper soil horizons, and (v)
undergo plant uptake (Bowell, 1994; Chen et al., 2008a; Fendorf et al., 2004; Fox and
Doner, 2003; Roussel et al., 2000; Xu et al., 2009; Liu et al., 2005).
Arsenite and Arsenate as Acids
Given the high pKa1 for arsenite, the dominant species will be H3AsO3 (Goldberg,
2002):
(1) H3AsO3 = H2AsO31-
+ H+ pKa1 = 9.2
and
(2) H2AsO31-
= HAsO32-
+ H+ pKa2 = 12.7.
Simulation of arsenite speciation with a total arsenic concentration of 0.01 mole/liter
is displayed in Figure 1. Clearly, in acidic and neutral soil pH environments H3AsO3
will be the dominant species, with H2AsO31-
only a major species in strongly alkaline
soil environments.
6 Michael Aide and Donn Beighley and David Dunn
4 5 6 7 8 9 10 11
pH
0
0.001
0.002
0.003
0.004
0.005
0.006
0.007
0.008
0.009
0.01
Co
ncen
trati
on
(M
ole
/Lit
er)
H2AsO3 H3AsO3
Figure 1. Simulation of species distribution involving arsonite [As(III)]across pH at
an initial arsenicconcentration of 0.01 mole/liter.
Arsenate will more readily behave as an acid, with approximately equal molar
concentrations of H2AsO4- and HAsO4
2- near a pH of 7. The acid-base behavior of
arsenate may be represented as:
(3) H3AsO4 = H2AsO41-
+ H+ pKa1 = 2.3
(4) H2AsO4- = HAsO4
2- + H
+ pKa2 = 6.8
(5) HAsO42-
= AsO43-
+ H+ pKa3 = 11.6.
Simulation of arsenate speciation with a total arsenic concentration of 0.01 mole/liter
is displayed in Figure 2.
3 4 5 6 7 8 9 10 11 12
pH
0
0.001
0.002
0.003
0.004
0.005
0.006
0.007
0.008
0.009
0.01
Co
ncen
trati
on
(M
ole
/Lit
er)
H3AsO4H2AsO4
HAsO4AsO4
Figure 2. Simulation of species distribution involving arsenate [As(V]across pH at an
initial arsenic concentration of 0.01 mole/liter.
Arsenic In The Soil Environment: A Soil Chemistry Review 7
Dissociation constants for methylated species (referenced in Jackson and Miller,
2000) include 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).The simulated pH-
dependent speciation for MMA and DMA are displayed in Figure 3 and Figure 4,
respectively.
3 4 5 6 7 8 9 10 11 12
pH
0
0.001
0.002
0.003
0.004
0.005
0.006
0.007
0.008
0.009
0.01
Co
ncen
trati
on
(M
ole
/Lit
er)
H2-MMAH-MMAMMA
Figure 3. Simulation of species distribution involving monomethylarsonic acid or
MMA across pH at an initial arsenicconcentration of 0.01 mole/liter.
3 4 5 6 7 8 9 10 11 12
pH
0
0.001
0.002
0.003
0.004
0.005
0.006
0.007
0.008
0.009
0.01
Co
nc
en
tra
tio
n (
Mo
le/L
ite
r)
H-DMADMA
Figure 4. Simulation of species distribution involving dimethylarsinic acid or DMA
across pH at an initial arsenicconcentration of 0.01 mole/liter.
Arsenite and Arsenate as Oxidation and Reduction Species
Half-cell equations are readily obtained from the standard free energies of
formation(Wagman et al., 1982). For arsenate to arsenite reduction the following
reactions are valid:
8 Michael Aide and Donn Beighley and David Dunn
A predominance diagram (Figure 5) illustrates the relative stability regions for the
various protonated arsenite and arsenate species. Oxic, suboxic and anoxic
partitioning in the predominance diagram was patterned after Essington (2004).
3 4 5 6 7 8 9
pH
-10
-5
0
5
10
15
20
pe
O2-H2O
H2O-H2
Oxic
Suboxic
Anoxic
HAsO42-
As(OH)3
H2AsO4 1-
Figure 5. Predominance diagram showing zone of species predominance given pe and
pH as master variables.
The likelihood for arsenite formation in anoxic soil environments is more readily
favored in increasingly acidic soil environments. The protonation-deprotonation pH
transition for arsenate is near pH 7, showing that oxic soils support the arsenate
species. Using Minteqa2/Prodefa2, a geochemical assessment model for
environmental systems simulation (Allison et al., 1991), Aide (unpublished material)
employed drainage water concentrations (unpublished data) for background
concentrations of Ca2+
, Mg2+
, K+, Na
+, NH4
+, NO3
-, SO4
2-, H2PO4-HPO4, with the
CO2partial pressure set at atmospheric concentrations. The pH was set at pH 5.0 to
simulate arsenic speciation at pe levels of (i) 0 (Eh = 0.0 mv) for anoxic conditions,
(ii) 7 (Eh = 414 mv) for suboxic conditions, and (iii) 12 (Eh = 710 mv) for oxic
conditions. Ionic strength and activitiy coefficients were determined using the Davis
Equation, and the As(III)-As(V) oxidation-reduction couple was employed. For
simulated oxic conditions, H3AsO3 (3.87 x 10-23
mole/l) and H2AsO41-
(2.95 x 10-5
)
were the dominant species, with arsenate the predominant species. On transition from
oxic to suboxic and then to anoxic conditions, the arsenite species become
increasingly more dominant, culminating with the anoxic system having H2AsO31-
at
1.61 x 10-9
mole/l and H3AsO3 at 2.997 x 10-5
mole/l, with H2AsO41-
at 2.29 x 10-11
mole/l.
Hundal et al. (2013) documented arsenic in alluvial soils in Punjab (India) and
observed that soils amended with arsenic and incubated for selected time periods
under anaerobic conditions demonstrated arsenate reduction to arsenite. Arsenite was
Arsenic In The Soil Environment: A Soil Chemistry Review 9
observed to be a more mobile species than arsenate. Orpiment (As2S3) was proposed
as a potential arsenic precipitation product as was ferrous sulfite (FeS2) and vivianite
(Fe3(PO4)2 8H2O). The reductive dissolution of Fe-oxyhydroxides did not result in
increased arsenic solubility, a feature attributed to arsenic adsorption on remaining
Fe-oxyhydroxides.
Hu et al. (2013) performed rice field trials having aerobic, intermittent, conventional
and flood irrigation regimes. Acid extractable soil arsenic was shown to transition to
increasing concentrations from the more oxic aerobic irrigation regime (aerobic) to
the more anoxic irrigation regime (Flood), whereas cadmium demonstrated an
opposite trend. Rice cultivars did show differences in the arsenic and cadmium
concentrations, with hybrids typically have a greater arsenic accumulation.
Mn-oxyhydroxides exist in the soil environment and have been implicated in the
oxidation of arsenite to arsenate (He and Hering, 2009; Ackermann et al., 2010;
Bravin et al., 2008; Radu et al., 2007). Mn-oxyhydroxides are frequently present in
smaller abundances than Fe-oxyhydroxides; however, they are thermodynamically
favored to oxidize arsenite to arsenate. Birnessite (δ-MnO2), cryptomelane (α-MnO2),
and pyrolusite (β-MnO2) are three Mn-oxide polymorph forms. Pyrolusite typically
has a better ordered crystalline structure and a higher point of zero charge (pH=6.4),
thus is a better candidate for interaction with arsenic oxyanions. Arsenate oxidation
may be assumed to follow a stoichiometry given by Nesbitt et al. (1998):
MnO2 + H3AsO3 + 2H+ = Mn
2+ + H3AsO4 + H2O.
Ying et al. (2012) observed arsenic adsorption onto Fe and Mn oxides. Subsequently
Ying et al. (2013) performed a flow through reactor experiment involving arsenic
presorbed on sand grains coated with ferrihydrite and birnessite. These authors noted
that As, Mn and Fe migrated from the reduced aggregate interiors to the more aerated
exterior, where Mn-oxide formation was coupled with As(III) oxidation.
Radu et al. (2008) proposed that the first step in arsenite oxidation with MnO2
involves As(III) inner sphere surface complex formation with the displacement of
OH- and H2O as ligand substitution. Transfer of two electrons from As(III) results in
the release of Mn2+
and arsenate. The presence of small quantities of Mn3+
in the
original MnO2usually results in a molar ratio release of Mn/As equal to or exceeding
unity.
Arsenite oxidation to arsenate by birnessite was investigated to assess whether Fe2+
could assist in arsenate sequestration (He and Hering, 2009). These authors observed
that the absence of Fe2+
the arsenate production resulted in an arsenate concentration
increase in the aqueous phase; however, in the presence of Fe2+
the arsenate was
almost completely sequestered. Arsenite addition to soils has been shown to result in
the rapid oxidation of arsenite to arsenate, wherein both species may be adsorbed
(Manning and Suarez, 2000).
Arsenic oxidation and reduction may be strongly mediated by microbial populations
(Saltikov and Olson, 2002). Dissimilatory arsenate-reducing bacteria are able to
effectively reduce arsenate to arsenite by using arsenate as a terminal electron
acceptor in anoxic soils or sediments. A gram-negative dissimilatory arsenic-reducing
bacterium (Citrobacter sp. NC-1) was recently isolated from As-bearing soil that was
shown to rapidly and completely reduce arsenate solutions (Chang et al., 2012).
10 Michael Aide and Donn Beighley and David Dunn
Focardi et al. (2010) documented anaerobic bacterial strains capable of using arsenate
and sulfate as electron acceptors, yielding arsenic sulfide precipitates. Recent research
suggests that aquatic macrophyte roots harbor active populations of Fe reducers that
substantially influence carbon flow and the many aspects of biogeochemical cycling
(King and Garey, 1999). Yoshida et al. (2008) examined cylindrical soil nodules
composed of Fe-oxyhydroxides. They isolated microbial 16s rDNA associated with
Fe-oxidizing bacteria, suggesting that the pedogenic nodules were created along roots
whose rhizosphere harbored these Fe-oxidizing bacteria.
Gleization and Arsenic Availability
In the United States rice production relies on either (i) drill-seeded, delayed flood
regimes or (ii) permanent flood regimes. Permanent flood regimes use airplane
seeding of pre-germinated seed into standing water. Recently, center-pivot irrigation
systems and furrow irrigated systems are being evaluated for commercial rice
production. Both drill-seeded and delayed flood and permanent flood regimes result in
soils having a suboxic layer, typicallyone to several centimeters in soil later thickness,
superimposed on a deeper anoxic soil layer. With the onset of oxygen depletion and
with an energy source (CH2O), nitrate reduction proceeds via the denitrification
process:
5CH2O + 4NO3- = 5HCO3
- + 2N2 + 2H2O + H
+
As soils become increasingly anoxic, manganese oxide dissolution may proceed,
followed sequentially by Fe-oxyhydroxide dissolution:
CH2O + 4Fe(OH)3 + 7H+ = HCO3
- + 4Fe
2+ + 10H2O
Arsenic may be co-precipitated with Fe-oxyhydroxides and the reductive dissolution
of these Fe-oxyhydroxides may promote the release of arsenate:
CH2O + 4Fe(OH)3(H2AsO4-) + 7H
+ = HCO3
- +4Fe
2+ + 4H2AsO4
- + 10H2O
The arsenate may be subsequently reduced to arsenite:
CH2O + 2H2AsO4- + H
+ = HCO3
- + 2H3AsO3.
Thus, the natural soil process of gleization and its acceleration in suboxic to anoxic
rice culture acts to convert arsenic from a relatively low state of bioavailability to an
easily plant assimilated species. Herbal and Fendorf (2006) formed Fe-oxyhydroxides
via microbial reduction and documented that released arsenic may be re-sequestered
by these newly synthesis reduction products. Similarly Pedersen et al. (2006)
documented that As(V) was strongly sorbed onto Fe(II) catalyzed formation of
ferrihydrite and lepidocrocite formation and that arsenic was rendered non-labile on
transition of these reduction products to crystalline Fe(III) oxides.
Microbially Assisted Arsenic Oxidation and Reduction Processes
The microbial populations in soil and water resources likely account for the majority
of life forms on planet earth. Prokaryotic organisms in the soil environment project a
profound influence on the type of soil processes that are viable, both in terms of
which processes may readily proceed and their reaction kinetics (Cullen and Reimer,
1989; Meng et al., 2003; Campbell et al., 2006; Blodau et al., 2008; Tufano et al.,
2008; Fendorf et al., 2010).
Arsenic In The Soil Environment: A Soil Chemistry Review 11
Microbial Inter- and Extracellular Bonding (Sequestration) Arsenic does not have a necessary metabolic or nutrient role in microorganisms
(Huang (2014), thus no specific arsenic uptake pathway has been identified. Arsenic
uptake appears to occur using existing transporting systems evolved primarily for
other chemical species, most likely phosphate. Prokaryotic organisms have evolved
metabolic strategies to either exclude arsenic or bind arsenic within the cell.
The extracellular polymeric substance surrounding many prokaryotic species is
composed of anionic functional groups (uronic acids and proteins) and cationic
functional groups (amino sugars) that chemically function as molecular sieves. Huang
(2014) identified recent research suggesting a complex chemistry involving the
bacterial matrix and cell wall constituents with arsenic species which reduces the
intercellular arsenic uptake potential (Illustration 1). Arsenate uptake by the
prokaryotic cell may simply be followed by arsenate efflux from the cell or arsenate is
rapidly reduced to arsenite (detoxification reduction). Arsenite concentrations in the
prokaryotic cell may be protein bound to reduce their metabolic influence
(intercellular sequestration) or oxidized to arsenate and ejected from the cell
(respiratory oxidation). Arsenic species may be methylated ((CH3)xAsH(3-x)) and
ejected from the cell as either gaseous or solubilized species (Huang, 2014).
An emerging area of research involves prokaryotic organism genetic isolation or
genetic engineering to support arsenic methylation. Chen et al. (2013) documented the
potential of Pseudomonas putida to support arsenic methylation. Methyl arsenic
species formed by bacteria having an arsM gene may result in the transfer of gaseous
methyl arsenic species or the diffusion of these methyl arsenic species into an aqueous
phase (Jia et al., 2013). Engineered bacterial cells having an ArsR metalloregulatory
protein are selective towards As(III) and accumulate arsenic. Arsenic dissolution may
proceed in the presence of Geobacter sp OR-1 bacteria possessing a respiratory As(V)
reductase gene (Ohtsuka et al., 2013). Huang and Metzner (2007) observed that
dimethylarsenic acid shows an order of magnitude greater membrane permittivity than
monomethylarsonic acid, thus potentially explaining the greater presence of
dimethylarsenic acid in natural catchment waters. Srivastava et al. (2011)
demonstrated that the soil fungi Rhizopus sp, Trichoderma sp, and Neocosmospora
support arsenic bio-volatilization.
Detoxification Reduction
As (V) to As(III)
Methylation As(III) to Me-As(III)and Intracellular Sequestration
Bacterial Cell Wall and Membrane
Volatilization and Dissolved Methyl As Species
As(VI) and As(III) sorption in Extracellular Polymeric Substance
EffluxTransport
Respiratory oxidation and Reduction Involving As(V) and As(III)
Illustration 1. The overview of microbial arsenic interactions.
12 Michael Aide and Donn Beighley and David Dunn
Microbial Arsenic Oxidation, Reduction and Methylation Reactions Major arsenic microbial soil transformations include oxidation reduction and
methylation-demethylation reactions. Ajith et al. (2013) proposed that microbial Mn
oxidation and reduction reactions support arsenic mobilization involving pathways
similar to those proposed for Fe-oxyhydroxides. Iron(II) activated goethite inhibited
arsenate reduction and arsenite oxidation to arsenate was augmented by the presence
of ferrous ions. Lafferty and Loeppert (2005) proposed that arsenic mobility is species
dependent and the proposed mobility order was methyl-As(III) >> methyl-As(V) >
As(III) > As(V). Dixit and Hering (2003) noted that arsenic mobility was limited by
arsenite to arsenate oxidation.
Kocar at al. (2006) investigated the dissimilatory reduction of Fe(III) and As(V) and
reported that Fe(III) inhibited As sulfide sequestration. Similarly Saalfield and
Bostick (2009) showed that iron reduction in the presence of sulfide promoted
magnetite, elemental sulfur and trace amounts of Fe- and As-sulfides that limited
arsenic mobility. Sun et al. (2009) reported on an arsenic remediation strategy
involving nitrate, wherein Fe(II) and As(III) in anoxic soil conditions yielded As(V)
adsorbed onto synthesized Fe(III)-oxyhydroxides.Gibney and Nusslein (2007)
observed that arsenic(V) reduction to arsenic(III) did not substantially proceed until
the available nitrate pool was depleted via denitrification processes.
Arsenite and Arsenate Precipitation In the soil environment Fe-As precipitates include: scorodite (FeAsO4●2H2O),
phenmacosiderite (Fe4(AsO4)3(OH)3●6H2O) and parasymplesite (Fe3(AsO4)2●8H2O).
Calcium arsenate precipitation reactions yielding raventhalite (Ca3(PO4)2● 10H2O)
are similar to precipitation reactions involving calcium and phosphate to form calcium
hydroxyapatite or octacalcium phosphate (Dungkaew et al., 2012). Gemeinhardt et al.
(2006) focused on enhancing arsenic retention in contaminated soils using FeSO4
amendments. Sequential extractions indicated that arsenic was immobilized into the
Fe-oxyhydroxide (non-crystalline and crystalline forms) and residual fractions. The
authors speculated that arsenic immobilization was attributed to the (i) precipitation of
FeAs (scorodite), (ii) inner sphere adsorption onto Fe-oxyhydroxides, or (iii) co-
precipitation and occlusion with newly synthesized Fe-oxyhydroxides.
Parisio et al. (2006) in New York documented that reduced Fe leachates having
arsenic originating from solid waste landfills frequently showed a “severe effects
level” towards aquatic life. Upon entering oxygenated environments the formation of
iron precipitates (Fe-floc) were shown to be enriched in arsenic.
The Surface Adsorption Process
Surface adsorption is a chemical process involving the accumulation of dissolved
substrate (adsorbate) at the interface of a solid (adsorbent). In the purest context, the
adsorbate is distinguished by its discrete positioning with respect to specific sites on
the adsorbent and its dimension normal to the interface is that of the dimensions of the
adsorbate. Conversely, precipitation is the three-dimensional growth of a structure.
Frequently, investigators may not be able to distinguish adsorption from precipitation
and the characteristic term “sorption” is applied to the process.
Arsenic In The Soil Environment: A Soil Chemistry Review 13
Charge creation on the adsorbent may be permanent charge or pH-dependent charge.
Permanent charge typically arises from isomorphic substitution and therefore is a
characteristic of the adsorbent. Phylosilicates typically are the more abundant
examples in the soil environment, especially 2:1 layer silicates; such as,
montmorillonite, illite, vermiculite, beidellite, and nontronite, Adsorbents having pH-
dependent charge (variable charge surfaces) typically possess surface hydroxyl groups
[≡SOH, where S is a metal] that may experience protonation or deprotonation
reactions. Examples of soil solids having pH-dependent charge capabilities include:
edge regions of phyllosilicates, crystalline and noncrystalline (amorphous) metal
oxides, metal hydroxides and metal oxyhydroxides. The reactivity of the surface
hydroxyl groups is primarily a function of the structural metal atoms bound to the
surface hydroxyl groups, the valence and coordination number of the structural metal
atoms.
As an example, surface hydroxyl groups in goethite are characterized as Type A,
Type B, and Type C: where,.
Type A: ≡FeOH-0.5
(terminal, where hydroxyl is directly coordinated to Fe),
Type B: ≡Fe3OH+0.5
, (surface hydroxyl is shared with three Fe)
Type C: ≡Fe2OH0 (surface hydroxyl is shared with two Fe).
Note that the charge density of the two and three Fe atoms in Type B and Type C
surfaces would not permit protonation of the surface hydroxyl groups.Stoichiometry
of the surface groups may be expressed in two ways, depending on the theoretical
approach of the investigator. The more classical method, with goethite as an example,
involves:
≡FeOH0 + H
+ = ≡FeOH2
+ K1 = 10
+6.2
≡FeOH0 = ≡FeO
- + H
+ K2 = 10
-11.8
Simulated protonated and deprotonated Fe-site distributions are illustrated in Figure 7.
The second approach uses the electrostatic valence principle, which involves the bond
strength (s) as determined by the cation charge (+3) divided by the coordination
number (+6): s=0.5. Given that the Fe atom has +0.5 units of charge directed along 6
coordination directions, the surface charge is determined as:
≡FeOH2+0.5
= ≡FeOH-0.5
+ H+, where K =10
-8.5.
This formalization permits the surface site acidity to be a function of metal
coordination and valency (M-O bond strength), the bond strength to bond length ratio,
and the metal’s electronegativity. Simulated protonated and deprotonated Fe-site
distributions for the electrostatic valence principle are illustrated in Figure 6.
2 4 6 8 10 12 14
pH
0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
Fe-S
peci
es/T
otal
Fe-
Site
s
FeOHFeOH2FeO
Figure 6. Simulated protonation-deprotonate Fe-sites on goethite using a classical
approach.
14 Michael Aide and Donn Beighley and David Dunn
2 4 6 8 10 12 14
pH
0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
Fe-
Sp
ecie
s /T
ota
l Fe-
Sit
es
Fe(OH2) +0.5Fe(OH) -0.5
Figure 7. Simulated protonation-deprotonate Fe-sites on goethite using the
electrostatic valency principle.
Ligand exchangeadsorption mechanisms involve direct coordination of the adsorbate,
with expulsion of either OH- or H2O from the adsorbent. The process may be
expressed as:
≡Metal-OH2+0.5
+ H2PO41-
= ≡Metal-OPO3H-0.5
+ H2O.
Adsorption of Arsenite and Arsenate Species
The arsenite and arsenate species experience pH-dependent adsorption and co-
precipitation on/with Fe-oxyhydroxides, most notably ferrihydrate (β-FeOOH),
lepidocrocite (γ-FeOOH), goethite (α-FeOOH), and hematite (Fe2O3) (Miretzky and
Cirelli, 2010). Surface protonation of goethite may be represented as(Luxton et al.,
2006):
≡Fe(OH)-0.5
+ H+ ↔ ≡Fe(OH2)
+0.5 pK = -9.59.
Surface protonation allows the interface to acquire sufficient positive surface charge
densities to promote arsenite adsorption. Arsenite adsorption may be either
monodentate or bidentate:
Monodentate:
≡Fe(OH)-0.5
+ As(OH)3 ↔ ≡FeOAsO2-2.5
+ H2O + 2H+ pK= +11.35
≡Fe(OH)-0.5
+ As(OH)3 ↔ ≡FeOAsO(OH)-1.5
+ H2O + H+ pK= +0.85
≡Fe(OH)-0.5
+ As(OH)3 ↔ ≡FeOAs(OH)2-0.5
+ H2O
pK= -5.24
Bidentate
2≡Fe(OH)-0.5
+ As(OH)3 ↔ ≡(FeO)2As(OH)-1
+ H2O pK=-15.47
Arsenate adsorption on Fe-oxyhydroxides may be represented as:
≡Fe(OH)-0.5
+ H2AsO4-↔ ≡FeHAsO4
-0.5 + H2O
Arsenic In The Soil Environment: A Soil Chemistry Review 15
The optimal pH for the adsorption of arsenite on Al- and Fe-oxyhydroxides ranges
from pH 7 to 10, depending on experimental protocols and the presence of phosphate,
silicic acid, naturally occurring organic acids, and other competing anions (Goldberg,
2002; Luxton et al., 2006; Grafe et al, 2001 and 2002; Pigna et al., 2006; Zhang and
Selim, 2008), whereas the optimal pH for the adsorption of arsenate on Al- and Fe-
oxyhydroxides varies across the pH range of 4 to 7 (Bowell, 1994; Jain and
Loepperts, 2000; Vetterlein et al., 2007; Saeki, 2008; Xu et al., 2008; Toor and
Haggard, 2009; Jackson and Miller, 2008; Smith et al., 2002 and 2009). In general,
arsenate adsorption increases in acidic media, reaches maximal adsorption at pH in
the range of 3 to 7; whereas arsenite adsorption reaches maximal levels in the pH
range of 7 to 8 (Jain and Loeppert, 2000; Goldberg, 2002). Arsenate typically exhibits
greater adsorption than arsenite in acidic to neutral pH levels; whereas arsenite may
exhibit greater adsorption in alkaline environments. Conversely, Oscarson et al.
(1983) observed that arsenite is adsorbed to a greater extent than arsenate on
noncrystalline Fe-oxyhydroxides. Jain and Loeppert (2000) showed arsenate
adsoption on ferrihydrate was greatest in acidic media and arsenite adsorption was
maximized at higher pH levels. Quaghebeur et al. (2005) performed both batch and
flow-through desorption on columns packed with kaolinite to compare their ability to
desorb arsenate. Not surprisingly, the batch protocol underestimated arsenate
desorption.
Goldberg (2002) noted that arsenate and arsenite adsorption on reference
phyllosilicates (kaolinite, illite and montmorillonite) was pH-dependent with arsenate
adsorption less evident on transition to alkaline media, whereas arsenite adsorption
exhibited less pH dependency with moderate adsorption increases on transition from
acidic to alkaline media.Goldberg (2002) also showed that arsenate adsorption on
amorphous aluminum oxides and amorphous iron oxides was relatively constant from
acidic environments to near pH levels exceeding pH 9; afterwhich, arsenate
adsorption declined. Arsenite adsorption showed a maximum adsorption near pH 8 for
amorphous aluminum oxides and exhibited little pH dependence on amorphous Fe-
oxides.
The conversion of arsenite to arsenate in the presence of Fe-oxyhydroxides is easily
shown to be thermodynamically favorable from the application of free energies of
formation. Jackson and Miller (2000) investigated arsenate adsorption on
noncrystalline Fe-oxyhydroxides with and without the presence of goethite. Using
oxalate and hydroxylamine as selective digestive agents for the noncrystalline Fe-
oxyhydroxides, they observed that in the absence of goethite arsenate concentration
were greater after Fe-oxyhydroxide dissolution, whereas in the presence of goethite
the arsenate appeared to be re-adsorbed onto the goethite and the arsenate
concentrations did not appreciably increase. They also reported that phosphate
addition inhibited the adsorption of arsenate onto goethite.Cornu et al. (2003)
observed the adsorption of arsenate onto kaolinite and also kaolinite pretreated with
humic acids. Their observations also involved contrasting calcium and sodium nitrate
ionic strength electrolytes. In this study the sodium adjusted media exhibited an
arsenate pH dependency with both kaolinite and humic acid coated kaolinite, with
kaolinite having greater arsenate adsorption, especially in acidic media. Conversely,
16 Michael Aide and Donn Beighley and David Dunn
in calcium adjusted media, kaolinite did not exhibit any arsenate pH-dependent
adsorption, whereas humic acid coated arsenate showed only an incident pH-
dependency.
Jackson and Miller (2000) evaluated various concentrations of phosphate (pH 3 and 7)
to extract arsenite, arsenate, dimethylarsinic acid, monoethylarsonic acid, arsanilic
acid and roxarsonesorbed onto goethite and amorphous Fe-oxides. Phosphate was
demonstrated to displace arsenite and arsenate. In the presence of goethite, a portion
of the arsenite was oxidized to arsenate. Khaodhiar et al. (2000) prepared iron oxide
coated sand (Fe2O3) and used completely mixed batch reactors to establish isotherm
adsorption curves involving arsenate and arsenate-chromate systems. Arsenate was
established to form inner-sphere complexes with the Fe-oxides, noting that the
concentration of the NaNO3 electrolyte had little influence on arsenate adsorption.
Arsenate adsorption was strongly adsorbed at acidic to slightly acidic pH values and
adsorption decreased with increasing pH. The equilibrium sorption constants for
arsenate on the iron oxide coated sand were:
SOH + AsO43-
+ 3 H+ = SH2AsO4 + H2O pK = -26.26
SOH + AsO43-
+ 2 H+ = SHAsO41-
+ H2O pK = -20.22
SOH + AsO43-
= SOHAsO43-
+ H2O pK = -7.48
Grafe et al. (2001) investigated arsenite and arsenate adsorption on goethite spanning
pH range of 3 to 11. These authors showed that arsenate adsoption was greatest at pH
3 and 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. Sun and Doner, (1996, 1998) reported that
arsenite adsorption is inner sphere based on Fourier transform infrared spectroscopy.
Arsenic and the Rhizosphere in Anaerobic Soils
Aquatic plants, including rice, produce aerenchyma in the root cortex that facilitates
oxygen diffusion towards the root tips (Wu et al.,2011). Oxygen diffusion in rice
aerenchyma supports a partial oxidation of a thin soil layer (rhizosphere) adjacent to
the root surface, permitting radial oxygen diffusion towards the bulk soil. Wang et al.
(2013) observed Fe-oxyhydroxide synthesis within and extending from the root
epidermis into the rhizosphere. The synthesis of Fe-oxyhydroxides (Fe-plaque) has
the potential to sequester metal(loid)s via adsorption and/or precipitation reactions.
Wang et al (2013) documented that the concentrations of iron, cadmium and arsenic
associated with Fe-plaque increased with greater radial oxygen loss activities. Chen et
al. (2005) observed that Fe-plaque was selective for the preferential adsorption of
arsenite relative to arsenate. Wu et al. (2013) employed a greenhouse experimental
design with rice cultivars and induced arsenic stress attributed to soil additions of
arsenate. Arsenic stress produced a slight increase in Fe-plaque formation; however,
the radial oxygen loss rates declined. With increasing arsenic stress seed DMA
concentrations increased as a percentage of total arsenic whereas relative and absolute
inorganic arsenic concentrations increase in stem and root tissues. Wu et al. (2011)
Arsenic In The Soil Environment: A Soil Chemistry Review 17
cultured 20 rice genotypes with arsenic bearing irrigation water, documenting
differences in aerenchyma, arsenic accumulation and speciation. As expected,
genotypes differed in aerenchyma, which was correlated with radial oxygen loss.
Radial oxygen loss was inversely related to total and inorganic arsenic accumulation,
suggesting that varietal selection may be important in reducing arsenic accumulation.
Adsorption of Arsenite and Arsenate Species in the Presence of Competing
Anions
Competing anions may either: (i) previously or preferentially attach to a potential
binding site and negate access for arsenic bonding, or (ii) displace arsenic by
competitive competition for the binding site. The most widely investigated counter
anions include: phosphate, sulfate, carbonate, silicic acid, dissolved organic materials
(DOM). Sulfate, carbonate and DOM have been shown to be relatively less effective
in displacing arsenic (Jain and Loeppert, 2000). Weng et al. (2009) reported that
adsorbed soil organic matter facilitated arsenic sorption. Using goethite as the
bonding surface, Luxton et al. (2005) showed that silicic acid (H4SiO4) was able to
effectively displace arsenite by forming an inner sphere complex involving silicic acid
and goethite by ligand exchange with hydroxyl functional groups. Swedlund and
Webster (1999) and also Meng et al. (2000) demonstrated that H4SiO4 may displace
arsenite from ferrihydrate.
In Australia, Smith et al. (1999) surveyed ten well-drained soils, observing that after
arsenic amendments, arsenate adsorption was greater than arsenite. They further
observed that phosphate inhibited arsenate adsorption. In a subsequent study, Smith et
al. (2002) studied the influence of phosphate, calcium and sodium on arsenate
adsorption in Alfisols, Vertisols and Oxisols, They reported that phosphate inhibited
arsenate adsorption in soils testing low in Fe-oxyhydroxides, whereas phosphate did
not influence arsenate adsorption in soils having higher Fe-oxyhydroxide abundances.
They attributed the phosphate inhibition on arsenate adsorption in soils having smaller
Fe-oxyhydroxide abundances as being due to competition for reactive sites. Smith and
Naidu (2009) using flow through reactors demonstrated that arsenate adsorption was
initially rapid, having 58 to 91 percent of the arsenate adsorbed within 15 minutes.
Subsequent adsorption continued as a slower rate of adsorption. These authors also
suggested that arsenate adsorption was inhibited by phosphate. In Australia, Smith et
al. (1999) surveyed ten well-drained soils, observing that after arsenic amendments,
arsenate adsorption was greater than arsenite. They further observed that phosphate
inhibited arsenate adsorption.
Liu et al. (2014) investigated silicic acid (H4SiO4) additions to soils in greenhouse
conditions and cultured to two rice cultivars. Silicic acid additions increased
dimethylarsinic acid (DMA) solubility, whereas inorganic arsenic solubility was only
marginally increased. Silicic acid was proposed to inhibit DMA adsorption or support
DMA desorption. Silicic acid additions promoted DMA concentrations in rice stem,
leaf and seed, reinforcing a concept that DMA concentrations in rice plant
components are entirely attributed to soil chemistry influences and plant physiology is
not involved in arsenic methylation. Swedlund and Webster (1999) demonstrated that
18 Michael Aide and Donn Beighley and David Dunn
silicic acid inhibits arsenate/arsenite adsorption on ferrihydrate. Luxton et al. (2006)
observed that silicic acid inhibits arsenite adsorption on goethite.
Spectroscopic Investigations Involving Arsenite and Arsenate Bonding
Mechanisms
Arsenite and arsenate adsorption may result in both monodentate and bidentate
bonding structures (Fendorf et al., 2004). Fendorf et al. (1997) and Waychunas et al.
(1993) employed extended x-ray absorption fine structure spectroscopy to document
that arsenate adsorption on goethite and synthetic Fe-oxyhydroxides is inner sphere.
Mononuclear and binuclear bridging has been also proposed for arsenate adsorption
reactions (Miretzky and Cirelli, 2010; Jackson and Miller, 2000; Fendorf et al., 2004;
Goldberg, 2002).
Rice Irrigation and Arsenic Availability
Traditionally, United States rice production relies on either (i) drill-seeded, delayed
flood irrigation or (ii) permanent flood irrigation. Permanent flood regimes use
airplane seeding of pre-germinated seed into standing water. Recently, center-pivot
irrigation systems and furrow irrigated systems are being evaluated for commercial
rice production. Both delayed flood and permanent flood irrigation result having a
suboxic soil layer, typically having a thickness of several centimeters, superimposed
on a deeper anoxic soil layer. Thus the suboxic soil layer may have a less negative
redox potential that may permit nitrification. With the onset of oxygen depletion and
with an energy source, subsequent nitrate reduction could proceed via the
denitrification process.
In China Hu et al. (2013) optimized water management to lower rice uptake of both
cadmium and arsenic without appreciable yield loss. Using flooded, conventional,
intermittent and aerobic irrigation strategies, these authors observed that the
conventional and flood irrigations increased arsenic and decreased cadmium uptake
compared to the intermittent and aerobic irrigation strategies. These authors further
suggested that maintaining flood until full tillering stage, then switching to
intermittent irrigation was the best compromise for maintaining yield and reducing
both arsenic and cadmium uptake. The mean arsenic concentrations were 0.20 to 0.28
mg As / kg-dry weight for the aerobic and intermittent irrigation strategies, whereas
the conventional irrigation (0.27 to 0.3 mg As / kg-dry weight) and flood irrigation
(0.3 to 0.48 mg As / kg-dry weight) demonstrated greater arsenic uptake.
Lombi et al. (2009) estimated total and speciated arsenic concentrations in husk, bran
and endosperm from rice cultured in contaminated Chinese soil. Total arsenic was
0.54 mg – As/ kg (husk), 6.24 mg As / kg (bran) and 12.42 mg As / kg (endosperm).
Inorganic arsenic constituted 74% of the arsenic in the husk, 75% of the arsenic in the
bran and 54% of the arsenic in the endosperm. Liang et al. (2010) also investigated
total and speciated arsenic concentrations in Chinese rice. Using numerous rice
samples the total arsenic concentrations ranges from 65.3 to 274.2 µg As / kg, with a
mean of 114.4 µg As / kg.
In Taiwan, Syu et al. (2014) investigated arsenic uptake among rice genotypes
cultivated in arsenic contaminated soils. Iron plaque associated with the rice roots
Arsenic In The Soil Environment: A Soil Chemistry Review 19
sequestered arsenic; however, no correlation between the quantity of arsenic
associated with the iron plaque and rice tissue arsenic concentrations were observed.
Liu et al. (2005) using hydroponics observed that iron plaque showed an adsorption
affinity for arsenate relative to arsenite. Using a microcosm approach, Chen et al.
(2008a) observed that nitrate reduced the concentrations of ferrous iron, whereas
ammonium enhanced ferric iron reduction. These authors concluded that nitrate
decreased iron plaque formation, resulting in increased arsenic uptake by rice.
Hossain et al. (2009) observed that ferrous iron reduced arsenic concentrations in rice
grain and increased rice yields. Addition of phosphate increased arsenic uptake. In
China, Liu et al. (2004) established iron plaque on rice seedling roots in a solution
culture experiment, observing that 75 to 89 percent of the total arsenic was
sequestered by the iron plaque. Arsenic root concentrations were equivalent among
the rice genotypes; however, rice shoots showed significant arsenic concentration
variation. Huang et el. (2012) employed a microcosm experiment and documented the
rate of iron plaque degradation post-harvest, noting at 76 percent of the arsenic
sequestered by the iron plaque was released to the soil solution in 27 days.
Areas for Productive Future Research
Alternative irrigation strategies to reduced arsenic accumulation in rice
Rice breeding programs to select cultivars that limit arsenic accumulation
Soil biology investigations to quantify the microbial roles involved in arsenic
speciation
Global initiatives to quantify arsenic concentrations in drinking and irrigation
waters
Quantitative modelling of the arsenic pathways in soil
Understanding the formation and effectiveness of Fe-plaque on arsenic uptake
by aquatic plants
Routine soil tests to predict arsenic plant availability.
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