Model analysis of the colloid and radionuclide retardation experiment at the Grimsel Test Site

Journal of Colloid and Interface Science 298 (2006) 467–475 www.elsevier.com/locate/jcis

Model analysis of the colloid and radionuclide retardation experiment at the Grimsel Test Site Susumu Kurosawa a,∗ , Scott C. James b , Mikazu Yui a , Motomu Ibaraki c a Japan Atomic Energy Agency (successor of Japan Nuclear Cycle Development Institute (JNC)), Tokai-mura, Naka-gun, Ibaraki 319-1194, Japan b Sandia National Laboratories, PO Box 5800, Albuquerque, NM 87185-0735, USA c The Ohio State University, Columbus, OH 43210-1308, USA

Received 28 September 2005; accepted 17 December 2005 Available online 19 January 2006

Abstract The colloid and radionuclide retardation experiments performed at NAGRA’s Grimsel Test Site in Switzerland are part of an international collaboration program designed to collect in situ data on the impacts of colloids on radionuclide transport. In this work, breakthrough behaviors of trivalent americium (i.e., 241 Am and 243 Am) both in the absence and presence of bentonite colloids are analyzed with COLFRAC—a code that models colloid-facilitated solute transport in discretely-fractured, porous media. Model fits to the experimental results indicate that Am sorbed onto mobile colloids, which enhance Am transport relative to a non-sorbing tracer, 131 I. Modelling results suggest that Am is kinetically sorbed onto both naturally occurring and exogenous bentonite colloids. Results also indicate that desorption of Am from colloids is slow with respect to the duration of the experiment. In addition, early colloid breakthrough compared to a conservative tracer suggests the effects of hydrodynamic chromatography. Overall, Am breakthrough curves suggest enhanced mobility due to co-transport with both naturally occurring and bentonite colloids. © 2005 Elsevier Inc. All rights reserved. Keywords: Radionuclide and colloid transport; In situ experiment; Fractured media; Model analysis; Kinetic sorption reaction

1. Introduction Prior to the late 1980s, studies of the saturated zone modelled contaminant transport as a two-phase system, where contaminants were either mobile with their movement defined by the flow of groundwater, or motionless because of sorption onto the matrix. McCarthy and Zachara [1] suggested that a third phase exists that may help to better describe and predict contaminant movement in the subsurface. This third phase is composed of mobile colloids, often with composition similar to the local media. Aqueous contaminants are capable of reacting with colloids thereby shielding these contaminants from stabilization through natural attenuation and retardation mechanisms such as matrix diffusion and sorption. Evaluation of the effects of colloid particles on radionuclide transport in ground* Corresponding author. Fax. +81 29 282 9328.

E-mail address: [email protected] (S. Kurosawa). 0021-9797/$ – see front matter © 2005 Elsevier Inc. All rights reserved. doi:10.1016/j.jcis.2005.12.036

water is an important issue in performance assessment (PA) for geological disposal of high-level radioactive waste (HLW). Laboratory studies have been carried out with combinations of radionuclides and colloids in a fractured-rock medium to experimentally determine the influence of colloids on radionuclide transport [2–4]. Numerical PA models have been developed since the early 1990s to analyze radionuclide and colloid co-transport in fractured, porous media [5–7]. Abdel-Salam and Chrysikopoulos [8,9] have also studied the various physicochemical processes that contribute to contaminant transport in the presence of colloids and reported conditions for which colloids enhance contaminant mobility and increase contaminant dispersion. Unfortunately, there has been little modelling of in situ experiments. This work seeks to better understand the mechanisms involved in in situ radionuclide and colloid cotransport by fitting model results to experimental data. Experiments for the colloid and radionuclide retardation (CRR) project began in 1998 to study the in situ transport behavior of actinides and fission products in the absence and

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presence of exogenous bentonite colloids in a water-conducting fracture (shear zone) at NAGRA’s Grimsel Test Site (GTS) in Switzerland. The CRR experiments are the result of an international collaborative research program funded by NAGRA (Switzerland), ENRESA (Spain), ANDRA (France), FZK-INE (Germany), USDOE/SNL (USA) and JNC (Japan) [10–13]. The experiments specifically study the effects of both natural and bentonite colloids on the transport behavior of radioisotopes: U, Th, Pu, Am, Np, Sr, Cs, I, and Tc. Modelling colloid-facilitated radionuclide transport was one of the key goals of the CRR project. As part of this project, the primary objectives of this work are to investigate the parameters controlling sorption of radionuclides onto colloids and to describe any corresponding colloid-facilitated radionuclide transport in this fractured, porous media. The dual permeability model, COLFRAC [14,15], is used to solve the flow and transport equations in two dimensions. This model was selected because it incorporates advective–dispersive–reactive aqueous phase radionuclide transport in the fractures and the rock matrix, colloid transport in fractures within the experimental shear zone, and sorption/desorption of radionuclides onto colloids. The numerical formulation for radionuclide sorption allows for either equilibrium or kinetic sorption reactions onto colloids according to the Henry isotherm [16]. Although several models are available to model colloid facilitated contaminant transport, COLFRAC was selected because it includes all relevant process to be modelled here and because it had been used successfully in the past [17]. 2. CRR experiments 2.1. Experiment overview [10–13] The CRR experiments were carried out at NAGRA’s GTS underground rock laboratory, which is located 1730 m above the sea level under a 450 m thick overburden of crystalline rock in the central Swiss Alps. The test site groundwater is a 2− Na+ /Ca2+ –HCO− 3 –SO4 type with a pH of 9.6 and Eh below −300 mV. The electrical conductivity was 103 µS/cm and the ionic strength was 1.2 × 10−3 M. The hydraulic and tracer migration experiments were performed in a shear zone at the site. The layout for the CRR experiments is illustrated in Fig. 1. A dipole flow field was established between two boreholes separated by a linear distance of 2.23 m. The test shear zone was characterized for the CRR experiments as well as for other experimental tracer migration experiments [18–21]. A dipole flow field for tracer test runs was established between the injection borehole (10 ml/min) and the extraction borehole (150 ml/min). The induced hydraulic gradient is coincident with that of the background. The boreholes were equipped with multi-packer systems and flow was induced with an HPLC pump and measured with a magnetic inductive flow meter. In the CRR experiments, two tracer injections (runs) were performed with 131 I, 85 Sr, 232 Th, 238 U, 237 Np, 238,242 Pu, and 243 Am in the absence of exogenous colloids and 131 I, 85 Sr, 137 Cs, 99 Tc, 232 Th, 233 U, 237 Np, 238,244 Pu, and 241 Am in the presence of 20 mg/l of exogenous bentonite colloids. The ben-

Table 1 Radionuclide recoveries in the CRR experiments [10,13] Recovery (%) Run #1 (without bentonite colloids) 131 I 85 Sr 232 Th 238 U 237 Np 238 Pu 242 Pu 243 Am

100 87
Run #2 (with bentonite colloids) 131 I 85 Sr 137 Cs 99 Tc 232 Th 233 U 237 Np 238 Pu 244 Pu 241 Am Bentonite colloids

92 88 70
Note. D.L.: detection limit.

tonite colloids used in this experiment are being considered for use in an engineered barrier for HLW disposal. They were prepared from a crushed natural bentonite from Spain called FEBEX [22]. The major mineral phase (92%) of the FEBEX bentonite is a calcium montmorillonite. The bentonite material was sieved (dry size fraction <64 µm) and washed with Milli-Q deionized water and finally equilibrated with GGW. Bentonite colloids were collected from the sediment of five consecutive centrifugation steps (4000 rpm for 40 min). The radionuclide mixtures were prepared by re-suspending the colloids in previously-collected GGW and then adding the radionuclides under anoxic conditions (in an Ar glovebox with <1 ppm O2 ). In these experiments, 100 ml of each mixture were injected into the test shear zone. Radionuclide concentrations in the extracted groundwater were measured by ICP mass spectrometry (ICPMS) and α- and γ -spectrometry. The bentonite colloids in the extracted groundwaters were analyzed by laser induced breakdown detection (LIBD). 2.2. Breakthrough curves of radionuclides Radionuclide breakthrough curves for run #1 (no bentonite colloids) and run #2 (added bentonite colloids) are plotted as functions of elution time in Figs. 2A and 2B, respectively. Fig. 2B includes the breakthrough curve for bentonite colloids. Table 1 summarizes radionuclide recoveries for each experimental run. Note that recovery of the tri- and tetravalent actinides (Am and Pu) in run #2 is at least twice that observed in run #1. Fig. 2A shows that the peak maxima (indicated by the vertical black bars) for tri- and tetra-valent Am, Pu, and Th arrive earlier than that for the non-sorbing tracer, 131 I, in run #1. This could be interpreted as colloid facilitated transport via formation of pseudocolloids with natural colloids in GGW

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469

Fig. 1. Schematic of the CRR experiments (modified from [10,13]).

Fig. 2. Breakthrough curves for radionuclides in CRR experiments (modified from [10,13]).

or the formation of homogeneous actinide colloids. Fig. 2B reveals that the peaks of the breakthrough curve for the bentonite colloids and those for tri- and tetra-valent Am, Pu, and Th are also shifted to an earlier peak elution time compared to 131 I in run #2. Along with the increased recovery rates, this indicates that the migration behavior of the tri- and tetra-valent actinides was influenced by the bentonite colloids. Interpretation of these results is detailed by Möri [10] and Möri et al. [13]. 3. Stabilities of radionuclides and colloids For both mixtures, Table 2 shows the injected activity, mass concentration, colloidal fraction (fraction of the radionuclide concentration associated with colloids), and solubilities calculated for each radioelement in these experiments. The presumed solid phase (precipitate or solubility limiting solid) of each radionuclide is also reported [23]. Note that the concentrations of Th, Tc, U, Np, and Pu in the tracer mixtures are greater than their calculated solubilities in GGW. This contradiction is explained by the hypothesis that homogeneous or heterogeneous radiocolloids (pseudocolloids) were formed by the radionuclides in the mixtures. Homogeneous radiocolloids are composed of radionuclide species such as PuO2 or polymerised actinide hydroxides. Pseudocolloids are formed by the sorption of a radionuclide onto a colloid. Also of note is that the concentrations of Am are less than its solubility limit. The low concentrations of Am may be due to the formation of pseudo-

colloids that are not counted. To measure the colloidal fraction in each injectant, the mixtures were subjected to ultracentrifugation for 1 h at 90,000 rpm. For Am, the colloidal fractions may be artificially low due to the presence of particles that cannot be sedimented by centrifugation (e.g., exceptionally small colloids, low-density colloids, or gel-like aggregates). Natural GGW colloids have an average size of 202 ± 12 nm and the average size of the bentonite colloids was 116 ± 12 nm. Both the natural and bentonite colloid populations in the GGW measured by LIBD showed no significant changes in either concentration or average size over the course of the experiments [10,11,13]. 4. Model analysis Only Am breakthrough curves are analyzed in this work because the sorption behavior of Pu, Np, U, and Tc are sensitive to oxidation state and the solid–liquid distribution coefficients (Kd) obtained by sorption experiments may be different under the reducing conditions of the in situ migration experiments. In addition, actinides (with the exception of Am) may form homogeneous colloids because their injected concentrations in these experiments are greater than the reported solubility limit of GGW. Finally, the transport behavior of Sr and Cs may be explained by sorption onto fracture surfaces without significant interaction with colloids [24].

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Table 2 Radioisotope activity, mass concentration, and assumed oxidation states in radionuclide mixtures for run #1 and run #2 [10,13] In situ experiments

Calculations

Injected radionuclide Radioactivity: M0 (Bq for 100 ml)

Concentration: C0 (mol/l)

Colloidal fraction (%)

Solid phase

Calculated solubility (mol/l)

Run #1 (without bentonite colloids) 131 I 7.46 × 104 85 Sr 9.52 × 104 232 Th 1.06 × 10−3 238 U 2.82 × 10−1 237 Np 5.82 × 102 238 Pu 6.70 × 102 242 Pu 3.50 × 101 243 Am 1.06 × 103

1.24 × 10−12 1.28 × 10−11 1.12 × 10−8 9.50 × 10−7 9.44 × 10−7 4.44 × 10−11 9.94 × 10−9 5.93 × 10−9

0 0 20–30 0–12 0-10 5–58 5–58 6–58

– – Th(OH)4 (am) U(IV) (solid) Np(OH)4 (am) Pu(OH)4 (am)

– – 10−8 10−8 10−8 10−10

AmOHCO3 (am)

10−7

Run #2 (including 20 mg/l FEBEX bentonite colloids) 131 I 5.56 × 104 85 Sr 8.24 × 104 137 Cs 6.07 × 105 99 Tc 6.55 × 101 232 Th 1.03 × 10−3 233 U 7.22 × 103 237 Np 6.72 × 102 238 Pu 7.20 × 102 241 Am 2.04 × 103 244 Pu 1.11 × 10−1

9.23 × 10−13 1.11 × 10−11 1.38 × 10−8 1.04 × 10−8 1.10 × 10−8 8.69 × 10−7 1.09 × 10−6 4.77 × 10−11 1.15 × 10−8 6.70 × 10−9

0 0 8 12 94 6 0–1 84 99 84

– – – TcO2 Th(OH)4 (am) U(IV) (solid) Np(OH)4 (am) Pu(OH)4 (am) AmOHCO3 (am) Pu(OH)4 (am)

– – – 10−8 10−8 10−8 10−8 10−10 10−7 10−10

Note. –: Not reported in [10] or [13]; am: amorphous.

and apertures, while the many competing transport phenomena modeled can complicate the parameter estimation process.

4.1. Model COLFRAC is a two-dimensional, finite element model that uses the standard Galerkin scheme to solve the governing equations for groundwater flow and dissolved solute and colloid transport [14,15]. COLFRAC describes the sorption of radionuclides onto colloids using either an equilibrium or kinetic reaction according to Henry isotherms [16], which may influence colloid-facilitated radionuclide transport. For an equilibrium Henry isotherm, sorbed radionuclide mass is described as: Sc = Kdc × Cf ,

(1)

where Sc is the amount of radionuclide on the colloids, Kdc is the distribution coefficient for the colloids, and Cf is the concentration of the radionuclide in the fracture. For a kinetic Henry isotherm, the rate of change of sorbed radionuclide mass is: ∂Sc (2) = Ks(Kdc × Cf − Sc ), ∂t where Ks is the kinetic rate constant for the colloids. In COLFRAC, the preceding equation is re-arranged to yield: ∂Sc (3) = α × Cf − β × Sc , ∂t where α = Ks×Kdc is the forward kinetic reaction rate constant and β = Ks is the backward kinetic reaction rate constant. COLFRAC allows only monodisperse (single-size) colloid plumes. Also, it should be noted that breakthrough curves may be sensitive to the fracture geometry, number of fractures, orientation,

4.2. Fracture geometry NAGRA’s investigations revealed that the shear zone is a reactivated mylonite in a weakly foliated granodiorite. The rock is brittle and a heavily brecciated shear zone extends around it for up to 30 cm [11]. Within the shear zone, water flow is concentrated in up to six channels, each 0.1–4 mm wide, which are predominantly situated in mylonite (porosity 0.1–0.4%) [11]. General bulk hydraulic conductivity values of 10−10 –10−9 m/s for the shear zone and 10−12 –10−11 m/s for the rock matrix were measured at the GTS [21]. The parameters estimated by COLFRAC are grouped into two categories: (1) those related to the geometry of the system that are independent of the tracer runs (e.g., groundwater flow channels, fracture aperture, porosity, and hydraulic conductivity) and (2) those related to the transported constituents (e.g., distribution coefficients, kinetic reaction rates, and colloid velocities). The results from COLFRAC are dependent upon the specified model geometry. Fig. 3 illustrates the geometry of the fracture system used in all COLFRAC runs. This geometry was determined through trial and error to best match the characteristics and shape of the breakthrough curve for the nonreactive tracer, 131 I. Fig. 4 illustrates COLFRAC breakthrough curves using fitted geometry parameters and the experimental data for 131 I. Table 3 lists estimated geometry parameters for the test shear zone that yielded the best fit to the 131 I breakthrough curve. Underlined values in Table 3 were reported by NAGRA and are characteristic of the GTS [11,21].

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Fig. 3. Geometry of the fracture system used for the COLFRAC code.

Fig. 4. COLFRAC fit to 131 I CRR experimental data (run #1). Table 3 Geometry data for the shear zone used in COLFRAC Geometry horizontal fractures fast flow slow flow vertical fractures

6 channels [11] 2 channels (aperture 1.1 mm) 4 channels (aperture 0.7 mm) 9 channels (aperture 1.0 mm)

Porosity

0.4% [11]

Hydraulic conductivity shear zone matrix

10−9 m/s [21] 10−11 m/s [21]

Table 4 Kd values for Am sorption onto granodiorite, bentonite colloids, and natural colloids in GGW

241 Am, 243 Am

Granodioritea (m3 /kg)

Bentonite colloidsb (m3 /kg)

Natural colloidsa (m3 /kg)

1.3 × 10−1 [11]

2.1 × 103 [11]

6.4 × 105

a Contact time: 2 weeks. b Contact time: 1 week.

4.3. Sorption data for radionuclides used in the analysis [11] Sorption of the CRR-relevant radioelements onto Grimsel granodiorite and bentonite colloids was investigated experimentally. The Kd of radionuclides have been measured and

471

used in model analyses of the radionuclide transport experiments and may be used for subsequent predictive modelling of the in situ migration experiments. The sorption experiments were carried out as batch experiments in an anoxic N2 atmosphere (O2 < 1 ppm) by two laboratories (FZK-INE and Ciemat, Spain) to simulate the GTS environment. The laboratories investigated different radionuclides: FZK-INE focused on the sorption of the actinides 241,243 Am, 237 Np, and 238,244 Pu onto Grimsel granodiorite and bentonite colloids and Ciemat studied the sorption behavior of 137 Cs, 233 U, and 99 Tc onto Grimsel granodiorite and bentonite colloids. In granodiorite batch experiments, the liquid/solid ratio of GGW to crushed granodiorite was 4 ml/g. The sizes of the crushed granodiorite ranged between 250 and 800 µm and the specific surface area was around 0.1 m2 /g. The specific surface area of bentonite colloids is presumed to be no less than 7 × 102 m2 /g [25], although their surface area was not measured directly in these experiments. In colloid batch experiments, sorption of radionuclides onto 20 mg/l of FEBEX bentonite colloids in GGW was studied. These experiments included natural colloids from the GGW and were conducted under an Ar atmosphere. Radionuclide concentrations in the batch sorption experiments were measured by ICP-MS. Table 4 presents the Kd values of Am obtained from these batch experiments. Due to the large surface area per unit mass of bentonite, radionuclide sorption onto bentonite colloids is notably larger than that measured for the granodiorite. Implications are discussed below. 4.4. Analytical results and discussion The results of the CRR experiments for 241 Am and 243 Am, as shown in Figs. 2A and 2B, were analyzed using COLFRAC subject to the geometry parameters listed in Table 3. Table 5 lists additional parameters required by COLFRAC for analysis of the CRR experiments. 4.4.1. 243 Am transport in CRR run #1 Initially, it was assumed that Am transport was only affected by sorption onto fracture surfaces. However, scoping calculation to fit COLFRAC results to the experimental data indicated that no reasonable fit could be obtained using either equilibrium or kinetic sorption reactions between Am and fracture surfaces. Thus, it was hypothesized that naturally occurring colloids in GGW affect Am transport. Natural colloid concentrations in GGW were measured to be 0.1 mg/l [11]. Degueldre et al. [26] provide a summary of the chemical composition of natural colloidal particle in GGW observed by electron microscopy. They comprise poorly crystallized silica and mica/clay, occasionally with smaller amounts of Fe-rich phase. Therefore, transport of Am might be facilitated through sorption onto the natural colloids in GGW. To investigate pseudocolloid formation by sorption of radionuclides onto natural colloids in GGW, a blank batch test was performed (radionuclides in GGW). Fig. 5 shows the concentration of Am in GGW for both ultra-centrifuged and non-centrifuged samples [11]. The difference in the Am concentration between

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Table 5 COLFRAC parameters for Am transport in the CRR experiments Parameter

Symbol

Value

Matrix retardation factor Effective diffusion coefficient Groundwater velocity in the fracture Colloid velocity in the fracture Longitudinal dispersivity for the fracture Free solution diffusion coefficient Decay constant Longitudinal dispersivity for colloids in the fracture Free solution diffusion coefficient for colloids Filter coefficient

Ra τ Dd b qf qm αl

868 2.6 × 10−7 m2 /d 39 m/d qf × 1 or qf × 1.4 0.22 m

Dd c λd αlc

1.7 × 10−6 m2 /d 0.0 y−1 0.22 m

Ddm d λ

3.3 × 10−7 m2 /d 0.0 m−1

a R = 1 + ρKd/θ , ρ (dry density of the matrix) = 2670 kg/m3 [10–13], Kd (distribution coefficient for granodiorite) = 1.3 × 10−1 m3 /kg (see Table 4), and θ (matrix porosity) = 0.4% (see Table 3). b This coefficient was based on a realistic value for granite, which was reported in the second progress report for geological disposal of HLW in Japan [27]. It was not directly measured during the CRR experiments. c This value is based on uranine data [12]. d The value is calculated from the Stokes–Einstein equation. Bentonite colloids were assumed to be spherical with diameters of 116 nm.

Fig. 5. Concentration of ultra-centrifuged and non-centrifuged americium as a function of time in GGW [11].

ultra- and non-centrifuged samples corresponds directly to the concentration of Am sorbed onto the natural colloids present in GGW. The Kd value is calculated as: Kd =

Ci − Cf V × , Cf m

(4)

where Ci is the initial concentration, Cf is the final aqueous concentration, V is the liquid volume and m is the mass of the natural colloids. Based on the data shown in Fig. 5, Kd of Am onto natural colloids in GGW is 6.4 × 105 m3 /kg. It should be noted that the Kd value for Am sorption onto natural colloids is two orders of magnitude higher than the Kd for bentonite colloids, and six orders of magnitude higher than Am sorption onto granodiorite (see Table 4). Table 2 indicates that the colloidal fraction of Am sorbed onto bentonite colloids (the formation of pseudocolloids) is higher than that onto natural colloids proving that colloidal fraction is not proportional to Kd. This implies that

Fig. 6. The sorption kinetics of americium onto natural colloids in GGW.

colloid concentration is critical to colloidal fraction. This is likely due to the increased opportunity to form pseudocolloids in the presence of 20 mg/l of bentonite colloids than in the presence of 0.1 mg/l of natural colloids. Thus, the availability of colloids should be considered when investigating radionuclide and colloid sorption rates. Pursuing such an approach for all radionuclides and solids is difficult and beyond the scope of this work. Fig. 6 shows the change in mass of Am sorbed onto natural colloids in GGW using the Kd derived from the data shown in Fig. 5. The sorption behavior of Am onto natural colloids may be experimentally derived by recasting Eq. (2) as: dS = Kse (Kd × C − S), (5) dt where S is the mass of the radionuclide sorbed onto natural colloids and C is the time-dependent concentration of the radionuclide in solution. The kinetic rate constant of Kse = 1.2 × 105 h−1 was obtained from the batch measurement of Am sorption onto natural colloids by fitting Eq. (5) to the data in Fig. 6. It should be noted that the kinetic rate constant strongly depends on the final data point collected at 340 h because little change in concentration is observed over the previous 300 h. Nevertheless, because few data points are used to obtain this kinetic rate constant, its value should be used with caution. Note that Kse obtained from the batch measurement corresponds to α = Ks × Kdc , the forward kinetic rate constant used in COLFRAC in Eq. (3). Hence, the backward kinetic reaction rate constant of β = Kse /Kd is determined to be 1.9 × 10−1 kg/(m3 h). The large value for α indicates that desorption time is slow with respect to the duration of the experiment. In this work, the transport of Am was modelled to include the effects of both instantaneous equilibrium and kinetic sorption onto colloids described by Eqs. (1) and (2), respectively. Fig. 7 illustrates the results of model analyses for Am transport without colloids (curve a), transport in the presence of natural GGW colloids with equilibrium sorption (curve b), and kinetic sorption reaction (curve c). Because none of these curves satisfactorily fit the experimental data, the effect of hydrodynamic chromatography (HDC) on colloid transport was con-

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Fig. 8. Concentration of americium as a function of time in the sorption experiment of americium onto bentonite colloids [11]. Fig. 7. COLFRAC results compared to CRR experimental data for 243 Am (CRR run #1). (a) 243 Am: no sorption onto colloids, (b) 243 Am: equilibrium sorption onto natural colloids, (c) 243 Am: kinetic sorption onto natural colloids, (d) 243 Am: kinetic sorption onto natural colloids including HDC.

sidered in the model analysis. HDC occurs when double-layer effects and van der Waals forces exclude colloids from the slowest moving portion of the velocity profile nearest the fracture surfaces resulting in colloids that travel faster than the average flow rate [28–30]. HDC can also occur by colloid size exclusion from regions of small aperture size [9,31,32]. Grindrod [33] reported that the advective velocity of colloids could be no more than 1.3 or 1.4 times faster than the groundwater flow velocity, although Sirivithayapakorn and Keller [31] noted accelerations of up to 5.5 times the pore water velocities if colloids were restricted from the slowest flowpaths. In this work, the velocity of natural colloids was set equal to 1.4 times the average groundwater flow rate to best match the experimental data. As shown in Fig. 7, the modelling results that include kinetic sorption of Am onto natural colloids in GGW and the effects of HDC (curve d) closely agree with the experimental results. 4.4.2. 241 Am transport in CRR run #2 In run #2, the radionuclide mixture included 20 mg/l of bentonite colloids and COLFRAC was again used to interpret the effect of bentonite colloids on the transport behavior of Am. Bentonite colloids are expected to further enhance the 241 Am transport rate, because their concentration is 200 times higher than the natural GGW colloids concentration. However, as noted previously, the Kd value of 2.1 × 103 m3 /kg for Am sorption onto bentonite colloids in GGW, which includes natural colloids, is two orders of magnitude lower than the value of Kd = 6.4 × 105 m3 /kg measured for natural colloids (derived from Fig. 5). The mass of the natural colloids was neglected when the Kd for bentonite colloids was calculated because it is insignificant compared to the mass of the bentonite colloids. Fig. 8 shows the change in aqueous concentration with time for Am in the bentonite batch sorption experiment [11]. Fig. 9 shows the change in mass of Am sorbed onto bentonite colloids using Kd derived from the data shown in Fig. 8. A kinetic rate of Kse = 5.0 × 102 h−1 was obtained by fitting Eq. (5) to the data

Fig. 9. Sorption kinetics of americium onto bentonite colloids.

in Fig. 9 as done for the natural colloids in run #1. Using this Kse , α = 5.0×102 h−1 and β = 2.4×10−1 kg/(m3 h) were obtained as the forward and backward kinetic rates, respectively, for Eq. (3). The flow velocity of bentonite colloids was again set 1.4 times higher than the average groundwater flow velocity. All other model parameters used to analyze run #2 were the same as those used for run #1 (see Table 5). Fig. 10 shows the experimental data (open circles) and model-estimated breakthrough curve for Am in the presence of bentonite colloids for run #2 (curve b). For reference, both the best-fit results for 243 Am from run #1 (curve a) and 131 I from run #2 (open diamonds) are also presented. Excellent agreement is shown between the experimental data and the model results under the assumption that 241 Am is sorbed kinetically onto bentonite colloids, which show effects of HDC. 5. Conclusions The results of radionuclide migration tests performed as a part of the CRR project provide an excellent opportunity to study in situ colloid-facilitated radionuclide transport. Analytical results indicate that under the conditions present in the test shear zone, both natural and bentonite colloids influence the transport behavior of radionuclides. Specifically,

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Dr. Linda McKinley, NAGRA, for helpful comments on the manuscript. Furthermore, we would like to thank the following organizations, NAGRA (Switzerland), ENRESA (Spain), ANDRA (France), FZK-INE (Germany), for significant contributions to the CRR project. This study was conducted under the collaboration between JNC (Japan Nuclear Cycle Development Institute) and SNL (Sandia National Laboratories, USA). This manuscript was substantially improved through the insightful comments of three anonymous reviewers. Sandia is a multiprogram laboratory operated by Sandia Corporation, a Lockheed Martin Company, for the United States Department of Energy’s National Nuclear Security Administration under contract DEAC04-94AL85000. References Fig. 10. COLFRAC fit to 241 Am with bentonite colloids (CRR run #2). (a) 241 Am: kinetic sorption onto natural colloids (0.1 mg/l) including HDC (for run #1), (b) 241 Am: kinetic sorption onto bentonite colloids (20 mg/l) including HDC (for run #2).

analysis of Am breakthrough provided the greatest evidence of colloid-enhanced radionuclide transport. Compared to nonsorbing tracers, the transport of Am should be significantly retarded by strong sorption onto rock surfaces; however, the CRR experiments revealed that the transport of Am could be enhanced by sorption onto mobile colloids, yielding an earlier peak arrival time than that for a non-sorbing tracer, 131 I. The results of the CRR experiments show that the transport of Am in the presence of colloids was enhanced relative to the transport behavior of a conservative tracer. Data were interpreted with COLFRAC by adjusting model parameters such that modelled breakthrough curves coincide with experimental data. Furthermore, radionuclide breakthrough results are best explained in the model using a kinetic process for sorption of radionuclides onto colloids. Results also suggest that sorption of Am onto colloids is reversible, follows the Henry isotherm, and that desorption is slow with respect to the duration of the CRR experiments. Finally, modelling results indicate that colloid velocity in this fracture system may be faster than the average groundwater velocity due HDC (exclusion of the colloids from the slowest flowing portions of the fracture system). Overall, it is recognized that one of the most important issues relevant to colloid-facilitated radionuclide transport is the reversibility of the radionuclide-colloid sorption reaction. This work indicates that Am sorption onto colloids is reversible, but the acquisition of an accurate sorption rate constant is difficult for these short-term experiments. The finite time scale of laboratory experiments cannot provide insight into long-term processes in natural systems. Future studies should focus on the underlying mechanisms related to the long-term kinetics of radionuclide species at the solid/water interface. Acknowledgments We would like to thank Dr. W. Russell Alexander, NAGRA, project manager of the CRR project, Dr. Ian G. McKinley, and

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