Introduction

90Sr, the β-emitter radionuclide with a half-life of 28.8 years, average biological half-life ∼18 years and decay energy 546 KeV, is the foremost isotope of strontium (Alfredo et al. 2014; Glasstone and Dolan 1977; Guimaraes et al. 2014; IAEA 2006; Keceli 2015; Smiciklas et al. 2015). The significant amounts of 90Sr radionuclides produced from nuclear reactors and/or fallout from weapon testing remained in the spent fuels were identified as a serious threat to the sustainable world (IAEA 2006; Mckinley et al. 2007; Zhijian 2008). The mobility of radiation emitted by nuclear wastes (containing strontium isotopes) can only be prevented by their decay and decontamination. For this, deep geological disposal has remained in-practice wherein the radioactive wastes are kept prolonged in isolation in the multi-barrier repository (IAEA 2006; Zhijian 2008).

In such repositories, bentonite has remained a prime candidate buffer material, often surrounded by the packaging material either of concrete at low pH < 4 or cement at high pH > 10 (IAEA 2006; Kwon et al. 2013; Mckinley et al. 2007; Till and Grogan 2008; Zhijian 2008). Owing to the exceptional characteristics of high swelling potential, low permeability, self-sealing capacity, higher cation exchange capacity, specific surface area and active mineralogy, the bentonite exhibits high sorption capacity (commonly determined as a distribution coefficient, k d) to facilitate the proper functioning of waste repository (Caglar et al. 2009; Mckinley et al. 2007; Zhijian 2008). The development of engineered multi-barrier waste repositories was pioneered albeit the adversity caused by natural and/or manmade intrusion, due to potential leakage in waste packaging, fire/explosion by the excessive pressure build-up inside the container, which would result in damage to the repository (IAEA 2006; Mckinley et al. 2007). This situation may lead to direct contact between the bentonite buffer material and acidic/alkaline pore fluid of radioactive wastes. In the past, fluid leakage from repositories had led to severe 90Sr contamination in soils/sediments/groundwater at a number of sites worldwide, including Hanford, New Mexico and Oak-Ridge (USA), Mayak in Russia and Fukushima in Japan (Kamei-Ishikawa et al. 2013; Keceli 2015; Wallace et al. 2012). Moreover, the solid strontium gets solubilized in water and may cause a serious threat to human health. Therefore, its limit in drinking water has been restricted to 1.5 ppm with an average annual drinking water limit of 182.8 mCi 90Sr (IAEA 2006; Gllowiak et al. 1977). Exposures above the permissible limit have chronic effects on the human body which can lead to somatic and genetic changes, and further develop several types of cancer in the bone, nose, lungs and skin (IAEA 2006; Smiciklas et al. 2015).

Therefore, with a necessity to understand the possible failure of a multi-barrier repository system, the relative changes in bentonite properties (before and after strontium contamination) should be investigated. Though the previous studies revealed that the longevity of bentonite buffer material primarily depends on the mineral contents, pH of pore solution and the temperature generated within the repository (IAEA 2006; Guimaraes et al. 2014; Guven 1990), it was observed that sorption properties of bentonite-contaminant interaction were substantially influenced by the physicochemical properties (i.e. cation exchange capacity, CEC; specific surface area, SSA; pH), the contaminant properties (i.e. pH, type and concentration of ionic species) and the mineralogy of bentonite (\Arnepalli et al. 2010; Guven 1990; Karnland et al. 2007; Pathak 2014; Pathak et al., 2014a, b, 2015). A greater influence on sorption onto bentonite was observed when the isomorphic substitution occurred in tetrahedral sheets by replacing tetravalent with trivalent ions and in octahedral sheets by the replacement of trivalents with divalent ions (He et al. 2016; Seliman et al. 2014); whereas, the nano-porous iron oxyhydroxides of large surface area (Srivastava et al., 2012, 2013) in bentonite often supports the phenomena for outer-sphere cation exchange (Wallace et al. 2012). Notably, previous studies on strontium sorption dealt with virgin bentonite and none of them determined the subsequent effect of contamination onto the contaminated bentonite soil (CBT). Therefore, such study would be helpful to understand the changes occurring in sorption properties of the bentonite buffer material in repositories, and sorption reversibility of the contaminated bentonite at the time of adversity. The need of this investigation can be underlined as the 90Sr in buffer material forms weakly bound surface complexes, which may be remobilized to a substantial extent if the ionic strength of groundwater increases or, with any saline intrusion.

Moreover, the present piece of work would be more relevant in a current scenario of India’s civil nuclear energy program with a flourishing target of 24 GW by 2030 from the present capacity of ∼6 GW (India Energy Outlook 2015). This would unprecedentedly generate the large amount of radionuclide wastes to fulfil the clean energy requirements for the rapid industrialization and economic growth of India. Hence, the search for suitable buffer (bentonite) material is ongoing, and one such plausible sample (WIn-BT) was used in this study. The present paper deals with a systematic study on the significant changes occurring in bentonite properties due to the acidic and alkaline attack of strontium.

Experimental investigation

A commercially available bentonite, WIn-BT, collected from the western part of India was used in this study. To simulate the contamination onto bentonite buffer material in laboratory, the WIn-BT was exposed to a model contaminant (in non-radioactive form) solution of 0.1, 0.01 and 0.001 M SrCl2. At a liquid-to-solid ratio, L/S = 5 where 20 g soil mixed with 100 ml of solution and stirred for 1 h, the pH (1, 4, 10, 13) of the uniform slurry was separately maintained by the addition of H2SO4/NaOH solution under stirring; the soil pH (without any acid/alkali addition) was nearly 8.5 (±0.2). After 1 h of mixing, the slurries were kept in closed containers for 30 days to dry. For an uncontaminated reference buffer material, WIn-BT was also mixed with distilled water and kept for 30 days. In this way, 15 types of contaminated bentonite soils designated as CBT and one uncontaminated bentonite soil designated as UCBT were prepared in the laboratory. The sorption reversibility of UCBT and CBT were subsequently studied in the laboratory by considering a similar contaminant (SrCl2) at an initial pH of 6.5 for batch tests (Pathak et al., 2014a, b, 2015). A wide range of concentrations C i = 0.7, 0.5, 0.4, 0.3, 0.2, 0.1, 5 × 10−2, 1 × 10−2, 5 × 10−3, 1 × 10−3 and 1 × 10−4 M of SrCl2 were selected, and the k d values were computed by employing the Freundlich isotherm (Pathak 2014; Pathak et al., 2014a, b, 2015), which is used to describe the heterogeneous system (He et al., 2016), as presented in Eq. 1.

$$ {C}_{\mathrm{s}}={k}_{\mathrm{d}}\cdot \sqrt[ n]{C_{\mathrm{e}}} $$
(1)

where k d is the distribution coefficient (in g1–1/n L1/nkg−1), n is a constant accounted for soil heterogeneity and C e (in g/L) and C s (in g/kg) are the concentrations of the contaminant in solution at equilibrium and sorbed onto soil. The k d and 1/n can be calculated from the intercept and slope of plot C s versus C e (discussed in the Results and discussions section). Notably, the values of 1/n obtained from the plot were ∼1 showing linearity, henceforth unit of the k d value can be taken as litre/kilogramme. The overall process of sample preparation and computation of the k d values are outlined in Fig. 1.

Fig. 1
figure 1

A schematic flow sheet for preparation of UCBT, CBT and computation of the k d

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Physical characteristics of UCBT and CBT

The ASTM D 5550-06 guidelines(ASTM D 5550–06, 2008)  were used for determination of specific gravity, G, of the soil by employing an ultra-pycnometer (Quantachrome, USA). Each experiment was performed in triplets, i.e. soil was analysed three times in the test and the average values are presented in Table 1.

Table 1 The properties of the soil-contaminant system
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Chemical, mineralogical and morphological characteristics of UCBT and CBT

A water quality analyser (Model PE 136, Elico Ltd., India) was used to acquire the chemical characteristics pH, electrical conductivity (σ) and total dissolved solid (TDS) of the UCBT and CBT at L/S = 20 (as shown in Table 1). Notably, σ and pH of distilled water were measured to be 4.88 μS/cm and 7.0 ± 0.2, respectively, at 25 °C. The ethylene glycol monoethyl ether (EGME) method was adopted to determine the SSA (in m2/g) of UCBT and CBT in two replicates, as follows:

$$ \mathrm{SSA}={W}_{\mathrm{a}}{\left(0.000286{W}_{\mathrm{s}}\right)}^{-1} $$
(2)

where, W a is the amount of EGME adsorbed on the soil (in g) and W s is weight of the dry soil (in g). An average of the obtained results of SSA is listed in Table 1. Whereas the CEC of soil is indicative of type of ions sorbed onto it, their replacement by other ionic species was determined by the sodium acetate method. For this, triplets of each sample were measured and their average is presented in Table 2.

Table 2 The cation exchange capacity of UCBT and CBT at different pH
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The influence of contaminant on soil particles can be estimated by measuring its net surface charge or zeta potential (ζ), determined at L/S = 100 (results shown in Table 1) using an automated electrophoresis instrument Zeta PALS, BIC (USA). Chemical bonding in soil samples was identified by using the acquired spectra of Fourier transform infrared (FTIR; Model: MAGNA 550, USA), for which the KBr was used to prepare the soil sample pallets. Samples were analysed in the mid-IR range from 4000 to 400 cm−1 (spectra shown in online resource Fig. S1). The mineralogical characterization was carried out by using an X-ray diffraction spectrometer (Philips, Eindhoven, the Netherland) with Cu-Kα as a radiation source under a slow scan of 2 °/min (refer Fig. 2 and online resource Table ST1). The morphological characteristics of the soil samples were studied by an environmental scanning electron microscope (SEM; Quanta 200, FEI, Netherland) and the micrograph is presented in online resource Fig. S2.

Fig. 2
figure 2

The XRD diffractograms of CBT and UCBT at different pH. Ae aerinite, Av allevardite, F faujasite, H halite, M montmorillonite, N nontronite-15A, I illite, P periclase, S saponite, Mu muscovite, Ca cacoxenite, Si sillimanite, A anorthite, Na natrolite, V vermiculite, Mi mica, Ma magnetite, Cl clinochlore IIb-4, Fo forsteriteferroan, B beidellite-12A, Ru rutile

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Thermal and mechanical characteristics of UCBT and CBT

The thermal response of soil was observed by the thermo gravimetric and differential thermal analysis (TG-DTA) using a Perkin Elmer (USA) Diamond TG/DTA. Measurements of weight loss against the original input soil (20 mg for each sample) up to 600 °C were taken under isothermal heating in air (flow rate 50 mL/min) and a heating rate of 5 °C/min (results shown in online resource Fig. S3). The endothermic and exothermic behaviour were analysed via the recorded difference in temperature between soil sample and reference material (alumina), results shown in online resource Fig. S4.

The mechanical properties (viz., hardness, H and residual young’s modulus, Er) of soil were determined by load-displacement relationships (Oyen and Cook 2009). A nanoindentation (Model TI 900 TriboIndenter, Hysitron, USA) experiment was performed on UCBT and CBT at three different loads of 1000, 2000 and 3000 μN (Srinivas et al. 2013). Subsequently, the hardness and residual modulus were calculated by Eqs. 3 and 4, respectively, as follows:

$$ H=\frac{P}{A_{\mathrm{c}}} $$
(3)
$$ {E}_{\mathrm{r}}=\frac{\sqrt{\pi}}{2\beta \sqrt{A_{\mathrm{c}}}} S $$
(4)

where P is the applied load (in μN), A c is the projected contact area between the indenter and the sample (in nm2), π and β are dimensionless correction factors for the indenter tip shape and S is the contact stiffness.

Two different indentation area of each sample were selected and the numbers of indentations of each area were 12. The average result of 12 indentations of each area was shown as replicates I and II as given in online resource Table ST2. Further, to generalize the discrimination in soil heterogeneity, the average results of hardness and residual modulus of applied load taken at different concentrations of contaminated bentonite, are shown in Fig. 3.

Fig. 3
figure 3

The mechanical properties (a) Residual Young’s Modulus and (b) hardness for CBT at different pH

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Results and discussions

Effect of physicochemical-mineralogical properties

The cation exchange capacity (CEC) is one of the main factors, which has been considered for manifesting the number of ionic charges on soil particles and the sorption characteristics of soil-contaminant systems (Pathak et al., 2014a, b, 2015). A high CEC values results from the isomorphous substitution of Si4+ and Al3+ with low valance cations (Na+, K+, Ca2+, Mg2+) in the soil structure. Due to its larger SSA, the bentonite is a candidate buffer material (Caglar et al. 2009; Seliman et al. 2014). The context can easily be understood from Tables 1 and 2, as the characteristics of CBT such as G, SSA, CEC and ζ are greatly affected by the variance of pH and contaminant concentration. The soil characteristics as a function of contaminant concentration in CBT are shown in Fig. 4, which indicates that increasing concentration of contaminant in CBT (from 0.001 to 0.1 M SrCl2) at pH ≥ 4 resulted in a decreasing G, SSA and CEC; on the contrary, soil properties tended to be enlarged with increasing CBT concentration at pH 1. Furthermore, the values of ζ were found to be ≤ ±30 mV which specified the soil suspensions were not stable; however, in CBT(0.001MSrCl2) at pH 10, ζ was −33.23 mV and displayed stable soil particles. The ζ determines an isoelectric point (IEP) and estimates the interaction between soil-contaminant solution as a function of pH and concentration of the electrolyte species (Duman and Tunc, 2009; Moayedi et al. 2011). Keeping this in view, pH vs. ζ relationship was plotted (refer online resource Fig. S5) and IEP (i.e. ζ = 0 mV, represent neutral sites) were estimated at pH 1.2, 1.3 and 2.6 for CBT(0.001MSrCl2), CBT(0.01MSrCl2) and CBT(0.1MSrCl2), respectively. It was noticed that pH greater than IEP, the mobile counter ions (AlO4)5−, OH− and (MgO4)6− in the diffuse layer yielded negative ζ (refer Table 1). However, when pH decreases from IEP then surface charge on CBT get changed due to sorption of counter ions and displays screening ion effect. The ζ values +1.84 to +10.04 mV were obtained at pH 1 in consequence of H+ and Sr2+ which sorbed on outer Helmholtz plane of the Stern layer. In the electrostatic field, these ions move towards slipping plane and alters the surface charge from negative to positive (Duman and Tunc, 2009; Wang et al. 2010). A further uptake of contaminant onto CBT could decrease the amount of protons by replacing with Sr2+. In this circumstance, the surface active co-ions can move towards diffuse layer and revers the surface charge, which can be corroborated to the available sorption sites onto CBT.

Fig. 4
figure 4

The relationship between concentrations of SrCl2 at CBT with soil characteristics, i.e. specific gravity, G; specific surface area, SSA; cation exchange capacity, CEC and zeta potential, ζ at different pH

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The change in chemical bonding of UCBT after being contaminated can be discussed by analysing the FTIR spectra (refer online resource Fig. S1). A slight shift in peak position from 3348 to 3410 cm−1 from UCBT to CBT, respectively, corresponds to the OH stretching of H-bonded water. The change is due to the replacement of divalent Ca2+ and Mg2+ with Sr2+. The band at 3604 cm−1 can be attributed to adsorbed water (stretching mode of hydroxyl group) and here mainly due to the Al─Al─OH and Al─Mg─OH (Caglar et al. 2009; Madejova 2003). A band at 1631 cm−1 corresponds to the bending mode of adsorbed water (H─O─H) in soil with SiO2/Fe2O3 minerals. The intensive peaks at 1024 and 520 cm−1 correspond to the vibration stretching of Si─O─Si and bending of Si─O─Al, respectively. The Fe─O vibrational bands were observed at 457 and 1484 cm−1. The change in FTIR spectra of UCBT and CBT revealed the acidic and alkaline attack on bentonite which led to the alteration of mineralogical properties of bentonite and was further evident from XRD analysis (as shown in Fig. 2 and online resource Table ST1). The major mineral of UCBT was found to be montmorillonite (Seliman et al. 2014). The XRD analysis indicated that as the concentration of contaminant increased with pH, the degree of crystallinity of CBT also increased (Fig. 2). Further, at low concentration of CBT(0.001M SrCl2) and pH ≥ 4, slight changes were observed in the mineral phases. However, peaks of nontronite and allevardite were identified at pH 1. Nontronite is an iron-rich smectite mineral whose formation may have been facilitated by the iron content in the sample (refer Table 2). The formation of allevardite indicated the interstratified mica and montmorillonite layers in the structure. The mineralogical alteration observed in bentonite under the studied conditions clearly displayed the plausible changes in the sorption characteristics of bentonite (Guven 1990; Galambos et al., 2012).

Furthermore, the surface morphology of UCBT and CBT explain the arrangement of soil particles (online resource Fig. S2). Vertical rod-shaped particles (due to SrSO4 and/or SrCl2 crystals) on soil surfaces were observed at pH 1 with phase-to-phase arrangement in a dispersed manner. As pH increased from 4 to 10, the flocculated soil particles increases the repulsive forces between the soil surface and vary with the ζ value. Interestingly, at pH 13, the soil particles displayed a stratified flocculated structure. Hence, a decreased thickness of the diffuse double layer resulted in shrinking soil particles of CBT (Gumaste et al. 2014); consequently, increased specific gravity and reduced SSA were also observed (as presented in Table 1). This highlights the fact that contaminant exposure onto bentonite decreases the sorption capacity of CBT than that of the UCBT.

Effect of thermal and mechanical properties

The thermal stability of bentonite was investigated by thermogravimetric and differential thermal analyses (TG-DTA). The TGA graph (online resource Fig. S3) revealed rapid weight loss between 27 and 183 °C which was due to the loss of water molecules (Lalikova et al. 2010). A further increase in temperature up to 600 °C showed a slight weight loss due to dehydroxylation of organic compounds (Lalikova et al. 2010; Wendlandt 1975). This would indicate that the disposal of radioactive waste at higher temperature (>100 °C) may cause inertness and decrease the sorption capacity of bentonite. The two peaks in the temperature ranges 90–100 °C and 140–160 °C of the DTA curve (online resource Fig. S4) agree with the loss of water (Smykatz-Kloss 1982). The requirement of more energy above 400 °C is ascribed to the loss of hydroxyl ions from organic compounds, owing to the endothermic behaviour of soil. Thus, the organic compounds in the soil were degraded at elevated temperature (>400 °C), and which led to a decrease in sorption.

Further, the results of nanoindentation experiments revealed substantial changes in mechanical properties (hardness and residual Young’s modulus) of CBT and UCBT (online resource Table ST3). Hardness and residual Young’s modulus of UCBT are represented as horizontal dotted lines, i.e. 0.12 and 3.51 GPa, respectively, in Fig. 3a, b. It can be concluded that the hardness and residual stress of CBT decreased with an increased concentration of contaminant from 0.001 to 0.1 M, thus reducing the binding energy of ions present in CBT, might be responsible for alteration in SSA and CEC values. Moreover, an increase in binding energy can break the ionic bonds (Parkhomenkho 1967). Moreover, the hardness of the material was highly dependent on its phases viz., crystalline and amorphous (Hancock et al. 2002). The crystalline phase of CBT demonstrated by the sharp peaks of XRD (Hancock et al. 2002) as shown in Fig. 2 was more prominent with 0.1 M strontium contaminant which resulted in a substantial decrease in hardness (Fig. 3a). The hardness of the crystalline phase was less than in the amorphous phase of materials. Furthermore, the residual modulus (E r) corresponding to the resilient stress in bentonite (Oliver and Pharr 1992) was subjected to effects of contaminant at different pH and concentrations. At pH 1 and 10, the E r decreased with an increased SrCl2 concentration. However, remaining CBT at pH 4, 8.5 and 13 exhibited stress randomness where E r values decreased from 0.001 to 0.01 M, and increased for 0.1 M SrCl2 (Fig. 3b). Based on this, the alteration in soil properties occurs at higher to lower concentration of contaminants. Such an alteration can be attributed to the variation in elasto-plastic behaviour of soil (Oliver and Pharr 1992), mainly due to the chemical attack onto soil at different pH.

Effect of soil properties on sorption mechanism

The minerology of bentonite significantly affects the sorption mechanism of the bentonite-contaminant system. As can be seen from Fig. 5, the octahedral Al3+ and tetrahedral Si4+ sheets of the bentonite mineral are interacted by the isomorphous substitution of higher valence cations with the lower valences. These substitutions create a negative charge excess for the unsatisfied valences (Poernomo 2010; Seliman et al. 2014). Subsequently, the strontium (as Sr2+) gets sorbed by the replacement of exchangeable cations (commonly the Na+, K+ and Ca2+) and via ion-pairing with the negative charge excess created by anionic (AlO4)5− and (MgO4)6− (as shown in Fig. 5). This leads to chemisorption; however, physisorption via outer-sphere complexation of Sr2+ onto bentonite surface can also take place (Wallace et al. 2012; Poernomo 2010). Notably, the net negative charge excess compensated by metal cations (including Sr2+) establishes the electroneutrality onto bentonite surfaces.

Fig. 5
figure 5

A typical representation of sorption mechanism for the bentonite-strontium system

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Nevertheless, the CBT further exposed to contaminant (SrCl2) solution at different concentrations also revealed some other phenomena. Figures. 6, 7 and 8 showed the relationship between C s verses C e at different pH and types of CBT. The concentration range above the horizontal dotted line indicated the contribution of contaminant for quantifying the k d in a soil-contaminant system, whereas below that line, sorption reversibility due to a salting-in effect took place which is also known as desorption (Pathak et al. 2015). The CBT(0.1M SrCl2) contacted with 0.1–0.7 M contaminant solution exhibited the sorption quite well, however, by contacting a lower contaminant solution (i.e. <0.1 M SrCl2, the increased concentration of contaminant in the solution) at equilibrium was an evidence for sorption reversibility or desorption. A continuing trend was also observed from Figs. 6, 7 and 8 for the 0.01 and 0.001 M concentration of CBT. Figures 6a and 8a demonstrated that at pH 1, CBT(0.1M SrCl2) and CBT(0.01M SrCl2) displayed a greater k d as compared to the CBT(0.001M SrCl2). The fact to be emphasized is that commonly at a lower pH, the protonation of specific sorption sites onto bentonite surface prevents the electrostatic sorption of Sr2+ (Wallace et al. 2012; Krauskopf and Bird 1995). This phenomenon clearly describes the insignificant k d at pH 4. In contrast, the high k d value at pH 1 can be corroborated with the permanent structural negative charges in clay minerals (Chen and Hayes 1999; Coppin et al. 2002; Parkman et al. 1998) and the presence of vermiculite (Langley et al. 2009; Wang et al. 2010). Additionally, precipitates of SrSO4 co-exist with CBT(0.1M SrCl2) at pH 1 but redissolved when contacted with the contaminant solution, and thus, yielded a high k d value. Subsequently, the sorption k d = 2.92 L/kg of CBT(0.001M SrCl2) at pH 8.5 was consistent with the maximum sorption observed for Sr2+ on soil ascribed to the contribution of anionic hydroxyls (Wallace et al. 2012). The lower k d values obtained for the CBT(0.01M SrCl2) and CBT(0.1M SrCl2) clearly indicated the effect of concentration (Wallace et al. 2012; Pathak et al. 2015). At alkaline pH 10, the addition of NaOH to fix the pH creates a significant area of uncertainty for the distribution of Na+ between the interlayer and inter-particle space (Galambos et al. 2012), and therefore may have decreased the k d value. Interestingly, on further increase in pH up to 13, the k d value for CBT(0.1M SrCl2) comparatively increased, which can be attributed to the anionic metal species onto bentonite surfaces. A plot for the k d of CBT at varied concentrations as a function of pH is presented in Fig. 9. It revealed that the CBT(0.001M SrCl2) exhibited a higher k d (2.92 L/kg) value at soil pH 8.5 though CBT(0.01M SrCl2) has a maximum k d (1.23 L/kg) at acidic pH 1. On the other hand, the CBT(0.1M SrCl2) showed higher k d at the extreme acidic and alkaline conditions as well (i.e. 2.37 L/kg at pH 1 and 1.10 L/kg at pH 13).

Fig. 6
figure 6

The distribution coefficients for CBT-SrCl2 system at pH 1 and 4 a 0.001 M, b 0.01 M and c 0.1 M

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Fig. 7
figure 7

The distribution coefficient for CBT-SrCl2 system at pH 8.5 and 10 a 0.001 M, b 0.01 M and c 0.1 M

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Fig. 8
figure 8

The distribution coefficient for CBT-SrCl2 system at pH 13 and UCBT-SrCl2 at 8.5 a 0.001 M, b 0.01 M and c 0.1 M

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Fig. 9
figure 9

The correlation of the distribution coefficient of CBT with respect to pH

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For an ideal candidate buffer material, the capacity to uptake contaminant/s is one of the prime factors to be determined along with the characteristics of sorbent material. Hence, the calculated sorption capacity of strontium onto the investigated samples was compared with other sorbents viz. activated charcoal, GMZ bentonite and local soil as reported in the literature (Chegrouche et al. 2009; He et al. 2016; Chegrouche et al. 2009; Hong et al., 2016). Notably, it was found that at soil pH, the sorption capacity of UCBT, i.e. 5.51 × 10−3 mol/m2 was well in the range as compared to other sorbents such as activated charcoal 5.71 × 10−3 mol/m2, GMZ bentonite 1.94 × 10−2 mol/m2 and local soil 1.50 × 10−6 mol/m2 (Chegrouche et al. 2009; He et al. 2016; Skupinski and Solecki 2014). It has been demonstrated that WIn-BT can be used as a potential buffer material in waste repositories.

Conclusions

Sorption properties of bentonite (WIn-BT) changed quite substantially due to the acidic and alkaline attack of strontium. The study revealed that sorption was highly influenced by the mineralogical changes that occurred due to pH variation, as well as changes in surface charge density, binding energy of ions and strontium concentration. Spectral and thermal analysis of the UCBT and CBT at varied pH and concentrations of SrCl2 clearly demonstrated the alteration in soil properties. Due to the contaminant exposure to soil, substantial changes in the k d values of UCBT and CBT were quantified. The maximum k d of UCBT and CBT(0.001M SrCl2) at soil pH 8.5 were 8.99 and 2.92 L/kg, respectively; whereas for CBT(0.1M SrCl2) and CBT(0.01M SrCl2), the maximum k d 2.37 and 1.23 L/kg, respectively, were obtained at acidic pH 1. A significant k d value (1.10 L/kg) observed at pH 13 for CBT(0.1M SrCl2) was due to the ion-pairing of Sr2+ with the anionic metal species onto bentonite. Based on the experimental findings, WIn-BT can be used as candidate buffer material at soil pH. However, at a higher concentration of contaminant, the sorption would have a negative effect with a declined k d. Moreover, the sorption capacity of UCBT (5.51 × 10−3 mol/m2) was in the range of a potent buffer candidate for waste repositories as compared to other sorbents.