Cesium retention and release from sulfur polymer concrete matrix under normal and accidental conditions

Journal of Hazardous Materials 381 (2020) 121180

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Cesium retention and release from sulfur polymer concrete matrix under normal and accidental conditions

T

Piotr Szajerski , Agnieszka Bogobowicz, Andrzej Gasiorowski ⁎

Institute of Applied Radiation Chemistry, Lodz University of Technology, Wroblewskiego 15, 93-590, Lodz, Poland

ARTICLE INFO

ABSTRACT

Editor: D. Aga

This paper proposes an efficient two-stage process for stabilization and solidification of the Cs-137 isotope in a sulfur polymer concrete (SPC) matrix. Lignite slag (SL) and fly ash (FA) were applied as active fillers for cesium immobilization. To study the release of Cs-137 isotope and determine the tracer activity in the leachates, we applied a slightly modified ANSI/ANS 16.1 protocol and the gamma spectrometry technique. The measured effective diffusion coefficients for the Cs-137 isotope were between 0.84·10−9 and 3.10·10−9 cm2·s−1. Normalized leaching rates were within the range of 1.74·10−5 – 3.85·10−5 g·cm−2·d−1, fulfilling acceptance criteria for radioactive wasteforms. As well as standard leaching under static conditions, we also studied dynamic leaching of SPC samples at increased temperatures and leaching in an aggressive environment. The Cs-137 effective diffusion coefficients were found to increase by 3 – 4 orders of magnitude (10−6 – 10−5 cm2·s−1), while the normalized leaching rate reached values of up to 2.36·10−3 g·cm−2·d−1 after 28 days of leaching. The proposed cesium immobilization mechanism is based on the formation of cesium silicate and aluminosilicate phases, together with effective matrix sealing during the SPC manufacturing process.

Keywords: Sulfur polymer concrete Leaching Diffusion Industrial waste Cesium immobilization

1. Introduction Radioactive waste materials are considered to be one of the most hazardous substances produced by humans. Without special treatment, radioactive contaminants can be easily spread in the environment, causing a serious threat to human health and all living organisms. Since the early days of the nuclear industry, special attention has therefore been focused on the development of safe radioactive waste stabilization and storage technologies, under the watchful eye of the public and government institutions. Of the different technologies currently available for radioactive liquid waste treatment, several techniques are commonly used in industry and include physico-chemical methods (adsorption and ion exchange, extraction, chemical precipitation and advanced oxidation processes) (Gu et al., 2018; Ojovan and Lee, 2014; Mao et al., 2017; Xia et al., 2018; Liu et al., 2016; Majidnia and Idris, 2015), membrane based methods (ultra, micro and nano filtration, reverse and forward osmosis, membrane distillation) (Ambashta and Sillanpää, 2012; Chmielewski et al., 1999), electrochemical methods (electrodialysis and electroadsorption) (Akhter et al., 2018), as well as evaporation techniques, biological methods and integrated (combined)

processes (Ojovan and Lee, 2014; Zhang et al., 2019a). Solid residues from the treatment of liquid radioactive waste must be subsequently stabilized in durable final wasteforms, which should be resistant to environmental and accidental conditions. Applications of the typical, classical materials used for the solidification of radioactive wastes, such as cement and asphalt, are limited by their poor thermal and radiation stability. In practice, their applications are restricted to low and intermediate-level waste. Cement based materials are porous, relatively susceptible to leaching and possess limited corrosion resistance against acids, chlorides, sulfates and carbon dioxide, as well as some microorganisms. Processing of spent nuclear fuel requires a few steps: cooling, separation from uranium and plutonium and finally vitrification in a borosilicate or phosphate glass matrix (Ojovan and Lee, 2014; Roth and Weisenburger, 2000; National Research Council, 1996a; Olanrewaju, 2010). All of these operations generate streams of radioactive contaminants which flow into processing media (primary cooling circuit water, water in spent fuel cooling ponds, exhaust gases from the vitrification process or into liquids used for surface decontamination) (Ojovan and Lee, 2014; Zhang et al., 2019a; National Research Council,

Abbreviations: SPC, sulfur polymer concrete; FA, fly ash; SL, slag; PWR, pressurized water reactor; SPB, sulfur polymer binder; STY, styrene; DCPD, dicyclopentadiene; DEC, decene; QS, quartz sand; CFL, cumulative fraction leached; SEM, scanning electron microscopy; EDS, energy dispersive spectrometry; DSC, differential scanning calorimetry ⁎ Corresponding author. E-mail addresses: [email protected] (P. Szajerski), [email protected] (A. Bogobowicz), [email protected] (A. Gasiorowski). https://doi.org/10.1016/j.jhazmat.2019.121180 Received 16 February 2019; Received in revised form 3 September 2019; Accepted 6 September 2019 Available online 06 September 2019 0304-3894/ © 2019 Elsevier B.V. All rights reserved.

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1996b; Ishigure et al., 1989). These streams are in most cases categorized as intermediate or low level waste (ILW, LLW). Cesium isotopes, especially Cs-137 and Cs-134, are the most abundant radionuclides found in the radioactive waste streams generated by the nuclear industry (nearly 3.5 mol% in spent nuclear fuel) (Ojovan and Lee, 2014). As with other fission and activation products, cesium isotopes can be found not only in liquid waste from spent nuclear fuel reprocessing and cooling water from the primary reactor circuit, but also in waste generated by other non-nuclear industrial activities (Zhang et al., 2019a). Cesium isotopes belong to the group of alkali metals, which is a cause of their high mobility. Therefore, cesium isotopes, especially Cs-137, are commonly used as tracers to investigate radionuclides immobilization techniques and characterization of materials for radioactive waste stabilization. In the last few years, many new materials have been developed which enable efficient separation of isotope Cs-137 from contaminated waters and soils (Alamudy and Cho, 2018; Kim et al., 2017; Zhang et al., 2019b; Nilchi et al., 2011). Many of these streams are immobilized in cementitious and bitumen matrices. However, other alternative immobilization methods, such as new compositions based on hydraulic cement (Koťátková et al., 2017; Gougar et al., 1996), ceramic matrices (Bohre et al., 2017; Trocellier, 2000; Stanek et al., 2012), polymers (Ojovan and Lee, 2014; Ojovan, 2011) or zeolites (Alby et al., 2018; Delkash et al., 2015; Rożek et al., 2019; Hosseini Asl et al., 2019), are being considered. A particularly interesting solution is sulfur polymer concrete technology (SPC). The first formulations of SPC were elaborated during the 1980s (ACI Committee, 1988; Vroom, 1992), and the idea of immobilizing hazardous waste in SPC matrices was presented shortly afterwards (Mohamed and El Gamal, 2007; Lin et al., 1995; Mattus and Mattus, 1994). Since then, the possibility of immobilizing radioactive waste in SPC has been the subject of intense research (Mattus and Mattus, 1994; Darnell, 1993; Mayberry et al., 1993a, b; van Dalen and Rijpkema, 1989). There is little data on cesium immobilization in sulfur polymer concrete matrices, but Dalen and Rijpkema report some experimental results of a study in which cesium-bearing wastes, in the form of incinerator ashes or borate waste, were immobilized in SPC (van Dalen and Rijpkema, 1989). According to the authors, the diffusion coefficients for Cs+ ions in SPC matrices were in the range of 10−14 – 10−13 m2·s−1. These results may be compared with those for hydraulic cement based concrete, as reported for example by Sullivan (2004). Due to the reduced contact surface between the leaching agent and the matrix, sulfur polymer concrete has enhanced performance in comparison with Portland cement concrete, in terms of immobilization of contaminants. Moreover, sulfur polymer technology allows the final product to be obtained after only 24 h since preapartion. Different types of fillers may also be used for SPC preparation, which strongly facilitates the use of problematic industrial wastes, such as fly ashes and slags, which as yet have a limited range of possible uses. There is increasing demand for new applications of waste and residual materials from industry, which often have interesting properties and can be successfully applied as alternative materials in common technical practice (Shiota et al., 2017, 2018). In this study, we investigate the leaching behavior of Cs-137 isotope immobilized in SPC matrices based on modified sulfur and industrial waste fillers, under static, dynamic and aggressive conditions. The novelty of the proposed solution lies in its application of lignite combustion by-products, as well as the sequential, two-stage immobilization procedure, whereby after initial stabilization of the radioactive contaminants on a fly ash or slag filler the matrix is solidified and sealed with a polymeric sulfur binder. Although very promising results have already been obtained in the case of Co-60 immobilized in SPC matrices (Szajerski et al., 2019a), it is necessary to verify the ability of SPC matrices based on lignite combustion residues to immobilize other radioactive isotopes generated by the nuclear industry, particularly Cs137, which is one of the most abundant fission products generated during the operation of light water reactors.

2. Materials and methods 2.1. Materials and chemicals Elemental sulfur, min. 99.95% (S, Siarkopol Tarnobrzeg Sp. z o.o., Poland), and the co-monomers styrene (STY), min. 99.0%, dicyclopentadiene (DCPD), min. 95.0%, furfural (FUR), min. 98.0% and 1decene (DEC), min. 97.0% (all from Sigma-Aldrich Polska Sp. z o.o., Poland) were used for sulfur modification. Technical grade lignite slag (SL) and fly ash (FA), provided by PGE Gornictwo i Energetyka Konwencjonalna S.A. Oddzial Elektrownia Belchatow (Poland) were used as mineral fillers. Radioactive tracer Cs-137 (type ER 3) used for SPC labeling was purchased from the Czech Metrology Institute (Czech Republic). 2.2. Synthesis of sulfur polymer binder (SPB) A two-stage process was applied for the preparation of Cs-137 labeled SPC. The first stage was the synthesis of the sulfur polymer binder (SPB), which involved melting sulfur (at around 120 °C) followed by the continuous introduction of a co-monomer or mixture of co-monomers, while the reaction mixture was stirred and maintained at a temperature of 130 – 140 °C for about 6 h. The SPB was then poured into a steel form and left to cool, forming solid modified sulfur. The SPB compositions used are presented in Table 1. 2.3. Preparation of Cs-137 labeled sulfur polymer concrete (SPC) samples The mineral fillers (SL, FA) used to prepare the SPC samples were dried overnight at 105 °C. The fillers were used without any chemical pre-treatment, and were in the native form obtained from the power plant. The coarse fraction of the aggregates in the slag waste was ground in a porcelain mortar. In order to speed up evaporation of the solvent and reduce agglomeration of the filler, the Cs-137 tracer was applied to the fillers by the dropwise addition of an isotope solution in water/acetone mixture (1/1, v/v), with a specific activity of about 20 kBq·g−1, in a Cs+ (CsCl) carrier. The main purpose of applying the cesium tracer to the fillers was to initially immobilize/adsorb the Cs-137 tracer. The studies on cesium immobilization in SPC described in this work were preceded by adsorption studies of Cs-137 on pure lignite slag and fly ash used as fillers in SPC. The results clearly showed cesium immobilization on these fillers to be effective, even in solutions of high ionic strength (Gasiorowski and Szajerski, 2016). The dried (overnight at 80 °C) labeled fillers were again homogenized in an agate mortar and used to prepare SPC samples consisting of modified sulfur (25%, w/w), filler with Cs-137 tracer (30%, w/w) and 0 – 2 mm quartz sand (45%). All the components were gently mixed in powdered form, then the mixture was slowly heated up to around 130 °C in a glass beaker using a heating mantle, until melting of the SPB in batch (about 30 g). This temperature was maintained for ca. 10 – 15 min. During heating, the batch was mixed continuously, and after complete homogenization the molten mixture was transferred to PTFE casting molds, preheated to 130 °C and pressed to form SPC samples. Table 1 Modified sulfur binders used for SPC preparation. Component

Elemental sulfur (S) Dicyclopentadiene (DCPD) Furfural (FUR) Decene (DEC) Styrene (STY)

2

Fraction in SPB (w/w): SP4

SP5

SP6

SP7

0.95 0.05 – – –

0.90 0.05 0.05 – –

0.90 0.05 – 0.05 –

0.90 0.05 – – 0.05

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activities of the Cs-137 tracer (ai) released during the leaching process were determined according to Eq. (1),

Table 2 Composition of SPC samples labeled with the Cs-137 tracer; (SP4-7 – sulfur polymer binders, SL, FA, QS – lignite slag, lignite fly ash, quartz sand); activity given at date of leaching start. Sample

SL_SP4Cs SL_SP5Cs SL_SP6Cs SL_SP7Cs FA_SP4Cs FA_SP5Cs FA_SP6Cs FA_SP7Cs SL_SP6Cs-DYN FA_SP6Cs-DYN SL_SP6Cs-HCl FA_SP6Cs-HCl

Fraction in sample (w/w): SP4

SP5

SP6

SP7

SL

FA

QS

0.25 – – – 0.25 – – – – – – –

– 0.25 – – – 0.25 – – – – – –

– – 0.25 – – – 0.25 – 0.25 0.25 0.25 0.25

– – – 0.25 – – – 0.25 – – – –

0.30 0.30 0.30 0.30 – – – – 0.30 – 0.30 –

– – – – 0.30 0.30 0.30 0.30 – 0.30 – 0.30

0.45 0.45 0.45 0.45 0.45 0.45 0.45 0.45 0.45 0.45 0.45 0.45

m, g

ACs-137, Bq

1.35 1.61 1.53 1.67 1.42 1.93 1.51 1.77 2.08 1.49 1.54 1.48

1148 1369 1301 1420 1208 1641 1284 1505 1769 1267 1310 1259

ai =

Is

Ibkg d

ln2 t

e T1/2

(1)

em

where Is and Ibkg are the sample and background count rates for 661.66 keV emission, in cps, εd is the detection efficiency of the system, ωem is the 661.67 keV photon emission intensity, T1/2 is the half-life of the Cs-137 tracer, in days and Δt is the time interval between the start of the experiment and sample measurement time, in days. The fractional activities were corrected for the decay from the start of the experiment. 2.6. Calculation model for Cs-137 tracer release from leached SPC samples A diffusive model of tracer transport from solid matrix was used to analyze the results of Cs-137 release from SPC during the leaching experiments. This model mimics environmental conditions, in which the waste matrix is exposed to long term contact with a leaching agent and the main driving force for tracer release is the diffusion of cesium within the volume of the wasteform material. Such analysis requires a detailed description of the sample geometry. The initial dependence between the tracer fraction released from the matrix, the experimental conditions and the time of leaching is Fick’s second law, presented in Eq. (2) (Crank, 1975; Červinková et al., 2007; Rahman et al., 2007),

The compositions of the samples are detailed in Table 2. 2.4. Cs-137 leaching under static, dynamic and aggressive conditions Leaching tests for Cs-137 labeled SPC samples were carried out under static and dynamic conditions. Static leaching experiments were performed according to a slightly modified ANSI/ANS 16.1 protocol (ANSI/ANS-American National Standards Institute/American Nuclear Society, 2019), closely related to the older procedure recommended by the IAEA ISO 6961-1982 (ISO, 1982). The modification was related to sample collection times, which were set according to the ISO test. The sample surface (S) to leaching agent volume (V) ratio was set to S/ V = 0.1 cm−1. For each experiment, the sample was placed in a polyethylene container filled with deionized water as a leaching agent. Deionized water was used as a leaching agent (Milli-RO, Milli-Q system, Millipore). The water was replaced after 1, 3, 7, 14, 28, 56 and 90 days, according to the ISO protocol. Dynamic leaching tests were performed for selected samples with cyclic solvent renewal. Each tested sample was placed in a Soxhlet extractor equipped with a reflux condenser. For dynamic leaching experiments, the sample surface (S) to leachant volume (V) was set to around 0.05 cm−1. The water was replaced after 1, 3, 7, 14, 28 and 56 days. Leaching of the Cs-137 tracer from the SPC samples under aggressive conditions was carried out in 1.0 M HCl solution. The leaching solution was replaced after 0.5, 2 and 8 h and after 1, 3, 7, 14, 28, 56 and 90 days. At the time of each replacement of the leaching agent, in all performed leaching experiments the selected samples were characterized by weighing and measuring their geometrical dimensions, in order to check for possible swelling or weight loss.

2C C =D 2 t x

(2) −3

where C is the tracer concentration, in Bq·m , D is the tracer diffusion coefficient in the investigated material, in m2·s−1, t is the diffusion time, in s and x is the diffusion coordinate in the investigated material, in m. During the leaching test period, for a known initial condition (C(x,0) = 0) and given boundary conditions (C(∞,t) = Co and C(0,t) = 0, where Co is the initial tracer concentration in the solid wasteform), the analytical solution of Eq. (2) leads to an expression connecting the Cumulative Leach Fraction (CLF) of the tracer released from the waste material and the leaching time, as expressed by Eq. (3) (Crank, 1975; Rahman et al., 2007; Mendel, 1973; Mattigod et al., 2001; Serne et al., 1995; Kim et al., 2001; Kosson et al., 2002; Papadokostaki and Savidou, 2009; Rahman and Zaki, 2011; Hinsenveld, 1993),

CLF =

ai D =2 e Ao

1/2

S 1/2 t + V

(3)

where ai and Ao are the fractional activity leached at time tn-tn-1 and initial activity of the Cs-137 tracer in the sample, in Bq, De is the effective diffusion coefficient of the tracer in the investigated material, in m2·s−1, S is the total surface of the sample, in m2, V is the volume of the leaching medium, in m3, t is the leaching time, in s and α is a parameter describing the fraction of the tracer released initially (not subjected to diffusive transport, but released during the initial contact between the leaching medium and the waste material, and located usually on the surface of the sample or in close proximity to the surface). The dependence given by Eq. (3) allows for determination of parameters describing diffusive transport of the tracer within the sample volume. It connects the experimental dependence of the cumulative fraction of the Cs-137 tracer leached from the material vs. the square root of time, Σai/Ao = f(t1/2), and allows for slope calculation, k = 2(De/π)1/2(S/V). Thus, the effective diffusion coefficient can be calculated after rearrangement according to Eq. (4):

2.5. Measurement of Cs-137 activity in leachates Cs-137 activity released during leaching was determined using gamma spectrometry. A 3 inch NaI(Tl) crystal detector was used (76B76/3 M, Scionix, Holland) coupled with a Flir ScintiSpec multichannel analyzer (Flir Radiation GmbH, Germany). Leachate samples (65 cm3 in volume) were counted inside the 5 cm thick wall lead shield and 4 mm zinc plate cylindrical liner. The collected γ-spectra were analyzed using winTMCA software, taking into account background counts and the contribution of the Bi-214 isotope (609.3 keV photon energy). Detection efficiency was determined using Cs-137 standard solutions of different activities (measured under the same geometrical conditions as the samples), and the tracer concentration in the standards and leachates was analyzed in terms of the Ba-137 m photon emission rate (661.66 keV, 84.99% emission intensity). The fractional

De =

k2 4

V S

2

(4)

Normally, final wasteforms are multi-barriered packages with durable external walls. However, after destruction of the outer liners of the final wasteform, the waste matrix is directly exposed to interaction with leaching agents. As the waste matrix is the primary barrier against the 3

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spread of radionuclides, the best measure of immobilization efficiency is the rate, at which they are released from the waste matrix, and this rate for each radionuclide can be quantified using Eq. (5) (Ojovan and Lee, 2014).

LR =

ai m Ao S t

ions on the surface of the fly ash filler, as also suggested by preliminary adsorption studies of isotope Cs-137 with slag and fly ash fillers (Gasiorowski and Szajerski, 2016). Both fillers were selected on the basis of the results of previous immobilization studies on Co-60, in which phosphogypsum filler was additionally tested, but was found to be ineffective. Our former results revealed, that in case of Co-60 immobilization in the sulfur polymer concrete matrix based on the slag and fly ash fillers, the Cumulative Leach Fractions obtained were found to be at most 0.61 and 0.26%, respectively, whereas for the phosphogypsum based SPC reached even 80% after 90 days of leaching (Szajerski et al., 2019a). Basically, for the Cs-137 isotope immobilization we applied the similar experimental procedure as for the Co-60 tracer. Graphical representations of the data showing Cs-137 release kinetics are provided in Fig. 2(A1-2) for the slag based sample (SL_SP4Cs) and in Fig. 2(B1-2) for FA based (FA_SP4Cs) SPC samples. Fig. 2(A1) and (B1) show nearly the same leaching characteristics. The cumulative fractions leached from each sample differ slightly and reach values of 2.2 and 1.6% for SL and FA based concretes, respectively, and in comparison with the results for divalent cobalt ions are roughly 3.5 and 6 times higher (Szajerski et al., 2019a). For the purpose of this data analysis, useful dependence must correlate the Cumulative Leach Fraction (CLF) and the square root of time. These dependencies are presented in Fig. 2(A2 and B2) for the samples SL_SP4Cs and FA_SP4Cs. The data presented in Table 3 provide further insight into the experimental results for all the tested samples. For all the concretes, similar behavior was observed, with CLF values of between 1.08 and 2.21%. Based on the data in Table 3, there is no clear evidence that Cs-137 immobilized in a fly ash based SPC exhibits significantly different leaching behavior in comparison with the slag based matrices. This is supported by the values for the fractional activities of Cs-137 in the collected leachates. Analysis of the presented kinetic traces based on the linear fit of the experimental data included in Table 3 allows for calculation of the cesium ion diffusion rate within the matrix volume and of its diffusion coefficients. Effective diffusion coefficients are presented in Table 4 and are of the order of 10−10 – 10−9 cm2·s−1, with the lowest value for the sample filled with fly ash with SP5 sulfur polymer binder (FA_SP5Cs) De = 8.44·10−10 cm2·s−1 and the highest for FA_SP6Cs with De = 3.10·10−9 cm2·s−1. The criterion for potential utilization of the wasteform matrix for radioactive waste immobilization is the material release rate from the matrix. The normalized leaching rate (LR) values for the investigated samples were calculated according to Eq. (5). Typical dependencies of the LR coefficients vs. specimen leaching time for SL_SP4Cs and FA_SP4Cs are presented in Fig. 3. Detailed data on the normalized leaching rate for all the investigated samples are given in Table 5, which presents LR values obtained after 28 and 56 days. In most countries, national regulations consider the normalized leaching rate as the main factor in the evaluation of radioactive waste matrix immobilization efficiency. The lower the value of the LR factor, the better the performance of the matrix, as it provides a higher level of safety during long-term storage. For intermediate level waste, normalized leaching rate values after 28 days of leaching in deionized water should be below 10−3 g·cm−2·d−1 (Ojovan and Lee, 2014; J. Laws, 2015; Rahman et al., 2014). For all the investigated samples, whether based on slag or fly ash filler, this condition was successfully met. Given the data presented in Table 5, it may be observed that after 28 days of leaching the normalized leaching rate was between 1.74·10−5 and 3.85·10−5 g·cm−2·d−1 and remained in roughly the same range after 56 days (1.89·10−5 – 3.39·10−5 g·cm−2·d−1). Based on these results, the investigated matrices can be considered as a potential material for the immobilization of low and intermediate level wastes. Although the leaching rate after 28 days is a valid criterion for evaluating waste matrix usability, careful analysis of the leaching rate values after 56 days of leaching indicates that the cesium release rate from the wasteform is stabilized or can even be further decreased over a

(5)

where m is the weight of the sample, in g, and Δt = tn-tn-1 is the fractional leaching time, in days. Other symbols are as explained previously. The criterion for deciding whether the investigated material could potentially be used as a radioactive waste stabilization matrix is the normalized tracer leaching rate from the material after 28 days of leaching, LR, in g·cm−2d−1. Under Polish regulations, the acceptance criteria for materials used as wasteforms for radioactive waste immobilization are established by the leaching rate after 28 days. This parameter must not be higher than 10−5 g·cm−2·d−1 for high level waste (HLW), 10−3 g·cm−2·d−1 for intermediate level waste (ILW) and 10−2 g·cm−2·d−1 for low level waste (LLW) (J. Laws, 2015). Similar limiting values for leaching rates can be found in other countries, such as Russia (Ojovan and Lee, 2014; Rahman et al., 2014). 2.7. Differential scanning calorimetry of sulfur polymer binders Differential scanning calorimetry (DSC) measurements were performed using the Differential Scanning Calorimeter Q Series (TA Instruments, USA), with a 10 deg·min−1 heating rate in aluminum crucibles in a temperature range of 0 – 200 °C. 3. Results and discussion 3.1. Static leaching of Cs-137 labeled SPC samples The results of leaching experiments performed on Cs-137 labeled SPC samples revealed very good immobilization characteristics for cesium. Sample gamma ray spectra for leachates from a SL_SP4Cs sample after 1 and 90 days of leaching time are presented in Fig. 1. A significant reduction in the 661.66 keV photon count rate may be observed after 90 days, in comparison with leachate obtained after 24 h. Comparable effects were observed for both the slag and fly ash filler based samples. The reduction in leachate activity was more pronounced for the FA based samples. However, only moderately worse cesium retention performance was observed for SL filler. The Cs-137 activities measured in the FA based SPC contacted leaching solutions after each leachant exchange were nearly half those of SL filler based samples obtained leachates. This was the result of strong adsorption of cesium

Fig. 1. Gamma ray spectra for SL_SP4Cs leachates after 1 and 90 days of static leaching; spectra counting times 7200 and 14,400 s, respectively, for samples collected after 1 and 90 days. 4

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Fig. 2. Leaching kinetics for Cs-137 tracer release from samples of sulfur polymer concrete based on lignite slag (SL_SP4Cs: A1-2) and fly ash (FA_SP4Cs: B1-2).

longer period of leaching (Table 5). Additionally, the fact that the highly mobile monovalent Cs-137 tracer was used in this study gives grounds to believe that even better immobilization efficiency might be observed for di- and trivalent cations, which are commonly present in radioactive effluents generated in primary cooling circuits due to fission and activation processes. On the basis of the studies presented here, it is difficult to speculate on the influence of the conspicuous co-monomer or co-monomers on the final properties of the investigated samples. Analysis of the CLF values for all investigated SPC samples (Table 3) revealed that, in the case of the slag filler based concretes, the lowest cumulative Cs-137 activity was released from the SP7 sulfur polymer binder based sample (SL_SP7Cs, 1.52%), which was prepared using styrene and dicyclopentadiene co-monomers (cf. Table 1). In the case of fly ash filled sulfur polymer concretes, the lowest CLF value (1.08%) was found for the FA_SP5Cs sample, which was prepared using sulfur polymer binder modified with dicyclopentadiene and furfural co-monomers. About 50% more activity (in comparison to the FA_SP5Cs sample) was released from the SP4 binder based fly ash filled concrete (FA_SP4Cs), reaching a value of CLF = 1.57%.

Table 4 Effective diffusion coefficients in SPC samples measured for Cs-137 isotope. Sample

Effective diffusion coefficient, De, cm2·s−1

SL_SP4Cs SL_SP5Cs SL_SP6Cs SL_SP7Cs FA_SP4Cs FA_SP5Cs FA_SP6Cs FA_SP7Cs

(2.70 (1.75 (2.66 (1.19 (1.36 (8.44 (3.10 (2.97

± ± ± ± ± ± ± ±

0.17)·10−9 0.22)·10−9 0.15)·10−9 0.10)·10−9 0.09)·10−9 0.50)·10−10 0.11)·10−9 0.20)·10−9

Analysis of the effective diffusion coefficients provides further insight into the factors affecting cesium immobilization efficiency. As with the results for the CLF parameter, the best performance (lowest values for the diffusion coefficient) were observed for the SP5, SP7 and SP4 binder based concretes (cf. Table 4). Again, these binders were prepared using dicyclopentadiene, styrene and furfural (cf. Table 1). The values for the normalized leaching rate presented in Table 5

Table 3 Partial and cumulative fractions of Cs-137 leached from SPC samples. Sample

Fraction leached after time (days), ai/Ao: 1

SL_SP4Cs SL_SP5Cs SL_SP6Cs SL_SP7Cs FA_SP4Cs FA_SP5Cs FA_SP6Cs FA_SP7Cs

3 −3

4.65·10 6.26·10−3 4.31·10−3 3.49·10−3 2.80·10−3 1.03·10−3 3.00·10−3 2.84·10−3

Cumulative Leach Fraction CLF = Σai/Ao

7 −3

3.97·10 3.69·10−3 3.98·10−3 2.62·10−3 3.20·10−3 2.21·10−3 3.08·10−3 3.10·10−3

14 −3

2.06·10 1.64·10−3 1.87·10−3 1.46·10−3 1.45·10−3 8.98·10−4 1.62·10−3 1.90·10−3

28 −3

2.01·10 2.34·10−3 1.95·10−3 1.23·10−3 1.70·10−3 1.28·10−3 1.82·10−3 2.46·10−3

56 −3

2.88·10 2.23·10−3 2.05·10−3 2.18·10−3 1.70·10−3 1.65·10−3 2.85·10−3 2.96·10−3

5

90 −3

3.54·10 2.22·10−3 3.77·10−3 1.79·10−3 2.38·10−3 1.68·10−3 3.94·10−3 3.13·10−3

2.97·10−3 2.57·10−3 4.03·10−3 2.46·10−3 2.51·10−3 2.05·10−3 4.21·10−3 3.84·10−3

2.21·10−2 2.10·10−2 2.20·10−2 1.52·10−2 1.57·10−2 1.08·10−2 2.05·10−2 2.02·10−2

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Fig. 3. Normalized leaching rates for selected SPC samples based on lignite slag (SL_SP4Cs: A) and fly ash (FA_SP4Cs: B) under static conditions.

time, CLF = f(t0.5). The results of these operations are presented in Fig. 4(A2 and B2). They indicate a purely diffusive release mechanism during dynamic leaching. Both the slag and fly ash based samples, SL_SP6Cs and FA_SP6Cs, exhibited similar leaching behavior. Much higher cumulative leach fractions were observed compared with the CLF values obtained under static conditions. For the slag based sample, the CLF value reached 0.489 and for the SPC sample, with fly ash filler, up to 0.516. Detailed data for dynamic conditions are provided in Table 6. The results of the leaching tests performed under dynamic conditions differed significantly from those of the static tests, mainly due to the temperature of the leaching agent. Under static conditions, the samples and leachant were maintained at 21 ± 1 °C, but during dynamic tests the average temperatures were much higher: around 82 °C for the slag based SPC sample (SL_SP6Cs) and up to 87 °C for the fly ash concrete (FA_SP6Cs). These high temperatures were an effect of the energy delivered to the system, of energy withdrawn using a condenser and of heat exchange between the experimental setup and the environment, and occurred despite intensive cooling of the system. Temperature profiles, measured on the surfaces of the samples, are presented in Fig. 5(A and B). They show cyclic temperature variations during dynamic leaching experiments. For sample SL_SP6Cs, the leachant renewal period was found to be around 3310 s, whereas for fly ash filled concrete it was about 1520 s. Both values are only indicative, as the renewal period depended on many factors related to the experimental system. Although not the same, the static and dynamic leaching conditions were similar (except for the temperature and two times lower S/V ratio in case of leaching under dynamic conditions) and the results obtained during these experiments can be compared. The data presented in Table 6 and in Fig. 4(A2 and B2) allow for determination of the Cs-137 diffusion coefficients under dynamic leaching conditions. This time, the CLF vs. square root of time reveals a purely linear correlation, which suggests that leaching of cesium tracer occurred by a purely diffusive mechanism. In the case of leaching under static conditions, the initial tracer release significantly departs from the linear dependence of the CLF vs. square root of time. This observation can be explained by the rapid release of loosely bound and easily accessible cesium fractions when they are exposed to the leaching agent. In contrast, during dynamic leaching the partial activities of the Cs-137 tracer are proportional to the square root of time and the isotope release process is faster, which suggests a much more rapid diffusion rate for cesium in the SPC matrix. As a consequence, the whole data range may be used to calculate the diffusion coefficients. The values of the calculated effective diffusion coefficients for two selected SPC samples, SL_SP6Cs-DYN and FA_SP6Cs-DYN, are presented in Table 7. The values

Table 5 Normalized leaching rate (material flux) for SPC samples measured after 28* and 56** days of leaching under static conditions. Sample

Normalized leaching rate, LR g·cm−2·d−1 after 28 days

SL_SP4Cs SL_SP5Cs SL_SP6Cs SL_SP7Cs FA_SP4Cs FA_SP5Cs FA_SP6Cs FA_SP7Cs

(2.98 (2.09 (3.60 (1.74 (2.20 (1.88 (3.85 (3.24

± ± ± ± ± ± ± ±

0.08)·10−5 0.06)·10−5 0.09)·10−5 0.06)·10−5 0.06)·10−5 0.05)·10−5 0.08)·10−5 0.07)·10−5

after 56 days (2.06 (1.99 (3.16 (1.96 (1.91 (1.89 (3.39 (3.27

± ± ± ± ± ± ± ±

0.05)·10−5 0.05)·10−5 0.07)·10−5 0.04)·10−5 0.05)·10−5 0.05)·10−5 0.07)·10−5 0.07)·10−5

* samples collected between days 28 and 56 of leaching. ** samples collected between days 56 and 90 of leaching.

indicate that SP5 and SP7 binders are the best options for SPC preparation, taking as a criterion LR after 28 days of leaching. All these considerations lead to the conclusion that the best immobilization efficiency may be observed in the case of dicyclopentadiene based sulfur polymer binders, but only when the co-monomer is styrene or furfural. Concretes based on SP4 (with only dicyclopentadiene as the comonomer) as well as SP6 (with dicyclopentadiene and decene as comonomers) as binders exhibited worse leaching characteristics under all other circumstances, except those discussed above. Nonetheless, it should be stressed that globally all the investigated sulfur polymer concretes differed only slightly and were comparable in terms of their immobilization efficiency towards Cs-137 isotope. Neither the filler (slag or fly ash) nor the sulfur polymer binder differentiated the SPC samples significantly in terms of Cs-137 tracer release. 3.2. Dynamic leaching of Cs-137 and release from SPC concretes Cesium is a difficult radionuclide to stabilize (Ojovan and Lee, 2014). To study its behavior in a harsher environment, leaching was carried out under dynamic conditions. The experimental system consisted of a reflux condenser, which allows for cyclic renewal of the leachant (deionized water), so that the sample was continuously in contact with fresh portions of the leaching agent. Two SPC samples were selected for testing under dynamic conditions. The results are presented in Fig. 4(A1 and B1). Analysis required linearization of the CLF values vs. time. According to the analytical model described by Eq. (3) one can obtain a very well fitted dependence for the cumulative leach fraction vs. square root of 6

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Fig. 4. Dynamic leaching kinetics for Cs-137 tracer release from concretes based on lignite slag (SL_SP6Cs: A1-2) and fly ash (FA_SP6Cs: B1-2).

shown are four orders of magnitude higher compared with the diffusion coefficients obtained for samples leached under static conditions. For SL_SP6Cs-DYN and FA_SLSP6Cs-DYN, the effective diffusion coefficients were found to be 1.40·10−5 and 1.54·10−5 cm2·s−1, respectively, whereas for the same samples leached under static conditions the values were 2.66·10−9 and 3.10·10−9 cm2·s−1, respectively (cf. Table 4). The suitability of the waste matrix as well as the safety and immobilization efficiency of the Cs-137 isotope in the SPC matrix were evaluated taking into account material flux from the matrix volume. To estimate the material flux, normalized leaching rates for samples subjected to dynamic leaching were calculated using Eq. (5). Data obtained for the previously selected samples, SL_SP6Cs-DYN and FA_SP6Cs-DYN, are presented in Fig. 6(A and B). Both leaching rate profiles present similar behavior in terms of the LR parameter over time. As may be expected, due to the four orders of magnitude higher diffusion coefficients under dynamic conditions, the normalized leaching rates are much higher in comparison to those of samples leached during static tests. In dynamic conditions, the observed normalized leaching rate values for SL_SP6Cs-DYN and FA_SP6Cs-DYN were found to be 1.73·10−3 and 1.57·10−3 g·cm−2·d−1, respectively, whereas under static conditions the corresponding values were around 10−5 g·cm−2·d−1, two orders of magnitude lower. Detailed data for normalized leaching rates under the static and dynamic conditions with

calculation uncertainties are presented in Table 8. 3.3. Release of Cs-137 from SPC in aggressive environment (1.0 M HCl) Analyses of Cs-137 release were also performed on data obtained from leaching experiments carried out using an aggressive leachant (1.0 M HCl). The initial sampling intervals were reduced and leaching agent exchanges were additionally carried out 0.5, 2 and 8 h prior to the standard protocol. The data obtained are presented in graphical form in Fig. 7. In 1.0 M HCl solution, the cumulative leach fractions in both samples subjected to contact with the leachant reached very high values, approaching 1. This indicates much more efficient Cs-137 release mechanisms when the SPC matrix is exposed to acid attack. The exact values are included in Table 6, where CLF = 0.830 and 0.927 for slag and fly ash based concretes, respectively. Analysis of the kinetic profiles indicates a more complex release mechanism than in the cases of static or dynamic leaching. Such values may suggest a dissolution mechanism for Cs-137 release, so in order to verify this assumption simple SPC samples weight analysis was performed each time the leachant was replaced. After being removed from the leaching solution, the samples were gently dried using paper towel and then weighted. The weight changes during aggressive leaching were found to be of minor importance, with weight variations of up to +3% of the initial

Table 6 Partial and cumulative leach fractions of Cs-137 released from SPC concretes under dynamic and aggressive conditions. Sample

Fraction leached after time (days), ai/Ao: 1

3

Dynamic leaching: SL_SP6Cs-DYN 0.079 0.052 FA_SP6Cs-DYN 0.075 0.061 Leaching under aggressive conditions (1.0 M HCl) SL_SP6Cs-HCl 0.084 0.064 FA_SP6Cs-HCl 0.108 0.068

Cumulative Leach Fraction, CLF = Σai/Ao

7

14

28

56

90

0.052 0.060

0.059 0.077

0.094 0.081

0.153 0.161

n.a. n.a.

0.489 0.516

0.085 0.086

0.163 0.157

0.150 0.168

0.196 0.248

0.088 0.092

0.830 0.927

7

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Fig. 5. Temperature profiles during dynamic leaching of the cesium tracer from concretes based on lignite slag (SL_SP6Cs: A) and fly ash (FA_SP6Cs: B). Table 7 Effective diffusion coefficients in SPC samples measured for isotope Cs-137 released under dynamic and aggressive conditions. Sample

Table 8 Normalized leaching rate (material flux) for SPC samples measured after 28* and 56** days of leaching under dynamic and aggressive conditions.

Effective diffusion coefficient, De, cm2·s−1

Dynamic leaching: SL_SP6Cs-DYN (1.40 FA_SP6Cs-DYN (1.54 Leaching under aggressive conditions (1.0 M HCl) SL_SP6Cs-HCl (8.63 FA_SP6Cs-HCl (1.26

Normalized leaching rate, LR g·cm−2·d−1

Sample

after 28 days

± 0.09)·10−5 ± 0.10)·10−5

after 56 days

Dynamic leaching: SL_SP6Cs-DYN (1.73 ± 0.03)·10−3 FA_SP6Cs-DYN (1.57 ± 0.03)·10−3 Leaching under aggressive conditions (1.0 M HCl) SL_SP6Cs-HCl (1.82 ± 0.04)·10−3 FA_SP6Cs-HCl (2.36 ± 0.07)·10−3

± 0.38)·10−6 ± 0.03)·10−5

mass in the case of the slag filled sample (this increase in weight was probably due to penetration of the leaching solution inside the sample), and up to −2% of the initial mass in the case of the fly ash based concrete. The slag based samples were of higher porosity in comparison with the fly ash based SPC, which could result in less efficient removal of the absorbed liquid phase. When CLF exceeds significantly 0.5, the semi-infinite model described by Eq. (3) cannot be used. Instead, one should apply a model based on Eq. (6) (Crank, 1975).

n.a. n.a. (6.78 ± 0.14)·10−4 (7.24 ± 0.15)·10−4

* samples collected between days 28 and 56 of leaching. ** samples collected between days 56 and 90 of leaching.

CLF =

ai =1 Ao

8 2

De 2t S 4 V

exp

2

(6)

Linearization of the model leads to the expression described by Eq. (7).

ln

2

8

1

ai Ao

=

De 4

2

S V

2

t

Fig. 6. Normalized leaching rates of Cs-137 tracer in dynamic conditions for concretes based on lignite slag (SL_SP6Cs: A) and fly ash (FA_SP6Cs: B). 8

(7)

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Fig. 7. Leaching kinetics under 1.0 M HCl conditions for Cs-137 tracer release from concretes based on lignite slag (SL_SP6Cs: A1-2) and fly ash (FA_SP6Cs: B1-2).

When samples are leached under aggressive conditions, one cannot assume a purely diffusive transport mechanism for the cesium tracer. The role of a dissolution mechanism should also be considered. In spite of the possible multi pathway release mechanism, the data presented in Fig. 7(A2 and B2) indicate good approximation of the experimental points by the diffusive model described by Eq. (7). For both concrete samples, a very good correlation of ln[π2(1-Σai/Ao)/8] vs. time is observed (goodness of fit in Pearson's test > 0.995). The diffusion coefficients calculated according to Eq. 7 are presented in Table 7 and were found to be lower than those for the samples leached under dynamic conditions. Final evaluation of cesium immobilization efficiency was performed by calculating the material release rate from samples in contact with HCl solution. The dependencies of the material flux vs. leaching times were calculated as the normalized leaching rate values. The experimental data obtained are presented in Fig. 8(A and B) for SL_SP6Cs

(HCl) and FA_SP6Cs (HCl) concretes, respectively. Contrary to the results presented previously in Figs. 3 and 6, a decreasing tendency is observed for the normalized leaching rate vs. time of the concretes leached in HCl solution. For the samples leached under static and dynamic conditions, the data suggest stabilization of the LR parameter over time. Such behavior is presumably a result of depletion of the cesium tracer within the matrix body. This explanation is supported by the fact that the CLF for both acid leached samples reached values close to 1. However, in the case of slag based SPC concretes, a continuously decreasing tendency may also be observed for the leaching rate (Fig. 3A), which can be the result of increased porosity (in comparison with fly ash based SPC) and more facile leachant exchange between the inner pore volume and the external volume of the leaching solution. Potentially, the higher leaching rate and easier Cs-137 release in the HCl solution treatment procedure could allow for the recovery of cesium isotopes from the waste matrix, when transmutation technology

Fig. 8. Normalized leaching rates of Cs137 tracer under aggressive conditions (1.0 M HCl) for SPC based on lignite slag (SL_SP6Cs: A) and fly ash (FA_SP6Cs: B). 9

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becomes standard procedure for the neutralization of radioactive isotopes stabilized in SPC wasteforms deposited at interim storage sites (Takao et al., 2019).

Table 9 Phase transitions of SP4-SP7 sulfur polymer binders determined by DSC method.

3.4. Cs-137 immobilization efficiency Activation enthalpy of diffusion is a useful quantity, providing insight into the diffusion mechanism, based on the temperature dependence of the diffusion coefficient (D) vs. temperature (T), as expressed by Eq. (8),

DT = Do e

Ha RT

T1 T2 D Rln 1 T1 T2 D2

Temperature range, °C

Average melting temperature, °C

SP4 SP5 SP6 SP7

90 82 90 85

113.2 103.6 112.8 100.6

– – – –

125 120 125 120

It has been reported that self-diffusion coefficients for lanthanides increase significantly, by 2 – 3 orders of magnitude when shifting to temperatures close to the melting phase transition (Tiwari and Mehrotra, 2008). In the same paper, the authors report increases in ionic conductivity in alkali metal halides (Li, Na, K) of 2 – 4 orders of magnitude, when the melting temperature is approached, and an even greater change (7 orders of magnitude) in the self-diffusion coefficient in the case of transition metals when the temperature is increased to close to the melting transition. Although in the case of SPC only the sulfur binder phase was susceptible to softening and melting, the accumulation of cesium ions on the surface of the filler favored its release at elevated temperatures and accelerated transport outside the boundaries of the solid wasteform. All these findings suggest that SPC is suitable for limited applications, in a maximum temperatures range of 10 – 15 degrees below the beginning of melting transition temperature region. This is lower than 100 °C given by Mayberry (Mayberry et al., 1993a) and even lower than the 88 °C suggested by McBee and Weber (1990). Moreover, it reveals that accelerated transport of the cesium radionuclide (and probably also of other isotopes) begins before the melting phase transition temperature region.

(8)

where DT is the diffusion coefficient, in m2·s−1, at temperature T, in K, Do is the temperature independent pre-exponential factor, in m2·s−1, ΔHa is the activation enthalpy of diffusion, in J·mol−1 and R is the universal gas constant (R = 8.314 J·mol−1 K−1) (Mehrer, 2007). For a given set of samples, one can easily estimate the activation enthalpy of diffusion for the Cs-137 tracer immobilized in sulfur polymer concrete matrix from two independent temperatures, according to Eq. (9),

Ha =

SPB

(9)

where D1 and D2 are diffusion coefficients at temperatures T1 and T2, respectively. For SL_SP6Cs concrete, substitution D1 and D2 (cf. Tables 4 and 7) and T1 = 294.0 K (21.0 °C in static leaching) and T2 = 354.9 K (81.9 °C, dynamic leaching) leads to a value for activation enthalpy of diffusion ΔHa = 122 kJ·mol−1. The same procedure gives a slightly lower value ΔHa = 114 kJ·mol−1 for Cs-137 immobilized in the fly ashfilled sample (FA_SP6Cs). Both values are slightly above 1 eV per Cs+ ion immobilized in SPC matrix (1 eV per atom = 96.47 kJ·mol−1). Such high values for activation enthalpy of diffusion are observed only in the case of the so-called configurational diffusion mechanism, which can take place in nanoporous and zeolitic materials (Hensen et al., 2008; Inglezakis and Zorpas, 2012). It is unlikely that the investigated SPC samples, based on slag and fly ash fillers, would have exhibited such properties. In the literature, data for enthalpy of diffusion of Cs+ ions in a simple aqueous solution are in the range of 16.8 – 21.0 kJ·mol−1 and in Ordinary Portland Cement matrix 12 – 14 kcal·mol−1 (50.3 – 58.7 kJ·mol−1) (Papadokostaki and Savidou, 2009). Since the figures we obtained are significantly higher, one should consider possible reasons. Shifting the leaching temperature from 21 °C up to between 80 and 90 °C, increased the diffusion coefficients by four orders of magnitude, in the case of both FA and SL based SPC (cf. Tables 4 and 7). These results indicate dramatic changes regarding cesium immobilization conditions. The measured diffusion coefficients, in the order of 10−5 cm2·s−1 (1.40·10−5 and 1.54·10−5 cm2·s−1 for SL and FA based concretes, respectively) are only slightly lower and of the same order of magnitude as those measured for Cs+ ions in aqueous solutions or gels, usually between 1.7·10−5 – 2.2·10−5 cm2·s−1 (Sato et al., 1996; Chakrabarti and Kanjilal, 2010; Gokarn and Rajurkar, 2006). This proves that the SPC matrix state approaches liquid or semi-liquid conditions. To verify this hypothesis by determining the phase transitions in the SPB binders used (SP4-SP7), DSC measurements were performed. The results are presented in Table 9. If the dynamic leaching temperature is compared with the melting phase transition temperature region for SP6 sulfur polymer binder (90 – 125 °C), it can easily be seen that the dynamic leaching experiments were performed close to the melting range of the SP6 binder. This proximity favors softening of the matrix and explains both the sudden change in the calculated values for the diffusion coefficients and the abnormal values found for enthalpy of diffusion. These assumptions are confirmed by available data regarding diffusion phenomena studied at close to melting phase transition in solids.

3.5. Mechanism of Cs-137 immobilization under static, dynamic and aggressive conditions Our results show that the investigated matrices are effective for Cs137 immobilization under normal exploitation conditions, in terms of cesium leach rates, which are similar or even better than those found previously by van Dalen and Rijpkema (1989). In their study, the authors report cesium diffusion coefficients of between 2·10−10 and 3·10−9 cm2·s−1 (our results are between 8.4·10−10 – 3.1·10−9 cm2·s−1, cf. Table 4), and cesium normalized leaching rates of between 8·10−4 and 1·10−3 g·cm−2·d−1 (our results after 28 days of leaching are in the range of 1.74·10−5 – 3.85·10−5 g·cm−2·d−1, cf. Table 5). However, at increased temperatures, as well as during leaching in an aggressive environment, the immobilization efficiency towards Cs+ ions drops significantly and the Cs-137 release rate speeds up. To understand the mechanisms behind these observations, one must take into account both the interactions between Cs+ ions and the fly ash or slag fillers and the interactions between the sulfur polymer binder and the fillers under various chemical and physical conditions. The surface complexation of cesium based on chemical and electrostatic interactions should be considered, as the first immobilization step. Numerous papers indicate that aluminosilicate and silicate phases play a crucial role in the retention of Cs+ ions (Fernandez-Jimenez et al., 2005; Shi and Fernández-Jiménez, 2006; Jang et al., 2017; Jiang et al., 2017; Fujii et al., 2018; Khandaker et al., 2018). These materials can be improved by the process of alkaline activation, which is responsible for the increased fractions of aluminosilicate phases in fly ash and slag derived products (Alby et al., 2018; Khandaker et al., 2018; Duxson et al., 2005; Sakthivel et al., 2013). The presence of aluminate and silicate phases in SL and FA fillers has been confirmed by XRF and SEM-EDS measurements. The respective contributions of the silicate and aluminate phases have been established as 68.2 and 7.9% in slag and 61.0 and 25.7% in fly ash (Szajerski et al., 2019a). The presence of 10

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these phases in the fillers has also been confirmed by XRD analysis (Szajerski et al., 2019a). Similar observations are reported in many other papers (Hosseini Asl et al., 2019; Khandaker et al., 2018; Karayannis et al., 2017; Rungchet et al., 2017; Bicer, 2018; Guo et al., 2018; Wu et al., 2018; Fujiwara et al., 2017; Vereshchagina et al., 2018). Efficient binding of cesium has been observed under many circumstances, in which various compositions of fly ash and slag derived materials (ceramics, geopolymers and Portland cement based concretes) (Hoyle and Grutzeck, 1989; Jang et al., 2016; Yang et al., 2017a; Mon et al., 2005; Lieberman et al., 2015; Goñi et al., 2006; Vereshchagina et al., 2013; Yang et al., 2017b; Duy Quang et al., 2018) and generally materials containing silicate and aluminosilicate phases have been investigated as potential radioactive waste immobilization matrices (Chen et al., 2018; Kurihara et al., 2018; Montagna et al., 2011; Jing et al., 2018). Further confirmation of the high efficacy of the two-stage immobilization approach proposed in this work can be found in a report by the Japanese Atomic Energy Agency, in which the authors describe a procedure for the immobilization of lead in SPC (Sone et al., 2008). Although the results for lead powder based SPC were better than those for cement based matrices, the method was not found to be very effective for the stabilization of lead oxide and liquid radioactive waste residues. The authors used lead powder, lead oxide and solid residues obtained from low-level liquid waste processing (containing significant amounts of nitrates), without stabilization before their incorporation into the SPC matrix. Probably for this reason, and also because of the much higher solubility of lead oxide and especially of nitrate-bearing liquid radioactive waste residues in comparison with metallic lead powder, the results proved unsatisfactory. The second stage of our procedure, in which the filler is stabilized using sulfur polymer binder, is basically the same as the methods proposed in the Japanese study and other reports. Finally, together with the adsorption of Cs+, the synergistic effect of the application of the liquid sulfur polymer must be considered as an important factor promoting immobilization. Low porosity must also result in a significant reduction in the specific surface area of the wasteform during contact with the leaching agent (Szajerski et al., 2019b; López et al., 2011). Modified sulfur is able to form a tight, dense coating on the filling phases. This is due to the hot mixing process, during which the liquid binder phase acts as an agglutinating and encapsulating agent. Similar mechanisms have been proposed in various previous works (López et al., 2011; Mohamed and El Gamal, 2009, 2011).

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4. Conclusions The purpose of this work was to investigate the immobilization of cesium isotope in sulfur polymer concrete matrices based on lignite slag and fly ash mineral fillers. The results indicate very good immobilization efficiency toward cesium under normal exploitation conditions. However, in accidental conditions (increased temperature, aggressive leaching agent) immobilization efficiency drops significantly in comparison with standard leaching procedures. The effective diffusion coefficients for the Cs-137 isotopes were between 0.84·10−9 and 3.10·10−9 cm2·s−1, with normalized leaching rates after 28 days of leaching in the range of 1.74·10−5 – 3.85·10−5 g·cm−2·d−1. The immobilization mechanism seems to involve the synergistic effects of cesium adsorption on the filler surface, in the form of silicates and aluminosilicates and sealing of the SPC matrix by the sulfur polymer binder. Although the proposed approach allows for effective immobilization of cesium and provides a stable final wasteform under normal conditions, during accidental situations immobilization efficiency drops significantly as a result of softening of the matrix or a combined digestion/dissolution mechanism, leading to a higher release rate of the Cs-137 isotope. 11

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