**Clarification of Adsorption Reversibility on Granite that Depends on Cesium Concentration**

Keita Okuyama and Kenji Noshita *Hitachi Research Laboratory, Hitachi, Ltd. Japan* 

#### **1. Introduction**

118 Municipal and Industrial Waste Disposal

Gombert , D. (2007) *Appendixes for Global Nuclear Energy Partnership Integrated Waste* 

Kato, H.; Kato, O. & Tanabe, H. (2002) *Review of Immobilization of Techniques of Radioactive* 

Kornweitz, H. & Levine, R. D. (1998) Formation of molecular iodine in high-energy fourcenter CH3I+CH3I collisions, *Chem. Phys. Lett.* Vol. 294, pp. 153-161. Lee, J. Y.; Olson, D. H.; Pan, L.; Emge, T. J. & Li, J. (2007) Microporous metal-organic

Metropolis, N.; Rosenbluth, A. W.; Rosenbluth, M. N.; Teller, A. H. & Teller, E. (1953)

NEA OECD (2006) *Advanced Nuclear Fuel Cycles and Radioactive Waste Management*. NEA No.

Pasternak, A.; Anderson, A. & Leech, J. W. (1977) Bond charge model for lattice-dynamics of

Peters, M. & Ewing, R. C. (2007) A science-based approach to understanding waste form

Rovnyi, S. I.; Pyatin, N. P. & Istomin, I. A. (2002) Catching of I-129 during processing of spent nuclear fuel from power plants, *Atomic Energy*. Vol. 92, pp. 534-535. Stephenson, D.J.; Fairchild, C.I.; Buchan, R.M., & Dakins, M.E. (1999) A fiber

Sudik, A. C.; Côté, A. P.; Wang-Foy, A. G.; O'Keeffe, M. & Yaghi, O. M. (2006) A metal-

Wang, Y. & Gao, H. (2006) Compositional and structural control on anion sorption

Wang, Y.; Bryan, C.; Gao, H.; Pohl, P.; Brinker, C. J.; Yu, K.; Xu, H.; Yang, Y.; Braterman, P. S.

Wang, Y.; Gao, H.; Yeredla, R.; Xu, H. & Abrecht, M. (2006) Control of surface functional

Xu, H. & Wang, Y. (2000) Crystallization sequence and microstructure evolution of Synroc samples crystallized from CaZrTi2O7 melts, *J. Nuclear Mater.* Vol. 279, pp. 100-106.

iodine, *J. Phys. C: Solid State Phys.* Vol. 10, pp. 3261-3271.

hazard, *Aerosol Science and Technology*. Vol. 30, pp. 467-476.

blocks, *Angewandte Chemie*, Vol. 45, 2528-2533.

mass transfers in nanopores, *Geology*, Vol. 31, 387-390.

Grambow, B. (2006) Nuclear waste glasses - How durable? *Elements*, Vol. 2, pp. 357-364. Jubin, R. T. (1994) *The Mass Transfer Dynamics of Gaseous Methyl-Iodide Adsorption by Silver-*

*Exchanged Mordenite*. Oak Ridge National Laboratory, ORNL-6853.

GENP-WAST-WAST-AI-RT-2007-000324.

*Materials*, Vol. 17, 1255-1262.

1087-1092.

5990.

395-401.

19-26.

209-217.

*Iodine for Geological Disposal. JAERI-Conf* 2002-004.

*Management Strategy Waste Treatment Baseline Study*. Idaho National Laboratory,

frameworks with high gas sorption and separation capacity, *Advanced Functional* 

Equation of state calculations by fast computing machines, *J. Chem. Phys.* Vol. 21,

durability in open and closed nuclear fuel cycles, *J. Nuclear Materials*. Vol. 362, pp.

characterization of the natural zeolite, mordenite: A potential inhalation health

organic framework with a hierarchical system of pore and tetrahedral building

capability of layered double hydroxides (LDHs), *J. Colloid Interface Sci.* Vol. 301, pp.

& Xu, Z. (2006) *Potential Applications of Nanostructured Materials in Nuclear Waste Management*. Sandia National Laboratories, Albuquerque, NM. SAND2003-3313. Wang, Y.; Bryan, C.; Xu, H. & Gao, H. (2003) Nanogeochemistry: Geochemical reactions and

groups on pertechnetate sorption on activated carbon, *J. Colloid Interface Sci.*, 305,

Disposal of high-level radioactive wastes (HLW) is planned to be done in a repository located deep underground to isolate radionuclides from the biosphere. In case of a leakage accident of HLW, there will be no hazardous impact to humans because migration of the leaked radionuclides will be retarded by matrix diffusion and adsorption on the rock surface. Therefore, the geochemical retardation behavior of radionuclides in aquifers must be clarified, from the viewpoint of the performance assessment of HLW deep underground disposal.

Radionuclide-adsorbed sites in rock are classified into two general types: reversible adsorption sites, where desorption of once-adsorbed nuclides occurs (e.g., an ion-exchange reaction); and irreversible adsorption sites, where only adsorption occurs (e.g., a mineralization reaction). In the early stage of radionuclide leakage, migration of the radionuclides will be retarded by both reversible and irreversible adsorption sites. However, when the irreversible sites are filled, the radionuclides will no longer be retarded by them. Therefore, it is necessary to investigate the sorption reversibility to clarify the behavior of radionuclides (Fukui, 2004).

Cesium-137 (137Cs), which is one of the principal radioactive sources of HLW for 1000 years after geological disposal, is partially fixed in the interlayer of micas and it might be trapped irreversibly in these sites (Francis & Brinkley, 1976). Another report suggested that Cs adsorption on granite has a nonlinear relationship with its concentration in solution (Ohe, 1984). These findings imply it is possible to characterize Cs adsorption on granite. One of the important characterization factors is the adsorption amount on reversible and irreversible sites, which may change with Cs concentration; however, it has not been reported.

The purpose of this study is to clarify Cs adsorption reversibility on granite that depends on Cs concentration by adsorption and sequential extraction experiments, using a variety of chemical reagents for various inlet Cs concentrations (1.0 × 10-3 - 1.0 × 100 mol/m3). For the experiments, a narrow flow channel was formed on a granite specimen (Okuyama, et al., 2008). Breakthrough curves (BTCs) were obtained by injecting a Cs solution labeled with

Clarification of Adsorption Reversibility on Granite that Depends on Cesium Concentration 121

Oxide Content (wt%)

Oxide Content (wt%)

10 mm

Fig. 1. Photograph of a granite specimen showing side edge surfaces coated with an epoxy

**Granite specimen**

**Epoxy resin**

Quartz 37.9 Plagioclase feldspar 33.0 Potassium feldspar 18.8 Biotite 9.0 Chlorite 1.0 Prehnite 0.2 Carbonate mineral 0.1 Total 100.0

SiO2 69.2 Al2O3 15.9 K2O 4.30 CaO 2.97 Na2O 2.90 Fe2O3 2.54 P2O5 0.60 MgO 0.48 TiO2 0.29 MnO 0.05 Total 99.24

Table 1. Granite chemical composition

Table 2. Granite mineral composition

resin

134Cs into the channel. We estimated the amounts diffused into the rock matrix and absorbed on the rock surface. In order to verify the estimated sorption amount, we obtained the retardation coefficient by analyzing BTCs, and compared the value with that calculated using the distribution coefficient (*Kd*) obtained by a batch sorption experiment. After injecting 134Cs solution, extraction reagents (HCl, CaCl2, and KCl solutions) were injected into the flow channel in sequence. We obtained the desorption curves and investigated the chemical speciation of Cs in granite. In particular, we focused on the dependence of adsorption amount on reversible and irreversible sites on Cs concentration.

#### **2. Experiment**

#### **2.1 Materials**

Biotite granite from the Makabe area of Japan was used in this work. The chemical composition was measured by fluorescent X-ray analysis and the results are listed in Table 1. The mineral composition was also graphically determined using a polarization microscope and these results are listed in Table 2. The porosity of a granite specimen was measured by the water saturation method (Skagius & Neretnieks, 1986), and 0.73 % was obtained. The cation-exchange capacity was determined by the Peech method (Peech, 1965), and 5.9 meq/100g was obtained.

#### **2.2 Adsorption and sequential extraction experiments**

In order to clarify sorption reversibility of Cs on granite, we carried out adsorption and sequential extraction experiments. All specimens were cut as a parallelepiped with dimensions of 20 mm × 4 mm × 3 mm and their surfaces were polished with abrasive paper (#1200). The side edge surfaces were thickly coated with an epoxy resin to avoid evaporation of the test solution during experiments (Fig.1). A container was filled with distilled water (pH adjusted to 6 using HCl and NaOH) and set in a glass bell jar. Granite specimens were immersed in the water. Then, the bell jar was evacuated to remove air from the micropores of the specimens. The specimens were kept for 24 hours immersed in the distilled water in the bell jar under reduced pressure.

Experimental apparatus is shown in Fig.2. It consisted of an injection unit, a reaction unit, and a storage unit. For the reaction unit, a narrow flow channel (20 mm length, 4 mm width, and 160 µm depth) was formed from a fluoroplastic spacer (160 µm thick) with a slit (20 mm long, 4 mm wide). Two fluoroplastic base plates sandwiched the spacer and granite specimen, then everything was held together by applying pressure to downward from the top base plate and upward from the bottom base plate. A photograph of the spacer is shown in Fig.3. The roughness of the granite specimen and the bottom base plate surfaces was overcome by elastic deformation of the fluoroplastic spacer, so no leakage of solution occurred. The bottom base plate had inlet and outlet ports (0.5 mm inner diameter), for injecting or draining out radionuclide or extraction solutions, and fluoroplastic tubing was attached to each port. A blank experimental run in which the granite specimen was replaced with a fluoroplastic plate was carried out to verify the surfaces of the fluoroplastic spacer and base plates were non-reactive, then the actual experiments were done using the following procedures.


Table 1. Granite chemical composition

134Cs into the channel. We estimated the amounts diffused into the rock matrix and absorbed on the rock surface. In order to verify the estimated sorption amount, we obtained the retardation coefficient by analyzing BTCs, and compared the value with that calculated using the distribution coefficient (*Kd*) obtained by a batch sorption experiment. After injecting 134Cs solution, extraction reagents (HCl, CaCl2, and KCl solutions) were injected into the flow channel in sequence. We obtained the desorption curves and investigated the chemical speciation of Cs in granite. In particular, we focused on the dependence of

Biotite granite from the Makabe area of Japan was used in this work. The chemical composition was measured by fluorescent X-ray analysis and the results are listed in Table 1. The mineral composition was also graphically determined using a polarization microscope and these results are listed in Table 2. The porosity of a granite specimen was measured by the water saturation method (Skagius & Neretnieks, 1986), and 0.73 % was obtained. The cation-exchange capacity was determined by the Peech method (Peech, 1965),

In order to clarify sorption reversibility of Cs on granite, we carried out adsorption and sequential extraction experiments. All specimens were cut as a parallelepiped with dimensions of 20 mm × 4 mm × 3 mm and their surfaces were polished with abrasive paper (#1200). The side edge surfaces were thickly coated with an epoxy resin to avoid evaporation of the test solution during experiments (Fig.1). A container was filled with distilled water (pH adjusted to 6 using HCl and NaOH) and set in a glass bell jar. Granite specimens were immersed in the water. Then, the bell jar was evacuated to remove air from the micropores of the specimens. The specimens were kept for 24 hours immersed in the

Experimental apparatus is shown in Fig.2. It consisted of an injection unit, a reaction unit, and a storage unit. For the reaction unit, a narrow flow channel (20 mm length, 4 mm width, and 160 µm depth) was formed from a fluoroplastic spacer (160 µm thick) with a slit (20 mm long, 4 mm wide). Two fluoroplastic base plates sandwiched the spacer and granite specimen, then everything was held together by applying pressure to downward from the top base plate and upward from the bottom base plate. A photograph of the spacer is shown in Fig.3. The roughness of the granite specimen and the bottom base plate surfaces was overcome by elastic deformation of the fluoroplastic spacer, so no leakage of solution occurred. The bottom base plate had inlet and outlet ports (0.5 mm inner diameter), for injecting or draining out radionuclide or extraction solutions, and fluoroplastic tubing was attached to each port. A blank experimental run in which the granite specimen was replaced with a fluoroplastic plate was carried out to verify the surfaces of the fluoroplastic spacer and base plates were non-reactive, then the actual experiments were done using the

adsorption amount on reversible and irreversible sites on Cs concentration.

**2. Experiment 2.1 Materials** 

and 5.9 meq/100g was obtained.

following procedures.

**2.2 Adsorption and sequential extraction experiments** 

distilled water in the bell jar under reduced pressure.


Table 2. Granite mineral composition

Fig. 1. Photograph of a granite specimen showing side edge surfaces coated with an epoxy resin

Clarification of Adsorption Reversibility on Granite that Depends on Cesium Concentration 123

evaporation of effluent solution during fractionation, effluent solution was stored in a storage unit (0.5 mm inner diameter, 7.6 m length). After an adsorption experimental run, the stored solution was flushed out with a high flow velocity (1.3 × 10-3 m/s), and collected in a small vial. Collection quantity was obtained precisely by weighing the vials before and after runs. The concentrations of 134Cs in the effluent of the flow channel were determined with a germanium (Ge) semiconductor detector. The error of concentration measurement was less than 5%. The experimental apparatus, excluding the radioactivity detectors, was assembled in a glove box filled with air to keep dust particles off and all experimental runs were done at 25 ± 2 C. After these procedures were carried out, extraction reagents were injected (Table 3) into the flow channel in sequence under the same operating conditions for the sequential extraction experiments. A series of experimental runs was carried out by

Fig. 3. Photograph of a fluoroplastic thin spacer (160 µm thick) with the slit which forms the

We estimated Cs adsorption amount on granite by BTCs. In order to verify this estimation, we evaluated the retardation coefficient by analyzing BTCs, and compared it with the value calculated using the distribution coefficient (*Kd*) obtained by a batch sorption experiment. The experimental procedures for the batch sorption experiment were as follows. A granite specimen was crushed and sieved to obtain particle sizes in the range of 0.125 - 0.25 mm.

Extractant Composition

(mol/m3)

HCl 1.0×10-2 5 CaCl2 5.0×102 5 KCl 5.0×102 5

10 mm

pH

changing the Cs concentration (1.0 × 10-3 - 1.0 × 100 mol/m3).

flow channel on the granite plate

Table 3. Extraction reagents

**2.3 Batch sorption experiment** 

Fig. 2. Experimental apparatus for adsorption and sequential extraction experiments

In an experimental run, Cs solution labeled with 134Cs (specific activity adjusted to 9.8 × 108 Bq/m3) was injected into the flow channel at constant flow velocity (2.6 × 10-5 m/s). The flow velocity was maintained for 24 hours by using an injection pump in all experiments. Average residence time in the flow channel was 12.8 minutes. To avoid evaporation of effluent solution during fractionation, effluent solution was stored in a storage unit (0.5 mm inner diameter, 7.6 m length). After an adsorption experimental run, the stored solution was flushed out with a high flow velocity (1.3 × 10-3 m/s), and collected in a small vial. Collection quantity was obtained precisely by weighing the vials before and after runs. The concentrations of 134Cs in the effluent of the flow channel were determined with a germanium (Ge) semiconductor detector. The error of concentration measurement was less than 5%. The experimental apparatus, excluding the radioactivity detectors, was assembled in a glove box filled with air to keep dust particles off and all experimental runs were done at 25 ± 2 C. After these procedures were carried out, extraction reagents were injected (Table 3) into the flow channel in sequence under the same operating conditions for the sequential extraction experiments. A series of experimental runs was carried out by changing the Cs concentration (1.0 × 10-3 - 1.0 × 100 mol/m3).

Fig. 3. Photograph of a fluoroplastic thin spacer (160 µm thick) with the slit which forms the flow channel on the granite plate


Table 3. Extraction reagents

122 Municipal and Industrial Waste Disposal

**Injection unit Reaction unit Storage unit**

Water-saturated in distilled water (pH = 6) under a vacuum for 24 h.

Pressurization

Pressurization

Fig. 2. Experimental apparatus for adsorption and sequential extraction experiments

In an experimental run, Cs solution labeled with 134Cs (specific activity adjusted to 9.8 × 108 Bq/m3) was injected into the flow channel at constant flow velocity (2.6 × 10-5 m/s). The flow velocity was maintained for 24 hours by using an injection pump in all experiments. Average residence time in the flow channel was 12.8 minutes. To avoid

Fluoroplastic tube 0.5 mm inner diameter,

**Spacer**

**Base plate**

**Epoxy resin**

**Base plate**

**Outlet port** (0.5 mm inner diameter)

7.6 m length

**Detail of reaction unit**

**Flow channel** (20 mm × 4 mm ×160 µm)

> **Inlet port** (0.5 mm inner diameter)

**Granite specimen**

134Cs solution

or extraction reagents

#### **2.3 Batch sorption experiment**

We estimated Cs adsorption amount on granite by BTCs. In order to verify this estimation, we evaluated the retardation coefficient by analyzing BTCs, and compared it with the value calculated using the distribution coefficient (*Kd*) obtained by a batch sorption experiment. The experimental procedures for the batch sorption experiment were as follows. A granite specimen was crushed and sieved to obtain particle sizes in the range of 0.125 - 0.25 mm.

Clarification of Adsorption Reversibility on Granite that Depends on Cesium Concentration 125

is thickness of granite specimen (m), *De* is effective diffusion coefficient (m2/s), *DL* is longitudinal dispersion coefficient for the flow channel (m2/s), *ε* is porosity of granite ( - ), *L* is length of the flow channel (m), *Rf* is retardation factor of granite surface ( - ), *Rm* is retardation factor of granite matrix ( - ), *t* is elapsed times (s), *V* is flow velocity in the flow channel (m/s), *x* is distance along the flow channel (m) and *z* is vertical distance from the

*X* 

**Calculation region**

(4)

**resin**

flow channel (m).

**Granite** 

*Z* 

**Flow channel**

*d*

channel (m2).

*b* Advection and diffusion

*L*

**Base plate**

**specimen Epoxy**

diffusion

**Base plate**

*VC D*

*z*

*x*

Fig. 4. Analysis object; vertical section showing the granite plate and flow channel

The boundary conditions around the flow channel were as shown in Eqs*.* (4) and (5):

*x C*

 

*f <sup>f</sup> <sup>L</sup>*

*x 0*

where, *F* is the inlet flux of radionuclide (mol/s) and A is cross sectional area of the flow

*0, <sup>0</sup> <sup>x</sup> L, <sup>t</sup> <sup>0</sup>*

*0, <sup>0</sup> <sup>z</sup> d, <sup>t</sup> <sup>0</sup>*

Other boundary conditions around the granite specimen were as shown in Eqs. (6) – (8):

*Cm x, d,t*

*Cm 0, z,t*

**Spacer**

*A F*

*, t 0*

*C x,0,t C (x,t), 0 x L, t 0 <sup>m</sup> <sup>f</sup>* (5)

(6)

(7)

The sieved particles were rinsed with distilled water to remove the fine powder fragments. Specific surface area was measured by the BET method, and the mean value was 1.0 × 10-2 m2/kg. Then three grams of the sieved particles was saturated with distilled water (3.0 × 10-5 m3, adjusted to pH = 6) spiked with a radionuclide solution of 134Cs (specific activity 4.9 × 107 Bq/m 3); the saturation was done under a vacuum for 24 hours to fill the granite particle pores. The initial Cs+ solution concentrations ranged from 1.0 × 10-4 mol/m3 to 1.0 × 102 mol/m3. The particle-containing solutions were continuously stirred for seven days. Then the solid particles were removed by filtration through a membrane filter (pore size 0.45 μm) and the 134Cs concentration was measured with the Ge semiconductor detector. The *Kd* value was calculated by the conventional procedure as follows (Holland and Lee, 1992):

$$\text{Kd} = \frac{\text{C}\_{\text{int}} - \text{C}\_{\text{eq}}}{\text{C}\_{\text{eq}}} \cdot \frac{w}{w} \tag{1}$$

where *Cin* is initial radionuclide concentration (mol/m3), *Ceq* is equilibrium radionuclide concentration (mol/m3), *v* is solution volume (m3) and *w* is rock (granite) weight (kg).

#### **3. Analysis**

We obtained the retardation coefficient by parameter identification method for BTCs using an advection-diffusion equation. The analysis object was a vertical section through the granite specimen and the flow channel as shown in Fig.4. In the calculation, the flow channel was not deep (160μm); thus we could apply a one-dimensional advectiondispersion equation. On the other hand, the permeability of granite was much small thus a two-dimensional diffusion equation was modeled. The test time was several days at the longest; therefore the radioactive decay of 134Cs (half life 2.07 years) could be neglected.

The governing equations were thus formulated as follows:

For the flow channel

$$R\_f \frac{\partial \mathbf{C}\_f}{\partial t} = -V \frac{\partial \mathbf{C}\_f}{\partial \mathbf{x}} + D\_L \frac{\partial^2 \mathbf{C}\_f}{\partial \mathbf{x}^2} + \frac{D\_e}{b} \frac{\partial \mathbf{C}\_m}{\partial \mathbf{z}} \bigg|\_{\mathbf{z}=0} \tag{2}$$
 
$$\{t > 0, \ 0 < \mathbf{x} < \mathbf{L}\}$$

For the granite matrix region

$$R\_m \frac{\partial \mathbf{C}\_m}{\partial t} = \frac{D\_e}{\varepsilon} \frac{\partial^2 \mathbf{C}\_m}{\partial \mathbf{x}^2} + \frac{D\_e}{\varepsilon} \frac{\partial^2 \mathbf{C}\_m}{\partial \mathbf{z}^2} \tag{3}$$

$$\left(t > 0, \ 0 < \mathbf{x} < L, \ 0 < \mathbf{z} < d\right)$$

where, *b* is depth of the flow channel (m), *Cf*(*x*,*t*) is radionuclide concentration in the flow channel (mol/m3), *Cm*(*x*,*z*,*t*) is radionuclide concentration in the porous granite (mol/m3), *d*

The sieved particles were rinsed with distilled water to remove the fine powder fragments. Specific surface area was measured by the BET method, and the mean value was 1.0 × 10-2 m2/kg. Then three grams of the sieved particles was saturated with distilled water (3.0 × 10-5 m3, adjusted to pH = 6) spiked with a radionuclide solution of 134Cs (specific activity 4.9 × 107 Bq/m 3); the saturation was done under a vacuum for 24 hours to fill the granite particle pores. The initial Cs+ solution concentrations ranged from 1.0 × 10-4 mol/m3 to 1.0 × 102 mol/m3. The particle-containing solutions were continuously stirred for seven days. Then the solid particles were removed by filtration through a membrane filter (pore size 0.45 μm) and the 134Cs concentration was measured with the Ge semiconductor detector. The *Kd* value was calculated by the conventional procedure as follows (Holland

> *w v*

(1)

*z 0*

(2)

(3)

*e m*

*2 <sup>m</sup> <sup>2</sup> <sup>e</sup>*

*C*

 

*b D*

*ε D*

*C*

*2 f 2*

*x C*

 

*L*

*t 0, 0 x L*

*2 <sup>m</sup> <sup>2</sup> <sup>m</sup> <sup>e</sup> <sup>m</sup> <sup>z</sup>*

*x C*

*t 0, 0 x L, 0 z d*

where, *b* is depth of the flow channel (m), *Cf*(*x*,*t*) is radionuclide concentration in the flow channel (mol/m3), *Cm*(*x*,*z*,*t*) is radionuclide concentration in the porous granite (mol/m3), *d*

*ε D*

*<sup>f</sup> z*

 

*x C -V*

*D*

*eq <sup>C</sup> eq <sup>C</sup> in <sup>C</sup> Kd* 

where *Cin* is initial radionuclide concentration (mol/m3), *Ceq* is equilibrium radionuclide concentration (mol/m3), *v* is solution volume (m3) and *w* is rock (granite) weight (kg).

We obtained the retardation coefficient by parameter identification method for BTCs using an advection-diffusion equation. The analysis object was a vertical section through the granite specimen and the flow channel as shown in Fig.4. In the calculation, the flow channel was not deep (160μm); thus we could apply a one-dimensional advectiondispersion equation. On the other hand, the permeability of granite was much small thus a two-dimensional diffusion equation was modeled. The test time was several days at the longest; therefore the radioactive decay of 134Cs (half life 2.07 years) could be

The governing equations were thus formulated as follows:

*t C*

*R*

*f f*

*t <sup>C</sup> <sup>R</sup>*

and Lee, 1992):

**3. Analysis** 

neglected.

For the flow channel

For the granite matrix region

is thickness of granite specimen (m), *De* is effective diffusion coefficient (m2/s), *DL* is longitudinal dispersion coefficient for the flow channel (m2/s), *ε* is porosity of granite ( - ), *L* is length of the flow channel (m), *Rf* is retardation factor of granite surface ( - ), *Rm* is retardation factor of granite matrix ( - ), *t* is elapsed times (s), *V* is flow velocity in the flow channel (m/s), *x* is distance along the flow channel (m) and *z* is vertical distance from the flow channel (m).

Fig. 4. Analysis object; vertical section showing the granite plate and flow channel

The boundary conditions around the flow channel were as shown in Eqs*.* (4) and (5):

$$\left. \text{V} \mathbf{C}\_f - D\_L \frac{\partial \mathbf{C}\_f}{\partial \mathbf{x}} \right|\_{\mathbf{x}=0} = \frac{F}{A}, \quad \left( t > 0 \right) \tag{4}$$

$$\mathbf{C}\_{m}(\mathbf{x},0,t) = \mathbf{C}\_{f}(\mathbf{x},t), \quad \text{( $0 < x < L$ ,  $-t > 0$ )}\tag{5}$$

where, *F* is the inlet flux of radionuclide (mol/s) and A is cross sectional area of the flow channel (m2).

Other boundary conditions around the granite specimen were as shown in Eqs. (6) – (8):

$$\frac{\partial \mathbb{C}\_m(\mathbf{x}, d, t)}{\partial \mathbf{z}} = 0, \quad \left(0 < \mathbf{x} < L, \ t > 0\right) \tag{6}$$

$$\frac{\partial \mathbb{C}\_m(0, z, t)}{\partial \mathbf{x}} = 0, \quad \left(0 < z < d, \ t > 0\right) \tag{7}$$

Clarification of Adsorption Reversibility on Granite that Depends on Cesium Concentration 127

*st*

*Lx00,t st*

where, *α* is dispersion length (m). *Cst*(*x*, *t*), *Lst* and *Vst* are radionuclide concentration

Flow volume [×10-7 m3

Fig. 5. Experimental BTC values of 3H with the calibration curve using the results of the

The BTC of 3H, which is a non-sorbing species, is shown in Fig. 5. In a blank experimental run, we confirmed that 3H was not absorbed on the fluoroplastic tubing, thus the outlet concentration agreed with the inlet concentration (*Cst* / *C0* = 1). The transient region to the value of *Cst* / *C0* = 1 represented the effect of dispersion in the storage unit. The value of *α* was determined by a parameter identification method and *α* = 0.023 m was obtained. By substituting the values of *α* into Eq. (12), we obtained *DLst* = 3.0 ×10-5 m2/s for storage unit; thus this value was used for BTC analysis. Equations (2) and (3) were solved, subject to

We evaluated Cs adsorption amount on granite by BTCs. BTCs was obtained by injecting the Cs solution labeled with 134Cs into the flow channel at constant flow velocity. Figure 6 plots BTCs of 134Cs for various inlet Cs concentrations, showing the change in normalized effluent concentration *C*/*C0*, which is the effluent concentration *C* divided by the inlet

0 4 8 12

]

*x <sup>C</sup> -V*

*st*

(mol/m3), length (m) and flow velocity (m/s), in the storage unit, respectively.

*st*

*t C*

0.4

0.2

0.0

0.6

Effluent 3H concentration [-]

numerical analysis

Eqs.(4)-(8), using a finite volume scheme.

**4. Results and discussion** 

**4.1 Adsorption of 134Cs** 

0.8

1.0

1.2

*2 st 2*

*Vst0Lst DD α* (12)

(11)

*x <sup>C</sup> <sup>D</sup>*

 

*Lst*

Calibration curve


Table 4. BTCs calculation condition and parameters

$$\frac{\partial \mathbb{C}\_m(\mathcal{L}, z, t)}{\partial \mathbf{x}} = 0, \quad \left(0 < z < d, \ t > 0\right) \tag{8}$$

The calculation parameters are given in Table 4. The effective diffusion coefficient (*De*) was related to the diffusion coefficient in free water (*D0*) and the formation factor (*FF*) (Skagius, et al., 1982) as given by Eq.(9):

$$D\_{\varepsilon} = \varepsilon \cdot \frac{\delta}{\pi^2} \cdot D\_0 = FF \cdot D\_0 \tag{9}$$

where, *δ* is constrictivity ( - ) and *τ*2 is tortuosity ( - ). The value of *FF* for 3H was used instead of that for 134Cs because non-sorbing characteristics of 3H lead to direct determination of *FF* as shown in Eq. (10):

$$FF = \left(\frac{D\_e}{D\_0}\right)\_{\text{3}\_H} \tag{10}$$

The values of *De* and *D0* for 3H are 1.5 ×10-11 m2/s (Okuyama, et al., 2008) and 2.00 ×10-9 m2/s (Chemical Society of Japan, 2004), respectively. By substituting the values of *D0* for Cs as 2.04 ×10-9 m2/s (Chemical Society of Japan, 2004) into Eq. (9), we obtained *De* = 1.5 ×10-11 m2/s for Cs; we used this value for BTC analysis.

When the Cs-containing solution flowed through the flow channel and storage unit, it was dispersed in the longitudinal direction. The length of the flow channel (20 mm) was much shorter than that of the storage unit (7.6 m length), thus we neglected dispersion in the flow channel. In order to obtain longitudinal dispersion coefficient for the storage unit *DLst*, we calculated the value of the dispersion length *α* by analyzing the one-dimensional advectiondispersion equation as Eqs. (11) and (12):

$$\frac{\partial \mathbb{C}\_{\rm st}}{\partial t} = -V\_{\rm st} \frac{\partial \mathbb{C}\_{\rm st}}{\partial \mathbf{x}} + D\_{\rm Lst} \frac{\partial^2 \mathbb{C}\_{\rm st}}{\partial \mathbf{x}^2} \tag{11}$$

$$\begin{pmatrix} t > 0, & 0 < \mathbf{x} < L\_{\mathrm{st}} \end{pmatrix}$$

$$D\_{\mathrm{Lst}} = D\_0 + \mathbf{a} \cdot \mathbf{V}\_{\mathrm{st}} \tag{12}$$

where, *α* is dispersion length (m). *Cst*(*x*, *t*), *Lst* and *Vst* are radionuclide concentration (mol/m3), length (m) and flow velocity (m/s), in the storage unit, respectively.

Fig. 5. Experimental BTC values of 3H with the calibration curve using the results of the numerical analysis

The BTC of 3H, which is a non-sorbing species, is shown in Fig. 5. In a blank experimental run, we confirmed that 3H was not absorbed on the fluoroplastic tubing, thus the outlet concentration agreed with the inlet concentration (*Cst* / *C0* = 1). The transient region to the value of *Cst* / *C0* = 1 represented the effect of dispersion in the storage unit. The value of *α* was determined by a parameter identification method and *α* = 0.023 m was obtained. By substituting the values of *α* into Eq. (12), we obtained *DLst* = 3.0 ×10-5 m2/s for storage unit; thus this value was used for BTC analysis. Equations (2) and (3) were solved, subject to Eqs.(4)-(8), using a finite volume scheme.

#### **4. Results and discussion**

#### **4.1 Adsorption of 134Cs**

126 Municipal and Industrial Waste Disposal

*0, <sup>0</sup> <sup>z</sup> d, <sup>t</sup> <sup>0</sup>*

(8)

*Cm L, z,t*

The calculation parameters are given in Table 4. The effective diffusion coefficient (*De*) was related to the diffusion coefficient in free water (*D0*) and the formation factor (*FF*) (Skagius,

> *<sup>0</sup> <sup>0</sup> <sup>2</sup> <sup>e</sup> <sup>D</sup> FF <sup>D</sup> τ*

where, *δ* is constrictivity ( - ) and *τ*2 is tortuosity ( - ). The value of *FF* for 3H was used instead of that for 134Cs because non-sorbing characteristics of 3H lead to direct determination of *FF*

> 

The values of *De* and *D0* for 3H are 1.5 ×10-11 m2/s (Okuyama, et al., 2008) and 2.00 ×10-9 m2/s (Chemical Society of Japan, 2004), respectively. By substituting the values of *D0* for Cs as 2.04 ×10-9 m2/s (Chemical Society of Japan, 2004) into Eq. (9), we obtained

When the Cs-containing solution flowed through the flow channel and storage unit, it was dispersed in the longitudinal direction. The length of the flow channel (20 mm) was much shorter than that of the storage unit (7.6 m length), thus we neglected dispersion in the flow channel. In order to obtain longitudinal dispersion coefficient for the storage unit *DLst*, we calculated the value of the dispersion length *α* by analyzing the one-dimensional advection-

*0 H e D <sup>3</sup> <sup>D</sup> FF*

 

*<sup>δ</sup> <sup>D</sup> <sup>ε</sup>* (9)

(10)

Symbol Parameter Value A Cross sectional area of the flow channel 6.4 ×10-7 m2 *b* Depth of the flow channel 1.6 ×10-4 m *d* Thickness of granite specimen 3.0 ×10-3 m *De* Effective diffusion coefficient 1.5×10-11 m *DL* Longitudinal dispersion coefficient for the flow channel 0 m2/s *DLst* Longitudinal dispersion coefficient for the storage unit 3.0 ×10-5 m2/s *ε* Porosity of granite 0.73 % *L* Length of the flow channel 2.0 ×10-2 m *Lst* Length of the storage unit 7.6 m *t* Elapsed times 86400 s. *V* Flow velocity in the flow channel 2.6 ×10-5 m/s *Vst* Flow velocity in the storage unit 1.3 ×10-3 m/s

Table 4. BTCs calculation condition and parameters

et al., 1982) as given by Eq.(9):

as shown in Eq. (10):

*x*

*De* = 1.5 ×10-11 m2/s for Cs; we used this value for BTC analysis.

dispersion equation as Eqs. (11) and (12):

We evaluated Cs adsorption amount on granite by BTCs. BTCs was obtained by injecting the Cs solution labeled with 134Cs into the flow channel at constant flow velocity. Figure 6 plots BTCs of 134Cs for various inlet Cs concentrations, showing the change in normalized effluent concentration *C*/*C0*, which is the effluent concentration *C* divided by the inlet

Clarification of Adsorption Reversibility on Granite that Depends on Cesium Concentration 129

In order to verify this estimation of sorption amount, we determined the retardation coefficient by parameter identification method for BTCs using Eqs. (2), (3), and (11), and compared it with the retardation coefficient (*Rm*) calculated using the value of *Kd* obtained by a batch sorption experiment. By inserting *ε*, *Kd*, and density of granite *ρ* (2.98 ×103 kg/m3)

The *Rm* values obtained by the two methods were almost identical, therefore the area enclosed by the lines of *C*/*C0* = 1 and BTC represented the Cs diffusion into the granite

Analyzing BTC values

10-6 10-3 100 103

Cs concentration in solution [mol/m3]

After injecting 134Cs solution, extraction reagents (HCl, CaCl2, and KCl solutions as shown in Table 3) were injected into the flow channel in sequence. We obtained the desorption curves and investigated the chemical speciation of Cs in granite, in particular the ratio of the density of reversible and irreversible adsorption sites depends on Cs concentration in solution. Desorption curves are shown in Fig. 8 as the change in normalized desorbed Cs divided by the total adsorption amount. For each curve, desorption amount of Cs was significantly decreased to 1 % or lower when elapsed time was 24, 48, and 72 hours, so Cs

Fig. 7. Retardation coefficients obtained by analyzing BTC values and batch sorption

matrix or the Cs absorbed on the rock surface (Okuyama, et al., 2008).

10-8

10-6

10-4

Retardation coefficient

experiment results

**4.2 Desorption of 134Cs** 

*R*

*m* [-]

10-2

100

*ε ρ Kd*

Batch sorption experiment results

*<sup>m</sup> 1R* (13)

into Eq.(13), *Rm* were calculated.

concentration *C0*, as a function of elapsed time. A quasi-plateau region was observed for each curve and the breakthrough values in this region did not reach the equilibrium *C*/*C*0 = 1. This is because some of the Cs may be still driven away by diffusion into the granite matrix or they were absorbed on the rock surface. From the mass balance standpoint, we assumed that the area enclosed by the lines of *C*/*C0* = 1 and the BTC represented the Cs diffusion into the granite matrix or the Cs absorbed on the granite surface.

Fig. 6. Experimental BTCs of 134Cs for various inlet Cs concentrations with a curve using the result of the numerical analysis

In order to verify this estimation of sorption amount, we determined the retardation coefficient by parameter identification method for BTCs using Eqs. (2), (3), and (11), and compared it with the retardation coefficient (*Rm*) calculated using the value of *Kd* obtained by a batch sorption experiment. By inserting *ε*, *Kd*, and density of granite *ρ* (2.98 ×103 kg/m3) into Eq.(13), *Rm* were calculated.

$$R\_m = 1 + \frac{\rho \cdot \text{Kd}}{\varepsilon} \tag{13}$$

The *Rm* values obtained by the two methods were almost identical, therefore the area enclosed by the lines of *C*/*C0* = 1 and BTC represented the Cs diffusion into the granite matrix or the Cs absorbed on the rock surface (Okuyama, et al., 2008).

Fig. 7. Retardation coefficients obtained by analyzing BTC values and batch sorption experiment results

#### **4.2 Desorption of 134Cs**

128 Municipal and Industrial Waste Disposal

concentration *C0*, as a function of elapsed time. A quasi-plateau region was observed for each curve and the breakthrough values in this region did not reach the equilibrium *C*/*C*0 = 1. This is because some of the Cs may be still driven away by diffusion into the granite matrix or they were absorbed on the rock surface. From the mass balance standpoint, we assumed that the area enclosed by the lines of *C*/*C0* = 1 and the BTC represented the Cs

0.4

divided by

*C0* [-]

0.2

0.0

0.4

divided by

*C0* [-]

0.2

0.0

0.6

Effluent 134Cs concentration

Fig. 6. Experimental BTCs of 134Cs for various inlet Cs concentrations with a curve using the

0.8

1.0

(b) *C0* = 1.0 ×10-1 mol/m3

50 25201510

Calibration curve

Time [h]

(d) *C0* = 1.0 ×10-3 mol/m3

Time [h]

50 1510 20 25

Calibration curve

0.6

Effluent 134Cs concentration

0.8

1.0

diffusion into the granite matrix or the Cs absorbed on the granite surface.

0.4

0.2

0.0

0.4

0.2

0.0

result of the numerical analysis

0.6

Effluent 134Cs concentration

divided by

*C0* [-] 0.8

1.0

(a) *C0* = 1.0 ×100 mol/m3

Time [h] 50 201510 25

Calibration curve

(c) *C0* = 1.0 ×10-2 mol/m3

Time [h]

50 10 15 2520

Calibration curve

divided by

*C0* [-]

0.6

Effluent 134Cs concentration

0.8

1.0

After injecting 134Cs solution, extraction reagents (HCl, CaCl2, and KCl solutions as shown in Table 3) were injected into the flow channel in sequence. We obtained the desorption curves and investigated the chemical speciation of Cs in granite, in particular the ratio of the density of reversible and irreversible adsorption sites depends on Cs concentration in solution. Desorption curves are shown in Fig. 8 as the change in normalized desorbed Cs divided by the total adsorption amount. For each curve, desorption amount of Cs was significantly decreased to 1 % or lower when elapsed time was 24, 48, and 72 hours, so Cs

Clarification of Adsorption Reversibility on Granite that Depends on Cesium Concentration 131

10-11

10-11

(c) KCl (d) Residue

Desorption amount of Cs at various quasi-plateau Cs concentration of BTCs is summarized in Fig. 9. The fractions extracted with HCl, KCl and CaCl2 gave straight lines and the slopes were about 0.7, which could be described by a Freundlich isotherm. On the other hand, the residue fraction was directly proportional to the Cs concentration at low Cs concentration (lower than 1 × 10-1 mol/m3), and showed signs of leveling off at 1 × 100 mol/m3; thus Cs adsorption on granite surface may become saturated at Cs concentration of 1 ×100 mol/m3. The chemical speciation of Cs desorbed by each regent was described as follows. Biotite has two types of adsorption sites; variable-charge sites, and permanent-charge sites. The charge of variable-charge sites changes with changing pH, thus the fraction of Cs extracted with HCl indicates adsorption amount on variable-charge sites. On the other hand, it has been

Fig. 9. Desorbed amounts of Cs at various quasi-plateau Cs concentration of BTCs

10-10

10-9

Slope = 1.0 10-8

Slope = 0.11

*C0* [mol/m3]

10-3 101 10-1 10-5

*C0* [mol/m3]

10-3 101 10-1 10-5

Residue Cs [mol]

10-6

10-7

10-10

10-9

10-8 Slope = 0.72

Extracted Cs by CaCl2

(a) HCl (b) CaCl2

 [mol] 10-6

10-7

10-11

10-6

10-7

10-11

10-10

10-9

Extracted Cs by KCl [mol]

10-10

10-9

10 Slope = 0.72 -8

10 Slope = 0.73 -8

10-3 101 10-1 10-5

*C0* [mol/m3]

*C0* [mol/m3]

10-3 101 10-1 10-5

Extracted Cs by HCl [mol]

10-6

10-7

was adequately desorbed by each reagent. Desorption amount of Cs changed drastically between the reagents, thus it was conceivable that desorbed Cs for each reagent indicated the presence of multiple adsorption sites.

Fig. 8. Experimental desorption curves of 134Cs for various inlet Cs concentrations

was adequately desorbed by each reagent. Desorption amount of Cs changed drastically between the reagents, thus it was conceivable that desorbed Cs for each reagent indicated

(b) *C0* = 1.0 ×10-1 mol/m3

Time [h]

CaCl2 (24 h)

KCl (24 h)

HCl (24 h)

10

0

10

0

Desorbed 134Cs [%]

20

Desorbed 134Cs [%]

20

HCl (24 h)

0 24 48 72

CaCl2 (24 h)

KCl (24 h)

(d) *C0* = 1.0 ×10-3 mol/m3

Time [h]

0 24 48 72

the presence of multiple adsorption sites.

10

0

10

0

Desorbed 134Cs [%]

20

Desorbed 134Cs [%]

20

HCl (24 h)

HCl (24 h)

0 24 48 72

CaCl2 (24 h)

KCl (24 h)

Time [h]

(a) *C0* = 1.0 ×100 mol/m3

CaCl2 (24 h)

KCl (24 h)

Fig. 8. Experimental desorption curves of 134Cs for various inlet Cs concentrations

(c) *C0* = 1.0 ×10-2 mol/m3

Time [h]

0 24 48 72

Fig. 9. Desorbed amounts of Cs at various quasi-plateau Cs concentration of BTCs

Desorption amount of Cs at various quasi-plateau Cs concentration of BTCs is summarized in Fig. 9. The fractions extracted with HCl, KCl and CaCl2 gave straight lines and the slopes were about 0.7, which could be described by a Freundlich isotherm. On the other hand, the residue fraction was directly proportional to the Cs concentration at low Cs concentration (lower than 1 × 10-1 mol/m3), and showed signs of leveling off at 1 × 100 mol/m3; thus Cs adsorption on granite surface may become saturated at Cs concentration of 1 ×100 mol/m3.

The chemical speciation of Cs desorbed by each regent was described as follows. Biotite has two types of adsorption sites; variable-charge sites, and permanent-charge sites. The charge of variable-charge sites changes with changing pH, thus the fraction of Cs extracted with HCl indicates adsorption amount on variable-charge sites. On the other hand, it has been

Clarification of Adsorption Reversibility on Granite that Depends on Cesium Concentration 133

When the irreversible sites are saturated, the Cs will no longer be retarded by them. Thus, it is concluded that the ratio of the density of reversible and irreversible adsorption sites must be thoroughly considered to demonstrate the safety of HLW deep underground disposal

We carried out adsorption and sequential extraction experiments using three chemical reagents (HCl, CaCl2, and KCl solutions) for various inlet Cs concentrations to investigate the chemical speciation of Cs in granite. In particular, we focused on the dependence of adsorption amount on reversible and irreversible sites on Cs concentration. The sum of fractions extracted with HCl, CaCl2 and KCl, which indicated reversible adsorption, was 60 - 80 %. Irreversible adsorption was relatively large at high Cs concentration. However, Cs adsorption became saturated, and the fraction of irreversible adsorption decreased. Cs adsorption on granite surface was saturated at Cs concentration of 1 ×100 mol/m3. In the early stage of radionuclide leakage, Cs will be retarded by the presence of reversible and irreversible adsorption sites. However, when the irreversible sites were saturated, Cs will not be retarded by them. Thus, the sorption reversibility must be thoroughly considered to demonstrate the behavior of Cs in granite, in particular for high

Brown, D. L., Haines, R. I., Owen, D. G., Stanchell, F. W. & Watson, D. G. (1984). Surface

Cornell, R. M. (1993). Adsorption of cesium on minerals: A review. *J. Radioanalytical and* 

Francis, C. W. & Brinkley F. S. (1976). Preferential adsorption of 137Cs to micaceous minerals

Fukui, M. (2004). Affinity of trace elements to sandy soil and factors affecting on the

Holland, T. R. & Lee, D. J. (1992). Radionuclide getters in cement. *Cement and Concrete* 

Chemical Society of Japan. (Ed.). (2004). *Kagaku Binran, Basic 5th ed.,* Maruzen Company,

Ohe, T. (1984). Ion exchange adsorption of radioactive cesium, cobalt, manganese, and

Okuyama, K., Sasahira, A., Noshita, K., & Ohe, T. (2008). A method for determining both

Peech, M. (1965). Exchange acidity: Barium chloride-triethanolamine method. *Methods of* 

micro-reactor technique. *Applied Geochemistry*, 23, pp. 2130-2136

strontium to granitoid rocks in the presence of competing cations. *J. Nucl. Sci.* 

diffusion and sorption coefficients of rock medium within a few days by adopting a

*soil analysis*, *Part 2*, American Society of Agronomy, Madison, Wisconsin, pp. 910–

in contaminated freshwater sediment. *Nature*, 260, pp. 511-513

migration through media. *KURRI KR*, 99, pp. 21-26

Ohtaki, H. (1990). Hydration of Ions. *Kyoritsu Shuppan,* Tokyo, pp. 55

studies of the interaction of cesium with feldspars. *American chemical society*, 246,

option.

**5. Conclusion** 

Cs concentration.

**6. References** 

pp. 217-227

*Nuclear Chemistry*, 171, 2, pp. 483-500

*Research*, 22, pp. 247-258

*Technol.*, 67, 1, pp. 92-101

Tokyo, pp. 64-65

911

reported that Cs is strongly absorbed on biotite grains; in particular it is distributed onto the interlayers of these grains2), which is permanent-charge sites. The hydrated ionic radius of K+ (0.13 nm) (Ohtaki, 1990) are almost identical to that of Cs+ (0.12 nm) (Ohtaki, 1990), thus Cs+ which has been taken into the interlayers induces interlayer spacing that is large enough to permit diffusion of K+; thus K+ would displace adsorbed Cs on the biotite interlayer permanent-charge sites. The hydrated ionic radius of Ca2+ (0.31 nm) (Ohtaki, 1990) is much larger than that of Cs+ and Ca2+ diffusion into the interlayers is restricted (Cornell, 1993); thus Ca2+ would displace adsorbed Cs on the edge of the biotite interlayer or at other mineral sites such as feldspar (Brown, et al., 1984). From the above, although the chemical speciation of Cs desorbed for the three regents was different, the adsorption mechanism of Cs for each was an ion-exchange reaction. The adsorption mechanism of the residue fraction might be fixed in biotite2). In this study, we defined that as irreversible adsorption.

The fraction of Cs desorbed by the three reagents is shown in Fig. 10. The sum of fractions extracted with HCl, CaCl2 and KCl, which indicated reversible adsorption, was 60 - 80 %. Irreversible adsorption was relatively large at high Cs concentration. However, Cs adsorption was saturated at 1 ×100 mol/m3, and the fraction of irreversible adsorption decreased.

Fig. 10. Fraction of Cs desorbed by three extraction reagents (HCl, CaCl2, KCl)

When the irreversible sites are saturated, the Cs will no longer be retarded by them. Thus, it is concluded that the ratio of the density of reversible and irreversible adsorption sites must be thoroughly considered to demonstrate the safety of HLW deep underground disposal option.

#### **5. Conclusion**

132 Municipal and Industrial Waste Disposal

reported that Cs is strongly absorbed on biotite grains; in particular it is distributed onto the interlayers of these grains2), which is permanent-charge sites. The hydrated ionic radius of K+ (0.13 nm) (Ohtaki, 1990) are almost identical to that of Cs+ (0.12 nm) (Ohtaki, 1990), thus Cs+ which has been taken into the interlayers induces interlayer spacing that is large enough to permit diffusion of K+; thus K+ would displace adsorbed Cs on the biotite interlayer permanent-charge sites. The hydrated ionic radius of Ca2+ (0.31 nm) (Ohtaki, 1990) is much larger than that of Cs+ and Ca2+ diffusion into the interlayers is restricted (Cornell, 1993); thus Ca2+ would displace adsorbed Cs on the edge of the biotite interlayer or at other mineral sites such as feldspar (Brown, et al., 1984). From the above, although the chemical speciation of Cs desorbed for the three regents was different, the adsorption mechanism of Cs for each was an ion-exchange reaction. The adsorption mechanism of the residue fraction

might be fixed in biotite2). In this study, we defined that as irreversible adsorption.

HCl CaCl2

KCl

1.0 × 10-3 1.0 × 10-2

Fig. 10. Fraction of Cs desorbed by three extraction reagents (HCl, CaCl2, KCl)

*C0* [mol/m3

]

1.0 × 10-1 1.0× 100

0

20

40

60

80

Extracted Cs [%]

100

The fraction of Cs desorbed by the three reagents is shown in Fig. 10. The sum of fractions extracted with HCl, CaCl2 and KCl, which indicated reversible adsorption, was 60 - 80 %. Irreversible adsorption was relatively large at high Cs concentration. However, Cs adsorption was saturated at 1 ×100 mol/m3, and the fraction of irreversible adsorption decreased.

Residue : Irreversible adsorption

Reversible adsorption

We carried out adsorption and sequential extraction experiments using three chemical reagents (HCl, CaCl2, and KCl solutions) for various inlet Cs concentrations to investigate the chemical speciation of Cs in granite. In particular, we focused on the dependence of adsorption amount on reversible and irreversible sites on Cs concentration. The sum of fractions extracted with HCl, CaCl2 and KCl, which indicated reversible adsorption, was 60 - 80 %. Irreversible adsorption was relatively large at high Cs concentration. However, Cs adsorption became saturated, and the fraction of irreversible adsorption decreased. Cs adsorption on granite surface was saturated at Cs concentration of 1 ×100 mol/m3. In the early stage of radionuclide leakage, Cs will be retarded by the presence of reversible and irreversible adsorption sites. However, when the irreversible sites were saturated, Cs will not be retarded by them. Thus, the sorption reversibility must be thoroughly considered to demonstrate the behavior of Cs in granite, in particular for high Cs concentration.

#### **6. References**


**7** 

**Removal of Selected Benzothiazols with Ozone** 

Benzothiazole derivatives are widely used as industrial chemicals in the leather and wood industries, as bio-corrosion inhibitors in cooling systems, ingredients in anti-freezing agents for automobiles, and mainly as vulcanisation accelerators in rubber production. Correspondingly, these xenobiotic compounds are widely distributed in the environment and they have been detected in industrial wastewaters, as well as in soils, estuarine sediments, and superficial waters (Valdés & Zahor, 2006). They cause environmental concern when released into watercourses (Valdés et al., 2008). These compounds inhibit micro-organisms' activity in conventional biological wastewater treatment systems and most of them are not readily biodegradable (de Wewer & Verachter, 1997; de Wewer, 2007). Moreover, they can be absorbed onto cell membranes, leading to bioaccumulation (Gaja & Knapp, 1998). According to Knapp et al. (1982), 7.0 mg.l-1 BT causes a 50% and 54 mg.l-1 a 100% inhibition of ammonia oxidation, while nitrate utilization is not affected (de Wever and Verachter, 1997). Tomslino et al. (1966) proved 75% inhibition of ammonium nitrogen oxidation at the concentration of 3 mg.l-1 MBT. HOBT causes 100%

Unfortunately, conventional biological wastewater treatment processes are not able to effectively remove such contaminants since these are resistant to biodegradation (Valdés & Zahor, 2006). Thus, the development of efficient treatment/pre-treatment processes is required in order to eliminate their discharge into the aquatic ecosystem. Advanced oxidation processes (AOPs) might be a viable option for the decontamination of biologically recalcitrant wastewaters (Kralchevska et al., 2010). An important group of AOPs are ozone based oxidation procesess, e.g. ozonation at elevated pH, combinations of ozone with UV,

Experimental part of this work was focused on the removal of benzothiazole derivatives by ozone. The results of ozonation trials carried out with the model wastewaters containing single MBT and BT pollutants, the mixture of BT and MBT, the mixtuere of benzothiazole derivatives contained in an industrial wastewater from sulfonamides production as well as with the real wastewater are presented. The conventional ozonation process was

inhibition of oxidation of ammonium nitrogen at 60 mg.l-1 (Hauck, 1972).

**1. Introduction** 

hydrogen peroxide etc.

Jan Derco1, Michal Melicher1 and Angelika Kassai2 *1Institute of Chemical and Environmental Engineering,* 

*Slovak University of Technology, Bratislava,* 

*2Water Research Institute, Bratislava,* 

*Slovak Republic* 

