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Review article
Arsenic accumulation in rice: Consequences of rice genotypes and
management practices to reduce human health risk
Shofiqul Islam
a,b,c
, Mohammad Mahmudur Rahman
a,b
, M.R. Islam
c
,RaviNaidu
a,b,
⁎
a
Global Centre for Environmental Remediation (GCER), Faculty of Science and Information Technology, The University of Newcastle, Callaghan, NSW 2308, Australia
b
Cooperative Research Centre for Contamination Assessment and Remediation of the Environment (CRC CARE), The University of Newcastle, Callaghan, NSW 2308, Australia
c
Department of Soil Science, Bangladesh Agricultural University, Mymensingh 2202, Bangladesh
abstractarticle info
Article history:
Received 10 June 2016
Received in revised form 16 August 2016
Accepted 7 September 2016
Available online xxxx
Rice is an essential staple food and feeds over half of the world's population. Consumption of rice has increased
from limited intake in Western countries some 50 years ago to major dietary intake now. Rice consumption rep-
resents a major route for inorganic arsenic (As) exposure in many countries, especially for people with a large
proportion of rice in theirdaily diet as much as 60%. Rice plants are more efficient in assimilating As into its grains
than other cereal crops and the accumulationmay also adversely affect the quality of riceand their nutrition. Rice
is generally grown as a lowland crop in flooded soils under reducing conditions. Under these conditions thebio-
availability of As is greatly enhanced leading to excessive As bioaccumulation compared to that under oxidizing
upland conditions. Inorganic As species are carcinogenic to humans and even at low levels in the dietpose a con-
siderable risk to humans. There is a substantial genetic variation among the rice genotypes in grain-As accumu-
lation as wellas speciation. Identifying the extent of genetic variation in grain-As concentration and speciation of
As compounds are crucial to determiningthe rice varieties whichaccumulate low inorganic As. Varietalselection,
irrigation water management, use of fertilizer and soil amendments, cooking practices etc. play a vital role in re-
ducing As exposure from rice grains. In the meantime assessing the bioavailability of As from rice is crucial toun-
derstanding human health exposureand reducing the risk.
© 2016 Elsevier Ltd. All rights reserved.
Keywords:
Arsenic
Rice genotypes
Bioaccumulation of As
Management practices for mitigation
Human health risk
Contents
1. Introduction.............................................................. 140
2. Arsenicandriceplants......................................................... 141
2.1. SourcesandtranslocationofAsinriceplants............................................ 141
2.2. Arsenicphytotoxicitytoriceplants ................................................ 142
2.3. Bio-accumulationofAsinriceplants................................................ 143
2.4. Arsenicbio-accumulationandricegenotypes............................................ 144
2.5. Arsenicspeciationandricegenotypes............................................... 145
3. ManagementpracticestocontrolAsbio-accumulation........................................... 145
3.1. Fertilizersandothersoilamendments............................................... 145
3.2. Watermanagement....................................................... 147
4. HumanhealthexposuretoAsthroughrice................................................ 147
4.1. RiceconsumptionandexposuretoAs............................................... 147
4.2. RiceprocessingandtheireffectonAscontents........................................... 148
4.3. BioavailabilityandbioaccessibilityofAsfromricegrains....................................... 149
4.4. Associatedhealthriskfromtheingestionofrice........................................... 150
5. Concludingremarks........................................................... 150
Acknowledgements............................................................. 151
References................................................................. 151
Environment International 96 (2016) 139–155
⁎Corresponding author at: Global Centre for Environmental Remediation (GCER), Faculty of Science and Information Technology, The University of Newcastle, Callaghan, NSW 2308,
Australia.
E-mail addresses: ravi.naidu@crccare.com,ravi.naidu@newcastle.edu.au (R. Naidu).
http://dx.doi.org/10.1016/j.envint.2016.09.006
0160-4120/© 2016 Elsevier Ltd. All rights reserved.
Contents lists available at ScienceDirect
Environment International
journal homepage: www.elsevier.com/locate/envint
1. Introduction
Arsenic (As) is one of the most challenging environmental problems
in drinking water as well as food crops especially in rice which affects
millions of people worldwide viz. in Bengal Delta, China, Taiwan, parts
of South America and South East Asia (Naidu et al., 2006; Nordstrom,
2002; Smedley and Kinniburgh, 2002). A significant number of popula-
tions rely upon groundwater for its drinking water supply as well as ir-
rigation water supply for growing agricultural crops especially paddy
rice and consequently exhibit extreme health effects of chronic As expo-
sure (Bhattacharyya et al., 2003b). Arsenic in rice is a serious concern for
N3 billion people across the world who consume rice as a staple food
and millions of people may be at risk of developing As-related health
problems (Meharg and Zhao, 2012). Fig. 1 shows the global rice produc-
tion and percapita consumption inselected countries. Arsenic contents
in rice varies widely, with most reported concentrations found in the
range 0.0 2 to 0.90 mg kg
−1
(Meharg and Zhao, 2012). Recent studies in-
dicated that rice genotypes have wide variations in total grain-As con-
centrations and As speciation around the world (Al Rmalli et al., 2005;
Huq et al., 2006; Meharg et al., 2009; Rahman et al., 2011; Rahman et
al., 2009; Torres-Escribano et al., 2008; Williams et al., 2007b; Zavala
and Duxbury, 2008; Zhu et al., 2008). Total As content in rice grains
from Bangladesh varies from between 0.058 and 1.835 mg kg
−1
(Meharg and Rahman, 2003). While this initial study shows that the
As in rice grains can be as high as 1.835 mg kg
−1
, recent studies show
that As in rice grains was usually below 1.0 mg kg
−1
(Rahman et al.,
2009). Moreover, rice has been identified as an important source of in-
organic As which may vary from 10% to 90% of total As (Williams et al.,
2005; Williams et al., 2007c). Rice contains about ten times more As
than other crops, and rice produced in regions such as Bangladesh,
India, China and the US, often contains even higher levels of inorganic
As (Meharg et al., 2008; Williams et al., 2005, 2007a).
Bioaccumulation of As by rice plantshave been linked to a number of
soil and environmental factors together with the nature of rice species.
A number of field studies have shown substantial genetic variation
among the rice genotype in grain-As concentration as well as As
speciation. In rice, there are three interacting loci on chromosome 6
and chromosome10 which maintain the genetic variation to regulate
As tolerance (Dasgupta et al., 2004). Other factors include soil-As
loading, water management practices, nutrient management in soils
etc. Accumulation of As from paddy soils and irrigation water poses a
potential health risk to humans especially inorganic As species. They
are classified as human carcinogens (IARC, 2004b) and several epidemi-
ological studies confirmed the relationship between As exposure via
drinking water and various health effects including skin and kidney
disease, heart disease, diabetes mellitus,neurological, respiratory com-
plications and gall bladder and lung cancers (Chen et al., 2015;
Hopenhayn-Rich et al., 1996; Islam et al., 2015b; Sommella et al.,
2013). Recent studies show that phosphate fertilizer is a major source
of As in areas affected with chronic kidney disease in Sri Lanka
(Jayasumana et al., 2014, 2015).
The association between a low level of As exposure and human
health effects still needs to be confirmed. Given the many case examples
of adverse impacts of As, there is a need to extend the existing studies
that focus on an assessment of the presence of As in rice grains to the
bioavailability of As from rice with a view to understanding their expo-
sure and risk. A number of review articles were published on different
aspects of Asand rice (challenges of As in soil plant systems, geograph-
ical variation on As accumulation, management practices etc.). For this
review we indexed articles in the Web of Science between 1956 and
2016 related to the keywords, As and rice, As and their phytotoxicity
in rice, As and rice genotype, As management in paddy fields, water
management and rice grain As, fertilizer management grain As in rice,
As and human health risk from rice, bioavailability of As from rice, and
cancer risk from ingestion of rice and retrieved them to obtain the max-
imum number of relevant articles (Fig. 2). There was no language re-
striction and we manually reviewed the references from the original
research articles. We aimed to identify all related studies in our topics.
The exclusion criteria we followed were: publications containing no
original research (reviews, editorials, proceedings, non-research letters
etc.), research related to performance of instruments, research related
to As chemistry other than rice, research lacks of data on As exposure
from rice and any duplications of the searched research articles in differ-
ent categories. The main focus of this manuscript is that it deals with As
and rice genotypes, growth and yield vs As contamination in irrigation
water and soils, and how rice genotypes control As uptake and their
speciation. It also discusses the major management practices that min-
imize As uptake by rice plants. Additionally, we also discuss about rice
consumption and risk of As exposure to human health in the context
of bioavailability and incremental lifetime cancer risk.
Fig. 1. The top rice producing and consuming countries of the world.
(data from USDA world riceproduction and consumption: https://www.worldriceproduction.com)
140 S. Islam et al. / Environment International 96 (2016) 139–155
2. Arsenic and rice plants
2.1. Sources and translocation of As in rice plants
Arsenic contamination occurs naturally in alluvial and deltaic sedi-
ments, as well as volcanic rocks andthermal springs and the weathering
of such deposits can lead to mobilization of As (Nordstrom, 2002;
Smedley and Kinniburgh, 2002; Welch et al., 2000). There are some an-
thropogenic sources of As that are released in the environment through
mining, pesticide application, wood preservation, combustion of coal
etc. Elevated concentrations of As are often derived from natural
sources. Bangladesh is an example of As occurring naturally in alluvial
sediments that is being mobilized into the groundwater. The natural
sources of As in paddy fields are biogeochemical processes. The anthro-
pogenic sources are from the use of As-contaminated irrigation water,
mining activity, use of arsenical pesticides and fertilizers
(Bhattacharyya et al., 2003a; Smedley and Kinniburgh, 2002). In pore
water the concentration of As and their speciation are governed by
soil redox potential, soil pH, soil organic matter status, nutrient concen-
tration especially silicon (Si) and phosphorus (P), iron oxide and clay
mineral contents (Bogdan and Schenk, 2009; Dixit and Hering, 2003;
Smedley and Kinniburgh, 2002; Williams et al., 2011).
Rice is grown as a lowland crop in flooded paddy soils under re-
ducing conditions where As availability is higher than under oxidiz-
ing conditions. In the flooded soil condition the mobility of arsenite
[As(III)] is higher than arsenate [As(V)] due to the reductive dissolu-
tion of iron oxide or hydroxides (Takahashi et al., 2004). This reduc-
tive mobilization of As under the anaerobic conditions greatly
enhances the bioavailability of As to rice, leading to excessive As
bioaccumulation in rice grains and other plant parts (root and
straw). Also widespread use of As-contaminated groundwater for
the irrigation of paddy rice causes an additional health hazard to
people who rely on this staple (Meharg et al., 2009; Mondal and
Polya, 2008; Rahman et al., 2009; Williams et al., 2006). Paddy fields
which are irrigated with As-contaminated groundwater act as net
sinksofAsfromgroundwaterandasmall amount returns to or re-
plenishes the aquifer (Neumann et al., 2011). These paddy soils and
As-contaminated irrigation water are both linked to elevated
concentrations of As in the rice grains (Heikens et al., 2007;
Panaullah et al., 2009; Xie and Huang, 1998). Furthermore, the depo-
sition of As in irrigated soils poses a serious threat to sustainable
agriculture in impacted areas (Heikens, 2006). However, grain-As
concentrations are impacted by the combined influences of soil char-
acteristics, environmental conditions, and crop management (Cheng
et al., 2006; Khan et al., 2009; Liu et al., 2006; Panaullah et al., 2009;
Xie and Huang, 1998; Xu et al., 2008)aswellasricevarieties(Ma
et al., 2008; Norton et al., 2009b; Zhang et al., 2008).
There is limited information on the uptake and translocation of As in
rice plants (Fig. 3). The uptake and translocation of As in rice plants is
greatly influenced by As species present in the rhizosphere (Arao et
al., 2011). The presence of Si and P in soil as well as in pore water also
impacted the As uptake by rice plants (Bogdan and Schenk, 2008).
Two main pathways identified to uptake As in rice plants include phos-
phate transport pathways since As(V) is an analogue of phosphate and
new evidence reveals that As(III) (silicic acid analogue) and undissoci-
ated methylated As species dimethylarsinic acid (DMA) and
monomethylarsonic acid (MMA) enter into the root by aquaporin chan-
nels which are used as silicate transport (Li et al., 2009a; Ma et al., 2008;
Wu et al., 2011b; Zhao et al., 2009). There is a significant variation in the
uptake efficiency of different As species. The uptake efficiency of meth-
ylated As species is comparatively lower than inorganic As species but
the translocation efficiency is higher for methylated As species (Raab
et al., 2007). Research shows that transfer of As in rice grains is greater
than other cereals (Williams et al., 2007c). This is due to the high soil/
shoot ratio and this difference in the transfer ratio is probably due to dif-
ferences in As speciation and dynamics in aerobic and anaerobic soils
(Williams et al., 2007c). In rice, the export of As from the shoot to the
grain appears to be under tight physiological control as the grain/
shoot ratio decreases by more than an order of magnitude (from 0.3
to 0.003 mg kg
−1
) and as As levels in the shoots increase from 1 to
20 mg kg
−1
.Norton et al. (2010b) examined the relationships between
P and As, and Si and As in a wide range of cultivars grown in As contam-
inated field trials in Bangladesh and China. They observed no correlation
between shoot and grain speciation, with the inorganic form compris-
ing 93 to 97% of As in the shoot and 63 to 84% in the grains (Norton et
al., 2010b).
Fig. 2. Selection process used for reviewing As accumulation in rice: consequences of rice genotypes and management practices to reduce human health risk, 1965–2016.
141S. Islam et al. / Environment International 96 (2016) 139–155
2.2. Arsenic phytotoxicity to rice plants
Arsenic is highly phytotoxic with the rice plants showing intolerance
to elevated levels of soil-As resulting in a decrease in plant growth and
crop yield (Barrachina et al., 1995; Carbonell-Barrachina et al., 1998).
Various reports investigated the effect of As on growth of rice and
most of them indicated that As has a significant negative effect on rice
growth andyield. There are reports of rice grown either in soil or in so-
lution culture with As or use of As-contaminated irrigation water which
significantly reduces plant height, effective tiller number, shoot bio-
mass, grain and straw yield reduction (Abedin et al., 2002a, 2002b;
Frans et al., 1988; Hossain et al., 2007; Islam et al., 2004a; Khan et al.,
2006; Liu, 1987; Marin et al., 1992; Milam et al., 1988; Tang and
Miller, 1991; Tsutsumi, 1980; Wang and Forsyth, 2006). Rice yield was
reduced by 10% at 12.9 mg As kg
−1
soil application, and 50% at
52 mg As kg
−1
application; no yield leading to death of plants occurred
at 109–157mgAskg
−1
soil application (Yan-Chu, 1994)and66%yield
loss when mean soil solution As concentration was raised to 1.5 mg L
−1
(Onken and Hossner, 1995). These findings demonstrate the c ritical role
that soil solution As can play on rice plant growth andhence yield. Thus
soil management strategies that change moisture levels such as via
flooding or irrigation can impact rice growth. This is further evident
from the studies conducted by Milam et al. (1988) who found a signifi-
cant yield difference when rice grown was under field conditions
followed by two irrigation options viz. continuous flooded and drained
during mid-season. For instance, application of As@ 0, 2, 4 or 6 lb acre
−1
results in an average grain yield of 1.32 ton acre
−1
with continuous
flooding and 1.78 ton acre
−1
with mid-season drainage (Milam et al.,
1988). However, As has also been found to enhance plant growth at
concentrations below the toxicity threshold value. For example, applica-
tion of As with the irrigation water up to 0.25 mg L
−1
enhanced the
plant height, panicle length, filled grains panicle
−1
, thousand grain
weight and finally the grain yield of Boro rice but further doses de-
creased the plant growth, yield and yield components (Islam et al.,
2004a). Above 60% and 40% grain yield reduction for rice varieties
Iratom 24 and BRRI dhan28, respectively, were found with 20 mg L
−1
of As compared to control and the reduction in straw yield was also
significantly higher for both rice varieties (Hossain, 2005). Maturity of
rice grains wasgreatly delayed in the high As (57 mg kg
−1
)plotswith
about 40–50% remaining green at the milk stage; as a result grain
yield was very poor, only 2.0–2.5 t ha
−1
(rough rice), which was almost
half the standard yield, which is about 4.0–5.0 t ha
−1
(Panaullah et al.,
2006). A field study in China, observed that grain-As concentrations
up to 0.72 mg kg
−1
when grown in soils containing 68 mg As kg
−1
(Xie and Huang, 1998). The concentrations of As in stems plus leaves
were more closely related to soil total and available As than those of
roots or grain (Kang et al., 1996). Another study from China indicated
that As concentration in rice increased significantly with the addition
of As in soil under field conditions with an isolation chamber (Wang
et al., 2006).
Rice showed a physiological disorder called the straighthead disease,
showing delayed heading date, shortened plant height and dramatically
reduced grain yield which has been reported to be associated with As
(Gilmour and Wells, 1980). The severity of straighthead increased sig-
nificantly with the increase of soil As concentration. Straighthead
caused approximately 17–100% sterile spikelet formation and about
16–100% reduction of grain yield (Rahman et al., 2008a). Literature sug-
gests that As affects the photosynthetic pigments, chlorophyll-aand
chlorophyll-bwhich is directly associated with rice yield and growth.
Both chlorophyll-aand chlorophyll-bcontents in rice leaf decreased sig-
nificantly with the increase in soil As concentration of up to 30 mg kg
−1
.
No rice plant survived up to maturity stage in soil treated with 60 and
90 mg As kg
−1
(Rahman et al., 2007a). The varieties Cocodrie, Mars,
Kaybonnet, and Bengal were highly susceptible to straight head with
ratings from 7.2 to 8.0 and grain yield reductions from 80 to 96%.
Wells, Lagrue, Drew, Cypress, and Japan 92.09.31 were susceptible
with ratings from 5.9 to 6.7 and yield reductions from 49 to 73% (Yan
etal.,2005).Arsenicspikedsoilshaving a concentrationof 50 mgAskg
−1
showed straighthead symptoms of rice causing 17–100% sterile
spikelet's in BRRI dhan29 (Rahman et al., 2008a). Higher As also causes
grain sterility in rice as reported in a field and glasshouse study in
Bangladesh (Islam et al., 2004a). Grain yield and the occurrence of
straighthead disease were cultivar-dependent and influenced by soil
As level and water management practices (Hua et al., 2013).
Fig. 3. Possible mechanisms of As uptake in rice plants.
(adopted and modified from Ali et al., 2009)
142 S. Islam et al. / Environment International 96 (2016) 139–155
Straighthead resistant cultivars yielded more and had lower grain-As
than the susceptible ones. Elevated soil As with continuous flood man-
agement significantly reduced the grain yield of susceptible cultivars
by N89% due to substantially increased straighthead,which was induced
by increased As content in grains. This study demonstrates that the se-
lection of less As-susceptible cultivars and intermittent flood water
practice could be an effective means to lower the As accumulation in
grains and minimize the occurrence of the As-induced straighthead
symptom and yield reduction.
2.3. Bio-accumulation of As in rice plants
The uptake or bioavailability and phytotoxicity of As is dependent on
a number of factors including the source andthe concentration of theel-
ement (NAS, 1977), nature of As species (Carbonell-Barrachina et al.,
1998; Marin et al., 1992), soil properties such as clay content, pH and
redox equilibrium conditions (Johnson and Hiltbold, 1969; Marin et
al., 1993; Naidu et al., 2006) and type and amount of organic matter
present (Mitchell and Barre, 1995). Adsorption–desorption and micro-
bial processes are also predominant factors controlling the bioavailabil-
ity of As (Naidu et al., 2006). The As content of rice plant parts generally
follow the pattern: root Nstraw Nhusk Ngrain (Lei et al., 2012; Rahman
et al., 2007b; Smith et al., 2008; Wang et al., 2006; Xie and Huang,
1998). Moreover, the concentration of As in all plant parts increased
with an increase in soil As (Marin et al., 1992; Odanaka et al., 1987).
From a solution culture experiment it was observed that most of the
As(V), As(III) and MMA accumulated in the root while DMA was readily
translocated to the shoot (Marin et al., 1992; Odanaka et al., 1987). This
was furtherdemonstrated by Marin et al. (1993) who found rapid trans-
location of DMA up the plant with higher concentrations of As in rice
shoots than in roots when rice plants were grown in DMA solution at
0–1.6 mg As L
−1
. Irrespective of As chemical forms, root As concentra-
tion was 10.5 mg kg
−1
in the 0.05 mg As L
−1
treatment, which in-
creased to 212.7 mg kg
−1
in the 0.8 mg As L
−1
treatment (Marin et
al., 1992). Increasing concentration of As in irrigation water significantly
increased As concentrations in root, straw, and rice husk (Abedin et al.,
2002a; Carbonell-Barrachina et al., 1998; Hossain et al., 2007). A similar
relationship between soil solution As and plant response was found in
rice (Onken and Hossner, 1995). Application of As-contaminated irriga-
tion water tothe 1st crop Boro rice (dry season rice) had significant re-
sidual effects on the 2nd rice crop T. Aman (rain fed rice) and also
increased the levels of As accumulation (Islam et al., 2004a). Arsenic
present in irrigation water was found to enhance the bioaccumulation
of As in rice plants (Bhattacharya et al., 2010)andhigheraccumulation
of As was noticed in the root as compared to the straw, husk, and grain.
The As concentration of rice produced in As-contaminated areas is 2 to 3
times higher than that produced in non-contaminated districts of Ban-
gladesh (Hironakaand Ahmad, 2003) with total As concentrations rang-
ingfrom0.11to0.34mgkg
−1
in contaminated soils.
The range in concentrations of inorganic As in all unpolished rice
was from 0.26 to 0.52 mg kg
−1
dry weight. In the case of aromatic
rice, it had a low level of As compared to other typesof rice from differ-
ent regions of Bangladesh (Ahmed et al., 2011). Rice samples from the
districts of Pabna, Chapai Nawabganj, Rajbari, Faridpur and Gopalganj
of the Gangetic floodplains of Bangladesh showed that 16% of the
grain samples had no detectable As while on the other hand 14% grains
had As level N1mgkg
−1
(Islam et al., 2004b). Comparing varietal ef-
fects, the grain-As concentrations in IR 8 and BRRI dhan29 rice were
higher in comparison with BRRI dhan28 and Parija and grain-As concen-
trationswere always lower than in straw As (Islam et al., 2004b). Arse-
nic accumulation by rice plants growing with either As(III) or As(V)
increased with increasing As treatment, irrespective of water regimes
(Huq et al., 2006). However, the accumulation was greater in the
As(III) treated soil than that in the As(V)-treated one.
In a study from West Bengal (India), high levels of As in the ground-
water led to bioaccumulation of high levels of As in rice b0.04–
0.605 mg kg
−1
(Roychowdhury et al., 2003). Another study has
shown that the concentration of total As in cooked rice ranged from
0.33 mg kg
−1
in the Jalangi block, Murshidabad to 0.38 mg kg
−1
in
the Domkal block, Murshidabad district west Bengal India
(Roychowdhury et al., 2002). Paddy rice from North 24-Parganas India
was detected at 0.50 mg kg
−1
As (Signes-Pastor et al., 2009). Aman
rice (wet season rice) in Bangladesh has been studied previously by dif-
ferent groups (Duxbury et al., 2003; Williams et al., 2006). The mean
concentration of As in Aman rice was 0.125 mg kg
−1
with a range of
0.072–0.170 mg kg
−1
and in Boro rice (Irrigated dry season rice) it
was 0.183 mg kg
−1
(Duxbury et al., 2003), and the range of As levels
in another study was 0.180–0.310 mg kg
−1
(Williams et al., 2006). Ar-
senic levels in the rice grains of Bangladesh were typical of other regions
of the world when rice was grown in an area where ground water and
paddy soils contained low levels of As (Meharg and Rahman, 2003). In
general, lower As content was found for Aman rice compared to Boro
rice (Duxbury et al., 2003). However, it has also been reported that
Aman rice from As affected areas have higher As content than Aman
and Boro rice from low As affected areas (Roberts et al., 2011;
Williams et al., 2006). The statistics of rice total As contents produced
in different countries is presented in Table 1.
Table 1
Total As concentration (mg kg
−1
) in rice producedin different countries of the world.
[modified from Meharg et al. (2009) and Rahman et al. (2009)]
Countries Total As References
No. of
sample
Range
(mg kg
−1
)
Mean
(mg kg
−1
)
Bangladesh 214 0.002–0.557 0.143 Rahman et al. (2009)
144 0.020–0.330 0.130 Meharg et al. (2009)
4b0.005–0.020 0.011 Al Rmalli et al. (2005)
78 (Boro) 0.108–0.331 0.183 Duxbury et al. (2003)
72 (Aman) 0.072–0.170 0.117
10 0.040–0.270 0.136 Das et al. (2004)
15 0.030–0.300 0.130 Williams et al. (2005)
13 0.058–1.835 0.496 Meharg and Rahman
(2003)
133 (Boro) 0.040–0.910 –Williams et al. (2006)
189 (Aman) b0.040–0.920 –
6 (Aromatic) 0.038–0.073 0.056 Rahman et al. (2014)
35 0.050–2.050 −0.080 Islam et al. (2004b)
India 133 0.180–0.310 0.070 Meharg et al. (2009)
11 0.041–0.605 0.232 Roychowdhury et al.
(2003)23 0.079–0.546 0.033
8 0.120–0.663 0.358 Chowdhury et al. (2001)
China 11 0.283–0.725 0.501 Xie and Huang (1998)
124 0.020–0.460 0.140 Meharg et al. (2009)
2 0.460–1.180 0.820 Sun et al. (2008)
43 0.052–0.253 0.129 Ma et al. (2016)
165 0–0.665 0.116 Huang et al. (2015)
446 0.033–0.437 0.143 Li et al. (2015)
21 0.065–0.274 0.114 Liang et al. (2010)
Taiwan 5 0.010–0.630 0.100 Lin et al. (2004)
–0.010–0.140 0.050
2 0.190–0.210 0.200 Schoof et al. (1998)
5 0.063–0.170 0.130
280 0.100–0.630 –Lin et al. (2004)
Thailand 53 0.060–0.500 0.150 Adomako et al. (2011)
54 0.010–0.390 0.140 Meharg et al. (2009)
Vietnam 31 0.032–0.465 0.208 Phuong et al. (1999)
Turkey 50 –0.202 Sofuoglu et al. (2014)
Japan 26 0.070–0.420 0.190 Meharg et al. (2009)
France 33 0.090–0.560 0.280
Italy 38 0.070–0.330 0.150
Spain 76 0.050–0.820 0.200
Egypt 110 0.010–0.580 0.050
USA 163 0.030–0.660 0.250
134 0.100–0.660 –Williams et al. (2007b)
33 0.026–1.000 0.210 Heitkemper et al. (2009)
Australia 21 0.188–0.438 0.270 Rahman et al. (2014)
Pakistan 9 (Basmati) 0.073–0.088 0.082
143S. Islam et al. / Environment International 96 (2016) 139–155
The mean and median grain-As levels for the South Central US were
0.30 mg kg
−1
and 0.27 mg kg
−1
, respectively (Williams et al., 20 07b). In
that study, 22 rice grain samples were analyzed from Camargue
(France) reporting that the mean and median grain-As levels were
0.32 and 0.34 mg kg
−1
, respectively. The total As was measured in
901 polished rice samples originated from 10 different countries
(Meharg et al., 2009). The data showed that the lowest As content
was in Egyptian rice (0.04 mg kg
−1
) and the highest As was in US rice
(0.25mgkg
−1
). Although, Bangladeshi rice contained 0.13 mg kg
−1
of
total As, 61% of this was present as inorganic As (Meharg et al., 2009).
Rice grown in pots with high-As soil produced on average 17-fold
higher grain-As than the same cultivars grown in a paddy field with
low-As soil (Kuramata et al., 2011). Another result showed that when
the soil and ricecultivar were the same, pot-grown ricegrains contained
larger percentagesof DMA than field grown rice (Khan et al., 2010). This
differences may be due to the redox potential differences between pot
and field soil.
2.4. Arsenic bio-accumulation and rice genotypes
There is a considerably wide variation among the rice genotypes to
As sensitivity. Three interacting loci on chromosomes 6 (2 loci) and 10
(1 locus) have been identified to regulate As tolerance in rice
(Dasgupta et al., 2004; Zhang et al., 2008). A number of field studies
have shown substantial genetic variation in grain-As concentration as
well as As speciation. Germination testand seedling growth by exposing
eight Bangladeshi rice varieties to As(III) and As(V) showed that germi-
nation wasslightly inhibited at 0.5 and 1 mg L
−1
. At 2 mg L
−1
, inhibition
was N10% (Heikens, 2006). Root growth was inhibited by ~ 20% at
0.5 mg L
−1
and As(V) was more toxic than As(III) (Meharg and
Rahman, 2003). Another investigation showed the significant genotypic
difference in response to As(V) toxicity in rice on root elongation, As(V)
uptake kinetics, physiological and biochemical response and As specia-
tion (Geng et al., 2005). Uptake kinetics data showed that P-deprived
genotype 94D-54 had a little higher As uptake than P-deprived 94D-
64, but the difference was not large enough to cause acute toxicity in
P-deprived 94D-54. A significant genotypic variation was detected in
the As concentrations in rice grains. There are also significant geno-
type-environment (location) interactions of the concentrations of As
in grains, suggesting the importance of cultivar choice in producing
rice with low As in grains for a given location (Cheng et al., 2006). Lei
et al. (2012) evaluated the effect of As-contaminated soil on uptake
and distribution of As in 34 genotypes of rice (including unpolished
rice, husk, shoot, and root). Themean As concentrations in rice plant tis-
sues were different among the 34 rice genotypes. The highest As con-
centrations were accumulated in rice root (196.27–385.98 mg kg
−1
dry weight), while the lowest was in unpolished rice (0.31–
0.52 mg kg
−1
dry weight). The statistics of rice varietal differences on
total As contents is presented in Table 2.
The rice varieties did not show significant differences in As accumu-
lation in straw, husk, brown and polish grain when the concentrations
of As in soil was low. However, at higher concentrations, different rice
varieties showed significant differences. Arsenic translocation from
root to shoot (straw) and husk was higher in hybrid varieties compared
to those of non-hybrid varieties (Rahman et al., 2007c). The uptake and
translocation of As from roots to grains involves a number of steps and
checkpoints with these giving rise to potential differences among geno-
types. Rice genotypes also show significant differences in iron plaque
formation of a reddish brown coating on the root surface (Dwivedi et
al., 2010; Geng et al., 2005; Liu et al., 2011b) as a result rhizosphere
characteristics which may also play an important role. Significant nega-
tive correlations have been showed between As concentrations in straw
or grain with root porosity or radial oxygen loss (ROL) from the roots
among 20–25 rice cultivars (Mei et al., 2009; Wu et al., 2011a). The cul-
tivars able to release more oxygen may maintain the rhizosphere at a
higher redox potential and form more iron plaque on the root surface,
which may reduce As uptake (Wu et al., 2011a). Mei et al. (2012) re-
cently focused on genotypes with higherROL which havea strong abil-
ity to reduce As accumulation in shoots and increase As tolerance by
reducing As mobilization in the rhizosphere and by limiting As
translocation.
There are significant genotypic differences in As concentrations of all
organs, and polished grains are significantly affected by genotype and
soil type (Ye et al., 2012). This result indicated that As concentration
in grain was partially governed by As uptake and the transfer from
root to grain. Some genotypes, such as Japonica rice had consistently
low grain-As concentrations. An experiment with 216 rice genotypes
showed that the averages of As contents for Indica rice were higher
than those of Japonica rice. The ranges of As contents in Indica rice and
Japonica rice were 0.021–0.296 mg kg
−1
and 0.005–0.274mgkg
−1
,re-
spectively (Jiang et al., 2012). When rice is grown in low As affected
areas the results indicated that Boro (0.072 mg kg
−1
) and Aman
(0.086 mg kg
−1
) rice contained 2 to 4-fold lower levels of As compared
to high As affected areas of Bangladesh but inorganic As content (70%)
was the same as the highly affected areas (Al-Rmalli et al., 2012). This
suggests that irrespective of total As content it is the bioaccumulation
Table 2
Summarizing the varietal differences on grain-As accumulation.
Study type Rice genotypes Total grain As
(mg kg
−1
)
References
Glasshouse study BRRI dhan28 0.23 ± 0.05 Rahman et al. (2007a)
BRRI dhan29 0.16 ± 0.08
BRRI dhan35 0.20 ± 0.03
BRRI dhan36 0.18 ± 0.05
BRRI hybrid dhan1 0.14 ± 0.04
BRRI dhan29 0.21–0.86 Khan et al. (2009)
BRRI dhan33 0.23–0.75
BRRI hybrid dhan1 0.70 ± 0.01 Rahman et al. (2007b)
BRRI dhan28 0.60 ± 0.01
BRRI dhan29 0.86 ± 0.03 Talukder et al. (2012)
BRRI dhan32 0.37 ± 0.01
YY-1 0.66 ± 0.03 Liu et al. (2006)
94D-64 0.32 ± 0.02
KY1360 0.32 ± 0.00
Gui630 0.43 ± 0.10
94D-54 0.41 ± 0.04
94D-22 0.36 ± 0.01
Hybrid 0.89–1.03 Ye et al. (2012)
Japonica 0.24–0.63
Indica 0.43–0.66
Field study Local landraces
with red bran
0.57 ± 0.08 Norton et al. (2009b)
Local landraces
with brown bran
0.42 ± 0.08
Red minikit 0.17–0.57 Bhattacharya et al. (2010)
Megi 0.11–0.43
Aman:BR23,
BRRI dhan33
0.10–0.22 Ahmed et al. (2011)
Boro: BR3,
BRRI dhan35
0.22–0.34
IR-68144-127 1.68 ± 0.38 Dwivedi et al. (2010)
IR-68144-120 0.97 ± 0.14
CN1643-3 0.68 ± 0.02
CN1646-2 0.78 ± 0.02
IR-36 0.41 ± 0.19
IR-64 0.41 ± 0.02
Gotrabhog
(IET-19226)
0.52 ± 0.06
Zhe 733 0.45 ± 0.07 Hua et al. (2011)
Rondo 1.25 ± 0.12
Cocodrie 1.45 ± 0.98
9 non-glutinous 0.10–0.66 Kuramata et al. (2011)
1 glutinous 0.14
Hybrid middle rice 0.31–0.47 Lei et al. (2012)
Hybrid late rice 0.35–0.51
Normal late rice 0.39–0.50
Indica rice 0.02–0.30 Jiang et al. (2012)
Japonica rice 0.01–0.27
144 S. Islam et al. / Environment International 96 (2016) 139–155
potential of plants that control grain Ascontent. These investigators also
reported that the As level in aromatic rice (0.049 mg kg
−1
) from the
Sylhet region was over 40% lower than that of non-aromatic rice
(0.081 mg kg
−1
); aromatic rice also contained higher levels of essential
elements and consumption of aromatic ricemay increase Se and Zn in-
take by 46% and 23% respectively (Al-Rmalli et al., 2012).
Arsenic concentration varied between rice subpopulations and
ranges between 3 and 34 fold (Norton et al., 2012) and inorganic As cor-
related strongly with total As among a subset of 40 cultivars harvested
in Bangladesh and China. Genetic variations have shown a large deter-
mining factor for grain-As concentration. The rice genotype Bala is toler-
ant to As contamination while the genotypes Cocodrie, Mars,
Kaybonnet, Bengal, and Azucena are very much sensitive to As contam-
ination (Yan et al., 2005). Other field trials at Faridpur and Sonargaon,
Bangladesh, showed 4–4.6 fold variations in total grain-As among 76
cultivars including local landraces, locally improved cultivars and par-
ents of permanent mapping populations (Norton et al., 2009b).
Environmental factors made the largest contribution to the variation in
grain-As (61%), followed by genotype (6%) and genotype × environment
interaction (19%) (Norton et al., 2009a). These results indicate a
genotype × environment interaction across diverse environments,
which is not surprising considering that As bioavailability in soil is greatly
influenced by soil properties, the source of As contamination, water
management and other environmental factors. Results showed that there
was a significant genotype effect on the percentages of inorganic As and
DMA in grain, but the influence by the environmental factor was greater
(Norton et al., 2009a, 2009b). Another report showed that variety and envi-
ronmental interaction had significant effects on grain-As concentrations in
both wet and dry seasons and the relative variability due to the environ-
ment was greater than that due to varietal differences in both seasons
(Ahmed et al., 2007). Environment and genotype × environment interac-
tion accounted for 70–80% and 10–21% of the total variation, respectively
while the variety explained only 9% of the total variation in As. A large en-
vironmental effect has been reported from a trial of 38 Bangladeshi cultivars
grown at ten experimental sites across different agro-ecological zones
of Bangladesh (Ahmed et al., 2011). Environment factors accounted for
69–80% of the variation in grain-As concentration, whereas genotype
and genotype × environment interactions accounted for only 9–10% and
10–21% of the variability. Another report showed that total grain-As and
As species [As(III) and DMA] varied widely among 25 diverse rice cultivars.
It indicates that As concentration and speciation are mostly dependent on
genotype, which accounted for about 70% of the variation in total grain-
As (Pillai et al., 2010). All the studies identified a number of rice cultivars
that accumulated relatively low levels of As across sites and it may be
concluded that genetic stability is greater across seasons than across diverse
sites (Ahmed et al., 2011; Norton et al., 2009a, 2009b; Pillai et al., 2010).
2.5. Arsenic speciation and rice genotypes
Similar to its adverse impact on human health that varies with na-
ture of speciation, toxicity of As to rice depends to a large extent on its
speciations with significant difference between organic and inorganic
As species. Therefore, the speciation and localization of As species and
their distribution in rice grains are key factors controlling the bioavail-
ability of As. As discussed above, the distribution of As varies between
the various parts of the grain (husk, bran and endosperm) and are char-
acterized by element specific distribution patterns. Themajor As species
in roots, stems and leaves are As(III) and As(V), while As(III) and DMA
(comprising 85 to 94% of the total As) are the major As species identified
in the grain (Smith et al., 2008). Speciation of As in husk, bran and the
endosperm shows As(III)-thiol complexes as the predominant form
(Lombi et al., 2009). Depending on nature of As speciation, rice can be
classified into two groups —those rich in inorganic As-type and those
that are rich in DMA-type, depending on the speciation of As in the
grain. The dominant form of As speciation in rice grains appears to de-
pend on the total rice As content. Zavala et al. (2008) found that when
rice contains low levels of As, the dominant species is As(III) but when
rice contains high levels of As, the dominant form is DMA (Zavala et
al., 2008). In the same report researchers also indicate much difference
in the speciation of As in rice from the US and Europe and Asia. These re-
searchers methylated As (a less toxic form) in the rice from the US
whereas rice grown in Europe and Asia contains the predominant
form which is toxic inorganic As. Inorganic As has been detected in Eu-
ropean, Bangladeshi, and Indian rice at 64 ± 1% (n= 7), 80 ± 3% (n=
11), and 81 ± 4% (n= 15), respectively. A study from Australia showed
that Asian rice contained relatively lower total As than Australian and
Italian rice on sale in Australia. In Asian rice inorganic As is predominant
(86–99%),whereas in Australian rice the average inorganic As and DMA
in Australian-grown rice were found to be about 58–63% and 18–26% of
the total As, respectively (Rahmanet al., 2014). In another study an Aus-
tralian rice variety Quest represents DMA that was 85–94% of the total
As (Smith et al., 2008). The statistics of As species concentration in
rice produced in different countries is presented in Table 3.
The proportion of organic and inorganic As in rice is also cultivar de-
pendent (Williams et al., 2005). A study reported the higher amount of
inorganic As in Boro rice (mean 82%) than in Aman rice (mean 66%) in
Bangladesh (Williams et al., 2006). Brown rice has a higher proportion
of inorganic As than white rice (Meharg et al., 2008) and in white rice
As was generally dispersed throughout the grain, while in brown rice
As was found to be preferentially localized at the surface. Another
study showed that inorganic As is the major species in all rice cultivars
(Kuramata et al., 2011) with few genotypic differences in the levels of
total As and inorganic As in the grain. In soils with high As levels, the
total grain-As increased, markedly with increased levels of DMA. Speci-
ation of As in rice is also affected by root aeration and variation of geno-
types (Wu et al., 2011a).
There is a significant genotype effect in the percentage of inorganic
As and percentage of cacodylic acid respectively, in rice grains. An anal-
ysis of polished rice from various production regions of China showed
that the inorganic As species was predominant, accounting for approx-
imately 72% of the total As in rice, with a mean concentration of
82.0 ng g
−1
(Liang et al., 2010). The concentrations of total grain-As
and As species [As(III) and DMA] have been shown to vary widely
among 25 diverse rice cultivars (Pillai et al., 2010). Arsenic concentra-
tion and speciation are mostly dependent on genotype, which accounts
for about 70% of the variation in total grain-As. However, the proportion
of organic species (DMA and MMA) in rice grains may vary depending
on the source of grain and method of extraction. Nowadays not only in-
organic As but also organic forms should receive more attention as the
toxicity of the trivalent As metabolites MMA and DMA were found to
be highly toxic to human cell lines (Petrick et al., 2000; Styblo et al.,
2000).
3. Management practices to control As bio-accumulation
3.1. Fertilizers and other soil amendments
For successful crop production and reducing As toxicity, fertilizer
amendments play an important role. Rice contributes N80% of dietary
energy and isa source of many minerals, including selenium (Se), man-
ganese, iron and zinc (Zn). There are many reports on the effect of sili-
con (Si) and Se on As accumulation. Silicon was reported to play a
significant role in reducing As toxicity, because the two elements act
as metabolic antipodes. A recent study suggests that a high level of As
in soils decreases Se and Zn levels in rice (Norton et al., 2010a). Many
forms of Si fertilizers are available, some of which are industrial by-
products, such as basic slag from the steel manufacture processes
(Takahashi, 2002). Silicon fertilizers have been widely used in different
countries in the world especially in Japan to increase rice yield. Silicon
fertilization can effectively decrease As accumulation in rice in the
areas affected by As under field conditions. Addition of 20 g kg
−1
of sil-
ica gel (SiO
2
) to soil decreased As concentrations in straw and grain by
145S. Islam et al. / Environment International 96 (2016) 139–155
78% and 16%, respectively (Li et al., 2009b). In the same experiment, Si
addition decreased the concentration of inorganic As in grain by about
60% but increased the DMA concentration by 33% (Li et al., 2009b).
This study also showed significant differences between two rice geno-
types in grain-As speciation. A recent study investigating the effects of
Si fertilization on As uptake and speciation in rice plants with different
radial oxygen loss (ROL) showed that Si addition significantly reduced
shoot and root total As and also decreased the inorganic As in shoots
and had the tendency to increase DMA concentrations (Wu et al.,
2015). Another investigation demonstrated the potential effects of Si
fertilizers in reducing As in rice grains and found that the application
of high NP + S-KSi 9000 (9000 kg ha
−1
) significantly reduced the As
concentration in rice grains by up to 20%, compared with the control
(Wang et al., 2015a). Silica nutrition acts asa central player that restricts
photosynthetic impairmentin As-treatedplants and in addition to lim-
iting As uptake via modulation of the expression of genes with prime
importance in As uptake and translocation (Sanglard et al., 2016).
Nitrate fertilization also has a positive role in reducing As uptake. A
pot microcosm experiment showed that nitrate addition reduced As up-
take by the rice plant. Nitrate may inhibit Fe(III) reduction and/or stim-
ulate nitrate-dependent Fe(II) oxidation, leading to As co-precipitation
with, or adsorption to, Fe(III) minerals in the soil (Chen et al., 2008).
Considerable controversy exists in the literature regarding the effect of
P on the availability and uptake of As by rice. For instance, P has been
shown to increase, decrease or have no effect on the uptake of As. A
number of pot and field experiments have shown that additions of
phosphate fertilizer to soil decreased As uptake by rice (Abedin et al.,
2002a; Hossain et al., 2009; Talukder et al., 2011; Wu et al., 2011b).
However, in some cases phosphate was found to be effective in ex-
changing adsorbed As(V) or As(III) from the soil solid phase and from
the iron plaque on the root surface, thus increasing the As availability
to rice plants. Another report indicates that P influences As mobility
and bioavailability which depends on the charge components of the
soil (Bolan et al., 2013). Results showed that the addition of P resulted
in an increase in As desorption, and the effect was more pronounced
in the case of allophanic soil. In the case of both As-spiked soils and
field contaminated sheep-dipsoil, application of P increased the desorp-
tion of As, thereby increasing its bioavailability. Phosphorus concentra-
tion may promote As translocation from roots to shoots. High P
concentrations decreased the percentages of As distribution in iron
plaque from around 70% to 10%, while it increased the percentages of
As in roots and shoots (Geng et al., 2005). Phosphate fertilization or sil-
icon amendment of paddy soils could therefore be possibilities for re-
ducing the uptake of As(V) and As(III) in the rice plant, respectively
(Hu et al., 2005;Li et al., 2009b). However, a higher level of P application
than plant requirements is not appropriate due to environmental con-
cerns (Lee et al., 2015).
Iron-oxide minerals have a significant impact on As solubility, reten-
tion, release and As dynamics in flooded rice culture (Loeppert et al.,
2005; Takahashi et al., 2004). Iron plaque diminishes the inhibition ef-
fect of phosphate on paddy rice's As(V) uptake (Chen et al., 2005). Sev-
eral recentstudies have shown that the formation of iron plaque on the
rice root surface leads to the sequestration of As and reduction in As
transfer to shoot and grain (Chen et al., 2005; Liu et al., 2004, 2005;
Voegelin et al., 2007). The accumulation of As on plaque in close associ-
ation with the root might increase the potential bioavailability of As
(Voegelin et al., 2007). Iron oxides or hydroxides are an important com-
ponent in the soil–paddy rice system with a strong influence on As bio-
geochemistry. Hossain et al. (2009)) found that when As was added to
the soil, an addition of ferrous sulphate alleviated As toxicity in rice
plants, possibly by enhancing the formation of iron plaque on the root
surface (Hu et al., 2007). Sometimes sulphur itself plays an important
role in As detoxification and mobility from root to shoot. The nutritional
status of sulphur can influence As detoxification and mobility within a
plant due to the role of thio-rich compounds (e.g. phytochelatins) in
complexing and sequestering As(III) (Zhao et al., 2010). Plants deficient
Table 3
Arsenic species concentration (mg kg
−1
)inriceproducedindifferentcountries.
Countries Types of rice As species Total As References
As(III) As(V) DMA MMA
Bangladesh White 0.083
(0.010–0.210)
–0.019
(0–0.050)
–0.131
(0.030–0.300)
Williams et al. (2005)
Brown 0.280 –0.170 0.010 0.610 Meharg et al. (2008)
India White 0.027
(0.020–0.040)
–0.066 0.0007 0.046
(0.030–0.050)
Williams et al. (2005)
Brown 0.040 –bLOD bLOD 0.070
Red 0.050 –0.010 bLOD 0.080
China White 0.114
(0.051–0.302)
0.040 0.040
(0.009–0.147)
0.013
(0.007–0.013)
0.230
(0.019–0.586)
Zhu et al. (2008)
Brown 0.210 –0.090 0.010 0.360 Meharg et al. (2008)
Taiwan White 0.247
(0.110–0.510)
–0.037
(0.030–0.050)
0.032
(0.015–0.060)
0.383
(0.190–0.760)
Williams et al. (2005)
Thailand Light yellow 0.080 –0.060 bLOD 0.110 Williams et al. (2005)
Japan White 0.071 0.013 0.011 0 0.095 Narukawa et al. (2008)
Spain –0.080 –0.050 bLOD 0.170 Williams et al. (2005)
USA White 0.076
(0.020–0.100)
0.042
(0.032–0.051)
0.077
(0.050–0.260)
bLOD 0.277
(0.170–0.400)
Williams et al. (2005)
White 0.092
(0.079–0.101)
–0.0137
(0.136–0.141)
–0.329
(0.308–0.350)
Zhu et al. (2008)
White 0.110 –0.155
(0.040–0.302)
bLOD 0.280 Meharg et al. (2008)
Brown 0.105
(0.060–0.140)
–0.090
(0.010–0.150)
bLOD 0.225
(0.110–0.340)
Williams et al. (2005)
Brown 0.170 ––0.010 0.440 Meharg et al. (2008)
Brazil White 0.078
(0.040–0.156)
0.034
(0.016–0.062)
0.093
(0.039–0.258)
0.008
(0–0.029)
0.223
(0.109–0.376)
Batista et al. (2011)
Boiling white 0.087
(0.045–0.127)
0.043
(0.024–0.060)
0.065
(0.017–0.139)
0.010
(0–0.051)
0.215
(0.108–0.367)
Brown 0.146
(0.139–0.151)
0.042
(0.037–0.051)
0.127
(0.070–0.206)
0.011
(0–0.018)
0.348
(0.271–0.428)
Canada Wild 0.045
(0.010–0.080)
–0.010 bLOD 0.065
(0.020–0.110)
Williams et al. (2005)
146 S. Islam et al. / Environment International 96 (2016) 139–155
in sulphur were found to have a greater root-to-shoot translocation of
As (Liu et al., 2010; Zhang et al., 2011). Another result shows that rice
plants treated with sulphur fertilizers show an apparent increase of
iron plaque formation and small decreases in the concentrations of As
in rice tissues (Hu et al., 2007). Thus the proper supply of S nutrition
may be helpful in the prevention of As accumulation in the aerial parts
of a plant as well as As induced toxicity (Dixit et al., 2016).
Organic matter has many important roles in crop production. How-
ever, there are some controversies regarding the effect of organic matter
on As uptake and mobilization. Mobilization of As from paddy fields is
governed by organic matter (Williams et al., 2011). Generally microbes
utilizing the organic matter consume oxygen that leads to a decrease in
redox potential, which in turn leads to As dissolution from FeOOH
(Rowland et al., 2009; Smedley and Kinniburgh, 2002). It has another
two rolesin As availability insoils, desorbingAs species from soilsurface
exchange sites (Weng et al., 2009), and dissolved organic matter
complexing As species (Liu et al., 2011a; Williams et al., 2011). Recently,
it was reported that total grain-As is higher when rice was grown in soil
with high As contamination and high organic matter, with the increase
of organic Asspecies. The results indicate that the application of organic
matter in paddy soils with elevated soil As may lead to an increase in the
accumulation of As within rice grains (Norton et al., 2013). Biochar is a
pyrogenic carbon material produced by the combustion of biomass
under oxygen limited conditions. It has an important role in soil reme-
diation for example by reducing As mobility and bioavailability in soils
because of its unique properties (high surface area and cation exchange
capacity) (Ahmad et al., 2014; Zhou et al., 2013). It also improves soil
fertility and helps carbon sequestration (Mohan et al., 2014; Wang et
al., 2015b). There are several methods developed to modify the biochar
to increase its sorption capacity to As. However, the methods used are
relatively complex and costly. Thus, additional investigations are need-
ed to develop simple and cost-effective methods to modify biochar.
3.2. Water management
Water management techniques may also prove to be highly effective
in combating the problem of excessive accumulation of As in rice grains.
Besides saving water it also means less input of As into the paddy field
from irrigation of As-contaminated groundwater. Selecting appropriate
water management practices and rice cultivars according to the level of
As contamination in soils will benefit food security as well as high yields.
In the dry season, irrigation is required for rice crops. In As-affected
areas like Bangladesh As-contaminated groundwater is widely used to
irrigate rice crops (Saha and Ali, 2007), which has resulted in elevated
As concentrations in soils and rice grains (Hossain et al., 2008; Islam
et al., 2007; Meharg and Rahman, 2003), and significant yield losses
due to As phytotoxicity (Khan et al., 2009; Panaullah et al., 2009;
Tripathi et al., 2013). A study was conducted on the impact of As con-
taminated irrigation water on soil-As content and rice productivity
over two dry seasons in the area of the Faridpur district, Bangladesh
with soil-As levels ranging from about 10 to 70 mg kg
−1
.Asimple
mass balance calculation using the current water As level of
0.13 mg As L
−1
suggested that 96% of the added As was retained in
the soil and yield declined progressively from 7–9to2–3tha
−1
(Panaullah et al., 2009).
It has been estimated that 900 to 1360 tons of As per year is brought
onto the arable land of Bangladesh due to irrigation with As-contami-
nated groundwater (Ali, 2003). In paddy soils, the reductive mobiliza-
tion of As under the anaerobic conditions greatly enhances the
bioavailability of As to rice that leads to excessive accumulation of As
in rice grains and straw. In comparison, soils maintained at aerobic con-
ditions generally have very low levels of As in the soil solution. Under
aerobic conditions the least As accumulation was found in rice straw
and grain (Liet al., 2009b) but anaerobic conditions enhanced As uptake
in rice plants (Talukder et al., 2012). Research showed that aerobic soil
conditions resulted in a 10-fold lower amount of As uptake among a set
of several hundred global cultivars as compared to a flooded field
(Norton et al., 2012). Similar results from a greenhouse experiment,
maintaining soil under aerobic conditions decreased As concentration
in rice grains and straw by 10–20 fold, and 7–63 fold, respectively, com-
pared with continuously flooded rice (Li et al., 2009b; Xu et al., 2008).
Further investigation of the effect of imposing a period of aerobic condi-
tions during either the vegetative or reproductive growth showed that
it decreased grain-As by 80% and 50%, respectively (Li et al., 2009b).
The pot study also showed that an aerobic treatment 3 weeks before
and after heading was effective in reducing grain-As concentration
(Arao et al., 2009). It was also observed that intermittent flooding re-
duced As uptake (23.33% in root, 13.84% in shoot and 19.84% in leaf)
at panicle initiation stage, instead of continuous flooding (Rahaman et
al., 2011).
Under aerobic soil conditions, As(V) is the dominant species, where-
as under submerged soil conditions the predominantspecies is As(III). A
comparative study using flooded versus intermittent flooding by Hua et
al. (2011) showed that the combination of water management practices
that allow periodic aeration of the soil and use of cultivars that are low
accumulators of As can reduce As in the grain. A recent field study
showed that in flooded treatment, As(III) in the pore water wasthe pre-
dominant As species, accounting for 87.3–93.6% of the total As, whereas
in the non-flooded and Alternate Wetting and Drying (AWD) treat-
ments, As(V) was the dominant species, accounting for 90–96% and
73–83%, respectively (Das et al., 2016). In another field study in Bangla-
desh, the site employing intermittent irrigation showed lower grain-As
content than the site under continuously flooded conditions (Stroud et
al., 2011). A field study at Stuttgart, Arkansas, US showed that grain
contained 41% lower As in an intermittently flooded paddy than in a
continuously flooded paddy (Somenahally et al., 2011). There is a 13%
increase in grain yield over the conventional cultivation method and,
importantly, As concentrations in grain and straw are decreased by
62% and 86%, respectively (Talukder et al., 2011). A glasshouse study
in Bangladesh indicates that the use of AWD techniques increases the
grain yield at around 40% in As contaminated soils over water manage-
ment options (Islam et al., 2 015a). These literatures provide proof of the
concept that water management can be a highly effective tool in con-
trolling As bioavailability in paddy soils and the subsequent accumula-
tion in rice grains. Arsenic concentrations in rice grains are also
cultivar dependent and influenced by water management (Hua et al.,
2011). Therefore, the selection of less As-responsive rice cultivars and
the use of saturated water management in paddy fields could be an ef-
fective means of minimizing As accumulation in rice grains, thus reduc-
ing health risks of As exposure.
4. Human health exposure to As through rice
4.1. Rice consumption and exposure to As
Drinkingwater was considered the major exposure route of As but a
number of studies have established that rice is also a significant expo-
sure pathway of daily As intake (Halder et al., 2012; Jackson et al.,
2012; Mondal and Polya, 2008; Rahman et al., 2011, 2009; Williams et
al., 2006). Rice can be a major food source in the daily diet of the
Asian population and can reach 67% of the daily Asian diet (Adomako
et al., 2011; Brandon et al., 2014; EFSA, 2009; Li et al., 2011). Fig. 4 illus-
trated the process of As contamination in rice and the human health
risk. Recently, dietary exposure studies have been reported in many
countries. The estimated daily intake varies significantly from country
to country. In Asian countries, the consumption of rice is very high at
around 588 g day
−1
(Correll et al., 2006; FAOSTAT, 2007) and it contrib-
utes a relatively high amount of As compared to other foods (Schoof et
al., 1999). Depending on the age group, according to EFSA, rice
accounted for 5% to 15% of the dietary intake of inorganic As (EFSA,
2014). According to Mondal et al. (2010) rice is the major route of inor-
ganic As exposure in some blocks of West Bengal India and rice itself
147S. Islam et al. / Environment International 96 (2016) 139–155
may contribute 12 to 34% of the total As. The regulation sets out limits
for the adult population at 0.20 mg kg
−1
for white rice, and
0.25 mg kg
−1
for brown rice. Based on EFSA (2014) concentration
data 0.089 mg kg
−1
contribute 0.465 μgkg
−1
BW day
−1
As in high
rice consumers (300 g raw rice).
Rice grown in Bangladesh, the world's hot spot for As poisoning, con-
tains about80% inorganic As, and people there eat 450 g day
−1
(Potera,
2007). But the average daily rice consumption for adult ranges from 400
to 650 g (Ahsanand Del Valls, 2011; Rahman et al., 2006). This is one of
the highest per capita rice consumption figures in the world. Consump-
tion of As-contaminated rice may contribute as much as 60% of the daily
dietary As intake based on conservative As concentrations in riceof Ban-
gladesh (Meharg, 2004). When rice ingestion is assumed at
0.432 kg day
−1
for a 60 kg adult, assuming bioavailability of 90%
when cooked (Juhasz et al., 2006), then a grain inorganic As level of
0.053 mg kg
−1
and 0.267 mg kg
−1
would equate to 16% and 81% of
the MTDI, respectively. A survey data showed that the average rice con-
sumption rates for adult males and females were 432 and 420 g day
−1
(uncooked basis) and there is no significant difference between adult
males and females and this consumption rate may contribute 0.049
and 0.047 mg inorganic As day
−1
for adult males and females, respec-
tively (Rahman et al., 2011). Ohno et al. (2007) reported that adult
males and females consume 776 and 553 g of cooked rice daily, based
on their study from Bangladesh and rice contributes 56% of the total
As intake which is greater than the total As intake from drinking
water (13%). Another study from Bangladesh showed that adults
(male or female) consumed 495 g day
−1
of rice and this rice and
water consumption contribute 0.888 and 0.706 mg inorganic As day
−1
for adult males and females, respectively (Rahman et al., 2009). Fig. 5 in-
dicates the daily intake of As from food (rice and vegetables) and drink-
ing water. According to Smith et al. (2006b) rice and drinking water
contribute 1.186 mg As daily in the study of Bangladesh. The intake of
As through rice as a food source was 0.20 to 0.35 mg day
−1
per adult
(Rahman et al., 2008b). In India and Bangladesh the consumption of
420 g rice combined with 2 L of drinking water containing 0.05 mg of
As L
−1
, would equate to a dietary exposure for the people of India and
Bangladesh of 0.121 and 0.155 mg As day
−1
, respectively (Williams et
al., 2005). Roychowdhury et al. (2003) reported 0.693 and 0.560 mg
As daily intake for adult males and females, respectively in the
Murshidabad district in West Bengal India and the daily intake of total
As from water and food was 4.5 times higher than the tolerable daily in-
take with food itself contributing 27%.
The mean daily intake of inorganic As in a Cambodian study was
0.00187mgkg
−1
BW and none of the individuals exceeded the
BMDL
0.5
of 0.003 mg kg
−1
BW from rice consumption alone (Gilbert
et al., 2015). In the Chinese population, the daily intake of inorganic As
was approximately 0.042 mg day
−1
, with rice asthe largest contributor
of inorganic As, accounting for approximately 60% of the total As (Li et
al., 2011). But in Taiwan the average intake value was 0.079 to
0.104 μgkg
−1
BW day
−1
(Chen et al., 2015). A study by Torres-
Escribano et al. (2008) in Spainshowed that the ricewith the highest in-
organic As (0.253 mg kg
−1
) would contribute 0.148 mg inorganic As in-
take from rice which is 99.2% of the TDI value. In Finland, the estimated
inorganic As intake from long-grain rice and rice-based baby foods were
close to the lowest BMDL
0.1
value of 0.0003 mg kg
−1
BW day
−1
set by
the European Food Safety Authority (EFSA) for every age group
(Rintala et al., 2014). There are average rice consumption figures of
15 g day
−1
in France, 24 g day
−1
in the United States (US), and even
218 g day
−1
in China by comparison (Meharg et al., 2009). In Brazil
the inorganic As intake from rice consumption was estimated at 10%
of the PTDI values with a daily consumption 88 g of rice (Batista et al.,
2011).
4.2. Rice processing and their effect on As contents
Traditionally rice is cooked with a substantial amount of excess
water. Arsenic content in cooked rice is the main determinant in case
of actual dietary exposure. Many studies have reported a significant de-
crease of As in brown andpolished rice after washing several times. Re-
cently, a study by Naito et al. (2015) reported that washing three times
may reduce 71–83% of the total and inorganic As in polished rice.
Sengupta et al. (2006) reported that washing of long grain rice 5 to 6
times may remove 28% of total As whereas washing long grain white
rice 3 times removed 8–17% of total As (Mihucz et al., 2007). It differs
significantly from that of raw rice depending on the As content in
cooking water and processing methods (Pal et al., 2009; Rahman et al.,
Fig. 4. Schematic illustration of Ascontamination in rice and the human health risk.
148 S. Islam et al. / Environment International 96 (2016) 139–155
2006). The effect of cooking on speciation of As in grain lacks clarity
given widely different results in the literature. Several reports indicate
that the differences in As contentin raw and cooked ricewould be either
rice varietal differences or concentration of As in cooking water (Huq
and Naidu, 2003; Meharg and Rahman, 2003). The type of rice, percent-
age of water absorbed and the manner of preparation also affects the
total As in cooked rice (Ackerman et al., 2005), however the duration
of cooking also may affect the total As concentration in cooked rice
(Rahman et al., 2006). In fractions of parboiled and non-parboiled rice
grains the order of As concentrations was; rice husk Nbran brown
rice Nraw rice Npolish rice (Rahman et al., 2007c). From a household
survey processing of rice (parboiling and milling) reduced As concen-
trations by an average of 19% (Duxbury et al., 2003).
Absorption of As-contaminated water during the rice cooking pro-
cess may significantly increase the amount of As in cooked rice
(Ackerman et al., 2005), which is often overlooked when calculating
daily As intake values. Cooking rice in excess water efficiently reduces
the amount of inorganic As in the cooked rice on an average by 40%
from long grain polished, 60% from parboiled and 50% from brown
rice (Gray et al., 2015). Another report from Rahman et al. (2006) indi-
cates that when rice is cooked with excess water and the gruel was
discarded, the concentration of As in cooked rice decreased. They attrib-
uted this tothe release ofwater soluble As during cooking at high tem-
perature and decantation of cooking water after cooking. This was
further confirmed by cooking rice with limited water and the gruel
(cooking water) was absorbed in rice. This led to a slight increase in
As concentration in cooked rice (Rahman et al., 2006). The mean As con-
centration in 14 raw Aman rice samples was 0.153 mg kg
−1
with a
range of 0.074–0.302 mg kg
−1
but after cooking the mean value is
0.139 mg kg
−1
so there was no significant difference in the mean As
concentrations between raw and cooked rice as As free pond water
was used for cooking rice (Rahman et al., 2011).
Research shows that the cooking of rice with water does not make
significant changes in total and inorganic As contents. On average 87%
of inorganic As is present in cooked rice (Smith et al., 2006a)but
Ohno et al. (2007) reported that it ranges from 97 to 102%. Pal et al.
(2009) reported that cooking procedures did not change the nature of
As in cooked rice. They showed that 95% of the recovered As present
in cooked rice is inorganic when rice from contaminated areas and tra-
ditional cooking methods were used. On the other hand, when white
rice was cooked without washing, the As levels in raw and cooked rice
were almost the same and this indicate that total As and inorganic As
in rice was decreased by washing only (Naito et al., 2015). Sengupta et
al. (2006) showed that total As content in cooked rice depends strongly
on the method of cooking when cooked with low As water. There was
up to a 57% reduction of total As from rice by subsequent washing 5 to
6 times and high-volume (water: rice = 6:1) cooking followed by
discarding excess water. Raab et al. (2009) reported that decanting of
excess water after high-volume cooking effectively decreases total As
and inorganic As in rice. In general the larger the water:rice cooking
ratio, the greater the removal of more inorganic As. At a water-to-rice
cooking ratio of 12:1, 57 ± 5% of inorganic As could beremoved. Recent-
ly percolating technology proved highly effective in removing inorganic
As from the cooking rice, with up to 85% removed from individual rice
types (Carey et al., 2015).
4.3. Bioavailability and bioaccessibility of As from rice grains
Knowledge of As bioavailability is important to minimize the uncer-
tainty in the risk of As. Bioavailability is often used as the main indicator
of the potential risk that chemicals pose to the environment and human
health (Naidu et al., 2008). The objectives of all bioavailability studies
are to obtain the best possible estimate of the amount of available As
that poses a potential risk to human health. However, most of the
methods suffer from draw-backs, e.g. if the systems are artificially con-
structed and may not always simulate actual human physiological con-
ditions.In an attempt to overcome these limitations, human volunteers
or experimental animals have been used (Caussy, 2003). There are a
limited number of in vivo studies using animals but they are almost all
soil related with incidental high levels of As used (Bradham et al.,
2011; Brattin and Casteel, 2013; Denys et al., 2012; Laird et al., 2013).
Arsenic is relatively rapidly excreted from the body primarily through
the urine (Cascio et al., 2011) and most of the inorganic As, As(III) and
As(V), are metabolized to DMA and MMA prior to excretion through
the urine. Therefore, monitoring the urinary excretion of As is one of
the easy ways to determine bioavailability. Direct measurement of
human exposure from rice has been assessed by urine sampling. There
is only one, limited rice eating trial where rice consumption was mea-
sured and urinary excretion monitored in humans (Meharg and Zhao,
2012).
Naidu and his co-researchers assessed As bioavailability in cooked
rice (glasshouse grown rice and supermarket bought rice) using an in
vivo swine model (Juhasz et al., 2006).Arsenic speciation shows that su-
permarket bought rice contains 100% inorganic As whereas 86% of the
glasshouse grown rice contains dimethylarsinic acid (Juhasz et al.,
2006). Only 33 ± 3% of As was bioavailable in greenhouse grown rice
cooked in water contaminated with sodium As(V) as the low absolute
bioavailability of dimethylarsinic acid whereas in supermarket bought
rice, As bioavailability was high (89 ± 9%). The bioavailability using in
vivo swine models showed that the bioavailability in rice is highly de-
pendent on As speciation, whichin turn can vary depending on rice cul-
tivars, As in irrigation water and cooking water. These findings suggest
that rice genotypes rich in DMA are likely to pose less risk from expo-
sure to As and subsequent toxicity.
The bioaccessibility of As from cooked rice with an in vitro dynamic
digestion process and parboiled rice, which is most widely consumed in
South Asia, showed a higher percentage of As bioaccessibility (59% to
99%) than non-parboiled rice (36% to 69%) and most of the As bioacces-
sible in thecooked rice(80% to 99%) was easily released during the first
2hofdigestion(Signes-Pastor et al., 2012). The estimation of the As in-
take through cooked rice based on the As bioaccessibility highlights that
a few grams of cooked rice (b25 g dry weight per day) cooked with
highly As-contaminated water is equivalent to the amount of As from
Fig. 5. Intake of Asfrom food (rice and vegetables) and drinking water per day.
(adopted from Rahman et al., 2013)
149S. Islam et al. / Environment International 96 (2016) 139–155
2 L of water containing the maximum permissible limit
(0.0010mgAsL
−1
). With As(V) and DMA being the predominant spe-
cies present in cooked rice, very similar trends in the bioaccessibility of
these As species were observed as upon digestion of the pure As stan-
dards range between 85% and 90% at pH 1.8 (Alava et al., 2013). The
use of an enzymatic approach indicates that on average 94% of As is lib-
erated during enzymatic extraction (Ackerman et al., 2005). The mean
As concentration is 0.275 ± 0.161 mg kg
−1
(n= 31) and rice samples
with relatively high total As (N0.20 mg kg
−1
,n= 18) has shown bioac-
cessibility ranging from 53% to 102% (He et al., 2012). Laparra et al.
(2005) evaluated the bioavailability of inorganic As in cooked rice to
better understand thepossible health risks derived from the consump-
tion of rice. The contents of total As in raw rice (0.05–0.53 μgg
−1
) in-
creased considerably after cooking (0.88–4.21 μgg
−1
), as a
consequence of the presence of As(V) in the cooking water. The total
As content of the bio-accessible fraction (1.06–3.93 μgg
−1
) demon-
strates the high bio-accessibility of As from cooked rice (N90%)
(Laparra et al., 2005).
4.4. Associated health risk from the ingestion of rice
It is evidentfrom existing literature thatthe presence of As in the en-
vironment has adverse consequences to human health. Arsenic is a
chronic carcinogen and exposure to elevated levels causes to serious ill-
ness including different types of cancers (IARC, 2004a, 2004b). Inorgan-
ic As in drinking water has been studied and is linked to human
carcinogenesis and exposure. Fairly constant levels of a mixture of As
metabolites are generally excreted in urine, i.e. 10–30% inorganic As,
10–20% MMA, and 60–70% DMA (Vahter, 1999) and it is associated
with various internal cancers viz. liver, bladder, kidney, and lungs as
well as other health problems, including skin cancer and diabetes
(Guo et al., 1997; IARC, 2004a). Inorganic As is metabolized by consecu-
tive reduction and oxidative methylation in the liver and is largely ex-
creted via urine (Yamauchi et al., 1989). This process is considered to
be a detoxification mechanism because the major methylated metabo-
lites MMA and DMA are easily excreted and are less acutely toxic than
the inorganic species (Gebel, 2002; Thomas et al., 2001). In a clinical
study of 18,000 persons in Bangladesh and 86,000 persons in West Ben-
gal India from As-affected areas showed that of these, 3695 (20.6% in-
cluding 6.11% children) in Bangladesh and 8500 (9.8% including 1.7%
children) in West Bengal had arsenical dermatological features
(Rahman et al., 2001). A large portion of the total population is highly
vulnerable to various internal cancers. However, the exposure time to
develop arsenicosis varies from case to case reflecting its dependence
on As level in drinkingwater and food, nutritional status, genetic variant
of the human being and other compounding factors (Anawar et al.,
2002).
Arsenic contamination of rice grains and the low concentration of
micronutrients in rice have been recognized as a major concern for
human health (Lombi et al., 2009; Williams et al., 2009). Therefore,
rice itself plays a potential exposure route to humans but there is very
limited human data as evidence for this impact on humans directly
from rice. Banerjee et al. (2013) reported that cooked rice with
0.2 mg kg
−1
As was associated with genotoxic effects measured in
micronuclei in urothelial cells. A study from India showed that the
chronic daily intake (CDI) and health risk index (HRI) were N1forrice
indicating the potential health risk from the middle Ganga plains of
India (Kumar et al., 2016). Several studies reported the incremental life-
time cancer risk from the consumption rice. Excess lifetime cancer risks
for rice consumption by country is presented in Table 4. The US EPA has
an upper limit of acceptable risk for cancer from any given source of 1 in
10,000 (Kavcar et al., 2009; Tsuji et al., 2007). This is a useful figure with
which to consider cancer risks from inorganic As of rice. Tsuji et al.
(2007) also calculated that at the 95% 6.1 μgday
−1
inorganic As inges-
tion rate for rice, at a slopeof 1.5 mg kg
−1
BW day
−1
, for a 65 kg person,
this equates to an excess skin cancer risk of 1.4 in 10,000 and using the
slope of 3.67 per mg kg
−1
day
−1
, this equates to an excess cancer rate of
3.4 in 10,000. The probabilistic risk assessment model calculated medi-
an increased lifetime cancer risk due to cooked rice intake was 7.62 per
10,000 and was attributable to eating rice and rice contributes about
44% to this median risk (Mondal and Polya, 2008). The risk levels of As
from rice for the people of Bangladesh, China and India had median
risks of 22, 15 and 7 in 10,000, respectively (Meharg et al., 2009). A
study from rural Bangladesh showed that cooked rice is the second
highest contributor towards the adverse health risk with an overall
risk 7.59 × 10
−4
(Khan et al., 2012). In the adult Chinese population,
the incremental life-time risk of cancer from food intake was 106 per
100,000 (Li et al., 2011). Chen et al. (2015) reported that the incremen-
tal lifetime cancer risks for the Taiwan population were 7.9 and 10.4 per
100,000 forfemales and males, respectively. Meanwhile, the non-cancer
health hazard index for the daily intake of rice showed that the toxic risk
due to As was 7.8 times greater than the reference dosage (Lee et al.,
2008). Thus, rice is a major potential source and exposure pathway of
As in the As affected areas where people are exposed to very low levels
of As via drinking water.
5. Concluding remarks
Naturally occurring As has been detected in many parts of the world,
and populations of south-east Asian countries are adversely affected
through drinking water and food. During the last 50 years there has
been a major shift in the food dietary intake by people largely related
to global migration that has seen many Asians relocate to Western
countries —both as students and as professionals. This has resulted in
asignificant change in food dietary intake with rice now being a staple
food for half of the world's populations. As a consequence, consumption
of inorganic As through rice poses significant health risks including can-
cers to humans. Paddy rice accumulate As from the soil and irrigation
water and the ingestion of this As-contaminated rice acts as an exposure
route to humans in any stage of life. Arsenic also has a negative effect on
the growth, yield and quality of rice. Recognition of the poisoning of
thousands of people in Bangladesh, West Bengal, India and parts of
China, has led to a substantial progress in the understanding of As bioac-
tivity, their exposure pathways, mechanism of bioaccumulation,
Table 4
Excess lifetime cancer risks from rice consumption by country.
Country Polished rice
consumption
(g day
−1
)
Median inorganic
As content of rice (mg kg
−1
)
Country specific rice inorganic
As intake (mg day
−1
)
Country specific rice excess cancer
rate per 10,000
References
Bangladesh 445 0.081 0.036 22.10 Meharg et al. (2009)
China 218 0.109 0.024 15.20 Meharg et al. (2009)
India 192 0.059 0.011 6.90 Meharg et al. (2009)
–– – 7.62 Mondal and Polya (2008)
Italy 17 0.071 0.001 0.70 Meharg et al. (2009)
Taiwan –– – 1.04 Chen et al. (2015)
US 24 0.088 0.002 1.30 Meharg et al. (2009)
–– 0.006 1.4–3.4 Tsuji et al. (2007)
150 S. Islam et al. / Environment International 96 (2016) 139–155
speciation pattern and their toxicity levels in soil-water-plant systems.
Arsenic speciation draws major attention because toxicity totally de-
pends on the speciation pattern and it varies from variety to variety
and location to location. However, environmental variation is more
prominent than genotypic variation. Potential strategies are required
to minimize As exposure from water, soil and food. Different manage-
ment practices such as varietal selection, fertilizer and nutrient manage-
ment, irrigation water management and cooking practices showed
significant variation on As bioaccumulation, speciation and their bio-
availability. There is an urgent need to develop a mitigation strategy
to minimize As accumulation in rice grains. So, it is crucial to reduce
As transfer from soil to rice grains to reduce human exposure risk
from rice. Because As intake from rice represents an important route
of exposure, especially for people consuming a large amount of rice in
their diet, mitigation measures to reduce As accumulation in rice are ur-
gently needed through genetic variation or identification of rice varie-
ties which accumulate low As and effective water management
strategies. Current knowledge of As contamination including bioavail-
ability and bio-accessibility should be used to minimize As exposure.
Acknowledgements
The first author thanks the Global Centrefor Environmental Remedi-
ation (GCER), Faculty of Science and Information Technology, University
of Newcastle, Australia for the scholarship support. Financial support
from CRC CARE is greatly acknowledged.
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