1Dept. Plant, Soils, and Biometeorology, Utah State Univ., Logan, UT, 84321 U.S.A. Email firstname.lastname@example.org
2Dept. Crop and Soil, Environmental Science, Virginia Polytech. Inst. & State Univ., VA 24601 U.S.A. Email email@example.com
3Dept. Plant and Soil Sciences, Univ. of Delaware, Newark, DE 19717 U.S.A. Email firstname.lastname@example.org
The bioavailability of arsenic (As) in aerobic systems is governed by its adsorption onto metal oxide – most frequently iron (Fe) oxide - solid phases. The focus of this study is on how naturally occurring soil constituents (i.e. humic substances, silica, and phosphate) enhance the bioavailability of arsenate (As(V) oxyanion) and what effect enhanced As bioavailability has on food crops. We found that in the natural pH range of soils (3 to 10), As(V) bioavailability was enhanced in the presence of humic, fulvic, and citric acids, which inhibited the adsorption of As(V) onto Fe oxide (goethite and ferrihydrite) surfaces. Others have found that phosphate also enhances As bioavailability. To determine the effect of increased As bioavailability on As uptake and phytotoxicity into edible food crops and As food chain transfer, lettuce and radish plants were grown either in nutrient solution culture or Fe-oxide coated sands. Exposure to 0.2 mM As(V) significantly reduced total radish biomass by 44%, while exposure to As(V) levels ranging from 0.01 to 0.06 mM did not affect total lettuce biomass. Enough As accumulated in radish roots including edible root tissue, to be potentially hazardous to humans. No As accumulated in the leafy edible portion of lettuce. It was all sequestered in the lettuce roots. We surmise that iron oxide plaques on the lettuce roots provided a sorption surface for As, thus preventing As movement into the plant.
Arsenic, arsenate, ferrihydrite, iron oxides, plant uptake
High levels of arsenic (As) have been reported in areas where arsenicals have been used as pesticides, herbicides, and fungicides, and in areas of mining (Tamaki and Frankenberger 1992; Bowell 1994; Carbonell-Barrachina et al. 1997). High levels of As have also been reported in well water where surrounding bedrock is naturally high in As. In fact, well water contamination in the Bengal Basin has been described as the worst mass poisoning of a population in history (Smith et al. 2000). Here, at least 28 million people consume water containing greater than 50 μg/L As, which is the current drinking water standard in the US and Bangladesh. Levels exceeding 50 μg/L As are considered hazardous to humans (Dhar et al. 1997; Ullah 1998).
In soil and aquatic environments, inorganic As exists in the +3 or +5 valance state as the oxyanions, arsenite (AsO33-) and arsenate (AsO43-), respectively. From this point forward we will be referring to arsenite as As (III) and arsenate as As (V). Their relative occurrence is a function of pH and electrochemical potential (Eh) (Masscheleyn et al. 1991). Typically, As (V) is present under aerobic conditions, while As (III) is the predominant form under waterlogged (anaerobic) conditions. The bioavailability of either form of As in natural waters is regulated by their sorption onto soil colloidal surfaces. Both As (V) and As (III) bind via ligand exchange mechanisms to variably charged surfaces of aluminum and iron oxides, which is dependent on pH. For example, sorption of As (V) on the iron oxide goethite tends to decrease with increasing pH . Adsorption to these surfaces is mostly favored at or around the pKa1 of the oxy-acid, and may be reduced by competing ligands such as phosphate (Jacobs et al. 1970; Fendorf et al. 1997; Grossl et al. 1997). In addition to phosphate, other soluble soil constituents such as soluble organic matter (humic, fulvic and citric acids) and silica can also inhibit adsorption of both arsenate and arsenite to metal oxide surfaces (Grafe et al. 2001 2002; Waltham and Eick 2002). However, it is not clear to what extent this inhibition process, by rendering As bioavailable, facilitates food chain transfer of As in agro-ecosystems, and contributes to possible crop production losses due to As phytotoxicity.
Generally, As(V) binds more strongly to soil constituent surfaces than As(III) and consequently, As(III) is considered more mobile in soils. It is also considered more toxic to living organisms. Arsenite has a high affinity for sulfhydryl groups present in amino acids such as cysteine, thus, it deactivates many enzymes in involved in intermediate metabolism processes (Ehrlich 1990). For example, As(III) binding to sulfhydryl groups destroys radicular root membranes upon contact, ultimately affecting water and nutrient uptake in plants (Carbonell-Barrachina et al. 1994; Carbonell-Barrachina et al. 1997; Carbonell-Barrachina et al. 1998). Arsenate acts as an analog of phosphate and disrupts the glycolytic pathway by uncoupling substrate-level phosphorylation (Stryer 1981; Carbonell-Barrachina et al. 1997). Thus, substitution of As(V) for phosphate restricts ATP production and the plant is effectively deprived of its source of energy. Additionally, As(V) inhibits enzyme activity and production of chlorophyllase and other photosynthesis related enzymes (Jiang and B.R. 1994; Jain and Gadre 1997).
Diet is the largest source of As exposure (excluding exposures due to occupation or As polluted environments) to the general population (Tao and Bolger 1998). The highest levels of total As are found in seafood, followed by rice and/or rice cereal, mushroom and poultry (Tao and Bolger 1998). The US daily diet contains less than 0.04 mg As but it can reach 0.2 mg As per day if the diet includes seafood (Klaassen 1996). However, most of the As found in seafood exists as organic, non-toxic forms (ATSDR 2000; Larsen et al. 1993). Although vegetable crops usually do not contain critically high As to warrant concern, there is a question of how much As may accumulate in edible portions of food crops growing on As contaminated soils. Carbonell-Barrachina et al. (1999) reported that high levels of As where associated with hydroponically grown turnips treated with both As(III) and As(V). They found the greatest amount of As on the outer skin of the turnip roots where concentrations reached 116 mg kg-1 (dry mass); whereas, the inner potion of the root only reached 42 mg kg-1 (dry mass). It appears that the root surface sorbs As from solution. They concluded that consumption of turnips from As contaminated soils is potentially dangerous to human health, especially if the peels are consumed as part of the edible portion of turnip.
In this paper we will focus mostly on As (V) since it is generally the form that accumulates in arable land soils. Our objective was to examine the effect of As (V) on radish and lettuce growth and subsequent As uptake and partitioning. Radish and lettuce are two edible food crops where exposure and transfer of As to edible portions is quite different – radish being a crop where the edible root tissue is directly exposed to As, while the edible leafy portion of lettuce can only receive As indirectly through translocation via roots. Experiments were conducted in hydroponic solution cultures under controlled environment conditions to determine radish and lettuce response to increasing levels of soluble As (V); and greenhouse studies where plants are being grown in ferrihydrite (an amorphous iron oxide) coated sands spiked with increasing levels of As (V) to assess its biogeochemistry in iron oxide laden systems.
Hydroponic Growth Studies
Radish (Raphanus sativus), Cherry Belle and lettuce (Lactuca sativus) (vigor pak: Waldman’s Green Lettuce) seedlings were placed in a 2 L brown Nalgene™ bottle with roots suspended from the mouth of the bottle by a Styrofoam disc. The bottles were filled to the bottom of the disc with a starter nutrient solution. An aeration tube was inserted into each bottle and the bottles were placed in a controlled environment growth chamber at 21 C receiving 12 h of light per day at a PPF of 400 μmol m-2 s-1. All plants were allowed to acclimate in nutrient solutions for 5 d prior to initiation of As treatments consisting of 0, 0.02, 0.1, and 0.5 mM of arsenate (As (V)) as Na2HAsO4. The pH, EC, Eh and dissolved oxygen content were monitored daily using hand held portable meters. Every two days pH was adjusted to 5.5 with 0.1 M HNO3 and nutrient refill solutions were applied and the amounts recorded. Treatments were arranged in a completely randomized factorial design with each treatment run in triplicate. After a 30 day growth period, radishes and lettuces were harvested. Plant roots were washed with copious amounts of doubly deionized water before being separated into above round and root portions. Separated plant parts were oven dried at 60oC and dry matter production was recorded before digestion. Eh, pH, EC, and DO levels were recorded and 10 ml nutrient solutions from each bottle were sampled and submitted for elemental analysis at the end of the growth period.
Dry plant parts were ground and placed into previously acid-washed 125 ml Erlenmeyer flasks, treated with 8 ml concentrated (15.8N) nitric acid, capped with plastic funnels, and allowed to stand overnight. The suspensions were heated for one hour at 120oC and allowed to cool. Appropriate amounts of 30% H2O2 were added to each digest, heated for 30 min at 120oC, and allowed to cool down and repeated two more times to insure complete digestion. After the final cooling period, 50 ml of doubly deionized water (DDI) were added to the contents of each flask and mixed well. A sub-aliquot was taken and diluted with DDI to 25 ml using previously acid-washed 25 ml volumetric flasks. The diluted digests were analyzed for total elemental content by Inductively Coupled Plasma spectrometry from Jarrel-Ash®.
Greenhouse study - Ferrihydrite coated sands
Radish and lettuce seedlings were transplanted into pots containing 300 grams of ferrihydrate coated sand prepared in the laboratory. Treatments including the addition of increasing levels of As(V) to a maximum level of 50 mg/kg As. Treatments were replicated 5 times. Controls consisted of the same treatments in pure quartzite sand. Additional studies will examine the effect of increasing phosphorus and other soil constituents (silica and humic substances) on As bioavailability in iron oxide coated sands.
Radish plant biomass decreased with increasing As(V). For example, relative to control treatments (0 mM As), total biomass decreased by 44% with the addition of 0.02 mM As (V). The impetus of the study was to determine the effect of As on edible tissue, which for radish would consist of the root portion of the plant. Arsenic phytotoxicity affects water mobility into and through plants with symptoms that include root and plant discoloration, leaf wilting, necrotic leaf tips and margins, and ultimately plant death (Woolson et al. 1971). All roots developed edible tissue at 0.02 mM As(V). No edible radish tissue formed at As(V) levels of 0.1 mM and above. In the absence of As, roots comprised 54% of the total radish plant biomass. The As (V) had no affect on partitioning, where growth declined equally between shoots and roots in response to increasing As(V) concentrations. Arsenic accumulation in radish roots did increase linearly with exposure to increasing As(V) concentrations (Figure 1).
Figure 1. Arsenic levels in radish root tissue as a function of As(V) treatment concentrations. A single least significant difference (LSD 0.05) was calculated to compare As treatments.
There was no statistically significant decrease in lettuce biomass with exposure to increasing As (V). No As accumulated in edible lettuce shoots: however, lettuce roots exhibited the same linear increase in As accumulation increasing As (V) concentrations. We surmise that the fibrous root system of lettuce may have provided a surface that facilitated the formation of iron oxide plaques, which would sequester and prevent As movement into the plant. We have found such plaques on root surfaces of other hydroponically grown plants and identified the plaque as ferrihydrite, an amorphous iron oxide (Mackowiak and Grossl 1999). We also found that these root plaques were a major sink for phosphate. Similarly, Carbonell-Barrachina et al. (1999) found that As was strongly sorbed to the surfaces of hydroponically grown turnip roots. In fact, root surfaces had a significantly greater amount of As than corresponding inner root tissue (Carbonell-Barrachina et al. 1999). Perhaps, iron oxide plaques may have formed on the turnip root surfaces providing a sink for the adsorption of As.
Arsenic (V) levels in soil solutions are regulated by the sorption/desorption of As(V) on iron oxide surfaces (Dzombak and Morel 1990; Grossl et al. 1997). A conceptual framework for the biogeochemistry of As(V) in an Fe-oxide rich system is illustrated in Figure 2.
Figure 2. Biogeochemistry of As(V) in an Fe-oxide rich system. X represents a soil constituent ion (i.e. phosphate, organic acids, silica, etc.) that blocks the adsorption of As(V) onto the Fe-oxide surface and renders As(V) bioavailable.
Based on our results and using established equilibrium constants for As(V) sorption on ferrihydrite listed in Dzombak and Morel (1990) we can estimate how much As might accumulate in food crops growing in As contaminated soils. Arsenate adsorption on ferrihydrite is represented by Equation 1 below:
≡FeOH + H2AsO4- = ≡FeHAsO4 + H2O
The equilibrium constant for equation 1 is 105.05 and ≡FeOH and ≡FeHAsO4 symbolize the ferrihydrite surface with and without adsorbed As(V), respectively. At pH 5.5 (the pH of hydroponic solutions), H2AsO4- is the predominant As (V) species present in solution. The equilibrium expression for equation 1 is (H2AsO4-)-1 = 105.05, where ( ) represent solution activity. Thus, at equilibrium As (V) activity would be 10-5.05 M in solutions of soils laden with ferrihydrite. Using the Davies equation and assuming that the ionic strength of a non-saline soil solution was 0.01 M, we calculated an activity coefficient (γ1 ) = 0.662 (Lindsay 1979). Since ci = ai/γi, where c is the concentration of an ion i, and a is the activity of an ion i, an As(V) activity of 10-5.05 M translates to a solution concentration of 13 μM. This value will likely increase at pH >8.5, since the mobility of As(V) increases in high pH soils (Pierce and Moore 1982; Xu et al. 1991; Darland and Inskeep 1997) which is attributed to decreased As(V) adsorption on iron oxides with increasing pH (Goldberg 1986; Grossl et al. 1997). Based on our results, As(V) was toxic to radish at 20 μM, but should not be phytotoxic in an acidic soil, where we predict an As (V) concentration of 13 μM as regulated by its adsorption/desorption on iron oxides. However, in our experiments As was applied only once, whereas in soil environments As(V) is continuously supplied. It is not known how a low but chronic dose of As(V) would compare to an acute exposure of As (V). This warrants further study. Using the linear relationship established between As accumulation in radish roots and As(V) treatment dose (Figure 1) we predicted that edible radish roots growing in a soil with an As(V) solution level of 13 μM would accumulate about 40 μg/g As on a dry weight (DW) basis. This is consistent with As levels measured in turnip roots grown in hydroponic solutions treated with the same level of As (V), where the outer skin of turnip roots contained 84 mg/kg As and the inner root contained 10 mg/kg As (Carbonnell-Barrachina et al. 1999). Assuming that typical root crops have water contents near 90% (Maynard and Hochmuth 1997), an As content of 40 μg/g dry weight would equate to an As fresh weight content of 4 μg As / g. If a serving of raw vegetables corresponds to about 100 grams fresh weight, then one serving of a root vegetable containing 4 μg As / g fresh weight would provide 400 μg As. This is almost 3 times the World Health Organization’s (WHO) provisional tolerable intake level for inorganic As which for an adult male weighing 70 kg would be 147 μg As per day. Eating a leafy vegetable, such as lettuce, growing on As contaminated soils poses less threat to human health than a root crop, such as radish. Our results indicate that for lettuce plants growing in a solution containing 13 μM As(V), all the As would be sequestered by roots and little, if any, transported to leaves.
Naturally occurring soil constituent ions (phosphate, organic acids, and silica) can inhibit the adsorption of As(V) onto Fe-oxide particle surfaces (Jacobs et al. 1970; Grafe et al. 2001 2002; Waltham and Eick 2002). This should maintain greater As levels in the soil solution and increase As availability to plant roots (Figure 2); yet, the extent of this phenomenon is unknown. We are concerned that competing ions may increase As bioavailability to a level that it could threaten the environment and ultimately human health. Results from the ongoing greenhouse study with ferrihydrite coated sand are forthcoming and may shed some light on this issue.
ATSDR (2000) Toxicological Profile for Arsenic. Agency for Toxic Substances and Disease Registry. U.S. Department of Health and Human Services, Atlanta, GA.
Bowell RJ (1994) Sulphide oxidation and arsenic speciation in tropical soils. Environmental Geochemistry and Health 16, 84.
Carbonell-Barrachina AA, Burlo F, Burgos-Hernandez A, Lopez E and Mataix J (1997) The influence of arsenite concentration on arsenic accumulation in tomato and bean plants. Scientia Horticulturae 71, 167-176.
Carbonell-Barrachina AA, Burlo F, Lopez E and Mataix J (1998) Tomato plant nutrition as affected by arsenite concentration. Journal of Plant Nutrition 21, 235-244.
Carbonell-Barrachina AA., Burlo F and Mataix Beneyto J (1994) Effect of arsenite on the concentrations of micronutrients in tomato plants grown in hydroponic culture. Journal of Plant Nutrition 17, 1887-1903.
Carbonell-Barrachina AA, Burlo F, Valero D, Lopez E, Martinez-Romero D and Martinez-Sanchez F (1999) Arsenic toxicity and accumulation in turnip as affected by arsenic chemical speciation. Journal of Agricultural and Food Chemistry 47, 2288-2294.
Dahr RK, Biswas BK, Samanta G, Mandal BK, Chakraborti D, Roy S, Jafar A, Islam A, Ara G, Kabir S, Khan AW, Ahmed SK and Hadi SA (1997) Groundwater arsenic calamity in Bangladesh. Current Science 73, 48-59.
Darland JE and Inskeep WP (1997) Effects of pH and phosphate competition on the transport of arsenate. Journal of Environmental Quality 26, 1133-1139.
Dzombak DA and Morel FMM (1990) 'Surface Complex Modeling - Hydrous Ferric Oxide.' (John Wiley and Sons: New York)
Ehrich HL (1990) 'Geomicrobiology.' (Marcel Dekkar: New York)
Fendorf S, Eick MJ, Grossl P and Sparks DL (1997) Arsenate and chromate retention mechanisms on goethite. 1. Surface structure. Environmental Science and Technology 31, 315-320.
Goldberg S (1986) Chemical modeling of arsenate adsorption on aluminum and iron oxide minerals. Soil Science Society of America Journal 50, 1154-1157.
Grafe M, Eick MJ, Grossl PR (2001) Adsorption of arsenate and arsenite on goethite in the presence and absence of dissolved organic carbon. Soil Science Society of America Journal 65, 1680-1687.
Grafe M, Eick MJ, Grossl PR, Sanders AM- (2002) Adsorption of arsenate and arsenite on ferrihydrite in the presence and absence of dissolved organic carbon. Journal of Environmental Quality 31, 1115-1123.
Grossl PR, Eick M, Sparks DL, Goldberg S, and Ainsworth CC (1997) Arsenate and chromate retention mechanisms on goethite. 2. Kinetic evaluation using a pressure-jump relaxation technique. Environmental Science and Technology 31, 321-326.
Jacobs LW, Syers JK, and Keeney DR (1970) Arsenic sorption by soils. Soil Science Society of America Journal. 34, 750-754.
Jain MJ and Gadre RP (1997) Effect of arsenic on chlorophyll and protein contents and enzymatic activities in greening maize tissues. Water, Air, and Soil Pollution 93, 109-115.
Jiang QQ and Singh BR (1994) Effect of different forms and sources of arsenic on crop yield and arsenic concentration. Water, Air, and Soil Pollution 74, 321-343.
Klaassen CD (1996) 'Casarett & Doull's Toxicology - The Basic Science of Poisons Fifth Edition.' (McGraw-Hill Health Professions Division: New York)
Larsen EH, Pritzl G and Hansen SH (1993) Arsenic speciation in seafood samples with emphasis on minor constituents: an investigation using high performance liquid chromatography with detection by inductively coupled plasma mass spectrometry. Journal of Analytical Atomic Spectrometry. 8, 1075-1084.
Lindsay WL (1979) 'Chemical Equilibria in Soils.' (John Wiley and Sons: New York)
Mackowiak CL and Grossl PR (1999) Iron phosphate precipitation in the absence and presence of soluble organic matter. In 'Annual Meeting Abstracts.' p.317 (American Society of Agronomy:Madison, WI)
Masscheleyn PH, Delaune RD and Patrick WH Jr. (1991) Effect of redox potential and pH on arsenic speciation and solubility in a contaminated soil. Environmental Science and Technology 25, 1414-1419.
Maynard DN and Hochmuth GJ (1997) 'Knott's Handbook for Vegetable Growers - Fourth Edition.' (John Wiley and Sons, New York)
Pierce ML and Moore CB (1982) Adsorption of arsenite and arsenate on amorphous iron hydroxide. Water Research 16, 1247-1253.
Smith AH, Lingus EO and Rahman M (2000) Contamination of drinking-water by arsenic in Bangladesh: a public health emergency. Bulletin World Health Organization 78(9), 1093-1103.
Stryer L (1981) 'Biochemistry.' (W H Freeman: New York)
Tamaki S and Frankenberger WT Jr. (1992) Environmental chemistry of arsenic. In 'Reviews of Environmental Contamination and Toxicology.' (Springer-Verlag: New York) 124, 79-110.
Tao SSH, Bolger M (1998) Dietary arsenic intakes in the United States: FDA Total Diet Study, September 1991-December 1996. Food Additives and Contaminants 16, 465-472.
Ullah SS (1998) Geochemical mapping and speciation of arsenic in groundwater of Faridpur municipality, Bangladesh. Journal Bangladesh Academy of Sciences 22,143-147.
Waltham CA, Eick MJ (2002) Kinetics of arsenic adsorption on goethite in the presence of sorbed silicic acid. Soil Science Society of America Journal 66, 818-825.
Woolson EA, Axley JH and Kearney PC (1971) Correlation between available soil arsenic, estimated by six methods, and response of corn (Zea mays L.). Soil Science Society of America Proceedings 35, 101-105.
Xu H, Allard B and Grimvall A (1991) Effects of acidification and natural organic materials on the mobility of arsenic in the environment. Water, Air, and Soil Pollution 57, 269-278.