Arsenic accumulation in maize crop (Zea mays): a review.

Science of the Total Environment 488–489 (2014) 176–187

Contents lists available at ScienceDirect

Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Review

Arsenic accumulation in maize crop (Zea mays): A review J.M. Rosas-Castor 1, J.L. Guzmán-Mar 1, A. Hernández-Ramírez 1, M.T. Garza-González 1, L. Hinojosa-Reyes 1 Universidad Autónoma de Nuevo León, UANL, Department of Chemistry Sciences, San Nicolás de los Garza, N.L. 66451, Mexico

H I G H L I G H T S • • • •

As(V) is the main chemical form absorbed by the corn and shows the highest toxicity Fe and P play a key role in the phytoavailability and translocation of As to maize High As levels lead effects as oxidative stress and damage to corn plant growth Information of As species in corn grain and corn-based products is limited

a r t i c l e

i n f o

Article history: Received 2 November 2013 Received in revised form 20 April 2014 Accepted 20 April 2014 Available online xxxx Editor: Dr. D. Barcelo Keywords: Arsenic accumulation Corn Cereal crop Phytoavailability Arsenic phytotoxicity

a b s t r a c t Arsenic (As) is a metalloid that may represent a serious environmental threat, due to its wide abundance and the high toxicity particularly of its inorganic forms. The use of arsenic-contaminated groundwater for irrigation purposes in crop fields elevates the arsenic concentration in topsoil and its phytoavailability for crops. The transfer of arsenic through the crops–soil–water system is one of the more important pathways of human exposure. According to the Food and Agriculture Organization of the United Nations, maize (Zea mays L.) is the most cultivated cereal in the world. This cereal constitutes a staple food for humans in the most of the developing countries in Latin America, Africa, and Asia. Thus, this review summarizes the existing literature concerning the conditions involved in agricultural soil that leads to As influx into maize crops and the uptake mechanisms, metabolism and phytotoxicity of As in corn plants. Additionally, the studies of the As accumulation in raw corn grain and corn food are summarized, and the As biotransfer into the human diet is highlighted. Due to high As levels found in editable plant part for livestock and humans, the As uptake by corn crop through water–soil–maize system may represent an important pathway of As exposure in countries with high maize consumption. © 2014 Elsevier B.V. All rights reserved.

Contents 1. 2. 3.

4. 5.

6.

1

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sources of arsenic pollution in maize fields . . . . . . . . . . . . . . Arsenic phytoavailability in agricultural soil for maize crops . . . . . . . 3.1. Effect of iron and manganese . . . . . . . . . . . . . . . . . 3.2. Influence of pH . . . . . . . . . . . . . . . . . . . . . . . 3.3. Effect of phosphate . . . . . . . . . . . . . . . . . . . . . . 3.4. Microorganisms . . . . . . . . . . . . . . . . . . . . . . . 3.5. Organic matter . . . . . . . . . . . . . . . . . . . . . . . Arsenic transfer from the soil to the maize grain . . . . . . . . . . . . Translocation of arsenic in the maize plant . . . . . . . . . . . . . . 5.1. Factors of arsenic mobility in the maize plant . . . . . . . . . . 5.2. Translocation of arsenic species in the maize plant . . . . . . . Arsenic phytotoxicity in maize crops . . . . . . . . . . . . . . . . . 6.1. Mechanism of arsenic toxicity . . . . . . . . . . . . . . . . . 6.2. Toxic effects of arsenic on the nutritional status of the maize plant

E-mail address: [email protected] (L. Hinojosa-Reyes). Tel.: +52 81 8329 4000x3434; fax: +52 81 8352 9025.

http://dx.doi.org/10.1016/j.scitotenv.2014.04.075 0048-9697/© 2014 Elsevier B.V. All rights reserved.

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

. . . . . . . . . . . . . . .

177 178 179 179 179 180 180 180 181 181 183 183 183 183 184

J.M. Rosas-Castor et al. / Science of the Total Environment 488–489 (2014) 176–187

7. Human exposure to arsenic through maize grain 8. Conclusions and remarks . . . . . . . . . . . Acknowledgments . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

1. Introduction Arsenic (As) is a highly toxic element found ubiquitously in nature that represents a potential environmental threat to human, animal and plant health. This metalloid has been classified by the U.S. Environmental Protection Agency (EPA) as a human carcinogen (EPA, 2012). Chronic exposure to a low dose of arsenic can cause skin lesions, neurological defects, atherosclerosis and cancer (Watts et al., 2010). The adverse health effects of As depend strongly on the dose, duration of exposure, and the nutritional status of the exposed population (Markley and Herbert, 2009). The issue of chronic arsenic intoxication appears to be a more important public health problem in certain regions of the world, such as Taiwan, Mexico, India, Germany, Argentina, and Chile (ATSDR, 1989). In terms of toxicity, inorganic forms of arsenic, arsenate [As(V)] and arsenite [As(III)], have a higher degree of toxicity than the organic forms, monomethylarsonic acid [MMA(V)] and dimethylarsinic acid [DMA(V)] (Wu et al., 2011). Although both inorganic arsenic species have been associated with issues related to cancer, As(III) exhibits 10 times more toxicity than As(V). The most common sources of exposure to arsenic in the environment are natural sources, such as volcanic rocks, marine sedimentary rocks, hydrothermal deposits and fossil fuels, such as coal and oil (O'Reilly et al., 2010; Zheng et al., 2011). Anthropogenic activities, including mining, industrial activities and the use of arsenical pesticides, have also contributed to increased arsenic levels in the environment (Páez-Espino et al., 2009; Zhao et al., 2009). Humans can be exposed to As through different pathways, with direct water intake and the As transfer through the crops–soil–water system representing the major pathways (Khan et al., 2009). Arsenic can be absorbed by the root system of plants and vegetables through agricultural soil and irrigation water, thereby gaining access into the food chain. Groundwater with elevated As concentrations has been recognized as a global concern. Currently, Bangladesh and West Bengal—India have the most serious groundwater arsenic problem in the world (Chowdhury et al., 2001; Bundschuh et al., 2012). In countries, such as Argentina, Australia, Chile, China, Hungary, Mexico, Peru, Taiwan, Thailand and the United States of America (Razo et al., 2004; Jasso-Pineda et al., 2006; Armienta and Segovia, 2008; Cao et al., 2009; Deng et al., 2009; O'Reilly et al., 2010; Watts et al., 2010), the total As concentrations (AsT) in groundwater and soil exceeding the limit established by the USEPA (10 μg L−1 and 20 mg kg−1, respectively) have been found (EPA, 2012). High As levels in farming zones can affect plant development and reduce agricultural production and could add substantial amounts of As to the diet intake through agricultural product consumption, thus posing additional risk to human health (Yu et al., 2009). Limited studies (Abedin and Meharg, 2002; Duxbury et al., 2003; Huq et al., 2006; Meharg et al., 2008) have been conducted to evaluate the presence and behavior of arsenic in the food chain. Thus, additional studies on the presence of arsenic in irrigation water, agricultural soil and agricultural products are required to develop a reliable database and to evaluate their interrelationship (Khan et al., 2009). Maize (Zea mays L.) is the most cultivated cereal in the world (FAOSTAT, 2004). The global production of maize according to the Food and Agriculture Organization of the United Nations (FAOSTAT) exceeded 883 million tons per year in 2011, a production greater than the corresponding production for wheat (704 million tons) and rice (723 million tons) (FAOSTAT, 2004). Mexico has one of the highest consumption per capita of maize in the world, with a total production of 0.8 millions of tons in 2011 (FAOSTAT, 2004). According to Food and Agriculture Organization data for the year 2005, per capita maize

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

177

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

. . . .

184 185 186 186

consumptions were 70, 104, and 120 kg in the United States, South Africa, and Mexico, respectively (FAO, 2005) Maize has multiple uses and constitutes a staple food for humans (grain maize) in most developing countries in Latin America, Africa and Asia. Maize has a high nutritive value and this cereal provides in a great number of rural zones approximately 50 and 70% of the protein intake and caloric consumption, respectively (Serna-Saldivar et al., 2008). Maize contains approximately 72% starch, 10% protein, and 4% lipid, supplying an energy density of approximately 365 kcal/100 g. For this reason, maize has become highly integrated into the global agriculture, human diet, and cultural traditions (Nuss and Tanumihardjo, 2010). In Mexico, 60% of the total maize production is used for human consumption, and an average consumption of tortillas (the most common maize-based food) can be as high as 325 g/day per person (Plasencia, 2004). Additionally, in countries, such as the USA and Brazil, maize is an important animal feed (FAOSTAT, 2004). Thus, understanding the distribution and speciation of arsenic within the edible portions of maize is essential for the estimation of the risk presented by arsenic in maize and for the establishment of effective strategies to reduce arsenic concentrations. Regulations for agricultural products in the world are typically related only to the total amount of As; however, the speciation of As is important because the availability, mobility, (phyto)toxicity, and human health outcomes rely on the levels of individual arsenic compounds rather than the total arsenic content (Watts et al., 2010). Arsenic occurs in the natural environment in the following four oxidation states: As(V, arsenate), As(III, arsenite), As(0, arsenic) and As(-III, arsine). The chemical structures of the primary arsenic compounds found in environmental compartments are given in Table 1. Arsenic speciation in the plant–soil system is dynamic with interconversion through redox cycling and methylation between the inorganic species arsenate and arsenite to the organic species (Abbas and Meharg, 2008). In groundwater, arsenic is commonly found as its inorganic species. However, once the As is present in agricultural soil, the organic forms can increase because of methylation reactions catalyzed by microbial activity. The distribution of As species in the environment plays an important role in their availability to maize plants due to the selective uptake of arsenic species. The mechanisms of arsenic toxicity differ greatly among chemical species. As(V) has the ability to mimic phosphate, therefore, entering into the cell via transporters intended for the uptake of this essential nutrient. Once within the cell, As can interfere with phosphate-based energy-generating processes, thus inhibiting oxidative phosphorylation, for example (Markley and Herbert, 2009). However, arsenite [As(III)] enters cells via a different route (aqua-glycerolporins) and targets a broader range of cellular processes, binding to the thiol groups of important cellular proteins, such as pyruvate dehydrogenase and 2-oxoglutarate dehydrogenase. Organic arsenic compounds have been identified in plant tissues; however, it is unknown whether the compounds are simply absorbed from soil or whether transformation occurs within the plant (Meharg and Hartley-Whitaker, 2002). Thus, the purpose of this review is to summarize the recent advancements related to the conditions involved in agricultural soil for As accumulation in maize crops and its metabolism and phytotoxicity in the plant. Additionally, studies of arsenic accumulation in maize and corn products are synthesized, and As biotransference into the human diet is highlighted. This critical analysis of the available literature will assist in the evaluation of arsenic exposure pathways through corn-based products in the arsenic-contaminated zones with high maize consumption and will form a basis for risk assessment.

178


Table 1 The chemical structure and toxicity of the primary arsenic species of environmental concern. pKac

LD50 (g kg−1)

Genotoxicity response (μg mL−1)

Arsenite [As(III)]

9.29

0.014a, 0.034b

1–2

Arsenate [As(V)]

2.24 6.96 11.5

0.020a, 0.041b

10–14

N1000

Monomethylarsonic acid [MMAs(V)]

3.6 8.2

0.7–1.8a, 0.7b

2500–5000

N3000

Dimethylarsinic acid [DMAs(V)]

1.78 6.14

0.7–2.6a, 3.3b

10,000

Compound

a b c

Chemical structure

In vitro genotoxicity (mM) N300

N300

Tamaki and Frankenberger, (1992). Eguchi et al. (1997). Simon et al. (2004).

2. Sources of arsenic pollution in maize fields Maize plants may be exposed to As through several sources. The plants essentially take up the arsenic that is naturally present in the soil or arsenic that is added through groundwater irrigation or by soil additives contaminated with arsenic (Fig. 1). Several studies have shown that countries with high corn production, such as China, Argentina, India, and Mexico (11.6, 2.8, 2.6 and 0.8 millions of tons in 2011, respectively) (FAOSTAT, 2004), significantly exceeded the world average soil background concentration of As in soil (10.0 mg kg−1) and the maximum acceptable limit for agricultural soil of 20.0 mg kg−1 recommended by the U.S. Environmental Protection Agency (EPA, 2012). The results are shown in Table 2. Similarly, the As concentration in groundwater from these areas presents higher values than the recommended maximum allowable As level for irrigation water (1.0 mg L−1) (EPA, 2012). Several studies have described a significant relationship between the As concentration in the irrigation water or in the agricultural soil and the total As

level accumulated by maize crops. For example, Prabpai et al. (2009) found that the accumulated As in the maize plant was significantly correlated (p = 0.03) with the total As concentration in the agricultural soil of Thailand (0.144 mg kg−1) amended with different ratios of residues from municipal solid waste landfills (2 mg As kg−1). Similarly, Gulz et al. (2005) found a linear relationship between the total As concentration in the soil and the As uptake by the maize root in calcareous and sandy loam soils. In the same study, a high correlation coefficient was calculated for soluble As in soil and the uptake of As in roots. Groundwater has also been identified as the primary source of As contamination in certain areas. For example, in Comarca Lagunera, Mexico, Rosas et al. (1999) found a positive correlation between the content of As(V), As(III) and the sum of both inorganic species in the groundwater with the extractable As in the agricultural soil at a depth of 0–30 cm (P = 0.01). The phytoavailability of As from agricultural soils amended with poultry manure or residues from municipal solid waste landfills to

Fig. 1. Sources of arsenic pollution in maize fields.


179

Table 2 The arsenic concentration in agricultural soil and ground water in selected maize consumer regions. Zone, country

San Luis Potosi, Mexico Comarca Lagunera, Mexico Chihuahua, Mexico Hidalgo, Mexico Baja California, Mexico Sonora, Mexico Tlahualilo, Durango, Mexico Gomez Palacio, Durango, Mexico West Bengal, India Hetao Plain, Inner Mongolia, China La Pampa, Argentina

Arsenic concentration

Ref.

Agricultural soil (mg As kg−1)

Groundwater (μg As L−1)

7.4–1932 30 35.3 97.3 – – – – – – –

8–7165 412 – 1090 b410 2–305 357 42.7 170 76–1093 0.3–1326

maize crops has been evaluated. D'Angelo et al. (2012) estimated the As transference from agricultural soil amended with poultry manure to maize. The presence of As in poultry manure was attributed to the use of roxarsone (4-hydroxy-3-nitrobenzenearsonic acid) as a broiler feed additive. Roxarsone remains in use as chicken feed in several countries, such as Argentina, Canada, Brazil, Chile, Mexico, Venezuela, Indonesia, Jordan, Malaysia, Pakistan, Philippines, Vietnam, and Australia. The same study reported high levels of As in the litter from chicken houses (0.6 to 43.4 mg As kg− 1); however, an insignificant difference was found in the As concentration between maize crops cultivated in agricultural soils non-amended and amended with poultry litter over a two year period. The effect of maize crops with the addition of poultry manure from chicken houses enriched with 3-nitro 4-hydroxy phenyl arsonic acid on agricultural soils was studied by Liebhardt (1976). The application of As-polluted poultry litter did not significantly affect the As accumulation in maize grains, the As levels in grain ranged from 40 to 60 μg kg−1 (Liebhardt, 1976). Prabpai et al. (2009) evaluated the effect of adding residues from municipal solid waste landfills to the soil on the maize yield and the content of heavy metals, including As in the soil. The agricultural soil amended with different ratios (20–80%) of residues from municipal solid waste landfills presented As levels in the range of 0.194–1.055 mg As kg− 1 dry weight (dw). The As concentration in maize grains cultivated in the amended soil ranged from 23 to 32 μg kg−1 dw. 3. Arsenic phytoavailability in agricultural soil for maize crops The phytoavailability of As present in agricultural soil can vary dramatically from one location to another depending, among other factors, on the soil type. In several studies, the water-soluble As level in the soil was better correlated with the As content in the plant than the total As content in the soil. Sadiq (1986) reported that the As concentration in maize crops cultivated in calcareous soils was significantly correlated (P b 0.05) with the water-extractable As content but not with the total As concentration and the diethylenetriamine pentaacetic acid (DTPA) extractable fraction. In studies performed by Gulz et al. (2005), the correlation between the total accumulated As in the maize plant and the water-soluble As fraction in the soil was higher than the total As content in the soil. Note that numerous studies focusing on arsenic uptake by plants evaluated only the effect of the total As dose applied to soils without considering its solubility. The amount of available As for crops has been closely linked to the As solubility in soil. Several factors, including pH, redox potential, organic matter content, interaction/competition with other elements and chemical forms of the pollutant, can affect the As solubility in agricultural soils (Marwa et al., 2012). Agricultural soil conditions can vary from one location to another; therefore, the phytoavailability of nutrients and pollutants can vary through the root system. The effect of parameters, such as iron (Fe), manganese (Mn), phosphorus (P), and organic matter contents, pH value and the presence of microorganisms, has been evaluated

Razo et al. (2004) and Jasso-Pineda et al. (2006) Rosas et al. (1999) and Del-Razo et al. (2002) Gutierrez et al. (2009) Rosas et al. (1999) and Espinosa et al. (2009) Armienta and Segovia (2008) Armienta and Segovia (2008) Rosas et al. (1999) Rosas et al. (1999) Gault et al. (2005) Deng et al. (2009) O'Reilly et al. (2010)

as they relate to the soluble As concentration in soil and/or its phytoavailability to maize plants.

3.1. Effect of iron and manganese Low contents of Fe and Mn (hydr)oxides, which are strong sorbents for As, can reduce the As mobility in deep soil horizons, therefore, increasing the amount of AsT in the topsoil (Sadiq, 1986). Additionally, certain researchers have found that Fe plays an important role in the As distribution on the crop field. Neidhardt et al. (2012) reported a moderate positive correlation (r ≥ 0.64) between the total As concentration and the total Fe content in the topsoil of a maize crop field and a notable decrease in the content of As(III) and AsT in irrigation water of agricultural zone canals after 4 h of the irrigation process (the initial concentrations of As(III) and AsT were 219 and 238 μg L−1, respectively); after 4 h of the irrigation process, the values were 71 and 110 μg L−1, respectively. Both of the results were attributed to groundwater exposure to atmospheric oxygen and the resulting change in the redox potential during the irrigation process. The As(III) and Fe(II) species are oxidized to As(V) and Fe(III), respectively. Because the Fe(III) species shows lower stability under common irrigation conditions than Fe(II), it tends to precipitate as Fe(oxyhydr)oxide. Subsequently, the As(V) ions can be adsorbed or/and co-precipitated on Fe (oxyhydr)oxide particles (Jiang et al., 2005; Neidhardt et al., 2012). The levels of Fe or Mn in the soil can also affect the As solubility in agricultural soil and its transference to vegetable tissues because of the formation of Fe or Mn plaques on the maize root system surfaces, reducing the As absorption and conferring relative immobility (Mallick et al., 2011). In the field study carried out on agricultural soil with Fe2O3 concentration lower than 6% w w−1 was reported the low ratio of total As concentration in Z. mays roots/total As in the topsoil (2.13 mg kg−1 dw/16.8 mg kg−1 dw) (Neidhardt et al., 2012).

3.2. Influence of pH The pH of the soil plays an important role in the mobility and phytoavailability of As. Depending on the soil texture and the nature of the mineral constituents, the specificity of As adsorption on the surface of certain soil components can be influenced by the pH. Under acidic conditions, As(V), the predominant As form in agricultural soil, is primarily sorbed onto iron and aluminum oxides, whereas this species is sorbed by calcium oxides under alkaline conditions (Pongratz, 1998). Woolson et al. (1971) revealed that As(V) adsorption at a lower pH value is weaker on calcium oxides than on iron and aluminum oxides. Other studies reported that the carbonate ion, under alkaline conditions, plays an important role in As immobilization in sediments over a wide range of redox conditions (Ryu et al., 2010). The low pH values in soil results in the increase of the soluble As fraction, which increases its phytoavailability.

180


3.3. Effect of phosphate The similarity of the chemical structures and properties of the phosphate ion and As(V) can explain the observation that the content of phytoavailable P in agricultural soil modifies the As transference to maize crops (Mallick et al., 2011). The phosphorus nutrition status in the plant itself and the inorganic P availability in the rhizosphere can strongly affect the As bioaccumulation in plants. Several studies have reported a negative correlation between the As uptake by maize roots and the P concentration in the soil solution. Abbas and Meharg (2008) evaluated the effect of the phosphate ion administered as a pre-treatment (PO34 levels of 0 and 5 mM) in hydroponic experiments using different maize varieties and As species concentrations (0 and 2.5 mM). The inorganic P concentration significantly reduced (P b 0.001) the influxes of As(III) and DMA(V) to 50 and 90%, respectively. Similarly, phosphorus uptake is affected by As content in agricultural soil. Studies conducted by Mallick et al. (2011) showed that the phosphate ion content in the root and leaves of the maize plant decreased at 3 and 7 days when As levels increased under hydroponic conditions using two varieties of maize and As concentrations ranging from 2.5 to 12.5 μg mL−1. The antagonistic effects between inorganic P and the total As content can be explained by the following: 1) exudation of organic acids by the maize root system under phosphorus deficiency increases the availability of P; thus, the available As fraction is increased (Marwa et al., 2012), 2) changes in the regulation of expression of membrane transporters that allow the membrane permeability to increase for phosphate ions (Abbas and Meharg, 2008), and 3) competition of arsenic and phosphorus for their mutual uptake by plants (Mallick et al., 2011). Contrary to the earlier finding, Sadiq (1986) observed in pot experiments using clay loam, sandy clay, and loamy sand soils a significant positive correlation (p b 0.01) between sodium bicarbonate-extractable P and the As concentration in maize crops. The sodium bicarbonate-extractable P content was similarly correlated with the As content in the irrigation water; thus, the positive correlation of extractable P with As present in maize can be explained by a common enrichment source of P and As. The proportion of the soluble P in soil can be affected by the content of clay, and therefore, the As bioavailability to corn crops. The adsorption capacity and affinity of agricultural soil for substances as PO34 and As(V) can vary with the mineralogical composition and soil texture. Certain studies have reported that the amount of available P is lower in clayrich soils than in soils with low clay content (Gulz et al., 2005). Gulz et al. (2005) found lower levels of soluble P in the silty loam (2.0– 3.2 mg kg−1) than in the sandy loam (3.2–4.2 mg kg−1) enriched with 130 mg kg−1 P (CaH2PO4)2. High clay content leads to the reduction of the anions such as PO34 and As(V); however, low phytoavailability can be counteracted with the organic acids produced from maize root due to P deficiency. 3.4. Microorganisms Microbial interactions in the rhizosphere can affect the P nutritional status in the maize plant, therefore, affecting the As transference to the root system (Bai et al., 2008; Yu et al., 2009). Bai et al. (2008) evaluated the As uptake by the maize plant during 10 weeks of growth after the inoculation of arbuscular mycorrhizal fungal strain, Glomus caledonium 90036 (M1) from agricultural soils in China. In this study, the P content in the maize plant was also determined. At high arsenic concentrations in the agricultural soil (287 ± 10 mg As kg−1), M1 increased the As levels in the maize roots and the content of the dw and P in the shoots and roots compared with non-inoculated maize plants. However, at lower As concentrations, 185 ± 4 and 24 ± 1 mg kg−1, the M1 treatment decreased the As content in the roots and shoots. These results indicate that the M1 treatment can alleviate As toxicity, stabilizing As and P in the agricultural soil (Bai et al., 2008). Yu et al. (2009) evaluated the uptake of As(V) involving different As additions in soil (0–100 mg As kg−1). Arbuscular mycorrhizal inoculation significantly increased (P b 0.01)

the P concentration in roots and shoots at all of the As application levels and decreased the total As content in the roots and shoots. For example, the detected As concentrations in the roots using an As addition of 100 mg kg−1 of soil was 121.7 ± 10.4 mg kg−1 and 175.2 ± 9.3 mg kg−1 for mycorrhizal and non-mycorrhizal treatments, respectively. The effect of mycorrhizal inoculation on As accumulation in maize crops using As(III) and As(V) as the arsenic source was also evaluated. The uptake of As(V) by the root was much lower with mycorrhizal than without mycorrhizal inoculation, and the difference for the As(III) uptake was negligible. Certain microorganisms in the rhizosphere can increase the As tolerance capacity due to their important role in the As detoxification process. The study reported by Wang et al. (2008) showed that Arbuscular mycorrhiza of Glomus mosseae and Acaulospora morrowiae affects the root As efflux. Deposition of As in the external mycelium suggests a possible role of mycorrhizal fungi in avoiding the As accumulation in corn crops. Xia et al. (2007) also described that the use of G. mosseae as arbuscular mycorrhizal tends to reduce the As uptake by corn plant. They found that the beneficial effect can be reduced by the addition of 100 mg kg− 1 P, that could produce mycorrhizal growth decrement. 3.5. Organic matter The organic matter can be originated in soil by decomposition of plants and animals. In addition to affecting the mobility and phytoavailability of As in the agricultural soil, the organic matter has an influence on the following: 1) the soil redox potential, 2) the available adsorption sites, and 3) the formation of aqueous complexes (Wang and Mulligan, 2006). Reducing conditions are generally provided by high organic matter concentrations (Ryu et al., 2010). High in organic matter content in soil leads the proliferation of microorganisms that tend to modify the redox potential in soil. Redman et al. (2002) reported the reduction of As(V) to As(III) by organic matter of the Inangahua River. Low redox potential reduced As uptake by corn crop. As(III) has a greater mobility than As(V) because the oxidized form presents a higher affinity for Fe(III) (hydr)oxides than the reduced species. The oxidation state of As can also affect its degree of accumulation in the maize plant. Generally, As(V) is preferably adsorbed in this crop over reducted As species (Mallick et al., 2011). In addition, organic matter has functional groups that can participate in As adsorption and reduce the amount of available As in the plant. Some factors such as pH, molecular weight, ionic strength, and functional group content, can affect the amount of ions bound to organic matter (Lin et al., 2004). Okieimen et al. (2011) evaluated the effect of three organic soil amendments; poultry manure, cow dung and sludge on As uptake by corn plant. They found that the As accumulation by maize decreased with the increment of organic matter in soil which was attributed to reduction of soluble As and the high P content present in organic matter. A study reported by Urunmatsoma et al. (Urunmatsoma et al., 2010) showed an increment on As mobility with the use of cow dung in agricultural soil. On the other hand, soluble organic compounds can take an important role in the distribution of soluble-adsorbed As in agricultural soil. Approximately, 80% of soluble organic matter is constituted by humic substances such as humic and fulvic acids, whereas the remainder (20%) by carbohydrates, hydrocarbons, carboxylic acids, amino acids and other organic acids (Wang and Mulligan, 2006). The formation of organic matter–metal complexes may strongly bind As through metalbridging mechanisms. Studies conducted by Lin et al. (2004) reported that Ca and Mn and, especially, Fe, Mn and Al participate in complex formation of As with humic substances. Soluble organic complex formation can reduce the As adsorption in agricultural soil and its phytoavailability to crops. Grafe et al. (2002) reported that organic matter, such as citric acid, decreased the As(V) adsorption on ferrihydrite, primarily between 2pH 3 and 5. Additionally, the presence of oxyanions as PO34 , SO4 and


181

Fig. 2. Arsenic transference from agricultural soil and irrigation water to maize tissues. Adapted from Zhao et al. (2010).

MoO24 in organic matter is another factor that can increase the As phytoavailability in agricultural soil due to competitive adsorption with As(V) on the surface soil particles (Manning and Goldberg, 1996). In studies realized by Moreno-Jiménez et al. (2013) reported that the amount of As soluble and their mobility in agricultural soils was significantly increased by using olive mill waste compost (P = 0.001).

of organic species was higher in plants inoculated with G. mosseae than in plants not inoculated (Yu et al., 2009). A method of measuring the As transference from agricultural soils to crops is the bioaccumulation factor (BAF). The BAF expresses the relationship between the As concentration in a plant and its respective value in the soil. Table 3 shows the BAF values calculated from the reported As content in root crops and the total As concentration in soil. The calculated BAF values in corn plants ranged from 0.1 to 2.4.

4. Arsenic transfer from the soil to the maize grain

5. Translocation of arsenic in the maize plant

The maize crop is grown in an aerobic environment in which As is primarily present in the oxidized As(V) form. As(V) enters the plants through P cell channels at the root level and, subsequently, is reduced within the maize root system to As(III) through complexation with phytochelatines and then is stored in cell vacuoles as an As(III)-tris thiolate complex (Mallick et al., 2011). Unreduced As can access the aerial part of the plant via the shoot, be transported via the xylem and the cell phosphate transporter, and stored as an As(III)-tris glutathione complex (Fig. 2). There is a controversy whether plants can produce methylated inorganic species as a detoxification process (Abbas and Meharg, 2008). The presence of organic species in corn crop tissues has been primarily attributed to their absorption from soil rather than to the plant metabolism. The results reported by Lomax et al. (2012) suggest that plants are unable to methylate absorbed inorganic species. The methylated As content in plant tissues has been associated with absorption of methylated forms produced by the microorganisms in the agricultural soil (Lomax et al., 2012). In corn crops, the concentration

Once As is adsorbed in the roots, it is transported to the upper portion of the crop. High As levels in the areal parts of the corn plants have been evaluated as exposure pathways for humans. As can be consumed by farm animals in corn silage, therefore, representing a risk for human health through the food chain. In countries, such as Switzerland, the tolerable limit of arsenic concentration in fodders is 4 mg kg−1; Table 3 shows that this value is not exceeded by the As concentrations found in the leaves and stems (Gulz et al., 2005). Rosas et al. (1999) found that the washed and unwashed corn silage cultivated in agricultural soils of field of Comarca Lagunera, Mexico, with arsenic concentration of 20 mg kg−1 presented an As concentration of 0.5 and 1 mg As kg−1, respectively. In Zimapan, Hidalgo, Mexico, the As concentration in corn silage reached higher levels from 2.1 up to 33 mg kg−1 (CastroLarragoitia et al., 1997). Table 3 shows the As concentrations reported in the different plant parts and their respective translocation efficiencies (expressed as a percentage) estimated from their ratio with the As level in the root. This

182

Table 3 Total As concentrations, bioaccumulation coefficients and translocation efficiency of arsenic in Zea mays [mean ± SD (lower limit–upper limit)]. Soil type/location

Sandy loam/pot experiment

Kilosa, Tanzania Kongwa, Tanzania Dodoma Urban, Tanzania Rorya, Tanzania Musoma Rural, Tanzania Morogoro Urban, Tanzania Morogoro Rural, Tanzania Makete and Mbeya Urban, Tanzania Loamy silt/Mongolia, China Sandy loam–clay loam/Hidalgo, México Hidroponic experimetsa Hidroponic experimetsb Khon Kaen, Thailand Khon Kaen Thailand Sandy clay loam/Torviscosa, Italy Khairpur Mir's, Pakistan Kot Diji, Pakistan n.c., not calculated. a, Azad kamal. b, Azad uttam. ⁎⁎ ,expressed in mg L−1.

Root

(mg kg−1/mg L−1)

(μg kg−1 dw)

5.0 7.0–13.0 20.0 93.0 165.0 210.0 240.0 189.0 212.0 240.0 0.4 (0.2–0.7) 0.2 (0.2–0.7) 0.3 (0.2–0.5) 0.7 (0.5–1.4) 7.1 (3.4–12.4) 0.4 (0.2–15.7) 1.2 (0.4–14.7) 1.7 (1.7–1.8) 16.8 ± 0.2 – 12.5⁎⁎ 12.5⁎⁎ 0.1 1.1 586.0–718.0 25 (20.5–31.0) 32.7 (18.7–54.0)

483.4 ± 301 – – 1.0 × 105 ± 5900 2.5 × 105 ± 36,100 4.2 × 105 ± 12,000 5.7 × 105 ± 27,300 1.5 × 105 ± 6700 2.1 × 105 ± 8700 2.8 × 105 ± 22,400 – – – – – – – – 2130 ± 1250 41.3 1.1 × 108 ± 3.0 × 107 1.3 × 108 ± 1.1 × 107 – – 1.3 × 105–1.6 × 105c 2050 ± 660 1740 ± 530

Stem

Leave

Grain

7.8 ± 2.0 – 3000 1000 ± 200 1000 ± 100 2000 ± 300 3000 ± 300 7000 ± 400 10,000 ± 700 9000 ± 300 20 (20–40) 30 (20–50) 70 (50–100) 40 (10–60) 80 (30–140) 4 (30–650) 40 (20–50) 40 (20–60) 210 ± 230 b28.4 – – – – 18,500–22,540c 406 ± 90 365 ± 100

8.1 ± 0.8 17.0–29.0 – 3000 ± 900 3000 ± 500 4000 ± 1000 5000 ± 2200 9000 ± 1700 9000 ± 600 11,000 ± 400 – – – – –

– 5.0–6.4 – – – – – 200 ± 10 200 ± 20 100 ± 10 b10.0 b10.0 b10.0 b10.0 b10.0 b10.0 b10.0 b10.0 60 ± 20 120.3 – – 23 32 – 302 ± 50 280 ± 40

Translocation Stem

Ref.

Leave

Grain

1.7 n.c. n.c. 2.9 1.2 1.0 0.9 6.0 4.2 4.0 n.c. n.c. n.c. n.c. n.c. n.c. n.c. n.c. 29.1 n.c. 28.6 26.0 n.c. n.c. n.c. n.c. n.c.

n.c. n.c. n.c. n.c. n.c. n.c. n.c. 0.1 0.1 3.6 × 10−4 n.c. n.c. n.c. n.c. n.c. n.c. n.c. n.c. 2.8 291.3 n.c. n.c. n.c. n.c. n.c. 15.0 16.1

(%)

– – 620 ± 410 b28.4 3.1 × 108 ± 1.5 × 106 3.4 × 108 ± 3.3 × 106 – – – – –

0.1 n.c. n.c. 1.1 1.6 2.0 2.4 0.8 1.0 1.2 n.c. n.c. n.c. n.c. n.c. n.c. n.c. n.c. 0.1 n.c. n.c. n.c. n.c. n.c. 0.2 0.1 0.1

1.6 n.c. n.c. 1.0 0.4 0.5 0.5 4.7 4.7 3.2 n.c. n.c. n.c. n.c. n.c. n.c. n.c. n.c. 9.9 n.c. n.c. n.c. n.c. n.c. 1.4 20.1 20.1

Schulz et al. (2007) D'Angelo et al. (2012) Rosas et al. (1999) Gulz et al. (2005)

Marwa et al. (2012)

Neidhardt et al. (2012) Prieto-García et al. (2007) Mallick et al. (2011) Prabpai et al. (2009) Fellet et al. (2007) Baig et al. (2010)


Quartz sand/pot experiment Maury silty loam/Kentucky, EUA Sandy clay loam/Comarca lagunera, Mexico Silty loam/pot experiment

BAF

AsTot Soil/water


parameter was calculated as the rate between the As concentration in different plant parts (stem, leaf and grain) and their respective As concentration in the root (Table 3). As shown in Table 3, the transfer efficiency values varied between the reviewed studies. In general, the As concentrations in the different maize plant parts increased in the following order: root N stem N leaf N grain. Gulz et al. (2005) did not report a significant translocation in corn plants from roots to shoots; however, other researchers reported a significant As transport in this crop. Marwa et al. (2012) evaluated the uptake and translocation of As in maize grown in Tanzanian agricultural soils with an As concentration ranging from 0.20 to 15.72 mg kg−1 to estimate the safety of maize used for human and animal consumption. The As content in the shoot and grain of the corn plant varied from 0.01 to 0.65 mg kg−1 and from 0.01 to 0.17 mg kg−1, respectively (Marwa et al., 2012). Similarly, Parsons et al. (2008) reported relatively high concentrations of As(V) and As(III) of 18 and 12 mg kg− 1, respectively, in buds, even when the As concentrations in the roots were 95 and 112 mg kg−1, respectively. As previously stated, the majority of these studies showed a decrease in the As content from the root to the aerial corn parts (leaves, stems, and seeds). However, findings by Prieto-García et al. (2007) in the municipality of Actopan in the Valley of the Mezquital, Hidalgo (Mexico) showed opposite results. The As concentrations in corn grain (0.1203 mg kg− 1) and corn cob (0.1046 mg kg−1) were higher than those in the corn plant close to the topsoil with As values in root (0.0413 mg kg− 1), stem (b0.0284 mg kg− 1), and leaves (b0.0284 mg kg−1). 5.1. Factors of arsenic mobility in the maize plant Z. mays generally revealed lower translocation capacities compared with other cereal crops. In the study conducted by Su et al. (2009), cereals, such as rice, wheat, and barley, grown in soils with 5 μM As(V) (0.375 mg kg−1) showed an As translocation efficiency of 9, 6, and 2%, respectively. However, the translocation factor can vary considerably, and it can be affected by the contribution of several parameters, such as the As-soluble fraction content in the soil, the presence of chelating agents in the roots, the crop age, and the soil type. Recently, Mallick et al. (2011) determined that the As translocation efficiency decreased with increasing As concentration in the hydroponic media. After three days of growth, the translocation values from the root to the leaves in two varieties of Z. mays, Azad kamal (AK) and Azad uttam (AU), were 34.0, 23.3, and 28.7% (AK) and 69.3, 38.5 and 26.0% (AU) for the high As(V) levels of 2.5, 7.5 and 12.5 mg mL−1, respectively. The relatively high translocation efficiency observed in this study can be attributed to the high As concentrations evaluated and the short period of growth. Other conditions, such as the chelating agent concentration, can affect the As translocation in maize crops. The As translocation in the maize plant can be reduced by As sequestration on Fe oxide plaques and thiol ligands of maize roots (D'Angelo et al., 2012). Arsenic accumulation in the above-ground maize tissues can also be restricted when roots are associated with arbuscular mycorrhizal fungi (Yu et al., 2009). Zheng et al. (2011) reported that the transport and distribution of As varied with the temporary status of the plant. The total As concentration in vegetative tissues increased predominantly during the two weeks after the flowering phase, and inorganic As was transported primarily within the caryopsis during the grain development. Similarly, certain studies have reported variability in As translocations caused by the type and properties of the agricultural soil. Gulz et al. (2005) found lower As translocations in maize crops grown in silty loam soil (0.4%) than in sandy loam soil (4.7%) for a soluble As concentration in both soils of 2.8 mg kg−1. The higher efficiency of translocation in sandy loam soil clearly has a critical effect on the final concentration in the kernel. Although the As concentration in maize roots cultivated in silty loam soil was 1.7 times higher than in sandy loam soil, the As concentration found in corn seed grown in sandy loam soil was greater (0.1 ± 0.001 mg As kg−1) because of its higher translocation efficiency.

183

5.2. Translocation of arsenic species in the maize plant A study conducted by Raab et al. (2007) under hydroponic conditions with 46 corn species showed that DMA(V) was more effectively translocated from the root to the shoot than As(V) and MMA(V) (3 and 10 times, respectively). Similar results have been reported for cereals, such as rice (Oryza sativa L.) (Marin et al., 1992). However, other studies reported that arsenite, the most toxic As species, was the As form found primarily in the plant xylem (86–97% of the total As content) of cereals, such as rice, wheat and barley. The As(III)/AsT ratio in the xylem sap was considerably higher when As(V) was used as the arsenic source in the root system (84, 45 and 63% for rice, wheat, and barley, respectively) (Su et al., 2009). Therefore, it is important to consider the distribution of the As species during the risk assessment of the use of corn silage as livestock feed. 6. Arsenic phytotoxicity in maize crops Arsenic is an essential ultratrace element and plays an important role in healthy plant growth. Abbas and Meharg (2008) reported that small additions of As increased the tolerance index values ((mean maximum root growth with As treatment) / (mean maximum root growth As control) × 100) in six evaluated corn varieties (single cross 10, single cross 30 k8, single cross 2030, single cross 3084, three-way cross 310 and three-way cross 323) under hydroponic conditions. With low As(III) concentrations (0.5 mg L−1), all of the corn varieties exceeded the 100% level of the Tolerance Index (TI, defined as exposed plant biomass/control plant biomass), and three corn varieties exceeded the 100% level with As(V) at the same concentration. A similar effect of trace As concentration has been observed for other crops, such as rye, potatoes, and wheat (Jacobs et al., 1970; Xu and Thornton, 1985; Carbonell-Barrachina et al., 1998). However, the increase in the As concentration becomes toxic for crops, causing necrosis, chlorosis, inhibition of growth and, finally, death (Gulz et al., 2005). In the study conducted by Ci et al. (2012) concentrations of 12.5 and 25 mg kg−1 in the soil promoted maize growth and the nutritional quality of the grain, whereas higher concentrations (50 and 100 mg As kg−1 of soil) presented toxic effects for the crop. The negative correlation between the As content in plant tissues and the crop development has been widely demonstrated. Woolson et al. (1971) observed a significant (p = 0.05) linear correlation between the logarithm of the As concentration in the soil soluble fraction and the growth reduction in corn. Certain Z. mays varieties differ in their As tolerance capacity. Mallick et al. (2011) evaluated the As resistance in two varieties of Z. mays and found that the fresh weight percentage and the root length were affected to a greater degree in A. uttam than in A. kamal and with a similar behavior; the total chlorophyll content at 7 days was lower in A. uttam than that in A. kamal. These results indicated that the A. kamal variety was more tolerant to As. The As phytotoxicity degree varies with its chemical form and the corn variety. The relative toxicity according to the TI under hydroponic conditions has been reported as As(V) N As(III) N N DMA (Abbas and Meharg, 2008). 6.1. Mechanism of arsenic toxicity The highest As(V) toxicity in plants arises primarily from the contribution of two factors. High As levels can reduce phosphate uptake from agricultural soil because of the structural similarity of As(V) to phosphate. In addition, the presence of As(V) in plants could lead to oxidative stress by altering the synthesis of adenosine triphosphate and altering the phosphate group of DNA. As(V) can also interact with the sulfydryl and thiol groups of proteins, cysteine, antioxidant enzymes, such as superoxide dismutase (SOD), and phytochelatin ascorbateglutathione (GSH). Mallick et al. (2011) reported that under a hydroponic As treatment (12.5 μg As mL−1 over 3 days), the SOD and GSH activity in the A. kamal corn variety increased considerably, and the

184


increment was higher in this variety than in A. uttam. These substances play an important role as the first defense line against generated radicals during oxidative stress (Mittler, 2002). Increased SOD production resulting from As toxicity has been reported in As hyperaccumulator plants and sensitive fern species (Srivastava et al., 2005), grass Holcus lanatus (Hartley-Whitaker et al., 2001) and corn crops (Mylona et al., 1998). Once absorbed, As(V) is reduced to As(III) through complexation by phytochelatins, such as GSH. The formed As complexes, As(III)-tris glutathione and As(III)-tris thiolate, are stored in vacuoles in the shoots and roots, respectively (Mallick et al., 2011). Phytochelatins play an important role in the As detoxification process in crops. Under As exposure, phytochelatins of different chain lengths (n = 2–4) can be synthetized in plants and linked to As (Fig. 2). Schulz et al. (2007) reported differences in phytochelatin production in the evaluated plant species and found that in corn crops, phytochelatin 3 was the component mostly induced by As exposure (248, 698, and 210 μg g− 1 for phytochelatins 2, 3, and 4, respectively). Similar results were reported for plants, such as Leonurus marrubiastrum, Glecomona hederácea, and Urtica dioica, in which phytochelatin 3 was the most induced form by arsenic (Schulz et al., 2007). 6.2. Toxic effects of arsenic on the nutritional status of the maize plant During the intoxication process by As, there is a change in the nutrimental state of the plant, especially in the content of elements, such as Mg, K and Ca. Najmowicz et al. (2010) reported that the Mg concentration increased with increasing As levels in the root and aerial parts of the corn plant. Potassium, a macronutrient required in high concentrations, constitutes up to 10% of the plant dry weight, and its content is significantly affected in crops poisoned by As. A decrease by 60% of the potassium level in corn after As treatment has been described by Parsons et al. (2008). This decrease can be associated with the agglomeration of plaques inside the roots system that reduced the nutrient translocation (Mallick et al., 2011). Potassium concentrations below 10 g kg−1 can be considered as deficient for certain plants showing functional alterations and, in severe cases, death of lateral and terminal meristems (Mallick et al., 2011). A higher calcium concentration in corn leaves in the two corn varieties at 7 days of As exposure compared with the control experiment has been reported (Mallick et al., 2011). Increases in Ca levels have been associated with a response of the oxidative stress of plant tissues. Inside the cell function, calcium acts as a free radical scavenger and as a second messenger, therefore, playing an important role in determining oxidative stability (Mallick et al., 2011). Regarding the As effect on P levels, Mallick et al. (2011) reported that the P concentration increased in plants at 7 days of As exposure,

a)

Water and other drinking beverages 38%

which was attributed to the ineffective scavenging of immobile nutrients, particularly phosphorus in corn plants, under oxidative stress conditions (Smith et al., 2003). Although minerals, such as Fe and Mg, exhibit participation in plaque formation on the root surface, these minerals show a slow ascent to the corn aerial parts; thus, the As effect on their levels in corn crops remains unclear (Mallick et al., 2011). 7. Human exposure to arsenic through maize grain Maize is a staple food for more than 200 million people in regions such as Latin America, sub-Saharan Africa, and Southeast Asia that most of them has been described as arsenic contaminated Zones (Nuss and Tanumihardjo, 2010). This cereal has multiple uses and high nutritive value providing an estimated 15% and 20% of the world's intake of protein and calories respectively, and thus it could represent an important source of As exposure for humans (Nuss and Tanumihardjo, 2010). In countries with high corn consume, high As concentrations in grain can present a risk for local consumers. Ci et al. (2012) reported that the As concentration in maize grain exceeds the maximum permissible concentration in China (0.7 mg kg−1) when crop grew in soil with As concentration higher than 50 mg kg−1. A provisional tolerable weekly intake (PTWI) of no more than 15 μg inorganic As kg−1 body weight is established by the Food and Agriculture Organization/World Health Organization (FAO/WHO) and the Expert Committee on Food Additives (JECFA) (FAO/WHO, 2013). Brahman et al. (2014) evaluated the As intoxication risk associated to the consumption of water and corn grain by the local population in Nagarparkar, Pakistan. They found a hazard index value for 7–15 group age higher than for the reference value which indicates that the local population has several risks to chronic As poisoning through consumption of water and corn grain. In countries, such as Mexico, in which corn consumption per capita has been reported at 315.1 g day− 1 (CNMI, 2001), the corn intake with inorganic As concentrations above 0.476 mg kg−1 for a person of 70 kg would represent a risk for the Mexican population. In countries, such as China, the maximum permissible concentration is 0.7 mg kg−1. In two rural populations of Mexico, Los Angeles, Durango and Lagos de Moreno, Coahuila, where groundwater was the primary source of As exposure, Del-Razo et al. (2002) estimated that the consumption of maize tortillas cooked from local corn grain represents the 32 and 7% of the total As intake by typical diet in the region, respectively (Fig. 3). The As concentration in tortillas ranged from 0.001 to 1.13 mg As kg−1 dw and the respective mean value were 0.14 and 0.54 mg As kg−1 dw in Los Angeles and Lagos de Moreno. In Los Angeles, the As concentration in tortillas corresponds to a contribution of 13.3% (19.9 μg day−1)

Tortilla 7%

b)

Tortilla 32%

Other foods 30%

Water and other drinking beverages 63%

Other foods 30%

Total As intake (µg day-1)

62.6

1059

Fig. 3. The As intake by consumption of typical food and beverages in rural population of: a) Los Angeles, Durango As and b) Lagos de Moreno, Coahuila.


of the PTWI, considering total arsenic as inorganic As, and the local corn tortilla consumption of 221 g day−1 (corporal weight: 70 kg) (Del-Razo et al., 2002). In the same study, but in Lagos de Moreno, the local daily As intake by consumption of corn tortillas reached approximately 50% of the PTWI (75.1 μg day−1) (Del-Razo et al., 2002). In other Mexican area such as Zimapan Hidalgo, the As concentrations in corn grain reached 0.55 mg kg−1 (Bundschuh et al., 2012). These values demonstrated the large contribution of corn consumption to the human total As intake and the importance of studying the health risk associated with the As chronic toxicity effect in countries in which corn constitutes a considerable portion of the diet. The relationship between the As intake and diabetes mellitus, a predominant disease in Mexico, has been studied. Coronado-Gonzalez et al. (2007), found that exposed people in Coahuila, Mexico, with total As concentration in urine normalized for creatinine ranged between 63.6 and 104 μg g−1 had higher risk of developing type 2 diabetes than people with lower As levels (odds ratio 2.16, P = 95%). This result suggests that this metalloid can be a risk factor for developing type 2 diabetes. Rosado et al. (2007) evaluated the relationship between As and cognitive performance in child exposed to high As levels in the environment (n = 602, 6–8 years old, 52% of child had As concentrations in urine higher to 50 μg L−1) in Torreon, Coahuila, Mexico. They found a negative correlation between the As levels and Visual–Spatial Abilities with the Peabody Picture Vocabulary Test, Figure Design, the WISC-RM Digit Span subscale, Visual Search, and Letter Sequencing Tests (P b 0.05). Gamiño-Gutiérrez et al. (2012) found that almost 50% of children population had MEC frequency values (biomarker associated with chronic diseases like cancer, cardiovascular diseases and neurodegenerative disorders) higher than the basal level for population non-exposed to genotoxic substances in Villa de la Paz, San Luis Potosi, Mexico. Those results confirmed that there is a genotoxic damage in local child population associated to the As exposure. In cereals as rice, a higher bioavailability has been reported for inorganic As than the organic forms high (Moreda-Piñeiro et al., 2011). Cereals, such as rice and wheat, contain inorganic As percentages ranging from 33 to 96% (Williams et al., 2005; Reyes et al., 2007; Sanz et al., 2007; Meharg et al., 2008; Zavala et al., 2008; Zhu et al., 2008; Batista et al., 2011; Jackson et al., 2012) and 93–95% (Reyes et al., 2007; Cubadda et al., 2010), respectively. Table 4 summarizes the reported values of the total and inorganic As content in corn grain and food derivatives found in the literature. The reported total As concentration ranged from 0.04 to 1.13 mg kg−1, and the inorganic As was 50–73%. In the study conducted by Ci et al. (2012) in hydroponic media, organic As compounds accounted for 61% of the total As content in grain. Certain As concentration values in raw corn grain are lower than the established maximum levels of 0.3 mg kg−1 for As in certain countries (Gulz et al., 2005). In countries, such as Switzerland, As levels greater than 0.20 mg kg−1 are considered unsafe for human consumption (Marwa et al., 2012). Table 4 reveals the lack of information about As species distribution in corn grain and food derivatives and the evaluation of their

Table 4 The arsenic concentration in corn grain and food derivatives. Food

Country

AT (mg kg−1)

Ai (%)

Ref.

Corn grain

0.04–0.06

–

Liebhardt (1976)

0.02–0.40 0.05–0.55

50–73 –

Corn grain Corn grain

Delaware, United States Northern Chile San Luis Potosi, Mexico Mongolia, China Hidalgo, Mexico

0.06 0.02 0.1203

– –

Tortilla Tortilla

Durango, Mexico Coahuila, Mexico

0.14 (0.05–0.82) 0.54 (0.16–1.13)

– –

Muñoz et al. (2002) Castro-Larragoitia et al. (1997) Neidhardt et al. (2012) Prieto-García et al. (2007) Del-Razo et al. (2002) Del-Razo et al. (2002)

Corn grain Corn grain

AT total arsenic; Ai, inorganic arsenic.

185

bioavailability. The toxicity varies considerably between As(III) and As(V) (Table 1), and the content of As in corn-based foods is not completely released to the lumen of the human gastrointestinal tract. In cereals, such as rice and wheat, the content of As(III), the most toxic As form, has been reported at approximately twice that of the As(V) concentration (rice: 41 ± 9 and 21 ± 19%; wheat: 61 ± 6 and 33 ± 4% for As(III) and As(V), respectively) (Williams et al., 2005; Reyes et al., 2007; Sanz et al., 2007; Meharg et al., 2008; Zavala et al., 2008; Zhu et al., 2008; Cubadda et al., 2010; Batista et al., 2011; Jackson et al., 2012). However, using an in vitro gastrointestinal digestion method, Laparra et al. (2005) found that the majority (N 90%) of the total arsenic in cooked rice is bioavailable. After gastrointestinal digestion, the bioaccessible fraction of inorganic As reached 63–99%, whereas As(V) was the predominant species. Additionally, Laparra et al. (2005) found that the cooking procedure of rice grain leads to an increase in the inorganic species content. Similar results were described by Batista et al. (2011) in which inorganic As content in rice, especially As(V), had greater bioavailability than organic forms. The percentages of the As species were 39.7, 17.8, 3.7, and 38.7% for As(III), As(V), MMA(V), and DMA(V), respectively. Using an in vivo swine model, Moreda-Piñeiro et al. (2011) found that the inorganic As in rice had a higher bioavailability (89 ± 9%) than the organic forms. Only 33 ± 3% of the DMA(V) and 16.7% of MMA(V) of the total arsenic in rice were bioavailable (MoredaPiñeiro et al., 2011). Therefore, the As speciation and bioavailability studies in corn grain and corn-based food products is a critical parameter for the estimation of exposure through the consumption of corn-based food in countries, such as Mexico.

8. Conclusions and remarks We have comprehensively compiled, compared and evaluated the literature concerning arsenic influx in corn crops and the uptake mechanisms, metabolism and phytotoxicity of As in corn plants. This information becomes important for understanding the conditions that affect the As transference from agricultural soils and aquifers to corn crops and for evaluating the As risk in natural ecosystems and to human health. Most of the field experiments of As uptake by corn plants were performed in agricultural soil and were based on the effects of the total As concentration without considering the As species distribution and As available fraction. Minerals, such as Fe and P, play an important role in As phytoavailability in agricultural soil and its translocation in corn crops. In addition, arsenic levels in corn crops can be indirectly affected by interactions in the rhizosphere of microorganisms, such as mycorrhizal, that alter the nutritional status of minerals, such as Fe and P. Similarly, the redox potential in the rhizosphere plays an important role in the As transference to the agricultural product because As uptake is significantly affected by the chemical form. As(V) is the predominant chemical form absorbed by the corn crop and shows the highest phytotoxicity. A high arsenic concentration in corn crops leads to oxidative stress and affects the plant development. Maize plants are able to reduce arsenate to arsenite, the more toxic As species for humans. As concentration in maize tissues increased in the order of grain b stem b leaf b broot. In field studies, the reported As concentration in corn grain varied widely from 0.04 to 1.13 mg kg−1, exceeding the maximum permissible concentration established in certain countries (200–300 μg kg−1). The inorganic As proportion has been determined to be 39–73% of the total arsenic. Limited information exists regarding the distribution of As species in corn grain and corn-based food products. To the best of our knowledge, there are no studies of As bioavailability for humans in raw or cooked corn. A greater bioavailability of As(III) and As(V) species than organic species has been shown in cereals, such as rice. Thus, it is important to evaluate the degree of As bioavailability in raw and cooked corn. This information will enable the development of strategies to

186


reduce the risk associated with As consumption through corn-based products. Finally, there are limited information about actions needed to control the toxicological risk by corn grain and corn-based food products consumption. The effect of microorganisms inoculation, phosphate and organic matter addition on agricultural soil has been evaluated with promising results; however, this needed to be investigated more thoroughly. The use of other sources of organic matter as fertilizers and complexing agents in agricultural soil as well as studies related to the stress reduction of maize plant preventing the release of organic acid exudates by root are lines of research that may contribute to reducing of As transference through water–soil–maize–human system.

Acknowledgments This study was supported by the Research fund of PAICyT-UANL and CONACyT through grants UANL-PAICyT-CN885-11 and CONACYT/CB/ 167372, respectively. M. Rosas-Castor would like to thank CONACyT for his doctoral grant.

References Abbas MHH, Meharg AA. Arsenate, arsenite and dimethyl arsinic acid (DMA) uptake and tolerance in maize (Zea mays L.). Plant and Soil 2008;304:277–89. Abedin MJ, Meharg AA. Relative toxicity of arsenite and arsenate on germination and early seedling growth of rice (Oryza sativa L.). Plant Soil 2002;243:57–66. Armienta MA, Segovia N. Arsenic and fluoride in the groundwater of Mexico. Environ Geochem Health 2008;30:345–53. ATSDR. Toxicological profile for arsenic. ATSDR/TP-88/02 GA: Agency for Toxic Substances and Disease Registry; 1989. Bai J, Lin X, Yin R, Zhang H, Junhua W, et al. The influence of arbuscular mycorrhizal fungi on As and P uptake by maize (Zea mays L.) from As-contaminated soils. Appl Soil Ecol 2008;38:137–45. Baig JA, Kazi TG, Shah AQ, Arain MB, Afridi HI, et al. Evaluating the accumulation of arsenic in maize (Zea mays L.) plants from its growing media by cloud point extraction. Food Chem Toxicol 2010;48:3051–7. Batista BL, Souza JMO, De Souza SS, Barbosa F. Speciation of arsenic in rice and estimation of daily intake of different arsenic species by Brazilians through rice consumption. J Hazard Mater 2011;191:342–8. Brahman KD, Kazi TG, Baig JA, Afridi HI, Khan A, et al. Fluoride and arsenic exposure through water and grain crops in Nagarparkar, Pakistan. Chemosphere 2014;100: 182–9. Bundschuh J, Nath B, Bhattacharya P, Liu C-W, Armienta MA, et al. Arsenic in the human food chain: the Latin American perspective. Sci Total Environ 2012;429:92–106. Cao HC, Luan ZQ, Wang JD, Zhang XL. Potential ecological risk of cadmium, lead and arsenic in agricultural black soil in Jilin Province, China. Stoch Env Res Risk Assess 2009; 23:57–64. Carbonell-Barrachina A, Aarabi MA, Delaune RD, Gambrell RP, Patrick WHJ. Bioavailability and uptake of arsenic by wetland vegetation: effects on plant growth and nutrition. J Environ Sci Health 1998;33:45–66. Castro-Larragoitia J, Kramar U, Puchelt H. 200 years of mining activities at La Paz/San Luis Potosi/Mexico-consequences for environment and geochemical exploration. J Geochem Explor 1997;58:81–91. Chowdhury UK, Rahman MM, Mandal BK, Paul K, Lodh D, et al. Groundwater arsenic contamination and human suffering in Wes Bengal, India and Bangladesh. Environ Sci 2001;8:393–415. Ci X-k, Liu H-l, Hao Y-b, Zhang J-w, Liu P, et al. Arsenic distribution, species, and its effect on maize growth treated with arsenate. J Integr Agric 2012;11:416–23. CNMI. Cámara Nacional del Maíz Industrializado. Retrieved 16 de octubre, 2012, from http://www.cnmaiz.org.mx/camara.html, 2001. Coronado-Gonzalez J, Del-Razo L, Garcia-Vargas G, San-Miguel-Salazar F, Escobedo-de-laPena J. Inorganic arsenic exposure and type 2 diabetes mellitus in Mexico. Environ Res 2007;104:383–9. Cubadda F, Ciardullo S, D'Amato M, Raggi A, Aureli F, et al. Arsenic contamination of the environment-food chain: a survey on wheat as a test plant to investigate phytoavailable arsenic in Italian agricultural soils and as a source of inorganic arsenic in the diet. J Agric Food Chem 2010;58:10176–83. D'Angelo E, Zeigler G, Beck EG, Grove J, Sikora F. Arsenic species in broiler (Gallus gallus domesticus) litter, soils, maize (Zea mays L.), and groundwater from litter-amended fields. Sci Total Environ 2012;438:286–92. Del-Razo LM, Garcia-Vargas GG, Garcia-Salcedo J, Sanmiguel MF, Rivera M, et al. Arsenic levels in cooked food and assessment of adult dietary intake of arsenic in the Region Lagunera, Mexico. Food Chem Toxicol 2002;40:1423–31. Deng Y, Wang Y, Ma T, Gan Y. Speciation and enrichment of arsenic in strongly reducing shallow aquifers at western Hetao Plain, northern China. Environ Geol 2009;56: 1467–77.

Duxbury JM, Mayer AB, Lauren JG, Hassan N. Food chain aspects of arsenic contamination in Bangladesh: effects on quality and productivity of rice. J Environ Sci Health A Tox Hazard Subst Environ Eng 2003;38:61–9. Eguchi N, Kuroda K, Endo G. Metabolites of arsenic induced tetraploids and mitotic arrest in cultured cells. Arch Environ Contam Toxicol 1997;32:141–5. EPA. Environmental Protection Agency of U.S. . Retrieved 14 de noviembre, 2012, from http://www.epa.gov/iris/subst/0278.htm, 2012. Espinosa E, Armienta MA, Cruz O, Aguayo A, Ceniceros N. Geochemical distribution of arsenic, cadmium, lead and zinc in river sediments affected by tailings in Zimapán, a historical polymetalic mining zone of México. Environ Geol 2009;58:1467–77. FAO. Food and Agriculture Organization of United Nations. Retrieved 23-March, 2012, from www.fao.org/home/en/, 2005. FAO/WHO. Food and Agriculture Organization/World Health Organization. Retrieved 13 de junio, 2013, from http://apps.who.int/ipsc/database/evaluations/chemical.aspx? chemID=1863, 2013. FAOSTAT. Food and Agriculture Organization of the United Nations. Retrieved 13 de mayo http://faostat.fao.org/site/613/DesktopDefault.aspx?PageID=613#ancor, 2004. Fellet G, Marchiol L, Perosab D, Zerbia G. The application of phytoremediation technology in a soil contaminated by pyrite cinders. Ecol Eng 2007;31:207–14. Gamiño-Gutiérrez SP, González-Pérez CI, Gonsebatt ME, Monroy-Fernández MG. Arsenic and lead contamination in urban soils of Villa de la Paz (Mexico) affected by historical mine wastes and its effect on children's health studied by micronucleated exfoliated cells assay. Environ Geochem Health 2012;35:37–51. Gault AG, Jana J, Chakraborty S, Mukherjee P, Sarkar M, et al. Preservation strategies for inorganic arsenic species in high iron, low-Eh groundwater from West Bengal, India. Anal Bioanal Chem 2005;381:347–53. Grafe M, Eick M, Grossl P, Saunders A. Adsorption of arsenate and arsenite on ferrihydrite in the presence and absence of dissolved organic carbon. J Environ Qual 2002;31: 1115–23. Gulz PA, Gupta S-K, Schulin R. Arsenic accumulation of common plants from contaminated soils. Plant and Soil 2005;272:337–47. Gutierrez M, Alarcón-Herrera MT, Camacho LM. Geographical distribution of arsenic in sediments within the Rio Conchos Basin, Mexico. Environ Geol 2009;57:929–35. Hartley-Whitaker J, Ainswort G, Meharg AA. Copper and arsenate induced oxidative stress in Holcus lanatus L. clones with differential sensitivity. Plant Cell Environ 2001;24: 713–22. Huq SMI, Joardar JC, Parvin S, Correll R, Naidu R. Arsenic contamination in food-chain: transfer of arsenic into food materials through groundwater irrigation. J Health Popul Nutr 2006;24:305–16. Jackson BP, Taylor VF, Karagas MR, Punshon T, Cottingham KL. Arsenic, organic foods, and brown rice syrup. Environ Health Perspect 2012;120:623–6. Jacobs LW, Keeney DR, Walsh LM. Arsenic residue toxicity to vegetable crops grown on Plainfield sand. Agron J 1970;62:588–91. Jasso-Pineda Y, Espinosa-Reyes G, Gonzalez-Mille D, Razo-Soto I, Carrizales L, et al. An integrated health risk assessment approach to the study of mining sites contaminated with arsenic and lead. Integr Environ Assess Manag 2006;3:344–50. Jiang W, Zhang S, Shan X, Feng M, Zhu YG, et al. Adsorption of arsenate on soils. Part 2: modeling the relationship between adsorption capacity and soil physiochemical properties using 16 Chinese soils. Environ Pollut 2005;138:285–9. Khan NI, Owens G, Bruce D, Naidu R. Human arsenic exposure and risk assessment at the landscape level: a review. Environ Geochem Health 2009;31:143–66. Laparra JM, Vélez D, Barberá R, Ferré R, Montoro R. Bioavailability of inorganic arsenic in cooked rice: practical aspects for human health risk assessments. J Agric Food Chem 2005;53:8829–33. Liebhardt WC. The arsenic content of corn grain grown on a coastal plain soil amended with poultry manure. Commun Soil Sci Plant Anal 1976;7:169–74. Lin H, Wang M, Li G. Complexation of arsenate with humic substance in water extract of compost. Chemosphere 2004;56:1105–12. Lomax C, Liu W-J, Wu L, Xue K, Xiong J, et al. Methylated arsenic species in plants originate from soil microorganisms. New Phytol 2012;193:665–72. Mallick S, Sinam G, Sinha S. Study on arsenate tolerant and sensitive cultivars of Zea mays L.: differential detoxification mechanism and effect on nutrients status. Ecotoxicol Environ Saf 2011;74:1316–24. Manning BA, Goldberg S. Modeling arsenate competitive adsorption on kaolinite, montmorillonite, and illite. Clays Clay Minerals 1996:44. Marin AR, Masscheleyn PH, Patrick WH. The influence of chemical form and concentration of arsenic on rice growth and tissue arsenic concentration. Plant Soil 1992;139: 175–83. Markley CT, Herbert BE. Arsenic risk assessment: the importance of speciation in different hydrologic systems. Water Air Soil Pollut 2009;204:385–98. Marwa EMM, Meharg AA, Rice CM. Risk assessment of potentially toxic elements in agricultural soils and maize tissues from selected districts in Tanzania. Sci Total Environ 2012;416:180–6. Meharg AA, Hartley-Whitaker J. Arsenic uptake and metabolism in arsenic resistant and nonresistant plant species. New Phytol 2002;154:29–43. Meharg AA, Lombi E, Williams PN, Scheckel KG, Feldmenn J, et al. Speciation and localization of arsenic in white and brown rice grains. Environ Sci Technol 2008; 42:1051–7. Mittler R. Oxidative stress, antioxidants and stress tolerance. Plant Sci 2002;7:405–10. Moreda-Piñeiro J, Moreda-Piñeiro A, Romarís-Hortas V, Moscoso-Pérez C, López-Mahía P, et al. In-vivo and in-vitro testing to assess the bioaccessibility and the bioavailability of arsenic, selenium and mercury species in food samples. Trends Anal Chem 2011; 30:324–45. Moreno-Jiménez E, Clemente R, Mestrot A, Meharg AA. Arsenic and selenium mobilisation from organic matter treated mine spoil with and without inorganic fertilisation. Environ Pollut 2013;173:238–44.

J.M. Rosas-Castor et al. / Science of the Total Environment 488–489 (2014) 176–187 Muñoz O, Díaz OP, Leyton I, Nuñez N, Devesa V, et al. Vegetables collected in the cultivated Andean area of northern Chile: total and inorganic arsenic contents in raw vegetables. J Agric Food Chem 2002;50:642–7. Mylona PV, Polidoros AN, Scandalios JG. Modulation of antioxidant responses by arsenic in maize. Free Radic Biol Med 1998;25:576–85. Najmowicz T, Wyszkowski M, Ciecko Z. Effect of soil contamination with arsenic and application of different substances on the manganese content in plants. J Elem 2010;15: 549–58. Neidhardt H, Norra S, Tang X, Guo H, Stüben D. Impact of irrigation with high arsenic burdened groundwater on the soil–plant system: results from a case study in the Inner Mongolia, China. Environ Pollut 2012;163:8–13. Nuss ET, Tanumihardjo SA. Maize: a paramount staple crop in the context of global nutrition. Compr Rev Food Sci Food Saf 2010;9:417–36. Okieimen FE, Uwumarongie-Ilori EG, Ikhuoria EU. Effect of organic amendments on metal accumulation by maize (Zea mays L.) in contaminated soil. Int J Agric Sci 2011;1: 366–72. O'Reilly J, Watts MJ, Shaw RA, Marcilla AL, Ward NI. Arsenic contamination of natural waters in San Juan and La Pampa, Argentina. Environ Geochem Health 2010;32:491–515. Páez-Espino D, Tamames J, Lorenzo V, Cánovas D. Microbial responses to environmental arsenic. BioMetals 2009;22:117–30. Parsons JG, Martinez-Martinez A, Peralta-Videa JR, Gardea-Torresdey JL. Speciation and uptake of arsenic accumulated by corn seedlings using XAS and DRC-ICP-MS. Chemosphere 2008;70:2076–83. Plasencia J. Aflatoxins in maize: a Mexican perspective. J Toxicol Toxin Rev 2004;23: 155–77. Pongratz R. Arsenic speciation in environmental samples of contaminated soil. Sci Total Environ 1998;224:1–3. Prabpai S, Charerntanyarak L, Siri B, Moore MR, Noller BN. Effects of residues from municipal solid waste landfill on corn yield and heavy metal content. Waste Manag 2009; 29:2316–20. Prieto-García F, Méndez-Marzo MA, Martínez-Pezina FH, Prieto-Méndez J. Presencia de metales pesados en cultivos del Valle del Mezquital, México. Rev Latinoam Recursos Nat 2007;3:100–10. Raab A, Williams PW, Meharg AA, Feldmann J. Uptake and translocation of inorganic and methylated arsenic species by plant. Environ Chem 2007;4:197–203. Razo I, Carrizales L, Castro J, Díaz-Barriga F, Monroy M. Arsenic and heavy metal pollution of soil, water and sediments in a semi-arid climate mining area in México. Water Air Soil Pollut 2004;152:129–52. Redman A, Macalady D, Ahmann D. Natural organic matter affects arsenic speciation and sorption onto hematite. Environ Sci Technol 2002;36:2889–96. Reyes MNM, Cervera ML, Campos RC, de la Guardia M. Determination of arsenite, arsenate, monomethylarsonic acid and dimethylarsinic acid in cereals by hydride generation atomic fluorescence spectrometry. Spectrochim Acta B Atomic Spectrosc 2007;62:1078–82. Rosado JL, Ronquillo D, Kordas K, Rojas O, Alatorre J, et al. Arsenic exposure and cognitive performance in Mexican schoolchildren. Environ Health Perspect 2007;115:1371–5. Rosas I, Baez A, Armienta A, Belmont R. Arsenic concentrations in water, soil, milk and forage in Comarca Lagunera, Mexico. Water Air Soil Pollut 1999;112:133–49. Ryu JH, Gao S, Tanji KK. Speciation and behavior of arsenic in evaporation basins, California, USA. Environ Earth Sci 2010;61:1599–612. Sadiq M. Solubility relationships of arsenic in calcareous soils and its uptake by corn. Plant and Soil 1986;91:241–8. Sanz E, Muñoz-Olivas R, Cámara C, Sengupta MK, Ahamed S. Arsenic speciation in rice, straw, soil, hair and nails samples from the arsenic-affected areas of Middle and Lower Ganga plain. J Environ Sci Health A Tox Hazard Subst Environ Eng 2007;42: 1–11.

187

Schulz H, Härtling S, Tanneberg H. The identification and quantification of arsenic-induced phytochelatins—comparison between plants with varying As sensitivities. Plant and Soil 2007;303:275–87. Serna-Saldivar SO, Amaya Guerra CA, Herrera Macias P, Melesio Cuellar JL, Preciado Ortiz RE, et al. Evaluation of the lime-cooking and tortilla making properties of quality protein maize hybrids grown in Mexico. Plant Foods Hum Nutr 2008;63:119–25. Simon S, Tran H, Pannier F, Potin-Gautier M. Simultaneous determination of twelve inorganic and organic arsenic compounds by liquid chromatography-ultraviolet irradiation-hydride generation atomic fluorescence spectrometry. J Chromatogr 2004;1024:105–13. Smith SE, Smith FA, Jakobsen I. Mycorrhizal fungi can dominate phosphate supply to plants irrespective of growth responses. Plant Physiol 2003;133:16–20. Srivastava M, Ma LQ, Singh N, Singh S. Antioxidant responses of hyper-accumulator and sensitive fern species to arsenic. J Exp Bot 2005;56:1332–42. Su YH, McGrath SP, Zhao FJ. Rice is more efficient in arsenite uptake and translocation than wheat and barley. Plant and Soil 2009;328:27–34. Tamaki S, Frankenberger WJ. Environmental biochemistry of arsenic. Rev Environ Contam Toxicol 1992;124:79–110. Urunmatsoma SOP, Ikhuoria EU, Okieimen FE. Chemical fractionation and heavy metal accumulation in maize (Zea mays) grown on CCA contaminated soil amended with rubber processing sludge manure. Int J Appl Environ Sci 2010;5:787–802. Wang S, Mulligan CN. Effect of natural organic matter on arsenic release from soils and sediments into groundwater. Environ Geochem Health 2006;28:197–214. Wang ZH, Zhang JL, Christie P, Li XL. Influence of inoculation with Glomus mosseae or Acaulospora morrowiae on arsenic uptake and translocation by maize. Plant and Soil 2008;311:235–44. Watts MJ, O'Reilly J, Marcilla AL, Shaw RA, Ward NI. Field based speciation of arsenic in UK and Argentinean water samples. Environ Geochem Health 2010;32:479–90. Williams PN, Price AH, Raab A, Hossain SA, Feldmann J, et al. Variation in arsenic speciation and concentration in paddy rice related to dietary exposure. Environ Sci Technol 2005;39:5531–40. Woolson EA, Axley JH, Kearney PC. The chemistry and phytotoxicity of arsenic in soils: I. contaminated. Field Soils Soil Sci Soc Am Proc 1971;35:97–100. Wu C, Ye Z, Shu W, Zhu Y, Wong M. Arsenic accumulation and speciation in rice are affected by root aeration and variation of genotypes. J Exp Bot 2011;62:2889–98. Xia Y-s, Chen B-d, Christie P, Smith FA, Wang Y-s, et al. Arsenic uptake by arbuscular mycorrhizal maize (Zea mays L.) grown in an arsenic-contaminated soil with added phosphorus. J Environ Sci 2007;19:1245–51. Xu J, Thornton I. Arsenic in garden soils and vegetable crops in. Cornwall, England: implications for human health. Environ Geochem Health 1985;7:131–3. Yu Y, Zhang S, Huang H, Luo L, Wen B. Arsenic accumulation and speciation in maize as affected by inoculation with arbuscular mycorrhizal fungus Glomus mosseae. J Agric Food Chem 2009;57:3695–701. Zavala YJ, Gerads R, Gürleyük H, Duxbury JM. Arsenic in rice: II. arsenic speciation in USA grain and implications for human health. Environ Sci Technol 2008;42:3861–6. Zhao FJ, Ma JF, Meharg AA, McGrath SP. Arsenic uptake and metabolism in plants. New Phytol 2009;181:777–94. Zhao FJ, McGrath SP, Meharg AA. Arsenic as a food chain contaminant: mechanisms of plant uptake and metabolism and mitigation strategies. Annu Rev Plant Biol 2010; 61:535–59. Zheng MZ, Cai C, Hu Y, Sun GX, Williams PN, et al. Spatial distribution of arsenic and temporal variation of its concentration in rice. New Phytol 2011;189:200–9. Zhu YG, Sun GX, Lei M, Teng M, Liu YX, et al. High percentage inorganic arsenic content of mining impacted and nonimpacted Chinese rice. Environ Sci Technol 2008;42: 5008–13.

Positional cloning in maize (Zea mays subsp. mays, Poaceae).

Comparative analysis of maize (Zea mays) crop performance: natural variation, incremental improvements and economic impacts.

Quantitative trait loci for mercury accumulation in maize (Zea mays L.) identified using a RIL population.

Influence of Rhizophagus irregularis inoculation and phosphorus application on growth and arsenic accumulation in maize (Zea mays L.) cultivated on an arsenic-contaminated soil.

Accumulation of Hydrocarbons by Maize (Zea mays L.) in Remediation of Soils Contaminated with Crude Oil.

Accumulation potentials of perfluoroalkyl carboxylic acids (PFCAs) and perfluoroalkyl sulfonic acids (PFSAs) in maize (Zea mays).

Ultrastructural studies on pollen embryogenesis in maize (Zea mays L.).

Ontogeny of the sheathing leaf base in maize (Zea mays).

Iron acquisition in maize (Zea mays L.) using Pseudomonas siderophore.

Plantlet regeneration from glume calli of maize (Zea mays L.).

Agrobacterium-mediated transformation of maize (Zea mays) immature embryos.

Comparative impact of genetically modified and non modified maize (Zea mays L.) on succeeding crop and associated weed.

Accumulation of 5-hydroxynorvaline in maize (Zea mays) leaves is induced by insect feeding and abiotic stress.

Growth, cadmium uptake and accumulation of maize (Zea mays L.) under the effects of arbuscular mycorrhizal fungi.

Effect of exogenous selenium supply on photosynthesis, Na+ accumulation and antioxidative capacity of maize (Zea mays L.) under salinity stress.

Genetic Effects Conferring Heat Tolerance in a Cross of Tolerant × Susceptible Maize (Zea mays L.) Genotypes.

Early responses to gibberellic acid in a dwarf maize mutant (Zea mays L. d 1).

A linkage map based on information from four F2 populations of maize (Zea mays L.).

A metallothionein-like gene from maize (Zea mays). Cloning and characterization.

Genetic analysis of arsenic accumulation in maize using QTL mapping.

Possible mitochondrial involvement in mechanism of cytoplasmic male sterility in maize (Zea mays L.).

Genomic Regions Associated with Root Traits under Drought Stress in Tropical Maize (Zea mays L.).

Cu-chitosan nanoparticle boost defense responses and plant growth in maize (Zea mays L.).

Codominant isoenzymic alleles as markers of genetic diversity correlated with heterosis in maize (Zea mays).