Competitive effect of the metallic canister and clay barrier on the sorption of Eu3+ under subcritical conditions

1 Competitive effect of the metallic canister and clay barrier on 1 the sorption of Eu3+ under subcritical conditions 2 Said El Mrabet1, Miguel A. Castro1, Santiago Hurtado2, M. Mar Orta1, 3 M. Carolina Pazos1, María Villa-Alfageme 3 and María D. Alba1† 4 5 1Instituto Ciencia de Materiales de Sevilla (CSIC-Universidad de Sevilla). 6 Avda. Américo Vespucio, 49. 41092 Sevilla, Spain 7 2Servicio de Radioisótopo del CITIUS (Universidad de Sevilla). 8 Avda. Reina Mercedes, 4. 41012 Sevilla, Spain 9 3 Dpt. Física Aplicada II (Universidad de Sevilla). 10 Avda. Reina Mercedes, s/n. 41012 Sevilla, Spain 11 12 Abstract 13 An in depth knowledge and understanding of high activity radionuclide (HLRW) 14 immobilization processes on the materials composing the engineered barrier (clay and 15 metallic canister) is required to ensure the safety and the long-term performance of 16 radioactive waste disposal procedures. Therefore, the aim of this study was to 17 understand the mechanisms involved in the retention of Eu3+ by two components of the 18 multibarrier system, the bentonite barrier and the canister. As such, a comparative study 19 of the interaction of trivalent Eu3+, used to simulate trivalent actinides, with both 20 bentonite and a metallic canister has been undertaken in this work. To this end, we 21 designed a minireactor into which the bentonite was introduced and compacted. The 22 minireactor-bentonite system was then submitted to a hydrothermal reaction with a 23 7.9·10-2 M solution of Eu3+ at 300ºC for 4.5 days. SEM and XRD results revealed that 24 † Corresponding author: E-mail: [email protected] 2 both bentonite and the container were involved in the immobilization of europium by 1 the formation of insoluble europium silicate phases. The presence of europium silicate 2 adsorbed on the surface of the metallic canister indicates the competitive effect of both 3 components of the engineered barrier (bentonite and metallic canister) in HLRW 4 immobilization. These results suggested that the canister could play a role in the HLRW 5 immobilization even during its corrosion process. 6 Keywords. Metallic canister, bentonite, engineering barrier, radionuclide waste, 7 actinide, lanthanide, rare earth elements, silicates, hydrothermal treatment. 8 9 10 3 1. Introduction 1 The complexity of radioactive waste management is one of the most significant 2 challenges faced by all countries involved in nuclear power generation. Likewise, it is 3 also a concern in many other countries that use radioactive materials for medical, 4 industrial, or research purposes. Much of this waste needs to be disposed of in ways that 5 minimise its negative impact. In particular, the most hazardous and long-lived waste 6 such as spent nuclear fuel or reprocessing waste arising from the operation and 7 dismantling of nuclear reactors must be contained and isolated in order to ensure the 8 long-term radiological protection of humans and the environment. Disposal of such 9 waste in deep underground-geological repositories has become an effective and 10 universally preferred option (Savage and Chapman, 1982; Allègre, 1999; McCombie et 11 al., 2000). However, such disposal requires the design of an isolation system consisting 12 of several barriers that are able to confine the radioactivity for a long period of time. 13 Most such repository designs are based on a multi-barrier system consisting of two 14 fundamental components, namely a natural and an engineered barrier (McCombie et al., 15 2000; Astudillo, 2001; Chapman, 2006). The engineered barrier includes a metallic 16 container, whose main function is to confine the radionuclides in the waste package 17 during the specified period and a backfill material, made up mainly of bentonite, that 18 stands between the container and host rocks in order to prevent the access of 19 groundwater to the HLRW waste and its subsequent migration to the geosphere (Bailey, 20 1980). 21 To improve the safety of the disposal system, each of those barriers alone should 22 be able to protect the repository systems from water intrusion over its design lifetime 23 and prevent the spread of hazardous radionuclides from the repository to the biosphere. 24 4 As such, container corrosion and the chemical interactions of corrosion products with 1 the clay buffer are of great interest for the long-term performance of a repository 2 (Wilson et al. 2006a and 2006b). After closure of the disposal repository, the low 3 permeability of the clay surrounding the container means that it progressively becomes 4 saturated with groundwater. The expected environment then becomes oxidizing due to 5 the presence of dissolved oxygen in the water and air trapped in the near field (Spink 6 and Wood, 1990). Once the oxygen trapped in the repository has been consumed 7 entirely, the canister undergoes anaerobic corrosion, thus resulting in the generation of 8 hydrogen and iron-corrosion products from the canister such as magnetite (Fe3O4), and 9 goethite (FeO(OH)) (Smart et al.2002; Carlson et al. 2007). Siderite (FeCO3) is also 10 formed under the expected geological conditions for groundwater with high 11 concentrations of CO32- ions and in the presence of compacted clay; iron silicates such 12 as berthierine may also be important (Ishidera et al. 2008; Savage et al., 2010). 13 Ultimately, the canister may be perforated as a result of such corrosion processes, thus 14 allowing radionuclide migration through the clay buffer via diffusion and retardation by 15 sorption within the clay (Xia et al., 2005). Most of the iron diffuses from the container 16 into the surrounding barrier (bentonite), where it precipitates and subsequently induces 17 deterioration of the buffer barrier, a process that may result in a loss of some crucial 18 properties of bentonite such as swelling capacities and sorption. Carlson et al. (2007) 19 have reported increased hydraulic conductivity due to interaction with steel corrosion 20 products. 21 Several studies concerning the corrosion of candidate metals for the container, as 22 well as the interaction of their corrosion products with bentonite, have been reported in 23 the literature (Guillaume et al. 2004; Perronnet et al. 2004; Chuanhe et al. 2011). 24 Likewise, systematic destabilization of the smectite to non-swelling sheet silicates when 25 5 in contact with iron metal and/or iron oxides has been extensively investigated 1 (Guillaume et al. 2003, 2004, Perronnet et al. 2004, Lantenois et al. 2005; Wilson et al. 2 2006a and 2006b). Thus, Lantenois et al. (2003) carried out experiments with different 3 smectites in order to determine the effect of crystallo-chemical features on the smectite/ 4 iron interactions. At 80°C, the results indicated that oxidation of the container by 5 smectites occurs only for dioctahedral smectites under basic pH conditions, whereas the 6 container corrodes by precipitating magnetite, but without smectite alteration, at pHs 7 ranging from slightly acid to neutral. Iron/clay interactions have also been modelled, 8 and it has been shown that most of the Fe diffuses from the canister into the bentonite, 9 where it sorbs or precipitates as magnetite (Samper et al. 2008). Detailed mineralogical 10 analysis of corrosion experiments by Milodowski et al. (2009a; 2009b) tends to suggest 11 other solid products. 12 As far as steel-bentonite interactions are concerned, the transformation of 13 bentonite into other minerals, as a result of its interaction with the corrosion products of 14 metallic container, has been observed (Julien et al 2005, Bildstein et al. 2006). These 15 corrosion products, together with the high pH value produced at the container-clay 16 interface and in the host rock, result in smectite destabilization and precipitation of 17 magnetite, hematite, goethite and serpentine-like minerals, thus resulting in the loss of 18 some of the bentonite properties, such as its swelling and cation-exchange capacity 19 (Bildstein et al., 2006; Carlson et al., 2007; Gaudin et al., 2009; Savage et al., 2010). 20 The effect of container corrosion on the stability of clay mineral therefore depends 21 mainly on factors such as pH, temperature, the crystal chemistry of the clay; the 22 water/solid ratio and the iron/clay mass (Gaudin et al., 2009). 23 6 Few studies of concerning the role of the backfill material in the kinetics and 1 mechanism of container corrosion have been conducted to date. An understanding of the 2 sorption/retention of radionuclides on the materials used to construct the engineered 3 barrier (clay and container waste) is therefore imperative in order to be able to 4 adequately assess the long-term performance of radioactive waste disposal facilities. 5 Likewise, very few studies have been carried out to evaluate the active participation of 6 the metal container in the retention of radioactive waste. In this sense, Parfitt et al. 7 (1980) studied the adsorption of Eu3+ by the metallic container and showed that the 8 presence of free iron oxides inhibits adsorption of Eu3+ on the clay surface. Moreover, 9 the adsorption of Eu3+ in hematite was found to be inhibited at pH> 5 in the presence of 10 organic matter (Fairhurst et al. 1995, Rabung et al. 1998). Similarly, Quinn et al. (2006) 11 found that the adsorption constant of various rare earth elements on the surface of 12 hematite is related to the physicochemical properties of these rare earth cations. 13 Davranche et al. (2004) studied sorption of the entire rare earth element (REE) series 14 onto iron oxyhydroxide, in the presence of humic acid, the observation of Yttrium 15 (YREE) sorption pattern was interpreted in terms of complexation with humate that was 16 dissolved in solution and sorbed onto hematite (Fairhurst et al., 1995; Rabung et al., 17 1998b; Davranche et al., 2004). Philippini et al (2006) studied the sorption of silicon on 18 various corrosion products such as magnetite Fe3O4, goethite (-FeOOH), siderite 19 (FeCO3) and pyrite (FeS2) and found that only three of the four studied corrosion 20 products are able to sorb silicon but in different proportions whereas the sorption 21 capacity by pyrite was estimated negligible. 22 Several studies have been undertaken to determine the ability of magnetite, 23 commonly formed on corroding steel surfaces, to absorb or reduce some radionuclides 24 (Granizo and Missana, 2006; Rovira et al., 2004). Savage et al. (2010) observed that 25 7 Magnetite occurs as a transitory phase, which may explain the very thin layers of 1 magnetite in some recent experimental studies of steel corrosion in compacted 2 bentonite. 3 In light of the above, this paper aims to study the competitiveness of bentonite 4 and the metal canister for retaining radioactive actinides. In this work, 152Eu was 5 selected to simulate trivalent actinides, Bentonite FEBEX material, as simulator of the 6 materials of the engineered barrier and the austenitic stainless steel 316L as simulator of 7 the metallic canister material (J. M. Gras, 2002; ENRESA, 2004). 8 9 2. Methods 10 2.1. Experimental design and materials 11 The clay mineral used in this study (Bentonite FEBEX) has been extensively 12 investigated in the recent past in many countries in Europe and around the word 13 (Tripathy et al., 2004). This bentonite was provided by the ENRESA Company (the 14 Spanish Company in charge of radioactive wastes management) and had the structural 15 formula ((Ca0.5Na0.08K0.11)(Si7.78Al0.22)(Al2.78Fe3+0.33Fe2+0.02Mg0.81)O20(OH)4.). Its main 16 mineralogical component is montmorillonite (smectite percentage higher than 90%) 17 together with small amounts of quartz (Fernandez et al., 2004). 18 Europium nitrate Eu(NO3)3·4.88H2O, which is commercially available from 19 Sigma-Aldrich, was used as the rare earth element source, as Eu3+ is commonly used as 20 a trivalent actinide analogue due to their chemical similarity (Chapman and Smellie, 21 1986; Buil et al., 2007). 22 8 A wide variety of metallic materials have been investigated as candidates for the 1 disposal of radioactive waste. Copper, titanium, stainless steels were found to be 2 suitable materials because they exhibit a high resistance attack in the expected disposal 3 environment (Rebak, 2006). Therefore, in the present study, hydrothermal experiments 4 were carried out in a stainless steel AISI-316 L reactor, (chosen as candidate container), 5 commercially available, the chemical composition is given in Table 1, and the design in 6 Fig. 1a. 7 In order to understand the competitive effect of the container material in the 8 processes by which the bentonite clay retains radioactive waste, a minireactor made 9 from the same material as the steel reactor was designed by us. Thus, 300 mg of the 10 powdered bentonite was placed into a cylindrical steel cell (minireactor) 8 mm in 11 diameter and 2 mm in height (Fig. 1a). The bentonite-minireactor set was then 12 compacted in a cylindrical die at a pressure of 4 tonnes for 5 min. Finally, the 13 compacted set was placed into the steel reactor and submitted to a hydrothermal 14 reaction with 40 mL of 7.910-2 M Eu3+ solution, this concentration allows that Eu3+ was 15 not a limiting reagent (Alba et al., 2011), at 300 ºC for 4.5 days. Despite, the expected 16 temperature in the disposal repositories which will not exceed 150ºC, many studies have 17 been carried out by simulating the deep geological disposal at temperatures up to 350ºC 18 to increase the reaction rate (Mather et al., 1982; Savage and Chapman., 1982; Allen 19 and Wood, 1988, Alba and Chain., 2007). Alba et al. (2009; 2011) observed that similar 20 chemical reaction between lanthanide cations and clays occurred at lowest temperature 21 but after longer reaction time. 22 To discriminate between the effect of the canister-bentonite competitiveness and 23 the effect of the hydrothermal treatment, a blank treatment was carried out. Thus, the 24 9 minireactor was submitted to a similar hydrothermal treatment but without bentonite 1 and europium solution. 2 2.2. Characterization methods 3 The X-ray diffraction patterns were obtained using an X'Pert Pro 4 PANALYTICAL diffractometer in the conventional θ − 2θ Bragg–Brentano 5 configuration using Cu Kα radiation. Diffractograms were registered from 3º to 70º 2θ 6 and in steps of 0.05º. The morphology and chemical composition of both the steel and 7 bentonite before and after hydrothermal treatment with the rare earth solution at 300ºC 8 for 4.5 days were investigated using a SEM-FEG HITACHI S- 4800 a scanning electron 9 microscope equipped with an Xflash 4010 (BRUKER) for energy dispersive X-ray 10 (EDX) analysis. The pH of the supernatant was measured at room temperature using a 11 Eutech Instruments PC 700 pH-meter before and after the hydrothermal treatment in 12 aerobic conditions. 13 14 3. Results and Disscussion 15 3.1. Sorption of Eu3+ on Bentonite FEBEX 16 Fig. 1 shows a comparative visual analysis of the system (minireactor and 17 bentonite) before and after the hydrothermal treatment. The visual analysis of the 18 bentonite after hydrothermal treatment at 300ºC for 4.5 days in the presence of 19 europium (Fig. 1b) showed the typical spongy appearance of clay. 20 Fig. 2 shows the XRD pattern of the bentonite before and after treatment with a 21 solution of 7.910-2 M Eu3+ at 300ºC for 4.5 days. The XRD pattern of untreated 22 16 Buil, B., Gomez, P., Garralon, A., Turrero, M.J., 2007. Rare-earth elements: a tool for 1 understanding the behavior of trivalent actinides in the geosphere. Mater. Res. Soc. 2 Symp. Proc., 985, 437-442. 3 Carlson, L., Karnland, O., Oversby, V.M., Rance, A.P., Smart, N.R., Snellman, M., 4 Vähänen, M., Werme, L.O., 2007. Experimental studies of the interactions between 5 anaerobically corroding iron and bentonite. Phys. Chem. Earth., 32, 334-345. 6 Chapman, N., 2006. 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Geotech., 41, 437-450. 16 Wilson, J., Savage, D., Cuadros, J., Shibata, M., Ragnarsdottir, K.V., 2006a. The efeect 17 of iron on montmorillonite stability. (I). Background and thermodynamic 18 considerations. Geochim. Cosmochim. Acta, 70, 306-322. 19 Wilson, J., Cressey, G., Cressey, B., Cuadros, J., Ragnarsdottir, K.V., Savage, D., 20 Shibata, M., 2006b. The efeect of iron on montmorillonite stability. (II). 21 Experimental investigation. Geochim. Cosmochim. Acta, 70, 323-336. 22 Xia, X., Idemitsu, K., Arima, T., Inagaki, Y., Ishidera, T., Kurosawa, S., Iijima K., Sato, 23 H., 2005. Corrosion of carbon steel in compacted bentonite and its effect on 24 neptunium diffusion under reducing condition. Applied Clay Science, 28, 89-100. 25 26 21 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 Table 1. Chemical composition (% by weight) of the stainless steel AISI 316 L used in this work Co V Si S P Mn Cr Fe Ni Cu Mo Cl 0.14 0.11 0.38 0.03 0.04 1.74 16.53 68.29 10.57 0.29 1.87 0.01 Table.2. Values of pH, redox potential (Eh) and conductivity of the initial aqueous solution and the solution after hydrothermal reaction at 300ºC for 4.5 days of FEBEX in contact with a solution of 7.9·10-2M de Eu3+. solution pH Eh (mV) Conductivity (mS.cm-1) initial 4.72 605 10.89 final 1.96 839 24.1 22 FIGURE CAPTIONS 1 Fig. 1. Illustrative picture of: a) the reactor and the minireactor used for the 2 hydrothermal treatments, and, b) the bentonite FEBEX and the minireactor after being 3 treated hydrothermally at 300ºC for 4.5 days in contact with a solution of 7.910-2 M 4 Eu3+ . 5 Fig. 2. XRD diffraction patterns of the bentonite FEBEX: a) untreated bentonite, and, b) 6 after being treated at 300ºC for 4.5 days in contact with a solution of 7.910-2 M Eu3+. 7 Q= quartz (PDF 00-01-649). 8 Fig. 3. a) SEM micrograph of the bentonite FEBEX. b) General view of the treated 9 bentonite at 300ºC for 4.5 days in contact with a solution of 7.910-2 M Eu3+. c) and d) 10 other zones from b) where small particles as well as block morphology are shown. Dots 11 indicates the places where EDX analysis was performed 12 Fig. 4. EDX spectra of the different zones indicated in Fig. 3.d. EDX spectrum of the 13 untreated bentonite FEBEX was included as reference. 14 Fig. 5. XRD diffraction patterns of the minireactor: a) after the blank hydrothermal 15 treatment. b) after hydrothermal reaction with bentonite at 300ºC for 4.5 days in contact 16 with a solution of 7.910-2 M Eu3+. A= austenite (PDF 00-33-397), *= Eu2Si2O7 (PDF 17 00-23-247), += Eu3SiO5 (PDF 00-20-403), h= halloysite (PDF 00-02-229), n=β-18 NaAl7O11 (PDF 00-21-1095), and, j=jadeite (PDF 00-22-1338). 19 Fig. 6. SEM micrographs of the minireactor: a) as-made, b) after the blank 20 hydrothermal treatment and, c-f) after hydrothermal reaction at 300ºC for 4.5 days with 21 23 bentonite FEBEX and a solution of 7.910-2 M Eu3+. Numbering levels indicate the 1 particles where the EDX analysis was performed. 2 Fig. 7. EDX spectra of: a) as-made, b) after the blank hydrothermal treatment, and, c-e) 3 after hydrothermal reaction at 300ºC for 4.5 days with bentonite FEBEX and a solution 4 of 7.910-2 M Eu3+; zones 1, 2 and 3, respectively, indicated in Fig. 6. 5 Fig. 8. a) SEM micrograph of a transverse section of the minireactor after the reaction 6 of the bentonite FEBEX with Eu3+ at 300ºC for 4.5 days. b) Intensity profile of the 7 elemental composition. 8 Fig. 9. pH-Redox potential (Eh) plot (Pourbaix diagrams) of the initial 7.910-2 M Eu3+ 9 solution (circle) and the supernatant recovered after hydrothermal treatment (triangle). 10 11 24 Fig. 1 1 2 3 4 5 Minireactor Bentonite Minireactor Reactor a b 25 1 Fig. 2 2 3 4 5 10152025303540455055606570 Q 2  Q x2 x5 [60] [001] 1.39 nm 1.51 nm a b 32 1 Fig. 9 2 3 4 5 6 7 8 9 10 11 12 0 2 4 6 8 10 12 14 -1500 -1200 -900 -600 -300 0 300 600 900 1200 1500 Eu(OH)3 Eu2+ Eh(mV) pH Eu3+