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Agriculture, Ecosystems and Environment
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Mechanism of arsenic uptake, translocation and plant resistance to
accumulate arsenic in rice grains
Lalith D.B. Suriyagoda
, Klaus Dittert
, Hans Lambers
Faculty of Agriculture, University of Peradeniya, Peradeniya, Sri Lanka
Department of Crop Science, Section of Plant Nutrition and Crop Physiology, University of Göttingen, Carl-Sprengel-Weg 1, 37075 Göttingen, Germany
School of Biological Sciences and Institute of Agriculture, The University of Western Australia, 35 Stirling Hwy, Crawley, Perth, WA 6009, Australia
A global data analysis shows that rice grain arsenic (As) concentrations increase with increasing soil As con-
centrations until about 60 mg As kg
soil and then decreases. Of the total grain As, 54% is composed of in-
organic As. Therefore, when considering the WHO-permissible grain inorganic As concentration, i.e. 0.2 mg As
, the permissible grain total As concentrations is 0.37 mg total As kg
grain. Soil total As concentration
when grain total As concentration reaches permissible level is 5.5 mg As kg
soil. Therefore, the suitable soil As
concentrations for screening rice cultivars in rice agroecosystems for As resistance is 5–60 mg As kg
has traits to reduce uptake and translocation of As to grains. Cultivars with higher root porosity, radial oxygen
loss, or formation of iron plaques bind more As to iron plaques, reducing As uptake (i.e. As avoidance). Once
taken up, glutathione/glutaredoxin-mediated As reduction, and phytochelatin-dependent complexation and
sequestration in vacuoles result in less translocation of As to the grain. Moreover, generation of reactive oxygen
species and the production of antioxidant enzymes further reduce As toxicity (i.e. As resistance). These resistance
mechanisms in rice agroecosystems are further enhanced when adequate concentrations of silicon and sulfur are
present in soils and tissues, and when plants are associated with arbuscular mycorrhizal fungi, particularly under
aerobic or intermittent-aerobic soil condition. Therefore, As concentrations in rice ecosystems decrease in the
order of: roots > leaves > grains, and in grains: hull > bran polish > brown rice > raw rice > polished
rice > cooked rice. Within the grain, As speciation is aﬀected by the location in the grain, forms of As species,
the grain-ﬁlling stage, geographic origin, ecosystem management and cultivars used. Indica type accumulates
more As in their grains than japonica type. Rice grain production, within safe limits of As, requires the con-
sideration of soil As dynamics including soil management, cultivar responses including uptake and translocation,
and post-harvest processing techniques.
Human exposure to arsenic (As) results from several pathways such
as drinking water, food, beverages, soil, inhalation of dust and atmo-
spheric particulates (Bhattacharyya et al., 2003; Kapaj et al., 2006; Kar
et al., 2006; Nriagu et al., 2007; Nath et al., 2008; Naidu and
Bhattacharya, 2009; Chatterjee et al., 2010). However, consumption of
rice is the primary source of As for humans in a non-seafood diet,
especially in the tropics (Lee et al., 2008; Torres-Escribano et al., 2008;
Halder et al., 2014). In populations not suﬀering from elevated As in
drinking water, chronic exposure to inorganic As may also occur
through the consumption of As-contaminated rice including baby food
containing rice (Williams et al., 2007; Lee et al., 2008; Meharg et al.,
2008; Halder et al., 2014; Lai et al., 2015; Signes-Pastor et al., 2016a).
The average daily consumption of rice by an Asian adult varies in the
range of 200–600 g, depending on the region (Duxbury et al., 2003;
Rahman et al., 2008; Zavala and Duxbury, 2008; Zhu et al., 2008;
Garnier et al., 2010), and in Ghana in Africa it is 32–232 g (Adomako
et al., 2011). Due to this variation in rice intake and in the con-
centration of As in rice, the potential daily intake values of As by an
adult is also highly variable among studies and regions, e.g., 19.6 μgin
India (Kumar et al., 2016), 100–350 μg in Bangladesh (Rahman et al.,
2008; Panaullah et al., 2008), and 69 μg in Cambodia (Phan et al.,
2014). Some of these estimates are much higher than the approximate
As intake through drinking water of 2 L per day at the acceptable WHO
limit of 10 μgL
inorganic As. It has also been reported from the
Bengal delta, India that daily dietary intake of inorganic As by an adult
from rice is 2.32 μgAskg
body wt. day
and this is more than the
Received 10 June 2017; Received in revised form 20 October 2017; Accepted 21 October 2017
Corresponding author at: Faculty of Agriculture, University of Peradeniya, Peradeniya, Sri Lanka.
E-mail addresses: email@example.com,firstname.lastname@example.org (L.D.B. Suriyagoda).
Agriculture, Ecosystems and Environment 253 (2018) 23–37
0167-8809/ © 2017 Elsevier B.V. All rights reserved.
WHO recommended potential daily intake value of 2.1 μgAskg
(Roychowdhury, 2008). Moreover, daily total As intake per
body weight reported for Cambodia was 1.46 μgkg
body wt. day
(Phan et al., 2014). Similarly, daily total As intake of 0.002,
body wt. day
in France (Jitaru et al., 2016), and 0.34,
body wt. day
in China for children and adults, re-
spectively have also been reported (Huang et al., 2013).
Carbonell-Barrachina et al. (2012) analysed the inorganic As con-
centration in food items for infants available in the markets of Spain
which were mainly manufactured in China, and found that approxi-
mately 23% of the pure infant rice samples showed inorganic As con-
centrations over 150 μgkg
(i.e. Chinese limit- USDA, 2006). When
daily intake of inorganic As by infants (4–12 months) was estimated
and expressed on a body-weight basis (μgd
body wt.), it was
higher for all infants aged 8–12 months than drinking water maximum
exposures predicted for adults (10 μgL
standard), indicating the in-
fants are at a higher risk of As toxicity than adults. Similarly, elevated
inorganic and total As concentrations in rice and rice-based infant food
were reported through market surveys conducted in Australia, Finland,
Kingdom of Saudi Arabia, Sweden and United Kingdom (Meharg et al.,
2008; Ljung et al., 2011; Rintala et al., 2014; Shraim, 2017). Epide-
miological studies show that chronic As poisoning can cause serious
health eﬀects including cancers, melanosis (hyper-pigmentation or dark
spots and hypo-pigmentation or white spots), hyper-keratosis (skin
hardening), restrictive lung disease, peripheral vascular disease (black
foot disease), gangrene, diabetes mellitus, hypertension and ischemic
heart disease (Mandal and Suzuki, 2002; Chatterjee et al., 2010; Maity
et al., 2012). Because of these health risks associated with As, it is
important to study the dynamics of As in soil and rice plants including
grains aimed at reducing As accumulation in soil, uptake by rice plants
and translocation to rice grains. Due to the risks associated with As in
rice grain, the maximum acceptable total As level in rice grains was
1 mg total As kg
grain (Sauvé, 2014). This level was based on pre-
vious risk assessments across a wide range of food products. However,
several recent studies on the levels of As in rice and rice-based products
have provided suﬃcient evidence and data to question this overly
generous threshold, speciﬁcally with regards to rice. Therefore, this
limit was recently revised to 0.2 mg inorganic As kg
Soil and grain As concentrations reported in this review cannot be
generalised as average or representative values for geographic regions.
They should be considered as upper limits that can be expected at As-
contaminated sites, as most of the experiments were conducted in ﬁelds
with high concentrations of As, or soils used to conduct pot experiments
were selected from areas with inherently high soil As concentrations or
contaminated with As to test speciﬁc hypotheses (Table S1). A similar
caution pertains to the data on daily As intake by humans in diﬀerent
regions. Moreover, as presented in Table S1, to date, there is an
abundant amount of literature on ‘As dynamics in rice’including As
translocation and speciation in rice plants, and molecular and bio-
chemical responses of rice to As toxicity generated through experiments
conducted with variable complexities from simple hydroponic experi-
ments to ﬁeld and market surveys across the globe.
2. As uptake by rice
The uptake of As by rice from soil strongly depends on the quantities
and speciation of As in the rhizosphere (Marin et al., 1992, 1993).
Under conditions of submerged rice cultivation in anaerobic soil, ar-
senate (As(V)) is reduced to arsenite (As(III)) in the soil solid phase,
followed by desorption of the latter from soil minerals due to its lower
sorption capacity as compared with As(V) (Yamaguchi et al., 2014).
Arsenite is taken up through a subclass of aquaporins (nodulin 26-like
intrinsic proteins: NIPs), and then enters the stele through the silicon-
(Si) uptake pathway (Ma et al., 2008; Panda et al., 2010). Among these
proteins, some members are Si transporters that load As(III) into the
xylem or secrete As(III) from the roots (Zhao et al., 2010). The NIP gene
family consists of 10–13 genes in rice (Forrest and Bhave, 2007) which
can be subdivided into three groups, NIP I, II and III on the basis of their
selectivity (Ma et al., 2008; Zhao et al., 2010). Aquaporin Lsi1 trans-
ports As(III) (Bienert et al., 2008; Ma et al., 2008), and also mediates
silicic acid inﬂux. It is expressed outer side of the plasma membranes of
the exodermis and endodermis cells where Casparian strips are formed.
Arsenite is an uncharged molecule with a diameter of approximately
411 pm (Ma et al., 2008), similar to silicic acid, and can be taken up by
transporters of silicic acid, but not by aquaporins in general. Another
silicic acid transporter also transporting As(III), Lsi2, is expressed at the
inner side of the plasma membranes of the exodermis and endodermis
cells, releasing As(III) into the cortex and into the stele (Ma et al.,
2008). In the cytosol, As(III) reacts with sulfhydryl groups of proteins,
aﬀecting many biochemical functions (Tripathi et al., 2012; Kumar
et al., 2015). Within the cell, As(III) can be detoxiﬁed via a reaction
with phytochelatins, which results in As(III)–phytochelatin complexes
that are ultimately sequestered in vacuoles (Briat, 2010). The transport
of that complex across the tonoplast is believed to be mediated by a C-
type ATP-binding cassette (ABC) transporter (Song et al., 2010, 2014),
which may therefore be of paramount importance for As resistance in
plants (Briat, 2010; Sanglard et al., 2016). Arsenite uptake follows
Michaelis-Menten kinetics, with a K
in the range of 3.7-22.9 μM As(III)
(Abedin et al., 2002; Chen et al., 2005).
Among diﬀerent forms of As in soil, As(V) is the predominant
phytoavailable form in aerobic soils; like phosphate, it strongly sorbs
onto mineral soil components (e.g., iron (hydr)oxides) (Takahashi et al.,
2004). It is an analogue of phosphate and taken up by the high-aﬃnity
phosphate uptake system (Ullrich-Eberius et al., 1989; Meharg and
Macnair, 1990; Zhao et al., 2010), and believed to be loaded into xylem
vessels by phosphate transporter (PHT) proteins (Zhao et al., 2010; Wu
et al., 2011). Recently, Begum et al. (2016) showed that As stress causes
a consistent decrease in tissue P concentration and expression of
phosphate transporters (OsPT8,OsPT4,OsPHO1;2) under both high and
low P supply to rice cultivar BRRI 33. Moreover, a simultaneous in-
crease in phytochelatin synthase (OsPCS1) expression and phytoche-
latin concentration in rice roots were also observed under As exposure.
Overall, the results suggests that As stress down-regulates phosphate
transporters, and enhances phytochelatin-mediated As sequestration to
vacuoles in root cells, limiting As translocation to shoots.
The organic species, dimethyl-arsenic acid (DMA) and monomethyl-
arsenic acid (MMA), are taken up at a much slower rate by the root than
inorganic As, due to the lower aﬃnity of transporters for organic As
(Abedin et al., 2002; Raab et al., 2007; Abbas and Meharg, 2008). MMA
uptake is also partly mediated by the silicic acid transporter Lsi1, while
the speciﬁc transport pathways of DMA is not yet clear (Li et al., 2009;
Carey et al., 2011). It was previously thought that plants are able to
methylate As to produce various forms of organic As (Nissen and
Benson, 1982; Wu et al., 2002), but recent studies have cast doubt on
this. Plants such as rice appear to lack the ability to methylate As, but
instead take up methylated As from the soil (Jia et al., 2013; Lomax
et al., 2012; Zhao et al., 2013). The magnitude of As uptake by rice
varies in the order of As(III) > As(V) > DMA > MMA (Raab et al.,
2007; Finnegan and Chen, 2012). This is further aﬀected by soil and
crop management strategies, and the activity of microorganisms, as we
will discuss below.
3. Translocation of As
3.1. Translocation of As from root to shoot
All major forms of inorganic (i.e. As(III) and As(V)) and organic (i.e.
MMA and DMA) As can be translocated from roots to shoots via the
xylem. Based on xylem sap analysis, Seyﬀerth et al. (2011) concluded
that oxidised As species are dominant in the xylem (86% as As(V) and
14% as DMA), whereas reduced species (71% as As(III), 29% as As tris-
L.D.B. Suriyagoda et al. Agriculture, Ecosystems and Environment 253 (2018) 23–37
) dominate in vacuoles of cells adjacent to the
xylem. Supporting this contention, using X-ray absorption spectroscopy
imaging, Smith et al. (2008b) demonstrated that As(V) is transported
from the root to shoots via the xylem. Once translocated to the shoot, As
(V) is taken up by shoot cells via phosphate transporters (Zhao et al.,
2010; Punshon et al., 2017).
After transport into the cells, As(V) is quickly reduced to As(III) or
other forms (Pickering et al., 2000; Carey et al., 2011). As(III) can be
complexed by phytochelatins, followed by sequestration into vacuoles
(Bleeker et al., 2006; Raab et al., 2007; Zhao et al., 2010). Phytoche-
latins are heavy-metal-binding peptides derived from glutathione (GSH)
and the synthesis of phytochelatins is induced by heavy metals or me-
talloids such as As. Song et al. (2010) identiﬁed phytochelatin trans-
porters responsible for transporting chelated As(III) into vacuoles in
Arabidopsis thaliana, but the exact transporters in rice are unknown.
This sequestration of As(III)-phytochelatin complexes into vacuoles
predominantly occurs in root cells and to a smaller extent in cells of
stems and leaves. This can be considered as an adaptation strategy in
rice plants to minimise translocation of As to grains.
In addition to complexation and sequestration in vacuoles in the
roots, As(III) is also transported to the shoot via the xylem (Carey et al.,
2010; Su et al., 2010; Zhao et al., 2010; Ren et al., 2014). Once in the
shoot, As(III) is taken up through aquaporins in leaf cells (Zhao et al.,
2010; Punshon et al., 2017).
Comparing inorganic and organic forms of As, the amounts of or-
ganic As species taken up by roots are much smaller, and is partly due
to the lower abundance in soil. However, organic forms of As are
translocated more readily within the rice plant than their inorganic
counterparts (Raab et al., 2007; Abbas and Meharg, 2008; Li et al.,
2009; Carey et al., 2010). Relatively good translocation of DMA to the
shoot may be due to its poor SH (sulfhydryl) coordination as a result of
altered molecular structure, in contrast to that of As(III) (Raab et al.,
2007). Therefore, the major forms of organic As taken up by rice, MMA
and DMA, can be translocated to shoots via the xylem (Zhao et al.,
2010). Then, at the shoot level, similar to As(III), MMA and DMA are
taken up through aquaporins into leaf cells (Zhao et al., 2010).
Therefore, overall the results suggest that, despite roots and shoots
having mechanisms to accumulate As in their vacuoles in the form of
phytochelatin-associated As(III) complexes, a fraction of As is taken up
and translocated into grains. The amounts and forms of As translocated
into grains greatly depends on the forms and concentrations of As
present in soil, forms and rates of As taken up by rice roots, capacity of
the rice cultivar to- reduce As (when taken up in oxidised forms), make
complexes with phytochelatins, sequester into cell vacuoles, and xylem
3.2. Translocation of As to rice grains
In a study of shoot-to-grain translocation of As species in excised
panicles, it was found that phloem transport accounted for 90% of As
(III) transport to the grain, and for 55% of DMA translocation (Carey
et al., 2010, 2011). This indicates that inorganic As is predominantly
translocated via the phloem, while DMA is translocated via both
phloem and xylem (Carey et al., 2010). It has also been found that a
larger fraction of ﬂag leaf DMA and MMA is translocated to rice grain;
As(V) is poorly translocated, and is rapidly reduced to As(III) within
ﬂag leaves and As(III) displays no translocation (Williams et al., 2005;
Norton et al., 2009, 2012; Carey et al., 2010, 2011). Therefore, grain As
accumulation results mainly from phloem transport (Carey et al., 2010,
2011; Song et al., 2014). However, to date, the identities of transporters
that move As species out of the phloem and into the grain are largely
unknown (Punshon et al., 2017).
4. As in grain
Based on a compilation of data across diﬀerent countries and types
of experiments, we derived a generic relationship for soil total As
concentration and grain total As concentration (Fig. 1A). For producing
Fig. 1A, soils artiﬁcially treated with As for experimental purposes, and
data generated from pot and ﬁeld experiments were included. To date
there is no critical limit of soil As concentration identiﬁed for many rice
producing soils across the world for safe rice production, except for a
few instances such as 15 mg As kg
of rice soil in Japan (Kitagishi and
Yamane, 1981; Japan, 2016), and 3.9 mg As kg
of agricultural soils
in general in Thailand (Punshon et al., 2017). However, as reported in
Fig. 1A, soil As concentrations used in diﬀerent studies ranged from
very low values to 120 mg As kg
of soil, allowing us to explore the
potential variation of grain As concentration to a large range of soil As
concentrations, irrespective of the soil, management and cultivar
characteristics. The main conclusions that can be drawn from this re-
(i) All rice varieties exhibited less than 1 mg total As kg
(which was the permissible limit of total As in rice grains (Sauvé,
2014)), when grown in rice soil with As concentrations below 15 mg As
(i.e. the maximum acceptable levels of As in Japanese rice soils
(Kitagishi and Yamane, 1981; Japan, 2016)). However, along with the
recent introduction of a new permissible level of inorganic As in rice as,
0.2 mg inorganic As kg
rice grain (WHO, 2016), and based on the
generic relationship that we developed for inorganic and total As con-
centration in rice (i.e. slope = 0.54; Fig 1B), the permissible level of
total As in rice grain possibly should be reduced to 0.37 mg total As
grain. Accordingly, the permissible level of total soil As con-
centration to produce rice grains within the new limits of inorganic and
y = 0.5399x
R² = 0.8309
Soil total As (mg kg
0 20 40 60 80 100 120
Fig. 1. Relationships between (A) grain total As concentration and soil total As con-
centration, and (B) grain inorganic As concentrations and grain total As concentration.
Data were compiled from diﬀerent experiments conducted using pot and ﬁeld trials as
stated in Tables 1 and S1. Permissible limit of total As in rice grains until 2016 was 1 mg
total As kg
grain (Sauvé, 2014) and the corresponding soil total As concentration was
15 mg As kg
(i.e. the maximum acceptable levels of As in Japanese rice soils (Kitagishi
and Yamane, 1981; Japan, 2016)). However, along with the recent introduction of a new
permissible level of inorganic As in rice of 0.2 mg inorganic As kg
rice grain (WHO,
2016), and based on the generic relationship presented in this study ( B with a
slope = 0.54), the permissible level of total As in rice grain possibly should be reduced to
0.37 mg total As kg
grain. Accordingly, the permissible level of total soil As con-
centration to produce rice grain within the new limits of inorganic and total As should be
reduced to 5.5 mg As kg
L.D.B. Suriyagoda et al. Agriculture, Ecosystems and Environment 253 (2018) 23–37
total As should be reduced to 5.5 mg As kg
soil DW. Therefore, ir-
respective of the rice variety, country of origin or management, rice
grain can be expected to have inorganic As concentrations below the
WHO-permissible limit for human consumption, when the soil total As
concentration is less than 5.5 mg kg
. While the critical soil As con-
centration identiﬁed in this study (i.e. 5.5 mg kg
) can be used as a
benchmark, it is advisable for the development of local critical upper
limits of soil As concentrations considering location speciﬁc cultivar,
soil and management. Therefore, those values may either be higher or
lower than the 5.5 mg kg
identiﬁed in this study.
(ii) A few rice varieties still exhibit less than 0.37 mg total As kg
of grain, even when grown in soils with greater than 5.5 mg As kg
soil. These varieties may have mechanisms to take up and/or partition
less As to grains, while the opposite would be the case with rice vari-
eties exhibiting high grain As concentrations.
(iii) As expected, when soil As concentrations increase, the varia-
bility of grain As concentrations also increases, indicating that plants
induce their As resistance mechanisms when exposed to soils with high
concentrations of As.
(iv) Certain rice plants tend to accumulate 2 mg total As kg
or more when grown in the range of 30–60 mg As kg
of soil, and then
grain As concentrations decline. This may be due to the inhibition of
plant growth and impaired cellular functions at higher soil As con-
centrations. Therefore, when conducting ﬁeld/pot experiments aiming
at the study of rice varietal responses such as mechanism of As re-
sistance and screening of rice germplasm to As sensitivity, growth
medium As concentration would need to be greater than 5 mg As kg
of soil. However, extremely high As concentrations may also cause
detrimental eﬀects on crop growth and development.
(v) The derived relationship between grain total As concentration
and soil total As concentration is highly scattered, particularly at higher
concentrations of soil total As. This may be due to the highly variable
buﬀering capacity of As in soils tested across experiments, and thus the
bioavailable As concentration. Despite the high relevance of bioavail-
able As concentrations in soils for plant uptake, most of the published
studies only analysed total soil As concentrations. Therefore, it is re-
commended to correlate the bioavailable As fraction in soil with grain
As concentration to draw ﬁrmer conclusions.
When the data from soil that was artiﬁcially enriched with As were
excluded (i.e. only the soils with their natural As concentrations were
considered), and a weighted average was calculated using the mean and
sample size data from the original studies, the mean grain As con-
centration for all countries was far below the previously considered
permissible limit of 1 mg total As kg
of grain (Table 1). Therefore, in
general rice from all countries was considered as safe to be consumed
by humans with respect to As concentration; however, depending on
the region and variability in soil As concentration, higher ranges of
grain As concentrations can be observed (Table 1). With the recent
introduction of new permissible inorganic As concentration in rice grain
as 0.2 mg inorganic As kg
rice grain (WHO, 2016), the new per-
missible total As concentration in rice grains suggested in this study is
0.37 mg total As kg
rice grain; a value lower than the previously
considered permissible level. Therefore, rice produced in certain re-
gions with high soil As concentration contributes to increase grain As
concentrations above the safe limits for human consumption as pre-
sented in Fig. 1A and Table 1.
Though shoot and grain As concentrations increase with increasing
soil As concentration, the grain/shoot As concentration ratio declines
with increasing of soil As concentration (Williams et al., 2007;
Adomako et al., 2009). This is due to a decline in As concentration from
roots > straw > husk > grains (Rahman et al., 2007, 2008, 2009;
Dahal et al., 2008; Roychowdhury, 2008; Smith et al., 2008a; Zhiyan
et al., 2008; Lu et al., 2010; Rahaman et al., 2011; Hsu et al., 2012;
Sarkar et al., 2012; Ye et al., 2012; Bhattacharya et al., 2013; Kar et al.,
2013; Ren et al., 2014). The greater storage of As in roots and lower
translocation to grains can be explained by the reduction of As(V) to As
(III) in roots, complexation with thiols, and sequestration in vacuoles
adjacent to the xylems (Zhao et al., 2009; Seyﬀerth et al., 2011).
Therefore, the As concentration in the grain is partially governed by
cultivar-speciﬁc characteristics such as the capacity to take up, trans-
locate, methylate, and store As in vacuoles in addition to soil char-
acteristics (Norton et al., 2009; Ahmed et al., 2011; Ye et al., 2012).
Speciation of diﬀerent As forms in rice grain is aﬀected by diﬀerent
factors as discuss below.
In rice grains, the As concentration decreases in the order of
hull > bran polish > brown rice > raw rice > polish rice (Rahman
et al., 2007, 2008). Moreover, As is not evenly distributed within those
tissues, e.g., using micro-synchrotron-radiation based methods, Kramar
et al. (2017) observed that As has the highest concentrations in the
embryo and some hot spots in the coating (up to 13 mg kg
fore, polished rice available in the market contains lower concentra-
tions of As than the full grain and is safer for human consumption in
terms of As availability.
Rice grains contain both inorganic and organic As species. Based on
a range of studies, basically from Asia, it has been found that As forms
in grain can vary in the order of inorganic As > organic As (Table 1).
Among inorganic forms, As(III) > As(V), while among organic forms,
DMA > MMA (Williams et al., 2005; Meharg et al., 2008;
Roychowdhury, 2008; Lu et al., 2010; Huang et al., 2012; Meharg and
Zhao, 2012; Halder et al., 2014; Ma et al., 2014; Hu et al., 2015).
Overall, As(III) and DMA are the predominant species in rice grain, with
As(V), MMA and occasionally, diphenylarsenic acid, methylpheny-
larsenic acid, tetramethylarsonium being minor species (Arao et al.,
2011; Huang et al., 2012; Meharg and Zhao, 2012). Using 121 com-
mercially purchased rice samples from the markets and supermarkets in
China, Taiwan, Japan, Germany and Switzerland, belonging to 12 rice
types (i.e. long grain, whole grain, japonica, indica, risotto, round-
sticky, long-sticky, basmati, jasmine, red, black and wild rice), Huang
et al. (2012) found that, on average, As(III) accounts for 90% of in-
organic grain As, regardless of geographic origin, rice type, grain size,
cultural practice and polish treatment. However, it is important to note
that all the results discussed above were generated using rice cultivars
grown in the Asian region. Contrasting results have also been reported
for rice cultivars grown elsewhere, and the organic forms of As were
either more abundant or similar to the inorganic forms of As in the rice
grain of some cultivars (Meharg et al., 2008; Carbonell-Barrachina
et al., 2012; Sommella et al., 2013; Moreno-Jiménez et al., 2014;
Rowell et al., 2014)(Table 1). Similarly, Smith et al. (2008a) found that
DMA comprised 85 to 94% of the total As concentration in the grain,
while the rest was mainly contributed by As(III). Syu et al. (2015) also
observed similar results in Taiwan, while variable inorganic and or-
ganic As forms in the range of 40–60% of total As were reported by Pan
et al. (2014) in China. Overall, the results indicate signiﬁcant en-
vironmental and/or management eﬀects on grain As speciation in ad-
dition to cultivar diﬀerences. Moreover, when As(III) was delivered to
excised rice panicles at 13.3 μM, As(III) was oxidised to As(V) within
the rice grain, while under exposure of 133 μM, As(III) remained stable
within the grain as free As(III) (Carey et al., 2010), indicating that the
rates of As(III) supply to rice grain and/or oxidation-reduction reactions
may also determine As speciation. Despite the diﬀerences in inorganic
and organic fractions of As presented among studies, when the global
data are arranged to study a generic relationship between the inorganic
and total As concentration in rice, a strong positive relationship is found
(Fig. 1B). According to this relationship, irrespective of the cultivar,
environment or management, an increase of 0.54 μg of inorganic As
occurs per 1 μg increase in total As concentration in rice grain.
L.D.B. Suriyagoda et al. Agriculture, Ecosystems and Environment 253 (2018) 23–37
An interaction between the location in the grain and form of As has
also been reported, i.e. inorganic As is located in the bran layer of rice
grain including ovular vascular trace (OVT), while DMA is found
throughout the outer layers into the endosperm (Sun et al., 2008; Carey
et al., 2010). The reason for this diﬀerence in location may be the
mobile nature of DMA in rice grains, while MMA and inorganic As
(mainly AS(III)) are less mobile, and, thus, remain close to the entry
point (OVT). It has also been found that most of the As in the bran
layers is present as oxidised As (69–88% as As(V) and 12–31% as DMA),
and in the embryo as a mixture of As(V) and As(III), whereas it is non-
detected in the endosperm (Seyﬀerth et al., 2011). Therefore, peeling
and polishing of rice before consumption would remove some of the As,
in particular more toxic inorganic forms, but not all of the As.
Arsenic speciation in rice grain is strongly inﬂuenced by soil
ﬂooding and addition of organic manure (Xu et al., 2008; Arao et al.,
2009; Li et al., 2009; Norton et al., 2013; Hu et al., 2015). Compared
with ﬂooding, aerobic treatments markedly decreases organic As
(mainly DMA), and thus increases the percentage of inorganic As in
grain, although the concentrations of inorganic As remain lower than
those in ﬂooded rice (Xu et al., 2008; Li et al., 2009; Hu et al., 2015).
For example, in aerobically grown rice, inorganic As comprises 88%
and DMA 11% of total As. Interestingly, under ﬂooding conditions,
inorganic As contributes only 38% and DMA 62% of the total As (Hu
et al., 2015). Norton et al. (2013) also observed an increase in organic
forms of As in the grain with an increase in the organic matter per-
centage in the growth medium. When comparing rice samples collected
from the markets of Brazil which were grown either organically or
using conventional methods, Segura et al. (2016) observed no diﬀer-
ence in total As concentration between the two types of rice (i.e.
147 mg As kg
). However, the inorganic As concentration was 45%
higher in rice grown organically than in that grown in a conventional
system. The observed diﬀerences in the percentage of organic and in-
organic As forms in rice grains between the two cultivation systems
would be due to the diﬀerences in nutrient management and associated
processes. Therefore, special care is required when irrigation methods
and organic matter addition are employed to mitigate As accumulation
Grain arsenic (As) characteristics reported for diﬀerent countries in a range of publications Note: Experimental data collected from diﬀerent countries and market data (only when the
country or origin was reported) were used.
Country Mean (μgkg
), and sample
size (in parentheses)
Inorganic As fraction, and
sample size (in parentheses)
Australia 120 (7) 30–188 –Williams et al. (2006);Smith et al. (2008a);Rowell et al. (2014);Shraim (2017)
Bangladesh 200 (316) 80–700 0.64 (92) Meharg and Rahman (2003);Islam et al. (2004);van Geen et al. (2006);
Williams et al. (2006);Ohno et al. (2007);Hossain et al. (2008);Rahman et al.
(2006, 2008, 2011);Sun et al. (2008);Meharg et al. (2009);Garnier et al.
(2010);Stroud et al. (2011);Talukder et al. (2011, 2012);Kippler et al. (2016)
Brazil 190 (73) 5–782 0.47 (60) Segura et al. (2016);Ciminelli et al. (2017)
Cambodia 201 (11) –0.83 (5) Phan et al. (2014)
Canada 65 (2) 20–110 0.61 (4) Heitkemper et al. (2001);Williams et al. (2005)
China 133 (2033) 54–1160 0.61 (197) Xie and Huang (1998);Cheng et al. (2004);Zhu et al. (2008);Huang et al.
(2006);Qin et al. (2006);Williams et al. (2006);Liu et al. (2005, 2007);Fu et al.
(2008);Sun et al. (2008);Mei et al. (2009);Meharg et al. (2009);Williams et al.
(2009);Lu et al. (2010);Qian et al. (2010);Fu et al. (2011);Wu et al. (2011);
Carbonell-Barrachina et al. (2012);Ye et al. (2012);Huang et al. (2013);Ma
et al. (2014);Pan et al. (2014);Hu et al. (2015);Geng et al. (2017)
Ecuador 86 (53) 42–125 0.87 (9) Otero et al. (2016)
Egypt 52.8 (118) 50–102 –Meharg et al. (2009);Rowell et al. (2014);Shraim (2017)
Finland 253 (8) 110–360 0.74 (8) Rintala et al. (2014)
France 283 (39) 183–535 0.86 (4) Meharg et al. (2009);Shraim (2017);Jitaru et al. (2016)
Ghana 110 (7) –0.83 (1) Adomako et al. (2011)
Hungary 139 (3) –– Mihucz et al. (2007)
India 146 (374) 41–995 0.71 (165) Roychowdhury et al. (2002); Sangupta et al. (2006); Roychowdhury. (2008);
Williams et al. (2005);Mondal and Polya (2008);Bhattacharya et al. (2010,
2013);Meharg et al. (2009);Pal et al. (2009);Stroud et al. (2011);Sarkar et al.
(2012);Halder et al. (2014);Rowell et al. (2014);Shraim (2017);Kramar et al.
(2017);Kumar et al. (2016);Jitaru et al. (2016)
Italy 188 (161) 81–283 0.52 (19) D'Ilio et al. (2002);Williams et al. (2005);Meharg et al. (2009);Rowell et al.
(2014);Jitaru et al. (2016)
Japan 264 (79) 117–445 0.76 (19) Narukawa et al. (2008);Meharg et al. (2009);Matsumoto et al. (2015)
Korea 247 (50) 90–410 –Lee et al. (2008);Kwon et al. (2017)
Nepal 180 (75) 60–330 –Dahal et al. (2008)
Pakistan 122 (15) 85–147 –
Philippines 70 (1) –– Williams et al. (2006)
Portugal 224 (23) 118–421 0.39 (23) Signes-Pastor et al. (2016b)
Spain 171 (376) 50–547 0.44 (328) Williams et al. (2005);Torres-Escribano et al. (2008);Meharg et al. (2009);
Carbonell-Barrachina et al. (2012);Moreno-Jiménez et al. (2014);Signes-Pastor
et al. (2016b)
Sri Lanka 122 (154) 12–540 –Rowell et al. (2014);Jayasumana et al. (2015)
Suriname 215 (2) 139–290 0.57 (2) Shraim (2017);Jitaru et al. (2016)
Taiwan 78 (233) 50–760 0.62 (31) Schoof et al. (1999);Lin et al. (2004);Jiang et al. (2014)
Thailand 146 (81) 100–241 0.77 (13) Williams et al. (2005, 2006);Meharg et al. (2009);Adomako et al. (2011);
Rowell et al. (2014);Shraim (2017);Jitaru et al. (2016)
Turkey 98 (25) 20–171 –Gunduz and Akman (2013)
UK 238 (22) 120–470 0.53 (27) Meharg et al. (2008);Carbonell-Barrachina et al. (2012)
Uruguay 215 (1) 0.57 (1) Jitaru et al. (2016)
USA 153 (375) 66–816 0.46 (36) Schoof et al. (1999);Williams et al. (2005);Heitkemper et al. (2001);Williams
et al. (2005);Meharg et al. (2009);Adomako et al. (2011);Somenahally et al.
(2011);Carbonell-Barrachina et al. (2012);Zavala et al. (2008);Rowell et al.
(2014);Shraim (2017);LaHue et al. (2016);Jitaru et al. (2016)
Vietnam 190 (2) 169–210 –Phuong et al. (1999);Rowell et al. (2014)
L.D.B. Suriyagoda et al. Agriculture, Ecosystems and Environment 253 (2018) 23–37
in the grain of rice grown in soils polluted with As.
Arsenic speciation of rice plays a pivotal role in environmental
evaluation when the impact of As toxicity on human health is assessed.
In general, the toxicity of As increases with decreasing oxidation states.
Meanwhile, the inorganic form of As is more toxic than the organic
form (Smith et al., 2008a, 2008b; Zavala et al., 2008; Huang et al.,
2012; reviewed by Zhao et al., 2013). As(III) and DMA are the major
inorganic and organic As forms in rice grains, respectively. Due to its
lower toxicity, the predominance of DMA in rice grains may be less of a
concern. However, precautions should still be required to reduce its
presence in rice, because the reduced trivalent form of this dimethyl-
arsenical is highly toxic (Styblo et al., 2000). In general, apart from the
genetic characteristics, growth environment and management condi-
tions such as quality of soil and irrigation water, method of irrigation
water application, use of fertilisers and organic manure and their
quality, and processing techniques can determine the quality of rice
grain available in the market with respect to As concentration and
Zheng et al. (2011) examined the temporal variation in As con-
centration in rice plants at diﬀerent stages of growth, and found that
unloading of As into the grain diﬀers for inorganic As and DMA, with
higher percentage of caryopsis As present as DMA before ﬂowering, and
this declines from ﬂowering to maturity. Therefore, the concentration
of DMA in grain decreases progressively with its development, due to
dilution, whereas the inorganic As concentration is stable (Zheng et al.,
2011). However, their results were based on a single rice cultivar and
this response likely varies among cultivars and with management
practices, and thus requires further attention.
Total grain As concentration (i.e. sum of both inorganic and organic
As forms) varies signiﬁcantly among countries, reaching even up to a
ﬁvefold diﬀerence (Table 1, between Egypt and France or Suriname).
Moreover, the minimum and maximum As concentrations reported
across countries varies over 15-fold (Table 1, between Sri Lanka and
France), and 11-fold (Table 1, between Egypt and China), respectively.
Apart from the diﬀerences between countries, a two- to 20-fold diﬀer-
ence in total grain As concentration even within a country can also be
observed (Table 1, for Egypt and China). Similar observations were
previously made by diﬀerent researchers (Huang et al., 2006; Qin et al.,
2006; Zhu et al., 2008; Meharg et al., 2009; Adomako et al., 2009,
2011; Fu et al., 2011). Diﬀerence in total grain As concentration be-
tween countries or studies within the same country can at least be
partly due to the diﬀerences in environmental conditions, in addition to
cultivar and management.
Despite the diﬀerence in grain total As concentration within and
among countries, there is substantial variation in As speciation among
rice produced in diﬀerent geographical regions (Zhao et al., 2013). For
example, when soil–shoot–grain transfer of As is considered, As is more
eﬃciently transferred to grains by rice cultivars grown in Asia than by
cultivars grown in EU and USA (Adomako et al., 2009). Also, rice
produced in Africa and Asia generally contains a high proportion of
inorganic As, whereas rice produced in the USA, Australia and Europe,
e.g., Italy and Spain, tends to have a high proportion of organic As
(Williams et al., 2005;Smith et al., 2008a;Zavala et al., 2008;Zhu
et al., 2008;Meharg et al., 2008; , 2009; Adomako et al., 2011;
Carbonell-Barrachina et al., 2012;Sommella et al., 2013;Zhao et al.,
2013;Moreno-Jiménez et al., 2014;Rowell et al., 2014)(Table 1).
Because of these diﬀerences, Zhao et al. (2013) found that rice pro-
duced in Asia shows a strong linear relationship between inorganic As
and total As concentrations with a slope of 0.78, while that in Europe
and the USA shows a more variable, but generally hyperbolic
relationship, with DMA being predominant in USA rice. In Australia,
DMA comprises 85 to 94% of the total As concentration in rice grain,
while the rest is mainly As(III) (Smith et al., 2008a). The generic re-
lationship developed for global data in this study shows that 54% of
grain total As comprises inorganic As (Fig. 1B). The reasons for this
geographical variation are not clear, but it may be related to the dif-
ferences in microbial community composition as methylated As species
are supposed to be generated in soil by the activity of microbes, and
aerobic rice culture tends to contribute more inorganic forms of As in
rice grain (Xu et al., 2008; Arao et al., 2009; Li et al., 2009; Norton
et al., 2013; Zhao et al., 2013; Hu et al., 2015). It may also be associated
with the diﬀerence in rice cultivars used in diﬀerent regions of the
world as we will discuss below. However, detailed investigations to
identify the reasons behind this diﬀerence in As species in rice grain are
5. Variation in resistance of rice cultivars to As
There is a signiﬁcant genotypic variation in As accumulation in rice
(Zhang and Duan, 2008; Norton et al., 2009; Tuli et al., 2010; Adomako
et al., 2011; Lou-Hing et al., 2011; Bhattacharya et al., 2013). For ex-
ample, Norton et al. (2009) showed a four–to ﬁve-fold variation in
grain As concentration among 76 rice varieties grown in Bangladesh.
Moreover, high-yielding hybrid rice varieties accumulate more As than
local rice varieties used in experiments conducted in West Bengal,
India, and Bangladesh (Bhattacharya et al., 2013). These As-accumu-
lating rice varieties even exceed the WHO-permissible grain As limit.
Therefore, growing rice cultivars that accumulate higher concentrations
of As in their grains should be avoided, especially during winter (Boro)
rice cultivation, as winter cultivation depends on groundwater con-
taminated with high concentration of As for irrigation in Bangladesh
and India. When comparing major rice types, Indica types accumulate
more As in their grains than Japonica types (Ye et al., 2012; Syu et al.,
2015), and less than the hybrids (Wu et al., 2016). In Sri Lanka, tra-
ditional rice cultivars accumulate less As in their grains than the higher-
yielding cultivars bred by the rice research stations in the country
(Jayasumana et al., 2015).
The causes of any diﬀerences in As accumulation in the grain among
rice cultivars are still not fully understood. Understanding the me-
chanisms underlying reduced As accumulation and the genes re-
sponsible for this is important in directing future breeding for cultivars
that can grow in high-As conditions, while showing low As concentra-
tion in the grain (Lou-Hing et al., 2011). There are a number of pro-
posed mechanisms for diﬀerential As accumulation in various organ-
isms. Methylation of As into less toxic forms (MMA and DMA) has been
shown in microbes and mammals (Challenger, 1945; Cullen and
Reimer, 1989; Qin et al., 2006). However, methylation has not been
described as a detoxiﬁcation pathway in plants. Rather, plants take up
methylated As forms generated by soil microorganisms. Metabolic
pathways for As detoxiﬁcation are via reduction, eﬄux from the cell, or
transport into vacuoles. Therefore, As accumulation and speciation in
rice grain among diﬀerent rice cultivars grown at diﬀerent levels of As-
contaminated soils, and the mechanisms associated with reduced As
uptake and accumulation in rice grains are important to be studied.
6. Inﬂuence of As on metabolic processes in rice
In spite of many studies on the eﬀects of As stress, the precise mo-
lecular mechanisms related to both the eﬀects of As phytotoxicity and
the defence reactions of plants against As exposure are not yet fully
understood (Chakrabarty et al., 2009). When rice is grown in As-con-
taminated soil, roots take up As, and at high internal concentrations, As
may cause toxicity to plant (Shri et al., 2009). Plants are in general
sensitive to As, showing decreased germination percentage, shoot and
root elongation, photosynthesis, growth and absorptive functions
(Abedin and Meharg, 2002; Shri et al., 2009). Roots are the ﬁrst barrier
L.D.B. Suriyagoda et al. Agriculture, Ecosystems and Environment 253 (2018) 23–37
to As translocation to the aerial parts. Recently, the toxicity threshold of
As(III) on intact rice seedlings was determined by Hoﬀmann and
Schenk (2011), who found the As(III) toxicity threshold of rice seed-
lings was 2.4 μM (AsIII), which reduced growth by 10%. As(III) con-
centrations greater than 21 μM induced wilting of seedlings, and re-
duced water uptake, transpiration rate and net photosynthetic rate. It is
suggested that this is due to increasing As(III) binding at the outer side
of the plasmalemma to proteins such as aquaporins impairing uptake of
water. Moreover, As(III) concentrations exceeding 5.3 μM in the solu-
tion also reduced the tissue concentrations and contents of Si, P, K, Cu,
Fe, Mn and Zn (Hoﬀmann and Schenk, 2011). Therefore, with in-
creasing As concentrations in the growth environment, even at very low
concentrations and within a narrow range, a wide range of cellular
processes are aﬀected including root elongation, photosynthesis, water
relations and nutrient uptake. Reduced tissue concentrations of Fe and
Mn would be due to binding of As with Fe and Mn oxides in the rhi-
zosphere, while a decrease in tissue concentrations of P and Si may be
due to competition with As at the uptake sites (Finnegan and Chen,
2012; Meharg and Meharg, 2015; Suda and Makino, 2016). Arsenite
binds to sulfhydryl groups, disturbing protein/enzyme structure and
functioning (Finnegan and Chen, 2012), including functioning of
aquaporins (Tyerman et al., 1999; Wan et al., 2004), reducing water
uptake (Hachez and Chaumont, 2010). Arsenate exerts its toxicity by
replacing phosphate in target molecules, such as DNA, RNA, and other
phosphorylated metabolites; inhibiting ATP synthesis; and targeting
important phosphorylation reactions in primary metabolism (Meharg
and Hartley-Whitaker, 2002;Tripathi et al., 2007; Finnegan and Chen,
There is signiﬁcant evidence that inorganic As exposure results in
the generation of reactive oxygen species (ROS) associated with the
conversion of As(V) to As(III); this then leads to the synthesis of en-
zymes such as superoxide dismutase (SOD) and catalase (CAT), and
antioxidants such as glutathione to keep ROS under control (Gupta
et al., 2009; Dixit et al., 2015). Reactive oxygen species generated as a
result of oxidative stress has been widely implicated in damage to cell
membranes and DNA (Sharma and Dietz, 2009; Ahmad et al., 2012). In
addition, As(V) may up-regulate genes that encode anti-oxidative en-
zymes (Abercrombie et al., 2008; Norton et al., 2008; Chakrabarty
et al., 2009). Therefore, As interferes with important metabolic pro-
cesses such as those that are related to photosynthesis and respiration,
and, depending on the severity, it can ultimately lead to plant death.
Leaf samples from two rice cultivars grown in hydroponics for two
weeks and then treated either with 0, 50, 150 or 300 μM As(III) for 24
or 96 h were used for RAPD analysis using 11 primers (Ahmad et al.,
2012). The genomic template stability (i.e. changes in RAPD proﬁle) of
the As-treated plants at all As concentrations and time durations was
signiﬁcantly aﬀected. This indicates the potential change in many bio-
chemical and physiological processes of the rice plant due to exposure
to As. However, as mentioned by Hoﬀmann and Schenk (2011),As
concentrations used in this experiment were far above the toxicity
threshold of rice seedlings, and, thus, plants may have undergone
permanent cellular damage. Similar As concentrations were used as the
highest As concentration in most of the hydroponic experiments con-
ducted (Table S1). The use of appropriate As concentrations to observe
gradual changes in molecular responses in rice plants is important.
7. Mechanisms of As resistance in rice
7.1. Aerenchyma, radial oxygen loss and Fe plaque formation
Properties of rice roots in reducing As uptake, despite its im-
portance, have largely been ignored until recently (Deng et al., 2010).
Mei et al. (2009) showed that the oxidising ability of roots inﬂuences As
accumulation in rice. The roots’oxidising ability can be estimated by
measuring radial oxygen loss (ROL), resulting from O
the shoot to the roots, followed by release into rhizosphere (Armstrong,
1980). When rice roots are grown in stagnant conditions, a ROL barrier
and aerenchyma develop (Colmer et al., 2006). Under stagnant condi-
tions, rice plants also produce roots with a larger speciﬁc surface area
(Deng et al., 2010). Conversely, when grown in aerobic soils, rice
possesses a diﬀerent root structure, which is characterised by longer
maximum root length, fewer adventitious roots, lower porosity, lower
speciﬁc root surface area, and a less pronounced barrier against ROL
(Colmer, 2003; Deng et al., 2010). The correlations among As accu-
mulation in grains and straw, rates of ROL, and porosity of roots using
25 rice cultivars were investigated by Mei et al. (2009), who found a
signiﬁcant negative correlations between As in grains or straw, and ROL
and porosity, and signiﬁcant positive correlations between rates of ROL
and porosities. Similar results were also found by Wu et al. (2011).
Therefore, genotypes with higher ROL have a strong ability to reduce As
Fig. 2. Schematic representation of rice plant responses to the exposure of
As in soil and tissues in order to minimise the translocation of As to rice
L.D.B. Suriyagoda et al. Agriculture, Ecosystems and Environment 253 (2018) 23–37
accumulation in shoots, and increase As resistance by reducing As
mobility in the rhizosphere and thus limiting As translocation to the
grain (Mei et al., 2009, 2012; Wu et al., 2011)(Fig. 2).
Due to ROL of the root system, Fe
in the rhizosphere is oxidised,
forming an oxyhydroxide plaque on the root surface (Armstrong, 1964;
Deng et al., 2010). This Fe hydroxide in the soil and solution has a very
strong binding aﬃnity for As (Meng et al., 2002; Liu et al., 2004; Deng
et al., 2010). Recently, Pan et al. (2014) showed that genotypes with
higher ROL oxidise more As(III) in the rhizosphere, and induce more Fe
plaque formation, which subsequently sequesters more As, reducing As
uptake and accumulation in rice grains. Similar results were obtained
by Yu et al. (2016) and Wu et al. (2017). The eﬀect of Fe plaque on As
sequestration and plant uptake is aﬀected by the As concentration in
the soil solution (Deng et al., 2010). Therefore, rice cultivars with high
porosities tend to possess faster rates of ROL, and have greater capa-
cities for the formation of Fe plaques and limiting As uptake and
transfer to aboveground tissues (Fig. 2).
When comparing diﬀerent rice cultivars and growth stages, cultivars
with more ROL tend to have greater eﬀects on rhizosphere Eh, pH,
quotients, Fe plaque formation, and As fractionation and
mobility compared with those with lower ROL (Mei et al., 2012).
Moreover the above eﬀects are more pronounced at the bolting stage
than at the tillering stage (Mei et al., 2012; Yu et al., 2016). However,
Fe-plaque formation on roots of rice is greatest at the tillering stage,
after which it gradually decreases (Mei et al., 2012). Garnier et al.
(2010) also conducted a comprehensive ﬁeld experiment to study the
temporal changes of As and Fe concentrations in soil water and in rice
roots, primarily the Fe plaque surrounding the roots, during the
growing season at two sites irrigated with groundwater containing
As and two control sites irrigated with water con-
taining < 15 μgL
As in Bangladesh. They found that at both sites
irrigated with contaminated water, the As concentration in soil water
increased from < 10 μgL
to > 1000 μgL
during the ﬁrst ﬁve
weeks of the growing season, and then gradually declined to < 10 μ
during the last ﬁve weeks. At the two control sites, concentrations
of As in soil water never exceeded 40 μgL
. At both contaminated
sites, the As concentration in roots and in the Fe plaque rose to
1000–1500 mg kg
towards the middle of the growing season. It then
declined to ∼300 mg kg
towards the end, a level still well above the
As concentration of ∼100 mg kg
in roots and plaques measured
throughout the growing season at the two control sites. The maximum
As concentration of grain observed at the sites irrigated with the con-
taminated water was 0.58 ± 0.05 mg kg
.Garnier et al. (2010) and
Yu et al. (2016) further highlighted that more attention should be paid
in future studies to processes regulating the transfer of As from soil to
the rice plant during the last month of the growing season, as this
period likely has a disproportionate eﬀect on the As concentration of
rice straw and rice grains. Though higher root porosity and ROL results
in the formation of Fe-plaque, accumulating more As in the Fe plaque
and lowering As uptake by rice roots, the bioavailability of As in soil is a
dynamic process with a large variability during the crop growing cycle.
Therefore, further attention should be paid to study the changes in As
uptake, partitioning, ROL/oxidising capacity of roots, degrees of Fe
plaque/Fe redox transformation, As speciation/fractionation and their
inter-relationships in diﬀerent cultivars during the growth cycle of rice.
Despite the processes discussed above, if soil is deﬁcient in Fe, the
formation of a Fe-plaque in the rhizosphere may causes Fe deﬁciency
and produce visible symptoms of Fe-chlorosis in leaves (Pestana et al.,
2003). The availability of Fe can be increased by the application of
chelating ligands to the solution, e.g., hydroxyiminodisuccinic acid
(HIDS), ethylenediaminetetraacetic acid (EDTA), ethylenediaminedi-
succinic acid (EDDS), and iminodisuccinic acid (IDS). Together with the
increased Fe uptake, an increase of the uptake of As in rice has also
been reported (Rahman et al., 2009, 2011). Therefore, the use of Fe to
reduce As uptake by rice through the formation of a Fe plaque has to be
addressed carefully, particularly in Fe-deﬁcient ﬁeld conditions without
making Fe deﬁcient to rice while minimising As uptake.
7.2. Biochemical adaptations
Arsenic resistance mechanisms in plants span a range of strategies
(Ahmad et al., 2012; Tripathi et al., 2012; Gupta and Ahmad, 2014).
Broadly, they can be either to assimilate less As (i.e. avoidance of As
uptake), as discussed in the previous section, and/or to restrict As
translocation to the shoot including seeds. One intrinsic/primary me-
chanism to combat As toxicity once taken up by rice is phytochelatin-
dependent detoxiﬁcation vis-à-vis induction of sulfate uptake and re-
duction pathways (Rausch and Wachter, 2005). Higher levels of phy-
tochelatins and phytochelatin-synthase activity along with coordinated
thiol metabolism were reported in rice which induces As tolerance
(Tripathi et al., 2012). Further, phytochelatin −As(III) complexion in
rice leaves reduces translocation of As from leaves to grains (Duan
et al., 2011)(Fig. 2). In addition, the potential role of various me-
tallothioneins in As detoxiﬁcation in rice has also been reported
(Gautam et al., 2012; Nath et al., 2014).
Glutaredoxin/glutathione/glutathione reductase (GRX/GSH/GR) is
one of the major redox systems involved in maintaining the cellular
redox state (Dubey et al., 2016). Glutaredoxin (GRX) acts as a trans-
hydrogenase or thioltransferase, which belongs to a group of ubiquitous
small heat-stable disulﬁde oxido-reductases with a molecular mass of
10–15 kDa (Fomenko and Gladyshev, 2002). In both prokaryotic and
eukaryotic systems, arsenate reductase (AR) uses the glutathione
(GSH)/GRX system as a reductant for As(V) to As(III) (López-Maury
et al., 2009); however, in the conversion of As(III) to As(V), apart from
the generation of reactive oxygen species (ROS), two electrons are
transferred, one from AR and other from GSH. Recently, Dubey et al.
(2016) investigated the expression of GRX genes under As(III) and As
(V) stress in rice cultivars diﬀering in As sensitivity, and the underlying
variation in detoxiﬁcation mechanism involving antioxidant pools.
They found a higher expression of the four GRX and two glutathione-S-
transferase (GST) genes in the sensitive cultivar than in the resistant
cultivar. This suggests that under As stress, GRX is synthesised more in
the sensitive cultivar than in the resistant one. Also, the expression of
four GRX genes was greater under As(V) than under As(III) stress.
Therefore, the higher As accumulation in the resistant cultivar is due to
a lower GST expression, which can be attributed to the absence of
thiolation and sequestration of As in roots, and translocation of more As
to shoots (Dubey et al., 2016). Hence, the overall resistance of a plant
towards As, would be determined by the eﬃciency of the reduction of
As(V) to As(III), scavenging the ROS generated followed by glutathio-
nylation of As(III) and eventually sequestering the glutathionylated As
(III) in to the vacuole (Fig. 2).
A secondary detoxiﬁcation strategy comprises protection against
oxidative stress by increasing the production of various antioxidant
enzymes such as catalase (CAT), superoxide dismutase (SOD), ascorbate
peroxidase (APX), guiacol peroxidase (GPX), glutathione reductase
(GR) and some other peroxidases (Shri et al., 2009; Rai et al., 2011;
Dave et al., 2013)(Fig. 2). It has also been found that the production
and activity of these antioxidant enzymes increases more in As-tolerant
rice cultivars, allowing them to accumulate more As than susceptible
cultivars do (Rai et al., 2011).
A number of metabolites are also being synthesised on exposure to
As. Speciﬁc amino acids such as proline, histidine, cysteine, glycine,
glutamic acid and methionine, and peptides such as glutathione and
phytochelatins increase in abundance (Dave et al., 2013). These amino
acids aﬀect the synthesis and activity of enzymes, gene expression and
redox-homeostasis related to heavy metal detoxiﬁcation (Rai, 2002).
Though amino acid synthesis related to stress resistance is enhanced,
Dwivedi et al. (2012) reported that high concentrations of soil As de-
crease amino acid concentrations in rice grain, and this response diﬀers
between the essential and non-essential amino acids, i.e. a reduction of
essential amino acid concentration was more pronounced with an
L.D.B. Suriyagoda et al. Agriculture, Ecosystems and Environment 253 (2018) 23–37
increase in grain As concentration than that observed for non-essential
amino acids (Dwivedi et al., 2010). Moreover, this reduction was
greater in As-accumulating rice cultivars (Dwivedi et al., 2012). The
decline in the level of amino acids is associated with an increase in the
levels of RNA, proteins and proline with As treatment of rice seedlings.
Therefore, exposure of rice plants to As may impair hydrolysis of RNA
and proteins due to the inhibition of RNase and proteases activities and
proline accumulation under As toxicity may serve as an enzyme pro-
tectant (Mishra and Dubey, 2006). Because of this, there is a danger in
fulﬁlling recommended daily intake of amino acids through rice with
the accumulation of As in rice grains.
Previous studies revealed novel insights into plant defence me-
chanisms and regulation of genes and gene networks in response to As
exposure. The diﬀerential expression of transcripts encoding glu-
tathione-S-transferases, glutathione synthase, superoxide dismutase,
sulfur metabolism, heat-shock proteins, metal transporters, and en-
zymes in the ubiquitination pathway of protein degradation as well as
several unknown proteins serves as molecular evidence for the phy-
siological responses to As stress (Norton et al., 2008; Chakrabarty et al.,
2009; Rai et al., 2011). Larger numbers of genes were diﬀerentially
expressed following exposure to As(V) when compared with As(III) and
were related to glutathione metabolism, transport, and signal-trans-
duction pathways indicating the occurrence of an As(V)-speciﬁc re-
sponse in rice (Chakrabarty et al., 2009; Ahsan et al., 2010; Rai et al.,
2011; Dubey et al., 2016). Similarly, one glutaredoxin gene
(Os01g26912) is expressed speciﬁcally in the shoot of As(III)-treated
plants, indicating the presence of As(III)- and As(V)-speciﬁc gene re-
sponses in rice (Chakrabarty et al., 2009).
Rai et al. (2011) studied the expression of genes associated with the
sulfur-assimilation pathway when plants were grown at 5 and 25 μMAs
(III). Among the many genes that were up-regulated, two GRX genes,
i.e. Os01g26912 and Os02g40500 were up-regulated when plants were
exposed to As(III) stress, whereas only one gene (Os02g40500) was up-
regulated when plants experienced As(V) stress. Similar up-regulation
was observed by Chakrabarty et al. (2009) in another study for the
sulfate transporter gene Os03g09970 under both As(III) and As(V)
stress. The members of this gene family encode proteins that diﬀer in
their intracellular locations, expression patterns, and kinetic properties
and help in transport of sulfate (Nocito et al., 2006), leading to en-
hanced production of S-rich metal-binding peptides (such as GSH and
phytochelatins) and amino acids (such as cysteine and methionine),
thus providing metal resistance and resulting in metal accumulation
(Mishra et al., 2009).
Protein proﬁles of two-week-old rice seedlings exposed to 0 (con-
trol), 50 or 100 μM As were investigated by Ahsan et al. (2010). A total
of 14 proteins showed reproducible changes in expression of at least
1.5-fold when compared with the control and showed a similar ex-
pression pattern in both treatments. Of these 14 proteins, eight were up-
regulated and six were down-regulated following exposure to As. The
increased expression of several proteins was associated with energy
metabolism. This suggests a higher energy requirement for enhanced
metabolic processes in leaves exposed to As. Similar changes in protein
proﬁle and gene expression in As(V)-resistant and −susceptible rice
seedlings were also observed by Gupta and Ahmad (2014).
Photosynthetic processes can be aﬀected due to the presence of As
in leaf tissues. The ribulose-1,5-bisphosphate carboxylase/oxygenase
(RuBisCO) large subunit was signiﬁcantly decreased under As stress.
Similarly, reduction in chlorophyll concentration, maximum and actual
quantum yields of photosystem II, and the electron transport rate in As-
treated rice plants were also observed (Rahman et al., 2007; Finnegan
and Chen, 2012; Ahmad et al., 2012; Bhattacharya et al., 2013; Gupta
and Ahmad, 2014; de Andrade et al., 2015; Li et al., 2015). Thus, the
down-regulation of RuBisCO and chloroplast 29 kDa ribonucleopro-
teins, chlorophyll concentration, and the reduced water uptake and
conductance at stomatal and mesophyll levels, might have collectively
contributed to a decrease in photosynthetic rate under As stress
(Hachez and Chaumont, 2010; Finnegan and Chen, 2012; Sanglard
et al., 2014). Despite the reduction in leaf photosynthetic rate and ca-
pacity, an increase in the reducing sugar concentrations in rice seed-
lings (Jha and Dubey, 2004; Choudhury et al., 2010), and total and
reducing sugar concentrations and carbohydrate concentration in the
grains (Bhattacharya et al., 2013) treated with As were also observed.
Supporting this contention the activities of sucrose degrading enzymes
viz., acid invertase and sucrose synthase were increased whereas, the
activity of sucrose synthesising enzyme, viz. sucrose phosphate synthase
declined (Jha and Dubey, 2004; Choudhury et al., 2010) leading to the
accumulation of soluble sugars. This may also be associated with a
greater reduction in growth than in photosynthesis itself, as many other
biochemical processes in rice plant are aﬀected by As exposure.
Many questions regarding the functional translated proteins, and
speciﬁceﬀects on the photochemical and gas exchange capacities in
rice in response to As stress remain unanswered. Proteomic technolo-
gies provide one of the best options for the functional analysis of
translated proteins of the genome. Data on genome sequences and in-
ferred protein sequences can be used to identify proteins and to follow
temporal changes in protein expression (Ahsan et al., 2010). Moreover,
further investigation of the redox proteome, phosphorylated proteins
and removal of the most abundant proteins (RuBisCO), which will
improve the number of diﬀerentially expressed low-abundance proteins
in two-dimensional gel electrophoresis, could provide better informa-
tion on the protein networks and on understanding the As-stress re-
sponse in plants.
7.3. Eﬀect of Si in plant tissues
Apart from the interaction of Si in soil in changing the soil As
bioavailability, there is evidence that Si interacts with As once taken up.
In hydroponic culture the negative eﬀects of 25 μM As on photo-
synthesis and carbohydrate status can largely be reversed by the ap-
plication of 2 mM Si (Sanglard et al., 2016)(Fig. 2). However, no major
metabolic reprogramming was observed, as denoted by minor, if any,
signiﬁcant changes in (i) the activities of a range of enzymes associated
with C metabolism; (ii) the levels of a wide range of organic acids and
amino acids; and (iii) the pools of NAD(P)H/NAD(P)
and the redox
states of ascorbate and glutathione. Therefore, the authors suggested
that the search for As-resistant plants under ﬁeld conditions should not
focus solely on oxidative stress, and hence the focus on photosynthesis
might be of greater signiﬁcance. Tripathi et al. (2013) demonstrated
that Si can mediate As(III) resistance by lowering As uptake and im-
proving the antioxidant defence system, whereas Sanglard et al. (2014)
showed that As(III) mediated reduction in stomatal and mesophyll
conductance could largely be reversed by Si.
Silicon can aﬀect inorganic As remobilisation from leaves to grains
(Carey et al., 2011). The Si concentration of rice leaves was strongly
positively correlated with grain As in the ﬁeld, with the As present
predominantly in its inorganic form (Norton et al., 2012). Therefore,
while silicic acid competes with As(III) during uptake by the roots,
plants that are genetically more eﬃcient at assimilating Si will have
greater loading of inorganic As into the grain (Meharg and Meharg,
2015). In summary, although a higher soil Si concentration can reduce
As uptake, a higher tissue Si concentration enhances As loading into
rice grains. Therefore, both soil and varietal characteristics should be
considered when using Si in As-aﬀected rice ecosystems.
7.4. Eﬀect of S in plant tissues
Sulfur plays a crucial role in regulating As tolerance through com-
plexation of As by S-containing ligands; glutahione (γ-Glu-Cys-Gly) and
phytochelatins (Tripathi et al., 2007; Duan et al., 2011; Dixit et al.,
2015). The As-thiol complexes are subsequently transported to vacuoles
and this process is known as As detoxiﬁcation (Song et al., 2010). Up-
regulation of sulfate transporters leads to a continuous supply of S,
L.D.B. Suriyagoda et al. Agriculture, Ecosystems and Environment 253 (2018) 23–37
which can be utilised in chelation and vacuolar sequestration of metals
or metalloids in plants (Dixit et al., 2015)(Fig. 2). Application of S at a
5 mM to the growth medium resulted in an increased root As accu-
mulation and prevented the translocation of As to shoots compared
with that during growth at 0.5 and 3.5 mM S, likely due to As com-
plexation through enhanced synthesis of thiolic ligands. Moreover,
transcript levels of Lsi1 and Lsi2 were reduced in the 5 mM S treatment,
and this reduction in expression levels is probably the reason for low
shoot As accumulation. Similar results were reported by Hu et al.
(2007) and Fan et al. (2013). These responses are supported by the
observation by Duan et al. (2011), as the concentration of phytoche-
latins is negatively correlated to grain As accumulation, indicating a
higher rate of As complexation in roots with the supply of S restricting
the As mobility to rice shoot. Hu et al. (2007) also observed a reduction
in rice shoot As concentration, and in the concentration of Fe in the Fe-
plaque in the rhizosphere with increasing rates of S application. High
concentrations of Fe
usually exist at the site of SO
sulfate-reducing bacteria, and S
thus produced immediately reacts
to form FeS (Murase and Kimura, 1997). Recently, it has
been reported that even in conditions of decreased S availability, plants
continue to rely on thiol metabolism to detoxify As eﬀectively, and
altered subcellular distribution of As contributes only partially
(Srivastava et al., 2016). Hence, the S-assimilation pathway is im-
portant in As detoxiﬁcation in rice.
Foliar sprays with 0.5 mM L-buthioninesulfoxime (BSO, a source of
S) on rice leaves at the grain-ﬁlling stage decreased GSH and phy-
tochelatin accumulation in rice shoots by 40–63% and 20–55%, re-
spectively without aﬀecting plant growth (Duan et al., 2011). More-
over, foliar sprays with BSO decreased shoot As concentrations,
suggesting that manipulation of phytochelatin synthesis may mitigate
As accumulation in shoots. Contrary to these reports, Reid et al. (2013),
who studied interaction of As(III) and S deﬁciency in Hordeum vulgare,
suggested that inﬂux of As(III) is faster than the rate of synthesis of
thiols, and questioned the importance of thiols in preventing As toxi-
city. Therefore, the role of S in As uptake and transfer in rice plants
needs to be studied further. Key areas include the eﬀectiveness of soil-
or foliar-applied S in reducing As translocation to grains, and cultivar
diﬀerences in up-regulating sulfate transporters, phytochelatin forma-
tion and chelation.
8. Mycorrhizas and As
Arbuscular mycorrhizal fungi (AMF) are fungal symbionts with a
profound inﬂuence on plant physiology by improving their nutritional
status and protection against heavy metals and pathogens (Smith and
Read, 2008; Suriyagoda et al., 2014). Plant performance can be
improved by arbuscular mycorrhizal (AM) associations under a range of
biotic and abiotic stresses such as drought, nutrient deﬁciencies or
imbalances, excessive levels of toxic elements or salinity (Pozo and
Azcón-Aguilar, 2007; Suriyagoda et al., 2014; de Andrade et al., 2015;
Albornoz et al., 2017). Under aerobic conditions, rice is able to form
associations with AMF, as is the case for upland or alternate-wetting
and drying (AWD) conditions in low-lands in many regions of the world
(Li et al., 2011, 2013; Chen et al., 2012).
AMF-rice associations allow the transfer of carbon from the plant to
the fungus, and of mineral nutrients from the fungus to the host plant.
Li et al. (2011) have shown that lowland rice cultivar ‘Guangyinzhan’
inoculated with Glomus intraradices and upland rice cultivar ‘Handao
502′inoculated with G. geosporum have a greater As resistance and
grain yield at 60 mg As kg
soil under aerobic conditions (Li et al.,
2011, 2013). Rice exhibiting high radial oxygen loss (ROL) favour AM
fungal colonisation and enhance the ratio of As(III)/As(V) concentra-
tion in their roots in the presence of AMF (Li et al., 2013). AMF play an
important role in decreasing As accumulation by transforming in-
organic As into less toxic organic forms, or diluting As concentration by
enhancing plant biomass (Christophersen et al., 2009; Smith et al.,
2010a, 2010b; Chen et al., 2012 and references therein) (Fig. 2). In an
experiment on the eﬀect of G. intraradices on As(III), As(V), DMA and
MMA uptake by lowland and upland rice, Li et al. (2011) found that AM
reduced the As(V) uptake by the low-aﬃnity P-uptake system, and that
of As(III) and MMA by the high- and low-aﬃnity P-uptake systems.
Several reports demonstrate that the diﬀerential As uptake in AMF
plants is related to the inﬂuence of AM on the modulation of phosphate
transporters in the roots (Christophersen et al., 2009; Smith and Smith,
2011). Supporting this contention, de Andrade et al. (2015) found that
the P/As ratio is higher in the shoots of AMF plants under both As(V)
and As(III) exposure, indicating the preferential uptake of P, rather than
As. Moreover, AM plants could immobilise most As in the roots and
prevent its translocation to the shoots (Li et al., 2011, 2013; Chen et al.,
2012). Overall, the results indicate that AMF in rice reduce As uptake
and translocation to grains (Fig. 2). Moreover, these eﬀects are more
prominent under aerobic soil conditions than in continuously ﬂooded
The inﬂuence of AMF Rhizophagus irregularis on CO
chlorophyll ﬂuorescence, chlorophyll concentration and growth of rice
exposed to As(V) or As(III) was studied by de Andrade et al. (2015).
Arbuscular mycorrhizal plants sustain photosynthetic eﬃciency and
growth even under elevated As concentrations in comparison with non-
mycorrhizal plants (Fig. 3). Similar results were found by Chan et al.
(2013), comparing the growth of an upland rice variety, Zhonghan 221,
in the presence of either a single or a mixture of AMF with non-my-
corrhizal plants. However, they did not ﬁnd signiﬁcant relationships
Fig. 3. Growth of mycorrhizal (M) and non-mycorrhizal (NM) rice
plants in a pot experiment in the absence (C) or presence of 50 μM
arsenate (AsV) or arsenite (AsIII). Growth was greater in M plants than
in NM plants, and the growth reduction was more severe for As
(III) > As(V) > C for both M and NM conditions. (Adapted from: de
Andrade et al., 2015).
L.D.B. Suriyagoda et al. Agriculture, Ecosystems and Environment 253 (2018) 23–37
between AMF colonisation rates and As concentrations in grains.
Therefore, the overall results suggest that the AMF-rice symbiosis shows
promise in ameliorating the reduction in growth and yield when plants
are grown in As-contaminated soils. Moreover, the importance of AMF
in alleviating As toxicity further increases in aerobic and/or AWD rice
cultures, as aerobic rice culture reduces As input and bioavailability in
soil and promotess AMF-rice symbiosis.
Being a Si hyperaccumulator, lowland rice takes up greater amounts
of As than most other plants (Meharg and Meharg, 2015). To in-
vestigate the eﬀect of AMF on the expression of Lsi1 and Lsi2 and on the
accumulation of As(III), rice plants were inoculated with Glomus in-
traradices (AH01) under diﬀerent As(III) concentrations (0, 2 and 8 μM)
(Chen et al., 2012). The relative mRNA expressions of Lsi1 and Lsi2
were less in AM plants. For example, at 2 μM As(III), Lsi1 and Lsi2 were
signiﬁcantly decreased, 30% (P < 0.05) and 50% (P < 0.01), re-
spectively, in AM plants, leading to a decrease of As(III) uptake per unit
of root dry mass in AM plants. Results from a wide range of experiments
conﬁrm the possibility of AMF enhancing As resistance of rice grown in
As-contaminated soils under aerobic or AWD conditions. However, the
AMF-As-rice relationship is highly complex, due to the presence of in-
teracting eﬀects with soil water management, P and Fe nutrition, and
rice cultivar and AMF speciﬁcity. Further studies on these speciﬁc to-
pics are warranted.
9. Grain As bioavailability
The total concentration and proportion of diﬀerent As species in
cooked rice can diﬀer from that in raw rice, and that diﬀerence greatly
depends on the As concentration in raw rice, cooking water and the
processes of cooking (Halder et al., 2014 and references there in). This
shows that the assessment of As exposure by only quantifying As con-
centration in raw rice may provide misleading information on actual
human health risk (Ackerman et al., 2005; Horner and Beauchemin,
Inorganic As is the predominant species in both raw and cooked rice
of most of the rice grown in the world, except for few deviations re-
ported in rice crops grown in USA and some parts of Europe (Table 1).
Also, the As concentration was higher in non-parboiled rice than that in
parboiled rice (Rahman et al., 2007). Moreover, cooking of rice with
water low in As (< 10 μgL
) can further decreases the total and in-
organic As concentration in cooked rice compared with that in raw rice
(Rahman et al., 2006; Rahman and Hasegawa, 2011; O’Neill et al.,
2013), while the opposite response has been observed when using As-
contaminated water for cooking (Roychowdhury, 2008; Kumar et al.,
2016). The As concentration of cooked rice can be reduced if rain water
is used for cooking (O’Neill et al., 2013), and/or when rice is cooked
with excess water that is low in As and the excess water is discarded
after boiling (Sengupta et al., 2006; Mihucz et al., 2007; Jitaru et al.,
It was thought that inorganic As presents a greater risk to human
health than DMA (Zavala et al., 2008), and it was therefore considered
desirable to select or breed rice cultivars with a low total grain As and/
or a low inorganic As proportion. Although originally viewed as a de-
toxiﬁcation product of inorganic As, work done by Kenyon and Hughes
(2001) indicates that DMA has unique toxic properties in itself, indu-
cing an organ-speciﬁc lesion (single-strand breaks in DNA) in the lungs
of both mice and rats and in human lung cells in vitro. Therefore, the
toxicity of diﬀerent As forms should be further evaluated.
Due to the lack of understanding of diﬀerent methods used to esti-
mate inorganic As in rice samples across the world and establishment of
standards, de la Calle et al. (2011) conducted a global analysis com-
piling results obtained with diﬀerent methods adopted in diﬀerent la-
boratories from diﬀerent parts of the world; they concluded that the
concentration of inorganic As determined in rice is independent of the
analytical method that is used in diﬀerent laboratories. Similar results
were recently reported by Huang et al. (2012 and references therein).
The need for developing standards of As in rice is highlighted by several
research groups (Arunakumara et al., 2013; Hite, 2013; Sauvé, 2014;
Lai et al., 2015). As a result, the maximum permissible level for in-
organic As in white/polished rice was identiﬁed as 0.2 mg kg
2016). Along with developing standards, rice and rice-based products in
markets should be intensively studied for their As concentration and
inﬂow of As-contaminated rice should be controlled (Rintala et al.,
2014; Rowell et al., 2014).
When considering diﬀerent types of experiments related to As and
rice interactions, as presented in Table S1, there is a fairly good dis-
tribution of information on the forms of As species studied in the
growth medium and in rice plant, types of studies conducted, duration
of the experiments and the geographic region from which the in-
formation was derived. However, when considering the duration of
plants exposed to As treatments, in over 33% of the experiments
(n = 24) plants experienced As exposure for less than 10 days. This was
particularly so for molecular biological and biochemical analyses con-
ducted using plants grown in hydroponic medium. Such short-term
exposure of plant to As may mainly induce acclimation responses. The
responses that were observed may not show the actual adaptive re-
sponses to As. Moreover, when considering the As concentrations used,
in over 60% of the experiments, the highest level of As was over 25 μM.
According to Hoﬀmann and Schenk (2011) As concentrations above
21 μM in hydroponics cause wilting of rice seedlings and reduce water
uptake, causing detrimental eﬀects to rice seedlings. Though rice cul-
tivars may diﬀer in their response to As concentration and the experi-
mental condition such as the volume of the containers and light in-
tensity, it is pivotal to consider these aspects in future experimentation,
because rice plants in the ﬁeld are highly unlikely to ever experience
such very high As concentration, even in the short term. Therefore,
when making decisions on the factors and the As levels to be studied in
future experimentation the above considerations should receive atten-
10. Concluding remarks
When rice is grown in As-contaminated soil, large amounts of As are
taken up and accumulate in the grain, occasionally exceeding the WHO-
permissible concentrations. This aﬀects a range of cellular processes
and growth. Rice cultivars respond to increasing As concentration by
reducing As uptake through greater radial oxygen loss and formation of
Fe-plaques that bind more As (As avoidance mechanisms), and through
internal detoxiﬁcation mechanisms such as glutathione/glutaredoxin-
mediated As reduction, phytochelatin-dependent complexation and
sequestration in vacuoles and the production of antioxidant enzymes as
tolerance mechanisms to As. These mechanisms vary, dependent on
cultivar × environment × management interactions. Grain As specia-
tion is aﬀected by location in the grain, forms of As species, grain-ﬁlling
stage and geographic origin. In addition, there are cultivar × envir-
onment × management interactions. Improving soil and tissue P, Si
and S nutrition, and associations with AMF decrease As uptake and
translocation to grains. Therefore, multidisciplinary basic and applied
research eﬀorts are required to alleviate As accumulation in rice and
avoid toxicity to humans consuming rice.
Georg Forster Research Fellowship through Alexander von
Humboldt Foundation to the ﬁrst author, and the editor and reviewers
for their valuable comments on an earlier version of this manuscript are
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