Journal of Nuclear Energy Science & Power Generation TechnologyISSN: 2325-9809

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Review Article, J Nucl Ene Sci Power Generat Technol Vol: 3 Issue: 3

Uranium Fixation and Removal from Different Soil Types: Review

MF Abdel-Sabour*
International Innovative Environmental Solution Center (IIESC), Cairo, Egypt
Corresponding author : Mamdouh Fathy Abdel-Sabour
International Innovative Environmental Solution Center (IIESC), Cairo, Egypt
Tel: +966-59310-1207; Fax: +966-2-699-6950
E-mail: [email protected]
Received: June 15, 2014 Accepted: September 12, 2014 Published: September 20, 2014
Citation: MF Abdel-Sabour (2014) Uranium Fixation and Removal from Different Soil Types: Review. J Nucl Ene Sci Power Generat Technol 3:3. doi:10.4172/2325-9809.1000126

Abstract

Uranium Fixation and Removal from Different Soil Types: Review

Environmental contamination caused by radionuclides, in particular by uranium and its decay products is a serious problem worldwide. The development of nuclear science and technology has led to increasing nuclear waste containing uranium being released and disposed in the environment. The objective of this paper is to develop a better understanding of factors affecting uranium fixation and removal from different soils with different soil characteristics. Soils contaminated with uranium at concentrations above regulatory limits pose an environmental and human health risk. Investigations about uranium content and fate in soils revealed several finding depending on soil factors, radionuclide source and nature, and existing vegetation in the area. In organic rich soils U seems to be more soluble and bioavailable. Data demonstrates the effectiveness of soil amendments (Hydroxyapatite, illite and zeolite) in reducing the mobility of U, which makes in-place immobilization an effective remediation alternative. In contrast to increase U solubility and leach-ability the maximum solubilization of U was observed with a combined soil acidification and citric acid addition, which may be necessary to maximize the phyto-extraction of U from soils with a pH >6.0.

Keywords: Remediation; U-speciation; U-chemistry in soils

 

Keywords

Remediation; U-speciation; U-chemistry in soils

Introduction

The movement of both essential and non-essential trace elements through agricultural ecosystems and food chains is complex. Such elements as As, B, Cd, Cr, Cu, Hg, Ni, Pb, Se, U, V, and Zn, are generally present in soils in low concentrations but concentrations may be elevated because of natural processes and/or human activities, such as fossil fuel combustion, mining, smelting, sludge amendment to soil, fertilizer application, and agricultural practices.
In the earth’s crust, uranium is generally found as oxides, such as uranium dioxide (UO2) or tri-uranium octa-oxide (U3O8). The mineral pitchblende, the main uranium ore, consists primarily of uranium oxides. In soil, uranium is primarily (80-90%) present in the +6 oxidation state as the uranyl cation (UO22+) [1]. Speciation of uranium in soil and aqueous systems is pH-dependent. Therefore, under acidic reducing conditions, UO22+ is the predominant uranium species in the soil; under neutral conditions, hydroxide complexes such as UO2OH+, (UO2)2(OH)22+,(UO2)2(OH)5+ and (UO2)2(OH)7- and phosphate complexes such as UO2HPO40 and UO2(HPO4)22- form; under alkaline conditions, carbonate complexes such as UO2CO30, UO2(CO3)22- and UO2(CO3)34- predominate [1].
Uranium is found in all rock types in varying small concentrations [2,3]. Uranium is widely dispersed in the earth’s crust, rocks and soils at a level of about 2-4 ppm by weight [4]. The natural concentration of uranium in the earth crust is about 10−6 g/g. Also, the uranium concentration in ocean water, plants, and animal organisms is around 10−7 g/g, as a result of the solubility of U(VI) compounds in water [5].
Worldwide uranium concentrations in soil have been reported to range from 0.3 to 11.7 mg/kg [7]. The average background concentration of uranium in soil is about 2 mg/kg [8]. Volcanic eruption is another natural phenomenon that may increase the concentration of natural uranium in the soil. The redistribution of uranium and uranium progeny to both soil and water occurs often naturally in environmental circuit.
Although a significant effort has been expended over the past 40 years to evaluate and quantify the transfer of U from soils to plants, more attention needs to be given to mechanisms within the soil and plant systems, which influence its solubility, chemical speciation, mobility, and uptake by and transport in plants.
Contamination of the soil can occur either from deposition of uranium originally discharged into the atmosphere, or from waste products discharged directly into or on the ground (e.g. water containing uranium from either underground or open-pit mines). Examples of industrial activities that may result in soil deposition include uranium mining and milling, uranium processing, phosphate mining, heavy metal mining, coal use and inappropriate waste disposal [9].
The objective of this paper is to develop a better understanding of factors affecting uranium fixation and removal from different soils with different soil characteristics.

Uranium in Contaminated Soils

Natural uranium exists in three different forms (isotopes), all of which are radioactive [10]. The two most abundant isotopes, uranium-235 (0.72%) and uranium-238 (99.27%), have radioactive half-lives of about 7×108 and 4.4×109 years, respectively [4]. The average radioactivity in soil of 234U from is 0.6-1 pCi/g. Since the activity of 234U accounts approximately for one-half of the total activity in natural uranium, the value may be multiplied by two to obtain the total uranium radioactivity in soils (approximately 1.2 pCi/g) [11].
There is an increasing trend of uranium accumulating in soils due to a number of deliberate or wrong practices. Public and political pressure to solve a problem situation of this nature occurs when critical toxic levels are reached. As a consequence, there would be a risk for ecosystems, agro-systems and health. It is suggested that knowledge of the mechanisms that control U behavior must be improved and can be used for risk valuation and suggestion of remediation treatments [12].
For example, soil samples were collected around a coal-fired power plant from 81 different locations in Hungary [13]. Brown coal, unusually rich in uranium, is burnt in this plant that lies inside the boundaries of a small industrial town and has been operational since 1943. Activity concentrations of the radionuclides 238U, 226Ra, 232Th, 137Cs and 40K were determined in the samples. Considerably elevated concentrations of 238U and 226Ra have been found in most samples collected within the inhabited area. Concentrations of 238U and 226Ra in soil decreased regularly with increasing depth at many locations, which can be explained by fly ash fallout. Concentrations of 238U and 226Ra in the top (0-5 cm depth) layer of soil in public areas inside the town were 4.7 times higher, on average, than those in the uncontaminated deeper layers, which mean there are approximately 108 Bq kg-1 surplus activity concentrations above the geological background.
In particular, uranium mining and milling have caused enormous damage to the environment by means of abandoned waste accumulation and improper disposal of the radioactive material, waste dump after uranium prospection. During the uranium recovery, many natural ecosystems were heavily polluted with radioactive elements, mainly through the seepage of acid drainage waters [14]. For a long period of time uranium was leached commercially in a large number of deposits using different in situ technologies [15-17], either alkaline leaching using solutions containing carbonate and hydrocarbonate ions, or acid leaching [2,18]. Such waters are still a persistent environmental problem at many abandoned mine sites, while soils around the water flow path are polluted with radioactive elements becoming unsuitable for agricultural use. Large amounts of uranium-containing (both high- and low-level) waste are generated from activities such as fuel fabrication, fuel reprocessing, research and development. All these negative impacts influenced the quality of the environment and affected mainly surface and ground waters, soils and simultaneously polluted great areas of land and threatened the catchments of drinking water. Therefore, it is strongly evident that the contamination caused by uranium has severe negative biological effects on important groups of the soil food web [19].
The potential risk of uranium soil contamination is a global problem as about every country can be affected by one or more activities mentioned above. Depleted, enriched and natural uranium contamination in soil and water has been identified at many sites worldwide, so that measures for preventing their assimilation by plants should be considered a preliminary step towards the remediation of contaminated areas [20-22]. Thus appropriate soil remediation techniques have to be considered [23,24].
Mineralogical identification of U (as a contaminants) provides important information concerning the nature of the contamination because once the mineral form is known, its properties can then be determined from geochemical data. A new densityfractionation technique was used to concentrate U particulates from U-contaminated soils. Results from neutron-activation analysis of each density fraction showed that the U had been concentrated (up to 11-fold) in the heavier fractions [25]. Mineralogical analyses of the density fractions of these soils using x-ray diffraction, scanningelectron microscopy, and an electron microprobe showed the predominance of an autunite [Ca(UO2)2(PO4)2.10-12 H2O]-like mineral with lesser amounts of uraninite (UO2) and coffinite (USiO4) as the U-bearing minerals in these soils. The presence of reduced forms of U in these soils suggests that the optimal remediation strategy requires treatment with an oxidizing agent in addition to a carbonate-based leaching to solubilize and remove U from these soils [25].
Litaor [26] indicated that soils in Colorado, USA, contaminated with U as a result of past waste-storage practices, accidental release of oils laden with U, and low-level airborne emissions were analyzed to determine the concentration and distribution pattern of U. Soils were sampled from 118 plots of 4.05-ha by compositing 25 evenly spaced samples from the top 0.64 cm. Uranium-234 activity ranged from 25.9 to 92.8 Bq/kg, 235U activity ranged from 0.1 to 25.1 Bq/kg, whereas 238U activity ranged from 30.7 to 286 Bq/kg. Spatial correlation was not observed for 234U, implying that it is randomly distributed in the soil of the study area. Uranium-235 exhibited a spotty and localized concentration pattern with no clear relationship between known burial and spill sites, and the present distribution of 235U in the soils. Proposed wind-dispersal mechanisms were not consistent with the spatial distribution of U isotopes. Although 238U showed a pattern of localized spatial distribution, most of its observed activity was well within the natural range of 238U activity in soils.
Studies of the long-term consequences of exposing terrestrial ecosystems to aged deposits of natural and depleted uranium (for 24 years) at Los Alamos Scientific Laboratory (LASL) have been conducted by Miera et al. [27]. The highest surface-soil (0 to 2.5 cm) uranium concentrations occurred at 0 to 10 m from the detonation point and averaged 4500 mug/g. Concentrations in surface soil 50 and 200 m from the detonation point were generally less than 15% of that value. The uranium distribution to 30-cm depths for the 0- to 50-m distant sampling locations and to 10-cm depths at 50- to 200-m sampling distances showed significant penetration into the soil profile. Alluvium collected 250 m from the E-F detonation area indicated that surface (0 to 2.5 cm) uranium concentrations were about 10% of those at the detonation point, and at 2.8 km they were twice background, averaging 5 mug/g. Ratios of plant/soil uranium concentrations varied from 0.05 to 0.08. Internal tissues from two species of small mammals had tissue/soil ratios of 10-3 and 10-4, although gastrointestinal contents of these mammals had mean uranium levels greater than 10% of soil concentrations.
The recommended Canadian Soil Quality Guidelines for the protection of environmental and human health are 23 mg/kg for agricultural land use, 23 mg/kg for residential/parkland land use, 33 mg/kg for commercial land use, and 300 mg/kg for industrial land use.

Depleted Uranium (DU)

Another problem is the contamination of soil and water with depleted uranium, which has increased public health concerns due to the chemical toxicity of DU at elevated dosages [9,28,29]. For this reason, there is great interest in developing methods for U removal from contaminated sources.
Uranium like other heavy metals is a threat to both health and the environment because of its pronounced toxicity. Significant amounts of uranium have been released in the last decade with armor piercing ammunition that was manufactured from DU, not only during major conflicts but also on numerous military shooting ranges all over the world (Bosnia, Kosovo, Afghanistan, Iraq, Lebanon and several Arab countries) [30].
The mass composition of depleted uranium (DU) is almost entirely 238U (99.8%) with nearly all the 234U (0.0006%) and approximately two thirds of the 235U (0.2%) removed. DU radioactivity is approximately 60% that of natural uranium [31]. In terms of chemical, physical and toxicological behavior, DU is the same as the metallic form of natural uranium [32-34].
A field study, organized, coordinated and conducted under the responsibility of the United Nations Environment Programme (UNEP), took place in Kosovo, Serbia in November 2000 to evaluate the level of depleted uranium (DU) released into the environment by the use of DU ammunition during the 1999 conflict [30]. Representatives of six different scientific organizations took part in the mission and a total of approximately 350 samples were collected. During this field mission, the Italian National Environmental Protection Agency (ANPA) collected water, soil, lichen and tree bark samples from different sites. The samples were analyzed by alpha-spectroscopy and in some cases by inductively coupled plasma-source mass spectrometry (ICP-MS). The 234U/238U and 235U/238U activity concentration ratios were used to distinguish natural from anthropogenic uranium. All water samples had very low concentrations of uranium (much below the average concentration of drinking water in Europe). The surface soil samples showed a very large variability in uranium activity concentration, namely from 20 Bq kg-1 (environmental natural uranium) to 2.3 x 105 Bq kg-1 (18000 mg kg-1 of depleted uranium), with concentrations above environmental levels always due to DU. The uranium isotope measurements refer to soil samples collected at places where DU ammunition had been fired; this variability indicates that the impact of DU ammunitions is very site-specific, reflecting both the physical conditions at the time of the impact of the DU ammunition and any physical and chemical alteration which occurred since then. The results on tree barks and lichens indicated the presence of DU in all cases, showing their usefulness as sensitive qualitative bio-indicators for the presence of DU dusts or aerosols formed at the time the DU ammunition had hit a hard target. This result is particularly interesting considering that at some sites, which had been hit by DU ammunition, no DU ground contamination could be detected.

Uranium Chemistry in soil

Uranium existing in soil can be dissolved in solution, or ion exchanged in reaction, complexed with soil organics or precipitate as pure or mixed solids. It can move into the water, air and the food supply. The immobility of these radioactive elements in uppermost soil layers represents a problem for environment and human health, since they can be easily integrated in the food chain [20,35]. The major part of radionuclides released into the environment will finally accumulate in either the upper layer of soils or interstitial system of sediments in aquatic systems [36].
The solubility of uranium in soil is dependent on several factors such as: pH, redox potential, temperature, soil texture, organic and inorganic compounds, moisture and microbial activity [37]. Soluble forms can migrate with soil water and taken up by plants or aquatic organisms or volatilized [36].
A number of investigations were performed concerning the natural attenuation of uranium in a tailings disposal site, which revealed that a number of radionuclides exhibit significant migration potential in the presence of aqueous, low molecular weight organic compounds immobile organic matter in the form of peat or organicrich horizons in soils and sediments that may provide excellent substrates for radionuclide retention [38-40]. When uranium is dissolved in groundwater, it can be attracted to natural iron coatings located on walls of rocks through which the water is flowing, and bind to the iron coating and then move into microscopic rock pore and then incorporated into the iron coating [40-42]. On the other hand, uranium is mobilized from rock by the weathering of uraninite (UO2). The action of surface waters and groundwater causes oxidative dissolution of uraninite to the soluble uranyl ion (UO22+). Worldwide, from 27 000 to 32 000 t uranium are released from igneous, shale, sandstone, and limestone rocks annually by weathering and natural erosion [43,44].
Uranium can exist in the +3, +4, +5 and +6 oxidation states. In soil, the U4+ valence state (typically solid UO2) of uranium occurs in strongly-reducing environments and is formed by the oxidation of organic matter or iron in the soil. Tetravalent uranium forms hydroxides, hydrated fluorides, and phosphates which are strongly adsorbed and very immobile in soils.
The U6+ valence occurs in oxidizing environments (UO22+) and is strongly adsorbed by soils, forming stable complexes with many ligands - notably carbonates and organic compounds. High ligand concentrations can result in a lower positive or negative charge and increase mobility of the complexed uranium [45].
Sparingly soluble contaminants are less likely to affect human health through food chain transfers, such as plant uptake or passage through animal-based foods, because mobility in these pathways is limited by solubility [46]. Direct ingestion or inhalation of contaminated soil becomes the dominant pathway. However, both of these can be selective processes. Clay-sized particles carry the bulk of the sparingly soluble contaminants, and mechanisms that selectively remove and accumulate clay from the bulk soil also concentrate the contaminants. Erosion is another process that selectively removes clays. This project examined the degree of clay and contaminantconcentration enrichment that could occur by these processes, using U, Th and Pb as representative contaminants and using clayey and loamy soil.

Mobility and Solubility of Uranium in Soils

Soil properties that affect uranium mobility (and subsequent uptake by biota) include aeration (water saturation, high biological or chemical oxygen demand), carbonate content (organic material content, pH, parent material, weathering), and cation exchange capacity (texture, clay content, organic matter, pH). A higher soil cation-exchange-capacity will retain more uranium, while carbonate in the soil increases the mobility of uranium through the formation of anionic U and CO3 complexes [47,45]. Uranium does not migrate substantially in loam compared to sandy soils [48]. Uranium migration in soil occurs over the period of a few months, depending on sorption, and may be upwards when there is a net water deficit or downwards as a result of net leaching [48]. Soil properties reported to increase mobility and plant accumulation of uranium include acidic soils with low adsorptive potential, alkaline soils with carbonate minerals, and the presence of chelates (citric acid) [49,50].
Based on primary accumulation mechanisms in soil/or sediments, U as other heavy metals can be classified into five categories: (i) adsorptive and exchangeable, (ii) bound to carbonate phases, (iii) bound to reducible phases (Fe and Mn oxides), (iv) bound to organic matter and sulfides, and (v) lattice metals [51]. In a fractionation experiment U forms were compared [52] in two soil types (clayey and sandy soil). Also, the variation of U forms due to soil treatment (spiking) were studied. In case of the clayey soil the initial U - fractions were 45.63% as residual form, 20.69% organically bound 16.36% Mn and Fe oxides bound, 9.76% Carbonate form, 7.41% exchangeable fractions and 0.15% water soluble fractions. These fractions varied significantly when the soil was spiked with 200 mg U / Kg soil to 46.88%, 23.19%, 9.97%, 16.07%, 3.79% and 0.10% for residual, organically, Mn- Fe oxide, carbonate, exchangeable and water soluble fractions respectively. These result showed significant reduction in U-ex fraction forms and Mn- Fe bound forms with significant increase in U- carbonate form due to U application. On the other hand, in case of the untreated sandy soil, the main U - fractions were 57.42% as residual form (relatively higher residual - U form in the clayey soil) 16.10% organically bound, 13.78% Mn and Fe oxides bound, 7.22% Carbonate form, 5.23% exchangeable fractions and 0.25% water soluble fractions The application of 200 mg U/Kg soil resulted in a significant changes in U - Fractions distribution as follows : 59.26%, 11.27%, 19.59%, 6.84%, 2.90% and 0.14% for residual, organic, Mn-Fe oxides, carbonate, exchangeable and water soluble fractions, respectively. They concluded that the solubility of soil uranium depends upon the soil’s physicochemical, mineralogical and micro-morphological properties; the nature of the uranium association; and the mineralogical, morphological, and compositional characteristics of the uranium-bearing phases.
The primary abiotic and biological processes that transform uranium in soil are oxidation-reduction reactions that convert U(VI) (soluble) to U(IV) (insoluble) [53]. Further abiotic and biological processes that can transform uranium in the environment are the reactions that form complexes with inorganic and organic ligands. Research has shown that there are at least three different forms of uranium in the contaminated soil:
• uranium (VI) phosphate minerals;
• reduced U(IV) phases;
• complexed U(VI) with soil organic matter;
• A small fraction of U(VI) absorbed onto soil minerals [54].
Ibrahim and Whicker [55] investigated environmental behavior of U-series radionuclides at a uranium mine-mill in Wyoming. Plant/ soil concentration ratios were in the order 238U > 230Th > 210Po > 226Ra > 210Pb. It was concluded that for sulfuric acid leached tailings, Ra and Pb are sequestered as sulfates which were highly insoluble relative to U and Th sulfates, resulting in reduced availability for plant uptake. Soil acidity and the saturation condition at the tailings impoundment edge tend to enhance radionuclide availability for plant uptake.
Soil characteristics affect the removal process, since soil behaves as a complex sorbent. Uranium preferentially adheres to soil particles, with a soil concentration typically about 35 times higher than that in the interstitial water. Concentration ratios are usually much higher for clay soils. The concentrations and distributions of uranium among particle size fractions of the soils vary significantly [49,56].
The mobility of uranium in soil and its vertical transport (leaching) to groundwater depend on properties of the soil such as pH, oxidation-reduction potential, concentration of complexing anions, porosity of the soil, soil particle size and sorption properties, as well as the amount of water available [47,57]. Retention of uranium by the soil may be due to adsorption, chemisorption, ion exchange or a combination of mechanisms [47]. Any soil property that alters the sorption mechanism will also alter the mobility of uranium in the soil. Complexation and redox reactions control the mobility of uranium in the environment [58].
In aqueous media only U(IV) and U(VI) are stable. Some compounds, such as UCI4, decompose in aqueous media to the U(VI). In acid solution and in the body, the oxygen-containing cation UO22+, where uranium has the oxidation state VI, is the predominant form. In general, hexavalent uranium compounds are the most soluble. The reduction of U(VI) to U(IV) by abiotic and biotic processes, as well as its re-oxidation has received considerable attention because the oxidation state of uranium has a significant effect on its mobility in the natural environment.
Uranium (IV) is stable under reducing conditions and is considered relatively immobile because U(IV) forms sparingly soluble minerals, such as uraninite (UO2) or a mixed valence oxide phase like UO2.25 or U02.33. Dissolved U(III) easily oxidizes to U(IV) under most reducing conditions found in nature. The U(V) aqueous species (UO3+) readily disproportionate to U(IV) and U(VI). Under reducing conditions, the speciation of U(IV) is dominated by the neutral aqueous species U(OH)40 (aq) at pH values greater than 2 [59,60].
The estimates of the solubilities and the speciation of uranium (nature and concentration species) are predicted from thermodynamic data, taking into account the presence of inorganic ligands in the ground waters studied, mainly [OH]-, [HCO3]-, [CO3]2-, [H2PO4]- [HPO4]2-, [PO4]3-, [SO4]2- (in case of disposal in rock salt formation) and the properties of these waters (redox potential) [61].
Uranium (VI) also forms soluble complexes with carbonate anions in natural waters. The aqueous speciation of U(VI) in carbonatecontaining waters at near neutral and basic higher pH values is dominated by a series of strong anionic aqueous carbonate complexes [e.g. UO2CO30 (aq), UO2(CO3)22- and UO2(CO3)34-]. In aqueous complex [UO2(CO3)2]2- is the predominant form of uranium between pH 7 and 8 in an oxidized environment. Numerous investigations of the adsorption of uranium on soils and minerals have shown that carbonate complexing appreciably reduces adsorption of uranium leading to its release from soils [62-65].
Eh-pH diagram for uranium shows the presence of solid phase at low Eh and predominance of dissolved uranium carbonate complexes at high Eh values. When Eh values are above 0.25V and pH between 7 and 8, uranium will be in the oxidized valence state (VI). Also, when Eh values are higher than 0.25V (usually for pH ranging between 1 and 5), uranium is in the valence state (VI), as uranyl ion [UO]2+. In alkaline medium, carbonate is the most significant ligand (in natural water) and the greater solubility of the U(VI) ion is in part due to its tendency to form anionic carbonate complexes [62-65]. The formation of carbonate complexes can change the stability field of U(VI). These U(VI) complexes may exist in alkaline conditions and high carbonate concentrations even in reducing conditions [66-68]. Uranyl hydroxy complexes such as UO2(OH)+ and (UO2)3(OH)5+ are also formed, but generally in smaller amounts except at high temperature or in carbonate-depleted alkaline water.
In reducing water, the U(IV) hydrolysis leads to U(OH)40 [59,60]. The solubility of reduced uranium is low and it has a strong tendency to hydrolyse, forming colloids, especially when environmental conditions change. High concentration of inorganic salts hinders the formation of colloids, while colloids already present may coagulate [67,69].
In addition to dissolved carbonate, uranium can also form stable complexes with other naturally occurring inorganic and organic ligands such as phosphate complexes [UO2HPO40 (aq) and UO2PO4] [70]. Complexes with sulfate [71], fluoride and possibly chloride are potentially important uranyl species where concentrations of these anions are high. However, their stability is considerably less than the carbonate and phosphate complexes [72].
Oxidation state is a fundamental property of U speciation that greatly influences U solubility and, thus, mobility. Whereas U(V1) forms soluble complexes in most surface water and groundwater, U(1V) forms highly insoluble solid phases such as uraninite (UO2(c).
Organic complexes may also be important to uranium aqueous chemistry, thereby increasing their solubility and mobility in soil. The uncomplexed uranyl ion has a greater tendency to form complexes with fulvic and humic acids than many other metals with a +2 valence [73]. In particular, the presence of organic substances and/or colloids in the groundwater increases the complexity of the system. Humic substances formed by the degradation of plants and animals constitute a heterogeneous category of compounds with a complex forming capacity due to the presence of carboxylic, hydroxy and phenolic groups [6]. Dissolved humic substances (humic and fulvic acids) proved to be strong complexing agents for many trace metals in the environment, forming also stable complexes or chelates with radionuclides [74,75]. These substances can be found as dissolved in surface waters as well as in ground waters, in concentrations ranging from less than 1mg (TOC)/L to more than 100 mg (TOC)/L. It has been shown that the binding of metals to humic acid apparently occurs at binding sites with relatively well-defined complex formation constants [75]. Uranium mineral precipitation and co-precipitation processes may also be important during remediation for some environmental conditions, and several uranium (co)precipitates may form, depending on the geochemical conditions [76,77].
The bioreduction and immobilization of soluble U(VI) to insoluble U(IV) minerals is a promising strategy for the remediation of uranium-contaminated soil and groundwater. While a mechanistic description is not fully resolved, it appears humic materials could interrupt electron transport to U(VI). The results of Lenhart et al. [78] suggested that humic materials could potentially decrease U(VI) reduction under certain conditions. Furthermore, humic materials could prevent U(IV) precipitation and thus facilitate the transport of U(IV)-humic complexes.
Solubility processes may also be particularly important for the environmental behavior of U(VI) under oxidizing conditions in those soils that become partially saturated with water or completely dry, when the concentration of uranium in the residue pore fluids may exceed the solubility limits for U(VI)-containing [60].

Uranium Removal

Radionuclides and heavy metals are retained by soil in three ways [79,80]:
• adsorption onto the surface of mineral particles;
• complexation by humic substances in organic particles;
• precipitation reaction.
• As was highlighted above, the mobility of uranium in soil is mainly controlled by complexation and redox reactions [81,82]:
• complexation leads to mobile species or precipitation of U bearing minerals;
• redox reactions change the solubility between the two major oxidation states: U(IV)-U(VI):
• reduction of U(VI) to U(IV) immobilizes uranium;
• oxidation of U(IV) to U(VI) mobilizes uranium because of the dissolution of U(IV) bearing minerals.
These reactions are the basis for certain removal technologies, their combination determining the mobility and fate of uranium. Furthermore, the techniques and methods for uranium removal from soil are selected according to the type of contaminants present, the behavior of the contaminants in the environment and the exposure pathways [83]. For sites with mixed contamination, it is often necessary to use several remediation technologies, sometimes in series, i.e. treatment trains, to effectively address risk from the radioactive, chemical and physical hazards that could be present.

Uranium Contaminated Soil remediation

Various methods for remediation of soils contaminated with radioactive elements are known but only few of them have been applied under large-scale conditions. Each one will direct decisionmakers to substantially different paths with regard to their subsequent choices, actions and potential results, making available significantly different technological options for application.
The use of plants to extract metals and radionuclides from contaminated soil and water has been explored as an economical approach [1]. In order to overcome these problems, one of promising strategies for treating the large scale, low- level contamination is the use of plants to extract metals (Phytoextraction) from soil [23,84-86]. This technique has repeatedly been suggested as a novel clean-up technology. It has the potential to provide cost effective, renewable alternatives to previously used remediation techniques while preventing the loss of topsoil which occurs through the excavation process [87]. As suggested by Robinson et al. [88], a different approach lies in the use of plants which are fast growing, deep-rooted, easily propagated and accumulates the target metal, combined with an increase of the phytoavailability of the metals in soil [89]. Research has focused, therefore, on crops such as maize (Zea mays), tobacco (Nicotiana tabacum), Indian mustard (Brassica juncea), oat (Avena sativa), barley (Hordeum vulgare), pea (Pisum sativa), poplar (Populus spec.) and sunflower (Helianthus annus) [90,91]. Unfortunately, the main disadvantage of the phytoremediation techniques is the long time required for cleanup of metal contaminated soils [92,93]. Many plant species have been screened to determine their usefulness for U and heavy metals phytoextraction. Researchers initially applied hyperaccumulators to clean metal polluted soils [94]. Sunflower (Helianthus annuus L.) and Indian mustard (Brassica juncea Czern.) are the most promising terrestrial candidates for metal removal in water. The roots of Indian mustard are effective in the removal of U [49] from hydroponic solutions. Steubing et al. [95] studied the potential of using higher plants as indicators of uranium distribution in soil at a site in Germany where uranium concentrations ranged from 5-1500 mug/g soil and reached a maximum of 1860 mug/kg in soil water. Results indicated that Sambucus nigra was the best indicator of uranium contamination. Chemical analysis of its leaves provided more detailed information regarding uranium distribution than soil analyses. The plants not only indicate the location of mineralization but also the migration pathway of U-containing soil water. The adsorption of contaminated water is the main source of the U-accumulation in the different plant organs.
Petrescu et al. [92] studied the bioaccumulation of U and Th in the vegetation; they found that Uptake of U is greater than that Th. One reason for this is that the elements similar to nutrients (such as U and Sr being similar with Ca) follow the same path as the nutrients they are similar to.
Shahandeh and Hossner [50] investigated thirty four plant species and its uptake capacity for uranium (U) accumulation from U contaminated soil. There was a significant difference in U accumulation among plant species. Sunflower (Helianthus annuus) and Indian mustard (Brassica juncea) accumulated more U than other plant species. Sunflower and Indian mustard were selected as potential U accumulators for further study in one U mine tailing soil and eight cultivated soils (pH range 4.7 to 8.1) contaminated with different rates (100 to 600 mg U(VI) kg-1) of uranyl nitrate (UO2(NO3)2.6H2O). Uranium accumulated mainly in the roots of plant species. The highest concentration of U was 102 mg U kg-1 in plant shoots and 6200 mg U kg-1 in plant roots. Plant performance was affected by U contamination rates, especially in calcareous soils. Plants grown in soils with high carbonate-U fractions accumulated the most U in shoots and roots. The lowest plant U occurred in clayey acid soils with high Fe, Mn and organic U-fractions.
Ebbs et al. [1] reported that the addition of citric acid and its salts selectively increase uranium mobility in soil and subsequently also plant uptake. The authors suggest that the strong mobilization of U by citric acid is due to the formation of citrate-uranyl complexes rather than to the decreased pH. They found a close correlation between the U and the Fe and Al concentrations in the soil solution after the addition of citric acid, which they explained by the dissolution of Fe and Al sesqui-oxides and hence release of U from soil material to the soil solution.
Adding 10.5 mg citric acid kg-1 to a potted soil that contained 310 mg U kg-1 enhanced uranium solubility 73-fold to 110 mg U kg-1, whereas adding HEDTA at the concentration of 5 g kg-1 had no significant effect [1]. The authors chose the different concentrations of these agents from positive results of other studies, which were made with U. In the same study, 10.5 mg citric acid kg-1 increased the U concentration in the dry matter of Swiss chard (Beta vulgaris L. subsp. vulgaris) 14-fold from 15 to 200 mg kg-1, whereas no effect was observed after the HEDTA application. Ebbs et al. [1] observed optimum U solubilization with citrate over the pH range from 4 to 5, with soluble U concentrations considerably higher than those at pH 6 or 6.8. In this study, at pH 5, adding 0.61 g kg-1 of potassium citrate increased the soluble U concentration 93-fold.
Author unpublished data indicated that sunflower and cotton shoots accumulated the highest U content among the five tested plant species, irrespective of soil type. Shoot concentrations of U were as high as 69.9 Bq kg-1 dry matter of sunflower, followed by cotton and napier grass, panikum then squash with a range of U between 4.2 to 69.9 Bq kg-1 dry matter in case of the alluvium soil. However in the sandy soil, sunflower U -shoots were > cotton > penakium > napier grass > Squash with a lower order of magnitude which could be explained by the lower U -content in sandy soil compared to the alluvial soil. In another pot experiment, the effect of soil amendment application on plant uptake was evaluated as practical ways to solubilize, detoxify, and enhance U accumulation by plants. Sunflower (Helianthus annuus) was selected as a potential U-accumulator in two selected soils different in texture and contamination levels. To enhance metals phyto-extraction, ammonium nitrate and organic chelators (EDTA, and citric acid) were added to soils at rates of 0, 5, 10 and 20 mmol kg-1. Experiment was run for 8 weeks growth period. The application of soil chemical treatments significantly increased U uptake in sunflower shoots and roots. Also it was clear that, increasing the rate of application up to 20 mmol kg-1 significantly enhanced U uptake in sunflower shoots and roots. It could be noticed that U accumulated in roots more than in shoots. Citric acid had enhanced U accumulation (in shoots and roots) more than EDTA and Amnitrate treatment. It could be also concluded that, the effectiveness of U remediation of soils by plants was strongly influenced by soil type. Soil properties determined the tolerance and accumulation of U in plants. The concentrations of soluble U in the soil could be enhanced to attain high U removal rates by increasing the metal accumulation of plants. This could be achieved by adding certain chelating agents to the soil. However, enhanced chelating agents may cause unavoidable leaching of chelated metals down the soil profile which could lead to rapid leaching of these toxic metals to groundwater.

Results

In General the results of these experiments confirm that both the Plant biomass production and metals accumulation were varied with contaminants content and species, chelator form as well as rate of application, and soil properties. The highest metals accumulation was found in plants growing on clayey soil and the lowest was in plant growing on sandy soils. Metals accumulation and translocation to the shoots were significantly increased as application of citric acids. Addition of citric acid at 20 mmol kg-1 soil to clayey soil led to increasing U concentration in shoots several-fold of magnitude on the other hand, adding ammonium nitrate had a little effect on metals translocation to shoots. Citric acid was the most effective chelating agent in plant accumulation for heavy metals.
Fresquez et al. [93] studied the uptake of radionuclides by beans, squash, and corn growing in contaminated alluvial soils at Los Alamos National Laboratory. Pinto beans (Phaseolus vulgaris), sweetcorn, and zucchini squash (Cucurbita pepo) were grown in a pot study using alluvial soils contaminated with various radionuclides. Soils as well as washed edible (fruit) and non-edible (stems and leaves) crop tissues were analyzed for tritium (3H), 137Cs, 90Sr, 238Pu, 239Pu, 240Pu, 241Am, and total uranium (totU). Most radionuclides, with the exception of 3H and totU, in soil and crop tissues were detected in significantly higher concentrations (P<0.05) than in soil or crop tissues collected from regional background locations. Significant differences in radionuclide concentrations among crop species (squash were generally higher than beans or sweet corn) and plant parts (non edible tissues were generally higher than edible tissue) were observed. Most soil/plant concentration ratios for radionuclides in edible and non-edible crop tissues grown in the studied soils were within default values in the literature commonly used in dose and risk assessment models. Overall, the maximum net positive committed effective dose equivalent of beans, sweet corn, and squash in equal proportions was 74 mrem/year (740 μS/year). This upper bound dose was below the International Commission on Radiological Protection permissible dose limit of 100 mrem/year (1000 μS/year) from all pathways and corresponds to a risk of an excess cancer fatality of 3.7 X10-5 (37 in a million), below the US Environmental Protection Agency’s guideline of 10-4.
In a comprehensive literature review of Characterization and remediation of soils contaminated with uranium studies, Gavrilescu et al. [86] concluded that natural attenuation of U involves enclosing the radionuclides in a mineral where they will not escape unless chemical conditions change dramatically. Chemical approaches are available for metal and radionuclide remediation, but are often expensive to apply and lack the specificity required to treat target metals against a background of competing ions. Electro-kinetics became more and more applied for the remediation of contaminated soils and/or groundwater. This technique has been called electrokinetic remediation, electro-remediation, electro-restoration, electroreclamation, electro-chemical decontamination or electro-migration.
Microbial processes are beginning to be used in the cleanup of radioactive and metallic contaminants of soils and sediments through biotransformation, biodegradation and bio-mineralization. Biological approach offer the potential for the highly selective removal of toxic metals coupled with considerable operational flexibility; they can be used both in situ and ex situ. Recently published studies have demonstrated direct microbial reduction of U(V1) to U(IV), and it has been suggested that this process could be potentially utilized as an in situ biological remediation strategy. Microorganisms can degrade soluble organo-uranium compounds in soil and rocks using ligands as a source of carbon and energy which enhances uranium precipitation and deposition. Fungi (Aspergillus ochraceus and Penicillium funiculosum) were able to take up large amounts of soluble uranium in their mycelium from rocks [96]. Microorganisms appear to act as sinks for uranium, which is accumulated and concentrated to high levels in cell walls [97]. Often aquatic fungal cultures tested, five (Alternaria tenulis, Chaetomium distortium, Fusarium sp. Saccharomyces cerevisiae and Trichoderma horzianum) were capable of biosorbing more than 90% of uranium present in aqueous solution at an initial concentration of 150 mg U/L [98]. Biosorption of the uranium was highly dependent on solution pH, with optimal uptake of the uranyl acetate occurring at pH 5.5. [98]. Streptomyces sp. either living or dead, are able to accumulate UO22+ ions, which binds to cell wall sites as well as to cytoplasmic structures within the cells of the bacteria. The uptake of uranium was examined for indigenous bacterial species (Bacillus) occurring in a uranium waste pile in Saxony, Germany [99]. The study demonstrated that bacteria species B. spaericus, B. cereus, and B. megaterium and their spores selectively accumulated uranium from contaminated waters. In vegetative cells, sorption of approximately 90% of the uranium (present in the water at 72.1 μg U/L) was observed, while spores tended to show slightly lower sorption with more variability across different Bacillus strains [99].

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