Abstract

 

Copper oxide nanoparticles (CuO NPs) are widely used in many industries. The increasing release of CuO NPs from both intentional and unintentional sources into the environment may pose risks to rice plants, thereby reducing the quality or quantity of this staple grain in the human diet. Not only has arsenic (As) contamination decreased rice yield, but As accumulation in rice has also been a great human health concern for a few decades. New technologies have succeeded in removing As from water by nanomaterials. By all accounts, few studies have addressed CuO NP phytotoxicity to rice, and the interactions of CuO NPs with As are poorly described. The present study 1) reviews studies about the environmental behavior and phytotoxicity of CuO NPs and As and research about the interaction of CuO NPs with As in the environment, 2) discusses critically the potential mechanisms of CuO NP and As toxicity in plants and their interaction, and 3) proposes future research directions for solving the As problem in rice. Environ Toxicol Chem 2018;37:11–20. © 2017 SETAC

INTRODUCTION

Rice has served as an important human food source for more than 8000 yr and has played an essential role in the process of Asian civilization and urbanization [1]. Historically, rice has been and continues to be the main food source for half of the world's population, because of its easy transportation and storage and relatively short cooking time [2, 3]. Globally, 42% of the 2868 kcal energy consumed daily by the average person comes from cereal crops in which rice is the dominant grain [4]. Rice also serves as a mainstay or supplementary food for people on restricted diets. For example, rice replaces simple carbohydrates, meat, and dairy products for lactose‐intolerant people or for those on a macrobiotic diet because of its low percentage of gluten and a slow and continuous release of glucose into the blood [5]. However, the production of rice is not keeping pace with increasing demand as the global population increases [6]. By 2050, the production of rice may be insufficient to feed the rapidly increasing human population [1].

Several pressures currently limit rice production: diminishing clean water availability as a result of weather system shifts, increasing salinity from sea‐level rise, conversion of agricultural land to other uses [6]. In addition, contamination and pollution from various anthropogenic activities disrupt or severely damage wetland functionality [7]. These include domestic, agricultural, and industrial activities such as surface runoff from concentrated animal feeding areas, effluent disposal of wastewater‐treatment plant, mine and factory discharges, and fertilizer and pesticide application. These activities can transport excessive nutrients, dissolved and suspended metals, and organic pollutants to wetland areas, including rice paddies. Particularly, arsenic (As) contamination has reduced rice yield and has become a great concern for causing a variety of adverse chronic human health effects [8].

To solve the problem of decreasing rice yield and poor quality caused by As and other contaminants, significant research is under way [9, 10, 11, 12, 13, 14, 15, 16, 17]. For example, the 3000 Rice Genomes Project is trying to identify the most important genes for rice production [18]. In addition, genetic engineering may discover or create new rice varieties—to increase yields and nutritional value and to make cultivars more resistant to diseases and pests and more tolerant to severe weather such as droughts and floods. For example, the genetically modified golden rice has the potential to solve the vitamin A deficiency problem, which has caused much of the death and disease in developing areas such as Africa and Southeast Asia [19]. However, the impact of genetically modified food has yet to be directly or widely tested, and whether it will be a lifesaver or not is as yet undetermined [18, 20, 21, 22]. Although scientific and regulatory agencies deem that biotech foods are safe, some environmental organizations have strongly opposed genetically modified crops including golden rice [22]. Resolving these differences will require input and agreement by a wide array of stakeholders such as scientists, rice producers and consumers, and regulatory agencies. In addition, nanotechnology application in agriculture may provide another promising approach to increase rice production and improve rice quality caused by As contaminants. However, some nanoparticles are phytotoxic [23, 24, 25, 26, 27, 28, 29, 30, 31, 32, 33, 34], and the effects of As combined with nano–metal oxides are unknown (Figure 1).

The present study reviews literature in the year range of 1900 to 2017. Key topic words (including “nanotechnology,” “nanomaterials,” “nanoparticles,” “copper oxide,” “arsenic,” “rice,” “agriculture,” “phytotoxicity,” “toxicity,” “dissolution”) were searched in multiple databases including Web of Science, Scopus, and PubMed. In addition, new publication alerts were created with the same key topic words to keep the information current. To the date of completion, 138 references closely related to the following topics were selected and reviewed: 1) the phytotoxicity of As and nanoparticles (NPs), particularly copper oxide (CuO) NPs; 2) the environmental behavior and interaction of CuO NPs with As in the context of environment and toxicity to plants; and 3) potential mechanisms of CuO NPs and As toxicity in plants and their interaction.

ENVIRONMENTAL BEHAVIOR OF CuO NPs AND As

Interaction of CuO NPs and As in the environment

Nanotechnology has been used in many industries including agriculture. Nanomaterials exhibit significantly different properties from their corresponding bulk materials and may interact differently with other chemicals relative to their bulk size [35, 36]. Adsorption of As by nanomaterials has been proposed as an alternative to conventional adsorbents for remediation [37]. This is probably because nanomaterials have enhanced properties and improved effectiveness compared with their bulk counterparts when interacting with As. Various metal oxide NPs, such as iron (hydr)oxides, alumina, titanium dioxide, zinc oxides, and CuO, have been used as nano‐adsorbents for As removal [37]. In particular, the maximum adsorption capacity of CuO NPs was much higher than that of its bulk counterpart. The maximum adsorption capacity of CuO NPs was reported to be 26.9 mg/g CuO NPs for As(III) and 22.6 mg/g CuO NPs for As(V), whereas the maximum removal of As from water by bulk CuO was approximately 369 μg As/g CuO [38, 39]. Moreover, As(III) adsorption to CuO NPs showed greater dependence on pH (6–11) and ranged from 62 to 83%, whereas As(V) was relatively independent of pH in the range and consistent from 90 to 97%. In addition, As(III) can be oxidized into As(V) when sorbed to the surface of CuO NPs [40]. Nevertheless, compared with conventional adsorbents, the sorption capacity of CuO NPs allows it to be used effectively without adjusting the pH or oxidizing As(III) into As(V). Moreover, the presence of competing anions (sulfate, silicate, and phosphate [P]) did not have a significant effect on As adsorption even at exceptionally high concentrations [39]. Desorption, regeneration, and reuse of CuO NPs also solve the problem of waste sludge or spent media disposal. In addition, As collected during the regeneration process of CuO NPs can be reused in the industry. The water chemistry, such as pH, major elements, and trace elements including Cu, of the treated water was seldom affected by the regeneration process of CuO NPs [40, 41].

However, no additional studies have been conducted to evaluate the interaction of As and CuO NPs in environmental systems including soils, uptake by plants, and phytotoxicity (Figure 1). Hypothetically, CuO NPs may decrease the bioavailability of As via adsorption, as mentioned previously in this section, thus potentially reducing As toxicity to plants; however, this is yet to be studied.

Production, application, and disposal of CuO NPs

Copper and its compounds are naturally present, and they have been widely used for approximately 10 000 yr [42]. Copper is an essential element for all known living creatures and is ranked as the third most important metal for human service because of widespread use in everyday life and in almost every industry [43].

Copper was first used as a fungicide in 1882 in Bordeaux, France, for protecting grape plants from Plasmopara viticola fungi. Known as “Bordeaux mixture,” the Cu‐containing mixture is still in use on several crops for preventing damage by various fungi [44, 45, 46]. However, because of their low water solubility and concomitant low bioavailability to plants, relatively large amounts of these traditional Cu‐containing agrochemicals are applied to the crops, which may cause phytotoxicity to the plant while protecting it from the phytopathogenic fungi [44]. Copper phytotoxicity manifests as seedling growth inhibition, reactive oxygen species (ROS) generation, gene alteration, DNA damage, and so on. [25, 26, 27, 28, 29, 30, 31, 32, 33, 34, 47, 48]. In addition, the widespread use of conventional Cu‐containing agrochemicals accounted for a noticeable proportion of serious environmental and human health problems reviewed in the Cu toxicological profile [49]. Moreover, Cu was identified at 921 of 1630 sites on the US Environmental Protection Agency (USEPA) National Priorities List in 2015 [50]. Therefore, it is essential to develop new products which have higher biological activity and contain less Cu in the formulation. Fortunately, the emergence and development of nanotechnology can make it a reality. Given that nanotechnology allows the precise control of nano‐scale manufacturing of the agrochemicals and the delivery vehicles, the stability of the active ingredients can be improved against transformation in the environment. Thereby, the excess amount of Cu‐containing mixture can be reduced by using nano‐sized materials (e.g., CuO NPs) with high surface activity, which increases the effectiveness of the agrochemical. For example, 3 different Cu‐based NPs including CuO NPs were tested to be more effective against Phytophthora infestans than the 4 registered Cu‐based agrochemicals including Bordeaux mixture [44]. In addition, CuO NPs improved the pest resistance of transgenic insecticidal crops by significantly enhancing the Bacillus thuringiensis toxin protein expression in the leaves and roots of B. thuringiensis transgenic cotton at a low concentration (10 mg/L) [34]. Furthermore, pollution caused by agrochemical runoff can also be reduced when using metal oxide nano‐forms, thus diminishing adverse environmental consequences [51].

Overall, the development of nanotechnology greatly promotes the application of Cu and its compounds in many industries. Copper‐based engineered nanomaterials (ENMs) are unique among the most widely used ENMs because Cu contains 3 oxidation states (Cu⁰, Cu¹⁺, and Cu²⁺) and exhibits many unique and useful physicochemical properties [52, 53]. In particular, CuO NPs are being widely used in many applications including high‐efficiency catalysts [54, 55], energy‐saving materials [56, 57], high‐temperature superconductors [58], gas sensors [59], antimicrobial agents [54, 60], environmental remediation [41, 61], and friction‐reduction and anti‐wear additives [62].

Research on new applications (e.g., biomarkers and biomedical use) has been undertaken. Conservative and optimistic estimates of the annual global market for CuO NPs from 2010 to 2025 were 200 to 830 and 330 to 1600 tons/yr, respectively; and the 16‐yr totals were 7075 and 14 320 tons, respectively [63]. On the whole, the application of CuO NPs will keep increasing. However, the application of CuO NPs and other nanomaterials in agriculture is relatively new and has not yet become a common practice, although there are nano‐fertilizers freely available on the market [64, 65]. Thus, environmental toxicologists and chemists have the opportunity to assess potential beneficial and adverse effects from using CuO NPs and other nanomaterials before they are widely adopted in agriculture. An article reviewed potential applications of nanomaterials in agriculture for several aspects: seed germination and plant growth, protection of plants and promotion of food production, detection of pesticide residue using nano‐sensors, and detection of plant pathogens with a nano‐diagnostic tool [64].

Like any other material, CuO NPs may be released to the environment during any stage of a product's life cycle including manufacturing, delivery, application, and disposal [66, 67, 68, 69, 70]. Therefore, CuO NPs may also enter the environment from sources other than direct use in agriculture. Nanoparticles are introduced into the environment by 2 general pathways: intentional and unintentional release. Intentional release includes remediation of contaminated soils and groundwater with NPs, such as CuO NPs being used as an antibacterial agent [54] and as an adsorbent to remove As from drinking water [39]. Atmospheric emissions, solid waste, and liquid sewage from production industry facilities are unintentional release pathways of CuO NPs into the environment [54, 71]. Wastewater effluents, direct discharges, or accidental spillages to the aquatic systems are all possible ways for CuO NPs to travel a long distance from their sources. In addition, the wind or rainwater runoff can transport CuO NPs and redistribute it in the environment. Keller et al. estimated the fate of global ENMs with a time‐integrated, mass balance approach [72, 73]. In 2010, of the 260 000 to 309 000 metric tons of total global ENM production, including approximately 200 metric tons of CuO NPs and other nano Cu compounds, 63 to 91% ended up in landfills, 8 to 28% in soils, 0.4 to 7% in water bodies, and 0.1 to 1.5% in the atmosphere. Eventually, most of that in the atmosphere will deposit on land and water surfaces.

Environmental behavior of CuO NPs

The high demand and application of CuO NPs will likely increase their release into the environment, but the environmental behavior of CuO NPs has not yet been well characterized. For example, most previous studies determined the dissolved Cu percentage after a specific time; however, the NP concentration supersaturated with CuO varied by several orders of magnitude (0.025–60%) [74]. And only a few studies reported the dissolution rates of CuO NPs [74, 75, 76, 77, 78].

The behavior of CuO NPs depends on the intrinsic physiochemical properties and the chemistry of the surrounding environment [79]. In aqueous systems, aggregation, sedimentation, and dissolution control the stability of CuO NPs; and stability is a key factor determining the transformation, transport, fate, and toxicity of CuO NPs in different environmental media. Generally, water properties like ionic strength, pH, salinity, total organic carbon, natural organic material (NOM), redox potential, and other chemical components influence the short‐ and long‐term behavior of CuO NPs [79, 80, 81]. Surface charge, controlling zeta potential, and being altered with the pH change influence aggregation and disaggregation of CuO NPs. The maximum hydrodynamic diameter of CuO NP aggregates occurs near pH 10, which was defined as the isoelectric point [79]. In addition, aggregation of CuO NPs was shown to have a strong positive correlation with the ionic strength of natural waters. This relationship was most pronounced within the concentration range (0.03–0.15 M) of CuO NPs and was independent of pH [79]. The dissolution of aqueous CuO NPs and the ionic Cu fraction were significantly influenced by complex‐forming ions and the presence of NOM [80]. These constituents also influence aggregation via electrostatic stabilization mechanisms and electrosteric repulsion because the adsorption of NOM to the CuO NP surface may reduce the positive charges of the particle [79]. Sedimentation of CuO NPs was enhanced by high salinity in the water column. Particularly, CuO NP sedimentation appeared to be controlled by P because P has a strong ability to covalently bond to metal oxides, making it a precursor in CuO NP sedimentation [80]. In addition, the fate and toxicology of CuO NPs were influenced when the surface charge of CuO NPs shifted from positive to negative, thereby altering their interactions with other ions and substances. Sulfidation of CuO NPs with soluble sulfide was also studied. A rapid sulfidation was observed, and the dissolution–precipitation mechanism was involved [82].

Bioavailability of metal‐based NPs is controlled by their environmental behaviors, which depend on many factors mentioned earlier in this section, especially the dissolution in the aqueous phase [83, 84]. Quantifying the bioavailability of nanomaterials and the released constituents is key to explaining the toxicity [77]. An intelligent nonexperimental modeling method, nano‐quantitative structure—activity relationships, was developed to predict the toxicity of metal‐based NPs. This method was based on 26 physicochemical properties of the metal and their cytotoxic effects in Escherichia coli. This model can be useful in evaluating the bioavailability and toxicity of metal‐based NPs in the future [85].

Environmental behavior of As

Arsenic occurs naturally in the environment and is generally combined with other elements, especially in minerals and ores. Arsenic minerals are often associated with base metals such as copper, lead, tin, and zinc and precious ores such as gold and silver. Primary natural sources of As include weathering, bioturbation, and dissolution of As minerals. Mining and processing (e.g., extraction and refining) of such ores have produced a wide scale of industrial pollution by inorganic As–containing waste, which remains a primary anthropogenic source of As in water bodies [41]. Other anthropogenic sources of As include solid by‐product disposal, water discharge from various industrial processes (e.g., coal combustion, wood preservation, glass production, and in situ extraction processes of oil and natural gas), and arsenical pesticide application in agriculture [49, 86, 87]. Moreover, As was identified at 1143 of 1630 sites on the USEPA National Priorities List in 2015, ranking first among the 848 substances found at hazardous waste sites [50].

Arsenic has a relatively high mobility over a wide range of redox conditions in aquatic systems [88]. The 2 common inorganic As forms, As(III) and As(V), usually coexist together, whereas As(V) predominates in an oxidizing environment, and As(III) predominates in reducing conditions [49]. Redox potential together with pH are the most important factors influencing As speciation [88, 89]. In oxidizing conditions, H₂As₄O^– is the dominant species at low pH (<6.9), whereas HAsO₄^2– becomes dominant at higher pH. In extremely acidic and alkaline conditions, H₃AsO₄ and AsO₄^3– may be present, respectively. In addition, the uncharged As acid (H₃AsO₃) dominates in reducing conditions at most environmentally relevant pH values (<9.2) [88, 90, 91]. Speciation, temperature, salinity, redox potential, pH, and ionic strength are important factors in determining As bioavailability, as are many minerals (e.g., FeS) which bind or sequester metals, thereby controlling As distribution among biotic and abiotic compartments [49, 92, 93, 94, 95].

The forms As(V) and As(III) have different affinities for minerals, which influence the mobility of As in the environment and bioavailability to plants: As(V) has a high affinity for iron oxyhyroxides, manganese oxides, aluminum (hydr)oxides, and clay minerals [11, 14, 96]. Under oxidizing conditions and when pH is below 8.5, As(V) tends to adsorb tightly onto those mineral constituents, which makes it relatively less mobile in the soil; As(III) has a lower affinity for these solid phases, which makes it more mobile. Under strongly reducing conditions, As can precipitate as sulfide minerals [88]. Both As(V) and As(III) formed surface complexes on FeS at pH 5.5 to 6.5 with high As loadings [97]. However, with the coexistence of P, which competes with As for sorption sites on FeS, the mobility of As can be increased [95]. Therefore, FeS minerals greatly influence the speciation, mobility, and partitioning of As in a sulfidic environment [95]. Moreover, As(III) has a greater tendency to partition into the solution phase in the presence of ferric (hydr)oxide [9, 11, 98]. Although the adsorption capacity of ferric (hydr)oxide is greater for As(III) than for As(V), the desorption rate of As(III) is much greater than that of As(V) [99]. This is because As(III) has an outer‐sphere and multiple inner‐sphere complexes, whereas As(V) has only one inner‐sphere complex [100, 101]. It was demonstrated that As concentration in flooded paddies was increased under anaerobic conditions by reductive dissolution of As (e.g., because of the activity of Fe‐reducing bacteria) [9]. Although not so common, reduction of As(III) to As(II) was also observed under a slightly acidic environment with high As loadings [102, 103]. All of the behaviors mentioned in this section influence the As bioavailability and accumulation by rice plants.

BIOACCUMULATION AND PHYTOTOXICITY OF CuO NPs AND As

Plants provide an important pathway for NP bioaccumulation into the food chain. The pore diameter of the plant cell wall ranges from 5 to 20 nm [104], functioning as a sieving process, which allows NPs with sizes less than the pore diameter to traverse the cell wall easily. In addition, NP interaction with cell walls may enlarge the pore size of the plant cell wall and facilitate the entrance of NPs [30]. Once through the cell wall, NPs may traverse the plasma membrane assisted by the embedded protein transport carrier, or they may enter the interior of the cell directly through ion channels [30]. When applied to leaves, NPs can also enter the plants through stomatal openings and trichomes [30]. The ability of NPs to enter leaf cells depends on the plant species. For example, CuO NPs aggregated on the epidermis of conventional cotton leaves, whereas it entered transgenic cotton leaf cells by endocytosis [34].

Nanomaterials (e.g., ZnO, Al₂O₃, and CuO NPs) can cause phytotoxicity after entering the plant [23, 24]. The wide array of ENMs may affect different plants through various routes, thereby causing a range of physiological effects. In addition, a single NM may have differential effects on several plant species [25, 26, 27, 28, 29, 30, 31, 32, 33, 34]. The bioaccumulation of Cu‐based ENMs was found to cause toxicity in many types of organisms, such as fungi [105, 106], mussels [107], amphipods [108], Daphnia magna [84, 109], worms [110], and plants [25, 26, 27, 28, 29, 30, 31, 32, 33, 34]. In particular, CuO NPs were reported to cause phytotoxicity to various plants. Exposure of Indian mustard (Brassica juncea L.) to CuO NPs (0, 20, 50, 100, 200, and 500 mg/L) inhibited both root and shoot growth in a dose‐dependent manner [32]. At 500 mg/L CuO NPs inhibited the growth (height, fresh weight, leaf surface area) of maize (Zea mays L.) [47]. It significantly inhibited the growth (height, root length, root number, and biomass production) of transgenic and conventional cotton plants at concentrations >10 mg/L [34]. And CuO NPs (10–100 mg/L) induced DNA (e.g., 7,8‐dihydro‐8‐oxoguanine) damage in agricultural and grassland plants (e.g., radish, perennial ryegrass, and annual ryegrass) [26]. Notably, in rice, CuO NPs caused a variety of phytotoxicity at target concentrations at different life stages of rice growth. At 5 mg/L CuO NPs severely inhibited the root growth of rice seedlings by generating ROS and influencing the expression level of 2 genes, OsCDC2 and OsCYCD, which are associated with root growth [27]. At concentrations of 0.5, 1.0, and 1.5 mM, CuO NPs decreased seed germination and shoot and root growth as well as causing oxidative damage to rice seedlings [48]. High concentrations (500 and 1000 mg/L) of CuO NPs caused adverse effects on rice growth during the whole life cycle [111]. Plant length and biomass were decreased from the tillering stage. Fatal effects occurred after plants were treated with a high concentration (1000 mg/L) of CuO NPs at the heading stage. The 2 higher concentrations of CuO NPs also decreased grain yield. Based on previous studies, potential mechanisms of NP toxicity are summarized: 1) DNA damage or gene alteration [26, 27], 2) ROS generation and oxidative stress [82], 3) NP penetration into the cell interfering with intracellular metabolism [112], 4) metal ions released from NPs hindering enzyme function, and 5) adsorption of NPs on the surface of an organism (e.g., seed), generating locally concentrated ions [113].

In contrast to CuO NPs, of which the toxicity mainly depends on the small size and concentration of the particle, the toxicity of As is speciation‐ and concentration‐dependent [114, 115]. Inorganic As species are generally more toxic than organic species, and As(III) is much more toxic than ionized As(V) [116]. As described in the previous section (Environmental behavior of As), As speciation can be quite complicated [117]. In the process of plant uptake and transport, deprotonated As species—As(III), monomethylarsonic acid (MMA), and dimethylarsinic acid (DMA)—can behave as silicic acid analogues and the arsenate as P analogues [17, 118]. Thus, As is absorbed when competing with silicates and phosphates for the same carriers in the root plasmalemma [14, 119]. A study on Spartina alterniflora indicated the potential mechanism of As phytotoxicity [120]. Chemical forms and concentrations of As affected the macro‐ and micronutrient uptake by the plant. As a result of sharing the same uptake system of the root with P, As(V) uptake was increased under insufficient P condition, whereas organic As species decreased the P uptake by damaging the root cell metabolism. In addition, organic arsenical depressed potassium (K) uptake, whereas it increased the sodium (Na) concentration in the plant root, which was coordinated with the antagonism between K and Na. The significant concentration reduction of several essential macronutrients—P, K, calcium, and magnesium—and micronutrients—boron, Cu, and Fe—may contribute to the phytotoxicity of MMA, which was the most phytotoxic species to this marsh grass. Notably, for rice, P is an essential and usually limiting macronutrient, thus making it efficient for rice to assimilate As analogues of the P moieties [121]. However, the presence of Fe plaque showed more complex impacts on As(III) and As(V) uptake by rice plant roots [14]. Once within the plant, As species undergo metabolism, complexation, symplastic transport, subcellular localization, and xylem transport during different life stages. Thereafter, As can be remobilized from shoots to grains via phloem during the grain filling process and accumulate in the grains [119]. Within the plant, As species exert toxicological action via inhibiting adenosine triphosphate formation and other phosphorylation processes, causing oxidative stress and binding to protein sulfhydryl groups, among others [122]. As a result, growth inhibition and grain yield reduction of the plants occur because of the toxicological action [13]. In addition, the bioaccumulation of As in rice grains was increased because of the elevated concentration of As in rice paddies and the high mobility of As under anaerobic conditions (e.g., flooded paddy) [11, 121, 123]. For example, rice from mine‐impacted regions had a higher total As concentration (a high percentage of inorganic As) than that from non‐mine‐impacted regions [124]. Moreover, seasonal rainfall influenced the As concentration in irrigated groundwater and paddy field soil; thus, the As concentration in rice grains also varied in a seasonal pattern [10].

Studies have shown that nanomaterials (e.g., graphene oxide, silicon NPs, MnO₂ NPs, nano‐Fe₂O₃, and nano‐Al₂O₃) amplified or alleviated the phytotoxicity of As [125, 126, 127, 128]. However, it is as yet unknown for plants whether As and CuO NP interaction increases or decreases the toxicity relative to each individual toxicant.

HUMAN HEALTH EFFECTS CAUSED BY As ACCUMULATION IN RICE

The As species of toxicological concern may be inorganic—arsenate including all As(V) species and arsenite including all As(III)—or organometallic (MMA, DMA, trimethylarsineoxide, tetramethylarsonium, arsenochloline, arsenobetaine, and arsenosugars). Generally, inorganic species cause more acute toxicity to animals than do organic As species [129]. Although inorganic As was widely used as a pesticide in the past, it has not been used in agriculture since 1993 [130], whereas some arsenicals are still in use, especially on nonfood crops such as cotton and turf [49]. Presently, approximately 90% of As overall is used as wood preservatives, the production of which was phased out in the United States by 2004, although the stockpiles could still be sold and used [49]. Arsenic may enter the air, water, and soil and then be accumulated by organisms including plants. Humans can be exposed to As by eating food, drinking water, and breathing air. Rice is considered to be a major source of inorganic As in the human diet, which poses great risks to human health because rice plants take in approximately 10 times more As than other grain crops [15]. Moreover, the major As species (up to 90% of the total content) in rice grain are inorganic (arsenate and arsenite) [131]. This is primarily caused by the anoxic condition in which rice plants are submerged in water and by the unique physiology of the rice plant, which allows it to scavenge As from the environment and accumulate it mostly in the inorganic form [8, 12]. Overall, As in the rice grain can be elevated by 3 major anthropogenic scenarios: irrigation using surface water with elevated As to rice paddies, industrial activities contaminating paddy soils, and conversion of soils previously treated with arsenical pesticides to rice paddies [121]. Among the 3 scenarios, irrigation‐related contamination of As is becoming worse because of the shortage of clean surface water supplies [132], which may also force more people to drink groundwater. Higher As concentrations are common in deeper geologic formations [132, 133]. Arsenic‐contaminated rice will add to exposures experienced by people who drink As‐laden groundwater.

Inorganic As compounds are considered class 1 carcinogens, causing various carcinomas including skin, bladder, lung, kidney, liver, and prostate [121]. Chronic As exposure has also been linked with heart disease, muscle cramps and skin lesions, whereas acute As exposure can cause gastrointestinal damage [49]. The Joint Food and Agriculture Organization of the United Nations/World Health Organization (WHO) Expert Committee on Food Additives derived the lowest benchmark dose (0.5) of inorganic As, which was 3.0 µg/kg body weight/d [134]. This value could range 2 to 7 µg/kg body weight/d based on dietary exposure, resulting in a 0.5% risk increase in lung cancer. Currently, together with the As problem in drinking water, As contamination in rice also presents a serious global concern with respect to human health [121]. The WHO proposed maximum safe concentrations of As in rice at 0.2 mg/kg for white rice in July 2014 [135] and 0.35 mg/kg for brown rice in July 2016 [136]. In the future, the WHO may target a lower concentration for brown rice when more data from all regions are available. Because of the dominant role of rice in the total diet, China introduced a more restrictive regulatory threshold value of As, 0.15 mg/kg rice [137]. Human uptake of As from rice grain consumption depends not only on the cooking process but also on the nature of As contamination (e.g., speciation, concentration, and distribution) in the original production area. Nevertheless, to decrease rice in the diet is an efficient way to avoid extra exposure overall, especially in developing countries where rice permeates the culture [138]. However, this poses nutritional challenges for large populations.

RESEARCH NEEDS

Arsenic is known to adversely impact rice plants by causing phytotoxicity, reducing the grain yield and accumulating in the grain, thereby causing human health effects through consumption of rice. However, few studies have emphasized the impact of CuO NPs on rice plants, and no studies have evaluated the toxicity of CuO NPs in combination with As. Overall, more research needs to be done, as follows. First, to characterize the mechanistic bioavailability and uptake of CuO NPs to rice plants and eventually to the rice grain. To do this, properties of media need to be characterized, and rice plant physiology must be well understood regarding the uptake, transport, and accumulation of CuO NPs in the plant. Second, to determine the toxicity of CuO NPs in combination with As to the rice plants. Environmental behaviors (such as chemical speciation, fate, and transport) of CuO NPs should be characterized. Proteomics and relative enzyme (e.g., superoxide dismutase and catalase) activities in the plant need to be evaluated. The interaction (additive, synergistic, or antagonistic) between CuO NPs and As to cause positive or negative effects at the cellular and whole‐organism levels also needs to be determined. Third, to develop feasible farming methods to minimize As uptake by rice and other food crops. Although CuO NPs can be used as sorbents to remove As in water, it is not determined whether they can be used to decrease As bioavailability to plants. In addition, there have been studies which determined ways to decrease As uptake in rice; however, these methods are not feasible for farming applications. Altogether, it is necessary to develop methods that will minimize As uptake by rice and reduce risks from consumption of rice.

Acknowledgment

The present study was supported by the C. Gus Glasscock, Jr. Endowed Fund for Excellence in Environmental Sciences in the College of Arts and Sciences at Baylor University.

REFERENCES