Solidification of 232Th-Contaminated Soils

C H A P T E R

12 Reclamation of Sites Impacted by Mining Activities: Stabilization/ Solidification of 232Th-Contaminated Soils Pietro P. Falciglia*,†, Stefano Romano†,‡, Federico G.A. Vagliasindi* *

University of Catania, Catania, Italy †Laboratori Nazionali del Sud—Istituto Nazionale di Fisica Nucleare, Catania, Italy ‡University of Catania, Catania, Italy

12.1 INTRODUCTION The rapid development of nuclear activities (industry, energy) has produced, in the last few decades, a drastic increase in the demand for uranium (U) mining metallurgy products. However, extraction of U and ore in milling facilities has produced a large amount of radioactive wastes, which also contain a series of long-lived radionuclides such as radium (Ra) and thorium (Th) isotopes, defined as TENORM (Technologically Enhanced Naturally Occurring Radioactive Materials). Their indiscriminate and improper deposition have resulted in heavy soil contaminations (Yan and Luo, 2015). Investigations on the monitoring of TENORMcontaminated sites have recently received particular attention worldwide; however, very limited remedial techniques are present in the literature for their restoration. The cement-based stabilization/solidification (S/S) technique has been shown to be a perfect candidate for this purpose, also due to the possibility of employing high-density binder/ materials with γ-radiation shielding properties. The aim of this chapter is to review the information available for the treatment of ­thorium-contaminated soils by means of the S/S technique, reporting the theoretical background, including the S/S process and recent related scientific and technoeconomic development for the S/S application in reclamation activities of 232Th-contaminated soils from mining activities.

Assessment, Restoration and Reclamation of Mining Influenced Soils http://dx.doi.org/10.1016/B978-0-12-809588-1.00012-8

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© 2017 Elsevier Inc. All rights reserved.

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12.  RECLAMATION OF SITES IMPACTED BY MINING ACTIVITIES

12.2  TENORM-CONTAMINATED SITES 12.2.1 Long-Lived γ-Ray Emitting Radionuclides Radioactivity is the process through which nuclei spontaneously emit subatomic particles. Three main types of radiation are emitted during radioactive decay processes (Ahmed, 2015): • α-rays (helium nuclei with two protons and two neutrons); • β-rays (electrons); • γ-rays (photons). When nuclei emit subatomic particles, their configuration, state, and even identity may change. For example, when a nucleus emits an alpha particle, the new nucleus has two protons and two neutrons less than the original one. Except for γ-decay, in which the nucleus retains its identity, all other decays transform the nucleus into a very different one. Nuclei able to emit radioactive particles are called radionuclides. Uranium (U) and thorium (Th) are radionuclides generally present in the earth’s crust at low mass concentrations of about ten parts per million; however, they can be concentrated by human activities in mineral deposits where their concentrations can significantly increase up to tens of percents (Arogunjo et al., 2009). Both U and Th are defined as Natural Occurring Radioactive Materials (NORM) and the parent radionuclides 238U, 235U, and 232Th are primordial with half-lives in the range 4500–14,000 million years (Metcalf, 1996). It is estimated that 87% of the radiation to which human targets are exposed is generated by natural sources and the remaining is due to anthropogenic radiation (Olise et al., 2010, 2014). Mining and processing of bearing ores for the extraction of U and other minerals, as well as ore crushing and milling processes, result in the production of several contaminated wastes (rock and tailings) generally contaminated with low specific activity and long radioactive half-life radionuclides. Radioactive contaminants present in these wastes are considered to be TENORM, due to their high concentration in the byproducts (Arogunjo et al., 2009).Their indiscriminate or improper disposal increases the risk of radionuclide transportation and contamination. TENORM are generally disposed of in unsuitable landfills without appropriate isolating systems and this results in contaminant leaching processes and, thus, in soil and groundwater pollution (Yan and Luo, 2015). Long-lived emitting radionuclide contaminations have a long-term radiological impact, increasing risks connected to their exposure and representing a major concern for the environment (Fuma et al., 2015). Among long-lived radionuclide contaminants, isotopes able to generate radiation in the form of gamma (γ) rays have high radiotoxicity, resulting in major environmental concerns (Falciglia et al., 2012; Lukšiene et al., 2012). Gamma (γ) radiation (γ ray) is an electromagnetic radiation of high frequency and energy generated by the decay from high-energy states of atomic nuclei (gamma decay) after alpha (α) or beta (β) particle emission. Both α and β particles have an electric charge and mass, and thus are quite likely to interact with other atoms in their path. On the other hand, γ rays are composed of photons, which do have mass and electric charge, and penetrate much further through matter such as biological tissue than either α or β radiation, thus being more biologically hazardous (Fig. 12.1). Contrary to radioactive contamination caused by extreme events such as Fukushima (Japan, 2011) or Chernobyl (Ukraine, 1986) nuclear power plant (NPP) disasters, where



12.2  TENORM-Contaminated Sites

331

FIG.  12.1  Penetration effects of different radioactive radiations.  Source: Thorium SVG image by Wikipedia contributor BatesIsBack.

c­ ontamination has the potential to affect a geographical area much greater than the NPP site and its immediate surroundings, TENORM-contaminated sites are often not perceived as very hazardous for human health. This is also caused by a lack of understanding within the wider community about radiological matters and TENORM in particular (Booth, 2015). Based on these considerations, it is clear that, in the case of soils contaminated with longlived gamma (γ) emitting TENORM, landfill disposal represents an unsafe solution, making the development of alternative, cost-effective solutions a key factor in the territory management and restoration strategies (Nakano and Yong, 2013).

12.2.2  232Thorium as Soil Contaminant Among TENORM, thorium is prevalently (99%) present in the form of thorium-232 (232Th) as thorite and thorianite, whereas after mining activities, 232Th is usually concentrated as thorium dioxide (ThO2) (Table 12.1). After most of the thorium is removed from the rocks, these are called "depleted" ore or tailings. According to the Environmental Protection Agency (EPA),232Th is the primary source of contamination at the Superfund sites due to extraction activities. 232Th has also been used to make ceramics, gas lantern mantles, and metals used in the aerospace industry and in nuclear reactions or used as a fuel for generating nuclear energy (ATSDR, 2015; Shtangeeva et al., 2005). These activities represent another minor source of 232Th-soil contamination. 232Th decays through a 10-step chain to 208Pb emitting α and β particles as well as γ-rays and, as a soil pollutant, presents both a chemical and radiological hazard (Fig. 12.2). It has a half-life of about 200 years and could be permanently accumulated in the human skeletal and central nervous system, leading to an irreversible health problem

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TABLE 12.1  Thorium (Th) Chemical-Physical Properties Parameter

Value

Group

Actinides

Atomic number

90 −1

Atomic mass (g mol ) −3

232.04

Density at 20°C (g cm )

11.72

Melting point (°C)

1750

Boiling point (°C)

4790

Key isotopes

230

232

1.405E + 10

Th half-life (y)

Th, 232Th

(Falciglia et al., 2013). 232-Thorium is a relatively reactive, metallic radioactive element. In the case of radionuclide-contamination, the evaluation of adverse health effects due to radionuclide exposure requires a slightly different approach than with other chemical contaminants. Radiation is a health risk because radioactive elements can emit energetic particles or electromagnetic radiation that can damage cells. Studies on thorium workers have shown that breathing thorium dust may cause an increased chance of developing lung disease and cancer of the lung or pancreas many years after being exposed. Changes in the genetic material of body cells have also been shown to occur in workers who breathed thorium dust. The presence of large amounts of thorium in the environment could result in exposure to more hazardous radioactive decay products of thorium, such as radium and thoron, which is an isotope of radon (ATSDR, 2015).

12.2.3  Remedial Alternatives for TENORM- and 232Th-Contaminated Soils In the last few decades, scientists, engineers, and regulators have devoted a great deal of effort to the development of innovative technologies aimed at reducing the impact of industrial activities on the environment and decontaminating the polluted areas (USEPA, 2006). Soils have an important capacity for the retention of contaminants, which avoids the spread of the contaminants, but, once the retention capacity of the soil is exceeded, radionuclides are released into the groundwater. Due to the dangers of radioactivity, the decontamination of soils and groundwater with radionuclides is a priority (Cameselle, 2015). Recently, a lot of effort has also been devoted to research and development on innovative technologies for the remediation of impacted soil and groundwater (USEPA, 2006). The main interest has shifted from ex situ remediation to in situ remediation technologies and several techniques were proposed for radionuclide-contaminated soils (Cameselle, 2015). However, despite the need for effective and economically competitive strategies for radionuclide-contaminated soils, aimed at protecting local populations against exposure to high radioactivity dose rates (Nakano and Yong, 2013), very limited techniques have been reported for this purpose. Soil flushing, bioremediation, and electrokinetic remediation are recent technologies tested for soil restoration; however, despite good



12.2  TENORM-Contaminated Sites

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FIG. 12.2  Thorium-232 decay series (Font: Argonne National Laboratory, EVS—Human Health Fact Sheet, 2005).

results obtained at bench-scale tests, the performance of the remediating technologies at field scale had poor removal results, long treatment times, and high costs. This reveals the difficulties of the implementation of the techniques to the field and the importance of the nature of the pollutants, the geochemistry of soils, and geotechnical properties of the soils (grainsize, permeability, etc.) in the success of the remediation technology. Considering the physical-chemical characteristics of radionuclides, only a few of these cited technologies can be applied for their successful removal in situ from soils, and often their combined application could be useful (Elless and Lee, 2002).

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Much lower effective and economical alternatives are available for 232Th-contaminated soils (Nakano and Yong, 2013). In this case, literature essentially provides two main groups of remedial treatments, both based on a biological approach: bioremediation and phytoremediation techniques. The biosorption of thorium by bacteria, actinomycetes and fungi have been screened in the last decade by several scientists (Desouky et al., 2011; Picardo et  al., 2009). These studies clearly showed that thorium-biosorption/accumulation could be due to the combined result of ion-exchange-complexation-microprecipitation-adsorption mechanisms and that microbial activities can be limited by several factors such as the pH of soil (Kazy et al., 2009). Some Th-accumulator plants were also screened in labscale phytoextraction treatments (Guo et al., 2010; Shtangeeva et al., 2005); however, during their application, Th-bioavailability was shown to be affected by several factors, such as the physiological characteristics of plant growth stages, the amount of acid agent added and the presence of microorganisms in the plant rhizosphere on thorium. Therefore, biobased remedial techniques may be ineffective or too lengthy, especially in the presence of high Th-concentrations in soils. Furthermore, contaminant removal-based treatments generally produce highly radioactive wastes or wastewater needing landfill disposal or a further treatment. In order to minimize the total amount of radioactive waste, immobilization of radioactive contaminants in impacted soils could represent a suitable cost-effective technique (Mallampati et al., 2015). Based on these considerations, it is clear that the possibility of reaching both the objectives to immobilize the radionuclides within a solid matrix and, at the same time, to shield the γ radiation emission, is a preferable choice. Due to its advantages and the opportunity to employ high-density shielding materials, stabilization/solidification (S/S) was shown to be an optimal solution in 232Th-contaminated soil treatments.

12.3  STABILIZATION/SOLIDIFICATION THEORETICAL BACKGROUND 12.3.1  Cement-Based Stabilization and Solidification Processes Stabilization/solidification (S/S) is a treatment process by which contaminated soils, sediments or waste materials are mixed with binders and specific additives with the aims of reducing the mobility of the contaminants by increasing the pH and fully or partially binding the contaminants in the solid matrix (stabilization), and of improving the physical properties (strength, compressibility, permeability and durability) of the final treatment products (solidification) (Falciglia et al., 2013) (Fig. 12.3). Specifically, the S/S technique is widely employed to deliver physical and dimensional stability, as well as better handling, and to produce more chemically stable waste constituents. Portland cement (PC) is reported as the most commonly used binder. However, low-cost byproducts can also be employed as cobinders to reduce the costs and increase the environmental sustainability of the treatment (lime, pulverized fuel ash PFA and clays) (Wang et al., 2015a,b). PC is a heterogeneous mixture of five mineral phases: 50%–70% alite (C3S), 20%–30% belite (C2S), 5%–12% alluminate (C3A), 5%–12% ferrite (C4AF), and 2% gypsum. The hydration of PC is a sequence of overlapping chemical reactions between dry binder compounds



12.3  Stabilization/Solidification Theoretical Background

PC based grout

Soil

335

Additive (adsorbent material)

Contaminant

FIG. 12.3  Schematic representation of stabilization and solidification processes.

and water, leading to continuous cement paste stiffening and hardening. The early behavior of hydrating PC is governed by reactions of the aluminate phases, while the setting and the early strength development behavior is mostly dependent on the hydration of silicates, particularly alite. The formation of hydration products and the development of microstructural features depend on solution processes and interfacial and solid-state reactions. The hydration products of PC are mainly made up of 20%–25% Ca(OH)2 (CH), 60%–70% calcium silicate hydrate gel (CSH) and 5%–15% other phases, including grains of still-unhydrated cement (LaGrega et al., 1994). In metal (stable isotopes or radionuclides) polluted soils stabilized/solidified with PC, three possible mechanisms may be responsible for the immobilization of the contaminant during the hydration phases. One mechanism may be precipitation resulting from the formation of silicate oxides. Another may be inclusion, either by physical encapsulation and/or by chemical inclusion. Physical encapsulation can be achieved by creating a solidified monolith, while chemical inclusion can be achieved through the incorporation of contaminant in binder hydration products, such as CSH gel phases, which play such an important role in the retention of metal. The main mechanisms that determine metal immobilization in the solid matrix are: sorption on clay and pozzolanic reaction products (Dermatas and Meng, 2003; Moon and Dermatas, 2007) and addition or substitution reactions with CSH (Falciglia et al., 2014a). Several studies on S/S treatment have been performed to better understand the ­fundamentals of heavy metal immobilization and leaching or to investigate the performance of PC-based S/S treatments of soils polluted by heavy metals, such as CrVI (Wang and Vipulanandan, 2000), Cr III (Jing et al., 2006), PbII (Lin et al., 1996; Moon and Dermatas, 2007), As (Moon et  al., 2008) or CdII (Shawabkeh, 2005). Authors have found that, in PCbased treatments, the percentage of binder used, curing age and metal contamination level significantly influenced the setting time, compressive strength, leachability and chemical and crystalline structure of the treated soils.

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12.3.2  Gamma (γ) Radiation Shielding Phenomena: High-Density Materials and Photon Interaction With Matter In the processes of γ-ray (photon) interaction with matter, the γ-ray attenuation through a given material or mixture of materials is a function of γ-ray energy, the elemental composition, the density and the thickness of the material. The γ-ray intensity attenuation is a nonlinear function of the thickness (x) of the material, described by the Lambert-Beer law (Knoll, 2000): I ( x ) = I0 e − µx

(12.1)

where I(x) is the transmitted intensity, I0 is the incident intensity and μ (cm−1) is the linear attenuation coefficient of the material; μ is a function of the γ emission energy (keV) and depends on the density of the material, ρ (g cm−3), and is often expressed as a function of the mass attenuation coefficient, μm (cm2 g−1):

µ = µm ρ

(12.2)

It is therefore clear, that, in S/S treatments, the possibility of employing high-density material such as PC or barite aggregates represents a key factor in terms of radioprotection (Akkurt et al., 2010a,b). In Falciglia et al., 2015 it is reported that the shielding effects of PC-based S/S treatments involved the following processes regarding γ-ray interaction with matter: photoelectric effect, Compton scattering and pair production. They are statistical processes and their relative contribution depends on the γ-ray energy and on the atomic structure of the medium. Thus, the greater the interaction the more likely are the shielding effects. However, if the maximum photon emission energy is about 1 MeV, only the first two effects must be considered in contributing to the γ-ray absorption, these being the pair production, an effect with a minimum energy of about 1 MeV, corresponding to the rest mass energy of the pair positron-electron. Moreover, in these physical conditions (energy and atomic structure of the medium) the photoelectric effect should be dominant, giving a complete absorption of the γ-ray interacting. The photoelectric effect occurs when matter emits electrons upon exposure to electromagnetic radiation, such as photons. When a material is exposed to a sufficiently energetic electromagnetic wave, photons will be absorbed and electrons will be emitted. The threshold frequency is different for different materials. The photoelectric effect occurs with photons having energies from a few electronvolts to over 1 MeV. At the high photon energies comparable to the electron rest energy of 511 keV, Compton scattering may occur. This is a type of scattering that gamma rays undergo in matter. The inelastic scattering of photons in matter results in a decrease in energy of photons, called the Compton effect. Part of the energy of the gamma ray is transferred to a scattering electron, which recoils and is ejected from its atom (which becomes ionized), and the rest of the energy is taken by the scattered, "degraded" photon. The increasing of γ-ray shielding properties can be achieved by means of the increase in the density of the S/S mix, obtained by replacing PC with higher-density material such as barite or magnetite mineral (Falciglia et al., 2014b).



337

12.4  TESTING PROCEDURES AND γRS PERFORMANCES OF S/S FINAL PRODUCTS

12.4  TESTING PROCEDURES AND Γ-RADIATION SHIELDING PERFORMANCES OF S/S FINAL PRODUCTS 12.4.1  Conventional Testing Protocol: Chemical and Physical Tests To verify the effectiveness of the S/S treatment, it is necessary to assess the characteristics of the treatment final products and compare them with specific performance criteria. Other particular aims of testing protocols may include (Bone et al., 2004): • to establish correlation between results from proposed compliance tests with characterization tests; • to compare the costs of the different mix designs in terms of raw materials; • to assess the long-term performance of the treated material. This is achieved by carrying out several tests, the results of which may be compared against performance criteria. Physical tests are used to predict mixing behavior, reagent needs and volume increases, and compare treated and untreated materials in terms of their strength and durability. Chemical tests are also used to determine the leaching behavior of the S/S material. They could be categorized as (Perera et al., 2004): • basic information tests, which measure basic material properties (e.g., grading, plasticity, particle density, total contaminant concentration); • performance tests, which relate to the properties of the material in use (e.g., strength, leachability). These categories include physical and chemical (predominantly leaching) tests, and may be used for understanding mechanisms, assessing compliance with reference criteria. Testing protocol on S/S treated soils generally includes the following relevant bench scale tests: leaching, unconfined compressive strength (UCS), hydraulic conductivity and durability test. Other tests are bulk density, specific gravity, water absorption, moisture content, setting time, heat of hydration and microstructural examination. Leaching tests are conducted to examine mass transfer from a solid (the S/S material) to a liquid (termed the “leachant” before contact with the solid, and the “leachate” afterwards). Their contact time depends on the characteristics of the contaminated material and the surrounding environment, because S/S matrices’ leachability is dependent on the physical and chemical properties of the contaminated material and the leachant (LaGrega et al., 1994). Extraction can be performed by means of different official procedures. For example, the Toxicity Characteristic Leaching Procedure (TCLP) is a commonly used standard single batch leaching test, which was developed by the United States Environment Protection Agency (USEPA) as a rapid regulatory compliance test for determining whether a waste is suitable for disposal in a landfill with municipal waste. In the case of radioactive contaminants, leaching measurements can be performed according to the International Atomic Energy Agency (IAEA) testing standard for solidified radioactive waste (IAEA, 2004). With this method, which has been applied by Falciglia et al. (2015) and Shaaban and Assi (2011), the leaching rate (LR) of S/S radionuclide-contaminated soil samples can be calculated as follows: LR =

∑A i

A0

i

⋅

M 1 ⋅ S ∑ i ti

(12.3)

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12.  RECLAMATION OF SITES IMPACTED BY MINING ACTIVITIES

where Ai is the radioactivity of the leached constituent during each leaching interval ti (Bq); A0 is the specific radioactivity initially present in the specimen at zero time (Bq); S is the exposed surface area of the specimen (cm2); M is the mass of the specimen (g); ti is the duration of the leaching period (day). In order to assess the leaching behavior of the contaminant before and after the S/S treatment, the ANSI/ANS-16.1 (1986) or ASTM C 1220-92 leachability test Standard Test Method for Static Leaching of Monolithic Waste Forms for Disposal of Radioactive Waste can also be considered (Marrero et al., 2004). The UCS test relates to the mechanical resistance of monolithic S/S products and, specifically, to the progress of hydration reactions in the product and durability of a monolithic S/S material, and is therefore a key variable. It is one of the most commonly used tests and there are numerous standard methods for its determination, all of which involve vertical loading of a monolithic specimen to failure (ASTM C109/C109M-99, 2001; ASTM D1633-00, 2002; BS 1881: Part 116, 1983; BS EN 12390: Part 3). Hydraulic conductivity indicates the rate at which water can flow through a material, which is a key variable for environmental behavior. The main methods for determination of hydraulic conductivity are reported in ASTM D5084-00 (2002) and BS 1377: Parts 5(5) and 6(6) (1990). Finally, durability test methods are applied to assess the long-term performance of the S/S products and in particular the resistance of the material to repeated cycles of weathering. Cured test specimens were subjected to 12 wet/dry (W/D) and freeze/thaw (F/T) cycles according to ASTM D4843-88 (2009) and ASTM D484290 (2001) methods, respectively (Perera et al., 2004).

12.4.2  Assessment of γ-Radiation Shielding Performances as γRS Index For a configuration where the radioactive source and shielding elements are different units, the linear attenuation coefficient (μ) can be simply adopted in order to assess the shielding feature of specific elements (i.e., compounds and mixtures, building materials or concretes) (par. 3.2). However, if radiation sources and shielding elements are mixed together, as in the case of S/S radionuclide-polluted soils, the shielding effects of each material are very difficult to assess (Fig. 12.4). In order to overcome this limitation, an experimental index, defined as gamma radiation shielding (γRS) of S/S monolithic samples, has recently been proposed. It was

reduced g-ray emission

g-ray emission

Radionuclide contaminated soil

S/S radionuclide contaminated soil

FIG. 12.4  γ-ray shielding effects of S/S treatment of a radionuclide-polluted soil.



339

12.5  Site-Specific S/S Technologies and Applications

demonstrated to be suitable to assess the effectiveness of the treatment in terms of the limitation of the radioactivity effects (Falciglia et al., 2015). γRS is defined in Eq. (12.4):

γ RS =

I Soil − I SS ⋅ 100 ( % ) I Soil

(12.4)

where ISoil is the γ-ray counting rate of the contaminated soil sample (CPS) measured before the treatment and ISS is the one corresponding to the S/S treated soil sample. The γ-ray counting rate (CPS) is generally measured using a gamma ray spectrometer, for example, with a high purity germanium (HPGe) detector. An HPGe detector provides an analog signal for each absorbed gamma ray, which is proportional to the energy of the impinging gamma ray. An analog-to-­ digital conversion allows one to record the energy value for each pulse and add up all the pulses onto a histogram (energy spectrum). Each energy value identifies the radioactive isotope by which the γ ray has been generated, while the number of counts for each isotope is proportional to the abundance of the radioactive source. Considering a theoretical scenario connected to a real site impacted by a radionuclide contamination and the potential S/S application, the gamma-­ radiation emission values of soil after S/S assessable by γRS can be used in order to calculate the absorbed radioactivity dose. Absorbed dose indicates the amount of radiation absorbed by an object or person (that is, the amount of energy that radioactive sources deposit in the materials they pass through) and is fundamental in order to assess the effectiveness of a selected S/S treatment in terms of shielding features. In fact, absorbed dose values, assessed by γRS and compared to regulatory radiometric limits make the full-scale applicability of the intervention official.

12.5  SITE-SPECIFIC S/S TECHNOLOGIES AND APPLICATIONS Based on the experimental investigations and the chemical and physical material properties collected from testing protocol and on the disposal scenario, a number of mix designs can be proposed for evaluation. In some situations, the nature of the contaminants may be very variable, and it may be preferable to consider using more than one mix design to treat different sites, either on a technical and/or economical basis. Main factors to consider when selecting a binder can include (Bone et al., 2004): • any possible interference effects of contaminants on the performance of the binder; • the moisture content of the material and whether it will affect the binder performance; • the workability of the binder-soil-water mixture having regard to the method of application and/or compaction; • the ability to permanently bind contaminants; • factors that could affect potential impacts on the surrounding environment. It is also recommended that the manufacturer and supply of binder remain constant from treatability studies through to full-scale treatment, to avoid the potential for inconsistent performance due to minor variations in binder properties. The selection of the process and equipment strictly depends on technical and economic considerations. The key factors taken into account when identifying suitable process and plant are (Bone et al., 2004): • material characteristics and ground conditions; • quantity of the contaminated matrix to be treated and rate of treatment;

340

12.  RECLAMATION OF SITES IMPACTED BY MINING ACTIVITIES

• extension of treatment area; • dust generation of plant with respect to environmental and health and safety aspects. Environmental monitoring activities need to be programmed into the works to ensure that remedial activities do not give rise to uncontrolled emissions to air, land or water. Monitoring may also be required to demonstrate that S/S remains effective for a predefined time scale depending on the disposal or reuse scenario (Bone et al., 2004). Due to its versatility, efficiency, time and costs, S/S has been widely used to immobilize and dispose of hazardous (Chen et al., 2009; Shi and Spence, 2004) or low- and i­ ntermediate-level radioactive wastes (Ojovan et al., 2011), landfill leachate (Hunce et al., 2012), sludge (Ucaroglu and Talinli, 2012), and sediments (Careghini et  al., 2010), as well as to remedy metal-­ contaminated soils in situ (Chen et al., 2009; Harbottle et al., 2007) or petroleum drill cuttings (Leonard and Stegemann, 2010). Basic cementation treatment of radioactive waste has been practiced for many years to immobilize contaminants, since cemented matrices are characterized by good compressive strength, thermal, chemical and physical stability. Moreover, the alkaline chemistry of hydraulic cements ensures low solubility for most radioactive waste radionuclides. The prominent advantages of cement immobilization are due to the low expense, availability and durability over time of hydraulic cements (Ojovan et al., 2011).

12.5.1  In Situ Technologies In situ treatment involves the processing of contaminated soil in its existing position, by direct addition of the binders/materials. In mechanical mixing, soil mix is carried out using mixing augers through which a grout is introduced and mixed with the contaminated soil, resulting in overlapping S/S soil-grout columns (Plante et al., 2008) (Fig. 12.5). More details of typical in situ mobile treatment units, mixing augers and potential scenarios of soil-mixing applications are reported in Al-Tabbaa and Perera (2006). In hydraulic mixing, liquid binders are added by means of a typical jet grouting system (Fig. 12.6).

Grout from plant Crane

Tube Turnable Auger

(A)

(B)

FIG. 12.5  Schematic of in situ S/S process with mechanical auger (A) and overlapping columns of S/S contaminated soil (B).  Modified from Plante, T., Gustafson, A., Guay, M., Corradino, K., 2008. Equipment and scale-up consideration for in situ solidification of MGP sites. In: Gasworks Europe Redevelopment, Site management and Contaminant Issues of Former MGP’s and Other Tar Oil Polluted Sites. Proceedings of MGP 2008 conference in Dresden, Germany, March 4–6.



12.5  Site-Specific S/S Technologies and Applications

341

FIG. 12.6  Schematic of in situ S/S process with jet grouting injection.

In situ methods can be of particular benefit on sites that have a restricted space to store materials or where it is preferable to avoid excavation, reducing risks to site workers. They also reduce the volume of spoil that requires handling, generally have low levels of noise production, and the ability to treat contamination close to existing structures without the need to excavate or control groundwater. An in situ mixing plant also has a high speed of implementation, facilitation of rapid redevelopment of the site and reduction of off-site disposal (Harbottle et al., 2007). Some issues that may limit the use of in situ treatment are the presence of cohesive soils, debris or rocks which may reduce auger penetration rate and depth of operation, the potential increase in bulk volume of treated soils, especially in the cases requiring an increase in the quantity of additives used, and, especially, operating in very low or high ambient temperatures may cause freezing of the feed slurry before injection or a delay in the set time of the treated matrices. The need to ensure the perfect alignment of columns and provide adequate overlapping to avoid an untreated soil region represents another important issue which must be taken into account (Bone et al., 2004).

12.5.2  Ex Situ Technologies Ex situ treatment involves processing of contaminated soil after it has been excavated and moved from its original location. In a typical ex situ processing, after a pretreatment (homogenization, debris removal), the soil is mixed with the appropriate binder/grout and other additives if necessary and then placed at its final disposal site. The mixing is generally carried out with mechanical mixers using either batch or continuous processes (Fig. 12.7). In a batch process the required amount of contaminated soils and binder(s)/material(s) are added and blended for a fixed amount of time. In a continuous process, they are added and blended continuously. The variation of contact time is achievable by controlling the feed and mixing equipment. More details of a typical ex situ S/S system are reported in Al-Tabbaa and Perera (2006). The final disposal options are on-site and off-site. A main advantage of ex situ treatment is the ability to improve the homogeneity of the contaminated soil (in terms of both chemical-physical features and contaminant level) to be treated. This improves the process control and the possibility of better achieving remedial objectives in terms of final properties of the S/S treated materials. The possibility of performing

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12.  RECLAMATION OF SITES IMPACTED BY MINING ACTIVITIES

Grout feed Mixer Conveyor belt Contaminated soil feed

FIG. 12.7  Ex situ S/S treatment plant. Modified from HBR Limited, Essex—UK.

a pretreatment option is also particularly important in order to avoid damage of the plant units and delay to the work program (Bone et al., 2004).

12.6  STABILIZATION/SOLIDIFICATION OF TH-CONTAMINATED SOILS: APPLICABILITY, LIMITS AND COSTS

232

12.6.1  Effects of Contamination Level and Binder/Material Mix on S/S Performances Portland Cement (PC)-based S/S was shown to be a suitable technique to successfully treat Th-contaminated soils, even if high 232Th contamination can heavily inhibit its application (Falciglia et al., 2014b). The binder-material mix adopted was also demonstrated to significantly influence the performance of the S/S final products. In particular, the setting processes were shown to be strongly dependent on the thorium concentration in soil. Falciglia et  al. (2014b), investigating the influence of different PC-based and PC-barite grouts on S/S performance, showed that 232Th-concentrations in soil higher than 2% lead to setting times longer than 72 h (Table 12.2). In fact, hydration of cement is highly modified by heavy metal insoluble forms which, if present at high levels, inhibit the hydration processes and formation of the related hydration products due to coating around binder grains (Malviya and Chaudhary, 2006). This results in the formation of a protective coating of gelatinous hydroxide around the cement grain surface that determines grains of still-unhydrated cement and, consequently, in a retarding of the cement set and in a worsening of the physical performances of S/S-treated matrices (Chen et al., 2009). For this reason, in the case of 232Th concentrations higher than 2%, in situ S/S treatment is unsuitable due to the occurrence of a potential delay in the S/S mass curing and the transitory increase of the water amount in soil that could enhance the leachability of the contaminants, resulting in a downward contaminant migration during the in situ mixing processes. For 232Th concentrations lower than 2%, PC-based S/S (with or without barite) presents a good applicability also for the in situ interventions (BS EN limit of 24 h). 232

343

12.6  STABILIZATION/SOLIDIFICATION OF 232TH-CONTAMINATED SOILS



TABLE 12.2  Setting Time Values for Cement-Barite Based S/S 232Th-Contaminated Soils Grout Mix Soil:Grout Ratio

232

Initial Setting Time (h)

Final Setting Time (h)

PC (4:1)

2.0

1.2

3.0

≥5.0

>72

2.0

1.1

≥5.0

>72

2.0

1.4

≥5.0

>72

2.0

1.3

≥5.0

>72

2.0

1.3

≥5.0

>72

Th (%)

PC (4:1)

PC (3:1)

PC (3:1)

PC + Barite (3:1)

2.7

Density

2.8

3.1

3.1

2.9

24.5

Porosity

Density (g cm−3)

23.5 2.5 23.0 2.4 22.5 2.3

2.2

Porosity (%)

24.0

2.6

22.0

0

10

20

40 50 30 Barite amount (%)

60

70

80

21.5

FIG. 12.8  Density and porosity of S/S products as a function of the barite amount in grout (soil:grout = 3:1).

Recent findings (Falciglia et al., 2016) showed that the increase in barite amount in PC grout resulted in the S/S product density increasing and porosity decreasing (Fig. 12.8). This is due to the high density and fine texture of barite, which acts as a filler resulting in a pore refinement and porosity reduction. The addition of barite powder to PC resulted in the filling of the pores and in the increasing of the cohesion and compactness of the paste (Kilincarslan et al., 2006; Shaaban and Assi, 2011). However, exceeding a fixed barite:PC:soil ratio in the S/S mix was observed to result in a density decrease and porosity increase, due to the addition of an inert nonhydraulically active material that inhibits the PC hydrated product formation. This causes the formation of voids and thus a generation of weak bonds between soil aggregates and cement, leading to an increase in mixture porosity (Navi and Pignat, 1996). This effect

344

12.  RECLAMATION OF SITES IMPACTED BY MINING ACTIVITIES

3.50

PC:BA 1:1 PC

3.25

Density (g cm−3)

3.00 2.75 2.50 2.25 2.00 1.75 1.50

10

15

20

(A) 27

25 30 35 Grout amount (%)

40

45

50

PC:BA 1:1 PC

26

Porosity (%)

25 24 23 22 21 20 19

(B)

10

15

20

25 30 35 Grout amount (%)

40

45

50

FIG.  12.9  Density (A) and porosity (B) of S/S products as a function of grout amount in S/S products (barite:PC = 1:1).

could be more marked in the case of high metal concentrations in S/S paste (Falciglia et al., 2014b). A significant increase in density and reduction in porosity were also observed with increasing the grout amount in the grout:soil mix (Fig. 12.9). A change in density and porosity of S/S products is clearly reflected in a compressive strength (UCS) variation (Fig.  12.10). Falciglia et al. (2016) showed a marked UCS increase, up to ~20%, when replacing PC with barite (soil:grout = 3:1) up to a 50.0% barite content. Following the same behavior observed for the density variation, a further increase in Ba content resulted in a UCS reduction. Specifically, the maximum barite addition led to the UCS value of 7.45 MPa, lower than that observed for

345

12.6  STABILIZATION/SOLIDIFICATION OF 232TH-CONTAMINATED SOILS



10.0 9.5

UCS (MPa)

9.0 8.5 8.0 7.5 7.0 6.5 6.0

0

10

20

30 40 50 Barite amount (%)

(A)

60

70

80

11.0 10.5 10.0

UCS (MPa)

9.5 9.0 8.5 8.0 7.5 7.0 6.5 6.0

(B)

0

10

20

30 40 Grout amount (%)

50

60

FIG. 12.10  UCS variation as a function of the barite amount in grout (soil:grout = 3:1) (A) or grout amount in S/S products (barite:PC = 1:1) (B).

the S/S samples made with PC only (7.85 MPa). When the replacement increased to 40.0%, the UCS of concrete was subjected to a slight reduction. The increase in grout amount in S/S mix resulted in a significant increase in the mechanical performance of the S/S products. UCS values increased from 7.21 to 10.89 MPa, considering a grout content increase from 16.6% to 50.0%. In terms of quality acceptance, all the “S/S recipes” investigated by the authors led to the complete fulfillment of these requirements (US EPA, Netherlands and France criteria). In terms of durability, the replacement of PC in favor of barite powder leads to an increase of about 5% and 7% of the weight loss in durability F-T and W-D tests, respectively (Table  12.3), remaining, however, much lower than the maximum value accepted of 30%.

346

12.  RECLAMATION OF SITES IMPACTED BY MINING ACTIVITIES

TABLE 12.3  Weight Loss (%) from Wet-Dry and Freeze-Thaw Tests for Cement-Barite Based S/S 232 Th-Contaminated Soils (28-Day Curing) Grout Mix Soil:Grout Ratio

PC (4:1)

PC (3:1)

PC + Barite (1:1) (3:1)

F-T

2.1

0.5

5.1

W-D

2.3

0.8

7.5

TABLE 12.4  Values of Contaminant Leaching Rate for Cement-Barite Based S/S 232Th-Contaminated Soils (28-Day Curing) Grout Mix Soil:Grout Ratio −2

−1

LR (g cm  day )

PC (4:1)

PC (3:1)

6.56 × 10

−4

PC + Barite (1:1) (3:1)

6.26 × 10

−4

5.10 × 10−4

Likewise, the percentage variation of the grout amount in the S/S soil mixing results in a durability performance change. In the case of F/T durability, a slight increase in the weight loss at the end of the test was observed with a decrease in the amount of grout used (Falciglia et al., 2015). Overall, the W/D durability was shown to be worse than the F/T one. The addition of barite was also shown to significantly influence the 232Th leaching behavior of stabilized/solidified soils (Falciglia et al., 2012, 2014b). The leaching rate (LR) of 232Th from the soils treated with PC was reported to be in the range 6.02–6.56 × 10−4 g cm−2 day−1, whereas in the presence of barite, LR is reduced to 5.10 × 10−4 g cm−2 day−1(Falciglia et al., 2015) (Table  12.4). In terms of 232Th percentage immobilization, in a more recent study, Falciglia et  al. (2016) observed that in the presence of only PC, a total contaminant immobilization was achievable considering a pH value in the 5–7 range and leaching time up to 2  years (Table 12.5). A lower immobilization was found considering more acidic conditions (pH: 1–3). Total 232Th-immobilization, especially at pH in the range 5–7, is due to the low Th solubility and the sorption of Th ions into C-S-H hydrates (Li and Pang, 2014; Wierczinski et al., 1998). Thorium dioxide is extremely insoluble in nonhigh-acidic aqueous media (Hubert et al., 2001) and pH is the major controlling factor in the release of Th (Sapsford et al., 2012). For pH < 5 the ThO2 solubility sharply increases due to the dissolution process, which is favored by the presence of H+ ions in solution (Neck et  al., 2003). pH values lower than 5 also result in an increase in Th4+ formation and in a decrease in hydrolysis products (e.g., Th(OH)3+, +2 Th ( OH )2 ) (Anirudhan et al., 2013) and this leads to an increase in Th release. Vandenborre et al. (2008) reported that at pH = 3, about 56% Th4+ and 44% Th(OH)3+ are expected to be +2 formed, whereas at pH = 4, Th4+, Th(OH)3+, and Th ( OH )2 percentages of 10%, 67%, and 23% are predicted, respectively. Immobilization processes also include the formation of silicate hydrates capable of entrapping the Th-hydrolysis products or precipitates such as Th(OH)4 (Talip et al., 2009). 232Th-immobilization can also occur with increasing binder (PC) content and barite percentage in grout. It has been shown that the partial replacement of PC with barite significantly decreases the 232Th leaching phenomena, highlighting the possibility of achieving an almost complete immobilization even at the lowest pH values (Falciglia et al., 2016). The reduction in 232Th leaching rate was attributed to barite fine texture and high surface area, capable of reducing the porosity of the S/S matrix in respect to the treatment using PC only. This was by means of the refinement in pore structure, and increasing its

12.6  STABILIZATION/SOLIDIFICATION OF 232TH-CONTAMINATED SOILS



TABLE 12.5 

232

Th-Immobilization Efficiency (%) of PC-Barite S/S Soils 232

Th immobilization 365 Days (%)

232

Grout (G)

G (%)

pH

PC

25

7

100.00

100.00

33.3

7

100.00

100.00

50

7

100.00

100.00

25

5

100.00

100.00

33.3

5

100.00

100.00

50

5

100.00

100.00

25

3

99.20

98.71

33.3

3

100.00

99.68

50

3

100.00

100.00

25

1

98.17

97.81

33.3

1

98.55

98.03

50

1

98.50

98.55

25

7

100.00

100.00

33.3

7

100.00

100.00

50

7

100.00

100.00

25

5

100.00

100.00

33.3

5

100.00

100.00

50

5

100.00

100.00

25

3

100.00

99.23

33.3

3

100.00

100.00

50

3

100.00

100.00

25

1

99.22

99.00

33.3

1

98.87

99.67

50

1

100.00

99.91

PC:barite 1:1

347

Th Immobilization 730 Days (%)

contaminant adsorption capability. Reduction of porosity is well-known, corresponding to a reduction in the amount of pollutants leached from S/S matrices (Shaaban and Assi, 2011). For pH higher than 4.5, 232Th-precipitation phenomena occur and this significantly increases the adsorption mechanism. Furthermore, the formation of Th(OH)4 and the presence of charged Th(IV) hydrolysis products play an important role in the adsorption mechanism (Talip et al., 2009). The lower adsorption capacity observed at the lower pH values was also due to the competition between the H+ ions in the medium and the positively charged cationic species in the solution. On the other hand, when the pH is increased, the surface positive charge on the adsorbent decreases, which lowers the coulombic repulsion forces between the sorbing metal

348

12.  RECLAMATION OF SITES IMPACTED BY MINING ACTIVITIES

ions, resulting in a more electrostatically attractive surface to cations (Anirudhan et al., 2013). Overall, the authors concluded that, despite the variation in 232Th leaching observed with the change in pH, PC barite-based S/S treatment ensures excellent 232Th-immobilization, even in an extreme acidic environment.

12.6.2  Effects of Contamination and Binder/Material Levels on S/S Product γ-Ray Shielding A change in S/S product physical properties, especially density, was shown to heavily influence their γ-radiation shielding performances. Falciglia et al. (2014b, 2015) reported that in S/S soil systems, γ-radiation shielding performances can be expressed as γ-ray shielding (γRS) index. Authors calculated very low γRS values when PC was used alone, but, barite addition in grout mix caused a marked increase of the γ-rays shielding properties. The higher shielding effectiveness of a treatment with PC and barite can be better understood by comparing a typical γ-ray energy spectra obtained by means of a γ-ray counting spectrometer including a NaI(TI) detector for a PC-based S/S and PC-barite-based S/S treatment (Fig. 12.11).

Counts of g ray (CPS)

120

Contaminated soil S/S treated soil

100 80 60 40 20 0 0

100

200

(A)

Counts of g ray (CPS)

120

300 400 Energy (keV)

500

600

Contaminated soil S/S treated soil

100 80 60 40 20 0 0

(A)

100

200

300 400 Energy (keV)

500

600

FIG. 12.11  Typical γ-ray energy spectra for a PC-based S/S (A) and PC-barite-based S/S (B) treatment obtained by means of a γ-ray counting spectrometer with NaI(TI) detector.

349

12.6  STABILIZATION/SOLIDIFICATION OF 232TH-CONTAMINATED SOILS



50

Barite amount in grout (S:G = 3:1)

45

Grout amount in S/S mix (PC) Grout amount in S/S mix (PC:BA) = 1:1)

40

g RS index (%)

35 30 25 20 15 10 5 0

0

5

10

15

20

25 30 35 40 45 Amount (%)

50 55

60

65

70

75

FIG.  12.12  γRS variation as a function of the barite amount in grout (soil:grout = 3:1) or grout amount in S/S products (barite:PC = 1:1).

The variation of density with barite or grout amount (Figs. 12.8 and 12.9) was shown to be strictly linked with the γRS variation. Falciglia et al. (2016) studied the influence of barite replacement in S/S grout on γRS variation. They reported that a maximum γRS index of 23.4% was achieved considering a barite content of 50%, corresponding to a γRS index increase of about twofold with respect to the case of PC only (Fig. 12.12). However, a barite content higher than 50% resulted in a marked decrease in γRS. A γRS of 14.7%, not significantly different from that observed in the case of PC without barite (9.6%), was obtained for a barite content of 75%. In the same study, a γRS increase from 9.7% to 33.1% with increasing the grout content from 16.6% to 33.0% was observed. However, a further increase in the grout percentage did not lead to a marked γRS increase, defining a threshold grout:soil ratio that should be considered in scaling-up and economic analysis. In fact, increasing the grout content in soil from 33.0% to 50.0% corresponded only to a slight γRS rise of ~4%. Another important point involves the possibility of considering the γRS values independently from the 232Th-contamination level present in the soil. In fact, shielding properties, expressed as a γRS index of S/S, were shown to be not dependent on the level of radioactivity and thus on radionuclide concentration in soil (Falciglia et al., 2015) (Fig. 12.13). This aspect is fundamental since it allows shifting all the γ-ray shielding performance information (γRS) to any contamination level. Knowing γRS values of a selected S/S treatment as a function of the emission energy allows the calculation of the radioactivity of an S/S treated soil also in the presence of high-level radioactive contaminants. Consequently, considering a theoretical scenario connected to a real radionuclide-contaminated site, the gamma-radiation emission values after S/S treatment, assessable by γRS, could be used in order to calculate the absorbed radioactivity dose. The comparison of the calculated absorbed dose values to regulatory radiometric limits allows the understanding of the effectiveness of the full-scale S/S applicability.

350

12.  RECLAMATION OF SITES IMPACTED BY MINING ACTIVITIES

74.81 keV

87.20

89.80

238.63

241.00

277.73

70 60

γ RS (%)

50 40 30 20 10 0

0

1

2

3

4

5

6

7

8

9

10

11

12

13

14

15

16

Th (%) 300.09

338.32

583.02

726.84

861.11

910.40

968.94

50

γ RS (%)

40 30 20 10 0

0

1

2

3

4

5

6

7

8

9

10

11

12

13

14

15

16

Th (%)

FIG. 12.13  γRS variation as a function of 232Th contamination level and decay energy for cement-barite-based S/S.

12.6.3 Costs In situ S/S is widely considered as a sustainable technique to remedy soil impacted with organic or metal contaminants. The literature reports variable costs for its application, depending on the specific site conditions and materials employed in the interventions; however, studies generally report very sustainable costs, especially if compared with contaminant removal-based treatments. Harbottle et al. (2007), in a study aimed at comparing the S/S technical sustainability with off-site disposal to landfill, reported an estimated cost of remediation of approximately £ 28 ton−1 of soil. Al-Tabbaa and Perera (2006) reported ex situ PC-based mixing technology costs in the £ 30–38 ton−1 range. More recently, Zhang et  al. (2015), in their preliminary estimation of amendment costs based on a pilot-scale field test, calculated costs up to about USD 22 ton−1, highlighting the high availability and cost-effectiveness of the materials employed in S/S of lead-acid battery contaminated soils. Higher costs have been reported by Wang et al. (2015a, 2015b) for a PC-based S/S



12.7  Potential Directions for Future Research

351

treatment. According to the online quotations from the largest regional supplier (Alibaba China, 2014), authors reported that costs were approximately USD 63 ton−1 for PC and USD 7.5 ton−1 for fly ashes.

12.7  POTENTIAL DIRECTIONS FOR FUTURE RESEARCH The scientific literature and technical reports have demonstrated PC-based S/S to be a suitable technique for the successful remediation of 232Th-contaminated soils. However, recent lab-scale investigations remarked that Th contamination level significantly influences the performance of the treated samples in terms of setting time, compressive strength, durability and leaching, and that Th concentrations higher than 2% significantly worsen the S/S applicability. On the contrary, the addition of barite aggregates to the cement grout leads to a slight improvement of the S/S performance in terms of durability, mechanical strength, contaminant leaching and radioprotection. In fact, it was found that the replacement of PC with barite, up to 50% of content, allows the achievement of a marked UCS increase, up to ~20%, and of almost the total 232Th-immobilization, even in extreme acidic environmental conditions. Barite addition also results in a significant γ-ray shielding increase, up ~2-fold with respect to the case of PC alone. A γ-ray shielding increase is also achievable with increasing grout content in S/S soil mixtures. γRS of 9.7%, 13.8%, 23.8%, 33.1%, and 37.1% were observed for increasing grout percentages of 16.6%, 20.0%, 25.0%, 33.0%, and 50.0%, respectively. However, grout contents higher than 33.0% does not lead to a further significant improvement of the S/S product γ-ray shielding performance, highlighting a potential threshold that must be considered in scaling-up and economic analysis. This suggests PC-barite based S/S as an optimal choice for the treatment of 232Th-contaminated soils, even if the finding of new promising S/S shielding mixtures employing novel high-density shielding materials (magnetite, iron powder) could be of interest and an important research area in order to increase the shielding performance and consequently reduce the material costs.

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12.  RECLAMATION OF SITES IMPACTED BY MINING ACTIVITIES

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A comparison of the technical sustainability of in situ stabilisation/ solidification with disposal to landfill. J. Hazard. Mater. 141, 430–440. Hubert, S., Barthelet, K., Fourest, B., Lagarde, G., Dacheux, N., Baglan, N., 2001. Influence of the precursor and the calcination temperature on the dissolution of thorium dioxide. J. Nucl. Mater. 297, 206–213.



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