Technical Basis Document No. 8: Colloids Revision 2 September 2003 1. INTRODUCTION This technical basis document provides a summary of the conceptual understanding relevant to the projected behavior of radionuclides potentially attached to colloids as they are transported through the engineered and natural features of the Yucca Mountain repository system. This is one in a series of technical basis documents that are being prepared for the different components relevant to predicting the likely postclosure performance of the Yucca Mountain repository. The relationship of colloid formation, stability, and transport to the other components is illustrated in Figure 1-1. Figure 1-1. Components of the Postclosure Technical Basis for the License Application Colloids and colloidal transport processes are treated as an integrated component of the postclosure technical basis because aspects of their behavior could impact the performance of the various engineered and natural features of the repository system. Therefore, although colloids are not a feature of the repository system in the same way as the waste package, unsaturated zone and saturated zone are, treating colloids as an integrated component assures proper consistency of their treatment across these features. The information presented in this technical basis document, along with the associated references, provides a summary-level synthesis of the relevant aspects of colloid-facilitated radionuclide September 2003 1-1 No. 8: Colloids Revision 2 transport and forms an outline of the ongoing development of the postclosure safety analysis that will be included in the license application (LA). This information is also used to respond to open Key Technical Issue (KTI) agreements made between the U.S. Nuclear Regulatory Commission (NRC) and the U.S. Department of Energy (DOE). Placing the DOE responses to individual KTI agreement and additional information needed requests within the context of the overall colloid transport processes, allows for a more direct discussion of the relevance of the agreement. This technical basis document and appendices are responsive to agreements made between the DOE and the NRC during Technical Exchange and Management Meetings on Evolution of the Near-Field Environment (Reamer 2001a), Radionuclide Transport (Reamer and Williams 2000), and Total System Performance Assessment and Integration (Reamer 2001b). The appendices to this document are designed to allow for a transparent and direct response to each KTI agreement. This technical basis document presents a summary and synthesis of the detailed technical information presented in the analyses and model reports and other technical products that are used as the basis for the description of the waste form, invert, unsaturated zone, and saturated zone features and the incorporation of those features into the postclosure performance assessment. Several analyses, model reports, and other technical products support this summary: • Site-Scale Saturated Zone Transport (BSC 2003a) • Saturated Zone Colloid Transport (BSC 2003b) • Saturated Zone Flow and Transport Model Abstraction (BSC 2003c) • Saturated Zone In-Situ Testing (BSC 2003d) • Site-Scale Saturated Zone Flow Model (BSC 2003e) • EBS Radionuclide Transport Abstraction (BSC 2001a) • Radionuclide Transport Models Under Ambient Conditions (BSC 2001b) • Unsaturated Zone Flow Models and Submodels (BSC 2001c) • Waste Form and In-Drift Colloids Associated Radionuclide Concentrations: Abstraction and Summary (BSC 2003f) • Dissolved Concentration Limits of Radioactive Elements (BSC 2003g) • In-Package Chemistry Abstraction (BSC 2003h) • Particle Tracking Model and Abstraction of Transport Processes (BSC 2003i) • In Situ Field Testing of Processes (BSC 2001d) • Advection Versus Diffusion in the Invert (BSC 2003j) • Engineered Barrier System: Physical and Chemical Environment Model (BSC 2003k). September 2003 1-2 No. 8: Colloids Revision 2 1.1 SUMMARY OF CONCEPTUAL UNDERSTANDING Colloids are very fine particles ranging in size from 1 nm to 10 µm in characteristic dimension. Colloid transport may affect the rate of migration of contaminants in the subsurface environment relative to the transport of the contaminants as dissolved species. Potentially faster movement of colloids is the result of complex physical, chemical, electrical, and hydrodynamic phenomena acting on the colloids. These are phenomena that typically have not been associated with the transport of dissolved species. The potentially faster transport characteristic of colloids is generally referred to as the hydrodynamic chromatography effect (Haber and Brenner 1993). Radioactive colloids (i.e., colloids that carry radionuclides) have been classified as either Type I colloids (either true or intrinsic) or Type II colloids (pseudocolloids). This classification was adopted to distinguish colloids resulting from nucleation or precipitation of dissolved radionuclides under chemically saturated conditions, which generate radionuclides in particulate form (true colloids), from colloids carrying radionuclides due to sorption of dissolved radionuclides onto natural or other colloids suspended in groundwater (pseudocolloids). This classification is used in this technical basis document. There are several potential sources of colloids in groundwater, such as clay minerals, metal oxides, viruses, bacteria, and humic macromolecules (Chrysikopoulos and James 2003). It has been suggested (Haber and Brenner 1993) that dissolved groundwater contaminants could have a large affinity to sorption onto the surface of suspended colloids. As a result, colloids can possibly act as fast carriers for contaminants in subsurface environments and can be significant contributors to the rate of migration of contaminants, such as radionuclides potentially released from a repository system. In order for colloids to provide a significant means for the transport of radionuclides from the waste form to the biosphere and, consequently, be a major contributor to dose, colloidal suspensions must: (1) contain high enough concentrations of colloids to carry a significant amount of radionuclides, either embedded in the colloids or attached on the surface of the colloids, relative to the concentration of dissolved radionuclides; (2) be stable for distances in the order of tens of kilometers and for time periods in the tens of thousands of years; and (3) not be appreciably filtered, reversibly or irreversibly, by the host rock. If one or more of these conditions are not met, colloids will not be significant to system performance of a high-level radioactive waste repository. The main conclusion of this technical basis document is that colloid-facilitated radionuclide transport is not an important contributor to predicted system performance compared to transport of dissolved radionuclides. This conclusion is consistent with the opinions expressed by the International Review Team, which conducted a peer review of the total system performance assessment (TSPA) for the site recommendation. In its report, the International Review Team concluded that overly conservative assumptions may have been used in the modeling of colloid transport through the unsaturated zone (OECD and IAEA 2002, p. 39) and the saturated zone (OECD and IAEA 2002, p. 43). As a result of the conservatisms used in the modeling of colloid transport of radionuclides, the International Review Team concluded that the importance of colloid transport’s contribution to dose could have been “over rated.” September 2003 1-3 No. 8: Colloids Revision 2 1.2 PURPOSE The purpose of this technical basis document is to present a screening of processes important to colloid-facilitated transport of radionuclides in the Yucca Mountain engineered and natural barrier subsystems that will demonstrate that the presence of colloids will not lead to the early and/or significant discharge of radionuclides to the biosphere, and hence, will not play a major role in the system performance. Screening arguments will be presented for each of the subsystems, from the waste forms within the waste package to the boundary between the saturated zone and the biosphere. The technical basis supporting the process screening comprises the general theory of colloid stability, filtration, and radionuclide partitioning as documented in the scientific literature, and the results of Yucca Mountain Project (YMP) science activities that have studied the relevant physical/chemical processes for colloid generation in the waste package and transport of colloids from the waste package through the invert, unsaturated, and saturated zones. 1.3 SUMMARY CONCLUSIONS To determine the importance of colloid-facilitated radionuclide transport to system performance, the pertinent processes regarding colloid transport in four repository subsystems are examined: waste package, invert, unsaturated zone and saturated zone. Some radionuclides (e.g., radionuclides of plutonium and americium) may be transported primarily in colloidal form during the 10,000-year regulatory period, while others are primarily transported as dissolved species (e.g., technetium, iodine, and neptunium). As supported by the discussions in Sections 3, 4, 5, and 6 of this technical basis document, different processes impact the concentration of colloids in different repository subsystems. Within the drift environment, the key controlling factors are those that determine the stability of the colloidal suspensions. These factors are related to chemical conditions (such as high ionic strengths for the smectite colloids and a pH-ionic strength combination for the iron oxyhydroxide colloids) that cause the suspensions of both types of those colloids to be unstable during the first 2,000 to 3,000 years following waste emplacement. The instability of the colloid suspensions results in the colloids settling out. These same factors will also be the main contributors to the instability of colloid suspensions in the invert and the perturbed near field. Within the unsaturated zone and the saturated zone, filtration mechanisms will be the major contributors to reducing the concentration of colloids. Within the unsaturated zone, the interface between adjacent geologic units with sharply contrasting hydrologic properties will act as a barrier inhibiting colloids from moving through the downstream unit. Figure 1-2 shows qualitatively the impact of the different phenomena that lead to a decrease in the number of colloids in the repository subsystems, from the waste package to the biosphere. The impact of the reducing concentration of colloids on radionuclide releases to the biosphere is shown qualitatively in Figure 1-3. Among all of the radionuclides, americium and plutonium are the ones of biggest concern with respect to colloidal transport because of their high affinity to attach to particle/solid surfaces. In Figure 1-3, the dissolved concentration of these two radionuclides is qualitatively compared to their concentration associated with colloids for the different repository subsystems, from the drift to the saturated zone. Because radionuclides that September 2003 1-4 No. 8: Colloids Revision 2 adsorb to natural colloids may also adsorb to immobile material coated with secondary minerals, these radionuclides are eventually removed from the transporting fluid. The colloids of concern represent only a small fraction of waste degradation or rock alteration products. Figure 1-3 shows that the concentration of smectite colloids qualitatively drops significantly within the waste package until it reaches a constant value. This constant value qualitatively represents the natural concentration of smectite colloids available for reversibly sorbing radionuclides. The drop in the concentration of smectite colloids within the waste package is due to higher temperature and high ionic strength rendering the colloids unstable. Mills et al. (1991) concluded that metals strongly adsorbing onto natural colloids may also strongly adsorb onto the immobile matrix. It is thus anticipated that dissolved radionuclides, such as technetium and iodine, which are highly soluble, will be significantly more important to system performance than americium or plutonium, which are strongly associated with colloids. The screening arguments discussed in Sections 3, 4, 5, and 6 of this technical basis document support the conclusion that colloid-facilitated transport of radionuclides during the regulatory period is not as significant as transport of dissolved species for each of the four subsystems discussed. Consequently, colloid-facilitated radionuclide transport will not be significant to system performance. The discussion in this technical basis document comprises a screening argument for processes governing colloid-facilitated transport of radionuclides at the Yucca Mountain repository. However, it is not suggested that colloid-facilitated transport be excluded from the TSPA. Rather, it is concluded that, on a risk basis, colloid-facilitated transport is not a significant contributor to system performance; therefore, uncertain technical issues associated with this transport mode are of relatively low importance compared to uncertain technical issues related to radionuclide transport as dissolved species. It is also concluded that, for purposes of the TSPA, the current treatment of colloid-facilitated transport of radionuclides is adequate. 1.4 ORGANIZATION OF TECHNICAL BASIS DOCUMENT In addition to this introduction, this technical basis document consists of the following sections: • Section 2—Colloid Stability, Filtration, and Radionuclide Partitioning: General Theory • Section 3—Waste Package • Section 4—Invert • Section 5—Transport of Colloids in Unsaturated Zone • Section 6—Colloid Transport in the Saturated Zone • Section 7—Summary • Section 8—References • Appendix A—Exclusion of Entrained Colloids in Thermal-Chemical Alteration (Response to ENFE 1.06, ENFE 4.04 and GEN 1.01 (Comment #35)) September 2003 1-5 No. 8: Colloids Revision 2 • Appendix B—Sensitivity Analysis of Colloid Transport Parameters (Response to ENFE 4.06 AIN-1 and GEN 1.01 (Comments 35 and 37)) • Appendix C—Screening Out Coupled Thermal-Hydrologic-Chemical Effects (Response to ENFE 4.03 and GEN 1.01 (Comments 35 and 37)) • Appendix D—Contrasting Colloid Concentrations in the Engineered Barrier System and Saturated Zone (Response to TSPAI 3.30 and GEN 1.01 (Comments 43 and 46)) • Appendix E—Sensitivity Studies to Test Importance of Colloid Transport Parameters and Models (Response to RT 3.07 and GEN 1.01 (Comments 35, 43, and 46)) • Appendix F—Transport of Dissolved and Colloidal Radionuclides Through Invert (Response to TSPAI 3.17 and GEN 1.01 (Comments 36 and 38)) • Appendix G—Screening Criteria for Attachment of Radionuclides to Colloids (Response to RT 1.03 AIN-1, ENFE 3.05 AIN-1 and ENFE 4.05 AIN-1) • Appendix H—Changes in Colloid Concentrations due to Shifts in pH and Ionic Strength (Response to TSPAI 3.42) 1.5 NOTE REGARDING THE STATUS OF SUPPORTING TECHNICAL INFORMATION This document was prepared using the most current information available at the time of its development. This technical basis document and its appendices providing KTI agreement responses that were prepared using preliminary or draft information reflect the status of the Yucca Mountain Project’s scientific and design basis at the time of submittal. In some cases, this involved the use of draft Analysis and Model Reports (AMRs) and other draft references whose contents may change with time. Information that evolves through subsequent revisions of the AMRs and other references will be reflected in the LA as the approved analyses of record at the time of LA submittal. Consequently, the Project will not routinely update either this technical basis document or its KTI agreement appendices to reflect changes in the supporting references prior to submittal of the LA. September 2003 1-6 No. 8: Colloids Revision 2 September 2003 Figure 1-2. Processes Impacting Colloid Concentration from Drift to Biosphere Resulting in a Consistent Decrease in Colloid Concentration 1-7 No. 8: Colloids Figure 1-3. No. 8: Colloids Revision 2 Schematic Representation of Radionuclide Concentrations (Plutonium and Americium) as Dissolved Species and Colloids from Drift to Biosphere September 2003 1-8 Revision 2 2. COLLOID STABILITY, FILTRATION, AND RADIONUCLIDE PARTITIONING: GENERAL THEORY In this section, the general theory on the stability, transport and filtration, and sorption of radionuclides onto colloids is presented. A summary section is included presenting an overall picture of the potential impact of colloids on the performance of the repository system. This Section 2 is primarily based on a review of publications in the open literature. As appropriate, relevant information from YMP science activities is included to support the conclusions presented. It should be noted that the open literature, as well as that pertaining to repository systems has dealt mainly with colloid transport through saturated porous and fractured media. 2.1 STABILITY OF COLLOIDAL SUSPENSIONS A key condition for colloid-facilitated transport of radionuclides to be a significant contributor to repository performance is the ability of the colloidal suspension to remain stable for distances of tens of kilometers and over tens of thousands of years. The stability of colloidal suspensions against flocculation (also referred to as agglomeration or coagulation) of the colloids and deposition is governed by intercolloidal repulsive electrical forces, and these are a function of colloid number, density, size, composition, and surface chemistry (Haber and Brenner 1993), and attractive London-van der Waals forces that depend on the intercolloid separation distance and colloid size (Guzy et al. 1983; Elimelech and O’Melia 1990). In the 1940s and 1950s, the Derjaguin and Landau, and Verwey and Overbeek (Elimelech and O’Melia 1990) theory was used to explain colloid stability. The Derjaguin, Landau, Verwey and Overbeek theory defined the total interaction energy between colloids as a function of the separation distance between colloids. The total interaction energy is the sum of the energy of the electric double layer and the London-van der Waals energy. In addition, physical effects, such as flow rate and gravitational forces, and chemical conditions (the latter determining the colloid surface charge) also influence the intercolloid interactions that could lead to flocculation and gravitational settling of the particles (Guzy et al. 1983; Elimelech and O’Melia 1990). While gravitational settling is directly proportional to the third power of particle size, larger colloids tend to form more stable suspensions in natural systems due to size-exclusion effects (Haber and Brenner 1993). However, this should not be interpreted to mean that colloids, particularly in the higher end of the size spectrum (i.e., about 10 µm in diameter), will be stable over the temporal and spatial scales of relevance to repository performance. Several researchers have concluded that it is difficult to envision colloidal suspensions remaining stable, and therefore mobile, for large distances and long times (Mills et al. 1991; van der Lee et al. 1994). Elimelech and O’Melia (1990) have concluded that even for very stable colloidal suspensions, the attachment of the colloids to the wall of the flow channel may be significant. The interaction between colloids is governed by the sum of the electric repulsion force due to identical surface charges (known as the electric double-layer force) and the attractive London-van der Waals force (Elimelech and O’Melia 1990). The electric double-layer force stems from approaching colloids having equal surface charges. Each colloid is surrounded by a layer of oppositely charged ions that counterbalances the colloid surface charge. The September 2003 2-1 No. 8: Colloids Revision 2 London-van der Waals attractive force is a function of distance between approaching colloids and the colloids’ size. It has been shown that the strength of the London-van der Waals force is inversely proportional to the distance between the colloids and the colloid size (Guzy et al. 1983). That is, the closer the colloids are to each other and the smaller they are, the higher the strength of the attractive force. There are several factors that affect the strength of the repulsive electric double-layer force; the zeta potential (a measure of the surface charge), colloid size, and ionic strength and temperature. Ionic strength and temperature have a considerable impact on the thickness of the electrical double layer (Guzy et al. 1983). The double-layer thickness is inversely proportional to the square root of temperature and to the square root of the ionic strength (Guzy et al. 1983, Equations 47 and 48). Thus, higher ionic strengths and higher temperatures lead to thinner electric double layers and hence weaker repulsive forces between colloids. As temperature and ionic strengths increase, the repulsive force decreases allowing the attractive London-van der Waals forces to dominate colloidal particle interactions. These results are consistent with others reported in the literature. For example, Langmuir (1997, p. 439) states that a colloidal suspension’s stability ratio is inversely proportional to temperature. In Waste Form and In-Drift Colloids Associated Radionuclide Concentrations: Abstraction and Summary (BSC 2003f), experimental results demonstrate that colloid suspensions become more unstable with increasing ionic strength. As discussed in Section 3, during the first one-third of the 10,000-year postclosure regulatory period, the high temperature will result in water evaporation and increased ionic strengths rendering colloids important to radionuclide transport unstable. This will decrease the concentration of colloids either carrying or able to carry radionuclides for the waste package. Even after 10,000 years, the temperature within the emplacement drifts is expected to be 20° to 25°C higher than the host-rock ambient temperature (BSC 2003l, Figure 6.5.2-1). 2.2 COLLOID FILTRATION As colloids travel along the flow path from the repository to the biosphere, they will undergo interactions with the host rock. Larger colloids are likely to be physically trapped within the pores in the rock matrix. If colloids are bigger than the pore size through which they are attempting to pass, the particles will not be able to get past the pore entrance. Smaller colloids that can enter the pore are likely to come in very close proximity with each other to overcome any possible repulsive forces between the colloids and flocculate, thus, forming larger particles that will then get trapped within the pores. The smaller colloids (i.e., diameter less than 1 µm) will be subjected to Brownian deposition onto the rock matrix (Haber and Brenner 1993), thus, reducing the number of colloids entrained in the flow. Colloid filtration is also called colloid attachment or colloid deposition. There are many different studies regarding the transport of and deposition of colloidal particles through fractures under saturated flow conditions; some examples are Haber and Brenner (1993), van der Lee et al. (1994), Chrysikopoulos and James (2003), Chrysikopoulos and Abdel-Salam (1997), and James and Chrysikopoulos (1999, 2000, 2003). There have been other studies aimed at understanding the kinetics of colloid interactions leading to deposition on the surfaces of immobile phases in September 2003 2-2 No. 8: Colloids Revision 2 porous media. Some examples are Guzy et al. (1983), Elimelech and O’Melia (1990), and Adamczyk et al. (1992a, 1992b). Even though the latter works were for porous media, the kinetic models apply equally to the interactions between colloids and fracture surfaces because (1) the ratio of the characteristic dimension of the colloids to the characteristic dimension of the immobile phase was considered to be much smaller than one, and (2) the interactions between colloids and the immobile phases are only important at separation distances much smaller than the colloid characteristic dimension. Consequently, as the colloids approach the immobile phase, the immobile phase looks and behaves like a planar surface from the perspective of the colloids. As colloids move down a fracture, they will be affected by (1) hydrodynamic and physical forces in the directions parallel to the fracture walls and (2) external physical, chemical, and electrical forces in the direction normal to the fracture walls. Hydrodynamic forces account for the drag that the flow exerts on the colloids leading to entrainment of the colloids in the flow and gravitational effects. External forces include gravitational effect, London-van der Waals attractive force, and electric double-layer force. In addition, the external forces account for lubrication effects that colloids are subjected to as they approach the immobile phase. Finally, colloids are subjected to Brownian motion, which is typically modeled using an equivalent diffusivity. The Brownian diffusivity is directly proportional to temperature and inversely proportional to colloid size (Guzy et al. 1983). The effects of Brownian motion are more pronounced for colloids of less than 1 µm in characteristic dimension. The theoretical studies regarding colloid transport through porous and fractured media are generally consistent regarding the effect of colloid size on transport times. Larger colloids are transported at a faster rate than smaller ones, with the latter showing a larger propensity to deposit onto the immobile rock. Comparisons of experimental and theoretical results for colloid deposition reported by Elimelech and O’Melia (1990) show increased deposition rates measured experimentally than predicted theoretically. Those experimental results also suggest that in natural waters colloid deposition is independent of size. Elimelech and O’Melia (1990) attribute this discrepancy to failure of the Derjaguin, Landau, Verwey and Overbeek theory to account for the dynamic interactions between colloids and the immobile phase. Colloid filtration will be an important phenomenon resulting in the decrease of the colloid concentration, particularly within the unsaturated zone directly below the repository. 2.3 RADIONUCLIDE PARTITIONING BETWEEN COLLOIDS AND DISSOLVED SPECIES Radionuclides can potentially be transported from the waste form to the biosphere as either dissolved species or in colloidal form. In order for colloid transport to be a significant contributor to the discharge of radionuclides to the biosphere, colloids must carry a significant mass of radionuclides. As discussed in Section 3, the most likely forms of radioactive colloids under the expected conditions at the Yucca Mountain repository are: (1) colloids with embedded radionuclides resulting from the degradation of defense high-level radioactive waste glass, referred to as “waste-form colloids”; (2) colloids formed as a result of corrosion of steel components within the September 2003 2-3 No. 8: Colloids Revision 2 repository, referred to as “iron oxyhydroxide colloids”; and (3) natural groundwater colloids, referred to as “smectite colloids.” The formation of colloids due to hydrolysis or polymerization of dissolved actinides (referred to as “true” or “intrinsic colloids”) is not considered because recent experiments have demonstrated that these colloids dissolve under the expected repository conditions (see Section 3 of this technical basis report). First, in the tests performed at Argonne National Laboratory for degradation of defense high-level radioactive waste glass (Ebert 1995, Sections 6.2.1 and 6.2.2), no evidence was found of the formation of true colloids. Second, degradation of commercial spent nuclear fuel (CSNF) may form true colloids close to the fuel surface where the fluid may be saturated with respect to uranium, but these true colloids are likely to dissolve in unsaturated fluid. Besides the mass of radionuclides that can be embedded in waste-form colloids from the degradation of defense high-level radioactive waste glass, dissolved radionuclides can be adsorbed onto the surface of colloids to form “pseudocolloids.” Only plutonium and americium are expected to be embedded in colloids resulting from the degradation of defense high-level radioactive waste glass. Thus, the discussion herein focuses on the sorption of dissolved radionuclides onto the surface of available colloids. A key consideration in the assessment of colloid contribution to radionuclide transport is determining the partition coefficients between interstitial solvent (i.e., dissolved in the groundwater), the immobile phase of the host rock (i.e., sorption within the porous rock and the fracture surfaces), and the surfaces of the mobile colloids (Haber and Brenner 1993). Transport of groundwater colloids is likely to be preferentially through fractures in the unsaturated zone (see Section 5) and in the fractured volcanics within the saturated zone (Section 6). The stability of a colloidal suspension increases with particle size because of the reduced number of particles per unit volume that can collide (i.e., the larger the colloids the more stable the suspension will be and the higher the probability that the colloids will remain entrained in the groundwater flow) (Haber and Brenner 1993). However, larger colloids are unlikely to be effectively transported through the unsaturated zone and the volcanic fractures in the saturated zone in the vicinity of Yucca Mountain due to simple geometric considerations (BSC 2003m, Section 6.18). Larger colloids are more likely to be physically trapped in the pore constrictions and do not penetrate the porous matrix. As noted in Radionuclide Transport Models Under Ambient Conditions (BSC 2003m), colloids are discharged to the unsaturated zone directly into fractures for the model elements corresponding to the repository. During transport through fractures as both dissolved species and in colloidal form, radionuclides undergo (1) sorption of dissolved radionuclides onto the fracture surfaces, (2) sorption of dissolved radionuclides onto the porous matrix following diffusion, (3) sorption of dissolved radionuclides onto the surface of colloids, (4) attachment of radioactive colloids onto the surface of fractures (filtration), (5) transport as dissolved species and as colloids, (6) diffusion of dissolved radionuclides from the fractures into the porous matrix, and (7) decay. The conservation equation for each radionuclide needs to account for all of these phenomena to adequately estimate the amount of radioactive species discharged to the biosphere. Diffusion of the larger, more stable colloids into the rock matrix is not considered an important phenomenon because of the very low probability of the larger colloids being able to physically penetrate into the pores (BSC 2003m, Section 6.18). September 2003 2-4 No. 8: Colloids Revision 2 Sorption of dissolved radionuclides onto the fracture surfaces and onto the rock matrix after diffusion has been traditionally treated using an equilibrium isotherm model and defined by a distribution coefficient. Diffusion of dissolved species from the fractures into the rock matrix has been treated using a flux boundary condition at the fracture–matrix interface based on Fick’s law of diffusion. These modeling approaches are well established and accepted (de Marsily 1986). Sorption of radionuclides onto the surface of colloids depends on a number of phenomena; namely: (1) electrostatic forces, (2) ion exchange, (3) surface reactions, and (4) co-precipitation (BSC 2003f). These are complex phenomena, and it has been assumed that the effect of these phenomena can be captured using the linear isotherm model for reversible sorption (BSC 2003f). However, for this assumption to be applicable, some important conditions must exist. First, the dissolved radionuclide–colloid system must be in thermodynamic equilibrium. Second, the rate of radionuclide sorption must be linearly correlated to the concentration of dissolved radionuclides. Third, the rate of sorption must be independent of other solutes present in the groundwater. Fourth, the entire colloid surface area must be available for sorption. The likelihood that these conditions are met under repository conditions is not particularly high (BSC 2003f). Haber and Brenner (1993) investigated the sorption of dissolved contaminants onto the surface of colloidal particles under both equilibrium and nonequilibrium conditions. They developed a mathematical expression for the rate of solute sorption from the solvent onto the surface of colloids for nonequilibrium conditions (Haber and Brenner 1993, Equation 2.9b). That expression demonstrates that, under nonequilibrium conditions, the rate of sorption increases linearly with the solute concentration and decreases linearly with the concentration of solute already adsorbed onto the surface of colloids. The key parameter in the Haber and Brenner expression for nonequilibrium sorption is a mass transfer coefficient. The asymptotic limit of the Haber and Brenner expression for high values of the mass transfer coefficient is the linear isotherm model. The Haber and Brenner nonequilibrium sorption expression indicates that under nonequilibrium conditions the use of the linear isotherm model to describe the sorption of solutes onto the surface of colloids will overestimate the mass of solutes adsorbed. In Waste Form and In-Drift Colloids Associated Radionuclide Concentrations: Abstraction and Summary (BSC 2003f) it was concluded that equilibrium conditions are not expected at the Yucca Mountain repository. This conclusion, combined with the work of Haber and Brenner (1993), shows that the mass of radionuclides that can be transported as pseudocolloids will be overestimated by using a linear isotherm model to describe the partitioning of radionuclides between dissolved species and colloids. The mass of dissolved radionuclides that can be adsorbed onto the surface of colloids is a function of the colloids’ surface area per unit volume (BSC 2003f). Larger colloids, the ones that could transport farther in the fractures, have a smaller surface-area-to-volume ratio than the smaller colloids. The smaller colloids are more susceptible to attachment to the fracture walls. September 2003 2-5 No. 8: Colloids Revision 2 The mass of radionuclides that can adsorb onto colloids will also be a function of the colloid concentration. The higher the concentration of colloids, the higher the colloidal surface area available to compete with the immobile rock surfaces for sorption of radionuclides. Analyses of colloid concentrations and size distributions in Yucca Mountain groundwaters have found colloid concentrations to be too low for significant colloid-facilitated transport of radionuclides (DOE 2002, pp. 4-335 to 4-336). Moreover, Triay et al. (1996), in a compendium of colloid concentration as a function of water chemistry, found that under thermal and chemical conditions similar to those expected at Yucca Mountain, colloid concentrations are unlikely to be high enough for colloids to effectively compete with the immobile rock for the sorption of dissolved radionuclides, particularly when the immobile rock is coated with secondary minerals. As dissolved radionuclides travel from the repository towards the biosphere, their concentration is expected to decrease due to the various retardation processes that the radionuclides undergo, such as diffusion on the rock matrix and sorption onto the surfaces of the fractures and within the rock matrix. These processes act as effective mass sinks for dissolved radionuclides. This means that there will be lesser amounts of dissolved radionuclides available for sorption onto colloids and a lower likelihood for the formation of radioactive colloids. It is possible that far away from the repository desorption of radionuclides from the surface of colloids will take place, as discussed in the following paragraph. Based on conclusions reached by Mills et al. (1991), van der Lee et al. (1994) questioned whether colloid-facilitated transport of radionuclides would be significant at a nuclear waste disposal site. Mills et al. (1991) evaluated the transport of colloids and metals in porous media. They analyzed the transport of only dissolved species (solutes) under a variety of transport conditions and repeated the analyses under the same conditions including colloids. The basic system Mills et al. (1991) analyzed consisted of a contaminant source located in an unsaturated zone and the transport of the contaminants to a receptor downgradient in a saturated zone. The solute-colloid transport analyses considered uncertainty in several key parameters: source duration (1,000 and 100,000 years), natural colloid concentration (0, 10, and 100 mg/l), infiltration rate (0.007 and 0.35 m/year), partition coefficient for adsorption of solutes onto colloids (103, 104, and 105 l/kg) and partition coefficient for adsorption of solutes to the immobile soil matrix (0, 10, 102, 103, 104, and 105 l/kg). Mills et al. (1991, p. 206) concluded that, “As colloids move through subsurface media, and if adsorbed metal species are present, then the colloids reach a ‘clean’ portion of the aquifer, solutes are rapidly adsorbed to the aquifer-soil matrix and rapidly desorbed from the colloids to reestablish a new equilibrium between solute and colloid-adsorbed species. Thus, the colloids are rapidly stripped of adsorbed metal species and continue migrating with relatively little metal adsorbed.” According to Mills et al. (1991), this desorption phenomenon explains their results in which, for many of the cases they analyzed, the impact of colloid transport on the total mobile phase concentration (dissolved metal concentration plus metal concentration adsorbed to mobile colloids), in terms of either increased concentration or early arrival at the receptor location, was not “remarkable.” Mills et al. (1991, p. 206) further concluded that, “Typically, when travel times to a source are thousands of years in the absence of colloids, the travel times still remain on the order of thousands of years in the presence of colloids.” September 2003 2-6 No. 8: Colloids Revision 2 2.4 SUMMARY AND CONCLUSIONS While the potential exists for fast transport and early release of radioactive colloids to the biosphere at the Yucca Mountain repository, several conditions must simultaneously be met for transport of radioactive colloids to be a significant contributor to dose. Those conditions are: • Colloids must be present in high enough concentrations and provide high surface area per unit volume to carry a significant amount of radionuclides, either embedded in the colloids or attached on the surface of the colloids, relative to the concentration of dissolved radionuclides. • Colloidal suspensions must be stable for distances in the order of tens of kilometers and for time periods in the tens of thousands of years. • Radioactive colloids must not be appreciably filtered, reversibly or irreversibly, by the host rock. If any one of these conditions is not met, the likelihood that colloid transport provides a significant transport mechanism for the release of radionuclides to the biosphere is low. The higher temperatures and higher ionic strength within and in the vicinity of the repository during the first one-third of the postclosure regulatory period are expected to result in unstable colloidal suspensions. As radionuclides adsorb onto the surface of colloids, they may neutralize negative colloid surface charges, thus rendering the colloids more susceptible to filtration by attachment to the surface of fractures, particularly if the fractures are coated with secondary minerals. Therefore, radionuclide sorption may have a beneficial effect on decreasing the potential for fast transport of radioactive colloids. As radionuclides move farther away from the repository, their concentration as dissolved species decreases; therefore, there will be less mass of dissolved radionuclides available for sorption onto colloids. Diffusion into and sorption onto the rock matrix lead to the decrease in the concentration of dissolved radionuclides. It has been concluded (Mills et al. 1991) that over long distances and long travel times, a decrease in dissolved contaminants concentration far away from the source can result in desorption of contaminants from the surface of colloids. This potential desorption process has led some researchers to conclude that when contaminant travel times are thousands of years in the absence of colloids, they are also thousands of years in the presence of colloids. Most contaminant transport studies reported in the literature pertain to saturated conditions. There is little information regarding the transport of colloids under unsaturated conditions. This seeming lack of information for unsaturated systems notwithstanding, some intuitive arguments based on physics can be made regarding colloid transport in unsaturated conditions. Assuming that the flow in the unsaturated zone below the proposed repository horizon is primarily films of groundwater adjacent to the fracture walls, then only colloids of a size considerably smaller than the films could be transported with the groundwater flow. Also, as stated earlier, the smaller September 2003 2-7 No. 8: Colloids colloids are more susceptible to Brownian deposition and, hence, to attachment to the fracture walls. Finally, the colloids within the groundwater film could be physically closer to the fracture walls to enhance attachment to the latter by other mechanisms. 2-8 No. 8: Colloids Revision 2 September 2003 Revision 2 3. WASTE PACKAGE This section of the technical basis document summarizes the current understanding of the formation of radionuclide-laden colloids, prediction of their concentration within the waste package and their release into the invert. The waste package is the first of the repository system subfeatures examined to determine the importance of colloid-facilitated radionuclide transport with respect to the postclosure performance of the Yucca Mountain repository, as shown in Figure 1-1. This section provides a summary-level synthesis of the processes and phenomena governing the formation, prediction of their concentration, and transport of radionuclide-carrying colloids within the waste package. The modeling assumptions to be used in the TSPA for the LA (TSPA-LA) are discussed. In preparing this section, information has been summarized from the following pertinent analysis and model reports, as well as other technical products: • Multiscale Thermohydrologic Model (BSC 2003n) • Engineered Barrier System: Physical and Chemical Environment Model (BSC 2003k) • In-Package Chemistry Abstraction (BSC 2003h) • Dissolved Concentration Limits of Radioactive Elements (BSC 2003g) • Waste Form and In-Drift Colloids Associated Radionuclide Concentrations: Abstraction and Summary (BSC 2003f). 3.1 DESCRIPTION OF RELEVANT PROCESSES AROUND WASTE PACKAGE The relevant physical/chemical processes for colloid generation and transport inside waste packages prior to any disruption by igneous intrusion are shown in Figure 3-1. 3.1.1 In-Drift and In-Package Chemistry The evolution of general physical and chemical environments is important for understanding colloid generations and stabilities inside the drift and the waste package. Detailed analyses of indrift thermal-hydrologic conditions are documented in the Multiscale Thermohydrologic Model (BSC 2003n, Figure 6.3-12). A typical evolution of temperature, relative humidity, and liquidphase saturation degree over time is illustrated in Figure 3-2. Relative humidity decreases to low values shortly after emplacement of the waste packages. As the waste packages cool over time, relative humidity begins to rise until it reaches 100 percent after a few thousand years. During the period with relative humidity less than 98 percent, any dilute species in groundwater flowing into the drift become concentrated by evaporation, which results in a high ionic strength solution, which, in turn, destabilizes colloid suspensions. The ionic strength of some representative seepage waters as a function of relative humidity is shown in Figure 3-3 (BSC 2003k). As long as the relative humidity is less than 98 percent, the representative solutions have an ionic strength greater than the upper limit of colloid stability (0.05 M) (BSC 2003f, Sections 6.3.2.2 and 6.3.2.3). Therefore, the formation of a stable colloid suspension during the first thousand years after closure is unlikely. Colloid transport can be possible only after the effects of evaporation cease to be significant after the thermal period when temperatures are below 90°C (Figure 3-2). September 2003 3-1 No. 8: Colloids Revision 2 Furthermore, the degradation of the waste form and the waste package can lead to the formation of concentrated solutions. In process modeling, in-package chemistry has been predicted for two scenarios—water vapor condensation and seepage water dripping—with an assumption of no water evaporation. For the scenario of seepage water, the reaction-path calculations show that the initial seepage-water compositions are quickly modified by water–package chemistry produced as components within the waste package degrade (BSC 2003h). Depending on the conditions, in most cases, the resulting solution can attain ionic strength higher than 0.05 M as the water–waste package reaction progresses (Figure 3-4). Under such conditions, a stable colloid suspension, if there is any, will become unstable. This prediction is consistent with the observation in the static-saturated tests that colloid concentrations increase with time, up to the point where the colloid concentrations reach maximum values and start to decrease (Figure 3-5). In summary, the formation of a stable colloid suspension during the first few thousand years of the regulatory period is unlikely because of evaporation during the thermal period. Colloids could be generated at a late stage of the regulatory period, when evaporation becomes less effective. However, these colloids can be destabilized by high ionic strength environments that can be potentially created during waste degradation. Therefore, the general physical/chemical environments have imposed restrictive constraints on colloid stability inside the drift. It is expected that in-drift colloid transport can be possible only for limited time periods and limited combinations of physical and chemical conditions. September 2003 3-2 No. 8: Colloids Figure 3-1. Relevant Physical and Chemical Processes for Colloidal Generation and Transport from Waste Forms and Waste Packages Revision 2 September 2003 3-3 No. 8: Colloids Source: BSC 2003n, Figure 6.3-12. NOTE: These waste packages bracket the entire range of temperature at this location. Figure 3-2. No. 8: Colloids Thermal-Hydrologic Conditions for the Mean Infiltration-Flux Case for a Range of Waste Packages at the P2WR8C8 Location in the Tptpmn (tsw34) Unit Revision 2 September 2003 3-4 Source: BSC 2003k, DTN: MO0308SPAPCESA.001. Figure 3-3. Source: BSC 2003h, Figure 11. Figure 3-4. No. 8: Colloids Revision 2 Ionic Strength (in Molality) of Some Representative Seepage Waters as a Function of Relative Humidity An Example of Ionic Strength Evolution during the Degradation of Commercial Spent Nuclear Fuels under the Assumption of No Water Evaporation September 2003 3-5 Revision 2 Source: BSC 2003f, Figure 3. DTN: LL991109751021.094. Figure 3-5. No. 8: Colloids Concentrations of Plutonium and Colloids as a Function of Defense High-Level Radioactive Waste Glass Corrosion Test Duration 3.1.2 Waste Types and Potential Colloid Types As explained later in Section 3.4, three types of colloids are modeled in the current waste package model abstraction: • Natural colloids in seepage water/groundwater include mineral fragments, humic substances, and microbes. Humic substances are not sufficiently abundant in Yucca Mountain groundwater to impact transport (Minai et al. 1992). Microbes are susceptible to filtration due to their large sizes (~1 to 10 µm). (See Section 3.2.6.). The mineral colloids are represented as smectite colloids (BSC 2003f). • Corrosion product colloids are derived from the corrosion of waste package and metallic invert materials. These colloids are primarily composed of iron oxyhydroxides. The concentration range of the colloids is estimated based on experiments performed at the University of Nevada, Las Vegas on scaled-down miniature waste packages (DTN: MO0212UCC034JC.002; BSC 2003f). • Waste form colloids are formed from the corrosion of defense high-level radioactive waste glass. Glass waste forms tested include Savannah River Laboratory and West Valley Demonstration Project glasses (CRWMS M&O 2001a; BSC 2003f). (Glass is not available from the yet-to-be-built Hanford vitrification plant.) Colloids produced from both waste forms are primarily smectite clays containing measurable plutonium, as well as discrete radionuclide-bearing phases including brockite (thorium-calcium orthophosphate) and an amorphous thorium-titanium-iron silicate, similar to thoriutite. September 2003 3-6 Revision 2 Long-term corrosion testing of CSNF and DOE spent nuclear fuel (DSNF) under hydrologically unsaturated, oxidizing conditions has been conducted. Results from this testing show both very low colloid concentrations and low fractions of uranium in the colloid mass (Mertz et al. 2003). Thus, colloids from both CSNF and DSNF are excluded in the colloidal modeling. Note that CSNF accounts for the major fraction of the wastes by metric tons of heavy metal to be emplaced in the Yucca Mountain repository. The formation of true or intrinsic colloids depends on the degree of saturation of the solution (i.e., the solution must be supersaturated) with respect to the corresponding radionuclide-bearing mineral phase. There has been no evidence of their formation in the tests performed at Argonne National Laboratory for degradation of defense high-level radioactive waste glass. Degradation of CSNF may form true or intrinsic colloids (e.g., schoepite), but they are likely to dissolve in groundwater in the unsaturated zone. In the Argonne National Laboratory tests (Mertz et al. 2003), it was observed that these metastable colloids dissolved upon the introduction of J-13 groundwater. J-13 groundwater was assumed to be a lower bound with respect to the temperature and chemistry of water inside the drift. Typical water inside the drift is expected to be at a higher temperature and have higher ionic strength than J-13 water. Consequently, colloids will dissolve more readily in the water inside the drift than in J-13 groundwater. Therefore, the possibility of formation of intrinsic colloids is eliminated from further consideration in the colloid modeling (BSC 2003f, Section 6.3.1). 3.1.3 Radionuclide Sorption onto Colloids The effectiveness of colloid-facilitated transport depends on the sorption capability of colloids for radionuclides of interest. For reversible sorption processes, a linear isotherm model is used: (Eq. 3-1) K m m = d d c d is the distribution coefficient. The ranges of Kd values are where mc is mass of radionuclide adsorbed on a unit of mass of solid; md is the concentration of dissolved radionuclide; and K derived from literature data, which are obtained mostly from noncolloidal systems. The results of experiments with plutonium and americium with colloidal hematite and goethite show that the rates of desorption of these radionuclides are significantly lower than the rates of sorption. Over a time period up to 150 days, the extent of desorption is considerably less than that of sorption. Based on these data (Lu et al. 1998) and observations from field studies (Brady et al. 2002), 90 to 99 percent of sorbed plutonium and americium onto iron oxyhydroxides are modeled as irreversibly adsorbed (BSC 2003f). Colloid Filtration–Colloids generated within the defense high-level radioactive waste glass and at its outer surfaces can be filtered. Removal of suspended colloids within fractured and granular material occurs because of two types of phenomena: (1) physical effects, and (2) surface chemical effects. Physical filtration of colloids generally means the retention of colloids moving with the suspending fluid in pores, channels, and fracture apertures that are too small or dry to allow passage of the colloids. Two types of physical filtration are recognized in the unsaturated areas (Wan and Tokunaga 1997): conventional straining and film straining. Conventional straining, sieving, and pore-clogging will filter colloids if they are between 5 and 10 percent of September 2003 3-7 No. 8: Colloids Revision 2 the size of grains in the media and thus likely larger than a pore throat diameter or fracture aperture. Where water saturation is low, colloids may be filtered by film straining if their size is greater than the thickness of the adsorbed water film coating the grains of the rock. The rate of colloid transport through thin water films depends upon the colloid size relative to the film thickness. Therefore, physical straining because of low saturation and surface chemical effects such as adhesion and flocculation through chance interception and diffusion will dominate. However, these processes are more important in the extensive unsaturated and saturated zones. Within the waste package, the filtration process is implicitly included in the colloid generation source term. Colloid Sorption at the Air–Water Interface–Both hydrophilic and hydrophobic colloids may be sorbed at the gas–water interface for partially saturated conditions (Wan and Wilson 1994). Similar to physical straining at low saturation, the concentration of colloids sorbed depends on the saturation degree. However, other factors include the affinity of colloids for the gas–water interface, the electrostatic charge, and the salinity of the aqueous phase (BSC 2003f, Section 5.7). Hydrophobic colloids have higher affinities for the air–water interface. Colloids with low negative charge exhibit a stronger affinity. Also, high salinity of the aqueous phase generally promotes sorption. Buck et al. (2003) found that hydrophobic colloids of meta-studtite and meta-schoepite generated from degradation of CSNF tend to irreversibly attach to the air– water interface. These colloids contain moderate levels of neptunium and plutonium as well as high levels of strontium, cesium, and technetium. Colloid sorption to stationary air–water interfaces generally retard colloid transport, except in cases of relatively high saturation where bubbles with attached colloids could be transported. However, high-saturation conditions are expected to be unlikely in the drift and the unsaturated zone. Therefore, within the waste package, the colloid sorption at the air–water interface is not considered in the colloid generated source term. 3.2 MODELING ASSUMPTIONS 3.2.1 Colloids from the Corrosion of Commercial and DOE Spent Nuclear Fuel Long-term corrosion testing of CSNF and DSNF under unsaturated, oxidizing conditions has been performed to examine the release of dissolved radionuclides, as well as radionuclides associated with colloids (BSC 2003f). Testing was designed to simulate a variety of Yucca Mountain repository relevant water-exposure conditions for several spent nuclear fuels with a range of fuel burnup and composition. Results from the unsaturated testing of CSNF and DSNF at Argonne National Laboratory indicated formation of alteration products containing very low concentrations of uranium-based colloids and dissolution of the uranium-based colloids in less than several months (BSC 2003f, Section 6.3.1.2). Uranium-based spent nuclear fuels will be prevalent in the repository, and the colloidal properties of a mixture of two uranium minerals, meta-schoepite [(UO2)4(OH)6.5H2O] and UO2+x, have been examined (BSC 2003f, Section 6.3.1.2.3). These colloids are stable under short duration tests with respect to dissolution and interparticle interactions at near neutral and higher pH values in solutions saturated with the respective mineral phases. Furthermore, these colloids dissolve after introduction to J-13 groundwater in short duration tests. Natural analog studies also suggest that colloidal uranium will not be a significant contributor to radionuclide transport under oxic pH-neutral environments September 2003 3-8 No. 8: Colloids Revision 2 in uranium deposits (BSC 2003f, Section 6.3.1.2.4). Thus, colloids from CSNF and DSNF were eliminated from further consideration in the colloid modeling and abstraction. Commercial Spent Nuclear Fuel 3.2.1.1 Assessment of the importance of potential colloid formation from CSNF is based on four major observations: (1) very low colloid concentrations were observed in the CSNF degradation tests, at least an order of magnitude less than concentrations observed in the defense high-level radioactive waste glass degradation tests (based on dynamic light-scattering measurements) (Mertz et al. 2003); (2) the fraction of uranium in the colloid mass was uniformly low in the CSNF tests, the only deviation from this occurring immediately following one of two changes in vessel configuration in which the uranium fraction increased but rapidly decreased to the approximate level of earlier values (Mertz et al. 2003); (3) suspensions of meta-schoepite and UO2+x colloids in J-13 groundwater appear to dissolve in short-term saturated tests (their stability in unsaturated solutions has not been tested (Mertz et al. 2003)); and (4) field studies at uranium-bearing deposits indicate generally that under oxidizing conditions at near-neutral pH, colloid particles contain little uranium, and there is little sorption of uranium complexes to colloids (BSC 2003f). One of the reasons hypothesized for the low colloid release in the CSNF tests was the test configuration in which the Zircaloy-4 support for the fuel fragments had 7-micron holes. However, the results from the unirradiated UO2 tests also show few colloids after the formation of alteration products (Wronkiewicz et al. 1997). The unsaturated tests on unirradiated UO2 had a test configuration with large 2 to 3 mm holes at the holder base allowing for the spallation of UO2+x particulate during initial corrosion. However, the formation of a dense mat of alteration products during the UO2 corrosion apparently reduced particulate release by trapping particulate in the altered products (Wronkiewicz et al. 1997). A similar mechanism whereby the alteration products minimize particulate release may be applicable to the CSNF unsaturated tests. The concentration of released particulates or colloids from the CSNF tests is very low except during movement of the fuel samples from one setup to another (Mertz et al. 2003). In that case, colloid and particulate concentrations increased temporarily but returned to very low concentrations after the disruption (Mertz et al. 2003). While this indicates that disruptive events may contribute to the release of particulates and colloids from CSNF, it also indicates that the longevity of the colloids in the leachate is very short relative to the regulatory period. DOE Spent Nuclear Fuel 3.2.1.2 Of the approximately 250 different types of fuel in the DSNF inventory, metallic uranium fuel comprises approximately 85 percent (by weight of heavy metal) of that inventory and, as the only waste form in significant quantities distinct from the other spent fuel types, was selected for corrosion testing. An irradiated uranium metal fuel from the N-Reactor at Hanford was tested in an experimental setup similar to that used at Argonne National Laboratory for testing CSNF. Additional details on the testing can be found elsewhere (DTN: MO0306ANLSF001.459). Corrosion testing of metallic uranium samples resulted in rapid oxidization (within a few months) of the uranium primarily to an oxide sludge consisting of UO2 and higher oxides of uranium (DTN: MO0306ANLSF001.459). Although the uranium fuel disintegrated rapidly, September 2003 3-9 No. 8: Colloids Revision 2 corrosion testing was continued to determine the effect of groundwater leaching on the fuel sludge. Results from the corrosion tests showed that the composition of the DSNF colloids evolve over time from an initially UO2-rich population, to a mixed colloid population containing UO2 and higher oxides of uranium as well as smectite clays, to a population that appears to be dominated by uranium-containing smectite clays. After approximately one year of testing, the total quantity of uranium in the sludge represented approximately all the original uranium fuel sample. The uranium associated with the colloids corresponds to 0.002 to 0.006 weight percent of the original uranium fuel sample. The quantity of uranium in the fraction attached to the stainless steel vessel was 0.1 to 0.3 weight percent of the original fuel sample (DTN: MO0306ANLSF001.459). The attached material is measured by washing the stainless steel vessel in HNO3; the attached material includes sorbed solutes, sorbed colloids, and precipitates (DTN: MO0306ANLSF001.459). During DSNF corrosion, plutonium is predominantly associated with the colloidal, particulate, and sorbed size fractions. The 239Pu/238U ratios in the colloid fraction are significantly larger than those in the other fractions (sorbed, particulate, and dissolved) and is the only fraction that showed enrichment of plutonium in comparison to that in the fuel prior to corrosion (DTN: MO0306ANLSF001.459). Results from the testing suggest that plutonium is significantly adsorbed to the surface of colloids (such as corrosion products of the waste package or groundwater clays). However, it does not form as an embedded radionuclide in waste form colloids (DTN: MO0306ANLSF001.459). 3.2.2 Filtration of Colloids Near the waste, colloids may form at corroded waste fuel pellets and at its outer surfaces. They could be filtered within fractures in fuel pellets or trapped at grain boundaries. Colloids forming within fuel rods whose cladding has been breached could be filtered at perforations in the cladding. Colloids formed and spalled from the defense high-level radioactive waste glass could be filtered at perforations in the stainless steel high-level radioactive waste canister. No differentiation was made between colloid filtration and generation in the waste degradation experiments. Physical and chemical filtration of colloids is thus implicitly part of the generation process of colloids for radioactive waste. This filtration is a possible reason for the observed lack of waste form colloids in leachate leaving the CSNF and DSNF degradation experiments at Argonne National Laboratory noted above in Sections 3.2.1.1 and 3.2.1.2. Colloids reaching the interior of the waste package (after escaping from fuel-rod cladding and high-level radioactive waste containers) could be filtered at perforations in the skin of the waste package. However, further filtration within the waste package and the drift is excluded from modeling. There have been no comprehensive studies of colloid filtration within the defense high-level radioactive waste glass. Hence, meaningful analysis of colloid filtration separate from colloid generation within the waste package is currently not feasible. Therefore, a conservative assumption is made: all colloids formed within the waste (the calculated colloid source term) are assumed to exit the waste package and invert (see Section 4 for a discussion of the invert) and enter the unsaturated zone without filtration (BSC 2003f). September 2003 3-10 No. 8: Colloids Revision 2 3.2.3 Microbes and Colloidal Organic Components 2, The occurrence of microbes in the repository has been evaluated in In-Drift Microbial Communities (CRWMS M&O 2000a). In assessing the potential effects on microbial populations within the engineered barrier system, In-Drift Microbial Communities (CRWMS M&O 2000a) considered the drift mineralogy; drift physical parameters; metals, alloys, and cement used in engineered barrier system components; waste dissolution rates and quantities; groundwater compositions and infiltration rates; and compositions and fluxes of gases (e.g., CO water vapor). Environmental limits on microbial activity considered include redox conditions, temperature, radiation, hydrostatic pressure, water activity, pH, salinity, available nutrients, and others. The abundance of water and phosphorous were found to be the two environmental components which, could limit the development of microbial communities in the Yucca Mountain repository under the probable physicochemical conditions likely to be present in the proposed repository configuration. However, water and phosphorous in the repository would potentially be available in sufficient quantities to allow for the growth of microbial communities at certain stages of the postclosure period (CRWMS M&O 2000a). Microbes can accelerate or retard the transport of radionuclides in a number of ways. First, in some systems, microbes can passively or actively bioaccumulate radionuclides across the cell membrane. If the microbes are mobile, they can facilitate transport of the radionuclides. Alternatively, the microbes may be readily filtered by the rock or may form biofilms, a part of biomass attached to rocks, in which case they retard transport. Second, microbes may produce exudates, for example organic complexants, which may enhance the solubility of radionuclides and affect their sorption characteristics if complexes are formed. Third, in the course of extracting energy or nutrients, microbes may generate colloids by degrading materials or may actually destroy colloids by consuming them or by facilitating agglomeration. For example, microbial oxidation of metallic iron can produce iron-oxide colloids and aggregates. Conversely, microorganisms can decrease the concentration of stable colloids by aggregating colloidal material that they use as a food source. This has been shown to result in a significant decrease in colloid concentrations (Hersman 1995). Fourth, in some systems, microbes are able to reduce the oxidation states of some multivalent radioelements (e.g., uranium, neptunium, and plutonium). Typically, reduced forms of radionuclides are less soluble and more strongly sorptive. They may also impact local groundwater chemistry in and around the waste package. Microbes and inorganic colloids in close proximity to one another may result in collisions, which allow mutual adhesion or adsorption and result in particle size growth (agglomeration) (Buffle et al. 1998; Hersman 1995). This increase in particle size can have several effects on the potential for colloids to facilitate radionuclide transport within the repository: (1) for a given range of pore sizes in the transport medium, the larger composite particles may become filtered more readily than the individual microbe or colloid; (2) the larger particles will tend to diffuse more slowly and/or may precipitate through gravitational settling; and (3) the interactions of relatively large populations of colloids and microbes can result in agglomerated particles sufficiently large that the suspension becomes unstable and the microbe and colloid particles flocculate. Reduced transport from increased particle size is even more significant in the situation where water is present as thin films within the pore spaces and on the engineered barrier system components. In all of these processes, the net result is particle size increase and reduced transport of colloids and associated radionuclides. Hersman (1995) conducted experiments with September 2003 3-11 No. 8: Colloids Revision 2 Yucca Mountain-native bacteria and bentonite clay. In one test, agglomeration of clay colloids in a sterile microbial growth medium was compared to agglomeration in a medium into which a bacterium was introduced. The test results showed greater agglomeration in the growth medium inoculated with bacteria. In a second test, the bacteria were cultured in the medium and the test repeated with those bacteria, and similar results were noted. The development of biofilms on a substrate has been shown to result in an increased tendency to attract suspended inorganic colloids. Sprouse and Rittmann (1990) and Rittmann and Wirtel (1991) demonstrated that biofilm colonization increased the colloid cohesion efficiency, á, in a system where the collector was granular activated carbon in a methanogenic fluidized bed and the colloids were milk solids approximately one micron in diameter. Lo et al. (1996) concluded from an experiment with Fe(III) oxide colloids in a biofilm reactor that the deposition of Fe(III) oxide colloids increased slightly with biofilms present. Further iron deposition on surfaces increased with increasing particle size, suggesting interception of the colloids and/or sedimentation. Although uncertainty surrounds the microbial effect, the experiments with bacteria suggest that microbial action will tend to increase the sizes of inorganic colloids, and promote gravitational settling and filtration. Thus, not including the effects of microbes in the colloid source term and transport analysis is considered conservative with respect to total system performance. 3.2.4 Intrinsic Colloids Intrinsic colloids are colloidal-sized assemblages (between approximately 1 nm and 1 µm in longest dimension) formed from the hydrolysis and polymerization of actinide ions dissolved in solution (BSC 2003f). They may form in the waste package and engineered barrier system during waste-form degradation and radionuclide transport. Intrinsic colloids are also called primary colloids, Type I colloids, Eigenkolloide, real colloids, and true colloids. The formation of intrinsic colloids is solubility limited and, thus, based on the solution chemistry. This fact prevents significant introduction of intrinsic colloids to the environment (CRWMS M&O 2001a, Section 6.1). There has been no evidence of their formation in Argonne National Laboratory defense high-level radioactive waste glass degradation tests (Ebert 1995, Sections 6.2.1 and 6.2.2). Degradation of CSNF may form intrinsic colloids (e.g., schoepite-like uranium hydroxides) close to the fuel surface where the fluid may be saturated with respect to uranium, but they are likely to dissolve in unsaturated fluid. Because experimental data supports the conclusion that intrinsic colloids are negligible, intrinsic colloids are not considered in the TSPA-LA. 3.2.5 Effect of Temperature on Colloids Coupled thermal, hydrologic, and chemical effects are likely to result in unstable colloid suspension and, therefore, reduce the number concentration of colloids in suspension. With increasing temperature and increasing ionic strength, both conditions expected as a result of thermal, hydrologic, and chemical effects, colloid suspensions become less stable. Specifically, the coupled thermal-chemical perturbation will take place in the first few thousand years after repository closure (Figure 3-2). The thermal event produced by radiation heat will create a September 2003 3-12 No. 8: Colloids Revision 2 drying-out zone and induce water evaporation within the drift and the surrounding area, resulting in the increase in the ionic strength in the percolating solution. The chemical degradation of waste package will further concentrate the solution (Figure 3-4). The increase in ionic strength will reduce the stability of both smectite and iron oxyhydroxide colloids inside the drift. For example, the rate of particle coagulation is described by the stability relationship, W, defined as (BSC 2003f): (Eq. 3-2) W = exp(Vmax/kT) where, Vmax is the height of the energy barrier preventing particle coagulation; k is the Boltzman constant; and T is temperature. Thus, increasing the temperature will cause a colloid suspension to become less stable. This is because the elevated temperature will enhance Brownian motion of the particles and therefore increase the probability of interparticle collisions. In addition, the thermal event and its induced chemical perturbations are limited in space and transient in time. Because of the drying-out effect, a significant release of radionuclides from waste packages can be possible only after the thermal event, when the coupled thermal-chemical perturbations become completely attenuated and enough seepage water becomes available for leaching radionuclides. It is possible that some chemical perturbations may result in an increase in colloid concentration. However, such perturbations are expected to be transient and will attenuate rapidly as the colloid suspension moves away from the disposal room. This conclusion is consistent with the results reported by Triay et al. (1996) who found that, under the perturbed conditions expected at Yucca Mountain, colloid concentrations are unlikely to be high enough to effectively compete with the immobile rock for the sorption of radionuclides. High temperatures and high ionic strengths are not favorable to colloid transport. Therefore, the screening out of coupled thermal-hydrologic-chemical effects on the transport of radioactive colloids in the TSPA is a reasonable assumption. 3.2.6 Colloid Sorption at the Air–Water Interface The concentration of colloids sorbed at the gas–water interface is a function of the following conditions (BSC 2003f, Section 5.7): • The interface surface area available for colloid uptake, which is a function of the total gas saturation • The affinity of colloids for the gas–water interface (hydrophobic colloids have higher affinities than hydrophilic colloids) • The electrostatic charge on the colloid—colloids with lower negative charge exhibit a stronger affinity • The salinity of the aqueous phase with higher salinity promoting sorption. Empirical evidence suggests that the sorption affinity of colloids at the gas–water interface may be stronger than to the rock matrix (Wan and Wilson 1994). September 2003 3-13 No. 8: Colloids Revision 2 Partially saturated conditions may be classified by considering degrees of saturation. At low water saturations, the surface area of the gas–water interface approximates that of the rock matrix. Overall, colloid migration is retarded, although colloids may still move through the adsorbed water films. At intermediate water saturations, there is still an interconnected gas phase, although gas flux may be lower. The interface may act as a static sorbing surface, but estimating the geometry and surface area is complicated. At high water saturations, the majority of the gas is present as small gas bubbles that may migrate, transporting sorbed colloids. Colloid migration rates depend more strongly on colloid size as lower saturation states are considered (CRWMS M&O 2001b; McGraw 1996). To examine the influence of colloid size on transport, McGraw (1996) investigated transport of monodisperse colloids (five different sizes, between 20 nm and 1,900 nm) under both saturated and unsaturated conditions in a quartz sand. The results indicated that under saturated conditions the time required for breakthrough of 50 percent of the original colloid concentration was the same as that of the breakthrough of a nonreactive tracer, indicating no relationship of colloid size to migration under saturated conditions. However, the times required for breakthroughs of the colloids under highly unsaturated conditions exhibited a strong relationship between the colloid breakthrough and the colloid size with fairly complete breakthrough of the 20-nm colloid and little or no breakthrough of the 1,900-nm colloid. Another set of experiments (McGraw 1996) compared four sets of hydrophobic and hydrophilic colloids (modified latex microspheres). The results indicated that transport of hydrophobic colloids depends on colloid size, water film thickness, and colloid charge density. In contrast, hydrophilic colloids were not affected by these variables and were rapidly transported through the system even under very low moisture contents. It was concluded that for hydrophobic colloids, the cumulative mass of colloids recovered relative to column input was logarithmically dependent upon the ratio of the water film thickness to colloid diameter. In contrast, for hydrophilic colloids, the cumulative mass of colloids recovered relative to the column input was linearly dependent upon the ratio of the water film thickness to colloid diameter; similar, but more pronounced than, the effect with the nonreactive tracer. These findings suggest that unsaturated porous media may not completely impede colloid migration when a water film is present, even for relatively large colloids, however, larger colloids will tend to be retarded more than smaller ones. Although the potential effects of degree of saturation on colloid transport are varied and complex, on balance colloids would be somewhat retarded under low-saturation both inside and outside the waste package. Because of this conclusion, sorption at the air–water interface is conservatively omitted within the waste package. 3.2.7 Selection of Radioisotopes Transport of radioisotopes on colloids is potentially important for radioisotopes that (1) have long half-life and low solubility; (2) can be entrained in, or sorbed onto, waste forms, engineered barrier materials, or geologic barrier materials that generate colloidal particles; (3) represent a major portion of the inventory; and (4) have large dose conversion factors. Considering these four criteria as part of radionuclide screening, Waste Form and In-Drift Colloids-Associated Radionuclides Concentrations: Abstract and Summary (BSC 2003f) evaluated eight September 2003 3-14 No. 8: Colloids Revision 2 radionuclides for sorption onto colloids: plutonium, americium, thorium, cesium, protactinium, neptunium, uranium, and strontium. Considering these four criteria as part of radioisotope screening, the colloidal concentration abstraction concluded that five radioisotopes attached to colloids should be evaluated: plutonium, americium, thorium, cesium, and protactinium. Four of these were analyzed in TSPA for site recommendation (CRWMS M&O 2000b; Leigh and Rechard 2001). Cesium has been added for TSPA-LA based on an updated screening analysis (BSC 2003f). Plutonium, americium, thorium, cesium, protactinium are assumed to be attached to colloids reversibly using a linear isotherm model. Plutonium and americium are assumed to be predominantly attached to iron oxyhydroxide colloids and smectite waste form colloids irreversibly. The following discussion summarizes the rationale used to include or exclude these radionuclides from the model for reversible attachment to colloids (BSC 2003f, Section 6.3.3.1). Rationale for Including Five Radioisotopes 3.2.7.1 Plutonium–Plutonium meets all four criteria. A large quantity of plutonium will exist in the repository (criterion 3). Plutonium is sparingly soluble (criterion 1) but sorbs strongly to oxide mineral surfaces (generally less strongly to silicates). Plutonium is observed to sorb strongly to soil mineral, and laboratory investigations have shown that it sorbs readily to colloids as well (criterion 2). Finally, plutonium has a large dose conversion factor. Americium–Americium also meets all four criteria. Americium will be a significant contributor to radioactivity during the first 10,000 years. Like plutonium, americium is sparingly soluble but strongly sorbs to mineral surfaces, including colloids. Laboratory investigations have shown that it sorbs strongly to colloids. Protactinium–Protactinium will be a significant contributor to radioactivity during the first 10,000 years. Because of this, and the fact that relatively little is known of the colloid behavior of protactinium, it was included in this analysis. Thorium–Thorium will be a significant contributor to radioactivity during the first 10,000 years. Because of this, and the fact that there is evidence that thorium sorbs strongly to oxides, it was included in this analysis. There is relatively little known of the colloid-related behavior of thorium. Cesium–135Cs has a long half-life and can attach strongly to certain sheet silicates (including clays) by means of ion exchange. For this reason, cesium has been observed to sorb to soil minerals, and it could potentially form pseudocolloids particularly with groundwater and defense high-level radioactive waste glass-derived clay colloids. Rationale for Excluding Neptunium, Uranium, and Strontium 3.2.7.2 Neptunium–Because neptunium will be the most significant contributor to radioactivity beyond the first 10,000 years, it was considered for inclusion in this analysis. Neptunium is more soluble under anticipated repository conditions than many of the other important radionuclides, and it sorbs considerably less strongly than, for example, plutonium and americium (see Table 3-1). The typical Kd values for neptunium sorption on Yucca Mountain-vicinity colloids are less than 100 mL/g. As demonstrated in Section 3.5, as long as Kd is less than 5000 mL/g, colloid-facilitated transport is not likely to be significant relative to transport of dissolved September 2003 3-15 No. 8: Colloids species. It would appear then that the mobility of neptunium is influenced mostly by its solubility. For these reasons, and to simplify the modeling, neptunium was not included in the reversible-sorption portion of the colloid-associated radionuclide transport analysis. Uranium–Uranium will be by far the most abundant radioactive element in the repository and primarily for this reason was considered for the analysis. Uranium is more soluble under anticipated repository conditions than many of the other important radionuclides, and it sorbs considerably less strongly than, for example, plutonium and americium. Kd values for uranium sorption on Yucca Mountain-vicinity colloids typically are less than about 1,000 mL/g (Table 3-1). As demonstrated in Section 3.5, as long as Kd is less than 5,000 mL/g, colloidfacilitated transport should not be significant relative to transport of dissolved species. As with neptunium, the mobility of uranium is thus influenced mostly by its solubility. Field observations at uranium deposits and mine sites have indicated that little or no colloid uranium transport occurs. For these reasons, and to simplify modeling, uranium was not included in the reversible-sorption of the colloid-associated radionuclide transport analysis.” Strontium–Because of its very short half-life, strontium was not considered important in the groundwater pathway either as a dissolved species or attached to colloids. Modeled Kd for Plutonium, Americium, Thorium, Neptunium, and Uranium Sorption onto Kd Values, mL/g 1 to 6 × 102 1 × 101 to 1 × 102 1 × 103 to 1 × 104 2 × 103 to 9 × 104 1 × 104 to 1 × 107 September 2003 Table 3-1. Yucca Mountain-Vicinity Colloids 3.2.7.3 Radionuclide Sorbate and Oxidation States at YMP U(VI) Np(V) Pu(V) 3-16 Th(IV) Am(III) Source: BSC 2003f, Table 9. Rationale for Irreversible Adsorption Defense high-level radioactive waste glass degradation experiments show that plutonium is probably irreversibly attached to smectite colloids generated during the experiments. Further, evidence from sorption experiments with plutonium and americium (Lu et al. 2000) with colloidal hematite and goethite show that the rates of desorption (backward rate) of plutonium and americium are significantly slower than the rates of sorption (forward rate). More importantly, over a significant time period (up to 150 days in some experiments), the extent of desorption is considerably less than the extent of sorption. Plutonium and americium are considered so strongly sorbed to colloids that, in essence, they can be considered irreversibly sorbed and are modeled in this manner within the engineered barrier system. Plutonium transport velocities in soils reflect the fact that plutonium binds strongly to soils, leaving very little, if any, soluble plutonium available for groundwater transport or plant uptake. Coughtrey et al. (1985) estimate exchangeable plutonium to be less than one percent. At Rocky Flats, plutonium in soil is largely bound to soil metal hydroxides. Litaor and Ibrahim (1996) used 0.01M CaCl2 as an extractant and measured plutonium in Rocky Flats soil to be 0.04 to 0.08 percent exchangeable. Bunzl et al. (1995) measured exchangeable 239Pu and 240Pu (0.5 to No. 8: Colloids Revision 2 Revision 2 1 percent) and 241Am (1.5 to 15 percent) from fallout-contaminated soils in Germany using 1M C2H7NO2 (ammonium acetate NH4C2H3O2) as the extractant. Laboratory experiments of plutonium sorption onto iron oxides have shown that only approximately one percent of the initially sorbed plutonium can be desorbed into solution, even after months of time have elapsed (Lu et al. 2000), which is broadly consistent with field observations. For these reasons, plutonium and americium are modeled as irreversibly attaching to corrosion (iron oxyhydroxide) colloids. No other radionuclides are considered to be irreversibly attached to colloids. 3.3 SOURCE OF DATA AND TESTING 3.3.1 Tests on Defense High-Level Radioactive Waste Glass Tests on defense high-level radioactive waste in borosilicate glass have been conducted at Argonne National Laboratory using two modes of water contacting the waste (CRWMS M&O 2001a). In the static-saturated test, glass samples were immersed in fluid for more than four years. In the dripping tests, fluid was dripped at specified rates onto glass samples. Fluid used was J-13 water and deionized water. It was observed in the static-saturated tests that colloids developed and increased in concentration with time, up to the point where the colloid concentration reached a maximum value and then decreased (Figure 3-5). From Figure 3-5, it is estimated that 1 × 10-7 M plutonium is equivalent to 5 ppm of total radionuclide-embedded colloids when both concentrations reach their maximum values. 3.3.2 Tests of Waste Package Corrosion Maximum colloid concentration values are estimated based on experiments performed at University of Nevada, Las Vegas. In these experiments, scaled-down miniature waste packages were exposed to J-13 groundwater in either a bathtub mode or a flow-through mode. The experimental results obtained indicate that the iron corrosion products are mainly composed of magnetite (Fe3O4), lepidocrocite (FeOOH), and goethite (FeOOH) (DTN: MO0302UCC034JC.003). The cumulative results have yielded average concentrations of colloidal size materials in the range of 20 mg/L within the initial four weeks of the experiments (DTN: MO0212UCC034JC.002). 3.3.3 Iron Oxyhydroxide Colloids in Nature Few data for iron oxyhydroxide colloids concentrations in nature have been found in the scientific literature, and the colloid concentrations reported vary greatly. At an iron rich ore body in South America, the Morro de Ferro natural analog site, the concentration was near 1 mg/L. Values ranging from 0.6 mg/L to 260 mg/L were measured in the vicinity of a mined uranium ore formation at Cigar Lake in Northern Saskatchewan (Vilks et al. 1993). Two Swedish groundwater colloid concentrations were measured to be 0.02 mg/L and 0.043 mg/L for saline and nonsaline groundwater samples, respectively (Laaksoharju et al. 1995). The experiments performed at University of Nevada, Las Vegas on miniature waste packages produced colloids ranging in concentration from near 0 mg/L to approximately 50 mg/L. As mentioned in Section 3.4, this latter information was used to define a uniform distribution between 0.05 and 50 mg/L. September 2003 3-17 No. 8: Colloids Revision 2 3.3.4 Stability of Smectite and Iron Oxyhydroxide Colloids The zero-point of charge of smectite is about pH 2 (Figure 3-6), below possible pH values anticipated within waste packages (BSC 2003h). Therefore, smectite colloids would remain stable if the ionic strength of the solution is low (less than 0.05 M). In the defense high-level radioactive waste glass tests conducted at Argonne National Laboratory, which were run at pH 9 and 11.5, smectite colloids were detected (Buck and Bates 1999). Tombacz et al. (1990) investigated the stability of smectite (also referred to as montmorillonite) suspensions as a function of pH and ionic strength in an NaCl solution. The results are summarized in Figure 3-6. At a neutral pH, iron oxyhydroxide colloids tend to be unstable and agglomerate. The related stability diagram is displayed in Figure 3-7. Source: BSC 2003f, Figure 4. NOTE: zpc = zero-point of charge. Figure 3-6. No. 8: Colloids Experimental Determination of Montmorillonite (a Variety of Smectite) Stability as a Function of pH and Ionic Strength (M) September 2003 3-18 Revision 2 Source: BSC 2003f, Figure 7. NOTE: zpc = zero-point of charge. Figure 3-7. No. 8: Colloids Schematic Representation of Iron Oxyhydroxide Colloid Stability as a Function of pH and Ionic Strength (M) Note that the pH of groundwater at Yucca Mountain is generally close to neutral. Within this pH range, iron oxyhydroxide colloids would have a minimal surface charge, thus reducing mutual repulsive forces and resulting in flocculation, even at a low ionic strength. This fact explains why iron oxyhydroxide colloids may occur in low concentrations in groundwater and is one of the reasons why smectite colloids are used to represent seepage/groundwater colloids. This fact also implies that any iron oxyhydroxide colloids released from waste packages into groundwater may become unstable and flocculate as they move through the disposal system. 3.3.5 Colloid Concentrations in Groundwater The range of colloid concentration in seepage/groundwater was derived based on groundwater sampling from the Yucca Mountain area. Literature data (Degueldre et al. 2000) and groundwater sampling at the Idaho National Engineering and Environmental Laboratory (BSC 2003f) was used as additional technical information that corroborated the site-specific data. The results are summarized in Figures 3-8 and 3-9. Practically no colloids were detected for the ionic strength above 0.05 M. The upper limit of colloid concentration is about 200 ppm, with a median value around 0.1 ppm (Figure 3-9). September 2003 3-19 Source: BSC 2003f, Figure 11. NOTE: The ordinate values are in particles per milliliter (pt/mL). Figure 3-8. Groundwater Colloid Concentration Data Collected in the Vicinity of Yucca Mountain Compared with Data Collected from Groundwaters around the World Source: BSC 2003f, Figure 12. Figure 3-9. Cumulative Distribution Function Showing the Probability of Occurrence of Colloid Concentration Levels (ppm or mg/L) in Groundwater Samples in the Yucca Mountain Area and Idaho National Engineering and Environmental Laboratory No. 8: Colloids Revision 2 September 2003 3-20 Revision 2 3.3.6 Radionuclide Distribution Coefficients (K ds) The effectiveness of colloid-facilitated transport depends on the sorptive affinity of radionuclides to a substrate, which can be described by the distribution coefficient (Kd) (Eq. 3-1). The Kd values used in the colloid transport modeling were obtained from various data sources and measurements including those from National Cooperative for the Disposal of Radioactive Waste (Switzerland), U.S. Environmental Protection Agency, and Yucca Mountain-specific projects (BSC 2003f). Based on these sources, professional judgment was used to develop uncertainty distributions (Table 3-2). The work of Lu et al. (1998) needs to be mentioned specifically, because it provides the best direct evidence suggesting the irreversible sorption of plutonium onto colloidal particles. The work shows that Pu(IV) and Pu(V) were rapidly adsorbed by colloids of hematite, goethite, smectite, and silica in natural and synthetic groundwater. After five days, hematite colloids sorbed all Pu(IV) and Pu(V) present in the solution, goethite sorbed 97 to 100 percent of plutonium, smectite sorbed 94 to 100 percent plutonium, and silica sorbed 46 to 86 percent of plutonium. Desorption of plutonium from colloids of hematite, goethite, and smectite was much slower than the sorption process. After 150 days, less than 0.02 percent Pu(V) was desorbed from hematite colloids. Little colloidal Pu(IV) was desorbed from hematite colloids during 150 days, even using sequential extraction under vigorous shaking conditions. Although desorption of Pu(V), as well as Pu(IV), from goethite and smectite colloids is relatively faster, only less than 1 percent of plutonium was desorbed from goethite, and 1.5 percent Pu(V) and 2.5 percent to 11 percent of Pu(IV) were desorbed from smectite after 150 days. 3.4 MODEL FOR COLLOID-FACILITATED TRANSPORT IN WASTE PACKAGE This section summarizes the algorithm used to incorporate the colloids source term abstraction in the TSPA-LA model, using a simplified model intended to retain the important principles and processes of the analyses. The logic implemented in the TSPA-LA model is provided, although specific programming details are not. September 2003 3-21 No. 8: Colloids Table 3-2 Kd Values Used for Reversible Radionuclide Sorption on Colloids in Calculations for the Total System Performance Assessment for the License Application Radionuclide Pu Am, Th, Pa Cs Source: BSC 2003f, Table 10. NOTE: The Kd values for Tc and I are very low and not listed here. No. 8: Colloids Colloid Iron Oxyhydroxide Smectite Iron Oxyhydroxide Smectite Iron Oxyhydroxide Smectite Kd Value Range (mL/g) 104 to 106 103 to 106 105 to 107 104 to 107 101 to 103 102 to 104 3-22 Kd Value Intervals (mL/g) <1 × 104 1 × 104 to 5 × 104 5 × 104 to 1 × 105 1 × 105 to 5 × 105 5 × 105 to 1 × 106 >1 × 106 <1 × 103 1 × 103 to 5 × 103 5 × 103 to 1 × 104 1 × 104 to 5 × 104 5 × 104 to 1 × 105 1 × 105 to 5 × 105 5 × 105 to 1 × 106 > 1 × 106 <1x 105 1 × 105 to 5 × 105 5 × 105 to 1 × 106 1 × 106 to 5 × 106 5 × 106 to 1 × 107 >1 × 107 <1 × 104 1 × 104 to 5 × 104 5 × 104 to 1 × 105 1 × 105 to 5 × 105 5 × 105 to 1 × 106 1 × 106 to 5 × 106 5 × 106 to 1 × 107 >1 × 107 <1 × 101 1 × 101 to 5 × 101 5 × 101 to 1 × 102 1 × 102 to 5 × 102 5 × 102 to 1 × 103 >1 × 103 <1 × 102 1 × 102 to 5 × 102 5 × 102 to 1 × 103 1 × 103 to 5 × 103 5 × 103 to 1 × 104 >1 × 104 Revision 2 Kd Value Interval Probabilities 0 0.15 0.2 0.5 0.15 0 0 0.04 0.08 0.25 0.2 0.35 0.08 0 0 0.15 0.2 0.55 0.1 0 0 0.07 0.1 0.23 0.2 0.32 0.08 0 0 0.13 0.22 0.55 0.1 0 0 0.2 0.25 0.5 0.05 0 September 2003 Revision 2 3.4.1 Sources of Colloids As already noted in Section 3.1, three sources of colloids are considered (Figure 3-10): (1) natural colloids in the seeping groundwater, (2) colloids generated from degradation of the waste package and other metallic materials in the repository, and (3) colloids generated from degradation of high-level radioactive waste. Colloids formed from the defense high-level radioactive waste glass and the naturally occurring groundwater colloids are represented by smectite colloids as explained in Sections 3.3.1 and 3.3.5. The three colloid types can transport reversibly sorbed radioisotopes, but plutonium and americium radioisotopes attached on iron oxyhydroxides colloids can also transport as “irreversibly” attached radioisotopes (Figure 3-10). Furthermore, colloids generated from spallation of the high-level radioactive waste glass have plutonium and americium radioisotopes engulfed during formation and are thus irreversibly attached as well. The colloidal concentration model uses concepts developed for the Compliance Certification Application for the Waste Isolation Pilot Plant (DOE 1996; Larson 2000; Stockman et al. 2000), in particular, categories of colloids tracked, mode of radioisotope uptake, and colloid stability. Figure 3-10. Embedded, Reversibly, and Irreversibly Attached Radioisotopes on Colloids September 2003 3-23 No. 8: Colloids Revision 2 3.4.2 Algorithm Overview The algorithm in the colloidal concentration abstraction for TSPA-LA consists of three steps (BSC 2003f) that are executed at each time step in the calculations (Figure 3-11). First, the component determines the potential colloidal mass available, usually as a function of ionic strength. Second, the component determines the stability of the colloidal mass as a function of ionic strength and pH, where ionic strength and pH are determined by the in-package chemistry (BSC 2003h) of the waste form model (CRWMS M&O 2000c; Rechard and Stockman 2001). When the colloids are unstable, a low (nonzero) limit of colloid concentration is used. Third, the component determines the fraction of the available mass of the dissolved inventory of cesium, americium, plutonium, protactinium, and thorium to be adsorbed onto the colloids. The dissolved inventory of the five radioisotopes is determined by the dissolved concentration component model (BSC 2003g) of the waste form model (CRWMS M&O 2000c; Rechard and Stockman 2001). The colloidal mass generated in Step 1 and its stability as a function of ionic strength and pH in Step 2 are generally bounding, but some uncertainty is included in parameter values for sorption in Step 3. The specific methods employed for these three steps depend on the source of the colloids. The description follows. The abstraction calculates the concentration of dissolved radionuclides, the concentration of plutonium and americium embedded in defense high-level radioactive waste glass smectite colloids, the concentration of radionuclides sorbed reversibly onto the three colloid types, the concentration of radionuclides sorbed irreversibly onto iron-hydroxide colloids, and the mass concentration of each colloid type (Figure 3-11). The colloidal concentration abstraction assumes that colloids are in the groundwater that enters a breached waste package and adsorb radioisotopes. Groundwater colloids potentially present in the repository include (1) microbes, (2) organic macromolecules (humic and fulvic acids), and (3) mineral colloids (primarily clay minerals, silica, and iron oxyhydroxide minerals). Microbial colloids were disregarded because the relatively large size of microbes makes them susceptible to filtration by the geologic media as already described in Section 3.2.3. Organic macromolecules were also disregarded, based on their low capacity to form chemical complexes at Yucca Mountain. Specifically, Minai et al. (1992) studied humic substances from J-13 well water and found that the complexation capacity was only about 2.3 × 10-10 eq/L (equivalents per liter) at pH 6.9 and 2.7 × 10-11 eq/L at pH 8.2, because of the presence of calcium in J-13 well water. September 2003 3-24 No. 8: Colloids Revision 2 September 2003 Source: BSC 2003f, Figure 17a. Figure 3-11. Algorithm for Determining the Availability and Stability of Colloids within the Waste Package 3-25 No. 8: Colloids Revision 2 Mineral colloids were considered because quartz, feldspars, silica, cristobalite, fused silica, amorphous silica, aluminosilicates, layer silicates, zeolites, plagioclase, carbonate, smectite clay, hematite, and goethite colloids occur in groundwater in the vicinity of Yucca Mountain (Kingston and Whitbeck 1991). The mineral colloids were represented by smectite clays because smectite was the dominant colloid type observed and strongly adsorbs radioisotopes. As discussed in Section 3.3, the YMP has compiled the colloid concentration in groundwater from igneous rock in saturated and unsaturated regimes from the vicinity of Yucca Mountain. This relationship was used for Step 1 of the procedure to estimate the groundwater colloidal mass concentration (Figures 3-9 and 3-11). For Step 2, the colloidal concentration determined in Step 1 was maintained provided the pH and ionic strength were below the approximately linear stability function measured by Tombacz et al. (1990), based on known properties of smectite colloidal suspensions (Section 3.3.4, Figure 3-6). Otherwise, the colloidal concentration was reduced to a minimum value of 10-6 mg/L if ionic strength is less than 0.05 M. The threshold value of the ionic strength of the 0.05 M in the waste package used to calculate the concentration of groundwater colloids is identical to the upper threshold used for determining stability and concentration of waste form colloids, since both waste form and groundwater colloids are modeled as smectite. To determine the mass of radioisotopes (plutonium, americium, thorium, protactinium, cesium) reversibly sorbed by the groundwater colloids for Step 3, a linear Freundlich model (Eq. 3-1) (Langmuir 1997) was used, using the dissolved concentration of radionuclides within the waste package as calculated by the solubility component of TSPA-LA model. The sorption coefficient (Kd) values for cesium, plutonium, and americium (and, by analogy, protactinium and thorium) were assigned piecewise uniform distributions, based on sorption experiments on smectite (Table 3-2). For cesium, the values range from 102 to 104 mL/g; plutonium ranged from 103 to 106 mL/g; for americium, the values range from 104 to 107 mL/g (Table 3-2). 3.4.3 Colloids from Iron Oxyhydroxide The colloidal concentration component assumes that iron oxyhydroxide (rust) colloids (primarily goethite, ferrihydrite, and hematite) form during corrosion of waste packages. For Step 1, a uniform distribution was assigned with a range of 50 mg/L and 0.05 mg/L, based on the miniature colloid experiments conducted at University of Nevada, Las Vegas (Section 3.3.2). For Step 2, the stability of iron oxyhydroxide colloids has been measured (Liang and Morgan 1990) and was used for developing the stability region as a function of pH and ionic strength. At and near the zero-point of charge, colloids are unstable, even at low ionic strengths. For iron oxyhydroxide colloids, the zero-point of charge ranges from about pH 5.5 to about pH 8.5 depending on the mineralogy and water chemistry. Colloids are stable at higher and lower pH values, similar to other mineral colloids, provided the ionic strength is less than 0.05 M. The result is a typical “U-shaped” stability curve (Figure 3-7). The minimum concentration of rust colloids in the unstable region was 10-3 mg/L. For Step 3, two types of sorption are modeled, irreversible sorption of a large fraction of plutonium and americium onto iron oxyhydroxide corrosion colloids and fixed corrosion September 2003 3-26 No. 8: Colloids Revision 2 products; and reversible sorption of a small fraction of plutonium and americium and reversible sorption of protactinium, thorium, and cesium onto iron oxyhydroxide colloids. Plutonium and americium are considered so strongly sorbed to colloids that, in essence, can be considered irreversibly sorbed and are modeled in this manner within the engineered barrier system (BSC 2003o, Attachment II). Plutonium transport velocities in soils reflect the fact that plutonium binds strongly to soils, leaving very little, if any, soluble plutonium available for groundwater transport or plant uptake. Coughtrey et al. (1985) estimate exchangeable plutonium (“soil available,” best estimate, Table 2a, p. 119) to be less than one percent. At Rocky Flats, soil plutonium is largely bound to soil metal hydroxides. Litaor and Ibrahim (1996) used 0.01M CaCl2 as an extractant and measured plutonium in Rocky Flats soil to be 0.04 to 0.08 percent exchangeable. Bunzl et al. (1995) measured exchangeable 239+240Pu (0.5 to 1 percent) and 241Am (1.5 to 15 percent) from fallout-contaminated soils in Germany using 1M C2H7NO2 (ammonium acetate NH4C2H3O2) as the extractant. Laboratory experiments of plutonium sorption onto iron oxides have shown that only about 1 percent of the initially sorbed plutonium can be desorbed into solution, even after five months of time have elapsed as already mentioned in Section 3.3.6 (Lu et al. 2000), which is broadly consistent with field observations. In the colloid abstraction, the difference in specific surface areas between the colloids and the stationary corrosion products is taken into account as well as the total surface areas (based on total masses). For example, plutonium will tend to attach to the colloids with higher specific surface areas, while the much larger mass of stationary corrosion products relative to total colloid mass results in most of the dissolved plutonium attaching to the stationary corrosion products. Honeyman and Ranville (2002) incorporated a stationary phase in their analysis of the effects of colloids on radionuclide retardation. The stationary phase competes strongly with the colloids for radionuclide sorption. By far most of the americium mass was calculated as attached to the stationary phase; at most, 0.001 percent of the total americium was associated with colloids (Honeyman and Ranville 2002, p. 140). In order to accommodate the field and laboratory observations, distribution of plutonium between colloids and the stationary corrosion products was implemented such that a large fraction of total plutonium is sorbed to corrosion products, a small fraction to colloids, and a very small fraction remains dissolved in the fluid. The distribution of irreversibly and reversibly sorbed plutonium and americium onto fixed and colloidal substrates is performed using a kinetic model that is described in EBS Radionuclide Transport Abstraction (BSC 2003o). Figure 3-12 shows the two domains of the conceptual model of radionuclide sorption onto the iron oxyhydroxide colloidal and stationary phases. The upstream domain is assumed to be degraded fuel rods, including secondary phases, in equilibrium with the aqueous phase at the radionuclide solubility limit (CRNdiss) predicted by the solubility component (BSC 2003g). The radionuclides of concern are the plutonium and americium isotopes, but since the material balance equations are written as a mass balance, the equations are valid for any solute species. There is no sorption considered in the upstream September 2003 3-27 No. 8: Colloids Revision 2 domain. Plutonium and americium are transported by both advection and diffusion downstream into the reaction mixing cell domain, where they can be involved in six1 separate reactions: 1. Reversible plutonium and americium sorption onto iron oxyhydroxide colloidal particles 2. Reversible plutonium and americium sorption onto the stationary phase iron oxyhydroxide corrosion products 3. Irreversible plutonium and americium sorption onto iron oxyhydroxide colloidal particles 4. Irreversible plutonium and americium sorption onto the stationary phase iron oxyhydroxide corrosion products 5. Reversible plutonium and americium sorption onto waste form colloids 6. Reversible plutonium and americium sorption onto groundwater colloids. Note: SPu refers to dissolved plutonium from spent fuel, other parameters as defined in figure. §Ù ¡Â ¡Â 1 Because plutonium and americium are also embedded into waste spalling from the high-level radioactive waste, they actually are involved in a total of eight separate reactions, but the embedded plutonium and americium are = determined separately and not included in the kinetic model. Figure 3-12. First Two Computational Cells of Conceptual Model of Reversible and Irreversible Sorption (Eq. 3-3) . September 2003 No. 8: Colloids on Mobile Iron Oxyhydroxide Colloids and Immobile Rust The mass in the fluid exiting the reaction mixing cell is constrained such that the mass of plutonium sorbed both reversibly and irreversibly onto colloids is between 90 percent and 99 percent of the total mass of plutonium exiting the system in all forms.aqueous, reversibly sorbed, and irreversibly sorbed. This conforms to observations in nature, such as the transport of plutonium from the Benham test site (Kersting et al. 1999). This flux ratio is expressed as: _mass 0.90 0.99 _ flux_out _ flux_out colloid total _mass 3-28 Revision 2 Also of interest is the ratio of the mass flux leaving the mixing cell to the mass flux entering the mixing cell. This ratio of mass out to mass in is given by ¦· (Eq. 3-4) flux _ out flux _ in mass _ mass _ = and is a measure of the retardation due to the corrosion products. It is expected that most of the plutonium mass entering the mixing cell is sorbed to the stationary iron oxyhydroxide phase and only a small fraction of it flows downstream to the unsaturated zone. Reversible sorption onto the colloids is calculated from the sampled mass concentration of corrosion colloids, Kd values between the fluid and iron oxyhydroxide colloids, and the dissolved concentration of radionuclides. The Kd values are defined with piecewise-uniform distribution describing the distribution of radionuclides (BSC 2003f, Table 10): the Kd for plutonium ranges between 104 and 106 mL/g. The Kd for americium, thorium, and protactinium ranges between 105 and 107 mL/g. The Kd for cesium ranges between 101 and 103 mL/g. 3.4.4 Colloids from High-Level Radioactive Waste The laboratory evidence suggests that as high-level radioactive waste glass degrades, colloids are generated and often contain ¡°embedded¡± or irreversibly attached plutonium and americium (CRWMS M&O 2001a). The embedded plutonium and americium are likely isolated from the aqueous system and thus not in equilibrium with the surrounding aqueous environment. Therefore, some plutonium or americium in the high-level radioactive waste is assumed to be embedded in the high-level radioactive waste colloids. On the other hand, few colloids have been observed in laboratory tests of CSNF, and those colloids observed did not have radioisotopes irreversibly attached (CRWMS M&O 2001a). Furthermore, DSNF is assumed to behave similarly to CSNF (Section 3.2.). Step 1¨CFor Step 1, colloidal concentration was set to a bounding value above the maximum value observed in the high-level radioactive waste experiments (1 ¡Á 10-7 M) at low ionic strength (CRWMS M&O 2001a). At ionic strength greater than 0.05 M, colloidal concentration was set at a low value (10-11 M). At intermediate ionic strength (between 0.01 and 0.05 M), irreversible colloidal concentration was calculated by interpolating between concentrations of 10-11 and 8 ¡Á 10-8 M. Step 1a¨CThe radionuclides plutonium and americium are modeled in Step 1a as embedded within waste form colloids, which are produced from corrosion of defense high-level radioactive waste glass. The concentrations are defined by ionic strength, based on experimental observations. In TSPA-LA, Ihi-thresh,coll,wf and Ilo-thresh,coll,wf are threshold values of the ionic strength of the fluid in the waste package used to calculate the concentration of embedded plutonium used in the abstraction. Americium is calculated as the product of the plutonium concentration times the Am/Pu ratio in the inventory, which is determined at each time step of the TSPA-LA model calculations. Stability of the defense high-level radioactive waste glass-derived colloids is September 2003 3-29 No. 8: Colloids Revision 2 determined in Step 1b. Colloid mass concentrations are calculated from the amount of plutonium embedded in smectite waste form colloids in Step 1c. Step 1b–The stability of waste form colloids from defense high-level radioactive waste glass is determined on the basis of the fluid ionic strength and pH, based on known properties of smectite colloidal suspensions. The Ihi-thresh,coll,wf and Ilo-thresh,coll,wf parameters are the same threshold values of the ionic strength of the fluid in the waste package used to determine plutonium concentration (Step 1a). They are used here to determine, along with pH, whether or not the colloids are stable. These threshold values may vary due to specific chemical and environmental conditions and are therefore sampled over a range. Step 1c–Calculation of the mass concentration of waste form colloids from defense high-level radioactive waste glass is done on the basis of experimental observations. RNcoll,wf,embed,max. It is an The colloid mass concentration is related to the embedded plutonium concentration (Equation 3-3). With reference to Equation 3-3, Mcoll,wf,both,max is the maximum colloid mass concentration corresponding to the maximum plutonium concentration, C intermediate result based on the experimental data showing the relationship between embedded plutonium concentration and colloid mass concentration. Mcoll,wf,both is a parameter that represents the total mass concentration of defense high-level radioactive waste glass colloids with both embedded and reversibly sorbed radionuclides. Step 1d–The calculation of the concentration of radionuclides (plutonium, americium, thorium, protactinium, cesium) reversibly sorbed on waste form colloids from defense high-level radioactive waste glass is based on the mass concentration of waste form colloids, Kd values describing the distribution of radionuclides between the fluid and smectite colloids, and the dissolved concentration of radionuclides as calculated by the TSPA-LA model. The CRNdiss is the concentration of radionuclide “RN”, determined by the TSPA-LA model as output from the solubility concentration model, which is used as an input to the colloid abstraction. Kd,RN,wf is a parameter derived from several sources and is an equilibrium sorption coefficient used to approximate the partitioning of dissolved radionuclide “RN” between colloids and fluid. Step 2–For Step 2, the simple linear stability relationship that was used for groundwater colloids (primarily smectite) was also used to determine the stability of the irreversible high-level radioactive waste colloid concentration as a function of pH because waste form colloids are composed primarily of smectite clay minerals (CRWMS M&O 2001a). When in the unstable region, the colloid concentration was set to 10-11 M. Step 3–For Step 3, the algorithm is more complicated for high-level radioactive waste because the radioisotopes can be both irreversibly and reversibly attached to the high-level radioactive waste colloids. The algorithm makes several assumptions: (1) only americium and plutonium are irreversibly attached; (2) the irreversibly bound isotopes of plutonium and americium are partitioned according to the isotopic mass fraction calculated at the time the colloids are formed; (3) no initial mass of decay daughters are assumed irreversibly bound; and (4) in any one time September 2003 3-30 No. 8: Colloids Revision 2 step, americium and plutonium do not attach reversibly until the maximum available mass of americium and plutonium in the high-level radioactive waste is irreversibly attached. That is, the concentration of irreversible americium and plutonium is first determined; then, the reversible available mass (including any available radioisotopes from DSNF) is calculated as the total mass available times the ratio of the irreversible concentration and the maximum irreversible concentration (Figure 3-11). The reversible concentration is determined with the linear Freundlich model. Both plutonium and americium used Kd values defined with piecewiseuniform distributions (Section 3.4.3). 3.5 UNCERTAINTY AND SENSITIVITY 3.5.1 Uncertainty of Colloid Concentration The parameters developed for the colloid transport model have various degrees of uncertainty. The sources of uncertainty arise due to (1) uncertainty associated with the use of non-sitespecific data in some cases to develop parameter values; (2) uncertainty associated with the prediction of actual physical and chemical conditions in the repository; (3) uncertainty associated with site- and environment-specific testing of colloid formation, stability, and transport; and (4) uncertainty associated with the use of supporting technical and corroborative information that may not have been developed specifically for these physical and chemical conditions. These uncertainties must be appropriately captured in the TSPA calculations by choosing conservative model formulations or appropriate upper/lower limits for parameter values. In the implementation of a TSPA, it is often necessary to use experimental data gathered over a few days to a few years and field observations recorded over a few days to a few decades to arrive at conclusions on what is anticipated to occur over the 10,000-year regulatory compliance period The parameters relating to irreversible association of radionuclides with colloids, which is influenced by kinetics, must be used cautiously when speculating about future conditions. Similarly, the extrapolation of laboratory-scale results to the scale of the repository setting also needs to be cautious. Corroboration with field data is generally required. Kersting et al. (1999) measured the total plutonium concentration at the ER-20 site, which is 1.3 km from the Nevada Test Site. The maximum plutonium concentration measured was about 10-14 M. Given the strong affinity of plutonium for solid surfaces, Kersting et al. (1999) argued that this low concentration indicated that a small fraction of plutonium could be potentially transported by colloids over a long distance. Whether this mechanism is plausible or relevant to the repository conditions (note that the physical/chemical conditions at Nevada Test Site can be very different from repository conditions), the low concentrations from Kersting et al. (1999) indicate that plutonium transported by groundwater can be attenuated very quickly over a short distance (~1.3 Km). The plutonium concentration was about six orders of magnitude lower than the solubility limits of 10-8 M experimentally determined as likely to be present in Nevada Test Site groundwater. Therefore, these observations indicate that there would be a potential for longdistance colloid-associated plutonium transport in the Yucca Mountain environment (given the fact that plutonium tends to be strongly sorbed on solid surfaces), but the contribution of such transport to system performance is unlikely to be significant. Uncertainties associated with the determination of colloid concentration in seepage/groundwater derive from (1) field sampling techniques, including differences in pumping rates at each well September 2003 3-31 No. 8: Colloids Revision 2 during extraction of the water samples, (2) other unknown factors affecting the quantities of particles suspended in the water samples, including the types of additives introduced in the wells during the drilling process itself, and (3) errors inherent to the laboratory methods used to measure the quantities of colloids suspended in the water samples (e.g., filter ripening, interference and detection limitations for dynamic light scattering measurement techniques). The sampling perturbation may result in over-estimating colloid concentrations. With respect to the spatial scale, there is the issue of the appropriateness of extrapolation of specific water sample measurements to represent the colloid concentrations in waters over a wider region or area. Temporal scaling issues could include potential seasonal variations in the quantities of colloids suspended in water samples extracted, or even shorter time-scale changes in colloid concentrations in water samples during the sampling of wells. To define this uncertainty, a cumulative distribution function was developed based on groundwater samples extracted in the vicinity of Yucca Mountain (Section 3.3.5). To evaluate the reasonableness of this distribution for the site-specific data, the distribution function compared with colloid mass concentrations reported in the literature and at Idaho National Engineering and Environmental Laboratory (Figure 3-9). As shown in Figure 3-9, the probabilistic distribution of colloid concentrations developed for YMP is reasonable. 3.5.2 Uncertainty in Stability of Colloids Uncertainty in the stability of colloids (smectite and iron oxyhydroxide) as a function of pH and ionic strength is associated mostly with the extrapolation of laboratory data reported in the literature or project-supported experimental work to actual conditions (i.e., solution chemistry) that would be present in the repository environment over the regulatory compliance period. Much of this uncertainty is accommodated by establishing conservative, but reasonable, bounding values and ranges of parameter values. This uncertainty is propagated through the TSPA-LA model by stochastic sampling of these distributions during Monte Carlo simulations employed in the model calculations. 3.5.3 Uncertainty in Partition Coefficients Uncertainty is associated with the development of sorption partition coefficients (Kd values) to describe the degree of sorption of specific radionuclides to colloids. Values reported in the literature are primarily the result of experimental work designed to establish Kd values for contaminants sorbed onto rocks, soils and other minerals, but literature specific to colloid-size minerals is not readily available. For this reason, the Kd value parameters established for smectite and iron oxyhydroxide colloids rely on limited experimental work conducted at Los Alamos National Laboratory (Lu et al. 1998). Corroborative data reported in the literature were evaluated to augment the Los Alamos National Laboratory data (Table 3-2). The transport of a radionuclide in the presence of a constant colloid concentration can be described by: ¡Ó ¡Ó . ¡Ó ¡Ó ¥ñ ) ( ) ( ) 1 [( ) ] ( ) ( ( m m V ) m cV c d c d s s c d + + = . ¥õ ¥õ D D ¡Ó ¡Ó ¡Ó ¡Óx ¥õ ¡Ó ¥õ ¡Ó ¡Ó ¥õ ¡Ó cm t m x c m x m x ¡Ó t x .. . . ÿ. . .. . + ÿ. . x (Eq. 3-5) ¡Ó ¡Ó 3-32 ¥õ ¡Ót No. 8: Colloids ¡Ó ¥õ ¡Ó September 2003 Revision 2 where ¥õ is the porosity; md is the concentration of dissolved radionuclide (M); c is the colloid concentration (g/L); mc is the moles of radionuclide sorbed on colloids (mol/g); ¥ñs is the density of the immobilized adsorbing phase (g/dm3); ms is the moles of radionuclide sorbed on the immobilized adsorbing phase (mol/g); V is the flow velocity (dm/y); D is the dispersion coefficient of the dissolved species (dm2/y); t is the time (y); and x is the spatial coordinate (dm). Substituting Equation 3-1 into Equation 3-5 for mc, one can obtain: )V cK )D cK 2m d d d d d = (Eq. 3-6) 2 m x ¡Óm t ¡Ó ¥ñ ¥õ ¡Ó ¡Ó ) 1 ( + 1 ( . ) 1 ( ) 1 ( cK K cK K x d d d s d . d ¥õ cK ) ' (Eq. 3-7) ¡Ó ¡Ó d s by: ¥õ Q . ¥ñs V T ¡Ã ' (Eq. 3-8) L V T < ' ¥õ1 ( + cK ) d ¥õ1 ( + + d include technetium, iodine, and cesium. For 1 cK >> , Equation 3-8 is reduced to: d ¥õ1 ( + + V d VT . L for .. . V ¡Ã 'T L for . ¥õ L (Eq. 3-9) Q ¥õ1 ( + ) + Therefore, the actual velocity of the advancement of the adsorption front ( ' V ) can be calculated by: V = Assuming that a fluid percolates through a porous medium of a length of L (dm), the total radionuclide release from the system over a time period T (y) can be approximately calculated AF V where A is the cross-section area of the system; and F is the incoming flux (mol/dm2/y). For L , Q = 0, i.e., no radionuclide will be released from the system. For the colloid-facilitated radionuclide transport to be negligible compared to the dissolved species transport, it is required that cK << 1. Using the upper colloid concentration limit of 200 ppm in the Yucca Mountain groundwater, it can be calculated that any radionuclide with a d K less than 5,000 mL/g can be ignored for colloid-facilitated transport. These radionuclides .. . . AF V ¥õ1 ( + .. . colloids to the immobilized adsorbing phase, + .. . Therefore, as long as Kd is large enough (greater than 5000 mg/L), the total radionuclide release from the system becomes insensitive to Kd and only depends on the concentration ratio of ¥õc /(1. ¥õ) ¥ñ , which is typically very small. s ¥õ1 ( + cK ) + (1. ¥õ) ¥ñK cK ) d K 1 ( . d s ¥õ) ¥ñ VT . ¥õc ¥õc + (1. ¥õ) ¥ñs 3-33 September 2003 No. 8: Colloids 3.6 MODEL CONFIDENCE BUILDING Sensitivity studies conducted to prioritize work for the LA (BSC 2002a) showed that the colloids had a small effect on the dose. Therefore, a relatively low level of confidence is required for the colloid model. The types and characteristics (including stability and concentration) of colloids formed from the degradation of the waste forms as used in the model abstraction is based on extensive observations of colloids from testing programs and natural groundwater. The validation activities for this model analysis and abstraction took into account the criteria established in the Yucca Mountain Review Plan, Final Report (NRC 2003). Post-modeldevelopment validation is accomplished through corroboration of model predictions and data used with data published in referred journals or literature and through corroboration by comparison to data from natural analog sites. Detailed validation arguments are documented in Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary (BSC 2003f). Corroborating/supporting data and information used to develop and validate the parameters are listed in Table 3-3. It is concluded that the validation activities performed for building confidence in the model elements and data developed for YMP colloidal transport have sufficient scientific basis and that the acceptance criteria documented in the Yucca Mountain Review Plan, Final Report (NRC 2003) will be met satisfactorily. Supporting (Corroborating) Information Used to Build Confidence in the Colloid Model Table 3-3. Supporting (Corroborating) Information Source Tombacz et al. 1990 Mertz et al. 2003 Short et al. 1988 Payne et al. 1992 Zänker et al. 2000 Vilks et al. 1993 Brady et al. 2002 DTN LA0002SK831352.003 DTN LA0002SK831352.004 Coughtrey et al. 1985 Litaor and Ibrahim 1996 Bunzl et al. 1995 No. 8: Colloids Revision 2 Data/Information Stability of smectite (which is used as a surrogate mineralogy for defense high-level radioactive waste glass colloids) at full range of conditions anticipated in TSPA-LA calculations Corroborative data supporting conclusion of no significant colloids generated from CSNF degradation Corroborative information and data regarding low colloid-associated uranium concentrations in the vicinity of mines Corroborative information and data regarding low colloid-associated uranium concentrations in the vicinity of mines Corroborative information and data regarding low colloid-associated uranium concentrations in the vicinity of mines Corroborative information and data regarding low colloid-associated uranium concentrations in the vicinity of mines Corroborative information and data regarding limited extent of dissolved uranium plumes Corroborative (non-Q) data for qualified groundwater data Corroborative information regarding irreversibility of plutonium on mineral particles Corrobo