Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary Rev 02, ICN 00

MDL-EBS-PA-000004

August 2005

EXECUTIVE SUMMARY 
This report describes the analysis and abstraction of the colloids process model for the waste 
form and engineered barrier system components of the total system performance assessment 
(TSPA) calculations to be performed with the total system performance assessment for the 
license application (TSPA-LA) model. This report includes a description of the types and 
concentrations of colloids that may be: (1) generated in the waste package from degradation of 
the waste forms and the corrosion of the waste package materials, (2) produced from other steel 
components of the repository, and (3) present in natural waters in the vicinity of Yucca 
Mountain. Additionally, sorption and desorption characteristics and mechanisms of colloids 
anticipated in the repository are addressed and discussed. The abstraction of the process model 
simplifies radionuclidecolloid behavior and radionuclide sorption by assuming that all colloids 
in the waste package are either pure smectite (montmorillonite) clay or pure iron oxides 
(hematite). The true nature of colloid composition, generation, particleparticle interactions, and 
flocculation in the waste package will be complex, involving the formation of many types of 
phases. This conceptual model is intended to build a defensible and transparent abstraction; 
nevertheless, the model predicts colloid populations in the waste package will remain low once 
the ionic strength of the contacting water reaches 0.05 M. In reality, this approach is 
conservative, as it does not take into account other factors that might reduce colloid suspension 
stability. 
Data used to develop model parameters and model process elements are based on field and 
Yucca Mountain Project-developed laboratory data, and open-literature information. 
Experimental data are relied upon to establish parameters associated with colloid formation from 
waste forms and steel corrosion products, concentrations of colloids in groundwaters in the 
vicinity of Yucca Mountain, and sorptiondesorption characteristics of radionuclides associated 
with colloids. The abstraction of colloid behavior also relies on fundamental properties of pure 
mineral phases taken from the literature. In some cases, assumptions are used in the colloid 
abstraction to aid in the establishment of model parameter value ranges. The rationale for these 
assumptions and their confirmation status are described. 
Colloids will be produced from the degradation of defense high-level waste (DHLW) glass, 
commercial spent nuclear fuel (CSNF), U.S. Department of Energy spent nuclear fuel (DSNF), 
and waste package components. The abundance of these colloids within the breached waste 
package will depend on the extent and rate of degradation and the degradation products formed. 
Colloid suspension abundance and stability also depend on environmental factors including the 
ionic strength, pH, and cation and colloid concentrations of waters entering the waste package 
from the surrounding drift. The model considers the potential effects of these factors and 
processes on colloid formation and subsequent suspension stability. Transport processes are not 
discussed in this report. These processes and their effects on the migration of radionuclides 
generated from waste degradation are reported in an analysis of potential radionuclide transport 
within engineered barrier systemthe results of this analysis are reported in EBS Radionuclide 
Transport Abstraction (BSC 2004 [DIRS 169868]). 
Colloid generation from nuclear waste forms will depend on the extent of waste form corrosion, 
nature of water contact, temperature, ionic strength, Eh, and pH. Waste form corrosion tests 
have been used to provide data on the colloid crystal structure and composition, concentration, 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
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and transuranic content. Radionuclide-bearing colloids can be produced by several mechanisms; 
precipitation of released species (sometimes termed condensation), sorption of transuranics onto 
natural seepage colloids and sorption onto iron oxide corrosion products from the waste package 
(pseudocolloid formation), and by spallation of the corrosion rind (primary colloid formation). 
One general observation from the numerous tests that have been conducted is that colloid 
populations are easily destabilized. 
Studies by van Middlesworth and Wood (1998 [DIRS 170927]) represent possibly one of the 
best validations for the colloid model. Alteration of the granite resulted in the precipitation of 
monazite and rhabdophane analogous to the laboratory tests on waste glasses. The analysis of 
filtered and unfiltered samples indicated that the majority of the mobilized rare earths and 
thorium were associated with colloids (albeit in minute quantities (10-10 to 10-11 mol/L Th), while 
the mobilized uranium fraction was mostly soluble. An important aspect of these studies was the 
application of ferric-hydroxide coprecipitation for the analysis of trace levels actinides and rare 
earths in solution, which supports the contention that attenuation of actinides in the waste 
package will occur by reaction with iron corrosion products. 
Analysis related to colloid formation and behavior within the repository drift outside the waste 
package includes the formation of steel corrosion product colloids (iron oxyhydroxides) and 
potential concentrations of seepage water colloids that might be available to transport 
radionuclides within the drift. As in the waste package analyses, colloids identified in 
groundwaters underlying the repository are used as surrogates for natural colloids in the seepage 
waters. Additional analyses within the drift outside the waste package include assessment of the 
suspension stability of the three general classes of colloids considered in this report (i.e., waste 
form colloids, steel corrosion product colloids, and natural seepage water colloids). 
Model outputs include data developed for direct use in the TSPA-LA model calculations. 
Additional outputs include parameters for use in other submodel components in the TSPA-LA 
model calculations (e.g., unsaturated and saturated zone components of the natural barriers 
system underlying the repository). Specifically, the developed model output can be categorized 
as: (1) colloid formation from DHLW glass; (2) CSNF- and DSNF-derived colloids; 
(3) formation of corrosion products, including colloids, from degradation of steel within the 
waste package and drift; (4) natural occurrence of groundwater (seepage) colloids; (5) stability 
and concentration of all modeled colloid types; (6) radionuclide sorption to colloids; 
and (7) microbial interactions. These categories comprise the final product assemblage that exits 
a failed waste package in the model, interacts with the invert, and enters the unsaturated zone. 
Specific parameters developed in this model analysis for use elsewhere in the TSPA-LA model 
include waste form colloid mass concentrations (derived from DHLW glass), corrosion product 
colloid mass concentrations, the distribution of sorption partition coefficients (Kd values) for 
selected radionuclides onto smectite clays and iron oxyhydroxide colloids, and range of specific 
surface areas for iron oxyhydroxide colloids. 

1. PURPOSE 
The purpose of this report is to present and describe the abstraction of the colloids process model 
for the waste form and engineered barrier system components of the Total System Performance 
Assessment (TSPA). The model describes the types and concentrations of colloids: (1) that 
could be generated in the waste package from degradation of the waste forms and the corrosion 
of the waste package materials, (2) produced from the steel components of the repository and 
their potential role in radionuclide transport, and (3) present in natural waters in the vicinity of 
Yucca Mountain. In addition, sorption and desorption characteristics and mechanisms of 
colloids anticipated in the repository are addressed and discussed. The abstraction of the process 
model simplifies radionuclidecolloid behavior and radionuclide-sorption by assuming that all 
colloids in the waste package are either pure smectite (montmorillonite) clay 
{(Na,Ca)0.3(Al,Fe)2-3(Si,Al)4O10(OH)2  nH2O} or iron oxides (hematite {Fe2O3} and goethite 
{FeOOH}). The model uses laboratory results on stability of pure-phase hematite and 
montmorillonite colloid suspensions with ionic strength and pH to determine the range of 
stability. The true nature of colloid composition and heterogeneity, generation, and flocculation 
in the waste package will be extremely complex, involving the formation of numerous types of 
phases, often depending on the composition of the various waste forms and waste package 
materials. Traexler et al. (2004 [DIRS 172072]) have listed potential colloidal products from 
waste form degradation but it is not yet possible to generate a model that includes all types of 
particles. However, phyllosilicate clays, related to montmorillonite (derived from the weathering 
of waste glass and the local rock), iron oxides and hydroxides, such as hematite (Fe2O3), goethite 
(FeOOH) and ferrihydrite (5Fe2O3  9H2O) (derived from the corrosion of the waste package 
components), and U(VI) alteration phases such as schoepite and uranophane (from oxidized 
CSNF and DSNF) will be the dominant phases at any exposed surfaces in the repository system. 
Hence, colloid phases consisting of these species will be the most likely to form. The model 
assumes attachment of actinides and cesium to either hematite or montmorillonite clay but does 
not include U(VI) colloidal phases. The formation of U(VI) colloids and other spent nuclear 
fuel-derived colloids are considered as alternative conceptual models. Their exclusion from the 
TSPA abstraction is based on probable behavior in the waste package environment. 
This simplification of the waste package and in-drift colloid system into two main types of 
colloidal particles is anticipated to bound the potential impact of colloid-facilitated radionuclide 
transport on repository performance, as montmorillonite and hematite will be the dominant 
colloidal phases in the waste package. 
1.1 OVERVIEW 
The colloid model abstraction is used in TSPA-LA to estimate the stability and concentration of 
colloid suspensions, as well as concentrations of radionuclides associated with the colloids, 
based on in-package and in-drift fluid chemistry and in-package dissolved radionuclide 
concentrations. As such, the colloid parameters are estimated based on fluid chemistry only (i.e., 
regardless of fluid quantity and spatial distribution). Feeds for in-package ionic strength and pH 
are received from In-Package Chemistry Abstraction (BSC 2005 [DIRS 174583]) and feeds for 
in-drift ionic strength and pH are received from the engineered barrier system physical and 
chemical environment (BSC 2005 [DIRS 173727]). Feeds for dissolved radionuclide 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
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concentrations are developed in Dissolved Concentrations of Radioactive Elements (BSC 2005 
[DIRS 173873]). 
Colloid properties of montmorillonite clay and hematite have been determined by testing (in 
aqueous solution) and by calculation and modeling (Liang and Morgan 1990 [DIRS 109524]; 
Sondi et al. 1996 [DIRS 170928]; Tombcz et al. 1990 [DIRS 112690]; Tombcz and Szekeres 
2004 [DIRS 174073]). Because in-package chemistry calculations, including fluid chemistry, 
are performed for scenarios in which water layers or films are modeled and, since colloid 
behavior is determined by fluid properties regardless of fluid quantity or shape, the methodology 
employed in the colloid model abstraction is deemed appropriate for its purpose. 
Although colloid transport under saturated conditions within the in-drift environment is not 
postulated for the regulatory assessment period because of evaporation due to the high thermal 
load, limited movement of colloids within thin water films could occur. The extent and 
longevity of these thin films depends upon the specific hydrologic conditions within the waste 
package and various locations in the surrounding engineered barrier system environment at a 
particular time, as does the degree of continuity of the films. Consequently, transport of colloids 
and associated radionuclides from the waste package to the invert may be impeded to some 
degree. Transport in the engineered barrier system is described in EBS Radionuclide Transport 
Abstraction (BSC 2004 [DIRS 169868]). 
1.2 BACKGROUND 
For the purposes of this report, colloidal systems are defined as those in which the component 
phases are dispersed in an aqueous medium ranging in size from approximately 1 nm to 1 m 
(Kim 1994 [DIRS 109521]). However, the effective experimental range for particle detection can 
be narrower depending on the analysis technique and is typically between 6 nm to 450 nm in 
diameter. 
The approach taken in this analysis and modeling effort has been to bound the probable behavior 
of colloid-associated radionuclides using accessible data. Knowledge on the weathering of alkali 
borosilicate waste glasses is well established from a broad range of national and international 
sources. With respect to data on colloids, at least 26 individual intermittent drip tests were 
conducted over several years on Savannah River Laboratories (SRL)-165A, SRL-200R, and 
Approved Testing Material (ATM)-10 (West Valley composition) actinide-doped glasses 
(Bates et al. 1995 [DIRS 170880]). About one hundred immersion tests have been conducted on 
actinide-doped borosilicate glasses (SRL-131A, SRL-202A, and SRL-200R) (Bates et al. 1995 
[DIRS 170880]; DTN: LL000123351021.117 [DIRS 143308]). Immersion corrosion tests on 
SRL-131A and SRL-202A glasses (DTN: LL000123351021.117 [DIRS 143308]) have been 
used as direct input in the model. The nature of these waste glass corrosion tests prevents pH or 
ionic strength control (i.e., the dissolution of the waste form affects both the pH and ionic 
strength of the leachate). Waste glass corrosion tests were performed at 90C. Newer 
production waste glasses tend to be more corrosion-resistant and will be less likely to produce 
copious concentrations of colloids than the glasses used in the development of this model 
abstraction (i.e., SRL-131A, SRL-200R SRL-202A, and ATM-10). Colloid data on spent 
nuclear fuel has been provided through the four intermittent long-term drip tests conducted on 
spent nuclear fuel fragments at 90C (Finn et al. 1994 [DIRS 100746]; Mertz et al. 2003 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
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[DIRS 162032]). Wilson (1990 [DIRS 100793]) ran a number of immersion tests on bare 
fragments of spent nuclear fuel and spent nuclear fuel fragments with partial cladding at 25C 
and 85C and McNamara et al. (2005 [DIRS 174068]) reported analyses of suspended colloidal 
particles from spent nuclear fuel corrosion tests run at 25C, 60C, and 90C. Data from 
Mertz et al. (2003 [DIRS 162032]) and Wilson (1990 DIRS 100793]) have been used as input to 
this colloid model report to examine colloid formation from spent nuclear fuel. 
The model has been validated through the use of glass corrosion data from Mnard et al. (1998 
[DIRS 171053]), spent nuclear fuel corrosion tests conducted by Grambow et al. (1996 
[DIRS 113253]), natural analogue studies (van Middlesworth and Wood 1997 [DIRS 170927]; 
Murakami et al. 1997 [DIRS 113272]), and tests on uranium oxide (UO2) ceramic pellets 
(Wronkiewicz et al. 1996 [DIRS 102047]). 
Traexler et al. (2004 [DIRS 172072]) and Zhao and Steward (1997 [DIRS 170892]) extensively 
reviewed colloid formation from nuclear waste forms and concluded that all waste forms, 
including DHLW glass, DSNF, and CSNF, will generate colloids during weathering and colloidfacilitated 
transport may play a role in radionuclide transport. Recent studies of plutonium 
migration in natural (nonkarst) hydrologic systems suggest relatively rapid transport times for 
plutonium (Kersting et al. 1999 [DIRS 103282]. A complex assemblage of processes would play 
a role in the potential transport of radionuclides from waste package containers emplaced within 
the repository through the engineered barrier system and the underlying natural barrier system 
(unsaturated and saturated zones). The models described in this report focus first on the 
processes and physical or hydrological factors that might lead to the formation of colloid-size 
waste products from the degradation of the high-level nuclear wastes materials and secondly on 
their subsequent stability within the waste package and surrounding engineered barrier system 
environment. Other analysis and models developed for the TSPA-LA address the processes and 
mechanisms responsible for the stability and transport or retardation of radionuclides associated 
with colloids released from the engineered barrier system to the underlying unsaturated zone and 
saturated zone environments. 
The colloid source term is defined here as the total of those radionuclides associated in some 
manner with colloids that: (1) are mobilized at the surface of the waste form, (2) are transported 
within the waste package to the waste package wall, (3) leave the waste package at a breach or 
breaches in the waste package wall, and (4) enter the drift. 
The model describing colloid formation through waste form degradation and the subsequent 
stability of colloidal suspensions within the drift is valid over the range of chemical and 
hydrological conditions anticipated within the waste package and drift during the postclosure 
regulatory assessment period. Discussion on the valid range of application for the various 
process models is included in Section 7. 
1.3 DOCUMENT DEVELOPMENT 
This report was developed, in part, by building upon previously developed concepts on colloid 
formation and suspension stability within the waste form (CRWMS M&O 2001 [DIRS 153933]) 
and in-drift (CRWMS M&O 2000 [DIRS 129280]) environments. 

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This revision to the report was developed in response to condition reports generated after 
completion of Technical Work Plan for Postclosure Waste Form Modeling (BSC 2005 [DIRS 
173246]). Accordingly, the TWP (BSC 2005 [DIRS 173246]) does not specifically list this 
report as fiscal year 2005 work, which is, thus, a deviation from that plan. However, this report 
is listed as one of the waste form products in various TWP (BSC 2005 [DIRS 173246]) tables 
and, thus, includes this work. Section 7 also contains a deviation from the postdevelopment 
validation activities listed in the TWP (BSC 2005 [DIRS 173246], Table 2-3). 
1.4 MODEL REPORT OVERVIEW 
Section 2 lists the procedures, technical work plan, and requirements under which this report was 
prepared. Section 3 discusses the software used in report preparation. 
Section 4 describes the development of inputs for the colloid model abstraction and lists the 
information used in Table 4-1 (direct inputQ data) and Table 4-2 (other input, including 
corroborative non-Q data). Yucca Mountain Review Plan, Final Report (NRC 2003 
[DIRS 163274]) acceptance criteria and the codes and standards that apply to this report are 
outlined in Section 4.2. 
Section 5 presents the assumptions used in the colloid abstraction with their rationale, 
confirmation status, and the report sections in which the assumptions are used. In those cases 
where data are inadequate to support a particular concept or model technique, a bounding 
assumption is used. 
Section 6 presents conceptual models and background information on colloid formation and 
behavior (including potential stability and their potential role in radionuclide transport). 
 Section 6.1 presents a description of the model and its objectives. 
 Section 6.2 addresses pertinent features, events, and processes (FEPs) evaluated for 
their significance to this analysis. 
 Section 6.3 presents the base-case colloid model elements in several subsections. The 
background information and rationale are provided for the approaches used, as well as 
for elements considered but not incorporated into the model. 
 Section 6.4 describes alternative models that were considered during the development 
of the colloid model and the rationale for not using them. 
 Section 6.5 presents the abstraction of the colloid model elements and the 
methodology and logic for implementing the abstraction in the TSPA-LA model 
calculations. The section begins with a summary description of the abstraction and a 
schematic overview figure (Figures 6-17 and 6-18a). Section 6.5.1 presents the inpackage 
abstraction in some detail, with figures; Section 6.5.2 similarly provides the 
in-drift abstraction. Section 6.5.3 presents an overview of the disposition of colloids 
and associated radionuclides downstream from the engineered barrier system in the 
unsaturated zone and saturated zone. 

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 Sections 6.6, 6.7, and 6.8 present the model results, description and analyses of the 
barrier capability, and an evaluation of the alternative models. 
Section 7 validates the colloids model. It also describes the waste form, corrosion products, and 
groundwater colloid parameters (used as surrogates for seepage water colloids in this analysis); 
radionuclide sorption characteristics; model uncertainty; and processes not incorporated into the 
model abstraction. 
Section 8 provides a summary of analysis, describes the developed output parameters and 
uncertainties, and presents conclusions drawn from the model results. 
Section 9 lists references cited, codes, standards, regulations, procedures, and source data. 
There are three appendices. One describes data for determination of natural abundances of 
colloids, a second describes the qualification of data used as direct input in the model 
development, and the third lists qualification plans. 

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2. QUALITY ASSURANCE 
Development of this report and the supporting analyses have been determined in the technical 
work plan (BSC 2005 [DIRS 173246], Section 8) to be subject to the Yucca Mountain Projects 
quality assurance program. Quality Assurance Requirements and Description (DOE 2004 
[DIRS 171539]) applies to the development of this report. Approved quality assurance 
procedures identified in the technical work plan (BSC 2005 [DIRS 173246], Section 4) have 
been used to conduct and document the activities described in this model report. The technical 
work plan also identifies the methods used to control the electronic management of data 
(BSC 2005 [DIRS 173246], Section 8), and there were no deviations from these methods. 
The report contributes to the analysis and modeling data used to support performance assessment 
by developing and providing source terms for waste form colloids, corrosion product colloids, 
and natural groundwater colloids for the TSPA-LA model; the conclusions do not directly impact 
engineered features or engineered and natural barriers important to waste isolation, as defined in 
AP-2.22Q, Classification Analyses and Maintenance of the Q-List. 

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3. USE OF SOFTWARE 
No computer software was used for either calculations or modeling in support of this analysis. 
No codes or routines were developed. 
Reference to parameters used in calculations by the TSPA-LA model is made in this report. The 
parameters are provided for use by the TSPA-LA model, but no total system performance 
assessment model calculations were performed for this report. The logic presented in this report 
is intended for implementation in the TSPA-LA model, but no computer runs with GoldSim 
(BSC 2003 [DIRS 161572]) were executed for this report. 
Graphs and figures were prepared with Microsoft Excel 2000 and Microsoft PowerPoint 2000. 
Excel was used for tabular presentation of data only and to execute simple mathematical 
operations in the evaluation of these data; no built-in functions or mathematical operations were 
employed. PowerPoint was used for presentation (i.e., figures) of data extracted from Excel 
spreadsheets. These are commercial off-the-shelf software and are not required to be qualified, 
per LP-SI.11Q-BSC, Software Management, Section 2.1. Only standard functions of these 
software applications (i.e., no software routines or macros) were used. 

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4. INPUTS 
Technical product inputs and sources are described as direct inputs used to develop the model 
are listed in Table 4-1. Corroborating and supporting data and information are used to validate 
the model are listed in Table 4-2 following the procedures described by LP-SIII.10Q-BSC, 
Models. 
4.1 DIRECT INPUT 
Direct technical inputs, listed in Table 4-1, address the types of colloids formed from waste form 
degradation, plutonium behavior in a Yucca Mountain groundwater, steel corrosion degradation 
products, and natural seepage water colloids that might potentially come in contact with waste 
forms and the surrounding engineered barrier system. Also included in the input tables are 
sources of data used for the development of sorption partition coefficients (Kd values) that are 
used to model radionuclide sorption to colloids. For purposes of description, the parameters 
developed in this analysis have been broadly grouped as follows: 
 Waste form colloid parameters (modeled colloid: smectite; corrosion tests on highlevel 
waste glass have shown that the dominant colloidal product formed is a smectite 
(montmorillonite) clay). Plutonium and americium colloids from spent nuclear fuel 
corrosion. 
 Radionuclide sorption partition coefficients (Kd values). 
 Corrosion product colloids (modeled colloid: iron oxyhydroxide/hematite). 
 Seepage water (groundwater) colloid parameters (modeled colloid: smectite 
(montmorillonite) clay). 
These parameter groups are described in the next section, and the parameters are listed and 
described in Table 4-1. The parameters listed in Table 4-1 are based technically on findings 
from applicable analytical, laboratory, and field research, including investigations specific to the 
Yucca Mountain region and the Yucca Mountain Project (YMP). More detailed discussion on 
the technical bases for the development of these parameters may be found in Sections 5 and 6. 
Data used to develop the model were not used to validate the model. 
Uncertainty associated with the model analysis, including development of parameter values and 
their implementation in the TSPA-LA model, is discussed in Section 6.6.2. Parameter 
uncertainties are addressed by providing ranges, probability distributions, and bounding 
assumptions as appropriate for each given parameter. Many of the parameters are derived from 
data directly relevant to colloid formation, colloid suspension stability, and capacity to associate 
with radionuclides. However, many of these data were not collected from the YMP site and may 
not fully represent the physical and chemical conditions of the repository. Departure from 
anticipated repository conditions is addressed by acknowledging uncertainties in the parameters 
and providing a means to incorporate these uncertainties into the total system performance 
assessment calculations. These uncertainties are addressed by providing ranges, probability 
distributions, and bounding assumptions as appropriate for each given parameter or process. 

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Table 4-1. Summary of Direct Input Data Used in Model Parameter Development 
Data Tracking Number (DTN) or 
Report Data or Parameter Description Unitsa Locations in 
Report 
Waste Form Colloids 
LL000123351021.117 [DIRS 143308], 
LL000905312241.018 [DIRS 152621] 
Colloid and plutonium concentrations 
associated with DHLW glass colloids 
mol/L Section 6.3.1 
Wilson 1990 [DIRS 100793] Plutonium and americium concentrations 
associated with colloids from spent nuclear 
fuel, description of uranium particles 
released  Figure 6-5a 
mol/L Section 6.3.1 
Liu et al. 1992 [DIRS 172864] Fraction of particles under 1 micrometer 
following oxidation of UO2 
 Section 6 
Sorption Coefficients 
LA0003NL831352.002 [DIRS 148526] 
LA0005NL831352.001 [DIRS 149623] 
LA0003NL831352.001 [DIRS 149172] 
Lu et al. 2000 [DIRS 166315] 
Kd values for plutonium and americium on 
hematite, goethite, and montmorillonite in J- 
13 groundwater Uptake with time, ionic 
strength, and temperature 
mL/g Section 6.3.3.1, 
Table 6-5a, 6-6 
LA0407AM831341.002 [DIRS 170621] 
Torstenfelt et al. 1982 [DIRS 171770] 
Kd values for cesium on iron oxides and clay 
minerals, and silicates 
mL/g Section 6.3.3.1, 
Table 6-5b, 
Table 6-6 
LA0004NL831352.001 [DIRS 150272] Fraction of Pu(V) and Pu(IV) sorbed to 
Fe2O3 and FeO(OH) colloids in J-13 well 
water 
 Section 6.3.3.2 
Corrosion-Generated (Iron oxyhydroxide) Colloids 
MO0312UCC034JC.008 
[DIRS 166367] 
Concentration of iron oxyhydroxide colloids 
from the corrosion of steel corrosion 
products in waste packages 
mg/L Sections 
6.3.1.3, 6.3.2.2 
Lu et al. 2000 [DIRS 166315] Experimental concentration of iron 
oxyhydroxide colloids [mass of FeOx 
colloid/water volume], cFeOx_c 
mg/L Section 6.3.3.2 
Lu et al. 2000 [DIRS 166315], Roden 
and Zachara [DIRS 171518], 
Poulton et al. 2004 [DIRS 171519] 
Experimental value of specific surface area 
of iron oxyhydroxide colloids, c FeOx S _ 
 
m2/g Section 6.3.3.2 
Liang and Morgan 1990 
[DIRS 109524] 
Threshold values of ionic strength for 
stability of iron oxide/oxyhydroxide colloid 
suspensions 
M Section 6.3.2.2 
Groundwater Colloids 
LA0002SK831352.001 [DIRS 149232] 
LA0002SK831352.002 [DIRS 149194] 
LA9910SK831341.005 [DIRS 144991] 
LA0211SK831352.002 [DIRS 161581] 
LA0211SK831352.004 [DIRS 161458] 
LA0211SK831352.001 [DIRS 161580] 
LA0211SK831352.003 [DIRS 161582] 
Colloid particle size distributions and 
concentration values for groundwater 
samples in the Yucca Mountain vicinity 
nm 
(size); 
pt/mL 
(conc.) 
Section 6.3.2.4 
and Attachment 
I 
Tombcz et al. 1990 [DIRS 112690] Threshold values of ionic strength for 
stability of smectite clay colloid suspensions 
M Section 6.3.2.3 
Lide 1995 [DIRS 101876] Particle density values for silicate and iron 
oxide minerals - used to estimate general 
particle density value for groundwater 
colloids 
g/cm3 Attachment I 
NOTE: a mol/L Pu = moles per liter of plutonium dispersion, not molarity in the strict sense. 

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Table 4-2. Summary of Corroborating/Supporting Information 
Reference Data or Parameter Description Units Locations in 
Report 
Waste Form Colloids 
Mnard et al. 1998 [DIRS 171053], 
Luckscheiter and Nesovic 2002 
[DIRS 170884], Advocat et al. 2001 
[DIRS 163198]), Pirlet 2001 
[DIRS 174351] 
Clay colloids with plutonium and 
americium, description of alkali 
borosilicate glass weathering 
mol/L Section 7.1 
Grambow et al. 1996 [DIRS 113253], 
Wronkiewicz et al. 1996 
[DIRS 102047], McNamara et al. 2005 
[DIRS 174068], Buck et al. 2004 
[DIRS 171479], Buck et al. 2004 
[DIRS 172668], Schindler et al. 2004 
[DIRS 174507], Hsu et al. 1988 
[DIRS 174350], Parks 1965 
[DIRS 174361] 
Plutonium colloids from spent nuclear 
fuel, description of uranium particles 
released from CSNF. Characteristics 
of U(VI) particles and PuO2 particles. 
mol/L Section 7.1 
Iwasaki et al. 1968 [DIRS 172518], 
Spino et al. 1996 [DIRS 174080] 
Size fraction of spent nuclear fuel 
particles in the high burn-up rim and 
the polygonization of fuel 
 Section 7.1 
Mertz et al. 2003 [DIRS 162032] 
CRWMS M&O 2001 [DIRS 154071] 
Plutonium and americium 
concentrations associated with 
colloids from spent nuclear fuel, 
description of uranium particles 
released 
mol/L Section 6.3.1 
Sorption Coefficients 
Beckman et al. 1988 [DIRS 144956] Cesium sorption to silicates and 
zeolites 
mL/g Section 7.1 
EPA 1999 [DIRS 147475]; Honeyman 
and Ranville 2003 [DIRS 161657]; 
Stenhouse 1995 [DIRS 147477]; 
EPA 2004 [DIRS 172215]; 
Degueldre et al. (1994 [DIRS 174069]) 
Kd values for plutonium, americium, 
thorium, and cesium on rock, 
minerals, and colloids; modeling 
results for partitioning of 
radionuclides between colloids and 
stationary phases 
mL/g Section 7.1 
Corrosion-Generated (Iron Oxyhydroxide) Colloids 
Laaksoharju et al. 1995 [DIRS 106449] Concentration of iron oxyhydroxide 
colloids observed in natural 
groundwaters 
mg/L Section 7.1 
Davis and Kent 1990 [DIRS 143280] Specific surface area estimates for 
iron phases 
m2/g Section 7.1 
Groundwater Colloids 
Tombcz and Szekeres (2004 
[DIRS 174073]); Sondi et al. 
(1996 [DIRS 170928]) 
Threshold values of ionic strength for 
stability of smectite clay colloid 
suspensions 
M Section 7.1 
LA0002SK831352.003 [DIRS 161771] 
LA0002SK831352.004 [DIRS 161579] 
Colloid particle size distributions and 
concentration values for groundwater 
samples in the Yucca Mountain 
vicinity 
nm (size); 
pt/mL 
(conc.) 
Section 6.3.2.4 
and Attachment I 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 4-4 August 2005 
4.1.1 Developed Parameters 
Waste Form Colloid ParametersThe parameters listed under the category Waste Form 
Colloids in Table 4-1 relate primarily to the expected colloid composition resulting from the 
degradation of DHLW glass, CSNF, and DSNF, determination of colloid suspension stability 
and estimated colloid mass concentrations, and estimated concentrations of radionuclides within 
these colloids. Parameters used to describe reversible radionuclide sorption include sorption 
partition coefficients (Kd values), colloid specific surface areas, and effective Kd values 
derived from surface complexation calculations. In addition, there are functions and 
intermediate parameters listed in Table 6-12 resulting from the model abstraction as 
implemented in the TSPA calculations. 
Waste form parameter ranges and distributions were established using the available knowledge 
of processes and principles, utilizing necessary conceptual simplifications, and consideration of 
measurement error. Parameter ranges and distributions are listed in Table 6-12. 
Radionuclide Sorption ParametersSorption partition coefficients (Kd values) are used in the 
abstraction to estimate concentrations on colloids of the reversibly sorbed radionuclides 
considered in the abstraction. The ranges and distributions of these Kd values are given in 
Table 6-12. 
Corrosion-Generated (Iron Oxyhydroxide) Colloid ParametersThe colloid parameters 
listed under Corrosion-Generated (Iron oxyhydroxide) Colloids in Table 4-1 are related to 
estimates of corrosion product colloid mass concentrations in the waste package and the invert 
and determination of their stability. Parameters used to describe reversible radionuclide sorption 
include sorption partition coefficients (Kd values), colloid specific surface areas, and effective 
Kd values derived from surface complexation calculations. In addition, listed in Table 6-12 are 
functions and intermediate parameters resulting from model abstraction for TSPA calculations. 
Parameter ranges and distributions were formulated using the available knowledge of processes 
and principles associated with iron oxyhydroxide colloids, utilizing necessary conceptual 
simplifications, and consideration of measurement error. Parameter ranges and distributions are 
listed in Table 6-12. 
Seepage Water Colloid ParametersThe colloid parameters listed under Groundwater 
Colloids in Table 4-1 define seepage water colloid mass concentrations entering the waste 
package and other components of the engineered barrier system. Groundwater parameter ranges 
and distributions were formulated using data on groundwater colloid concentration in wells 
evaluated in the vicinity of Yucca Mountain. Data sources are listed in Table 4-1 (direct input). 
4.1.2 Other Parameters Critical to the Implementation of this Abstraction in the TSPALA 
Model 
Chemical and Environmental ParametersWhen this colloids process model is implemented 
in the TSPA-LA model, it is fed certain chemical and environmental parameters from other 
submodels implemented in the TSPA-LA. Ionic strength and pH are developed in the 
in-package chemistry model (BSC 2005 [DIRS 174583]) and engineered barrier system physical 
and chemical environment model (BSC 2005 [DIRS 173727]) abstractions, and dissolved 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 4-5 August 2005 
radionuclide concentrations are developed in Dissolved Concentrations of Radioactive Elements 
(BSC 2005 [DIRS 173873]). In-drift chemical environmental parameters (e.g., seepage water 
pH and ionic strength) are provided in the engineered barrier systems physical and chemical 
environment model (BSC 2005 [DIRS 173727]). This abstraction does not directly use these 
parameters, rather the parameters are used to implement the colloid subprocess models within the 
TSPA-LA model calculations. 
4.2 CRITERIA 
Projects Requirements Document (Canori and Leitner 2003 [DIRS 166275]) identifies the highlevel 
requirements for the Project. The requirements that pertain to this model report, and their 
link to 10 CFR 63 [2005 DIRS 173273], are shown in Table 4-3. Yucca Mountain Review Plan, 
Final Report (NRC 2003 [DIRS 163274]) lists acceptance criteria pertaining to these 
requirements. Criteria that are applicable to this report are described below. 
Table 4-3. Applicable Project Requirements 
Requirement Number Title 10 CFR 63 Link 
PRD-002/T-014 Performance Objectives for the Geologic 
Repository After Permanent Closure 
10 CFR 63.113 [DIRS 173273] 
PRD-002/T-015 Requirements for Performance Assessment 10 CFR 63.114 [DIRS 173273] 
PRD-002/T-016 Requirements for Multiple Barriers 10 CFR 63.115 [DIRS 173273] 
The following requirements were taken from Yucca Mountain Review Plan, Final Report (NRC 
2003 [DIRS 163274], Section 2.2.1.3.4.3) and are based on the requirements 
of 10 CFR 63.114(a)(c) and (e)(g) [2005 DIRS 173273], relating to the radionuclide release 
rates and solubility limits model abstraction: 
 Acceptance Criterion 1System Description and Model Integration are Adequate 
 Acceptance Criterion 2Data are Sufficient for Model Justification 
 Acceptance Criterion 3Data Uncertainty is Characterized and Propagated Through 
the Model Abstraction 
 Acceptance Criterion 4Model Uncertainty Is Characterized and Propagated Through 
the Model Abstraction 
 Acceptance Criterion 5Model Abstraction Output is Supported by Objective 
Comparisons. 
Section 8.3 quotes the full text of the applicable acceptance criteria with pointers to the 
information within this report that pertains to the criteria. 
4.3 CODES, STANDARDS, AND REGULATIONS 
The work conforms to the guidance provided in American Society of Testing and Materials 
(ASTM) Standard C 1174-97 [DIRS 105725] as stated in Technical Work Plan for Postclosure 
Waste Form Modeling. (BSC 2005 [DIRS 173246], Section 3). 

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MDL-EBS-PA-000004 REV 02 4-6 August 2005 
4.4 INDIRECT INPUTS TO MODEL ANALYSIS 
Corroborating or supporting data and information sources used in the model analysis are listed in 
Table 4-2. 

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MDL-EBS-PA-000004 REV 02 5-1 August 2005 
5. ASSUMPTIONS 
The assumptions used in the colloid abstraction are presented here along with their rationale, 
confirmation status, and the report sections in which the assumptions are implemented. A 
bounding assumption was used when data were deemed to be inadequate to fully support a 
particular concept or model technique (Assumptions 5.1, 5.2, and 5.3). Additionally, the role of 
airwater interfaces was excluded, as this would have added complexity to the model 
(Assumption 5.4). 
5.1 DISTRIBUTION COEFFICIENTS OF THORIUM AND PROTACTINIUM ARE 
THE SAME AS THAT OF AMERICIUM 
Assumption: The Kd-value ranges and distribution for thorium and protactinium are those of 
americium primarily because of limited repository relevant data on thorium and protactinium. 
Rationale: Because the reported Kd-value ranges and distributions for americium are high 
(Table 5-1), this is a conservative assumption with respect to radionuclide sorption to colloids. 
The sorption range for protactinium for silicate materials (i.e., vitric rock, zeolites) and hematite 
is less than that of americium. Hence use of americium Kd values for protactinium is 
conservative and bounding. The Kd values for thorium and americium are similar; therefore, use 
of americium Kds will be suitable for thorium. 
Table 5-1. Sorption Distribution Coefficient Data from the Unsaturated Zone, Yucca Mountain, Nevada 
Element Rock Type 
Minimum Sorption 
Distribution 
Coefficient (mL/g) 
Maximum Sorption 
Distribution 
Coefficient (mL/g) 
Devitrified 5 70 
Vitric 30 200 
Zeolitic 30 200 
Pu 
Iron oxide 1,000 5,000 
Devitrified 0 100 
Vitric 0 100 
Zeolitic 0 100 
Pa 
Iron oxide 500 1,000 
Devitrified 100 2,000 
Vitric 100 1,000 
Zeolitic 100 1,000 
Am 
Iron oxide 1,000 5,000 
Devitrified 100 2,000 
Vitric 100 1,000 
Zeolitic 100 1,000 
Th 
Iron oxide 1,000 5,000 
Source: DTN: LA0003AM831341.001 [DIRS 148751]. 
Confirmation Status: This is a bounding assumption, as Pa(V) chemistry is likely to exist both as 
solution and colloidal species. The sorption behavior of americium and thorium is likely to be 
similar in the repository environment. 

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MDL-EBS-PA-000004 REV 02 5-2 August 2005 
Use in the Model: Section 6.3.3.1 and Table 6-6. 
5.2 EFFECT OF TEMPERATURE ON WASTE FORM COLLOIDS 
Assumption: Waste form colloid suspensions produced at 90C are stable at 25C. 
Rationale: As flocculation is a particle collision process, temperature will play a role in the rate 
of coagulation. If colloidal suspensions are stable at 90C, they should be stable at 25C. 
Confirmation Status: The assumption that colloid suspensions will remain stable as temperature 
drops is correct. However, without the high temperatures (i.e., 90C for crushed glass immersion 
tests) used in the experiments on waste forms, reaction rates would have been slow resulting in 
less colloid generation. There are no data available on colloid formation at 25C for DHLWG; 
however, data are available for spent nuclear fuel colloids between 25C and 90C. 
Use in the Model: Section 6.3.2. 
5.3 IRREVERSIBLE (EMBEDDED) ATTACHMENT OF PLUTONIUM AND 
AMERICIUM TO DHLW GLASS COLLOIDS 
Assumption: Plutonium and americium are associated irreversibly (or embedded) with DHLW 
glass colloids. 
Rationale: Plutonium and americium have been found associated with colloids and suspended 
particles generated during DHLW glass weathering [CRWMS M&O 2001 [DIRS 154071]; 
DTN: LL000123351021.117 [DIRS 143308]; Bates et al. 1992 [DIRS 100704]. 
Confirmation Status: The assumption that all measured plutonium and americium associated 
irreversibly (embedded) with DHLW glass colloids results in maximized plutonium and 
americium association with these colloids. Hence, this approach is conservative for the 
TSPA-LA model analysis. 
Use in the Model: Sections 6.3.1 and 6.3.3. 
5.4 COLLOID SORPTION AT THE AIRWATER INTERFACE 
Assumption: Colloid sorption at the airwater interface will not occur within the waste package. 
Rationale: Colloid sorption at the airwater interface may occur within the waste package but is 
not considered in the abstraction. Colloid attachment to airwater interfaces commonly occurs in 
unsaturated environments and may limit mobile colloid generation and migration (Wan and 
Tokunga 1997 [DIRS 108285]). This phenomenon is dependent on the interface surface area, 
electrostatic charge on the particles, and the salinity of the aqueous phase. At low water 
saturations, 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 the estimating 
geometry and surface area is complicated, more so under changing saturation state. At high 
water saturations the majority of the gas is present at small gas bubbles that may migrate, 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 5-3 August 2005 
transporting sorbed colloids. However, a proportion of the bubbles may become trapped and will 
effectively immobilize sorbed colloids. 
Neglect of colloid sorption onto the airwater interface in the TSPA calculations should be 
conservative. Wan and Tokunga (1997 [DIRS 108285]) have shown that Na-montmorillonite 
clay will be excluded from airwater interfaces at any given pH owing to their structure; in 
contrast, kaolinite and some actinide oxides have a strong attraction to airwater interfaces 
(Buck et al. 2004 [DIRS 171479]). 
Confirmation Status: Smectite clay sorption to airwater interfaces should not occur; hence, the 
process may be irrelevant for the clay colloidal particles under consideration in this model. If 
this process occurs with other suspended particles within the waste package, it may result in 
greater attenuation of colloids than currently predicted. 
Use in the Model: This assumption is used throughout this report. This process is not modeled 
and, therefore, a discussion on its application is not included in this report. 
5.5 PLUTONIUM AND AMERICIUM SORPTION TO IRON OXIDES 
Assumption: Kd values obtained for plutonium and americium on hematite are valid for other 
forms of iron oxyhydroxides. 
Rationale: Partition coefficients have been reported for plutonium and americium sorption onto 
hematite; however, the waste package will contain an array of different iron oxyhydroxide 
colloid phases for which specific Kd values are not available. The term iron oxyhydroxides in 
this model report includes hematite, goethite, ferrihydrite, and other iron oxides and hydroxide 
phases. Results from sorption experiments suggest that the concentration of sorbed plutonium 
with surface area is similar in different iron oxide phases, ferrihydrite, goethite, and hematite 
(Table 5-2). 
Table 5-2. Sorption Plutonium (V) of Ferrihydrite, Goethite, and Hematite 
Iron Oxide 
Plutonium 
Sorption (mol/m2) 
1.18  10-10 
2.44  10-10 
4.51  10-10 
9.98  10-10 
Ferrihydrite 
1.67  10-9 
1.12  10-10 
2.37  10-10 
4.49  10-10 
9.96  10-10 
Goethite 
1.57  10-9 

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MDL-EBS-PA-000004 REV 02 5-4 August 2005 
Table 5-2. Sorption Plutonium (V) of Ferrihydrite, Goethite, and Hematite 
Iron Oxide 
Plutonium 
Sorption (mol/m2) 
2.86  10-10 
5.73  10-10 
1.15  10-9 
2.41  10-9 
Hematite 
3.88  10-9 
Source: DTN: LA9910SK831341.003 [DIRS 142466]. 
Notes : Plutonium Sorption was as Pu (V) and was measured in 
moles of Pu(V) sorbed per square meter of iron oxide 
surface area, depending on which oxide was used 
from the iron oxide column (ferrihydrite, hematite or 
goethite. 
Confirmation Status: No confirmation required. The data shows similar plutonium sorption 
behavior for different iron oxides. 
Use in the Model: Sections 6.3.1 and 6.3.3. 

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MDL-EBS-PA-000004 REV 02 6-1 August 2005 
6. MODEL DISCUSSION 
This section describes the: (1) model objectives, (2) features, events, and processes considered in 
the analysis, (3) base-case model, (4) consideration of alternative models, (5) model formulation 
for base-case assessment, (6) base-case model results, (7) description and analysis of the barrier 
capability, and (8) evaluation of alternative models. The supporting information 
(indirect/corroborative data) used in the analysis is described in Section 4.1. 
6.1 MODEL OBJECTIVES 
The purpose of this colloid model abstraction is to provide colloid source terms to the TSPA 
system model. Colloids will likely be produced as a result of the alteration of DHLW glass and 
CSNF waste forms. The abundance of colloids within the breached waste package will depend 
on the degree of waste form alteration, on the quantity of steel corrosion product colloids 
formed, and on quantities of natural colloids present in seepage water. Colloid abundance and 
suspension stability also depend on the ionic strength, pH, and colloid content of waters entering 
the waste package from the drift. 
In this section the types and concentrations of colloids that could potentially be generated in the 
waste package from the degradation of the waste (waste form colloids) and corrosion of the 
waste package internal components (corrosion-generated iron oxyhydroxide colloids) are 
described. Concrete and cement are excluded from use in the emplacement drifts by project 
design criteria. The treatment here on man-made materials is focused only on steel corrosion 
products. Potential interactions are possible between inorganic colloids and microbes or 
organics. Also the observed types and concentrations of colloids naturally present in 
groundwater (groundwater colloids) that would potentially seep or flow into the engineered 
barrier system during the postclosure regulatory period are described. In addition, sorption and 
desorption of dissolved radionuclides to and from colloid surfaces are discussed as well as the 
nature of colloids anticipated in the repository environment. The abstraction is intended to 
capture the most important characteristics of radionuclidecolloid behavior for use in predicting 
the potential impact of colloid-facilitated radionuclide transport on repository performance. 
Colloid transport in the waste package and drift is discussed in EBS Radionuclide Transport 
Abstraction (BSC 2004 [DIRS 169868]). 
Output quantities from the waste form and engineered barrier system colloid abstraction to the 
unsaturated zone include DHLW glass colloid concentrations, corrosion product (iron 
oxyhydroxide) colloids, and natural groundwater colloids, discussed in Sections 6.3.1 and 6.3.2. 
Colloids potentially produced from degradation of CSNF and DSNF are not modeled, for reasons 
discussed in Section 6.3.1. Three types of radionuclide association with colloids are modeled 
and described in Section 6.3.3: (1) irreversibly sorbed (also termed embedded) plutonium and 
americium in the DHLW glass waste form colloids; (2) irreversible sorption of plutonium and 
americium in iron oxyhydroxide colloids; and (3) reversible sorption of cesium, americium, 
plutonium, thorium, and protactinium onto all colloid types. 
Model inputs from upstream data sources include ionic strength and pH of the in-package fluid 
resulting from reaction of water with the waste, ionic strength and pH of the seepage water 
entering the drift, estimates of colloid masses formed from waste form degradation, the 

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MDL-EBS-PA-000004 REV 02 6-2 August 2005 
concentrations of radionuclides generated from degradation of the waste, and Kd values for 
radionuclides on mineral phases and colloids. 
Model outputs feed the unsaturated zone transport models and include (as described above) 
DHLW glass colloids with attached radionuclides, irreversibly and reversibly sorbed 
radionuclides, and colloid masses. 
The uncertainty associated with the development of parameters representing major processes or 
features modeled in this analysis is described in general terms in Section 6.6.2. Additionally, 
specific details on the types of uncertainties associated with the various components of the model 
elements and the cause(s) for these uncertainties are provided within Sections 6 and 6.3. 
6.2 FEATURES, EVENTS, AND PROCESSES CONSIDERED IN MODEL 
As described in Technical Work Plan for Postclosure Waste Form Modeling. (BSC 2005 
[DIRS 173246], Section 2.5), this report addresses the key colloid-oriented FEPs for the waste 
form and engineered barrier system components of the repository that are screened in 
(i.e., Included FEPs) for TSPA-LA. Table 6-1 provides a list of FEPs that are included in TSPALA 
models (DTN: MO0407SEPFEPLA.000 [DIRS 170760]). This report summarizes details of 
FEP implementation in TSPA-LA, and Table 6-1 provides specific references to sections 
concerning FEPs within this report. 
Table 6-1. Features, Events, and Processes Included (Screened In) in TSPA-LA and Addressed in 
this Report 
FEP No. FEP Names Location in This Report 
2.1.09.02.0A Chemical interaction with corrosion products Sections 6.3.1.1, 6.3.1.3, 6.3.2.1, 6.3.2.3, 
6.3.3.1, 6.3.3.2 
2.1.09.16.0A Formation of pseudocolloids (natural) in EBS Sections 6.3.1, 6.3.1.4, 6.3.2.1, 6.3.2.4, 
6.3.3.1 
2.1.09.17.0A Formation of pseudocolloids (corrosion product) in EBS Sections 6.3.1, 6.3.1.3, 6.3.2.2, 6.3.3.1, 
6.3.3.2 
2.1.09.23.0A Stability of colloids in EBS Sections 6.3.2, 6.3.3, 6.5 
2.1.09.25.0A Formation of colloids (waste-form) by coprecipitation in 
the EBS 
Section 6.3.1.1 
NOTE: EBS = engineered barrier system; TSPA = total system performance assessment. 
6.3 BASE-CASE MODEL 
A discussion of the colloid model is presented here, organized by major concepts considered and 
the specific ideas and values developed and incorporated into the model. Accordingly, the 
following subjects are discussed: (1) colloid formation from DHLW glass; (2) CSNF and DSNF; 
(3) formation of corrosion products, including colloids, from degradation of steel within the 
waste package and drift; (4) natural occurrence of groundwater colloids; (5) colloid suspension 
stability and concentration of all modeled colloid types; (6) radionuclide sorption to colloids; and 
(7) microbial interactions. 

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MDL-EBS-PA-000004 REV 02 6-3 August 2005 
Although primary plutonium-bearing colloids are generated from spent nuclear fuel, the quantity 
appears to be small in tests with intact or partially intact cladding (Wilson 1990 [DIRS 100793]) 
and are presumed to sorb to iron oxyhydroxides corrosion products. 
Colloids formed from corrosion of introduced materials are modeled only as hematite and 
goethite colloids. As described in Section 6.1, materials introduced during repository 
construction may include various steel alloys, organic liquids, and cementitious materials. 
Concrete and cement are excluded from use in the emplacement drifts and, therefore, colloids 
derived from these materials are not considered. 
6.3.1 Colloid Formation and Occurrence 
Following is a discussion of colloid types, definitions, and pertinence to the repository. 
A summary of the colloid types modeled in this abstraction is provided at the end of this section. 
Figure 6-1 is a schematic illustrating colloid types. 
Colloids by definition have at least one dimension between 1 nm and 1 m. Often, particularly 
in reporting of experimental results, the upper end of the colloid size range is 450 nm and the 
lower limit is >2 nm, due to conventional dimensions of laboratory equipment (primarily filters). 
True ColloidsTrue colloids are colloidal-sized assemblages formed from the hydrolysis and 
polymerization of actinide ions dissolved in solution. They may form in the waste package and 
engineered barrier system during waste form degradation and radionuclide transport. True 
colloids are also called real colloids, Type I colloids, Eigenkolloide, and intrinsic colloids. The 
formation of true colloids will depend on the solution chemistry, pH, and temperature. The 
speciation and formation of plutonium intrinsic colloids from an initial Pu(V) solution were 
investigated by Romanovski et al. (1999 [DIRS 170890]) in surrogate Yucca Mountain J-13 well 
water under an air atmosphere at 25C. A large proportion of plutonium is present as soluble 
Pu(V) through the pH range of 1 to 8 at 25C. Degradation of CSNF may form real colloids 
(e.g., plutonium oxides and hydroxides) consisting of americium, plutonium, and rare earths 
(Grambow et al. 1996 [DIRS 113253]), but these are anticipated to sorb to iron 
oxide/oxyhydroxides, forming a pseudocolloid. Alternative conceptual models consider 
reversible and irreversible association of plutonium and americium with U(VI) phases formed 
during spent nuclear fuel corrosion. 
Waste Form ColloidsWaste form colloids, termed primary colloids, may provide a 
contribution to colloid-facilitated radionuclide transport. These colloids form from the 
nucleation of colloids from waste form dissolution and from spallation (flaking off) of 
colloid-sized waste form alteration products. Waste form colloids may contain radionuclides and 
have the potential for increasing mobile radionuclide concentrations to levels higher than 
achievable with real colloids or other pseudocolloids. Dissolved radionuclides and other ions 
may coprecipitate to form colloids. Coprecipitates may consist of radionuclides bound in the 
crystal lattice of a dominating mineral phase or may consist of radionuclides engulfed by a 
dominating mineral phase. Traexler et al. (2004 [DIRS 172072]) listed potential waste 
form-derived colloids including, schoepite {[(UO2)8O2(OH)12](H2O)12}; soddyite 
{(UO2)2SiO42H2O}, uranophane {Ca(UO2)2(SiO3OH)25H2O}; boltwoodite 
{KH(UO2)(SiO4)1.5H2O}; cesium-molybdenum uranyl hydroxide; clay particles with brockite 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-4 August 2005 
{(Ca,Th,Ce)PO4H2O} inclusions, asbolan {(CO,Ni)1-y(MnO2)2-x(OH)2-2y-2xnH2O, birnessite 
{(Na,Ca,K)x(Mn4+,Mn3+)2O41.5(H2O)}, weeksite {K2(UO2)2Si6O154(H2O)}, calcite {CaCO3}, 
dolomite{CaMg(CO3)2}, metal oxides, Na-silicates, Al-oxides, analcime {NaAlSi2O6H2O}, 
hydrated Ca-silicates, and apatite-type phases. 
The most abundant waste type is CSNF (uranium oxide) from commercially owned and operated 
electric power reactors. DSNF is a diverse collection of waste from reactors at DOE nuclear 
complex sites. Defense high-level waste glass is an alkali borosilicate glass-based waste form 
containing radionuclides. 
Spent nuclear fuel corrosion can result in the generation of plutonium- and americium-bearing 
colloids (Wilson 1990 [DIRS 100793]; Wilson 1990 [DIRS 121808]; Finn et al. 1994 
[DIRS 100746]; McNamara et al. (2005 [DIRS 174068]). Production of plutonium colloids was 
observed to be highest with bare fuel fragments. In the presence of metal surfaces (e.g., steel 
vessels or Zircaloy), a reduction in plutonium colloid concentrations was observed (Wilson 1990 
[DIRS 100793]; Mertz et al. 2003 [DIRS 162032]). Furthermore, the concentration of plutonium 
colloids was observed to decrease with time in all tests. In experiments where cladding was 
present, plutonium colloid concentrations were reduced by three to four orders of magnitude 
(Wilson 1990 [DIRS 100793]). These observations may be supported by the intermittent drip 
spent nuclear fuel corrosion tests (ATM-103 HDR and ATM-106 HDR) (Mertz et al. 2003 
[DIRS 162032]); however, other uncertainties need to be considered in these tests, such as 
amount of spent nuclear fuel reaction, the amount and nature of water contact, and the physical 
barriers that will limit colloid transport from the altered fuel surface to the collection vessel. 
Various colloidal phases (e.g., rare earth element-bearing phases) were detected in the leachates 
at early times (Finn et al. 1994 [DIRS 100746]); but, at later times, colloid release virtually 
stopped, possibly due to filtration, attachment to air-water interfaces, or sorption by the metallic 
holder used in the tests. There was evidence of material buildup consisting of spent nuclear fuel 
particles and radionuclide-bearing uranium (VI) phases on these Zircaloy holders 
(Finn et al. 1998 [DIRS 100392]). Unfiltered plutonium concentrations measured from the 
ATM-103 HDR and ATM-106 HDR following removal of the Zircaloy holder were slightly 
higher than those reported before the changes. These results indicate that small but measurable 
levels of plutonium colloids were released between 5.2 and 8.7 years after the physical 
impediments to colloid release were partially removed (Mertz 2003 [DIRS 162032]). Plutonium 
concentrations were between 1  10-10 and 1  10-9 mol/L; only slightly lower than the levels 
reported by Finn et al. (1994 [DIRS 100746]) for the initial samplings. These values indicate 
that some plutonium colloids may continue to be released from CSNF. In corrosion tests on 
crushed and sieved H.B. Robinson spent nuclear fuel under immersion conditions in deionized 
water, plutonium, americium, and neptunium were found associated with colloidal metastudtite 
{UO4  2H2O} (McNamara et al. (2005 [DIRS 174068]). The results from these various spent 
nuclear fuel corrosion tests suggest that colloid formation during spent nuclear fuel weathering 
may happen; however, the occurrence of spent nuclear fuel-derived colloids has been excluded in 
the base model because of the effect of metallic components in reducing colloid concentrations. 
Physical filtration of colloids is not explicitly called out in the model abstraction; however, 
filtration is likely to have occurred in all tests where colloids have been reported. Colloid 
filtration may result in the removal of colloids by pore clogging, sieving, or straining 
(Aguilar et al. 1999 [DIRS 161648]). There are two types of physical filtration (Wan and 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-5 August 2005 
Tokunaga 1997 [DIRS 108285]), conventional straining and film straining. Conventional 
straining filters colloids if they are larger than a pore throat diameter or fracture aperture. If 
unsaturated conditions occur, colloids may be filtered by thin-film straining where the particle 
size is greater than the thickness of the adsorbed water film. Physical clogging of Zircaloy filters 
by released uranium particles was observed in spent nuclear fuel intermittent drip tests 
(Buck et al. 1997 [DIRS 112904]; Finn et al. 1998 [DIRS 100392]). It is possible filtration in 
both the intermittent drip tests on spent nuclear fuel (Finn et al. 1994 [DIRS 100746], 
Mertz et al. 2003 [DIRS 162032]) and in spent nuclear fuel immersion tests with cladding 
materials (Wilson 1990 [DIRS 100793]) led to the low colloid concentrations observed in these 
experiments. Particle shape may also have an effect on clogging probabilities and it is likely that 
the more elongated particles, common to many of the U(VI) oxyhydroxides and U(VI) silicates, 
may have a greater tendency to cause clogging than other types of colloids. 
Investigations of the degradation of DHLW glass (DTNs: MO0407ANLGNN02.608 
[DIRS 171277]; LL000123351021.117 [DIRS 143308]; CRWMS M&O 2001 [DIRS 154071]; 
and Bates et al. 1992 [DIRS 100704]) concluded that waste form colloids have formed by 
precipitation of radionuclide-bearing phases within the glass-weathering rind that later became 
associated with colloid-sized clay phases. Spallation of fragments of clay (primarily) and 
radionuclide-bearing phases from the alteration layers formed on the waste form surfaces may 
also result in waste form colloid generation. 
PseudocolloidsPseudocolloids consist of nonradioactive particles with radionuclides sorbed to 
them. Natural pseudocolloids include mineral fragments formed from the host rock, microbes 
and microbe fragments, and humic and fulvic acids. Pseudocolloids form as a result of dissolved 
(aqueous) radionuclides or smaller radionuclide colloids, or both sorbing to existing colloids 
(e.g., clay, iron oxyhydroxides, silica). At the YMP, pseudocolloids may develop as a result of 
radionuclide association with waste form colloids formed during degradation of DHLW glass, 
radionuclide association with corrosion-product colloids formed during corrosion of waste 
package materials and other metallic materials, and radionuclide association with naturally 
occurring groundwater colloids. 
SeepageGroundwater ColloidsNaturally occurring colloids (referred to in this report as 
groundwater colloids) may be inorganic (mineral fragments) or organic (microbes and humic 
substances). At YMP, inorganic groundwater colloids, primarily consisting of clay, silica, and 
iron oxyhydroxides, may be transported in seepage water into the repository from the vadose 
zone above it or may be formed due to changes in chemical conditions near the repository. 
Pseudocolloids may form due to the sorption onto these natural colloids of radionuclides 
mobilized from degradation of the waste form. Pseudocolloids thus formed in the waste and 
engineered barrier system may influence radionuclide transport. At YMP, organic colloids are 
not important. Humic substances are not sufficiently abundant in YMP groundwaters to impact 
transport (Section 6.3.1.4). Microbes, as large colloids, are susceptible to filtration during 
transport. 
Colloids are modeled in the abstraction as follows: 
 Waste form colloids. Colloids from the corrosion of DHLW glass (Section 6.3.1.1) 
are modeled as smectite colloids with associated (embedded) plutonium and 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-6 August 2005 
americium and reversible sorption of plutonium, americium, protactinium, thorium, 
and cesium. Colloids from the corrosion of CSNF and DSNF are not modeled in the 
base case. 
 Corrosion product colloids. Colloids from the corrosion of steel components in the 
engineered barrier system are modeled as hematite colloids, as they will likely be a 
mixture or combination of hydrous ferric oxide, ferrihydrite, and Fe2O3 (Section 
6.3.1.3). 
 Seepage/groundwater colloids. Naturally occurring colloids are modeled as smectite 
colloids (Section 6.3.1.4). Groundwaters will also contain inorganic silica and iron 
oxides. However, the low sorptive properties of silica are such that they can be 
ignored, and waste package corrosion products will dwarf the quantity of iron oxides 
from natural groundwater. Organic colloids in groundwaterhumic and fulvic acids, 
microbes and their detritus and metabolic products, are not modeled. 
For ionic strengths greater than or equal to 0.05 M, the concentration of colloids is negligible and 
is set to low values for modeling purposes. Complete flocculation of hematite colloids occurs 
at >0.05M NaCl(aq) (Liang and Morgan 1990 [DIRS 109524]). Hence, the use of a maximum 
ionic strength of 0.05M for the stability of iron oxide colloid suspensions is likely to be 
conservative. Montmorillonite clay colloid suspensions have been shown to be unstable at ionic 
strengths >0.05M depending on pH (Tombcz et al. 1990 [DIRS 112690], Tombcz and 
Szekeres 2004 [DIRS 174073]). Montmorillonite suspensions coagulated in the presence of 0.01 
M, 0.025 M, and 0.04 M NaCl at pH values of 2, 4, and 8, respectively. The Schulze-Hardy 
particle destabilization effects are not taken into account in the model and may provide further 
conservatism to the TSPA-LA model analysis (van Olphen 1977 [DIRS 114428]). The Schulze- 
Hardy rule states that the critical coagulation concentration (see section 6.3.2) for a colloid is 
sensitive to the valence of the counter-ions, where high valence gives a low critical coagulation 
concentration.The effectiveness of monomeric ions as coagulants for hematite particles, under 
comparable pH conditions is given by the decreasing sequence: Phosphate > oxalate > Mg2+ > 
SO4 
-> Ca2+ > Na+. 
Source: Adapted from CRWMS M&O 2001 [DIRS 154071], Figure 1. 
NOTE: The two major types of modeled radionuclide-bearing colloids are depicted: (a) pseudocolloids (clay or iron 
oxide) and (b) waste form colloids (clay). The radionuclide (RNn+) associated with the pseudo and waste 
form colloids can be an ionic species, a real colloid, or a discrete radionuclide-bearing phase. 
Figure 6-1. Radionuclide-Bearing Colloid Types 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-7 August 2005 
6.3.1.1 Colloids from the Corrosion of Borosilicate Glass 
Colloids produced from the degradation of alkali borosilicate glasses have been observed to be 
primarily smectite clay occasionally containing discrete radionuclide-bearing phases (DTNs: 
MO0407ANLGNN02.608 [DIRS 171277], LL000123351021.117 [DIRS 143308]; 
CRWMS M&O 2001 [DIRS 154071]; and Bates et al. 1992 [DIRS 100704]). Depending on the 
glass composition, different minor phases have been observed; such as brockite 
(Bates et al. 1992 [DIRS 100704]) or rhabdophane (Buck and Bates 1999 [DIRS 109494]) in 
thorium- and phosphorus-containing glasses. These phases were not observed in SRL202A and 
SRL131A glass tests used to establish plutonium concentration with colloids for the TSPA-LA. 
The colloidal properties of any radionuclide-bearing phase associated with smectite clay colloids 
will be governed by the properties of that smectite clay. Therefore, the stability of smectite 
(montmorillonite) colloid suspensions will control the mobility of the associated radionuclidebearing 
phases. 
Figure 6-2 depicts the process of waste form weathering with the subsequent formation of 
colloids derived from the waste form. As the waste form corrodes, an alteration layer forms, 
grows, and evolves as a layer on its surface. In the case of DLHW glass, a smectite-like phase 
forms initially on the glass surface (CRWMS M&O 2001 [DIRS 154071], Section 6.2.1). 
Source: Modified from CRWMS M&O 2001 [DIRS 154071], Figure 2. 
NOTE: Schematic of colloid formation from waste-form corrosion whereby several processes are represented: 
(1) spallation of radionuclide bearing alteration phases from the waste form that are within the colloidal 
size range, and (2) nucleation of colloids from ions dissolved from the waste form and sorption of ionic 
radionuclide species or radionuclide-bearing phases. Particle growth occurs by precipitation on nuclei 
and is controlled by factors such as ionic strength, temperature, pH, and solubility. When particle 
diameters exceed 1 m or solution chemistry destabilizes the colloids, deposition, or coagulation and 
gravitational settling of radionuclide-bearing alteration phases occurs. 
RN = radionuclide. 
Figure 6-2. Schematic of Colloid Formation from Waste-Form Corrosion 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-8 August 2005 
In static-saturated tests, colloids developed and increased in concentration (population) with 
time, up to a point where the colloid concentration reached a maximum and produced an unstable 
dispersion, owing to the concomitant increase in ionic strength to a threshold above which the 
colloids flocculated and settled by gravity (see Figure 6-4). 
Based on results from product consistency tests on a plutonium-doped alkali borosilicate glass 
(SRL131A) (DTN LL000905312241.018 [DIRS 152621]), the following approximate 
relationship was developed from inspection of the data presented in Figures 6-3 and 6-4, which 
may be used to relate the plutonium concentration to colloid concentration: 
1 ppm colloids = 2  10-8 mol/L Pucoll (Eq. 1) 
Pucoll is the concentration of plutonium sorbed to colloids. Colloids from similar PCT-B 
experiments on DHLW glass have been identified as montmorillonite clay (Buck et al. 1994 
[DIRS 122116], Feng et al. 1994 [DIRS 101611]). To enable the use of bounding values for 
applying the relationship, the approximate maximum (and bounding) value for Pucoll of 
about 1  10-7 mol/L (Figure 6-4) was adopted to correspond with the approximate maximum 
(and bounding) colloid mass concentration of 5.0 ppm (Figure 6-3); although, this value may 
depend on the starting concentration of plutonium in the actual waste glass. 
Source: DTN LL000905312241.018 [DIRS 152621], MOL20010124.0346, SN1644, p. 32, columns B&H. 
NOTE: Total colloid concentration (from dynamic light scattering measurements) as a function of test duration for 
the SRL 131A glass at 2,000/m (T = 90C). The role of pH is not addressed in this figure. 
Figure 6-3. Concentrations of Colloids as a Function of DHLW Glass Corrosion Test Duration 
Division of maximum Pucoll by the maximum colloid mass concentration yields the relationship 
in Equation 1. It is, therefore, recommended that the TSPA LA model compute the waste form 
colloid concentration by first calculating the Pucoll concentration between 1  10-11 mol/L and 1  
10-7 mol/L and then linearly calculating the waste form colloid concentration between 5  10-4 
mg/L and 5.0 mg/L, such that the upper end of the Pucoll concentration (1  10-7 mol/L) is 
equivalent to 5 mg/L and lower end of Pucoll concentration (1  10-11 mol/L) is equivalent 
to 5  10-4 mg/L 

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MDL-EBS-PA-000004 REV 02 6-9 August 2005 
Evidence from other DHLW glass tests demonstrated that plutonium and americium behaved 
similarly (DTNs: MO0407ANLGNN02.608 [DIRS 171277], LL000123351021.117 
[DIRS 143308]; Bates et al. 1992 [DIRS 100704]). As the plutonium concentration used in the 
DHLW glass tests is similar to the highest anticipated concentrations in high-level waste 
production glasses, these results are directly relevant to establishing colloid-associated plutonium 
concentrations. The effect of the plutonium content from the weapons plutonium disposition 
lanthanum borosilicate (LaBS) glass on the total plutonium colloid concentration is unclear. The 
blended composition is likely to have a concentration similar to that used in alpha-doped tests on 
R7T7 reported by Pirlet (2001 [DIRS 174351] and Mnard et al. (1998 [DIRS 171053]). 
Although these tests had less vigorous reaction conditions, the maximum plutonium 
concentration was 1.5  10-7 M. The colloid-associated americium concentration was calculated 
in TSPA model calculations by multiplying the ratio of americium to plutonium in the 
radionuclide inventory by the plutonium concentration at each time step. 
Source: DTN: LL000905312241.018 [DIRS 152621]; MOL.20001130.0021, p. 70. 
NOTE: Plutonium-bearing colloid concentration (units = moles/liter) as a function of time for corrosion tests on 
glass samples SRL 202A and SRL 131A of SA/V of 2,000/m and 20,000/m (at 90C). 
Figure 6-4. Plutonium Concentration as a Function of Time in Corrosion Tests 
6.3.1.2 Colloids from the Corrosion of Commercial and DOE Spent Nuclear Fuel 
Experiments that either endeavor to simulate waste package conditions or contained metallic 
components indicated that spent nuclear fuelderived colloid concentrations can be low. 
Furthermore, there is a reasonable indication from corrosion tests that U(VI) phases will undergo 
retardation either by physical filtration (Buck et al. 1997 [DIRS 112904]; Finn et al. 1998 
[DIRS 100392]), attachment to Zircaloy (Wilson 1990 [DIRS 100793]), attachment to iron 
phases (Grambow et al. (1996 [DIRS 113253]), or attachment to air-water interfaces 
(McNamara et al. 2005 [DIRS 174068]; Buck et al. 2004 [DIRS 171479]). The generation of 
plutonium-rich phases from high burn-up rim polygonization on the spent nuclear fuel surface 
could result in irreversible colloids formation as these would have formed during reactor 
operations and may be kinetically stable. 
1.E-12 
1.E-11 
1.E-10 
1.E-09 
1.E-08 
1.E-07 
1.E-06 
0 200 400 600 800 1000 1200 
Time (Days) 
Plutonium Colloids Concentration 
131A_20K 
131A_2K 
202A_2K 
202A_20K

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-10 August 2005 
In following sections the formation of spent nuclear fuel-derived colloids from corrosion tests is 
discussed. The major types of tests included in the discussion are immersion tests with 
fragments and crushed solids and unsaturated drip tests with fragments and monolith samples. 
There are three main types of spent nuclear fuel of concern; commerical spent nuclear fuel 
(CSNF), defense (uranium metal) nuclear fuel (DSNF), and mixed oxide (MOX) fuel. 
Immersion Tests on Commercial Spent Nuclear Fuel 
Wilson (1990 [DIRS 100793]) conducted tests on five in-rod segments of spent nuclear fuel 
representing a range of potential degrees of cladding failure. Tests on bare fuel fragments 
exhibited the highest plutonium colloid release rates, while all tests on undefected, slit-defect, 
and hole-defect samples exhibited releases of plutonium colloids that were approximately 10,000 
times lower, indeed, most plutonium was in a soluble form in the Zircaloy-containing tests 
(Figure 6-5a and Table 6-2). The concentration of colloidal plutonium peaked at 1  10-6 M in 
the uncanistered fuel tests but remained above 1  10-8 M at 25C for the duration of the test. 
The CSNF testing results from Wilson (1990 [DIRS 100793]) at 85C where metallic 
components were present were similar to those from unsaturated drip tests on CSNF reported by 
Mertz et al. (2003 [DIRS 162032]) between 5 and 7 years where plutonium concentrations are 
below 10-9 mol/L. 
Source: Adapted from Wilson 1990 [DIRS 100793]. Appendix A. 
NOTE: Plutonium colloid concentration (>1.8-nm and <400-nm size fraction from filtration data) as a function of 
test duration for two types of low-burnup spent nuclear fuels. All tests conducted at 85C except HBR/BF- 
25 that was conducted at 25C. HBR= H.B. Robinson Fuel (30.2 MWd/KgM), TP= Turkey Point Fuel 
(27.7 MWd/KgM), BF= bare fuel, SD = Slit Defect, HD = Hole Defect, and UD = undefected. 
Figure 6-5a. Concentrations of Plutonium Colloids as a Function of Spent Nuclear Fuel Corrosion Test 
Duration 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-11 August 2005 
Table 6-2. Plutonium Colloid Concentrations Obtained from Immersion Tests on Spent Nuclear Fuel 
Time (days) Vol. 
(mL) 
0.4 m 
(pCi) 
1.8 nm 
(pCi) 
Colloid 
(pCi) 
[Pu] 
(moles/L) 
log [Pu] 
(mol/L) 
A.2 HBR/BF-25 
1 15 1.23E+04 6.62E+03 5.68E+03 2.51E-07 -6.60017 
7 15 8.29E+03 3.90E+03 4.39E+03 1.94E-07 -6.71205 
14 20 6.40E+03 2.85E+03 3.55E+03 1.57E-07 -6.80429 
33 30 2.94E+03 8.58E+02 2.08E+03 9.20E-08 -7.03603 
62 30 1.91E+03 4.48E+02 1.46E+03 6.46E-08 -7.18957 
119 30 1.23E+03 2.31E+02 9.99E+02 4.42E-08 -7.35495 
174 250 9.14E+02 1.75E+02 7.39E+02 3.27E-08 -7.48587 
A.3 TP/BF-85 
1 15 2.59E+03 1.20E+02 2.47E+03 1.06E-07 -6.97516 
7 15 5.32E+02 6.08E+01 4.71E+02 2.02E-08 -7.69465 
14 20 3.88E+02 9.55E+01 2.93E+02 1.25E-08 -7.90173 
33 30 6.76E+01 3.60E+01 3.16E+01 1.35E-09 -8.86817 
62 30 1.35E+01 1.44E+01 -9.00E-01 -3.86E-11 NULL 
119 30 4.50E+00 4.23E+00 2.70E-01 1.16E-11 -10.9365 
174 250 4.05E+00 1.35E+00 2.70E+00 1.16E-10 -9.93649 
A.4 HBR/BF-85 
1 15 1.97E+04 7.48E+02 1.90E+04 8.38E-07 -6.07686 
7 15 3.68E+03 3.18E+02 3.36E+03 1.49E-07 -6.82792 
14 20 1.45E+03 1.91E+02 1.26E+03 5.57E-08 -7.25449 
33 30 4.82E+03 1.40E+02 4.68E+03 2.07E-07 -6.68427 
62 30 9.28E+02 1.20E+02 8.08E+02 3.57E-08 -7.4471 
119 30 9.05E+02 1.31E+00 9.04E+02 3.99E-08 -7.3985 
174 250 6.76E+00 9.01E-02 6.67E+00 2.95E-10 -9.5304 
A.5 HBR/SD-85 
7 15 2.25E-01 2.25E-01 0.00E+00 0.00E+00 NULL 
14 20 
33 30 3.15E-01 2.70E-01 4.50E-02 1.99E-12 -11.7013 
119 30 1.26E+00 2.70E-01 9.90E-01 4.38E-11 -10.3589 
174 250 1.35E-01 1.35E-01 0.00E+00 0.00E+00 NULL 
A.6 HBR/HD-85 
62 30 1.80E-01 9.01E-02 8.99E-02 3.97E-12 -11.4008 
A.7 HBR/UD-85 
119 30 5.41E-01 2.70E-01 2.71E-01 1.20E-11 -10.9215 
Source: Wilson (1990) [DIRS 100793], Appendix A. 
NOTES: Activities reported for plutonium as 239+240Pu (pCi/mL) from cycle 1. Conversion for HBR 
(H. B. Robinson Fuel) 4.42064  10-11 = 0.374*0.0000163/0.577/1000/239. 
Conversion for TP (Turkey Point Fuel) 4.28691  10-11 = 0.374*0.0000163/0.595/1000/239. 
Unsaturated Drip Tests on Spent Nuclear Fuel 
Unsaturated drip tests on spent nuclear fuels ATM-103 and ATM-106 were performed for over 8 
years (see Figure 6-5b). Fuel fragments were contained in a Zircaloy holder (Zircaloy-4). The 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-12 August 2005 
holder consists of a Zircaloy tube sitting atop a Zircaloy retainer with fifty 7-m-diameter holes. 
The Zircaloy retainers in the high-drip tests on ATM-103 and ATM-106 were replaced with new 
Zircaloy retainers (with 10-m-diameter holes) at the end of the 7.3-yr sampling interval; and 
these were subsequently replaced by gold retainers with 200-m-diameter mesh size at the end 
of the 7.7-yr sampling interval. Fuel holders and fuel fragments were set on a ledge inside each 
Stainless Steel Type 304L reaction vessel. At the start of each testing interval, about 5 mL of 
equilibrated J-13 water was placed inside each vessel. Each reaction vessel was sealed with a 
copper gasket and placed into a 90C oven. Each reaction vessel was injected with equilibrated 
J-13 well water. The average concentration of uranium in equilibrated J-13 well water entering 
the unsaturated drip test system was 6.7  10-9 mol/L (Thomas 2003 [DIRS 163048] Table 2, 
p. 18). 
Source: Developed from Table 4 p. 40 ATM103 HDR; Table 5, p. 41 ATM106 HDR, between 0 and 4.7 yrs of 
reaction (CRWMS 2000 [DIRS 131861], Table 18 p. 38, ATM 103 and ATM 106 HDR between 5.2 and 
8.7 yrs of reaction (Thomas 2003 [DIRS 163048]). 
NOTES: The Zircaloy holder was replaced at 7.3 yr and replaced with a gold mesh at 7.7 yr. In the ATM-103 HDR 
test, 33% of the expected volume was recovered in the 5.2-yr sample and the vessel was run dry at 6.8 yr 
(Mertz et al. 2003 [DIRS 162032]).In the ATM-106 HDR tests water loss occurred for the 5.2-yr and 6.8-yr 
samples, with 33% and 10% of the expected water volume recovered (Mertz et al. 2003 [DIRS 162032]). 
Values represent the total concentration measured. The reasons for the increased release at later times 
may be related to the change in experimental design after 6.3 yrs. 
Figure 6-5b. Released Fraction per Day of 239Pu and 241Am as a Function of Test Duration for the CSNF 
High Drip-Rate Tests. 
ATM-103 and ATM-106 Drip Tests 
The concentration of filterable uranium for the high drip-rate test is shown in Figure 65b. The 
Zircaloy retainer (with 50, 7-m drilled holes) was replaced twice, once at 7.3 yrs with a similar 
type of material and then at ~7.7 years with a gold mesh, which had a much larger open area. 
This may have had two effects: (a) reduce the total contact time of the water with the fuel and (b) 
allow larger particles to fall through. Evidence from examination of the retainers indicates that 
they are coated with particles of corroded fuel and alteration products (Finn et al. 1998 [DIRS 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-13 August 2005 
100392]). Therefore, the retainer may have filtered out many particles that would have 
otherwise reached the collection vessel. In 5.2-yr and 6.8-yr samplings, the testing system had 
lost water, which may have reduced the release of colloidal particles (Mertz et al. 2003 [DIRS 
162032]). 
Unsaturated Drip Tests on Uranium Oxide UO2 Monoliths 
Particles of both alteration phases and UO2 were observed in the leachate from the UO2 drip tests 
started several years before the spent nuclear fuel drip tests. The open form of the sample holder 
in these tests may have permitted particles to more easily enter the collection vessel than the 
spent nuclear fuel tests. Uranium silicate particles were found in the leachates of seven-year UO2 
tests (Wronkiewicz et al. 1996 [DIRS [102047]). The particles were elongated, displaying 
morphology typical of a uranyl silicate with attached anhedral uranium oxide particles. 
The principal colloidal carrier phase was identified as uranophane, ideally 
[Ca(H3O)2(UO2)2(SiO4)23H2O]. 
Colloids from DOE Spent Nuclear Fuel 
Of the approximately 250 different types of fuel in the DSNF inventory, metallic uranium fuel 
comprises approximately 85% (by MTHM) of that inventory. The degradation of metallic 
uranium fuel under unsaturated drip test conditions has been reported. (DOE 2003 
[DIRS 166268], Section 3). The reaction of uranium metal with water vapor produces binary 
oxides (Haschke 1998 [DIRS 174075]). The general reaction is: 
U + xH2O . UOx +xH2 
Values of the oxygen:uranium ratio (x) of the oxide product, vary from 2.0 to 2.2, depending on 
temperature and water pressure. In the case of U + moist air/O2, x in UOx is in the range 2.0 
to 2.7. The U + air/O2 is enhanced by moisture, but the U + H2O reaction is suppressed by 
oxygen. Haschke (1998 [DIRS 174075]) has shown that the oxide scale formed between 35C to 
100C has a laminar morphology, consistent with schoepite, with a relatively high surface area 
(15 m2/g). Mertz et al. (2000 [DIRS 162161]) lists schoepite, U4O9, and UO2, and uranium 
silicates as weathering products from uranium metal drip tests. The corrosion products observed 
during the long-term weathering of N-reactor fuel elements in the Hanford K-basins have been 
identified as meta-schoepite and studtite (McNamara et al. 2005 [DIRS 174068]). 
In the unsaturated tests with equilibrated J-13 well water, ~10-nm spheres of uranium oxides 
were seen after one month, which later became attached to larger clay colloids (DOE 2003 
[DIRS 166268], p. 3-37). Clays and other phases may be forming in the leachate of unsaturated 
drip tests due to evaporation during the course of the test. Sequential filtration showed that the 
released uranium was 40% colloidal and 45% particulate after seven days, with the colloidal 
fraction dropping and the particulate fraction increasing with time. After one month, the 
colloidal fraction had dropped to 10% and then to zero in the 8.5- and 11- months samples. 
Plutonium showed similar trends with 75% of the released plutonium colloidal at seven days 
falling to 8% colloidal and 90% particulate in the 8.5- and 11-month samples. It is likely that 
the colloid population was destabilized by increasing ionic strength as a result of the evaporating 
equilibrated J-13 well water solution over time, as evidenced by the formation of clay. 

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MDL-EBS-PA-000004 REV 02 6-14 August 2005 
In static batch tests, plutonium was released almost entirely as colloids with colloid 
concentrations of 2  10-8 to 2  10-7 M plutonium compared to generally less than 2  10-9 M 
plutonium dissolved. In a static batch test, colloids were not seen above the equilibrated 
J-13 well water background until the 32-day sample when the colloids were about ten times 
background. This 32-day sample was reexamined after nine months and no colloids were 
detectable (DOE 2003 [DIRS 166268], p. 3-43). This is in contrast to the sample from the 
unirradiated drip test that showed the colloid suspensions are stable for one year (DOE 2003 
[DIRS 166268], p. 3-39), but it was not clear if these stable colloid suspensions were the original 
UO2 spheres or the later clay and uranyl silicate colloids. Agglomeration of colloids was seen in 
the static batch and unsaturated drip tests, leading the authors to conclude that although a large 
number of colloids are generated (1012 particles per liter) they quickly agglomerate into a smaller 
number of large particles (DOE 2003 [DIRS 166268], p. 3-64). Temperature is usually used to 
accelerate waste form reaction to allow observation of weathering within a reasonable time 
frame; unfortunately, tests were not performed at lower temperatures that may have resulted in 
more stable colloidal sols. Corrosion testing of uranium metal fuel under drip conditions with 
additional discussion of the uncertainties associated with the DSNF unsaturated testing is 
reported by Goldberg et al. (DTN: MO0306ANLSF001.459 [DIRS 163910]). The highest 
observed colloidal uranium concentration was 3.5 mg U/L, and the highest 239Pu/238U ratio (0.13) 
was seen in the colloidal fraction which corresponds to a plutonium concentration of 2  10-6 M. 
Colloids from Mixed Oxide Fuel 
Initial plutonium release from unsaturated drip corrosion testing of mixed oxide fuel (MOX) and 
high burn-up fuel (ATM109-A, -B, and -C) has been reported by Cunnane (2001 
[DIRS 154818]). These tests resulted in lower release of plutonium than the lower burn-up 
ATM-106 and ATM-103 HDR tests. Any colloids from MOX are anticipated to behave like 
CSNF colloids. The test vessels for the ATM109 (-A through -C) HDR experiments were found 
to be dry in over 50% of the samplings (Thomas 2003 [DIRS 163048]. A dry test may have 
curtailed colloid transport. Although the potential for rim disintegration should be greater in the 
high burn-up and MOX fuels, evidence from the initial 0- to 3-yr high drip-rate unsaturated drip 
testing did not indicate increased release. 
Effects of CSNF Oxidation on Colloid Formation 
This section addresses the potential effects of CSNF oxidation (in the dry and humid air 
environments to which it might be exposed in nominal, seismic, and igneous intrusion scenarios) 
on colloid formation. 
A consequence of fuel oxidation is the possible disaggregation of the fuel pellets into fuelgrain-
sized powders (Dehaudt 2001 [DIRS 164037] Section 7.2.5.3). This can occur as a result 
of even minor oxidation at the grain boundaries, in low to medium burnup CSNF, to which hot 
air and water vapor can gain access through connected porosity at the grain boundaries. The asfabricated 
fuel grain sizes are close to 10 micrometers (e.g., a mean size of 92.4 m is given in 
Dehaudt 2001 [DIRS 164019], Section 5.2.2.5.4) and change little with irradiation in lowtomedium 
burnup CSNF (Dehaudt 2001 [DIRS 164037], Section 7.2.3.3). However, the grain sizes 
do change in the restructured rim region of high-burnup fuels; in this region the average grain 
size can be less than a micrometer (Spino et al. 1996 [DIRS 174080], pp. 179 through 190). 

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Oxidation induced disaggregation of these grains could produce submicrometer particles. Also, 
grain fragmentation due to the specific volume increase associated with oxidation of spent 
nuclear fuel with grain sizes greater than a micrometer (i.e., most of the spent nuclear fuel 
inventory) could also lead to submicrometer powders. If such powders were produced due to 
fuel oxidation prior to water contacting the fuels, the possible suspension of these particles upon 
water contact could produce a source of CSNF colloids. This section discusses the available data 
on the oxidation of CSNF under the conditions to which it is likely to be exposed in the nominal, 
seismic and igneous intrusion scenarios and the likelihood that such oxidation will lead to 
submicrometer particles that could be suspended as colloids upon water contact. 
Summary of fuel exposure conditions: 
Nominal and seismic scenarios - the fuel temperature could be up to about 300C. Because the 
pressure (air and water vapor) will not exceed one atmosphere, the air to which the fuel can be 
exposed will be dry (i.e., low relative humidity) until the temperature decreases to 
approximately 100C. The relative humidity will slowly approach 100% when the temperature 
decreases to below 100C. 
Igneous intrusion scenario  the fuel could be exposed to temperatures up to about 1,000C. 
The gas phase is likely to contain water vapor during and immediately after the intrusion event. 
During the cool down period (several month periods when the heat associated with the hot 
magma is dissipated) the gas phase can be considered to be dry air until the temperature 
decreases to approximately 100C. As in the case of the nominal and seismic scenarios, the 
relative humidity will approach 100% when the temperature decreases below 100C. 
The following paragraphs summarize available data that are pertinent to assessing if the initial 
exposure of CSNF to the dry air oxidizing conditions followed by exposure to humid air 
(approaching 100% relative humidity), as would be expected for the above exposure scenarios, is 
likely to lead to formation of submicrometer particles that would be available for suspension as 
colloids when water contacts the oxidized fuel. Available data on the size distributions of fuel 
powders produced in the oxidation of CSNF are discussed first. The expected behavior of the 
oxidized fuel, upon contact with humid air and water are then addressed. 
The oxidation of CSNF has been extensively investigated and the available information for 
temperatures below 400C has been summarized in a fairly recent comprehensive review 
(Dehaudt 2001 [DIRS 164037], Section 7.2.). Hanson (1998 [DIRS 101672], p. iii) describes the 
oxidation of spent nuclear fuel in dry air as a two-step process of the form UO2.UO2.4.U3O8. 
At temperatures up to 250C, UO2 fuel that is exposed to dry air will oxidize to UO2.4 at a rate 
that is controlled by the rate of diffusion of oxygen into the fuel. The transition from 
UO2.UO2.4 does not result in appreciable fuel pellet density changes. At higher temperatures, 
the fuel is oxidized to U3O8. Hanson (1998 [DIRS 101672], p. iii and 5.9) found that the 
UO2.4.U3O8 reaction rate, which is controlled by nucleation and growth of U3O8, is strongly 
dependent on fuel temperature and burnup. The transition from UO2.4.U3O8 results in a volume 
expansion of greater than 36%. No oxidation to U3O8 has been observed for temperatures below 
250C. The density or specific volume changes associated with fuel oxidation can cause 
disaggregation of CSNF into grain-sized powders when the connected porosity in the fuel allows 

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MDL-EBS-PA-000004 REV 02 6-16 August 2005 
oxidation to occur at the grain boundaries and cause intergranular fracturing. The density 
changes can also result in transgranular fracturing depending on the rate of oxidation involved. 
Results from the corrosion of spent UO2 fuel stored at ambient temperatures in dry air 
for 10 years, and subsequently stored in a humid-air environment for fifteen years, have shown 
oxidation processes along the grain boundaries that weaken the intergranular bonding and cause 
grain-boundary decohesion (Leenaers et al. 2003 [DIRS 168991]). Examination of the fuel 
showed progression of a grain-boundary corrosion front that extended from the periphery to 
about the midradius of the pellets. These results were interpreted to be due to the combined 
effects of two reactions. One reaction consumes oxygen by direct oxidation of the UO2 fuel to 
produce U4O9. The other produces hydrogen by oxidation of the fuel to U4O9 through reaction 
with H2O2 produced by radiolysis; the net reaction involves oxidation of the fuel by water and 
can be written as one mole of water reacting with four moles of UO2 to produce one mole of 
U4O9 and one mole of hydrogen. The larger oxygen depletion observed in the humid-air exposure 
compared to dry-air exposure indicated that the oxidation of the fuel had progressed more rapidly 
in humid air than in dry air. 
Particle size distributions for unirradiated and irradiated UO2 fuel following high-temperature 
oxidation (300C to 900C) have been reported by Liu et al. (1992 [DIRS 172864]) and 
Iwasaki et al. (1968 [DIRS 172518]). Figure 6-5c shows the particle size distribution results 
(percent smaller than a given size) reported by Liu et al. (1992 [DIRS 172864]) for unirradiated 
fuel. Results from both studies show that only a small fraction of oxidized material is less than 1 
micrometer in diameter. Also, the results show that the particle size increases as the oxidation 
temperature increases. This was attributed to an increase in the plasticity of U3O8 with 
increasing temperature, which allows the oxidation-induced stresses to be relaxed and thus 
minimize intra- and transgranular fracturing. These data suggest that few, if any, colloid-sized 
particles would be produced when CSNF is oxidized at the higher temperatures to which it 
would be exposed following an igneous intrusion event. 
NOTE: The shaded region represents a conservative range of potential colloids from oxidized fuel. 
Figure 6-5c. Data Replotted from Liu et al. (1992 [DIRS 172864]) for Unirradiated Fuel. 

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MDL-EBS-PA-000004 REV 02 6-17 August 2005 
Data on the particle-size distribution that may result from the oxidation of the restructured fuel in 
the rim region of high-burnup spent nuclear fuels are not available. The average grain size of the 
recrystallised fuel is 290 nm with a standard deviation of 140 nm (Spino et al. 1996 
[DIRS 174080], Figure 7). In contrast to the grain boundaries in medium-to-low burnup fuels, 
the fact that the fully restructured fuel in the rim region contains the exsolved fission gas in 
pressurized bubbles suggests that there is little if any connected porosity at the grain boundaries 
in this region (Spino et al. 1996 [DIRS 174080], Figure 9 and associated discussion). The 
morphological changes associated with oxidation of this material are, therefore, likely to be more 
similar to those that have been observed in the oxidation of unirradiated sintered UO2 pellets 
than to the morphological changes that have been observed in the oxidation of irradiated fuel 
fragments (Dehaudt 2001 [DIRS 164037], Sections 7.2.5.4). This indicates that the oxidized 
powder size distributions will likely range upwards from the fuel grain size distribution. 
The complex effects of humidity (particularly for relative humidities at and near 100%) on the 
corrosion of irradiated and unirradiated UO2 fuels are summarized in a review of the oxidation of 
UO2 at temperatures below 400C (McEachern and Taylor 1998 [DIRS 113270], Section 2.2). 
Humidity can influence air oxidation in this temperature range by forming hydrated-uranyl 
phases (e.g., by hydrous disproportionation of U3O8), by supporting oxidative dissolution in 
water films that form on the surface at higher relative humidities, by enhancing grain-boundary 
oxidation in spent nuclear fuel, and by supporting radiolytic processes at grain boundaries. 
When the finer particles that were produced by oxidation at 300C were exposed to 98% relative 
humidity, the particle size was observed to increase (from a median size of 7.5 m to about 100 
m) (Liu et al. 1992 [DIRS 172864]). Although Liu et al. (1992 [DIRS 172864]) attributed this 
to agglomeration or adhesion of the fine particles to the sieves used for size determination; it is 
also plausible that the size increase was associated with oxidative dissolution of the fine-grained 
oxide particles and reprecipitation as uranyl oxide hydrate phases. Results from work by Taylor 
et al. (1995 [DIRS 125815]) discussed below indicate that upon exposure to humid air with a 
relative humidity near 100%, any submicrometer particles that are formed as a result of fuel 
oxidation are likely to grow due to oxidative dissolution and precipitation of coarser uranyloxyhydroxide 
phases in a surface layer of water that condenses onto the particles under these 
conditions. Hot cell observations by on unsaturated spent nuclear fuel drip tests suggest that the 
fuel underwent considerable disintegration after 3.7 yrs for ATM-103 (Finn et al. 1998 [DIRS 
100392], pp. 25 to 30), and after 4.1 yrs for ATM-106 (Finn et al. 1998 [DIRS 100392], pp. 40 to 
41). Loosely aggregated clumps fell apart on handling. Uncorroded spent nuclear fuel is a hard 
ceramic material. However, it is unknown whether this would result in particles in the colloidal 
size range. 
Taylor et al. (1995 [DIRS 125815]) investigated the composition and microstructure of corrosion 
products formed on unirradiated UO2 fuel exposed at 225C to humid air with a range of relative 
humidities. In dry air (relative humidity < approximately 30%), they observed oxidation to U3O7 
and U3O8 via the solid-state mechanisms discussed above. At moderate relative humidities (30% 
to 70%), they observed surface and grain-boundary corrosion and fine-grained U3O8 and 
dehydrated schoepite crystals sparsely scattered over the surface. These observations were 
attributed to oxidative dissolution and precipitation of U(VI) phases within a thin (few molecules 
thick) and perhaps patchy layer of water on the fuel surface. At higher relative humidities 
(extending to near 100% relative humidity), a continuous and coarse-grained layer of U3O8 and 

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dehydrated schoepite were found overlying a U3O7 layer. These results were attributed to 
dissolution and precipitation in a thicker film of water, which closely resembles a bulk liquidphase 
film. When the test atmosphere was moisture-saturated (i.e., 100% relative humidity), 
large (tens to hundreds of micrometers in length) dehydrated schoepite crystals were observed 
consistent with precipitation and growth from bulk solution. 
Based on the results discussed above the likely effects of CSNF oxidation and subsequent 
exposure to humid air on colloid formation in the nominal, seismic and igneous scenarios can be 
summarized as follows: 
 Oxidation and breakup of low- to medium-burnup fuels is expected to produce 
powders with particle sizes greater than one micrometer. Higher temperature 
(> 500C) is expected to produce particle sizes greater than the fuels grain size; 
lower temperature exposure is likely to produce powders with particle sizes close to 
the grain size of the unoxidized fuel. 
 For the fine-grained restructured fuel in the rim region of high-burnup CSNF, 
extrapolation of the experimental observations for unirradiated fuel and low- to 
medium-burnup fuel suggests that the oxidized fuel particle sizes are likely to be 
larger than the submicrometer grain size of the restructured fuel. However, this 
extrapolation is uncertain because directly applicable experimental data are not 
available for high-burnup CSNF. 
 If submicrometer particles are generated as a result of fuel oxidation, the subsequent 
exposure of these particles to humid air approaching 100% relative humidity will 
result in uranyl oxide phase formation. The particle sizes produced are likely to be 
similar to those produced in the oxidative dissolution and precipitation of unoxidized 
CSNF. Secondary phase particles from the spent nuclear fuel unsaturated drip tests 
have been reported to be (< 10 m to 200 m) (Finn et al. 1998 [DIRS 100392], 
p. 25). This is above the accepted colloidal size range. 
 After exposure of the oxidized fuel to humid air, few, if any, submicrometer particles 
are expected to be available for suspension as colloids. 
Based on the above considerations, exposure of CSNF to the dry air oxidizing conditions, 
followed by exposure to humid air, is unlikely to lead to formation of submicrometer particles 
that would be available for suspension as colloids when water contacts the oxidized fuel. 
Association of Plutonium and Americium with Spent Nuclear Fuel-Derived Colloids 
The assumption made in Section 5.3 that irreversible (embedded) americium and plutonium 
colloids occur during DHWL glass corrosion was based on a limited series of microscopic 
images of corroded borosilicate waste glasses that indicated the formation of discrete actinidebearing 
phases. The nature of these phases depends on the glass composition but typically was 
due to precipitation of the actinide with phosphate or titanate ligands. Similar precipitates of rare 
earths or actinides (uranium and thorium) with phosphates (e.g., rhabdophane), titanates (e.g., 
brannerite), or arsenates (e.g., agardite) are commonly observed in nature; however, no anionic 

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complexants are available from within spent nuclear fuel that would result in the generation of 
similar highly insoluble phases for immobilizing actinides. Although the plutonium-bearing 
oxide phases on the surface of corroded spent nuclear fuel have been observed (Buck et al. 2004 
[DIRS 172668]), there is no information on whether these phases can be released from a 
weathered fuel surface. One criterion for the occurrence of irreversibly attached (or embedded) 
actinide phases is that a suitable carrier colloid exists which has the requisite morphology to 
readily encapsulate the actinide phases. In the case of DHLW glass, the highly flexible and 
amorphous smectite clays may be ideally suited to completely covering the actinide colloids. 
However, the main potential carrier colloid for spent nuclear fuel is a U(VI) oxide phase. These 
are normally euhedral crystals, typically rigidly accicular or planar in nature and may be unable 
to encapsulate a smaller colloid in the same manner. These phases are much more likely to sorb 
plutonium or americium at specific surface and edge sites. Any discrete plutonium- or 
americium-bearing particle having formed during reaction with the aqueous phase and attached 
to a U(VI) phase would be exposed to the aqueous phase and would dissolve once the solution 
conditions permitted. Bitea et al. (2003 [DIRS 173041]) have shown that thorium oxide colloids 
dissolved when contacted with water with a thorium concentration below the solubility limit of 
thorium hydroxide at constant pH. 
Field Evidence of Colloids from Natural Analogues 
Natural analogues for colloid-facilitated radionuclide behavior have been reviewed by Triay et 
al. (1995 [DIRS 100789]) and Traexler et al. (2004 [DIRS 172072]). However, few of these 
reviewed natural analogue sites are relevant to the Yucca Mountain unsaturated oxidizing 
environment. At the Koongarra deposit in Northern Australia where the groundwater is 
dominated by Mg and bicarbonate (the pH is between 6 and 7.5), colloids include clay minerals, 
fine-grained quartz, and iron-oxide particle coatings. Uranium was principally dissolved, 
whereas, thorium and actinium were associated with particles. Observations of nanocrystals of 
U(VI) phosphates by Murakami et al. 1997 [DIRS 113272] in Koongarra groundwaters that were 
undersaturated with respect to saleite {Mg[(UO2)(PO4)]2(H2O)10} and metatorbernite 
{Cu[(UO2)(PO4)]2(H2O)8} may support a colloidal transport process. The most appropriate 
natural analogue for colloid behavior at Yucca Mountain is the Nopal I uranium deposit at Pena 
Blanca District, Mexico. Pearcy et al. (1995 [DIRS 110223]) show that uranium has migrated 
along micro- and meso-fractures in the Tuff rock to distances less than 1 mm. The uranium 
mineralogy at Nopal I consists of almost exclusively the uranyl silicate (uranophane) and uranyl 
oxyhydroxides produced from uraninite alteration. Stoichiometric uraninite (UO2) is unknown in 
nature and is always partially oxidized in addition to containing radiogenic lead and commonly 
thorium, calcium, and lanthanides. 
Conclusion 
Spent nuclear fuel-derived colloids are not included in the base-case model. Alternative 
conceptual models have been developed to handle spent nuclear fuel-derived colloids. The 
alternative conceptual models include those for reversible and irreversible association of 
plutonium and americium with spent nuclear fuel-derived colloids (see Sections 6.4.1 and 6.4.2 
for the discussion and screening arguments). 

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6.3.1.3 Colloids from the Corrosion of Waste Package and Metallic Invert Materials 
The occurrence of iron oxyhydroxide colloids from the corrosion of waste package and metallic 
invert materials is included in the colloid model. Conceptually (and for purposes of estimating 
specific surface area for irreversible sorption calculations; Section 6.3.3.2), iron oxyhydroxide 
colloids are likely to be comprised of three species, under the anticipated repository conditions: 
goethite, hematite, and ferrihydrite, or hydrous ferric oxide) (Langmuir 1997 [DIRS 100051], 
pp. 436 to 437). Goethite forms from hydrous ferric oxide by dissolution-reprecipitation, while 
hematite will form by direct solidsolid transformation (aging of hydrous ferric oxide) 
(Langmuir 1997 [DIRS 100051], p. 437). The iron oxyhydroxides will occur in three forms: 
immobile scale; large particles that will settle out; and colloid-sized particles. 
The iron oxyhydroxide phases may be expected to provide abundant sorption sites for dissolved 
(aqueous) radionuclides in amounts determined by the appropriate Kd value (Section 6.3.3.1). 
Stable iron oxyhydroxide colloid suspensions may serve to increase the mobility of sorbed 
radionuclides (Sections 6.3.2 and 6.3.3). Their stability is controlled by fluid conditions, 
including ionic strength and pH. 
Corrosion products released from carbon steel miniature waste packages were characterized 
(DTN: MO0302UCC034JC.003 [DIRS 162871]) and the quantity of corrosion products 
transported out of the waste packages was measured by micro-filtration 
(DTN: MO0312UCC034JC.008 [DIRS 166367]). Two configurations were used to introduce 
water into the carbon steel miniature waste packages and induce corrosion: (1) a bath tub 
configuration in which water was introduced from the top of the miniature waste package and 
exits from an opening on the side of the miniature waste package after it has accumulated along 
the bottom, and (2) a flow-through configuration in which water is introduced from the top of 
the miniature waste package and exits directly from the bottom (Zarrabi et al. 2003 
[DIRS 171238]). Water, chemically similar to J-13 well water (near Yucca Mountain), was used 
in the experiments and the rate of water introduction was similar to projected rates of water 
movement within the drifts at Yucca Mountain. To establish the total amount of corrosion 
material transported out of the carbon steel miniature waste packages, the effluent was passed 
through a microfiltration system capable of separating particles of 0.1 m and upward, which is 
below the range for colloids; hence, this system will under report colloid concentrations. 
Identification of the oxidation corrosion products is ongoing, although to date the data suggest a 
preponderance of cryptocrystalline corrosion products, including magnetite (Fe3O4), 
lepidocrocite {FeO(OH)}, and goethite {FeO(OH)} (DTN: MO0302UCC034JC.003 
[DIRS 162871]). 
The colloidal size fraction of these materials represents viable analogues for the nature and 
quantity of corrosion-generated iron oxide colloids that could potentially be produced and 
transported from the engineered barrier system early during the corrosion of the waste packages 
after breaching. The cumulative results of the miniature waste package corrosion tests to date 
have yielded average concentrations of colloidal-size material in the range of 20 mg/L within the 
initial four weeks of the experiments (DTN: MO0312UCC034JC.008 [DIRS 166367]). All the 
carbon steel miniature waste packages clogged after several weeks of water influx, probably due 
to the formation and accumulation of corrosion products (e.g., iron oxyhydroxide) of larger 
molar volume than the original materials prior to corrosion. 

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Uncertainty in the data generated by the miniature waste package experiments and the 
consideration of iron oxyhydroxide colloid concentrations reported in the literature for 
groundwaters in iron-rich geologic environments is discussed in Section 6.6.2. 
6.3.1.4 Natural Colloids Originating from Unsaturated Zone Seepage Water/ 
Groundwater 
Natural colloids occurring in seepage water or groundwater in the repository vicinity are 
modeled as smectite clays. While silica and iron oxyhydroxides will be present in seepage water 
or groundwater, these are excluded for two reasons. First, silica has a much lower sorptive 
capacity than smectite clay and can be ignored. Second, the abundance of iron oxyhydroxides in 
the waste package vastly exceeds concentrations that may occur in the seepage waters. 
Seepage water and colloids enter the drift and breached waste package. Three types of seepage 
water/groundwater colloids may be present at Yucca Mountain, including hydrophobic mineral 
fragments, humic substances, and microbes. 
Clays, silica, hematite, and goethite colloids occur in groundwater in the vicinity of Yucca 
Mountain (Kingston and Whitbeck 1991 [DIRS 113930]), and the colloid model calls for these 
colloids to enter a failed waste package and be available to interact with released radionuclides. 
The presence and potential influence of natural seepage water colloids on the formation of 
radionuclide-bearing colloids in the engineered barrier system are considered by modeling 
seepage water colloids as smectite; these are referred to as groundwater colloids throughout 
this document. 
The complexation capacity of humic substances in the Yucca Mountain environment is low and 
they are not included among the groundwater colloids considered in this abstraction (Minai et al. 
(1992 [DIRS 100801]). In the presence of calcium in J-13 well water, the effective 
complexation capacity of humic substances is 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 (Minai et al. 1992 [DIRS 100801], p. 199). 
Because of the relatively large sizes of microbes (up to tens of ms), they are susceptible to 
filtration in geologic media and; hence, microbe-radionuclide complexes, are not modeled in this 
abstraction. Discussion on the probable influences of microbes on the colloids that are 
considered in this analysis (inorganic waste form colloids, steel corrosion product colloids, and 
natural seepage water/groundwater colloids) is presented in Section 6.3.4. 
6.3.2 Colloid Suspension Stability and Concentration 
The clay colloid suspensions observed in the DHLW glass tests are considered to have similar 
stability characteristics as natural clay colloid suspensions because the waste form-derived clays 
exhibit similar crystallographic and chemical properties as natural clays (Buck and Bates 1999 
[DIRS 109494]; Buck et al. 1994 [DIRS 122116]; Feng et al. 1994 [DIRS 101611]). The iron 
oxyhydroxide colloids will form from internal corrosion of the waste package are similarly 
considered as analogous to natural iron oxyhydroxide colloids. 

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TemperatureTemperature effects have also been shown to influence the stability of colloid 
suspensions (Langmuir 1997 [DIRS 100051], p. 439). Increased temperature causes a colloid 
suspension to become less stable, owing to an increase in particleparticle collisions that lead to 
flocculation. For the purposes of this abstraction, temperature effects are ignored, and it is 
assumed that waste form colloid suspensions produced at 90C are stable at 25C (Section 5, 
Assumption 5.2). The net effect of this assumption on TSPA-LA is that colloid suspensions 
formed at the higher temperatures remain stable as the repository cools. 
Ionic strengthColloid suspensions become unstable as ionic strength in a solution increases 
due to compression of the diffuse counter-ion atmosphere and decreases in surface potential (i.e., 
decreases in repulsive forces). Further, the degree of compression of the electric double layer is 
dependent on the concentration and valence of ions of opposite sign from that of the surface 
charge (van Olphen 1977 [DIRS 114428]). The effect of valence is captured by the 
Schulze-Hardy rule, which incorporates the principle that states for a given counter-ion 
concentration, the higher the valence of the counter-ions, the more unstable the colloid 
suspension and the greater the tendency to flocculate (van Olphen 1977 [DIRS 114428]). 
Colloidal actinide species have been observed in concentrated MgCl2 brines, possibly owing to 
higher rates of colloid formation than those for removal. However, these colloidal fractions 
decreased with time (Geckeis et al. 1998 [DIRS 114457]). 
Critical Coagulation ConcentrationCritical coagulation concentration is defined as the 
minimum concentration of an electrolyte that results in the flocculation (coagulation) of a 
colloidal suspension. At this point the colloid suspension is assumed to settle out of solution by 
gravitational forces. In his treatise on clay colloidal dispersions, van Olphen (1977 
[DIRS 114428], p. 24) summarized empirically determined critical coagulation-concentration 
values (Table 6-3a). 
Table 6-3a. Critical Coagulation Concentration Values 
Counterion Valence Critical Coagulation Concentration (M) 
1 2.5 x10-2 to 1.5 x10-1 
2 5 x 10-4 to 2 x 10-3 
3 1 x 10-5 to 1 x10-4 
Source: van Olphen 1977 [DIRS 114428]. 
At pH values below the pHZPC (where zpc is the zero point of charge) for a mineral, the 
negatively charged counterions are important; above the pHZPC, positively charged counterions 
are important (van Olphen 1977 [DIRS 114428]). For comparison, concentrations of counterions 
in J-13 well water have been reported and are listed in Table 6-3b. 
Table 6-3b. Concentrations of Counterions in Input Water Compositions 
Ca-porewater Na-porewater J-13 Well Water 
Ion Concentration (moles/L) 
Na+ 1.3  10-3 4.1  10-3 2.0  10-3 
K+ 2.6  10-4 2.1  10-4 1.3  10-4 
Ca2+ 3.2  10-3 2.8  10-3 3.2  10-4 

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Table 6-3b. Concentrations of Counterions in Input Water Compositions (Continued) 
Ca-porewater Na-porewater J-13 Well Water 
Ion Concentration (moles/L) 
Mg2+ 6.2  10-4 1.1  10-4 8.3  10-5 
Cl- 7.2  10-4 8.3  10-4 2.0  10-4 
SO4
2- 1.2  10-3 1.1  10-3 1.9  10-4 
pH 7.6 7.4 7.41 
Source: BSC 2005 [DIRS 174583 ] Table 4-2, p. 18. (units converted from mg/L to moles/L) 
The concentrations of individual ions in J-13 well waters do not exceed empirical critical 
coagulation concentration (CCC) values. However, for the Ca- and Na- porewaters, the CCC 
value is exceeded. The conclusion is that colloid dispersions will probably be stable in J-13 well 
water. However, the predicted Ca- and Na- porewater compositions as described in BSC 2005 
[DIRS 174583] suggest colloids may be destablized. The ionic strengths of the Ca-pore water, 
Na- pore water, and J-13 well water, are < 0.005 M, <0.02 M, and <0.02 M, respectively. In the 
model, the ion effect (termed Schulze-Hardy) is conservatively ignored and the model has 
adopted an ionic strength based model (see section 6.3.2.1). 
6.3.2.1 Stability of Smectite Colloid Suspensions as a Function of pH and Ionic Strength 
There is no effective zero point of charge for smectite (Sondi et al. 1996 [DIRS 170928]); 
however, it may be reasonable to set a value between pH 2 and 3 (Langmuir 1997 
[DIRS 100051]). Hence, it is reasonable to expect that smectite colloid suspensions will remain 
in a stable pH range much of the time under anticipated repository conditions (unless ionic 
strength exceeds a certain threshold). Tombcz et al. (1990 [DIRS 112690]) investigated the 
stability of montmorillonite clay colloid suspensions as a function of pH and ionic strength in 
NaCl solution. It was found that suspensions became unstable and flocculated at pH 2, 4, and 8 
in 0.01 M, 0.025 M, and 0.04 M NaCl solutions, respectively (Figure 6-6a). Figure 6-6b 
illustrates the abstraction and extension of these data to pH 10 and ionic strength of 0.05 M for 
use in the TSPA-LA calculations. Schematic representation was abstracted from data presented 
in Figure 6-6a. The established relationship between ionic strength and smectite colloid 
suspension stability (STcoll, wf, pH) as a function of pH for use in the TSPA-LA model is shown in 
Figure 6-6b. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-24 August 2005 
Source: Adapted from Tombcz et al. 1990 [DIRS 112690]. 
NOTE: Data points represent combinations of ionic strength and pH at which montmorillonite colloid suspension 
stability is decreased. There is no effective zero point of charge (zpc) for montmorillonite. 
Figure 6-6a. Experimental Determination of Montmorillonite (Smectite) Colloid Suspension Stability as a 
Function of pH and Ionic Strength 
Figure 6-6b. Schematic Representation of Smectite Colloid Suspension Stability as a Function of pH 
and Ionic Strength 
NOTE: The zpc for smectite (montmorillonite) clay colloid suspensions is at such a low pH that it does not 
influence flocculation in the model. 

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MDL-EBS-PA-000004 REV 02 6-25 August 2005 
The range of chemical conditions in the laboratory testing represents a subset of the conditions 
expected in the repository. The pH in product consistency tests depends on the alkali 
borosilicate glass reaction and is not controlled (Buck and Bates 1999 [DIRS 109494]). The pH 
range anticipated in the waste-form and drift environments for the majority of the time is 
expected to be between 4.5 and 8.1. Since the pH never drops below 7 during the reaction of 
water with alkali borosilicate glass (Luckscheiter and Nesovic 2002 [DIRS 170884]), the range 
of expected conditions is adequately represented. 
6.3.2.2 Stability and Concentration of Iron Oxide/Oxyhydroxide Colloid Suspensions as 
a Function of pH and Ionic Strength 
The point of zero charge of iron oxyhydroxide colloids ranges between pH values of 
approximately 5 and 8.5 (Langmuir 1997 [DIRS 100051], p. 351, Table 10.3), depending on 
colloid and water composition, and in this range they will tend to be unstable and tend to 
agglomerate. At higher or lower pH, however, iron oxyhydroxide colloids are stable, depending 
upon ionic strength. Liang and Morgan (1990 [DIRS 109524]) demonstrated that for a given 
ionic strength hematite colloidal suspension stability increases as pH both increases and 
decreases away from the zero point of charge (Figure 6-8). The higher the ionic strength, the 
wider the pH range about which the experimentally derived stability ratio is zero. The pH of 
zero-proton condition for hematite is indicated in Figure 6-7 (i.e., pH 8.5). Dashed lines are 
drawn through the specific experimental points as a guide. The solid line represents a Darjaguin, 
Landau, Verwey, and Overbeek (DLVO) model calculation for ionic strength 0.05 M. Note that 
W represents the stability of the dispersion, in terms of the rate at which colloidal particles in the 
dispersion agglomerate. A high value represents a stable dispersion and a value of zero indicates 
rapid agglomeration. At ionic strengths of about 0.05 M or more, the stability ratio is low 
between pH values of about 6 and 11; furthermore, in the presence of divalent cations, organics, 
or phosphate, colloid suspension stability is lowered. 
Source: Extracted from Langmuir 1997 [DIRS 100051]. 
Figure 6-7. Schematic Diagram Adapted from Experimentally Derived Stability Ratio, W, of a Hematite 
Suspension Plotted as a Function of pH for Differing Ionic Strengths 

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MDL-EBS-PA-000004 REV 02 6-26 August 2005 
As described in Section 6.3.1.3, the iron will corrode and form iron oxides and oxyhydroxides. 
The iron oxides and oxyhydroxides will occur as immobile scale with the large particles 
settling out. For modeling purposes, two parameters are established, the expected maximum and 
minimum mass concentrations of iron oxide colloids, Mcoll,FeOx,sampled and Mcoll,FeOx,min, 
respectively. 
Research (Section 6.3.1.3) on carbon steel miniature waste packages has shown that corrosion 
proceeds rapidly, and colloid-sized particles have been observed to leave the miniature waste 
package in concentrations of approximately 15 mg/L to 20 mg/L. The release is continuous until 
it decreases and the aperture clogs with corrosion products. It is possible that in a repository 
setting clogging material may become dislodged by an event or changed condition, and release 
could resume, the overall process possibly occurring more than once. 
Accordingly, the mass concentration range was chosen to include the 15 mg/L to 20 mg/L range, 
and to allow for releases dwindling to a low quantity (0.05 mg/L), as well as for a pulse-like 
release (up to 50 mg/L). This maximum value is well within the range of colloid concentrations 
observed experimentally and in natural groundwater (i.e., Lu et al. (2000 [DIRS 166315]), 
performed experiments with colloid concentrations of up to 200 mg/L; and Honeyman and 
Ranville (2002 [DIRS 161657]), reported colloid concentrations in groundwater occasionally in 
excess of 100 mg/L). 
The distribution was chosen to emphasize the upper end of the range (i.e., the 5 to 50 mg/L 
portion) to reflect the steady state conditions and pulses, but also allow for occasional clogging 
of the aperture in the waste package. The emphasis of the upper end of the range is conservative 
as it tends toward predicting higher colloid concentrations to which dissolved radionuclides may 
sorb. This distribution is valid for ionic strengths less than 0.05 M. The colloid mass 
concentration at ionic strengths greater than 0.05 M is set to 1  10-3 mg/L. This concentration 
value is between one and two orders of magnitude lower than the lower bound of the distribution 
used for ionic strengths less than 0.05 M. This value is reasonable based on experimental 
observations that have shown a low stability ratio. This low-stability ratio essentially implies 
complete agglomeration of colloids at ionic strengths exceeding 0.05 M (Liang and Morgan 1990 
[DIRS 109524]). Therefore, the use of a nonzero value for colloid concentration above ionic 
strength 0.05 M is conservative. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-27 August 2005 
Source: Abstracted from Liang and Morgan 1990, [DIRS 109524], Figure 1. 
NOTE: Schematic representation (used in abstraction) of iron oxyhydroxide colloid suspension stability as a 
function of ionic strength and pH. At and near the zero point of charge, colloids are unstable, even at low 
ionic strengths. At higher ionic strengths the pH range at which colloid suspensions are unstable is 
greater. Above ionic strength 0.05 colloids are assumed to be unstable for all pH values. In the equations 
on the diagram, I refers to the ionic strength. 
Figure 6-8. Schematic Representation of Iron Oxyhydroxide Colloid Suspension Stability as a Function 
of pH and Ionic Strength 
6.3.2.3 Stability and Concentration of Waste Form Colloid Suspensions 
Colloids were formed in the course of static DHLW glass tests on several of the glasses for 
different SA/V (glass solid surface area to leachate volume ratio). It was observed that as the 
ionic strength increased, colloid concentration generally decreased, and ultimately a threshold 
value was reached above which the colloids were not observed or were observed in low 
quantities (Figure 6-4). As this threshold was approached, it was also observed for one glass that 
the colloid size increased due to aggregation of the colloids. The threshold at which flocculation 
occurred was approximately I = 0.05M. 
The stability of smectite colloid suspensions (including smectite colloids formed from DHLW 
glass degradation) is determined primarily by ionic strength, but also to an extent by pH, as 
described in Section 6.3.2.1 (see Figure 6-9a). A relationship between the plutonium colloidal 
concentration ([Pu]colloid) and ionic strength (I) in the pH range from 2 to 10 was determined by 
the following limits as shown in Equation 2: 
[Pu]colloid (mol/L) = -2.50  10-6  I + 1.25  10-7, for 0.01 <I < 0.05 (Eq. 2) 
This relationship is valid for ionic strengths greater than 0.01 M and less than 0.05 M. The 
concentration at ionic strengths greater than 0.05 M is set to 1  10-11 mol/L, representing a small 
nonzero number for modeling purposes (Figure 6-9b). This concentration value is reasonable 
based on experimental observations that have shown complete agglomeration of colloids at ionic 
strengths exceeding 0.05 M (CRWMS M&O 2001 [DIRS 154071]; Buck and Bates 1999 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-28 August 2005 
[DIRS 109494]). At ionic strengths less than 0.01 M, the plutonium concentration is modeled as 
1  10-7 M, also based on experimental observations (Figure 6-9b). 
NOTE: Glass reaction can result in an increase in ionic strength and pH. Ionic strength calculations reported in 
CRWMS M&O 2001 [DIRS 154071] did not include anions. 
Figure 6-9a. Plutonium Concentration as a Function of Ionic Strength for Corrosion Tests on Glass 
Samples, SRL202A and SRL131A. 
Americium concentration is determined by multiplying the calculated plutonium concentration 
by the ratio of americium to plutonium in the radionuclide inventory of the DHLW glass at each 
TSPA-LA calculation time step. Examples of probable DHLW glass compositions are provided 
by Fowler (2003 [DIRS 164361]). 

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MDL-EBS-PA-000004 REV 02 6-29 August 2005 
NOTE: Schematic relationship (used in abstraction) between plutonium concentration in DHLW glass and ionic 
strength. The function represents the bound of the DHLW glass corrosion experimental data (Figure 6-4); 
the maximum value represented, 1  10-7 mol/L, bounds the maximum concentration of colloids observed 
in the DHLW glass experiments. The low value, 1  10-11 mol/L, represents the steep decline in colloid 
concentration above ionic strength 0.05 M. 
Figure 6-9b. Schematic Relationship between Plutonium Concentration in DHLW Glass Colloids and 
Ionic Strength 
6.3.2.4 Stability and Concentration of Seepage Water Colloid Suspensions 
Colloids are thought to exist naturally in groundwater in subsurface environments with the 
composition and concentration of colloids being site specific and determined by the geologic 
nature of the subsurface. Transport of radionuclides by groundwater colloids is a potentially 
important transport process and is thought to be generally more prevalent than transport by true 
colloids (McCarthy and Zachara 1989 [DIRS 100778]). 
McCarthy and Degueldre (1993 [DIRS 108215]) and Degueldre et al. (2000 [DIRS 153651]) 
have characterized the colloids from groundwaters sampled at various locations around the world 
from crystalline and sedimentary rocks in saturated and unsaturated hydrologic regimes. They 
showed that the colloid concentration and suspension stability is dependent on several factors; 
however, colloid suspensions tend to be stable if the concentration of alkalis (Na and K) is below 
approximately 10-2 M and if alkali-earth elements (Ca and Mg) are below approximately 10-4 M 
(Figure 6-10). In the case of Ca- and Na- porewaters, described in In-Package Chemistry 
Abstraction (BSC 2005 [DIRS 174583]), the Ca2+ concentration is >10-4 M which would lead to 
colloid suspension instability in the waste package environment. While the DLVO calculations 
provide useful inferences on mechanisms of colloid destabilization, these empirical observations 
may be more useful in developing approaches to performance assessment calculations. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-30 August 2005 
Source: Adapted from Degueldre et al. 2000 [DIRS 153651]. 
NOTE: Here, concentrations of colloids are compared on the basis of alkali and alkaline-earth element 
concentration for colloid size greater than 100 nm. 
Figure 6-10. Colloid Concentrations versus Alkali and Alkaline-Earth Concentration for Groundwaters 
from around the World 
Data from the Yucca Mountain area and the Nevada Test Site indicated a highly variable colloid 
concentration over a relatively narrow range of ionic strength, in agreement with the findings of 
Kingston and Whitbeck (1991 [DIRS 113930]) and Degueldre et al. (2000 [DIRS 153651]). 
Figure 6-11 plots the groundwater colloid concentrations in water samples collected in the 
vicinity of Yucca Mountain. Data sources included in the figure are: 
Groundwater Colloid PopulationsDTNs: LA0002SK831352.001 [DIRS 149232]; 
LA0002SK831352.002 [DIRS 149194]; LA0002SK831352.004 [DIRS 161579]; 
LA9910SK831341.005 [DIRS 144991]; LA0211SK831352.001 [DIRS 161580]; 
LA0211SK831352.002 [DIRS 161581]; LA0211SK831352.003 [DIRS 161582]; 
LA0211SK831352.004 [DIRS 161458]; LA0002SK831352.003 [DIRS 161771]) and 
colloid concentrations in groundwaters sampled around the world by 
Degueldre et al. (2000 [DIRS 153651]) and from DTN: SNT05080598002.001 
[DIRS 162744] data. 
Ionic Strength DataDTNs: SNT05080598002.001 [DIRS 162744], 
LA0304PR831232.001 [DIRS 163196], and LA0304PR831232.002 [DIRS 163197]. 
The groundwater colloid samples analyzed represent colloid particle size distributions from 
50 nm to 200 nm in diameter while the data from groundwaters around the world 
(DTN: SNT05080598002.001 [DIRS 162744]) represent colloid particle size distributions from 
100 nm to 1,000 nm (1.0 m diameter). Evaluation of the colloid populations in the various size 
fraction classes for each groundwater sample did not reveal a systematic increase in the number 

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MDL-EBS-PA-000004 REV 02 6-31 August 2005 
of particles with decreasing particle-size class (Table I-2). Therefore, inclusion of particle-size 
classes smaller than 50 nm should not result in a substantially greater mass concentration in the 
water sample, especially since the smaller particles would have exponentially lower mass 
therefore, omission of the less than 50-nm particle-sized fraction in the mass concentration 
calculations is reasonable. 
Sources: Yucca Mountain site colloid data: DTNs: LA0002SK831352.001 [DIRS 149232]; LA0002SK831352.002 
[DIRS 149194]; LA9910SK831341.005 [DIRS 144991]; LA0211SK831352.001 [DIRS 161580]; 
LA0211SK831352.002 [DIRS 161581]; LA0211SK831352.003 [DIRS 161582]; LA0211SK831352.004 
[DIRS 161458]; LA0002SK831352.003 [DIRS 161771]; and LA0002SK831352.004 [DIRS 161579]. 
Groundwater colloid population data: DTN: SNT05080598002.001 [DIRS 162744]. 
NOTE: Ordinate values are presented in particles per milliliter. 
Figure 6-11. Groundwater Colloid Concentration Data Collected in the Vicinity of Yucca Mountain 
Compared with Data Collected from Groundwaters around the World 
Colloid particle number data (particles (pt)/mL) for groundwater samples from the Yucca 
Mountain area and the Idaho National Laboratory were converted to mass concentrations (mg/L) 
(Attachment I). The colloid data from DTN: SNT05080598002.001 [DIRS 162744] were not 
used to calculate the mass colloid concentrations (mg/L) but were only used to establish a 
general relationship between ionic strength and colloid populations in groundwaters around the 
world and, specifically, in the Yucca Mountain area (Figure 6-11). 
The groundwater colloid data from the Yucca Mountain site and the Idaho National Laboratory 
were pooled, and a discrete cumulative distribution function was established to evaluate the 
uncertainty in colloid concentration distribution. The colloid concentrations data included 
DTNS: LA0002SK831352.001 [DIRS 149232]; LA0002SK831352.002 [DIRS 149194]; 
LA9910SK831341.005 [DIRS 144991]; LA0211SK831352.001 [DIRS 161580]; 
LA0211SK831352.002 [DIRS 161581]; LA0211SK831352.003 [DIRS 161582]; 
LA0211SK831352.004 [DIRS 161458]; LA0002SK831352.003 [DIRS 161771]; and 
LA0002SK831352.004 [DIRS 161579]. Inclusion of the Idaho National Laboratory 
groundwater colloid data (DTN: LA0002SK831352.003 [DIRS 161771]) was deemed 
appropriate for inclusion in the data analysis as corroborative data even though these 
groundwater colloid data are not from the vicinity of Yucca Mountain (Attachment I). Water 
samples that had been filtered were not considered in the development of the uncertainty 
distribution. Since the colloid particles were measured by dynamic light scattering techniques, 

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MDL-EBS-PA-000004 REV 02 6-32 August 2005 
any potential error in measurement of colloid populations for the various particle size classes due 
to filter ripening (i.e., clogging of filter pores through time during the filtration process) was 
avoided by using only nonfiltered sample data. 
The data shown in Figure 6-11 reflect observed variability in groundwater colloid 
concentrations. The goal of the uncertainty distribution for groundwater colloid concentrations 
is to numerically capture our knowledge about uncertainty on a large scale. Based on the 
groundwater colloid concentration collected at the two sites (Yucca Mountain and the Idaho 
National Laboratory) and the large uncertainty associated with the collection of groundwater 
colloids, a reasonable representation of the uncertainty is shown in the discrete cumulative 
distribution function shown in Figure 6-11 (Table 6-4 shows the probabilities associated with the 
discrete cumulative distribution function). Major factors that contribute to uncertainty in the 
collection of groundwater colloids include: (1) collection techniques, (2) differences in pumping 
rates at each well, and (3) unknown factors including the types of additives introduced in the 
wells during the drilling process itself. 
The waste form and engineered barrier system components of the TSPA-LA analyses implement 
the uncertainty shown in Figure 6-11 only for ionic strength less than 0.05 M. For ionic 
strengths greater than or equal to 0.05 M, colloid suspensions are assumed to be unstable, based 
upon experimental work that has shown the flocculation of colloids at high ionic strengths (Liang 
and Morgan 1990 [DIRS 109524]; Tombcz et al. 1990 [DIRS 112690]), and a single low value 
(1  10-6 mg/L) is used for seepage colloids. 
Figure 6-12. Cumulative Distribution Function Showing the Probability of Occurrence of Colloid 
Concentration Levels in Groundwater Samples in the Yucca Mountain Area and Idaho 
National Laboratory 

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MDL-EBS-PA-000004 REV 02 6-33 August 2005 
Table 6-4. Uncertainty Distribution for Groundwater Colloid Concentrations for the TSPA-LA Analyses 
Based on Ionic Strength (I) of the Water Sample 
Ionic Strength 
(I) 
Groundwater Colloid 
Concentration (mg/L) 
Probability of 
Occurrence 
0.001 to 0.1 50 
0.1 to 1.0 25 
1.0 to 10 15 
10 to 50 8 
I < 0.05 M 
50 to 200 2 
I =0.05 M 1  10-6 100 
Review of the colloid data in Attachment I reveals that the largest proportion of groundwater 
colloid samples had colloid concentration populations between 0.001 and 0.1 mg/L and thus the 
values between these concentrations are modeled as having a sampling probability of 50% if the 
ionic strength of the solution is less than 0.05 M. Colloid concentration values of 0.1 mg/L to 
10 mg/L indicate a reasonable probability of occurrence of 40%. To account for colloid 
concentrations between 10 mg/L and 50 mg/L, a probability of 8% was included. Some of the 
corroborative (non-Q) data included values as high as 171 mg/L to 200 mg/L (Table I-1, 
Attachment I), but the uncertainty distribution for the pooled data indicates that these values 
have a low probability of occurring. 
6.3.3 Radionuclide Sorption to Colloids 
A description of radionuclide sorption to colloids is presented in this section. Nonradioactive 
colloids are only important to repository performance insofar as radionuclides sorb to them. 
Sorption as used here is a general term that is governed by: (1) electrostatic forces, (2) ion 
exchange, (3) surface reaction, and (4) coprecipitation. Sorption may occur by a number of 
mechanisms and may be considered reversible or irreversible in the time frame of transport. The 
rate of sorption, desorption, and transport determine the effective reversibility of the sorption. 
When sorption is fast and reversible, equilibrium conditions apply and the linear isotherm model 
is applicable for describing radionuclide uptake. When sorption is irreversible or desorption is 
slow relative to transport times, then local equilibrium models may fail, resulting in over- or 
under-predictions of transport. 
6.3.3.1 Reversible Sorption 
Perhaps the most common approach used in assessment of contaminant-rock interactions in the 
subsurface is the linear isotherm, or Kd value (distribution coefficient) approach, based on results 
of batch sorption experiments. The linear isotherm model relationship is defined as follows 
(Langmuir 1997 [DIRS 100051]): 
S = Kd C (Eq. 3) 
where 
S = mass of a solute adsorbed on a unit mass of solid (g/g) 
Kd = distribution coefficient (ml/g) 
C = concentration of the adsorbing solute in solution (g/ml). 

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The amount of solute adsorbed on a solid can also be defined on the basis of area as follows: 
S = Ka C (Eq. 4) 
where 
S = mass of a solute adsorbed on a unit area of solid 
Ka = distribution coefficient (area based) 
C = concentration of the adsorbing solute in solution. 
Three critical conditions must be met for this relationship to be applicable. First, the water 
contaminantrock system must be in thermodynamic equilibrium (i.e., sorption must be 
completely reversible). Second, contaminant uptake must scale linearly with contaminant 
concentration. Third, the presence of other solutes in the system cannot affect the sorption. 
Because of the nature of the formation of waste form colloids and the interactions of some 
radionuclides, particularly some actinides, with mineral surfaces, certain conditions of the linear 
isotherm model are not met; therefore, other approaches must be taken, as described below. 
A large quantity of data exists in published literature for sorption, and, in the past several 
decades, increasing attention has been given to understanding the mechanisms of sorption. In 
this section, the development of Kd values for uptake of cesium, plutonium, and americium on 
colloids is described, along with caveats. Uranium and neptunium were excluded from the 
analysis. Values for americium were assumed to be valid for thorium and protactinium. These 
radionuclide elements were chosen for analysis because, based on a screening exercise, they are 
considered by TSPA to be important to dose; high uncertainty surrounds their potential behavior 
as pseudocolloids; and daughter products of uranium may remain associated with uraniumbearing 
colloids. 
The radionuclides considered for reversible sorption, and the rationale for their selection, are 
listed below. Characteristics of the radionuclides considered were a combination of long 
half-life, relatively strong sorption characteristics, relatively large inventory in the repository, 
and observed field and laboratory behavior. 
 Plutonium. Important isotopes with respect to dose calculations include 239Pu, 242Pu, 
and 238Pu. Plutonium is sparingly soluble but sorbs strongly to oxide mineral surfaces 
(generally less strongly to silicates) (EPA 1999 [DIRS 147475]). The formation of 
pseudocolloids through the sorption of plutonium particularly to iron oxyhydroxide 
colloids from corrosion of steel in the repository would increase the effective mobility 
of plutonium. 
 Americium. Americium is sparingly soluble but sorbs strongly to mineral surfaces, 
including colloids. Laboratory investigations have shown that it sorbs strongly to 
colloids and mineral surfaces. Americium exhibits large Kd values often in the range 
of 1,000 ml/g to >100,000 ml/g. However, the concentrations of dissolved americium 
may be controlled by precipitation in many environments. Therefore, high reported 
sorption measurements may reflect precipitation reactions (EPA 2004 
[DIRS 172215]). 

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 Thorium. Thorium adsorption occurs at concentrations <10-9 M and will precipitate 
at concentrations >10-9 M based on the solubility of Th(OH)4 at pH 5.5. Thorium has 
been observed in waste glass tests as a colloid often associated with phosphate. 
Similar observations have been made in natural systems. 
 Neptunium. The NpO2
+ ion will be soluble under anticipated repository conditions 
and is less likely to participate in colloid formation than the other actinides. The 
sorption properties of Np(V) have a strong dependence on pH (EPA 2004 [DIRS 
172215]). Retention or sorption of neptunium in alteration products has been 
observed under some conditions (Mnard et al. 1998 [DIRS 171053]; Lu et al. 2000 
[DIRS 166315]; McNamara et al. 2005 [DIRS 174068]). However, with the 
anticipated pH range of 4.5 to 8.5 in the waste package, sorption of neptunium is 
expected to be minimal. The sorption edge (i.e., pH where 50% of the species is 
sorbed) for neptunium (1  10-6 M) on hematite (1 g/l) occurs at pH 8 (Hardin 1998 
[DIRS 100123], p. 6A-23, Figure 628). 
 Cesium. Few stable cesium complexes will exist in groundwater. Cs+ adsorbs 
strongly to mica-like clay minerals, forming inner sphere complexes. The degree of 
sorption/desorption will depend on the availability of clays and the concentration of 
ions that will compete for adsorption sites. 
 Protactinium. The +5 state is the normal oxidation state for protactinium and there 
is quite a close resemblance to the chemistry of niobium and tantalum; however, 
Pa(IV) is also stable in solution. The chemistry of protactinium is dissimilar to either 
plutonium or americium; however, the potential colloidal behavior of protactinium 
has been assumed conservatively to be similar to americium for modeling purposes. 
 Uranium. Uranium (235U, 233U, and 238U) will be the most abundant radioactive 
element in the repository. For this reason uranium was considered in the analysis. In 
solution, UO2
2+ forms soluble complexes with carbonate, oxalate, and hydroxide; 
UO2
2+ may also be susceptible to adsorption either by organic matter, iron 
oxyhydroxides, or by precipitation with various anions, including silicate, vanadate, 
arsenate, and phosphate (EPA 1999 [DIRS 147475]). The sorption edge for uranium 
(2  10-6 M) on hematite (1 g/l) has been reported to be near pH 5.5 (Hardin 1998 
[DIRS 100123] p. 6A-23, Figure 6-28). However, it is conservative to consider 
uranium to be a non-sorbing soluble species. Uranium was not included in the 
reversible-sorption portion of the colloid-associated radionuclide transport analysis 
To a large extent, the effectiveness of colloids at facilitating contaminant transport is due to their 
large surface area (relative to mass and volume) available for sorption. Depending on the size 
distribution of colloids in the aqueous phase, the impact of choosing a mass-based Kd value, or a 
surface-area-based Ka, may be significant. The greatest deviation exists in situations in which a 
large number of small colloids exist, which have a high surface-area-to-mass ratio. Available 
sorption information for validation of the model is mass based, and conversion to surface-areabased 
information is not straightforward; however, in most instances, the data used for 
determining Kd ranges was obtained from project-specific conditions for colloidal hematite and 
montmorillonite. In the case of cesium sorption, YMP-specific data for hematite and 

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MDL-EBS-PA-000004 REV 02 6-36 August 2005 
montmorillonite was unavailable. The much higher surface area of hematite colloids was 
accounted for by scaling the surface area of the hematite solids used in experiments by 
Torstenfelt et al. (1982 [DIRS 171770]). In contrast, as most sorption in montmorillonite occurs 
via ion exchange processes in the interlayer, no such scaling was necessary (Atun et al. 1996 
[DIRS 164865]. 
By considering the effect of aqueous chemical conditions on sorption, the selection of relevant 
Kd values is simplified. pH and ionic strength (particularly concentrations of Na+, Ca2+, and K+) 
primarily control the behavior of the mineral surface. The behavior of the sorbate is primarily 
controlled by its oxidation state, pH, and the partial pressure of carbon dioxide (pCO2). For 
minerals in which the sorption mechanism is primarily by ion exchange (e.g., clay minerals), 
ionic strength impacts sorption of alkali metals on clay minerals (especially at their edges), 
because cations compete with radionuclides for exchange sites (Atun et al. 1996 
[DIRS 164865]). 
The chemistry of plutonium is more complex than any of the other actinides, and it can exist in 
the III, IV, V, and VI oxidation states with lower solubility and mobility for the lower oxidation 
states III and IV (Runde et al. 2002 [DIRS 168432]; Romanovski et al. [DIRS 170890]). The 
actinide elements exhibit strong hydrolysis behavior and tend to form strong carbonate 
complexes in the absence of other complexing ligands. The actinides may hydrolyze at pH 
values above about 6 where they are present as neutral hydroxo complexes. At slightly alkaline 
pH values and above, they will exist as negatively charged anionic hydroxyl complexes. Under 
repository conditions, the presence of silicate, phosphate, and other species may alter the fate of 
the actinides. However, sorption is greatest in the near neutral pH range. Carbonate is a strong 
complexant, and at alkaline pH values, the actinides may form mono-, di-, and tri-carbonate 
complexes, depending on the actinide (and valence). The presence of these complexes tends to 
decrease sorption. 
To develop colloid Kd values, information provided by project data and published literature was 
considered. Lu et al. (1998 [DIRS 100946]; 2000 [DIRS 166315]) provide useful information on 
the sorption of Pu(IV) and Pu(V) on colloidal dispersions of hematite, goethite, montmorillonite, 
and silica. Information on Am(III) is provided for all but goethite. Sorption was measured as a 
function of time up to 4 or 10 days. Desorption was measured after 150 days because desorption 
tends to be much slower in these mineral-sorbate systems. The experimentally determined Kd 
values are summarized in Table 6-5a and plotted in Figures 6-13a and 6-13b for plutonium and 
americium, respectively. 
Table 6-5a. Experimentally Determined Kd Values for Plutonium and Americium Sorption Onto Hematite 
and Montmorillonite 
Kd Values, mL/g Phenomena 
Investigated J-13 Well Water Synthetic J-13 Well Water 
Plutonium-Hematite (DTN: LA0003NL831352.002 [DIRS 148526]) 
Sorption (forward, up to 10 days; colloid 
concentration not specified in DTN) 5.00  103 to 1.10  105 1.20  105 to 7.00  105 
Desorption (backward, 32 days; colloid 
concentration not specified in DTN) 3.40  105 3.20  105 
Colloid Concentration (forward, 10 mg/L to 
150 mg/L, 10 days) 7.50  103 to 3.70  104 7.70  103 to 2.60  104 

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MDL-EBS-PA-000004 REV 02 6-37 August 2005 
Table 6-5a. Experimentally Determined Kd Values for Plutonium and Americium Sorption Onto Hematite 
and Montmorillonite (Continued) 
Kd Values, mL/g Phenomena 
Investigated J-13 Well Water Synthetic J-13 Well Water 
Temperature (forward, 20C to 80C, 
10 days) 1.70  105 to 4.90  106 (not in DTN) 
Plutonium-Montmorillonite (DTN: LA0003NL831352.002 [DIRS 148526]) 
Sorption (forward, up to 10 days; colloid 
concentration not specified in DTN) 2.30  102 to 5.80  103 2.40  102 to 6.50  103 
Desorption (backward, 32 days; colloid 
concentration not specified) 6.70  103 6.50  103 
Colloid Concentration (forward, 10 mg/L to 
150 mg/L, 10 days) 5.20  103 to 2.70  104 3.20  103 to 2.40  104 
Temperature (forward, 20C to 80C, 
10 days) 9.50  103 to 2.00  105 (not in DTN) 
Americium-Hematite (DTN: LA0005NL831352.001 [DIRS 149623]) 
Sorption (forward, up to 10 days; colloid 
concentration not specified in DTN) 7.90  104 to 1.00  107 1.20  105 to 6.20  106 
Desorption (backward, 200 days; colloid 
concentration not specified) 1.90  105 8.40  104 
Colloid Concentration (forward, 10 mg/L to 
150 mg/L, 10 days) 1.90  106 to 1.60  107 1.80  106 to 2.80  107 
Temperature (forward, 20C to 80C, 
10 days) 1.80  105 to 6.50  105 (not in DTN) 
Americium-Montmorillonite (DTN: LA0005NL831352.001 [DIRS 149623]) 
Sorption (forward, up to 10 days; colloid 
concentration not specified in DTN) 7.10  103 to 1.0  105 2.30  104 to 1.10  105 
Desorption (backward, 200 days; colloid 
concentration not specified) 9.10  104 1.80  104 
Colloid Concentration (forward, 10 mg/L to 
150 mg/L, 10 days) 1.60  105 to 1.50  106 1.60  105 to 8.40  105 
Temperature (forward, 20C to 80C, 
10 days) 6.00  104 to 2.70  105 (not in DTN) 
Source: Lu et al. 2000 [DIRS 166315]; DTNs: LA0003NL831352.002 [DIRS 148526], LA0005NL831352.001 
[DIRS 149623]. 
Waste-form colloids, corrosion-generated colloids, and seepage watergroundwater colloids may 
become pseudocolloids through reversible sorption of aqueous radionuclides onto them. Kd 
values for combined adsorption and desorption of plutonium, americium, protactinium, cesium, 
and thorium on waste form, iron oxideoxyhydroxides, and groundwater colloids are used to 
determine effective reversibility of the radionuclides on colloids according to the relationship: 
RNadsorbed/desorbed = RNdissolved  Kd,RN  Mcolloid (Eq. 5) 
where RN = radionuclide, Kd,RN is the sorption coefficient (ml/g) for radionuclide RN, and 
Mcolloid is the concentration of colloids in moles/L. 
The Kd values in Table 6-6 are based on the following: 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-38 August 2005 
 Ranges are drawn from the following references: Lu et al. 2000 [DIRS 166315]; 
DTNs: LA0003NL831352.002 [DIRS 148526]; LA0005NL831352.001 
[DIRS 149623], LA0407AM83141.002 [DIRS 170621]; Torstenfelt et al. 1982 [DIRS 
171770]. 
 For a given radionuclide the maximum value of each range is the same to allow for 
the possibility that iron oxideoxyhydroxide will occur as colloids and as coatings on, 
or microcrystalline aggregates in association with, smectite colloids in the iron-rich 
waste package environment. Figure 6-13b illustrates that sorption is dissimilar for 
americium with hematite and montmorillonite; however, the model simplification, in 
this case, is conservative. The similar sorption characteristics of smectite and 
hematite for plutonium (Lu et al. 2003 [DIRS 170883]) and the dissimilar results for 
montmorillonite, support the idea that iron oxides can be associated with clays and 
impact their sorption characteristics. 
 In the case of cesium sorption on hematite, the distribution has been corrected for the 
low surface area solids used in the experimental data compared to the relatively high 
surface area colloids, although literature data suggests cesium sorption onto hematite 
is negligible (Torstenfelt et al. 1982 [DIRS 171770]). 
 Sorption measurements that result in high Kd values for americium 
(i.e., DTN: LA0005NL831352.001 [DIRS 149623]) may have been affected by 
americium precipitation reactions (EPA 2004 [DIRS 172215]); hence, the distribution 
curves for americium sorption on hematite have been lowered. 
 The Kd intervals and probabilities are intended to cover the entire ranges and 
emphasize the higher ends of the ranges but deemphasize the highest intervals. These 
distributions are used because (1) the ranges are based on data for colloids and larger 
minerals, and it is believed that Kd values for colloids will be higher than for larger 
minerals due to their higher specific surface area; and (2) because of the mechanisms 
of surface phenomena that actually prevail, Kd values tend to overpredict desorption 
over time of actinide metals from iron oxyhydroxide colloids (i.e., they underpredict 
retention by the colloids). 
 Kd values ranges and distributions for protactinium are assumed to be those of 
americium primarily because of little data on protactinium. Project data were 
available for protactinium sorption on zeolitic materials but this was not used. 
Although protactinium generally has a lower Kd than either plutonium or americium, 
the model has taken the conservative step of assuming protactinium sorption is 
identical to that of americium (Section 5.1). 
 Kd values ranges and distributions for thorium are assumed to be those of americium 
owing to similar sorption behavior (Section 5.1). 
The input distributions are plotted in Figures 6-13a, 6-13b, and 6-13c as cumulative distributions. 
The data in Table 6-7 for cesium sorption is bounded by other studies on silicate minerals 
reported by Beckman et al. (1988 [DIRS 144956]; Section 7.1). 

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MDL-EBS-PA-000004 REV 02 6-39 August 2005 
0 
0.2 
0.4 
0.6 
0.8
1 
1.00E+03 1.00E+04 1.00E+05 1.00E+06 
Sorption Coefficients for Plutonium (mL/g) 
Distribution Function 
Hematite 
Montmorillonite Clay 
Source: DTNs: LA0003NL831352.002 [DIRS 148526], LA0003NL831352.001 [DIRS 149172]. 
NOTES: The straight lines are the distribution ranges as shown in Table 6-6. 
Figure 6-13a. Cumulative Distribution of Sorption Coefficients obtained for Plutonium on Clay 
and Hematite Colloids 
0.00 
0.20 
0.40 
0.60 
0.80 
1.00 
1.00E+04 1.00E+05 1.00E+06 1.00E+07 1.00E+08 
Sorption Coefficients for Americium (mL/g) 
Distribution Function 
Hematite-Data 
Montmorillonite Clay-Data 
Source: DTN: LA0005NL831352.001 [DIRS 149623] (extracted from Table S000277_001). 
NOTES: The straight lines are the distribution ranges as shown in Table 6-6. 
Figure 6-13b. Cumulative Distribution of Sorption of Americium on Montmorillonite and 
Hematite Colloids 
Figure 6-13a shows the sorption for plutonium and Figure 6-13b shows the sorption data for 
americium. The straight lines describe the distribution fits as listed in Table 6-6. As iron oxides 
will sorb onto clays in a repository environment, the clay colloids in the waste package 
environment would have an effectively larger Kd range than the pure mineral (montmorillonite) 
sorption studies from Lu et al. (2000 [DIRS 166315]) show. Indeed, sorption studies by Lu et al. 
(1998 [DIRS 100946]) demonstrate that smectite, a poorly crystallized phase that contains iron 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-40 August 2005 
had similar sorption characteristics to hematite rather than the pure mineral phase, 
montmorillonite (Lu et al. (2000 [DIRS 166315]). The high Kd for americium sorption onto 
hematite reported by Lu et al. (2000 [DIRS 166315]) may have been affected by precipitation. 
This issue has been raised (EPA 2004 [DIRS 172215]) regarding use of high Kd values for 
americium. For this reason, the distribution range for americium sorption onto hematite has been 
reduced from the experimental data. 
In Figure 6-13c, the distribution of estimate values for a series of cesium sorption tests on Yucca 
Mountain silicate minerals (including zeolites) is displayed. As smectite clay is a common 
surface alteration product on silicate glasses, application of the data in Table 6-7 is appropriate 
for determining sorption to montmorillonite clay. Zeolites display much higher sorption values 
for cesium and although zeolites have been sometimes observed as suspended particles in 
leachates from DHLWG corrosion tests, they are not included in the model abstraction. Sorption 
of cesium onto hematite is negligible. Values published by Torstenfelt et al. (1982 
[DIRS 171770]) are 2 mL/g to 8 mL/g; however, the particles used in these sorption experiments 
ranged from 45 m to 63 m, assuming the density of hematite is 5.27 g/cm3. The effective Kd 
range of hematite was adjusted based on available surface areas that would be on colloidal 
hematite. Table 6-5b presents the ranges for hematite colloids. 
Table 6-5b. Characteristics of Particles from Cesium Sorption on Hematite Experiments 
Particle Size (m)a SA/gram (m2/g)a 
Effective 
SA/g(m2/g)b 
Kd 
(ml/g) 
Adjusted Range 
(ml/g) 
Kd
a Effective Kd 
45 to 65 0.018 to 0.025 0.054 to 0.076 2 20 200 
8 170 800 
Source: aTorstenfelt et al. 1982 [DIRS 171770]. 
NOTE: bA roughness factor of 3 was applied. 

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MDL-EBS-PA-000004 REV 02 6-41 August 2005 
0 
0.2 
0.4 
0.6 
0.8
1 
1.00E+01 1.00E+02 1.00E+03 1.00E+04 1.00E+05 
Sorption Coefficient for Cesium (ml/g) 
Distribution Function 
Other Silicates 
Zeolites 
Hematite 
Hemate-TSPA 
Smecite-TSPA 
Source: DTN: LA0407AM831341.002 [DIRS 170621] ; Torstenfelt et al. 1982 [DIRS 171770] 
NOTES: Batch sorption coefficient data for cesium on Yucca Mountain Tuffs (devitrified tuff, zeolitic tuff, and vitrified 
tuff) at 20C to 80C. Sorption was slightly dependent on temperature. The nonzeolitic materials were 
used to establish sorption distribution. Hematite colloid data by Torstenfelt et al. (1982 [DIRS 171770]) is 
included in the plot and has been adjusted to account for the higher surface area of colloids and has error 
bars accounting particle-size variability. 
Figure 6-13c. Cumulative Distribution Function Showing the Distribution of Cesium Sorption Coefficients 
Obtained on Hematite, Zeolite, and Other Silicates 
Table 6-6. Kd Values (mL/g) and Interval Probabilities Used for Reversible Radionuclide Sorption on 
Colloids in Total System Performance Assessment-for the License Application 
Calculations 
Radionuclide Colloid Kd Value Range 
(mL/g) 
Kd Value Intervals 
(mL/g) 
Kd Value Interval 
Probabilities 
Iron oxyhydroxide 104 to 106 
<1  104 
1  104 to 5  104 
5  104 to 1  105 
1  105 to 5  105 
5  105 to 1  106 
> 1  106 
0 
0.15 
0.2 
0.5 
0.15 
0 
Plutonium 
smectite 103 to 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 
0 
0.04 
0.08 
0.25 
0.2 
0.35 
0.08 
0 

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MDL-EBS-PA-000004 REV 02 6-42 August 2005 
Table 6-6. Kd Values (mL/g) and Interval Probabilities Used for Reversible Radionuclide Sorption on 
Colloids in Total System Performance Assessment-for the License Application Calculations 
(Continued) 
Radionuclide Colloid Kd Value Range 
(mL/g) 
Kd Value Intervals 
(mL/g) 
Kd Value Interval 
Probabilities 
Iron oxyhydroxide 105 to 107 
< 1  105 
1  105 to 5  105 
5  105 to 1  106 
1  106 to 5  106 
5  106 to 1  107 
> 1  107 
0 
0.15 
0.2 
0.55 
0.1 
0 
Americium, 
thorium, 
protactinium 
smectite 104 to 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 
0 
0.07 
0.1 
0.23 
0.2 
0.32 
0.08 
0 
Iron oxyhydroxide 101 to 103 
<1  101 
1  101 to 5  101 
5  101 to 1  102 
1  102 to 5  102 
5  102 to 1  103 
> 1  103 
0 
0.13 
0.22 
0.55 
0.1 
0 
Cesium 
smectite 102 to 104 
< 1  102 
1  102 to 5  102 
5  102 to 1  103 
1  103 to 5  103 
5  103 to 1  104 
> 1  104 
0 
0.2 
0.25 
0.5 
0.05 
0 
Source: DTNs : LA0003NL831352.002 [DIRS 148526] ; LA0005NL831352.001 [DIRS 149623] ; 
LA0407AM831341.002 [DIRS 170621] ; LA0003NL831352.001 [DIRS 149172] ; and Torstenfelt et al. 
1982 [DIRS 171770]. 
NOTE: In engineered barrier system calculations, the upper bound for plutonium and americium on iron 
oxyhydroxide was reduced by a factor of 100 to be compatible with the mechanistic sorption model 
described in Section 6.3.3.2. 
Table 6-7. Cesium Sorption Coefficients on Yucca Mountain Tuffs 
Rock 
Sample Sample Identification 
Temperature 
(C) Material Kd (mL/g) 
G4-270 J-13 G4 -270-C.967-20 20.0 Devitrified Tuff 1.24E+02 
G4-270 J-13 G4-270-C.968-20 20.0 Devitrified Tuff 1.25E+02 
G4-270 J-13 G4-270-C.962-20 20.0 Devitrified Tuff 1.26E+02 
G4-270 J-13 G4-270-C.1029-80 80.0 Devitrified Tuff 1.66E+02 
G4-270 J-13 G4-270-C.1030-80 80.0 Devitrified Tuff 1.71E+02 
G4-270 J-13 G4-270-C.956-20 20.0 Devitrified Tuff 2.13E+02 
G4-270 J-13 G4 -270-C.961-20 20.0 Devitrified Tuff 2.16E+02 
G4-270 J-13 G4 -270-C.955-20 20.0 Devitrified Tuff 2.20E+02 
G4-270 J-13 G4-270-C.1021-60 60.0 Devitrified Tuff 2.39E+02 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-43 August 2005 
Table 6-7. Cesium Sorption Coefficients on Yucca Mountain Tuffs (Continued) 
Rock 
Sample Sample Identification 
Temperature 
(C) Material Kd (mL/g) 
G4-270 J-13 G4-270-C.1022-60 60.0 Devitrified Tuff 2.45E+02 
GU3-1407 J-13 GU3-1407-C.966-20 20.0 Vitric Tuff 5.17E+02 
GU3-1407 J-13 GU3-1407-C.965-20 20.0 Vitric Tuff 5.26E+02 
GU3-1407 J-13 GU3-1407-C.1031-80 80.0 Vitric Tuff 6.72E+02 
GU3-1407 J-13 GU3-1407-C.959-20 20.0 Vitric Tuff 7.67E+02 
GU3-1407 J-13 GU3-1407-C.1032-80 80.0 Vitric Tuff 7.68E+02 
GU3-1407 J-13 GU3-1407-C.960-20 20.0 Vitric Tuff 9.50E+02 
GU3-1407 J-13 GU3-1407-C.1024-60 60.0 Vitric Tuff 1.03E+03 
GU3-1407 J-13 GU3-1407-C.972-20 20.0 Vitric Tuff 1.04E+03 
GU3-1407 J-13 GU3-1407-C.971-20 20.0 Vitric Tuff 1.05E+03 
GU3-1407 J-13 GU3-1407-C.1023-60 60.0 Vitric Tuff 1.06E+03 
G4-1506 J-13 G4-1506-C.957-20 20.0 Zeolitic Tuff 7.38E+03 
G4-1506 J-13 G4-1506-C.964-20 20.0 Zeolitic Tuff 8.27E+03 
G4-1506 J-13 G4-1506-C.963-20 20.0 Zeolitic Tuff 1.02E+04 
G4-1506 J-13 G4-1506-C.1027-80 80.0 Zeolitic Tuff 1.65E+04 
G4-1506 J-13 G4-1507-C.1028-80 80.0 Zeolitic Tuff 1.75E+04 
G4-1506 J-13 G4-1506-C.1019-60 60.0 Zeolitic Tuff 1.96E+04 
G4-1506 J-13 G4-1506-C.970-20 20.0 Zeolitic Tuff 2.97E+04 
G4-1506 J-13 G4-1506-C.1020-60 60.0 Zeolitic Tuff 3.52E+04 
G4-1506 J-13 G4-1506-C.969-20 20.0 Zeolitic Tuff 3.64E+04 
G4-1506 J-13 G4-1506-C.958-20 20.0 Zeolitic Tuff 4.23E+04 
Source: DTN: LA0407AM831341.002 [DIRS 170621]. 
6.3.3.2 Irreversible Sorption of Plutonium and Americium to Iron Oxyhydroxide 
Colloids and Stationary Phases 
Evidence from sorption experiments with plutonium and americium (Lu et al. 2000 
[DIRS 166315]) with colloidal hematite and montmorillonite show that the rates of desorption 
(backward rate) of plutonium and americium are slower than the rates of sorption (forward rate). 
Thus, plutonium and americium are assumed to be so strongly sorbed to colloids that they are 
considered irreversibly sorbed and are modeled as such within the engineered barrier system 
(BSC 2004 [DIRS 169868]). 
At the Rocky Flats site in Colorado, soil plutonium is largely associated with the negatively 
charged organic macromolecular fraction and not with the more abundant iron oxides and clays 
(Santschi et al. 2002 [DIRS 170923]; Ibrahim and Salazar 2000 [DIRS 170882]). Litaor and 
Ibrahim (1996 [DIRS 161667]) used 0.01M CaCl2 as an extractant and measured plutonium in 
Rocky Flats soil to be 0.04% to 0.08% exchangeable. Laboratory experiments of plutonium 
sorption onto iron oxides have shown that ~1% of the initially sorbed plutonium can be desorbed 
into solution, even after months of time have elapsed (Lu et al. 2000 [DIRS 166315]). 
In order to accommodate these observations in the iron oxyhydroxide colloid abstraction, the 
TSPA-LA model calculates irreversible sorption of plutonium and americium as functions of 

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MDL-EBS-PA-000004 REV 02 6-44 August 2005 
specific surface area (SA), mass of corrosion colloids, dissolved concentration of plutonium and 
americium, target-flux out ratio (FRN, Table 6-12), and other parameters (such as flow and 
diffusion rate) internal to the TSPA-LA model. This is done in such a way that a majority 
(90% to 99%) of plutonium and americium are irreversibly sorbed, due to incorporation of 
parameter uncertainties (see BSC 2004 [DIRS 169868]). The reversibly sorbed portion is 
determined according to an equilibrium Kd value model (per Section 6.3.3.1; see Table 6-6 for 
values, ranges, and distributions). 
This approach is mechanistically more advanced than the approach taken for TSPA-SR and in 
prior efforts. Previous treatments have relied solely on the application of Kds to describe uptake 
of dose-critical radionuclides onto colloids. A clearer understanding of the identity and 
abundance of iron oxyhydroxide colloids in the engineered barrier system has prompted the more 
fundamental modeling effort. At the same time, more recent analyses of actinide transport in the 
environment have provided evidence for irreversible sorptionthe tendency for sorbed material 
to desorb slowly from iron oxyhdroxide colloids. To accurately represent this, the iron 
oxyhydroxide colloid facilitated radionuclide transport model separates sorbed plutonium and 
americium each into two populationsan irreversibly sorbed fraction and a reversibly sorbed 
fraction. Accordingly, in the TSPA-LA model, the upper bound of the Kd range for the 
reversibly sorbed fraction is reduced by a factor of 100 (which constrains the Kd value to the 
lower bound, which is a fixed value). 
Sorption to immobile iron oxyhydroxide within the waste package is also modeled. Honeyman 
and Ranville (2002 [DIRS 161657]) 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. For the symmetrical system described in Section 6.3.3.1, by 
far most of the americium mass was calculated as attached to the stationary phase; at most, 
0.001% of the total americium was associated with colloids (Honeyman and Ranville 2002 
[DIRS 161657], p. 140). 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). On the one hand plutonium will tend to attach to the 
colloids with their higher specific surface area; on the other hand 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. 
Specific Surface Area (SA) Values for specific surface area (SA) for three different iron 
oxyhydroxides, goethite, hematite, and ferrihydrite (or hydrous ferric oxide) as tabulated from 
the literature are listed in Table 6-8. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-45 August 2005 
Table 6-8. Published Values for Specific Surface Area (SA) for Iron Phases 
SA (m2/g) 
Mineral/Phase 
Min Max 
Goethite (FeOOH) 31.0 153.3 
Hematite (Fe2O3) 1.8a 5.4 
Ferrihydrite (5Fe2O39H2O) 240.8 720.0b 
Source: Roden and Zachara 1996 [DIRS 171518], Table 1, p. 1,621; 
Poulton et al. 2004 [DIRS 171519], p. 3,705. 
NOTE: a Minimum value for hematite was obtained by dividing measured 
values by ~3 to provide a more bounding range. 
bConservative value estimated from multiplying N2/BET 
determined values by ~3. This high surface area may not be 
achievable in any repository environment. 
Surface area measurements on similar materials may sometimes differ by up to an order of 
magnitude depending on the method used; however, the maximum values used in this model are 
conservative. Values reported by Roden and Zachara (1996 [DIRS 171518]) and Poulton et al. 
(2004 [DIRS 171519]) (Table 6-8) were obtained by N2 gas-adsorption measurements. Nitrogen 
gas adsorption measurements may underestimate surface areas because of microporosity (Davis 
and Kent 1990 [DIRS 143280]). The surface area of 720 m2/g of ferrihydrite has not been 
verified by any experimental measurements and is unlikely to occur naturally. Using the site 
density range of 2.16 sites/nm2 to 16.8 sites/nm2 (Davis and Kent (1990 [DIRS 143280]), a 2- 
nm-diameter ferrihydrite particle with monolayer coverage of thorium could not sorb more than 
5 g of thorium per gram of iron oxide. This is discussed in more detail in Appendix IV. 
However, this value was chosen as it represents a conservative surface area. The SA for iron 
oxyhydroxide corrosion phases in the colloid model implemented in TSPA-LA calculations 
(Table 6-8) is: 
Surface area (SA) is set between 1.8 and 720 m2/g (log uniform distribution). 
Forward Rate Constant for Irreversible Sorption (k)A mechanistic model for irreversible 
sorption of plutonium and americium on iron oxyhydroxide colloids and corrosion products for 
the TSPA-LA is developed and described in EBS Radionuclide Transport Abstraction 
(BSC 2004 [DIRS 169868], Appendix B). The forward rate constant for irreversible sorption is 
computed as a function of flow rate, colloid suspension stability, and other system parameters. 
However, under certain chemical and hydrologic scenarios modeled in TSPA-LA, such as, under 
zero advective flux conditions, unstable colloid conditions (ionic strength greater than 0.05 M), 
or low flow rates, the mechanistic model fails to compute reasonable values for the forward 
sorption rate constant of radionuclides onto iron oxyhydroxide colloids (designated FeOx 
colloids). For these scenarios, it is necessary that the forward rate constant for adsorption be 
determined independently of flow and transport parameters. It is more appropriate to use a rate 
constant determined from the analysis of time-dependent sorption experiments. 
A series of YMP-specific sorption experiments performed by Lu et al. (2000 [DIRS 166315]) 
provide useful information on the sorption rate of Pu(IV), Pu(V), and Am(III) on colloidal 
dispersions of hematite and montmorillonite. Since the forward rate constant for irreversible 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-46 August 2005 
sorption to iron oxyhydroxide colloids is of interest, the sorption data for hematite and goethite 
colloids is considered. The sorption data for plutonium (Table 6-9) used to develop the k values 
are given in the DTN: LA0004NL831352.001 [DIRS 150272]. A total of eight experiments 
were performed, four each for Pu(V) and Pu(IV), using J-13 well water and synthetic J-13 well 
water on colloids of hematite (Fe2O3) and goethite (FeOOH). 
A first order reaction in which the sorption rate depends upon the plutonium concentration in the 
water is considered for calculating the reaction rate constant from the available data sets. This is 
expressed as: 
solution 
solution Rf 
dt 
df - = (Eq. 6) 
where, solution f is the time-dependent fraction in solution (complement of fraction sorbed), R is 
the linear rate constant, and t is the time duration. 
The solution of the rate equation is: 
) exp( ) 0 ( ) ( Rt f t f solution solution - = (Eq. 7) 
where, the initial fraction in solution is one, 1 ) 0 ( = solution f . 
Upon evaluation of the data presented in Table 6-9 and reported in the 
DTN: LA0004NL831352.001 [DIRS 150272], it was observed that a relatively high sorption 
rate occurs at early time (duration less than one hour) with a gradual decline in sorption rate at 
later times. Since the intent is to provide an estimate for a first order rate describing the 
irreversible sorption of plutonium onto iron oxide/oxyhydroxide colloids (and stationary 
corrosion products) over a period ranging from tens of years to of several hundred years, the data 
at initial time and the end of experimental time have been used to estimate the average linear rate 
constant. 
In Table 6-9, an exponential fit is performed on each of the eight datasets using the initial 
condition, 1 ) 0 ( = solution f , and the end of the experiment measurement, ) 5760 ( solution f . Equation 7 
was solved for R using the experimental values of (t) for t = 0 and t = 5760. Values of R [yr-1] 
are listed in the last column of Table 6-9. The computed values of R range from a minimum of 
105 (yr-1) to a maximum of 2,523 (yr-1). 
Table 6-9. Time-Dependent FeOx Colloid Sorption Fraction for Plutonium and Rate Constant Calculation 
Time (min) 
Fraction Pu 
Sorbed 
Fraction Pu in 
Solution 
Computed Rate, 
R [yr-1] 
Sample: Fe2O3.J13.PuV 
0 0 1 
10 0.539 0.461 
30 0.565 0.435 
60 0.594 0.406 
240 0.674 0.326 
360 0.74 0.26 
1,440 0.907 0.093 
2,880 0.948 0.052 473 

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MDL-EBS-PA-000004 REV 02 6-47 August 2005 
Table 6-9. Time-Dependent FeOx Colloid Sorption Fraction for Plutonium and Rate Constant 
Calculation (Continued) 
Time (min) 
Fraction Pu 
Sorbed 
Fraction Pu in 
Solution 
Computed Rate, 
R [yr-1] 
5,760 0.994 0.006 
Sample: Fe2O3.SYNJ13.PuV 
0 0 1 
10 0.933 0.067 
30 0.966 0.034 
60 0.957 0.043 
240 0.996 0.004 
360 0.998 0.002 
1,440 0.999 0.001 
2,880 1 0 
5,760 0.999 0a 2,523 
Sample: FeOOH.J13.PuV 
0 0 1 
10 0.303 0.697 
30 0.328 0.672 
60 0.36 0.64 
240 0.515 0.485 
360 0.595 0.405 
1,440 0.793 0.207 
2,880 0.869 0.131 
5,760 0.902 0.098 210 
Sample: FeOOH.SYNJ13.PuV 
0 0 1 
10 0.831 0.169 
30 0.831 0.169 
60 0.837 0.163 
240 0.911 0.089 
360 0.932 0.068 
1,440 0.987 0.013 
5,760 0.991 0.009 420 
Sample: Fe2O3.J13.PuIV 
0 0 1 
10 0.941 0.059 
30 0.982 0.018 
60 0.992 0.008 
240 0.993 0.007 
360 0.997 0.003 
1,440 0.994 0.006 
2,880 0.999 0.001 
5,760 1 0 1,261.44 
Sample: Fe2O3.SYNJ13.PuIV 
0 0 1 
10 0.995 0.005 
30 0.997 0.003 
60 0.997 0.003 
240 0.993 0.007 
360 0.996 0.004 
1,440 0.999 0.001 
2,880 1 0 
5,760 0.999 0a 2,523 

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MDL-EBS-PA-000004 REV 02 6-48 August 2005 
Table 6-9. Time-Dependent FeOx Colloid Sorption Fraction for Plutonium and Rate Constant 
Calculation (Continued) 
Time (min) 
Fraction Pu 
Sorbed 
Fraction Pu in 
Solution 
Computed Rate, R 
[yr-1] 
Sample: FeOOH.J13.PuIV 
0 0 1 
10 0.496 0.504 
30 0.558 0.442 
60 0.834 0.166 
240 0.733 0.267 
360 0.703 0.297 
1,440 0.648 0.352 
2,880 0.689 0.311 
5,760 0.621 0.379 105 
Sample: FeOOH.SYNJ13.PuIV 
0 0 1 
10 0.523 0.477 
30 0.627 0.373 
60 0.832 0.168 
240 0.834 0.166 
360 0.921 0.079 
1,440 0.918 0.082 
2,880 0.861 0.139 
5,760 0.929 0.071 263 
Source: DTN: LA0004NL831352.001 [DIRS 150272]. 
NOTE: aThese values were 0.001, but were changed to zero (0) due to consideration 
of irreversible sorption processes. 
Evaluation of the data in Table 6-9 indicates that sorption of plutonium onto the iron 
oxyhyroxide colloids was lower in natural groundwater samples (J13) than in synthetic 
groundwater (SYNJ13), (i.e., the fraction of plutonium in solution appears to be consistently 
higher in the natural groundwater experiments). Lu et al. (2000 [DIRS 166315]) reported that 
these differences might be due to the adsorption of dissolved organic carbon present in the 
natural groundwater, thereby reducing the amount of plutonium adsorbed. Additionally, 
differences in the quantities of multivalent cations in the two experimental waters may have also 
influenced adsorption. However, given that data at initial time and the end of experimental time 
were used to estimate the average linear rate constant, these differences between the two 
experimental waters do not impact the computed rate (R). 
The calculated rate (R) is most applicable at pH values of approximately 7.2 to 8.4. The Lu et al. 
(2000 [DIRS 166315]) experiments had a starting solution pH of 8.2 to 8.4 with a 0.5 to 1 pH 
unit drop during the experiments. 
The rate computed above need to be converted to a per unit surface area basis in order to be 
applied to other mineralwater systems. The rate can be expressed as: 
k c S R c FeOx c FeOx _ _ 
 = (Eq. 8) 

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MDL-EBS-PA-000004 REV 02 6-49 August 2005 
where 
c FeOx S _ 
 = specific surface area of FeOx colloids [area FeOx colloids/mass FeOx colloids] 
c FeOx c _ = FeOx colloid concentration [mass FeOx colloid/water volume] 
k = forward rate constant per unit mineral surface area [water volume/area FeOx 
colloids/time]. 
Equation 8 is solved for forward rate constant, k, as: 
c FeOx c FeOx c S 
R k 
_ _ 
 = (Eq. 9) 
Lu et al. (2000 [DIRS 166315]) provide the values for g m S c FeOx / 5 . 53  2 
_ = , and 
L mg c c FeOx / 200 _ = for the experimental design and these parameters are used as direct input to 
these analyses (Table 4-1). These experiments use different values than those developed for iron 
oxyhydroxide colloid concentrations (Table 4-1) and for the specific surface area (SA) of these 
colloids (Table 4-2) in the engineered barrier system. Iron oxyhydroxide colloid concentrations 
referred to in Table 4-1 are based on experimental findings associated with the degradation of 
carbon steel miniature waste packages (Section 6.3.1.3). The concentration reported by Lu et al. 
(2000 [DIRS 166315]) was specifically prepared in the laboratory for their experiments. 
Likewise, the specific surface reported by Lu et al. (2000 [DIRS 166315]) was unique to their 
experiments, while the range of values referred to in Table 4-2 are based on values reported in 
the literature (Section 6.3.3.2). 
Substituting the experimental fixed values for these two parameters in Equation 9 and using the 
values of R listed in Table 6-9, the values for k are calculated to range from 0.01 (m3 m-2 yr-1) to 
0.24 (m3 m-2 yr-1). Therefore, based on the dataset (DTN: LA0004NL831352.001 
[DIRS 150272]), a log uniform distribution is recommended for sampling the value of k in the 
TSPA-LA model with a minimum of 0.01 (m3 m-2 yr-1) and a maximum of 0.24 (m3 m-2 yr-1) 
when the linear rate is determined independently of flow and transport parameters. 
6.3.3.3 Radionuclides Associated (Embedded) with DHLW Glass Colloids 
The colloid model abstraction assumes plutonium and americium are associated irreversibly 
(embedded) within smectite colloids from the DHLW glass degradation experiments (Section 
5.1) As such they are treated as a separate colloid subtype and assuming that the sorbed actinides 
are an intrinsic part of the colloid, not in equilibrium with the aqueous system. This premise has 
been based on a limited set of microscopic images obtained from corrosion tests and is not 
supported by results from waste glass corrosion tests reported by Mnard et al. (1998 [DIRS 
171053]) on the plutonium-doped glasses and by product consistency testing DHLW glass data 
from DTN: LL000123351021.117 [DIRS 143308]. The actinide particles are assumed to be 
attached irreversibly (or embedded) to the carrier colloid. In this model, waste form colloid 
parameters are represented entirely by the DHLW glass degradation colloids with irreversibly 
associated radionuclides (Section 6.3.1.1). 
In Figure 6-14, a series of images from waste form corrosion tests are shown. Each image shows 
the accumulation of actinides and rare earth elements (i.e., uranium, thorium, plutonium, 

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MDL-EBS-PA-000004 REV 02 6-50 August 2005 
lanthanides, etc.), either in particles found in the leachate or at the surface corrosion rind. 
Microscopy images represent snap-shots in time and specific location that may not necessarily 
represent the overall sample conditions. These images demonstrate a similarity in corrosion and 
colloidal behavior of glass and spent nuclear fuel. These images may also imply that the 
actinides are embedded in the carrier colloids; however, it is not known whether these phases 
persist at longer times. Indeed, thorium not plutonium or americium was identified in the 
particle in Figure 6-14A and only uranium and rare earths were observed in the particle 
agglomerate shown in Figure 6-14B. Furthermore, both particles exceed the nominal 
dimensions for a colloid and should be referred to as suspended particles, not colloids. . 
However, in the case of DHLW glass, modeling americium and plutonium as irreversible 
colloids is conservative and bounding (Section 5.3). 
Source: CRWMS M&O 2001 [DIRS 154071],(A) p. 26, Figure 12; (B) p. 31, Figure 18; (C) p. 28, Figure 15; 
(D) p. 34, Figure 21a. 
NOTE: (A) Suspended particle in leachate from thorium-rich glass unsaturated drip test, (B) suspended particle in 
leachate from spent nuclear fuel drip test, (C) corrosion rind from glass drip test, and (D) corrosion rind 
from spent nuclear fuel drip test. 
Figure 6-14. Transmission Electron Microscope Images of Defense High-Level Waste Glass and 
Commercial Spent Nuclear Fuel Corrosion Rinds and Colloids 
6.3.4 Potential Effects of Microbes and Organic Compounds 
Microbes and organic compounds will not affect the generation of inorganic colloids. Any 
interactions will tend to destabilize colloids, possibly decreasing the mobility of associated 
radionuclides. 

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MDL-EBS-PA-000004 REV 02 6-51 August 2005 
The effects of microbes in colloid-facilitated transport in the repository can include: 
 Microbial oxidation of metallic iron can produce iron-oxide colloids and aggregates 
 Microorganisms may decrease the concentration of stable colloid suspensions by 
aggregating colloidal material that they use as a food source. This has been shown to 
result in a decrease in colloid concentrations of up to 91% (Hersman 1995 [DIRS 
100750]). 
Natural organic matter, such as fulvic compounds from soils, will stabilize inorganic colloids in 
natural waters (Buffle et al. 1998 [DIRS 161653]). However, certain types of natural organic 
matter have been shown to have a tendency to destabilize colloids (Wilkinson et al. 1997 
[DIRS 161732]). 
6.3.4.1 Potential Microbial Communities Within the Engineered Barrier System 
Qualitative analyses and bounding calculations address microbial growth in Evaluation of 
Potential Impacts of Microbial Activities on Drift Chemistry (BSC 2004 [DIRS 169991], 
Section 7.1). That study recommends that TSPA use 90% relative humidity as a threshold for 
microbial-induced corrosion, but states that the potential impact of microbial activity on drift 
chemistry does not warrant a direct feed to the TSPA-LA (BSC 2004 [DIRS 169991], 
Section 7.1). 
In assessing the potential effects on microbial populations within the engineered barrier system, 
Evaluation of Potential Impacts of Microbial Activities on Drift Chemistry (BSC 2004 
[DIRS 169991]) 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., CO2, 
water vapor). Environmental limits on microbial activity considered include redox conditions, 
temperature, radiation, hydrostatic pressure, water activity, pH, salinity, available nutrients, and 
others. 
The study evaluated the potential for radionuclide transport facilitated by suspended microbial 
cells, or biocolloids (BSC 2004 [DIRS 169991], Section 6.5.4). The study concluded that the 
potential affect of radionuclide bearing colloids is negligible, and the presence of microbial cells 
and their by-products may even help immobilization of radionuclides. The transport and 
facilitated radionuclide migration are expected to be limited due to low liquid saturation, high 
ionic strength and low microbial activity. 
6.3.4.2 Discussion 
The presence of organic polymers such as those produced by microbes and their byproducts will 
generally reduce the stability of inorganic colloid suspensions and result in enlargement of 
particles through heteroaggregation (Wilkinson et al. 1997 [DIRS 161732]). 
It is currently anticipated that the repository invert, comprised of crushed volcanic tuff with 
intergranular void spaces, will remain undersaturated with respect to groundwater. The adhesion 
of microbes to unsaturated porous media, like the invert, can be controlled by the level of fluid 

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MDL-EBS-PA-000004 REV 02 6-52 August 2005 
saturation and the extent of the airwater interface network within the medium. Since microbes 
tend to adhere strongly to the airwater interface, lower saturations would result in increased 
microbe retention and reduced transport (BSC 2004 [DIRS 169991], Section 6.5.4). 
Considerable uncertainty surrounds the role that microbes may play in the repository system. 
Thus, given the current knowledge, it is difficult to estimate their effects on the colloid source 
term, in particular the effect of microbes on colloid concentration and suspension stability. 
However, it is understood that microbial action will tend to increase the sizes of inorganic 
colloids, which promotes gravitational settling and filtration. Thus, not including the effects of 
microbes in colloid source term and transport analysis is considered reasonable, and it follows 
that neglect of the effects of microbes is reasonable with respect to TSPA dose assessments. 
6.4 CONSIDERATION OF ALTERNATIVE MODELS 
The following sections describe five models considered as alternatives to the base-case model. 
Two models consider the formation of spent nuclear fuel-derived colloids, another considers a 
two-site sorption kinetic model as an alternative to the assumption of irreversible sorption used 
in the base-case model, a fourth considers waste form corrosion rate for bounding colloid 
formation, and the fifth model considers release of particles from a weathered waste form surface 
under unsaturated flow conditions where attachment to airwater interfaces may limit colloid 
movement. These models are listed in Table 6-10, along with the principal bases and screening 
assessments. 
Table 6-10. Alternative Conceptual Models Considered 
Alternative 
Conceptual 
Model Basis for Model Screening Assessment 
Reversible 
U(VI) Phase 
Colloids from 
Spent Nuclear 
Fuel 
Observation of actinide-bearing colloids from 
corrosion tests on CSNF and UO2 
(McNamara et al. (2005 [DIRS 174068]) 
Model uses U(VI) oxide phases as the 
primary colloid. (Wronkiewicz et al. (1996 
[DIRS 102047]. U(VI) colloidal phase 
suspension stability based on data Schindler 
et al. (2004 [DIRS 174507]) 
This ACM is screened out for the following 
reasons: 
 Although U(VI) oxide and silicate phases 
may form stable colloid suspensions over 
a range of pH conditions in Yucca 
Mountain groundwaters, U(VI) colloids 
might be anticipated to undergo some 
form of ripening process over time that 
would result in the destabilization of their 
suspensions. 
 The sorption characteristics for these 
phases is uncertain and there is little data 
showing the association of plutonium and 
americium with U(VI) colloids. 

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MDL-EBS-PA-000004 REV 02 6-53 August 2005 
Table 6-10. Alternative Conceptual Models Considered (Continued) 
Alternative 
Conceptual 
Model Basis for Model Screening Assessment 
Irreversible 
Sorbed 
Radionuclides 
on Spent 
Nuclear 
Fuel-Derived 
Colloids 
Observation of plutonium-rich submicrometer 
particles on the surface of 
corroded spent nuclear fuel under drip and 
flow conditions (Buck et al. (2004 
[DIRS 172668]). 
Pattern of transuranic accumulation in the 
corrosion rind is similar to that found in 
altered DHLWG. 
Stability of particle suspensions is based on 
data from Bitea et al. 2003 [DIRS 174504], 
Hsu et al. 1988 [DIRS 174350], and 
Parks 1964 (DIRS 174361]). 
This ACM is screened out for the following 
reasons: 
 Evidence for the release of the plutoniumrich 
particles has not been obtained from 
the various tests performed on spent 
nuclear fuel. 
 The colloidal particles may be subject to 
dissolution. 
 There are insufficient data to generate a 
complete model. 
Kinetic 
Sorption 
Model 
Instead of assuming irreversible attachment 
this model provides a method for predicting 
sorption behavior beyond laboratory timescales. 
Allows interpretation of data from 
different time scales. 
Based on model developed by Painter et al. 
(2002 [DIRS 174071]) and extended twoand 
three-site model developed by Wittman 
et al. (2005 [DIRS 174895]) 
This ACM is screened out for the following 
reasons: 
 Slow kinetic sorption rates are difficult to 
determine in laboratory tests that have 
been run for <150 days. 
 The lack of sufficient sorption data on 
colloids for sufficient time periods causes 
uncertainty in the extrapolation beyond the 
experimental times. 
Rate of Colloid 
Generation 
Method for using markers for waste form 
corrosion for estimating colloid production. 
This ACM is screened out for the following 
reasons: 
 The model may overestimate colloid 
formation by not accounting for significant 
retention of actinides to the immobilized 
phase as has been observed in waste 
form corrosion experiments. 
Mechanisms 
of Colloid 
Generation 
Mobile colloid generation will require large 
perturbations to the system that may not 
occur within the waste package. 
Furthermore, attachment of particles to airwater 
interfaces may limit mobile colloid 
generation. 
This ACM is screened out for the following 
reasons: 
 Data for Yucca Mountain related 
conditions are unavailable. 
 The process of attachment to air-water 
interfaces will depend on the degree of 
particle hydrophobicity. Smectite clays 
and iron oxyhydroxides in general will not 
attach to air-water interfaces but various 
oxides and U(VI) oxides phases may. 
 For flow rates similar to those anticipated 
in the repository waste package 
environment, there is no literature 
evidence to indicate mobile colloid 
generation. 
NOTE: CSNF = commercial spent nuclear fuel. 

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MDL-EBS-PA-000004 REV 02 6-54 August 2005 
6.4.1 Reversible U(VI) Phase Colloids Derived from Spent Nuclear Fuel Corrosion 
PurposeThe colloids from spent nuclear fuel are excluded from the base-case colloid model 
and abstraction based on results from unsaturated drip testing on spent nuclear fuel between 5 
and 7 years where few colloids were observed (CRWMS 2001 [DIRS 153933], Mertz et al. 2003 
[DIRS 162032]). However, use of these tests results for a generally applicable model is 
uncertain for the following reasons: the amount of waste form reaction and the role of metallic 
test components in filtering particles are uncertain, and the duration and nature of water contact 
that might limit colloid and radionuclide transport in these tests are also uncertain. In contrast, 
the unsaturated drip tests on UO2 monoliths did demonstrate the intermittent occurrence of 
uranium-bearing colloids over a period of several years (Wronkiewicz et al. (1996 [DIRS 
102047]). The unsaturated drip tests on UO2 were similar in design to unsaturated drip tests on 
DHLW glass in terms of the physical barriers for material transport. The DHLW glass drip tests 
also exhibited intermittent release of suspended particles (CRWMS 2000 [DIRS 154071]; 
Fortner et al. [DIRS 114452]). 
Although the colloid concentrations from the unsaturated drip tests on spent nuclear fuel were 
extremely low, evidence from examination of the corroded solids and the Zircaloy retainer 
indicated the presence of fine particulates (Buck et al. 1997 [DIRS 112904], Finn et al. 1998 
[DIRS 100392]). Fully saturated immersion tests by Wilson (1990 [DIRS 121808]) and 
McNamara et al. (2005 [DIRS 174068]) have shown evidence of plutonium- and americiumbearing 
colloids, and resulted in effective plutonium concentrations between 10-7 M and 10-6 M 
for the duration of these tests. 
The nature of spent nuclear fuel-derived colloids is uncertain but a conceptual model can be 
developed from the various tests that have been conducted on spent nuclear fuel and using the 
same principles that have been used to establish the DHLW glass-derived colloid model. This 
alternative conceptual model assumes that U(VI) secondary phases will act as carriers for sorbed 
plutonium and americium. The model requires knowledge of the stability range of U(VI) 
colloidal phase suspensions and the sorption coefficients for plutonium and americium on these 
phases. 
Carrier Colloidal Phase for Spent Nuclear Fuel-Derived Colloids-The reason for considering 
carrier U(VI) phases as the particle to model for spent nuclear fuel-derived colloids is based on 
observations from corrosion tests and considering which particle has both a high surface area and 
the potential to sorb radionuclides or radionuclide-bearing precipitates. Colloids identified in 
intermittent spent nuclear fuel drip tests by Finn et al. (1994 [DIRS 100746]) include schoepite 
and soddyite. Similar phases have also been reported in DSNF corrosion tests by Mertz et al. 
(2000 [DIRS 162161]). Some colloids also contained rare earths (Figure 6-14). In other test 
samplings, clay colloids derived from the equilibrated J-13 well water leachate were observed 
with trace amounts of uranium. The presence of clay suggests evaporation in the leachate during 
the course of the test; however, under conditions of igneous intrusion, silicate-bearing particles 
could be considered. 
Wilson (1990 [DIRS 100793]), observed acicular particles of uranophane 
{Ca(UO2)2(SiO3OH)2(H2O)5} on filters collected from corrosion tests on spent nuclear fuel with 
J-13 well water. In static tests on CSNF deionized water, McNamara et al. (2005 
[DIRS 174068]) reported the occurrence of suspended particles of metastudtite. Buck et al. 

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MDL-EBS-PA-000004 REV 02 6-55 August 2005 
(2004 [DIRS 171479]) reported the presence of meta-schoepite attached to larger UO2 particles 
in immersion tests on crushed and sieved UO2. Any U(VI) oxide hydrate phase that forms 
initially from the reaction of spent nuclear fuel and water would be anticipated to be 
thermodynamically unstable with respect to a silica-bearing groundwater. U(VI) silicates will 
form if the spent nuclear fuel is contacted with a silica-bearing solution and these phases may be 
the most likely to form colloids. Uranophane could be thermodynamically stable in Yucca 
Mountain groundwater environment (Finch and Ewing 1992 [DIRS 113030]) but colloidal 
uranophane transport remains uncertain. Colloidal uranophane particles may be subject to 
dissolution. Mertz et al. 2003 [DIRS 162032] examined the stability of meta-schoepite and 
UO2+x colloid suspensions after introduction to sodium silicate-bearing (J-13) groundwater and 
obtained a pHpzc ranging between 2 and 5. The large range for the point of zero charge is 
uncharacteristic for a colloidal system. Schindler et al. (2004 [DIRS 174507]) calculated the 
pHpzc values of faces and edges of schoepite and dehydrated schoepite using a rigorous valence 
bond approach that took account of the structure of these phases. The (001) face of schoepite 
and dehydrated schoepite were shown to have pHpzc value of 7.02 and 6.1, respectively and the 
edges of these uranyl oxide hydrate minerals were shown to have pHpzcs between 6.5 and 7.1. 
There was a difference in these two studies suggesting that other processes may have 
complicated the results of Mertz et al. (2003 [DIRS 162032]). Mertz et al. 2003 [DIRS 162032] 
observed a reduction in colloid size and 84% of the uranium was found to be in the dissolved 
fraction; the absence of precipitated U(VI) silicates and the high proportion of dissolved uranium 
concentration may suggest that the solution was undersaturated with respect to both schoepite 
and uranophane. One important consideration in these studies is the morphology of U(VI) 
phases formed which depends strongly on pH, presence of other mineral surfaces, and 
temperature (Schindler et al. 2004 [DIRS 174507]). Morphology will also affect dissolution of 
the phases and their long-term behavior. 
The observations from CSNF corrosion tests suggest that U(VI) phases might be considered to 
be carrier colloids analogous to smectite-type phases from corroded DHLW glass. Data on the 
colloidal suspension stability (zeta potential with pH) of schoepite type phases reported by 
Schindler et al. (2004 [DIRS 174507] has been used to develop this alternative model. In 
Figure 6-15a, a stability plot for U(VI) colloidal phases (schoepite) has been developed. 
Possible Reversible Association of Plutonium and Americium with U(VI) Phases 
McNamara et al. (2005 [DIRS 174068]) investigated the formation of colloids from samples of a 
low-burnup commercial spent nuclear fuel immersed in deionized water. H.B. Robinson CSNF 
was crushed and sieved into a 100-m to 500-m size range and 1-g samples were reacted 
in 8 ml of deionized water for 2 years at 25C, 60C, 75C, and 90C. Suspended colloidal 
aggregates were observed to form at the air-water interface for each of the 5 samples. Results 
from the analysis of the suspended solutions and leachate from a 1.5-year test demonstrated 
CSNF-derived waste form colloid formation with uptake of plutonium, neptunium, americium, 
and curium, as well as strontium and cesium. The plutonium and americium colloid 
concentrations were 1.15  10-7 M and 1.04  10-9 M, respectively. By dividing the solution 
concentration by the total mass of solid uranium metastudtite (UO4  2H2O) (mwt = 338.06 g) 
these correspond to effective Kd (ml/g) values for sorption onto metastudtite of 6.4  105 and 
1.12  104 ml/g, respectively (see Table 6-11a). Although, the Kd values for plutonium and 
americium sorption onto metastudtite were high, only a small fraction of the total available for 

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MDL-EBS-PA-000004 REV 02 6-56 August 2005 
sorption was associated with the colloids, and most remained on the parent-corroding immobile 
solid. The total uranium filterable fraction amounted to 80 mg/L (80 ppm). There is insufficient 
information to argue whether there is evidence for irreversible sorption. It is likely, as with other 
plutonium and americium sorption studies that sorption may be slowly reversible. 
Table 6-11a. Radionuclide Concentrations Associated with Colloids from Immersion Tests at 25C on 
Spent Nuclear Fuel 
Element Total 
Uranium) 
237Np 
239Pu 
241Am 
244Cm 
Radionuclide Mass (g)/ g 
Uranium 
623 g 330 352 3.23 0.58 
Radionuclide Inventory ((g)/ 
g Uranium 
389 4,689 1,423 9.66 
Uptake into UOH Suspended 
Particles (%) 
85 7.5 0.2 6 
237Np 239Pu 241Am 243+244Cm 
Dissolved Concentration in 
Leachate (mol/L) 
1.44  10-6 1.4  10-9 2.3  10-9 1.2  10-9 2.3  10-11 
Sorption (Kd) (ml/g) NA 9.95  105 6.4  105 1.12  104 1.03  105 
Error in Sorption 
Measurement 
6.26  106 1.9  104 3.27  103 4.49  103 
Source: McNamara et al. (2005 [DIRS 174068]) Table 1, p. 171 ; Table 2, p. 172. 
NOTES: Solutions were at pH 6.3. The solution aliquots were analyzed at 1.5 years and the solids were 
examined at 2 years. The reported value for 237Np in the leachate has a 629% standard deviation 
reported, leading to a large uncertainty in the calculated Kd. The solid analysis by McNamara et al. 
(2005 [DIRS 174068]) provided a combined value for 243+244Cm while the leachate analysis listed only 
244Cm. This makes that the calculated Kd for curium questionable. 
NA = not applicable 

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MDL-EBS-PA-000004 REV 02 6-57 August 2005 
Source: Adapted from pHpzc Calculations reported by Schindler et al. (2004 [DIRS 174507], pp. 1,642 
through 1,643). 
NOTE: The ionic-strength dependence is unknown, but Schindler et al. (2004 [DIRS 174507]) reported stable 
suspensions in 1.0 M NaCl and the pH dependence of zeta potential with different types of U(VI) phases 
that may exhibit variable basal plane to edge surfaces areas. Particles have been assumed to be unstable 
above an ionic strength of 1.0 M. 
Figure 6-15a. Schematic Representation of Schoepite Colloid Suspension Stability as a Function of pH. 
For the reversible case, the Kd ranges for plutonium and americium sorption coefficient ranges 
onto U(VI) phases have been expanded in each direction from the value obtained experimentally 
to account for uncertainty (see Table 6-11b): 
Table 6-11b. Developed Sorption Coefficient Distributions for Plutonium and Americium Sorption onto 
U(VI) Colloidal Phases 
Min 
(ml/g) 
Max 
(ml/g) 
Plutonium 1  104 1  106 
Americium 1  103 1  105 
Because the surface charge on layered phases depends on the proportion of edge to surface 
(Tombcz and Szekeres 2004 [DIRS 174073]), the surface potential and, hence, the suspension 
stability will vary depending on the growth pattern of the phase. For example, the precipitation 
of schoepite from a uranyl nitrate solution versus slow corrosion of UO2 result in different 
morphologies and possibly different surface and sorption characteristics. The stability of U(VI) 
silicate and peroxide colloidal suspensions is unknown, but these phases are likely to exhibit 
negative surface charges over a range of pH. The stability of these colloid suspensions as a 
function of ionic strength and pH can be estimated based on surface charge density and valence 
deficiency approach as discussed by Schindler et al. (2004 [DIRS 174507]). 

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Alternative Model Description and Data NeedsThe model requires additional data on 
sorption behavior of different U(VI) phases with pH, ionic strength, time and concentration and 
the stability trends with pH and ionic strength of various U(VI) phases that may result following 
an igneous event. Long-term sorption kinetics could be interfaced with the alternative 
conceptual model for sorption described by Wittman et al. (2005 [DIRS 174895]). 
Assessment of Alternative ModelThe alternative model would provide a source term for 
colloids from spent nuclear fuel. The data available from spent nuclear fuel tests comparable to 
those used to establish colloid concentration from DHLW glass demonstrate the propensity for 
colloid generation from spent nuclear fuel; however, the total concentration of plutonium and 
americium associated with U(VI) colloids may be low. In the scenarios where groundwater 
colloids enter the waste form region, the reversible uptake of radionuclides by groundwater 
colloids may bound that of the U(VI) colloids. The colloidal phases may be stable with respect 
to dissolution in Yucca Mountain groundwaters but might be anticipated to undergo some form 
of ripening process over time that would result in the destabilization of their suspensions. 
Particles may also be more susceptible to filtration either by thin films or pores than clay 
colloids. The sorption characteristics for these phases are uncertain and there are few data on 
the association of plutonium and americium with U(VI) colloids. 
6.4.2 Irreversible Plutonium and Americium Colloids from Corroded Spent Nuclear 
Fuel 
Detailed analysis of corroded spent nuclear fuel reacted under moist-air conditions revealed a 
precipitate with high concentrations of plutonium and americium (Buck et al. 2004 
[DIRS 172668]). These regions were patchy in occurrence but appeared to contain colloidal 
sized particles. These immobilized colloidal aggregates could become mobilized aqueous 
colloids under favorable conditions. These types of particles might be considered irreversible 
colloids similar to those observed in DHLW glass tests. The long-term dissolution kinetics of 
these phases is unknown; however, results from the report on thorium oxide colloids by 
Bitea et al. (2003 [DIRS 173041]) suggest that these plutonium- and americium-oxide phases 
might be expected to dissolve once exposed to dilute solutions. 
The ionic-strength dependence of plutonium-rind irreversible colloids is based on the values used 
for other colloids, 0.05 M. The pH dependence is based on the results of Parks (1965 
[DIRS 174361] for a zero point of charge (pHpzc) of 8.6 to 9.0 similar to that observed for 
zirconium oxide (Bitea et al. 2003 [DIRS 174504]). The ionic-strength dependence was based 
on studies on rare earth oxides by Hsu et al. (1988 [DIRS 174350]) where 0.6-M sulfate 
solutions caused the destabilization of cerium oxide colloidal suspensions and studies of 
zirconium oxide colloids from Bitea et al. (2003 [DIRS 174504]) (see Figure 6-15b). Colloidal 
ZrO2 is a reasonable analogue for PuO2 because of its low solubility and chemical inertness; 
whereas, Ce(IV) oxides can be strong oxidizers. Using this information, a plot showing particle 
suspension stability with pH and ionic strength is presented in Figure 6-15c. The concentration 
of plutonium-bearing irreversible colloids is difficult to establish from corrosion tests, as it is not 
possible to know whether the plutonium is sorbed to another colloid or present as a discrete 
particle. Applicable data for developing plutonium colloidal concentration comes from the data 
from Finn et al. (1994 [DIRS 100746]) where actinide-bearing particles were observed at early 
times during the course of the unsaturated drip tests. Some of these were described as rare earth

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MDL-EBS-PA-000004 REV 02 6-59 August 2005 
rich particles that are similar to the actinide-bearing particles described in this alternative 
conceptual model. Early release of uranium and plutonium from ATM-106 fuel tests (Finn et al. 
(1994 [DIRS 100746], Table 4, p. 192) show the following: 5  10-5 g/L uranium, 3  10-7 g/L 
plutonium, and 1  10-7 g/L americium. The results from the ATM-103 tests were 4  10-8 g/L 
for plutonium and 3  10-6 g/L for americium. The filterable uranium content in both tests was 
0.05 mg/L. 
Source: Extracted from Bitea et al. 2003 [DIRS 174504], p. 61 Figure 6. 
NOTE: The pHpzc value for ZrO2 is close to 8.1; whereas, the pHpzc value for PuO2 is between 8.6 and 9.0, making 
ZrO2 a reasonable analogue for PuO2 colloid behavior. 
Figure 6-15b. Stability Ratio for ZrO2 Colloids (0.4 g/l) for Ionic Strength of 0.1 M, 0.001 M, and 0.001 M 
NaCl 
Source : Parks 1965 [DIRS 174361], Table 1, p. 180 ; Bitea et al. 2003 [DIRS 174504] ; Hsu et al. 1988 
[DIRS 174350], 
Figure 6-15c. Stability of PuO2+x Colloid Suspension with Ionic Strength and pH 

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MDL-EBS-PA-000004 REV 02 6-60 August 2005 
By using the uranium colloid concentrations reported by McNamara et al. (2005 [DIRS 174068]) 
and Finn et al. (1994 [DIRS 100746]) the total filterable uranium content in parts per million 
colloids can be determined. Table 6-11c shows the relationships that may be established for 
irreversible association of plutonium and americium: 
Table 6-11c. Possible Irreversible Colloid Concentrations Associated CSNF Corrosion 
McNamara et al. (2005 
[DIRS 174068]) 
ATM-101 
Finn et al. (1994 
[DIRS 100746]) 
ATM-106 
Finn et al. (1994 
[DIRS 100746]) 
ATM-103 
1 ppm 
colloids 
= 1  10-9 mol/L Pucoll 1  10-9 mol/L Pucoll 3  10-9 mol/L Pucoll 
= 1  10-11 mol/L Amcoll 8  10-9 mol/L Amcoll 2  10-7 mol/L Amcoll 
Notes: Errors in the radiochemical data from McNamara et al. 2005 [DIRS 174068] are described 
above. Measurements provided by Finn et al. 1994 [DIRS 100746] were obtained by mass 
spectrometry, which have a 50% error (Mertz et al. 2003 [DIRS 162032]). 
Alternative Model Description and Data NeedsThe model requires additional data on 
colloid release from fully corroded spent nuclear fuel, assessment of colloid concentrations, and 
actinide concentrations. 
Assessment of Alternative ModelThe model can be used to describe potential plutonium 
colloid generation from the corroded surface of spent nuclear fuel as well as polygonized rim 
material from high burn-up fuels. The phases that can form at high reactor operating 
temperatures, such as perovskites, might be expected to be kinetically stable under repository 
time frames. 
The model has not been included in TSPA owing to the immaturity of the model and lack of 
suitable data on colloid populations including the stability of the particle suspensions with ionic 
strength and pH, and concentration of associated plutonium and americium. 
6.4.3 Kinetic Sorption Model-An Alternative to Irreversible Attachment 
Painter et al. (2002 [DIRS 174071]) presented a two-site sorption kinetics model that allows 
evaluation of sorption data; in particular, the model permits analysis of adsorption on long time 
scales as compared to times for experimentally measured adsorption. The two site model is 
probably the simplest dynamical linear model that can represent both a fast adsorption (or 
desorption) for a near-term equilibrium distribution [(Kc)fast] and a delayed adsorption for a long 
term (Kc)slow that may be much larger. Both reversible and irreversible colloid sorption can be 
represented by the full analytical solution to the rate equations for a flow-through or batch 
reactor. The solutions are offered as guidance for interpreting sorption data and for predicting 
the long-term discharge concentrations through an aquifer. This model was applied by 
Wittman et al. (2005 [DIRS 174895]) to plutonium sorption data reported for smectite, 
montmorillonite, and hematite obtained from Lu et al. (1998 [DIRS 100946]; 2000 
[DIRS 166315]). In Figure 6-16a, three site simulations of sorption data are shown that 
demonstrate the importance of a slowly reversible nonzero desorption term, kR3 = 0.0017 h-1 (red 
lines) versus kR3 = 0 (blue lines). Figure 6-16b shows the fit of the three-site model to the data. 

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MDL-EBS-PA-000004 REV 02 6-61 August 2005 
Painter et al. (2002 [DIRS 174071]) demonstrated a simple way to gauge the impact of the 
reversible rate constant,  (called kRi by Wittman et al. 2005 [DIRS 174895]). The qualitative 
behavior of the aquifer can be represented as a flow-through reactor. Figure 6-16a shows the 
long-term presence of mobile contaminant (dimensionless Pu discharge) in a system as a 
function of time for colloid sorption assuming a possible mean flow-through residence time of 
4,000 years by Wittman et al. (2005 [DIRS 174895]). Xo is the fraction of Pu in solution, Xi are 
the fractions of Pu present in i independent forms on the mobile colloid. At t = 0 the 
contaminant is assumed to be entirely distributed between the liquid and the immobile surface. 
In the case of irreversible sorption, adsorption onto the colloid continues to increase the mobile 
fraction of the contaminant until the source contribution from the immobile phase and liquid 
(Xo) drops to low concentrations from dilution. For reversible sorption, adsorption onto the 
colloid reaches an equilibrium maximum long before dilution effects purge the contaminant from 
the system. A greater fraction remains in the immobile phase during most of the flow history. 
The irreversible case of Figure 6-16a assumes that kR3 is zero; in the reversible case  is assumed 
to be the value obtained from fitting the colloid sorption data (red lines). Even though the total 
mobile concentration would possibly be low in both reversible and irreversible cases, the peak 
concentration for the reversible case is about 1,000 times lower. Additionally, in the reversible 
case, the equilibrium behavior over most of the flow history supports the use of a distribution 
coefficient Kd (called Kc for colloids) for representing sorption in colloid assisted transport 
models. For the two-site model by Painter et al. (2002 [DIRS 174071]) Kc is the sum of Kc(fast) 
and Kc(slow) representing equilibrium for fast and slow sorption processes, respectively. 
(i) 
Source: Wittman et al. 2005 [DIRS 174895], Figure 5.1 p. 5.2. 
Notes: Data were used to fit data from Lu et al. (2000 [DIRS 166315]). (i) With irreversible sorption (kR3 = 0), 
plutonium association with colloids is only reduced through 239Pu decay; however, with a small reversible 
factor, the peak dose for plutonium-sorbed colloids decreases. 
Figure 6-16a. Kinetic Sorption Model Predictions for Plutonium Sorption onto Colloids. 
Exact analytical solutions of the minimal sorption kinetics equations were used by Wittman et al. 
(2005 [DIRS 174895]) in the analysis. Analytical solution offered a huge advantage in 
computing speed because the model could be evaluated at an arbitrary time-step. Best fits to the 

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MDL-EBS-PA-000004 REV 02 6-62 August 2005 
data were obtained by performing a least squares fit for the rate constants. A multidimensional 
secant method was used to search for a local least squares fit with resulting rate constants 
allowing the determination of distribution coefficients (Wittman et al. 2005 [DIRS 174895]). 
Source: Wittman 2005 [DIRS 174895], Figure 4.3, p. 4.4. 
NOTE: Model was used for Fitting Data Reported for Plutonium Sorption onto Iron Oxide with Exponential Fit used 
to Derive Rate Constant from Table 6-9. 
Figure 6-16b. Three-Site Sorption Model 
Painter et al. (2002 [DIRS 174071]) show that uncertainty in the value of the reverse rate cause 
large uncertainty in the peak plutonium and americium discharge because no sorption experiment 
has been run for sufficient time to allow an accurate determination of the reversible process and 
some extrapolation is necessary. In Table 6-11d, the developed sorption coefficients for clay and 
hematite colloids are listed. In both cases, the developed values are at the low range for the 
distributions in Table 6-6. 
Table 6-11d. Plutonium Sorption Coefficients Developed from Kinetic Sorption Model 
Colloid 
Concentration 
Sorption Coefficients, Kc(ml/g) 
Mineral Cc (mg/L) Kc1 Kc2 Kc3 Kc(total) 
Hematite 200 9.16  103 1.245  105 0 1.336  105 
Smectite 1,000 1.046  103 2.81  103 1.04  104 1.43  104 
Montmorillonite 200 2.0  102 7.96  103 0 8.16  103 
Source: Wittman et al. (2005 [DIRS 174895]), Table 5.1, p. 5.1 with units converted from L/g to ml/g. 
Alternative Model Description and Data NeedsImportant data requirements for this model 
are longer-term sorption experiments and improved characterization of the surface area, 
morphology, and composition of the starting materials. For example, the sorption results reported 
by Lu et al. (1998 [DIRS 100946] for smectite and the results reported for montmorillonite are 
vastly different despite few differences in the nature of these phases. This suggests that small 

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MDL-EBS-PA-000004 REV 02 6-63 August 2005 
differences in the composition and morphology of these two materials led to large differences in 
the sorption characteristics. Painter et al. (2002 [DIRS 174071]) has suggested the need for a 
better understanding of the microscopic exchange mechanisms of plutonium on colloids at 
different time scales. 
Assessment of Alternative ModelThe kinetic model represents a useful tool for examining 
sorption data and examining the evidence for slow reversible sorption of plutonium on colloids. 
The model clearly demonstrates that the data from Lu et al. (2000 [DIRS 166315]) can be fit 
with reversible components suggesting that irreversible sorption is conservative. The kinetic 
model has not been adopted because it is immature and the base model is conservative and 
bounding. 
6.4.4 Rate of Colloid Generation Model 
The following subsections describe an approach to estimating the rate of release of colloidal 
plutonium from degraded waste, based on the rate of waste degradation. This modeling 
approach is preliminary and is provided here as an alternative model for prediction of colloid 
release and transport in the waste package and drift and as a basis for potential subsequent 
discussion and project planning. It is based on a colloid generation rate model presented in 
Colloid-Associated Radionuclide Concentration Limits: ANL (CRWMS M&O 2001 
[DIRS 154071]. 
Rate of Colloid Formation from DHLW GlassThe normalized mass loss of plutonium, 
NL(Pu), was compared to the normalized mass loss of boron, NL(B), to indicate the amount of 
plutonium released relative to the amount of B released, which is directly proportional to the 
amount of DHLW glass corroded. The NL(Pu) was consistently lower than the NL(B) by at 
least one order of magnitude. NL(Pu)/NL(B) decreases for the longer duration tests at higher 
SA/V, up to about 10-3 for stable colloids-inhibiting conditions. 
Under low-ionic strength conditions, the colloid suspensions are stable in the fluid and the rate of 
formation of plutonium-bearing colloids, dmPu-coll/dt, is proportional to the amount of altered 
glass, Malt. As the glass alters, the rate of radionuclide-bearing colloids production decreases, 
initially following a power law, attributed to their sorption by the fixed clay alteration layer 
(CRWMS M&O 2001 [DIRS 154071], Section 6.4.2). The rate of colloid formation can be 
described by the relationship in Equation 10: 
( )b 
alt 
Pucoll M a 
dt 
dm - = (Eq. 10) 
The constants, a and b, derived for the various glasses may be a function of the SA/V, leachate 
composition (ionic strength, etc.), colloid composition, or other parameters. Table 6-11e 
summarizes the ranges and bounding values for a and b. 

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Table 6-11e. Ranges and Bounds for Constants, a and b, in Equation 10 
Static-Saturated Tests Drip Tests 
Constant Low High Low High 
Bounding 
Values 
a 8  10-12 4  10-6 2  10-6 5  10-6 10-4 
b 3 8 0.3 1.7 1 
Source: CRWMS M&O 2001 [DIRS 154071], Section 6.4.2. 
At the high-ionic strengths encountered in certain static tests, there was no correlation observed 
between plutonium and B production rates since the quantity of colloids was low. In the drip 
tests, ionic strength remained low because EJ-13 well water was periodically injected. 
Once the cumulative B release is greater than approximately 1 to 3 g m-2, the radionuclide 
release is controlled by spallation of the clay layer. Thus, radionuclide release is a two-step 
process: (1) alteration of the glass waste form and precipitation of colloids, and (2) erosion of 
colloid-sized fragments from the alteration products. 
Release of colloids due to spallation is given by the derived relationship: 
alt 
Pu M 
dt 
dm coll . = (Eq. 11) 
where coll Pu dm /dt is the mass release of plutonium suspended per time as colloidal particulates 
and . is a parameter that depends on, among other things, the mechanical condition of the altered 
waste form, the water flow rate over the waste form, and the water chemistry (CRWMS M&O 
2001 [DIRS 154071], Section 6.4.2). An empirical bound for . is 10-4 d-1. 
Calculation of Quantity and Rate of Production of Irreversibly Attached Radionuclides 
Parameters used to calculate mobilized radionuclide quantities and rates of production are listed 
in Table 6-11b. After the quantities of total mobilized radionuclides are calculated, they are 
distributed between radionuclides irreversibly attached to colloids, dissolved (aqueous) 
radionuclides, and radionuclides coprecipitated in immobile secondary phases. Since the 
experimental evidence indicates that most waste form colloids are radionuclides, radionuclidebearing 
phases, or both, irreversibly associated with smectite colloids, this fraction is given as 
the difference between the dissolved radionuclides and the total radionuclides mobilized by the 
waste degradation time step: 
RNcoll,irrev + RNlayer,irrev = RNmobRNdissRNsecondaryphase (Eq. 12) 
where RN = radionuclide. 
The irreversibly attached radionuclides are in turn distributed between the clay layers on altered 
DHLW glass and clay colloids produced from DHLW glass and spent nuclear fuel; the aqueous 
radionuclides are available for reversible sorption onto colloids. 

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If the cumulative release of boron is less than 1 g m-2, then the following relationship is invoked 
to calculate the quantity of radionuclides irreversibly attached to colloids produced for that time 
step. 
The equation for calculating the mass of radionuclides is: 
Mass of irreversibly attached RN = a * (Bcum)-b (Eq. 13) 
where Bcum represents the mass of DHLW glass degraded and a and b are empirically derived 
constants, and RN = radionuclide. 
Table 6-11f. Parameters Used in Description of Alternative Model 
Parameter Description Value /Range/ 
Distribution Source 
Binv, Tcinv, 
Puinv, Aminv 
Masses of B and radionuclides in 
unaltered waste form during time step 
 Would be calculated by TSPA-LA 
model 
Bmob, Tcmob, 
Pumob, Ammob 
Masses of B and radionuclides 
mobilized in WFdeg during time step 
 Would be calculated by the 
TSPA-LA model 
Bcum Cumulative mass of B mobilized in 
WFdeg for all previous time steps 
 Would be calculated by the 
TSPA-LA model 
Malt Amount of a altered glass  CRWMS M&O 2001 [DIRS 154071] 
a Empirically derived constant used in 
expression for precipitated smectite 
colloids 
10-4 (bound) CRWMS M&O 2001 [DIRS 154071] 
b Empirically derived constant used in 
expression for precipitated smectite 
colloids 
1 (bound) CRWMS M&O 2001 [DIRS 154071] 
. Empirically derived constant used in 
expression for spalled smectite colloids 
10-4 (bound) CRWMS M&O 2001 [DIRS 154071] 
NOTE: WF = waste form. 
If Bcum is greater than 3 g m-2, then the second model is invoked to calculate the quantity of 
radionuclides irreversibly attached to colloids formed from spallation of the alteration clay layer. 
The equation for calculating radionuclide mass mobilized due to spallation is: 
Mass of irreversibly attached spalled RN = . * Bcum (Eq. 14) 
where 
. is an empirically derived constant 
RN = radionuclide 
If Bcum is between 1 and 3, then either the first or second is chosen randomly (0.5 
probability). 
Assessment of Rate-of-Colloid-Generation ModelThis alternative abstraction takes into 
account many of the results from waste form corrosion experimental program as well as models 
derived from them. It incorporates an analytical approach that would couple the waste form 

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MDL-EBS-PA-000004 REV 02 6-66 August 2005 
degradation calculations by the TSPA-LA model more closely to the process model derived from 
the experiments (CRWMS M&O 2001 [DIRS 154071]). 
While it takes advantage of the most complete information available on waste form corrosion, at 
the same time it is tailored closely to the experimental conditions and configurations. There are 
aspects of the experiments that may make them inappropriate to apply to a more general case. A 
possible example is the rate of spallation of the clay layer from altered DHLW glass, which may 
be due to the specific dynamics surrounding the impacts of the drips falling on the sample in the 
DHLW glass drip tests (CRWMS M&O 2001 [DIRS 154071], Section 6.2.1). If the corroded 
contents of a failed waste canister assume a geometry that does not allow dripping of water onto 
the alteration products, spallation may not occur. If dripping does occur, of course, spallation 
may be a viable process, and an analysis of colloid production and transport would have to 
consider it. 
6.4.5 Mechanisms of Colloid Generation 
This alternative model, described by Buck (2004 [DIRS 171479]), investigates the processes 
whereby colloid generation at, and mobilization from, the surface of degrading waste is primarily 
related to flow rate at the waste surface and the attachment of colloids to airwater interfaces. 
The supporting concepts and data from the peer-reviewed literature that support this alternative 
model were developed in the context of deposition and remobilization of existing colloids under 
conditions of moderate to high fluid flow. The model suggests that unless high flows are present, 
mobile colloid generation would be negligible owing to the strong and irreversible attachment of 
hydrophobic colloids to air bubbles on the surface of waste package components. 
Under specific experimental environments, a flow rate dependence on colloid release has been 
observed (Buck 2004 [DIRS 171479]), although a full explanation for these empirical 
relationships has not yet been developed. For flow rates similar to those anticipated in the 
repository waste package environment, there is no literature evidence to indicate mobile colloid 
generation. In this alternative model, flow-rate dependence of the entrainment and mobilization 
of waste form colloids would be determined. In addition, the influences of fluid chemistry and 
of the actual phases formed and their physical association in a model waste form system would 
be investigated. This model will allow quantification of changes in particle release under 
different flow rates, fluid chemistries, and particleparticle interaction scenarios. 
In tests performed by Buck et al. (2004 [DIRS 171479]), metaschoepite formed during the 
weathering of UO2 powder was demonstrated to be hydrophobic. The particles were stabilized at 
polar-nonpolar interfaces (i.e., air-water or oil-water interfaces), similar to observations in many 
other mineral oxide colloidal systems where hydrophobic colloids become irreversibly attached 
to air bubbles (Abdel-Fattah et al. 1998 [DIRS 158692]). 
Assessment of the Colloid Generation Mechanisms Model For flow rates similar to those 
anticipated in the repository waste package environment, there is no literature evidence to 
indicate mobile colloid generation due to water flow. Omission of colloid partitioning to the air 
water interface is conservative because this effect would likely curtail colloid transport. 

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6.5 MODEL FORMULATION FOR BASE-CASE ASSESSMENT 
This section describes in detail the methodology used to incorporate the various model elements 
formulated in this model analysis into a colloids source term abstraction (i.e., a simplified model 
intended to retain the important principles and features of the analyses) for implementation in the 
TSPA-LA model. The logic and parameters implemented in the TSPA-LA model are provided, 
although specific programming details are not. 
Figures 6-17 and 6-18a provide an overview of the model methodology with cross-references to 
the underlying model elements developed in Section 6. Figure 6-17 illustrates how colloids are 
modeled in the waste package (Section 6.5.1), and Figure 6-18a in the drift (Section 6.5.2). 
Parameters used in the abstraction are described in Table 6-12. Sections 6.5.1 and 6.5.2 present 
the details of the abstraction, and Figures 6-16 through 6-19 illustrate the logic and flow of the 
abstraction approach. 
At each time step executed in the TSPA-LA calculations, the colloid model uses characteristics 
of the water in the waste package (ionic strength, pH, and dissolved radionuclide concentrations) 
to describe the formation and behavior of colloids and the distribution of the radionuclides in the 
waste package. The elements considered and incorporated into the calculations are plutonium, 
americium, protactinium, cesium, and thorium. Three types of colloids are considered in the 
modelwaste form, corrosion product iron oxidesoxyhydroxide, and seepage water 
groundwater. Waste form colloids generated from DHLW glass are included in the model, but 
colloids from commercial and DSNF are not. Colloids formed from the DHLW glass and the 
naturally occurring groundwater colloids are modeled as smectite (phyllosilicate) colloids. The 
iron oxyhydroxide colloids are formed during corrosion of steel waste package materials. 
The stability of colloid suspensions is defined by ionic strength and pH, with separate 
relationships for smectite colloids (Section 6.3.2.1, Figures 6-17 and 6-18a, Table 6-12) and iron 
oxyhydroxide colloids (Section 6.3.2.2, Figures 6-17 and 6-18a, Table 6-12). At each time step, 
the values of ionic strength and pH of the waste package solution are used as input to determine 
whether or not colloids are stable. For both types of colloids, when unstable, a low (nonzero) 
limit of colloid concentration is used (Figures 6-17 and 6-18a, and Table 6-12 for smectite; 
Table 6-12 for iron oxyhydroxide). Within the stability zone for DHLW glass-smectite colloid 
suspensions, a function is used that relates plutonium and colloid concentrations to ionic strength 
(Figure 6-9). For iron oxyhydroxide colloid suspensions, upper and lower limits of ionic 
strength are used in the definition of stability (Figures 6-4, 6-17, and 6-18a; Table 6-12); for 
groundwater smectite colloids, an upper limit of colloid concentration is used (Figures 6-12, 6- 
17, and 6-18a; Table 6-12). 
Reversible sorption onto the smectite colloids for each of the five considered radionuclides is 
modeled (Section 6.3.3.1). For iron oxyhydroxide colloids, irreversible sorption of plutonium 
and americium is also included (Section 6.3.3.2). Irreversibly attached (or embedded) plutonium 
and americium on DHLW glass-derived colloids are also included (Section 6.3.3.3). Within the 
waste package, the major fractions of total plutonium and americium available at each time step 
are sorbed to stationary corrosion products, with smaller fractions adsorbed to and dissolved in 
the aqueous phase. Further, a high percentage of the sorbed plutonium is irreversibly sorbed 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-68 August 2005 
onto colloids and stationary corrosion products. Distribution coefficients (Kd values) for each of 
the radionuclides are used for both smectite and iron oxyhydroxide colloids (Table 6-12). 
The model calculates the concentration of dissolved radionuclides, the concentration of 
plutonium and americium irreversibly attached (embedded) in DHLW glass smectite colloids, 
the concentration of radionuclides sorbed reversibly onto colloids, the concentration of 
radionuclides sorbed irreversibly onto iron oxyhydroxide colloids, and the mass concentration of 
each colloid type (Figure 6-17). 
The colloid and radionuclide concentration values along with the ionic strength and pH of 
solution serve as source term for the invert. Based on the values of ionic strength and pH in the 
invert, the model determines colloid stabilities at the invert fluid composition and redistributes 
the reversibly sorbed radionuclides and dissolved radionuclides based on the distribution 
coefficients and the total mass of each type of colloid (Figure 6-18a). Colloids flow unimpeded 
through the invert to the unsaturated zone. The irreversibly sorbed radionuclides are not 
redistributed. 
6.5.1 Waste Form Abstraction Implementation in TSPALicense Application 
The abstraction requires parameters calculated by the TSPA-LA model as output from 
in-package chemistry, solubility limits, and near-field chemical environment calculations: 
CRN,diss, Iwp, and pHwp (Table 6-12). 
The concentrations of colloids and colloid-associated radionuclides were calculated on the basis 
of abstractions of the laboratory and field observations and chemical principles described in 
Section 6.3. A stepwise approach was taken in the abstraction, and each step is set out below 
with a flow diagram and logic statements that describe the abstraction step. Parameters are 
defined in Table 6-12.

MDL-EBS-PA-000004 REV 02 6-69 August 2005 
Figure 6-17. Processes Within the Waste Package 
Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 

MDL-EBS-PA-000004 REV 02 6-70 August 2005 
NOTE: WP = waste package; RN = radionuclide; SW = seepage water; HLW = high-level waste; UZ = unsaturated zone. 
Figure 6-18a. Processes Within the Drift (Invert) 
Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 

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MDL-EBS-PA-000004 REV 02 6-71 August 2005 
6.5.1.1 Waste-form (smectite) colloids 
Step 1a. The radionuclides plutonium and americium are modeled in Step 1a (Figure 6-18b) as 
irreversibly sorbed within waste form colloids, which are produced from corrosion of DHLW 
glass. The concentrations are defined by ionic strength, based on experimental observations. 
Input Iwp 
IF I = Ihi-thresh,coll,wf 
THEN CRNcoll,wf.embed = CRNcoll,wf,embed,min 
ELSE 
IF (I > Ilo-thresh,coll,wf) AND (I < Ihi-thresh,coll,wf) 
THEN CRNcoll,wf,embed = function(STcoll,wf,is) 
ELSE CRNcoll,wf,embed = CRNcoll,wf,embed,max 
function(STcoll,wf,is): -2.5  10-6  I +1.25  10-7 
Figure 6-18b. Flow Chart and Logic Statements: Effect of Ionic Strength on the Concentration of Waste 
Form Colloid-Associated Radionuclides Generated DHLW Glass Degradation Tests 
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 irreversibly sorbed plutonium used in the 
abstraction. Americium is calculated as the product of the plutonium concentration times the 
americium to plutonium ratio in the inventory, which is determined at each time step of the 
TSPA-LA model calculations. Stability of the DHLW glass-derived colloid suspensions is 
determined in Step 1b. Colloid mass concentrations are calculated from the amount of 
plutonium associated with smectite waste form colloids in Step 1c. 

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MDL-EBS-PA-000004 REV 02 6-72 August 2005 
Step 1b. The stability of waste form colloid suspensions from DHLW glass is determined on the 
basis of the fluid ionic strength and pH (Figure 6-19), based on known properties of smectite 
colloidal suspensions. 
Input pHwp 
IF (Iwp = Ihi-thresh,coll,wf) OR (I = function(STcoll,wf,pH)) 
THEN CRNcoll,wf,embed = CRNcoll,wf,embed,min 
function(STcoll,wf,pH): pH/200 
Figure 6-19. Flow Chart and Logic Statements: Effect of pH and Ionic Strength on Waste Form Colloid- 
Associated Radionuclide Concentration Based on Stability Behavior of Montmorillonite 
Colloid 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 the colloids are stable. These threshold values 
may vary due to specific chemical and environmental conditions and are, therefore, sampled over 
a range (Table 6-12). 
Step 1c. Calculation of the mass concentration of waste form colloids from DHLW glass 
(Figure 6-20) is done on the basis of experimental observations. 
Mcoll,wf,both = Mcoll,wf,both,max  CRNcoll,wf,embed  CRNcoll,wf,embed,max 

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MDL-EBS-PA-000004 REV 02 6-73 August 2005 
Figure 6-20. Flow Chart and Logic Statements: Determination of Mobile Mass Concentration of Waste 
Form Colloids Mcoll,wf,both) 
The colloid mass concentration is related to the irreversibly sorbed plutonium concentration 
(Equation 1). With reference to Equation 1, Mcoll,wf,both,max is the maximum colloid mass 
concentration corresponding to the maximum plutonium concentration, CRNcoll,wf,embed,max. It is an 
intermediate result based on the experimental data showing the relationship between irreversibly 
sorbed plutonium concentration and colloid mass concentration (see DTNs in Table 6-12). 
Mcoll,wf,both is a parameter that represents the total mass concentration of DHLW glass colloids 
with both irreversibly sorbed and reversibly sorbed radionuclides (Table 6-12). 
Step 1d. The calculation of the concentration of radionuclides (plutonium, americium, 
protactinium, cesium, and thorium) reversibly sorbed on waste form colloids from DHLW glass 
(Figure 6-21) 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. 
CRNcoll,wf,rev = CRNdiss  Kd,RN,coll,wf  Mcoll,wf,both  1  10-6 
Figure 6-21. Flow Chart and Logic Statements: Sorption of Radionuclides on Waste Form Colloids 
The CRNdiss is the concentration of radionuclide RN, determined by the TSPA-LA model as 
constrained by 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. 
6.5.1.2 Corrosion-Generated (Iron Oxyhydroxide) Colloids 
Step 2a. The stability and mass concentration of iron oxyhydroxide colloid suspensions formed 
as a result of the corrosion of waste package components is determined on the basis of the fluid 
ionic strength and pH (Figures 6-17 and 6-20). 
IF 
BEGIN 
(I > Ihi-thresh,coll,FeOx) 

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OR ((I = Ilo-thresh,coll,FeOx) 
AND (I > function(STcoll,FeOx,pHhi)) OR (I > function(STcoll,FeOx,pHlo)) 
OR ((I < Ilo-thresh,coll,FeOx) AND (function(STcoll,FeOx,pHint))) 
END 
THEN 
Mcoll,FeOx = Mcoll,FeOx,min 
ELSE 
Mcoll,FeOx = Mcoll,FeOx,sampled 
function(STcoll,FeOx,pHhi): -0.02  pH + 0.17 
function(STcoll,FeOx,pHlo): +0.02  pH - 0.17 
function(STcoll,FeOx,pHint) (boolean): ((pH = 8) AND (pH = 9)) 
Figure 6-22. Flow Chart and Logic Statements: Effect of Ionic Strength and pH on Stability of Iron 
Oxyhydroxide Colloid Suspensions 
The Ihi-thresh,coll,wf,FeOx and Ilo-thresh,coll,FeOx are threshold values of the ionic strength of the fluid in 
the waste package used to calculate the stability and concentration of corrosion-generated colloid 
suspensions, similar to the waste form colloid stability determination in Step 1a. STcoll,FeOx,pHhi 
and STcoll,FeOx,pHlo are functions which incorporate threshold values of pH for determining the 
stability and concentration of corrosion-generated colloid suspensions in conjunction with the 
ionic strength thresholds. 

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Step 2b. In this step two types of sorption are modeled, irreversible sorption of plutonium and 
americium onto iron oxyhydroxide corrosion colloids and fixed corrosion products (Figure 6- 
23); and reversible sorption of plutonium, americium, protactinium, cesium, and thorium onto 
iron oxideoxyhydroxide colloids. The distribution of irreversibly sorbed plutonium and 
americium onto fixed and colloidal substrates is performed in the TSPA-LA model. It is a 
kinetic model that is described in EBS Radionuclide Transport Abstraction (BSC 2004 [DIRS 
169868]). Parameters used include specific surface area (SA), the sampled corrosion colloid 
mass concentration (Mcoll,FeOx,sampled), concentration of dissolved radionuclides (CRNdiss), and 
experimentally-based forward rate constant (k) applicable under certain scenarios such as 
unstable colloidal suspension conditions, low flow rates, etc. The experimentally based forward 
rate constant for irreversible sorption is computed as a function of flow rate, colloid stability, and 
other system parameters under other conditions (e.g., stable colloid suspension conditions and 
higher flow rates). 
Reversible sorption is calculated from the sampled mass concentration of corrosion colloids, Kd 
values describing the distribution of radionuclides between the fluid and iron oxide 
oxyhydroxide colloids, and the dissolved concentration of radionuclides as calculated by the 
TSPA-LA model. 
CRNcoll,FeOx = CRNdiss  Kd,RN,FeOx  Mcoll,FeOx  1  10-6 
NOTE: RN = radionuclide 
Figure 6-23. Flow Chart and Logic Statements: Sorption of Radionuclides on Iron Oxyhydroxide 
Colloids 
In Step 2b the irreversibly sorbed plutonium and americium are determined, along with 
reversibly sorbed plutonium, americium, protactinium, cesium, and thorium. The CRNdiss is the 
concentration of radionuclide RN, determined by the TSPA-LA model as constrained by the 
solubility concentration model, which is used as an input to the colloid abstraction. Kd,RN,FeOx is 
a parameter derived from several sources (Tables 6-5 and 6-6) and is an equilibrium sorption 
coefficient used to approximate the partitioning of dissolved radionuclide between colloids and 
fluid. The 1  10-6 portion of the relationship designates a correction factor to accommodate 
differences in units among the various parameters. 

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6.5.1.3 Seepage Water (Smectite) Colloids 
Seepage watergroundwater colloid-related input parameters used in the abstraction are listed, as 
well as information used to develop the parameters, in Table 6-12. 
Step 3a. The mass concentration of seepage water colloids is calculated on the basis of fluid 
ionic strength in Step 3a (Figure 6-24). 
IF Iwp= Ihi-thresh,coll,gw 
THEN Mcoll,gw = Mcoll,gw,min (1  10-6 mg/L) 
ELSE IF (Iwp < Ihi-thresh,coll,gw) 
THEN Mcoll,gw = Mcoll,gw,sampled 
Figure 6-24. Flow Chart and Logic Statements: Effect of Ionic Strength on Mass Concentration of 
Seepage Water Colloids 
Ihi-thresh,coll,gw is a threshold value of the ionic strength of the fluid in the waste package used to 
calculate the concentration of groundwater colloid suspensions (stability is determined in Step 
3b). It is identical to the upper threshold used for determining stability and concentration of 
waste form colloid suspensions, since both waste form and groundwater colloids are modeled as 
smectite. Groundwater colloids are modeled as present in groundwater seeping into the drift and 
as entering a breached waste package. Colloid concentrations are sampled if the ionic strength is 
below Ihi-thresh,coll,gw. This ionic strength threshold value may vary due to specific chemical and 
environmental conditions and, therefore, is sampled over a range. The groundwater colloid 
concentration range (and distribution) is an intermediate parameter derived from several sources 
containing groundwater concentration data versus ionic strength from the Yucca Mountain 
vicinity and elsewhere. 
Step 3b. The stability of groundwater colloid suspensions is determined in Step 3b on the basis 
of the fluid ionic strength and pH (Figure 6-25), based on known properties of smectite colloidal 
suspensions. 
IF (I = Ihi-thresh,coll,gw) OR (I = function(STcoll,wf,pH)) 
THEN Mcoll,gw = Mcoll,gw,min 
GoldSi

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function(STcoll,wf,pH): pH/200 
Figure 6-25. Flow Chart and Logic Statements: Effect of pH and Ionic Strength on Groundwater 
Colloids Stability Based on Stability Behavior of Montmorillonite Colloid Suspensions 
Ihi-thresh,coll,gw is the same upper threshold value of the ionic strength of the fluid in the waste 
package used to determine plutonium concentration in Step 1a. It is used here with pH to 
determine the stability of the smectite groundwater colloid suspensions. This threshold value 
may vary due to specific chemical and environmental conditions and is, therefore, sampled over 
a range (Table 6-12). 
Step 3c. The concentrations of radionuclides (plutonium, americium, protactinium, cesium, and 
thorium) reversibly sorbed on groundwater colloids are calculated in Step 3c (Figure 6-26), 
based on the mass of groundwater colloids, Kd values describing the distribution of radionuclides 
between the fluid and montmorillonite, and the dissolved concentration of radionuclides within 
the waste package as calculated by the TSPA-LA model. 
CRNcoll,gw = CRNdiss  Kd,RN,coll,gw  Mcoll,gw  1  10-6 
Figure 6-26. Flow Chart and Logic Statements: Radionuclide Sorption on Groundwater Colloids 
The CRNdiss is the concentration of radionuclide RN, determined by the TSPA-LA model as 
output from the solubility concentration model that is used as an input to the colloid abstraction. 
Kd,RNcoll,gw is a parameter derived from several sources (Tables 6-5 and 6-6) and is an equilibrium 
sorption coefficient used to approximate the partitioning of dissolved radionuclide between 
colloids and fluid. The 1  10-6 portion of the relationship designates a correction factor to 
accommodate differences in units among the various parameters. 

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MDL-EBS-PA-000004 REV 02 6-78 August 2005 
6.5.1.4 Outputs from the Waste Package to the Drift 
In Steps 1 through 3 the total colloid-associated radionuclide concentrations, total colloid mass 
concentrations and the remaining radionuclide concentrations in the fluid were calculated for 
waste form colloids derived from DHLW glass (Step 1), colloids derived from the corrosion of 
the waste package interior (Step 2), and seepage water/groundwater colloids that entered the 
waste package. In Step 4 these quantities are passed from the waste package interior to the drift 
environment. 
Step 4. Outputs from the waste package, which are also inputs to the drift, are colloid-associated 
radionuclide concentrations and colloid mass concentration (Figure 6-17). 
The colloid-associated radionuclide source term is calculated in Step 4 by passing to the invert 
the contributions to the source term from Steps 1d, 2b, and 3c (Figure 6-27). 
Figure 6-27. Flow Chart and Logic Statements: Transfer of Colloid-Associated Radionuclide Source 
Term to Invert 
The radionuclide component of the source term is made up of individual radionuclide mass 
concentrations irreversibly sorbed in and reversibly attached to waste form (smectite) colloids 
(CRNcoll,wf,embed and CRN,wf,rev), irreversibly and reversibly attached to corrosion-generated iron 
oxyhydroxide colloids (CRNcoll,FeOx,irrev and CRNcoll,FeOx,rev), and reversibly attached to groundwater 
(smectite) colloids (CRNcoll,gw). 
The colloid component of the source term is the sum of the mass concentrations of the waste 
form colloids (Mcoll,wf,both; both signifies the inclusion of smectite waste form colloids with 
irreversibly sorbed radionuclides, or reversibly associated with radionuclides, or both); the iron 
oxyhydroxide corrosion-generated colloids (Mcoll,FeOx,both; both signifies that corrosiongenerated 
iron oxyhydroxide colloids irreversibly or reversibly associated with radionuclides, or 
both, are included); and the smectite groundwater colloids. 

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6.5.2 In-Drift Abstraction Implementation in TSPALicense Application 
6.5.2.1 In-Drift Colloid Stability and Mass Concentrations 
Step 5 incorporates the outputs from the waste package in Step 4 with the invert fluid properties, 
ionic strength (Iinv) and pH (pHinv), from TSPA-LA calculations. 
Step 5. The primary function of Step 5 is incorporation of the invert fluid chemical properties 
and the calculation of resulting stabilities and mass concentrations of colloids entering the invert 
fluid from the waste package as shown in Figure 6-28. 
Figure 6-28. Inputs to Drift from Waste Package 
The concentrations of colloids and colloid-associated radionuclides within the drift are calculated 
on the basis of abstractions of the laboratory and field observations and chemical principles 
described in Section 6.3, with methods similar to those used to determine radionuclide and 
colloid concentrations within the waste package. A stepwise approach was taken in the 
abstraction, and each step is set out below with a flow diagram and logic statements that describe 
the abstraction step. Parameters are defined in Table 6-12. 
In the TSPA-LA model calculations, the colloids with reversibly and irreversibly sorbed 
radionuclides leave the failed waste package and enter the drift geochemical environment. The 
colloids exit in-package chemical conditions and enter invert chemical conditions, which are 
potentially different from in-package conditions. Once the colloids are subject to the invert 
chemical conditions, the stability and concentration of the colloid suspensions are recalculated, 
based on Iinv and pHinv. Therefore, if the colloid mass concentrations change (due to loss of 
colloid stability), the concentrations of sorbed radionuclides are recalculated as shown in 
Figure 6-27. 
6.5.2.2 Recalculation of Colloid-Associated Radionuclides Concentrations in Drift 
Step 6. In Step 6, the concentrations of radionuclides sorbed to colloids exiting the waste 
package are recalculated because of different chemical conditions in the invert. The new 
concentrations of dissolved radionuclides sorbed onto the colloids are determined. This step is 
illustrated by the flow chart in Figure 6-29 and by the logic statements that follow: 

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MDL-EBS-PA-000004 REV 02 6-80 August 2005 
Figure 6-29. Recalculation of Colloid-Associated Radionuclides in Drift 
CRNdiss,net is the net concentration of dissolved radionuclides after removal of radionuclides from 
fluid by sorption onto waste form, corrosion, and seepage water colloids. CRNtotal is, therefore, 
equivalent to CRNdiss in Steps 1d, 2b, and 3c. 
Calculation of radionuclide sorption on colloids: 
CRNcoll,wf,rev = CRNdiss  Kd,RN,wf  Mcoll,wf,both 
CRNcoll,FeOx,rev = CRNdiss  Kd,RN,FeOx  Mcoll,FeOx,both 
CRNcoll,gw = CRNdiss  Kd,RN,gw  Mcoll,gw 
6.5.2.3 Calculation of Colloid Source Term for Radionuclide Element (RN) 
Step 7. The colloid source term is calculated by summing the determined concentrations of 
waste form, iron oxyhydroxide, and seepage water or groundwater colloids and the 
concentrations of the radionuclides associated with the colloids, as shown in Figure 6-30. 
Figure 6-30. Source Term Summation for Colloid-Associated Radionuclide Concentrations in Drift 
The sum of the radionuclides sorbed to colloids reversibly sorbed and irreversibly sorbed) and 
net dissolved radionuclides (those remaining after sorption) are passed to the unsaturated zone 
for disposition in TSPA-LA calculations. 

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6.5.3 Implementation of Colloid-Facilitated Transport in the Natural Barrier System 
The maximum groundwater colloid concentration used in the unsaturated zone and saturated 
zone models is a sampled value based on the data presented in Section 6.3.2.4 and in 
Figures 6-11 and 6-12. The groundwater colloids are modeled as smectite clay; smectite colloids 
are available to form pseudocolloids through sorption of dissolved radionuclides. The 
radionuclides plutonium, americium, protactinium, cesium, and thorium are considered in the 
unsaturated zone and saturated zone models for colloid-facilitated transport. The colloids 
considered in the models include, in addition to groundwater colloids, primary waste form 
colloids (smectite), and corrosion-product colloids (hematite). The smectite waste form colloids, 
which originate from alteration of DHLW glass, contain plutonium and americium that are 
assumed to be irreversibly attached) and reversibly sorbed radionuclides. Corrosion-product 
colloids also contain reversibly and irreversibly sorbed radionuclides. All of these colloids and 
associated radionuclides are passed to the unsaturated zone from the engineered barrier system. 
Colloid parameters used in the waste form/engineered barrier system models are listed in 
Table 6-6. The smectite values in Table 6-6 are also recommended for the unsaturated zone and 
saturated zone models. 
Colloid transport in the unsaturated zone (BSC 2004 [DIRS 170041]) and saturated zone 
(BSC 2004 [DIRS 170006]) occurs by advective transport; colloid diffusion is considered 
negligible because colloids have free diffusion coefficients that are 2 to 4 orders of magnitude 
smaller than solutes. Colloids will transport wherever water flows, which in the unsaturated 
zone is a combination of fractures and the matrix, and in the saturated zone it is almost 
exclusively in fractures or through advective flow pathways in alluvium. The project has no sitespecific 
data on colloid transport under unsaturated conditions, so colloid transport parameters 
developed from tests conducted in saturated fractured volcanic tuffs are applied to the 
unsaturated zone (BSC 2004 [DIRS 170041]). This approach is considered conservative, as it is 
widely recognized that colloids tend to be more retarded under unsaturated conditions than under 
saturated conditions in media that is otherwise similar. This greater retardation in the 
unsaturated zone occurs because: (1) immobile airwater interfaces serve as additional surfaces 
to trap colloids and, (2) unsaturated zone porewaters tend to have higher ionic strengths than 
saturated zone porewaters, resulting in a greater tendency for colloids to attach to rock surfaces. 
Some of the plutonium and americium leaving the engineered barrier system is modeled as 
irreversibly sorbed to colloids. The irreversibly sorbed fractions of these radionuclides travel 
through the unsaturated zone and saturated zone at the same rate as the associated colloids. 
Irreversibly sorbed radionuclides on colloids leaving the engineered barrier system are split into 
two fractions: one retarded by colloid filtration and detachment processes and the other assumed 
to be completely unretarded. The fraction that travels unretarded depends on the groundwater 
travel time through the natural barrier system (unsaturated zone and saturated zone combined), 
with greater travel times resulting in smaller fractions. However, the unretarded fraction is 
<0.01, so the vast majority of the colloids experience filtration processes in the natural barriers 
that result in colloid retardation. The fraction of unretarded colloids is determined by a 
procedure described in Saturated Zone Colloid Transport (BSC 2004 [DIRS 170006]). The 
rationale for assuming that a small colloid mass fraction travels unretarded through the natural 
barriers is that (1) some finite fraction of colloids will travel unretarded in laboratory and field 

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transport tests, and (2) there have been several observations of small mass fractions of 
contaminants migrating with colloids over large distances in relatively short times in field 
settings (e.g., Kersting et al. 1999 [DIRS 103282]). 
For colloids that experience filtration in the natural barrier system, colloid retardation factors 
(Rc) are randomly sampled from cumulative distribution functions developed separately for the 
fractured volcanics and the alluvium. These cumulative distribution functions are constructed 
using retardation factors derived from laboratory and field colloid transport tests conducted 
under varying geochemical conditions with different colloid types and sizes. The value of Rc is 
dependent on several factors such as colloid size, colloid type and geochemical conditions 
(e.g., pH, Eh, and ionic strength). For any TSPA simulation, the cumulative distribution 
functions are sampled independently for the volcanics and alluvium. However, the value 
sampled for the volcanics is used for both the unsaturated zone and the volcanic portion of the 
saturated zone flow system, but the ranges are sampled independently. No correlation is 
assumed between retardation factors in the volcanics and alluvium. The cumulative distribution 
functions in the volcanics and the alluvium are truncated at the lower end so that the minimum 
retardation factor in the volcanics is 6 and the minimum in the alluvium is 8. These truncations 
result in the vast majority of colloids (with the exception of the small unretarded fraction) 
experiencing retardation in the natural barrier system. 
The transport of radionuclides that are reversibly sorbed to colloids is governed by equilibrium 
partitioning of the radionuclides between the solution phase, the immobile rock matrix, and 
mobile colloids. Plutonium radioisotopes are transported as one group; americium, protactinium, 
and thorium radioisotopes as a second group; and cesium radioisotopes as a third group. 
Groundwater travel times through the natural barriers are slow enough that equilibrium sorption 
can be assumed even for slow radionuclide sorption and desorption rates. Any radionuclide 
mass that desorbs from colloids at rates slow enough to be comparable to travel times through 
the system is inherently considered part of the irreversibly sorbed radionuclide mass fraction. 
Colloid-facilitated transport of reversibly sorbed radionuclides depends on colloid transport 
parameters, the mobile colloid mass concentrations, and radionuclide partition coefficients onto 
colloids and the immobile rock matrix. High colloid concentrations and large radionuclide 
partition coefficients onto colloids (relative to the rock matrix) favor colloid-facilitated 
radionuclide transport at rates that can exceed transport rates in the absence of colloids. Because 
reversibly sorbed radionuclides will readily desorb from retarded colloids to resume partitioning 
between the solution phase, mobile colloids, and the immobile rock matrix, the retardation of 
colloids works to counterbalance the otherwise enhanced mobility of the radionuclide. However, 
because the sorption of this fraction is assumed to be reversible, colloids still serve to facilitate 
transport, even if they are retarded due to filtration processes. 
6.6 BASE-CASE MODEL RESULTS 
This section provides an overview of the model abstraction results, with a general discussion of 
the range of results, important couplings among processes and their impact on performance 
measures, and data uncertainty and propagation of the uncertainties through the model. 

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6.6.1 Overview 
As discussed in Section 6.1, colloid suspensions may affect repository performance if they are 
generated, and are stable within the waste package, engineered barrier system, and unsaturated 
and saturated zones. Radionuclide-bearing colloids could readily transport from the engineered 
barrier system to the unsaturated zone and saturated zone and beyond. Output quantities from 
the waste form and engineered barrier system colloids models to downstream natural barrier 
system models and/or analyses, or both, include waste form colloid source terms, including 
DHLW glass colloid concentrations with irreversibly sorbed radionuclides. Also, the waste form 
and engineered barrier system colloids models determine the concentration of natural seepage 
water/groundwater colloids and corrosion product (iron oxyhydroxide) colloids that are 
propagated from the engineered barrier system onto the underlying natural barrier system. 
6.6.2 Uncertainty Associated with the Model Analysis 
As discussed previously in Section 4.1, the parameters developed in this model analysis are in 
some cases based on data that are not site- or Yucca Mountain Project (YMP)-specific. 
Additionally, supporting technical and corroborative information may not have been developed 
specifically from physical and chemical conditions approximating those of the repository. 
Therefore, departure from anticipated repository conditions must be accounted for by 
acknowledging uncertainties in the parameters developed and providing a means to incorporate 
these uncertainties into the TSPA calculations. In this model analysis, value ranges and 
probability distributions, and bounding assumptions deemed appropriate for each given 
parameter or process modeled have been established as a means to incorporate these 
uncertainties into the TSPA-LA model calculations. Another potential uncertainty issue is 
related to the circumstances and conditions of data acquisition and the departure from those 
circumstances and conditions assumed during the application of the data to the TSPA models. 
The three types of parameter range-related issues relate to: (1) temporal scaling, (2) spatial 
scaling, and (3) extrapolation from the physical and chemical conditions of data acquisition. 
Temporal ScalingIn the implementation of a TSPA, it is obviously necessary to use 
experimental and field data obtained over a much shorter period of time than the regulatory 
compliance period and beyond (i.e., periods extending from thousands to hundreds of thousands 
of years). Parameters based on fundamental physical or chemical principles, such as 
radionuclide half-life, should be robust over these extended time parameters. Other types of 
parameters such as association of dissolved radionuclide ions with sorption sites on the surface 
of colloids, which is influenced by kinetics, must be used cautiously when speculating about 
future conditions. This aspect of uncertainty is best addressed with the formulation and use of 
scientifically based models. 
Spatial ScalingThis refers to the extrapolation or application of a documented process 
occurring at laboratory scale (centimeters), for example, to the scale of the repository setting 
(kilometers). Theory and observations of colloid behavior at the molecular and microscopic 
level used in this model report have been corroborated or validated to the extent possible with 
empirical field observations. An example of data that could be used to validate model 
predictions with actual field observations is the data reported by Kersting et al. (1999 
[DIRS 103282]). In this study observations of colloidal plutonium transport over kilometer-scale 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-84 August 2005 
distances at a nuclear detonation site (Kersting et al. 1999 [DIRS 103282]) were used to 
demonstrate the potential for long-distance transport of colloid-associated plutonium within and 
from the repository to the accessible biosphere  these results can be used to assess the validity 
of plutonium transport models at a more localized scale (i.e., perhaps at the localized scale of 
meters or even smaller distances of transport). 
The model uses stability (zeta potential measurements) of ideal particles with pH and ionic 
strength to predict long-term behavior in a heterogeneous waste package system that will contain 
many different types of surfaces that will be dissimilar to montmorrillonite and hematite. 
Elimelech et al. (2000 [DIRS 171066]) have demonstrated several problems with this approach 
and the role that minor components play in colloid deposition. However, the model is 
conservative in the application of colloid stability (for example by not including the Schulze- 
Hardy phenonemon) and these effects are not considered to have an impact. 
Extrapolation From Physical and Chemical Conditions of Data AcquisitionExperiments, 
the laboratory and field may yield potentially useful data for TSPA models, but these may have 
employed several parameters (e.g., pH or redox conditions) with ranges not directly relevant to 
repository conditions. In this model, the colloid properties of iron oxide and smectite clay have 
been modeled over repository conditions (pH and ionic strength) and the model is considered 
valid. 
Uncertainty Associated with Waste Form Colloid FormationThere is uncertainty in the 
model for colloid generation and the nature of colloids generated from the waste forms 
considered in this analysis. Uncertainties associated with these data and the experiments used to 
obtain them include measurement uncertainties and uncontrolled experimental factors and the 
subsequent extrapolation of these results to potential repository conditions. For example, one 
contribution to this uncertainty for CSNF is the potential for the sequestration of colloids 
formed during the degradation process within rinds of waste degradation products. The 
formation of alteration product rinds on the surface of the waste forms could impede the release 
of colloids from the waste and result in an underestimation of the total quantity of colloids 
produced. Additionally, uncertainties during the measurement of colloids formed from the 
various waste forms can potentially arise in the course of determination of mass, calculation of 
volume, determination of particle sizes by filtration, and measurement of concentrations by the 
various detection methods employed. 
Uncertainty Associated with Corrosion Product Colloid FormationThere is limited 
information in the literature concerning the quantity of corrosion product colloids that could 
potentially form during the degradation of waste package materials and the corrosion of other 
steel components in the engineered barrier system. For instance, the role of manganese in 
potential waste package colloids is uncertain. The manganese loading in canister materials 
(Carbon Steel Type A516) is 1.045% (BSC 2005 [DIRS 174583]). This manganese may effect 
the distribution of plutonium and other colloids. Buck and Bates (1999 [DIRS 109494]) noted 
the formation of manganese weathering products in unsaturated drip corrosion tests on DHLW 
glass. Duff et al. (1999 [DIRS 170925]) have demonstrated in laboratory sorption tests with 
Yucca Mountain tuff that plutonium is predominantly associated with manganese oxides in the 
rock rather than with the more abundant hematite minerals. The waste package materials will 
contain about 1% manganese, but the role of this species in plutonium behavior is not considered 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-85 August 2005 
in the model; however, the iron oxides corrosion products will dominate the interfacial 
environment in waste package. 
Therefore, the values and ranges developed for the corrosion colloid source term parameters 
have a large degree of uncertainty. To reduce some of this uncertainty and increase the 
defensibility of the developed parameters, consideration was given to the observed quantities of 
iron oxyhydroxide colloids observed in scaled-down miniature waste package degradation 
experiments (Section 6.3.1.3). The use of appropriately scaled-down waste package 
configurations (i.e., surface area to total package volume scaled-down proportionally to 
approximately 1/70th the size of actual repository waste packages) and the use of water of 
similar chemical composition to the groundwater found at Yucca Mountain render the 
experiments directly relevant to conditions anticipated in the Yucca Mountain repository. 
Additionally, conservative bounding values have been established for the corrosion product 
colloid parameters recommended for use in the model in order to propagate the uncertainty 
associated with these parameters throughout the TSPA-LA analysis. 
Uncertainty Associated with the Measurement of Seepage Water/Groundwater Colloids 
The determination of colloid concentrations in groundwater samples presents many challenges 
that add to the uncertainty of developed parameters. Major factors that contribute to uncertainty 
in the concentrations of colloids in the groundwater samples used to establish parameters for use 
in the TSPA analysis include (1) field sampling techniques, including differences in pumping 
rates at each well 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, (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). At 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 reduce some of this 
uncertainty, mass concentrations considered to establish the groundwater colloid concentration 
parameters were pooled using primarily groundwater samples extracted in the vicinity of Yucca 
Mountain (a few samples collected from Idaho National Laboratory were also used in the 
analysis) and a discrete cumulative distribution function was established to evaluate the 
uncertainty in colloid concentration distribution (Section 6.3.2.4, Figure 6-12). The goal of the 
uncertainty distribution for groundwater colloid concentrations is to numerically estimate 
uncertainty on a large scale. A cumulative distribution function was developed to accommodate 
the variable colloid concentrations and the uncertainty in groundwater sample collection. This 
allowed for stochastic sampling, including high concentrations (i.e., exceeding 100 mg/L) in at 
least 2% of the model iterations, while probabilistically allowing for the selection of lower mass 
concentrations (0.001 mg/L to 1.0 mg/L) 75% of the time. 
Uncertainty Associated with the Determination of Colloid StabilityThe ionic strength and 
pH of the water are the major considerations influencing the stability of the colloid suspensions 
considered in the model analysis (Section 6.3.2, Colloid Stability and Concentration). As with 
other parameters and data used or established in this analysis, uncertainty in the stability of 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-86 August 2005 
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. Again, much of this uncertainty 
is accommodated by establishing conservative bounding values and ranges of parameter values 
whose stability is affected by pH and ionic strength. This uncertainty is propagated through the 
TSPA-LA model by stochastic sampling of these distributions during Monte Carlo simulations 
employed in the model calculations. 
Uncertainty Associated with Radionuclide Sorption Onto ColloidsBecause the specific 
surface area of a colloidal phase may be larger than that of the mineral phase from which the Kd 
value is determined; the Kd distributions have been increased to account for this effect. High 
redox potentials will have an effect on actinide sorption and chemical state. In most 
groundwaters, this is not an issue; however, the redox potential at the surface of a nuclear waste 
form may be high (i.e., >500 meV) (Runde et al. 2002 [DIRS 168432]). Within this range small 
changes in the redox potential can cause large changes in overall plutonium behavior and will 
consequently effect colloid formation and adsorption. However, as these processes would tend 
to reduce the amount of colloidal plutonium, the model is considered to have bounded this 
process. 
Model Implementation of UncertaintyImplementation of the uncertainties associated with 
the parameters established during this model analysis is accomplished through the use of 
parameter ranges and distributions, rather than single, deterministic parameter values in the 
model analysis bounding values were established for most of the parameters. Within the 
TSPA-LA model, implementation of the colloids model abstraction incorporates the parameter 
uncertainties through the use of Monte Carlo simulation techniques, thereby ensuring that 
parameter uncertainties are propagated throughout the abstraction during computation and, 
therefore, reflected in the calculated results. Implementation of the various components of the 
model abstraction during their implementation in the TSPA-LA model is detailed in Section 6.5. 
In general, the uncertainty associated with each of these submodel components is related to the 
incomplete knowledge of the system and its properties (epistemic). 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-87 August 2005 
Table 6-12. Summary of Intermediate Parameters and Results Produced in Abstraction 
Reference 
Parameter 
Name Data/Parameter Description Unitsa 
Locations in Model 
Report Value/ Range 
Distribution/ 
Uncertainty 
Waste Form Colloids 
SN0311T0504103.011 CRNcoll,wf,embed,max Highest observed or expected 
concentration of irreversibly 
attached (irreversibly sorbed 
(embedded) as a solid inclusion) 
plutonium associated with waste 
form colloids 
mol/L Figure 6-18a 1  10-7 mol/Lb Fixed value 
Epistemic 
SN0311T0504103.011 CRNcoll,wf,embed,min Lowest observed or expected 
concentration of irreversibly 
attached (embedded) plutonium 
associated with waste form colloids 
mol/L Figure 6-18a 1  10-11 mol/Lb 
(small nonzero 
number) 
Fixed value 
Epistemic 
N/A CRNcoll,wf,embed Concentration of irreversibly 
attached radionuclide element RN 
(plutonium, americium) associated 
with waste form colloids 
mol/L Section 6.5.1 and 
Figure 6-18a 
N/A; calculated at 
each time step 
N/A; calculated at 
each time step 
Epistemic 
SN0311T0504103.011 Mcoll,wf,both,max Highest observed or expected mass 
of waste form colloids per unit 
volume or mass of water 
mg/L Sections 6.3.1.1 
Figure 6-20 
5.0 mg/Lc Fixed value 
Epistemic 
CRWMS M&O 2001 [DIRS 
154071] 
Mcoll,wf,both,min Lowest observed or expected mass 
of waste form colloids per unit 
volume or mass of water 
mg/L Sections 6.3.1.1 
Figure 6-19 
5.0  10-4 mg/Lc Fixed value 
Epistemic 
N/A Mcoll,wf,both Mass of waste form colloids with 
reversibly sorbed or irreversibly 
attached (embedded) radionuclide 
element RN per unit volume or 
mass of water 
mg/L Section 6.5.1 and 
Figures 6-20 and 6-21 
N/A; calculated at 
each time step 
N/A; calculated at 
each time step 
Epistemic 
N/A; Tombcz et al. 1990 
[DIRS 112690]; Figure 4 
Ilo-thresh,coll,wf Ionic strength below which waste 
form colloids are stable 
mol/L Sections 6.3.2.1 and 
6.3.2.3 and 
Figure 6-6a 
0.01 mol/L Fixed value 
Epistemic 
N/A; Tombcz et al. 1990 
[DIRS 112690]; Figure 4; 
CRWMS M&O 2001 [DIRS 
154071] 
Ihi-thresh,coll,wf Ionic strength above which waste 
form colloids are unstable 
mol/L Sections 6.3.2.1 and 
6.3.2.3 and 
Figure 6-6a, 6-4, and 
6-9a 
0.05 mol/L Fixed value 
Epistemic 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-88 August 2005 
Table 6-12. Summary of Intermediate Parameters and Results Produced in Abstraction (Continued) 
Data Tracking 
Number (DTN) 
Parameter 
Name Data/Parameter Description Unitsa 
Locations in Model 
Report Value/ Range 
Distribution/ 
Uncertainty 
N/A STcoll,wf,is Function relating stability of waste 
form colloids to ionic strength, 
based on results of DHLW glass 
degradation experiments (based on 
plutonium concentration) 
and 
mol/L 
Section 6.5.1 
Figure 6-19 
[Pu colloid] = -2.50 
 10-6  I +1.25  
10-7 mol/L 
N/A; calculated at 
each time step 
Epistemic 
N/A STcoll,wf,pH Function relating stability of waste 
form colloids to ionic strength and 
pH, over limited pH range (pH 2 to 
10), based on results of 
montmorillonite stability experiments 
mol/L Section 6.3.2.1 and 
6.3.2.3 and 
Figures 6-6a, 6-4, and 
6-9a 
Section 6.5.1 and 
Figure 6-19 
I = pH/200 N/A; calculated at 
each time step 
Epistemic 
SN0306T0504103.006 Kd,Pucoll,wf Distribution coefficient for reversible 
sorption of plutonium onto waste 
form (smectite) colloids 
mL/g 
colloid 
Section 6.3.3.1 
Figure 6-21 
103 to 106 mL/g 
colloid 
Table 6-6 
Epistemic 
SN0306T0504103.006 Kd,Amcoll,wf Distribution coefficient for reversible 
sorption of americium onto waste 
form (smectite) colloids 
mL/g 
colloid 
Section 6.3.3.1 
Figure 6-21 
104 to 107 mL/g 
colloid 
Table 6-6 
Epistemic 
SN0306T0504103.006 Kd,Thcoll,wf Distribution coefficient for reversible 
sorption of thorium onto waste form 
(smectite) colloids 
mL/g 
colloid 
Section 6.3.3.1 
Figure 6-21 
104 to 107 mL/g 
colloid 
Table 6-6 
Epistemic 
SN0306T0504103.006 Kd,Pacoll,wf Distribution coefficient for reversible 
sorption of Pa onto waste form 
(smectite) colloids 
mL/g 
colloid 
Section 6.3.3.1 
Figure 6-21 
104 to 107 mL/g 
colloid 
Table 6-6 
Epistemic 
SN0306T0504103.006 Kd,Cscoll,wf Distribution coefficient for reversible 
sorption of cesium onto waste form 
(smectite) colloids 
mL/g 
colloid 
Section 6.3.3.1 
Figure 6-21 
102 to 104 mL/g 
colloid 
Table 6-6 
Epistemic 
N/A CRNcoll,wf,rev Concentration of reversibly attached 
radionuclide element RN (plutonium, 
americium, thorium, Pa, cesium) 
associated with waste form colloids 
mol/L Section 6.5.1 
Figure 6-21 
N/A; calculated at 
each time step 
N/A; calculated at 
each time step 
Epistemic 
N/A; calculated by the 
TSPA-LA model 
(in-package chemistry) 
Iwp Ionic strength of fluid within and 
exiting breached waste package 
mol/L Section 6.5.1 and 
Figure 6-18 
N/A N/A 
N/A; calculated by the 
TSPA-LA model (inpackage 
chemistry) 
pHwp pH of fluid within and exiting 
breached waste package 
pH units Section 6.5.1 and 
Figure 6-19 
N/A N/A 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-89 August 2005 
Table 6-12. Summary of Intermediate Parameters and Results Produced in Abstraction (Continued) 
Data Tracking 
Number (DTN) 
Parameter 
Name Data/Parameter Description Unitsa 
Locations in Model 
Report Value/ Range 
Distribution/ 
Uncertainty 
N/A; calculated by the 
TSPA-LA model (solubility 
limits) 
CRNdiss Concentration of dissolved 
radionuclide element RN per unit 
volume or mass of water (plutonium, 
americium, thorium, protactinium, 
cesium) 
mol/L Sections 6.5.1 and 
6.5.2 
N/A N/A 
N/A; calculated by the 
TSPA-LA model 
Iinv ionic strength of fluid within invert mol/L Section 6.5.2 
Figure 6-17 
N/A N/A 
N/A; calculated by the 
TSPA-LA model 
pHinv pH of fluid within invert pH units Section 6.5.2 
Figure 6-17 
N/A N/A 
Corrosion-Generated Colloids 
SN0309T0504103.010 Mcoll,FeOx,sampled Highest observed or expected mass 
of iron oxyhydroxide colloids per unit 
volume or mass of water 
mg/L Section 6.3.2.2 
Figure 6-22 
0.05 to 50 mg/L Uniform 
Epistemic 
SN0309T0504103.010 Mcoll,FeOx,min Lowest observed or expected mass 
of iron oxyhydroxide colloids per unit 
volume or mass of water 
mg/L Section 6.3.2.2 
Figure 6-22 
1E-3 mg/L Fixed value 
Epistemic 
SN0309T0504103.010 k Forward rate constant for plutonium 
and americium for irreversible 
sorption onto iron oxyhydroxide 
colloids and stationary corrosion 
products in the waste package for 
zero advective flux conditions or 
where colloids are unstable 
m3 m-2 yr- 
1 
Section 6.5.1.2 
0.01 to 0.24 m3 m- 
2 yr-1 
Log Uniform 
Epistemic 
N/A Mcoll,FeOx Mass of iron oxyhydroxide colloids 
per unit volume or mass of water 
mg/L Section 6.5.1 and 
Figure 6-22 
N/A; calculated at 
each time step 
N/A; calculated at 
each time step 
Epistemic 
N/A Mcoll,FeOx, both Mass of iron oxyhydroxide colloids 
per unit volume or mass of water 
with both reversibly and irreversibly 
sorbed plutonium and americium 
mg/L Section 6.5.1 and 
Figure 6-27 
N/A; calculated at 
each time step 
N/A; calculated at 
each time step 
Epistemic 
SN0306T0504103.006 Kd,Pucoll,FeOx Distribution coefficient for reversible 
sorption of plutonium onto 
corrosion-generated (iron 
oxyhydroxide) colloids 
mL/g 
colloid 
Sections 6.3.3.1 
Figure 6-23 
104 mL/g colloid 
(EBS) 
104 to 106 mL/g 
colloid (UZ/SZ) 
Table 6-6 
Epistemic 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-90 August 2005 
Table 6-12. Summary of Intermediate Parameters and Results Produced in Abstraction (Continued) 
Data Tracking 
Number (DTN) 
Parameter 
Name Data/Parameter Description Unitsa 
Locations in Model 
Report Value/ Range 
Distribution/ 
Uncertainty 
SN0306T0504103.006 Kd,Amcoll,FeOx Distribution coefficient for reversible 
sorption of americium onto 
corrosion-generated (iron 
oxyhydroxide) colloids 
mL/g 
colloid 
Sections 6.3.3.1 and 
6.3.3.2 
Figure 6-23 
105 mL/g colloid 
(EBS) 
105 to 107 mL/g 
colloid (UZ/SZ) 
Table 6-6 
Epistemic 
SN0306T0504103.006 Kd,Thcoll,FeOx Distribution coefficient for reversible 
sorption of thorium onto corrosiongenerated 
(iron oxyhydroxide) 
colloids 
mL/g 
colloid 
Section 6.3.3.1 
Figure 6-23 
105 to 107 mL/g 
colloid 
Table 6-6 
Epistemic 
SN0306T0504103.006 Kd,Pacoll,FeOx Distribution coefficient for reversible 
sorption of Pa onto corrosiongenerated 
(iron oxyhydroxide) 
colloids 
mL/g 
colloid 
Section 6.3.3.1 
Figure 6-23 
105 to 107 mL/g 
colloid 
Table 6-6 
Epistemic 
SN0306T0504103.006 Kd,Cscoll,FeOx Distribution coefficient for reversible 
sorption of cesium onto corrosiongenerated 
(iron oxyhydroxide) 
colloids 
mL/g 
colloid 
Section 6.3.3.1 
Figure 6-23 
101 to 103 mL/g 
colloid 
Table 6-6 
Epistemic 
SN0309T0504103.010 SA,FeOx,coll Specific (effective) surface area of 
corrosion-generated (iron 
oxyhydroxide) colloids 
m2/g 
colloid 
Section 6.3.3.2 
Figure 6-23 
1.8 to 720 m2/g Log Uniform 
Epistemic 
SN0309T0504103.010 FRN Target-flux out ratio: ratio of 
radionuclide mass associated with 
colloids (reversible and irreversible) 
to radionuclide mass associated 
with colloids and dissolved 
radionuclide mass 
------ Section 6.3.3.2 
Figure 6-23 
0.9 to 0.99 Uniform 
Epistemic 
N/A CRncoll,FeOx,rev Concentration of radionuclide 
element RN sorbed reversibly onto 
corrosion-generated (iron 
oxyhydroxide) colloids (plutonium, 
americium, thorium, protactinium, 
cesium) 
mol/L Section 6.5.1.2 and 
Figure 6-23 
N/A; calculated at 
each time step 
N/A; calculated at 
each time step 
Epistemic 
N/A CPuAm,coll,FeOx,irrev Concentration of plutonium and 
americium irreversibly sorbed onto 
corrosion-generated (iron 
oxyhydroxide) colloids 
mol/L Section 6.5.1.2, 
6.3.3.2and 
Figure 6-23. 
N/A; calculated at 
each time step 
N/A; calculated at 
each time step 
Epistemic 
N/A STcoll,FeOx,pHlo Function relating stability of iron 
oxyhydroxide colloids to pH and 
ionic strength at relatively low pH 
values 
mol/L Section 6.5.1 and 
Figure 6-22 
I = -0.02  pH 
+0.17 mol/L 
N/A; calculated at 
each time step 
Epistemic 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-91 August 2005 
Table 6-12. Summary of Intermediate Parameters and Results Produced in Abstraction (Continued) 
Data Tracking 
Number (DTN) 
Parameter 
Name Data/Parameter Description Unitsa 
Locations in Model 
Report Value/ Range 
Distribution/ 
Uncertainty 
N/A STcoll,FeOx,pHhi Function relating stability of iron 
oxyhydroxide colloids to pH and 
ionic strength at relatively high pH 
values 
mol/L Section 6.5.1 and 
Figure 6-22 
I = +0.02  pH 
0.17 mol/L 
N/A; calculated at 
each time step 
Epistemic 
N/A STcoll,FeOx,pHint Function relating stability of iron 
oxyhydroxide colloids to pH and 
ionic strength at intermediate pH 
values 
Boolean Section 6.5.1 and 
Figure 6-22 
Boolean (pH = 8 
AND pH = 9) 
N/A; calculated at 
each time step 
Epistemic 
N/A; Reference: Liang and 
Morgan 1990 
[DIRS 109524] 
Ilo-thresh,coll,FeOx Ionic strength below which iron 
oxyhydroxide colloids are stable 
mol/L Section 6.3.2.2 
Figures 6-8 
0.01 mol/L Fixed value 
Epistemic 
N/A; Reference: Liang and 
Morgan 1990 
[DIRS 109524] 
Ihi-thresh,coll,FeOx Ionic strength above which iron 
oxyhydroxide colloids are unstable 
mol/L Section 6.3.2.2 
Figures 6-8 
0.05 mol/L Fixed value 
Epistemic 
N/A; calculated by the 
TSPA-LA model (inpackage 
chemistry) 
Iwp Ionic strength of fluid within and 
exiting breached waste package 
mol/L Section 6.5.1 and 
Figure 6-22 
N/A N/A 
N/A; calculated by the 
TSPA-LA model (inpackage 
chemistry) 
pHwp pH of fluid within and exiting 
breached waste package 
pH units Section 6.5.1 and 
Figure 6-22 
N/A N/A 
N/A; calculated by the 
TSPA-LA model (solubility 
limits) 
CRNdiss Concentration of dissolved 
radionuclide element RN per unit 
volume or mass of water (plutonium, 
americium, thorium, protactinium, 
cesium) 
mol/L Sections 6.5.1 and 
6.5.2; 
Figure 6-23 
N/A N/A 
N/A; calculated by the 
TSPA-LA model 
Iinv I of fluid within invert mol/L Section 6.5.2 
Figure 6-28 
N/A N/A 
N/A; calculated by the 
TSPA-LA model 
pHinv pH of fluid within invert pH units Section 6.5.2 
Figure 6-28 
N/A N/A 
Groundwater Colloids 
N/A Mcoll,gw Mass of groundwater colloids per 
unit volume or mass of water 
mg/L Section 6.5.1 
Figure 6-24 
N/A; calculated at 
each time step 
N/A; calculated at 
each time step 
Epistemic 
SN0306T0504103.005 Mcoll,gw, sampled Mass of groundwater colloids per 
unit volume or mass of water 
sampled at each time step 
mg/L Section 6.3.2.4 
Figure 6-24 
See Table 6-4 Table 6-4 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 6-92 August 2005 
Table 6-12. Summary of Intermediate Parameters and Results Produced in Abstraction (Continued) 
Data Tracking 
Number (DTN) 
Parameter 
Name Data/Parameter Description Unitsa 
Locations in Model 
Report Value/ Range 
Distribution/ 
Uncertainty 
SN0306T0504103.005 Mcoll,gw, min Minimum mass of groundwater 
colloids per unit volume or mass of 
water sampled at each time step 
mg/L Section 6.3.2.4 
Figure 6-24 
1  10-6 mg/L N/A 
SN0306T0504103.006 Kd,Pucoll,gw Distribution coefficient for reversible 
sorption of plutonium onto 
groundwater (smectite) colloids 
mL/g 
colloid 
Section 6.3.3.1 
Figure 6-26 
103 to 106mL/g 
colloid 
Table 6-6 
Epistemic 
SN0306T0504103.006 Kd,Amcoll,gw Distribution coefficient for reversible 
sorption of americum onto 
groundwater (smectite) colloids 
mL/g 
colloid 
Section 6.3.3.1 
Figure 6-26 
104 to 107mL/g 
colloid 
Table 6-6 
Epistemic 
SN0306T0504103.006 Kd,Thcoll,gw Distribution coefficient for reversible 
sorption of Th onto groundwater 
(smectite) colloids 
mL/g 
colloid 
Section 6.3.3.1 
Figure 6-26 
104 to 107mL/g 
colloid 
Table 6-6 
Epistemic 
SN0306T0504103.006 Kd,Pacoll,gw Distribution coefficient for reversible 
sorption of protactinium onto 
groundwater (smectite) colloids 
mL/g 
colloid 
Section 6.3.3.1 
Figure 6-26 
104 to 107mL/g 
colloid 
Table 6-6 
Epistemic 
SN0306T0504103.006 Kd,Cscoll,gw Distribution coefficient for reversible 
sorption of cesium onto groundwater 
(smectite) colloids 
mL/g 
colloid 
Section 6.3.3.1 
Figure 6-26 
102 to 104 mL/g 
colloid 
Table 6-6 
Epistemic 
N/A CRNcoll,gw Concentration of reversibly attached 
radionuclide element RN associated 
with groundwater colloids 
(plutonium, americium, thorium, 
protactinium, cesium) 
mol/L Section 6.5.1 and 
Figure 6-26 
N/A; calculated at 
each time step 
N/A; calculated at 
each time step 
Epistemic 
N/A; calculated by the 
TSPA-LA model (inpackage 
chemistry) 
Iwp Ionic strength of fluid within and 
exiting breached waste package 
mol/L Section 6.5.1 and 
Figure 6-24 
N/A N/A 
N/A; calculated by the 
TSPA-LA model (inpackage 
chemistry) 
pHwp pH of fluid within and exiting 
breached waste package 
pH units Section 6.5.1and 
Figure 6-25 
N/A N/A 
N/A; calculated by TSPALA 
model (solubility limits) 
CRNdiss Concentration of dissolved 
radionuclide element RN per unit 
volume or mass of water (plutonium, 
americium, thorium, protactinium, 
cesium) 
mol/L Sections 6.5.1 and 
6.5.2; 
Figure 6-26 
N/A N/A 
N/A; References: 
Tombcz et al. 1990 [DIRS 
112690]; Figure 4 
Ihi-thresh,coll,gw Ionic strength above which 
groundwater colloids are unstable 
mol/L Sections 6.3.2.1 and 
6.3.2.4 
Figure 6-6b 
0.05 mol/L Fixed value 
Epistemic 

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MDL-EBS-PA-000004 REV 02 6-93 August 2005 
Table 6-12. Summary of Intermediate Parameters and Results Produced in Abstraction (Continued) 
Data Tracking 
Number (DTN) 
Parameter 
Name Data/Parameter Description Unitsa 
Locations in Model 
Report Value/ Range 
Distribution/ 
Uncertainty 
N/A; calculated by the 
TSPA-LA model (solubility 
limits and near-field 
chemical environment) 
CRNdiss Concentration of dissolved 
radionuclide element RN per unit 
volume or mass of water (plutonium, 
americium, thorium, protactinium, 
cesium) 
mol/L Sections 6.5.1 and 
6.5.2; 
Figure 6-26 
N/A N/A 
Invert Parameters 
N/A; calculated by the 
TSPA-LA model 
CRndiss,net Net concentration of dissolved 
radionuclides after removal of 
radionuclides by sorption to colloids 
mol/L Section 6.5.2.2 and 
Figure 6-29 
N/A N/A 
Epistemic 
N/A; calculated by the 
TSPA-LA model 
CRNtotal Sum of net dissolved and sorbed-tocolloids 
radionuclide concentrations 
mol/L Section 6.5.2.2 and 
Figure 6-29 
N/A N/A 
Epistemic 
N/A; calculated by the 
TSPA-LA model 
Iinv I of fluid within invert mol/L Section 6.5.2 
Figure 6-28 
N/A N/A 
Epistemic 
N/A; calculated by the 
TSPA-LA model 
pHinv pH of fluid within invert pH units Section 6.5.2 
Figure 6-28 
N/A N/A 
Epistemic 
NOTES: WP = waste package; N/A = not applicable; RN = radionuclide; UZ = unsaturated zone; SZ = saturated zone; EBS = engineered barrier system. 
a Concentration, mol/L, is treated here as moles/liter of dispersion and, therefore, is not molarity in the strict sense. 
b Uncertainty is linked to uncertainty for ionic strength, an input parameter in the TSPA-LA model (no additional uncertainty is added). 
c Mass is correlated with irreversible radionuclide concentration. 

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6.7 DESCRIPTION AND ANALYSES OF THE BARRIER CAPABILITY 
As discussed in Section 1, the abstraction of the process models reported in this document is 
intended to capture the most important characteristics of radionuclidecolloid behavior for use in 
predicting the potential impact of colloid-facilitated radionuclide transport on repository 
performance. Colloid information from this report will be considered in evaluation of the 
unsaturated zone as a barrier to radionuclide transport. As modeled in TSPA, colloids may act to 
facilitate transport of radionuclides from the waste package to the surrounding engineered barrier 
system and beyond to the underlying unsaturated zone component of the natural barrier system 
of the repository. 
6.8 EVALUATION OF ALTERNATIVE MODELS 
The alternative models described in Section 6.4 provide mechanisms to address colloid 
formation from spent nuclear fuel, a method for including kinetic-based sorption that allows for 
slowly reversible processes, an alternative model for waste form colloid generation, and a 
method for introducing thin-film filtration. The latter two models were considered to be of less 
practical use within the existing model structure. 
The spent nuclear fuel colloid stability models could be implemented using the available data in 
a manner similar to the clay and hematite colloids. There are data needs for radionuclide uptake 
into these additional phases that prevent implementation of the model. Spent nuclear fuelderived 
colloids may be limited by dissolution of U(VI) colloidal phases, more subject to 
physical filtration than other colloids, and their impact, at least in the igneous event scenario, will 
be bounded by groundwater colloids. 
The kinetic sorption model may provide a reasonable viewpoint for sorption by including a slow 
reversible term.. Additional data are needed from longer-term sorption studies. 

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MDL-EBS-PA-000004 REV 02 7-1 August 2005 
7. VALIDATION 
The waste form and in-drift colloids abstraction model is used to describe the possible 
consequences of colloid-facilitated radionuclide transport during postclosure. The types and 
characteristics (including stability and concentration) of colloids formed from the degradation of 
the waste forms as used in the abstraction is based on extensive observations of colloids from 
testing programs and natural groundwaters. 
Confidence-Building During Model Development to Establish Scientific Basis and 
Accuracy for Intended UseSection 2.6.1 of Technical Work Plan for Postclosure Waste 
Form Modeling (BSC 2005 [DIRS 173246]) lists six steps for confidence building during model 
development. These are reproduced here with pointers to where in the document each is 
addressed. 
Each model will contain documentation of decisions and activities implemented 
during model development to build confidence and verify a reasonable, 
corroborative, and credible technical approach using scientific and engineering 
principles was taken to: 
A. Evaluate and select input parameters or data, or both, 
The bases for selecting the input data used to determine and develop the colloid model are 
documented Section 4.1. The input data are obtained from controlled sources. Wherever 
possible, site-specific data were used, which builds confidence that the developed model is 
appropriate for its intended use. Model assumptions have been described in Section 5. Detailed 
discussion about model concepts can be found in Section 6. Thus, this requirement is met. 
B. Formulate defensible assumptions and simplifications, 
Discussion of assumptions and simplifications and their rationale are provided in Sections 5 and 
Section 6. Thus, this requirement is met. 
C. Ensure consistency with physical principles, such as conservation of mass, 
energy, and momentum, 
The colloid model is consistent with the physical behavior of colloids. The classical theory 
developed by Darjaguin, Landau, Verwey, and Overbeek (DLVO) (van Olphen 1977 
[DIRS 114428]) contends that colloid mobilization is activated by physical or chemical 
perturbations, which can change the forces between the colloid and surfaces. Colloid 
mobilization requires that the changes in solution chemistry or changes in the hydrodynamic 
shear forces produce repulsive forces between the attached colloids and the surface that exceed 
the attractive forces. However, large discrepancies will exist between predicted and observed 
detachment of particles under natural conditions, mainly because of the heterogeneous properties 
of the colloids and surfaces (both in chemical composition and geometries), nonuniform flow 
velocity, and chemical heterogeneity of the infiltrating solution. Furthermore, pH can result in 
increased stability of smectite clay colloid suspensions, which would not be predicted by a 
DLVO-based model. The waste package is a heterogeneous environment and has been 
simplified through the model; however, the application of the DLVO theory within the model 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 7-2 August 2005 
provides a physically reasonable, bounding, and conservative case for colloid behavior in the 
waste package. Thus, this requirement is met. 
D. Represent important future state (aleatoric), parameter, and alternative model 
uncertainties 
Uncertainties associated with the model analysis are discussed in Section 6.6.2. Alternative 
model uncertainties are discussed in Section 6.4. Thus, this requirement is met. 
E. Ensure simulation conditions have been set up to span the range of intended 
use and avoid inconsistent outputs, and 
Waste form corrosion data were limited to only certain chemical conditions. In the case of 
DHLW glass, this was due to the effect of the glass dissolution on the chemistry of the leachate. 
In other cases, including spent nuclear fuel experiments, tests were conducted in J-13 well water 
or other similar composition solutions. However, data on the characteristics of the two main 
types of colloids (smectite clay and hematite) is presented in the model abstraction over a wide 
range of pH and ionic strength. As high ionic strength and extreme pH values will reduce 
sorption, the model is bounding by using sorption values obtained in J-13 well water. The 
sorption coefficients were measured using dilute radionuclide concentrations to avoid the 
confounding process of precipitation. When these coefficients are applied under conditions 
where solubility is high (extreme pH values) they may overestimate the amount of sorption. The 
colloid concentration model spans the full range of ionic strength and pH. Thus, this 
requirement is met. 
F. Ensure model predictions (performance parameters) adequately represent the 
range of possible outcomes consistent with important uncertainties. 
Uncertainties associated with the model analysis are discussed in Section 6.6.2. Thus, this 
requirement is met. Thus, all six steps for confidence building during model development have 
been met. 
Confidence-Building After Model Development To Support The Scientific Basis Of The 
ModelPostdevelopment model validation is required by LP-2.29Q-BSC. For confidence 
building after model development, Technical Work Plan for Postclosure Waste Form Modeling 
(BSC 2005 [DIRS 173246], Table 2-3) specifies this model as Level II with eight validation 
activities and a validation criterion that corroborating data must match qualitatively or must be 
bounded by model predictions. 
As LP-SIII.10Q-BSC, Section 5.3.2(C) lists validation activities, this model report will refer to 
the technical work plan (BSC 2005 [DIRS 173246]) activities as questions to be answered to 
validate the model through the LP-SIII.10Q-BSC validation activities. The eight questions 
specified in the technical work plan (BSC 2005 [DIRS 173246]) are listed below, along with the 
applicable deviation from the technical work plan for question C. The technical work plan 
(BSC 2005 [DIRS 173246]) also refers to VA1 and VA3, which are from LP-SIII.10Q-BSC, 
Sections 5.3.2(C)1) and 5.3.2(C)3) respectively. They are repeated here for completeness. In 
addition, the technical work plan states (N/A) for several of the questions, as no procedure 
validation activity is applicable to the questions. 

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Validation Activity 1 - (VA1) Corroboration of model results with data acquired from the 
laboratory, field experiments, analog studies, or other relevant 
observations, not previously used to develop or calibrate the model. 
Validation Activity 3 - (VA3) Corroboration with information published in refereed journals 
or literature. 
The following questions from the technical work plan (BSC 2005 [DIRS 173246]) are to be 
answered through validation activities: 
Validation of Mathematical Form of Model 
Question A. Is the use of the stability criterion based upon ionic strength and pH a standard 
scientific practice and corroborated in the literature and is the dependence of 
colloid stability on pH and ionic strength consistent with Project and industry 
literature data (VA 1 and/or VA 3)? 
Question B. To increase confidence with colloid concentration parameters, is the use of 
colloid concentration derived from waste form degradation reasonable (VA N/A)? 
Question C. Is the irreversible sorption of plutonium and americium onto iron oxide colloids a 
standard scientific practice corroborated with literature and Project data (VA 1 
and/or VA 3)? 
There is no standard practice for modeling irreversible sorption. A deviation from the technical 
work plan is taken and the following more-appropriate question is asked during model 
development: Does the model of irreversible sorption of plutonium and americium onto iron 
oxide reasonably represent or bound the literature and project data (VA N/A)? Postdevelopment 
model validation is performed in EBS Radionuclide Transport Abstraction (BSC 2004 
[DIRS 169868]). 
Question D. Is the reversible sorption of radionuclides corroborated and accepted as 
established fact within industry literature (VA 1 and/or VA 3)? 
Validation of Model Results for Colloid and Plutonium Concentrations 
Question E. Are the concentrations of colloids and plutonium consistent with scientific 
principles and observed colloid concentrations in the Project- and peer-reviewed 
literature (VA 1 and/or VA 3)? 
Question F. Are the commercial and DOE-owned SNF models approaches consistent with 
commercial and DOE-owned SNF-specific test results and peer-reviewed 
literature on field observations near uranium mines and deposits (VA 1 and/or 
VA 3)? 
Question G. Are the ranges of corrosion product and groundwater colloid concentrations 
reasonable (VA 1 and/or VA 3)? 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 7-4 August 2005 
Question H. Are the distribution coefficient parameters for radionuclides onto colloids (Kd 
values) consistent with peer-reviewed literature and analytical modeling results 
(VA N/A)? 
Corroborating or supporting data and information used to develop and validate the parameters 
are listed in Table 7-1. 
Table 7-1. Supporting (Corroborating) Information Used to Validate the Colloid Model 
TWP 
Question 
Supporting (Corroborating) 
Information Source Data/Information 
A 
Tombacz and Szekeres 2004 [DIRS 174073] ; 
Davis and Kent 1990 [DIRS 143280]; 
Schindler et al. 2004 [DIRS 174507] 
Information on the surface properties of smectite clay 
colloids, iron oxides, and uranium (VI) phases 
E 
Mnard et al. 1998 [DIRS 171053] ; 
Luckscheiter and Nesovic 2002 [DIRS 170884] ; 
Pirlet 2001 [DIRS 174351]; Vernaz and Godon 
1992 [DIRS 174503] 
Corroborative information regarding the association of 
plutonium/americium with colloids during glass 
corrosion 
E 
Grambow et al. 1996 [DIRS 113253] ; 
Wronkiewicz et al. 1996 [DIRS 102047] ; 
McNamara et al. 2005 [DIRS 174068] 
Corroborative data supporting colloid concentrations 
with spent nuclear fuel and UO2 corrosion 
E Nuttall 1986 [DIRS 121867]; 
Kingston and Whitbeck 1991 [DIRS 113930] 
Colloid concentrations in groundwater 
F, G 
Murakami et al. 1997 [DIRS 113272]; 
Triay et al. 1995 [DIRS 100789] 
Corroborative information and data regarding colloidassociated 
uranium concentrations in the vicinity of 
uranium deposits 
F, G van Middlesworth and Wood 1998 
[DIRS 170927] 
Corroborative information and data regarding thorium 
and uranium colloid concentrations in natural waters 
H 
DTN: LA0002SK831352.003 [DIRS 161771] ; 
DTN: LA0002SK831352.004 [DIRS 161579] 
Corroborative data for groundwater colloids 
concentrations 
H 
Kingston and Whitbeck 1991 [DIRS 113930]; 
EPA 1999 [DIRS 147475]; EPA 2004 
[DIRS 172215]; Serne and Relyea (1982 
[DIRS 147681] 
Honeyman and Ranville 2002 [DIRS 161657]; 
Stenhouse 1995 [DIRS 147477] 
Distribution coefficients (Kd values) for radionuclides 
onto colloids. Modeling results for partitioning of 
radionuclides between colloids and stationary phases. 
7.1 VALIDATION OF MODEL 
7.1.1 Validation of Mathematical Form of Model 
The mathematical form of the colloid model consists of: (a) a stability criterion based upon ionic 
strength and pH, (b) concentration of plutonium and americium within the glass degradation 
colloids, (c) rate of plutonium and americium sorption onto iron oxide colloids, and (d) 
reversible sorption onto colloids. Each of these model components is validated below. 
Question A. Is the use of the stability criterion based upon ionic strength and pH a standard 
scientific practice and corroborated in the literature and is the dependence of 
colloid stability on pH and ionic strength consistent with Project and industry 
literature data (VA 1 and/or VA 3)? 

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MDL-EBS-PA-000004 REV 02 7-5 August 2005 
As ionic strength increases the stability of colloid suspensions decreases. The pHpzc for the types 
of colloids considered in the model will vary depending on particle morphology and phase. For 
instance, the pHpzc values for iron oxides can range from 6.8 to 9.3 depending on the phase 
(Table 7-2). The pH-dependent edge charge on montmorillonite clays can result in changes in 
particle-to-particle interactions and increases the stability with ionic strength (Tombcz and 
Szekeres 2004 [DIRS 174073]. At pH 8.5 and an ionic strength of 0.06 M, the stability ratio, 
Wexp, for montmorillonite suspensions was 10. There is no information on the stability of 
montmorrillonite at a pH above 8.5. However, as the model does not consider the Schulze- 
Hardy phenonemon, the model is conservative. Schindler et al. (2004 [DIRS 174507] 
demonstrated that the pHpzc on uranyl oxide phases will depend on composition and varies 
depending on the crystallographic plane exposed to solution. 
Table 7-2. pHpzc of Iron Oxides 
Mineral pH 
a-Fe2O3 (hematite) 8.5 
a-FeOOH (goethite 7.3 
Fe2O3H2O 7.5 
a-Fe3O4 (magnetite) 6.6 
Source: (Davis and Kent [DIRS 143280]). 
Conclusion: The use of stability criterion based upon ionic strength and pH is well documented 
and is a standard scientific practice within the literature as discussed in Sections 6.3.2.1 
and 6.3.2.2. The dependence of colloid suspension stability upon pH and ionic strength is 
reasonably consistent with project data and literature data. Therefore, this component is 
validated through corroboration with literature and project data (VA 1 and 3). 
Question B. To increase confidence with colloid concentration parameters, is the use of 
colloid concentration derived from waste form degradation reasonable (VA N/A)? 
The model provides parameters to account for colloids derived from waste form degradation. 
Immersion corrosion tests on crushed or powdered waste forms are run to allow observation of 
changes within experimental time frames that under disposal conditions will be many orders of 
magnitude slower. The fast reaction leads to condensation colloid formation processes in the 
leachate. In more complex designed tests, filtration or reduced water contact can lower colloid 
formation. There is also a difference between the theoretical colloidal size range and the 
achieveable experimentally measured colloid range and population. It is important to consider 
the testing conditions and analysis techniques when evaluating colloid formation from waste 
form corrosion tests. The use of product consistency test data for DHLW glass-derived colloids 
is conservative. The use of parameters for this purpose is reasonable and does not require 
validation. The validation of the model output is contained in Sections 7.1.2 and 7.1.3. 
Question C. Does the model of irreversible sorption of plutonium and americium onto iron 
oxide reasonably represent or bound the literature and project data (VA N/A)? 
The model of irreversible sorption of plutonium and americium onto iron oxide bounds the 
literature and YMP data provided in Section 6.3.3.2. 

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Question D. Is the reversible sorption of radionuclides corroborated and accepted as 
established fact within industry literature (VA 1 and/or VA 3)? 
Reversible sorption of radionuclides is a well documented and accepted as established fact 
within the literature. Therefore, this component is validated through corroboration with literature 
(VA 3). The appropriateness of the distribution coefficient parameters is discussed in 
Section 7.1.5. 
7.1.2 Validation of Model Results for Colloid and Plutonium Concentrations 
Question E: Are the concentrations of colloids and plutonium consistent with 
scientific principles and observed colloid concentrations in the Project and peerreviewed 
literature? (VA 1 and 3) 
DHLW Glass-Derived Colloids 
The DHLW glass test data set used in the model can be validated by comparison with data 
generated from similar alkali borosilicate glass published data from other testing programs. 
Corrosion tests on the plutonium-, americium-, and thorium-doped alkali-borosilicate waste 
glasses (Mnard et al. 1998 [DIRS 171053], Pirlet 2001 [DIRS 174351], Vernaz and Godon 
1992 [DIRS 174503]) suggest that plutonium was strongly bonded to all colloidal particles in the 
leachate. Furthermore, 70% to 100% of lanthanides and thorium were associated with colloids 
(on 0.45-m and 1.8-nm filters), while uranium and neptunium retention vary from 40% in a 
5  10-3 carbonate medium to 95% in a 5  10-3 M phosphate solution. The R7T7 glasses were 
doped with different quantities of actinides for testing (Mnard et al. 1998 [DIRS 171053]). The 
actinide level in the alpha-doped glasses was 0.85% and is much higher than those used in the 
SRL glass static tests (Bates et al. 1995 [DIRS 170880]) or anticipated in current U.S. high-level 
waste production glasses and the R7T7 French production waste glasses (Table 7-3; 
Advocat et al. 2001 [DIRS 163198]). DHLW glass with LaBS plutonium-bearing glass will 
result in a higher total plutonium content for the borosilicate glasses that will be closer to that 
used in corrosion tests reported by Pirlet (2001 [DIRS 174351]) and Mnard et al. (1998 
[DIRS 171053]). 
Tests reported by Pirlet (2001 [DIRS 174351]) on 239Pu-doped R7T7 glass are comparable to the 
static tests on SRL glasses used in model development and exhibited similar long-term behavior. 
However, the tests were performed with a glass surface area-to-leachate volume (SA/V) of 50/m, 
which may result in less overall reaction. The actinides were initially released congruently with 
other glass matrix elements; however, as the corrosion progressed, an amorphous gel layer 
formed on the glass surface and retained most of the actinides despite further glass reaction. 
Over 80% of the plutonium and 99% of the americium were found as particulates rather than 
solution species. Actinide retention factors were found to increase with increasing temperature, 
which would be consistent with decreasing plutonium-colloid stability with temperature. 

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Table 7-3. Actinide Levels in Borosilicate Glasses Used for Testing 
Oxide 
(wt %) 
R7T7- 
239Pu 
Dopeda 
SRL-202Ab SRL-131Ab R7T7 
Productionc 
DWPF 
Productiond 
PuO2 0.85 1.3  10-2 1.2  10-2 ~7.8  10-3 2.2  10-4 to 
5.8  10-2 
ThO2 0 0.28 0.01 N/A 0.011 to 0.55 
AmO2 0 4.4  10-4 4.5  10-4 N/A N/A 
NpO2 0 1.4  10-2 1.3  10-2 N/A N/A 
UO2 0 2.0 3.0 0.52 1.01 to 2.88 
Source: aMnard et al. 1998 [DIRS 171053], Table 1 ; Vernaz and 
Godon 1992 [DIRS 174503], p. 39, Table 1 ; 
bBates et al. 1995 [DIRS 170880], Table 1b ; 
cAdvocat et al. 2001 [DIRS 163198], Table 1; 
dFowler 2003 [DIRS 164361], Table 3, p. 12. 
NOTES: The levels of plutonium, uranium, and thorium anticipated to be present in 
HLW production glasses is similar to that used in glass corrosion tests used 
in the model development. 
N/A = not available. 
In spite of the large plutonium loading, Mnard et al. (1998 [DIRS 171053]) were unable to 
provide supporting evidence for precipitated actinides associated with phyllosilicate 
(montmorillonite) colloids because examination of microscopic characterization of the leachates 
was limited in scope. Plutonium concentrations in static leach tests conducted for one year in 
distilled water with an SA/V of 50 /m resulted in peak plutonium concentrations of 1.5  10-7 M 
with 239Pudoped R7T7 glasses (Pirlet 2001 [DIRS 174351]; however, results from the 
238Pu-doped glass tests did not appear to reach a saturation level after one year of reaction 
(Figure 7-1). The 238Pu glass reaction is dominated by the radiolytic field and is not relevant to 
long-term DHLW glass disposal. 
Source: Pirlet et al. 2001 [DIRS 174351], p. 48, Figure 1; Vernaz and Godon 1992 [DIRS 174503]), p. 43, Table 3. 
Figure 7-1. Total Plutonium and Americium Concentration from Static Tests on Actinide-Doped R7T7 
Borosilicate Glass Conducted at 90C with an SA/V of 50/m. 

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MDL-EBS-PA-000004 REV 02 7-8 August 2005 
Spent Nuclear Fuel-Derived Colloids 
Direct comparison of plutonium colloid data from spent nuclear fuel tests by Grambow et al. 
(1996 [DIRS 113253]), Wilson (1990 [DIRS 100793]), McNamara et al. (2005 [DIRS 174068]), 
and Mertz et al. (2003 [DIRS 162032]) is difficult because of experimental differences. Values 
reported by Grambow et al. (1996 [DIRS 113253]) for plutonium colloids ranged from 
10-12 mol/L to 10-8 mol/L between pH values 7 and 8 for spent nuclear fuel tests. These values 
are similar to those reported by Wilson (1990 [DIRS 100793]) for tests with cladding and by 
Mertz et al. (2003 [DIRS 162032]) for intermittent drip tests. In the intermittent drip tests, the 
water contact mode and the amount of reaction is not known (the latter problem is common to all 
spent nuclear fuel corrosion tests). It is possible that the intermittent drip tests under report the 
levels of actinide colloids because of filtration by the Zircaloy holder; however, in immersion 
tests by Wilson (1990 [DIRS 100793]), the presence of any Zircaloy cladding (including 
defected cladding) reduced the observed plutonium concentrations by three to four orders of 
magnitude. Indeed, with cladding, the majority of the plutonium was in solution and not present 
as colloids supporting a possible filtration mechanism. A change in the design of the holder for 
the intermittent drip tests on spent nuclear fuel (Mertz et al. 2003 [DIRS 162032]) increased the 
release of plutonium, although the amounts remained between 10-8 mol/L to 10-10 mol/L. Test 
data suggest that estimates of plutonium colloids from tests on bare fragments or pellets of fuel 
will be conservative. 
Under Yucca Mountain-relevant conditions, filtration of colloids is likely to occur in the Tuff 
matrix depending on the degree of saturation, and the diameter of pores in the rock compared to 
the dimensions of the colloids (Nuttall 1986 [DIRS 121867]). Kingston and Whitbeck (1991 
[DIRS 113930]) report colloid concentrations from 24 springs and wells in central and southern 
Nevada, in a region underlain by carbonate rocks. The majority of the sites were within 50 km 
of the Nevada Test Site (NTS). Colloid concentrations ranged from 0.28 mg/L to 1.35 mg/L; 
three sites have high colloid concentrations ranging from 6.48 mg/L to over 25 mg/L. However, 
there are questions regarding the methods used to obtain these solutions that may have resulted 
in overreporting colloid concentrations (Triay et al. 1995 [DIRS 100789]). 
ConclusionsThe concentration of plutonium associated with DHLW glass may depend on the 
initial concentration of plutonium in the glass; however, the concentration of plutonium colloids 
from 239Pu-doped R7T7, SRL131A, and SRL-202A tests does not exceed 5  10-7 M. The 
concentration of plutonium colloids from spent nuclear fuel is likely to be low in the presence of 
Zircaloy materials or iron. The concentration of colloids in natural groundwaters reported in the 
model is likely to be conservative owing to the collection methods. The corroborating data on 
plutonium colloids match qualitatively and are bounded by model predictions. 
7.1.3 Validation of CSNF and DSNF Model Results 
Question F: Are the commercial and DOE-owned SNF model approaches 
consistent with commercial and DOE-owned SNF-specific test results and peerreviewed 
literature on field observations near uranium mines and deposits? (VA3) 
The model does not include waste form colloids from the degradation of CSNF and DSNF. This 
approach has been examined by comparison with available literature data. 

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Wronkiewicz et al. (1996 [DIRS 102047]) observed suspended particles of uranophane 
(Ca(UO2)2(SiO3OH)25H2O) and uraninite (UO2) within the leachates of drip tests on 
unirradiated UO2; however, similar uranophane suspended particles were not observed on the 
two high drip rate tests on spent nuclear fuel (Mertz et al. 2003 [DIRS 162032]). This may have 
been due to the design differences between these intermittent drip tests, which resulted in holdup 
of particles on the Zircaloy holder used in the spent nuclear fuel intermittent drip tests. 
Studies by van Middlesworth and Wood (1998 [DIRS 170927]) at the Cretaceous Idaho batholith 
represents possibly the most intriguing natural analogue site for colloids derived from waste 
forms. At this deposit, hydrothermal alteration of the Casto granite containing allanite [(Ca, Mn, 
Ce, La, Y)2(Fe2+Fe3+,Al)3OOH(Si2O7)(SiO4)] to phosphate phases, such as monazite [(La, 
Ce)PO4] and rhabdophane [(La,Ce)PO42H2O], is almost directly analogous to the laboratory 
weathering experiments with alkali borosilicate glasses (Buck and Bates 1999 [DIRS 109494]; 
DTN: MO0407ANLGNN02.608 [DIRS 171277]). Many of these rare earth-containing 
phosphate minerals have been found to be associated with silicates, including clays 
(Taunton et al. 2000 [DIRS 170891]; van Middlesworth and Wood 1998 [DIRS 170927]). The 
analysis of filtered and unfiltered samples from the Idaho batholith indicated that majority of the 
rare earths and thorium was associated with colloids, while most of the uranium was soluble 
(Table 7-4). An important aspect of the study by van Middlesworth and Wood (1998 
[DIRS 170927]) was the application of ferric hydroxide coprecipitation for the quantitative 
analysis of trace levels uranium, thorium, and rare earths in their solutions which supports the 
contention that attenuation of actinides in the waste package will occur by reaction with iron 
corrosion products. 
Table 7-4. Colloid Data on Rare Earths and Actinides from Locations in the Cretaceous Idaho Batholith 
Total (Rare Earths) 
(mol/L) 
Thorium 
(mol/L) 
Uranium 
(mol/L) 
Weir Creek 
Unfiltered 3.9  10-9 1.4  10-10 4.2  10-11 
Filtered 1.1  10-9 <4  10-11 1.4  10-8 
Horse Creek 
Unfiltered 8.2  10-9 9.2  10-10 8.4  10-9 
Filtered 3.3  10-9 2.0  10-10 8.0  10-9 
Big Creek 
Unfiltered 1.0  10-8 6.3  10-10 2.1  10-10 
Filtered 5.6  10-9 2.7  10-10 1.4  10-10 
Sharkey 
Unfiltered 1.5  10-8 5.5  10-10 7.6  10-11 
Filtered 1.1  10-9 <4  10-11 6.3  10-11 
Source: van Middlesworth and Wood 1998 [DIRS 170927] 
(units converted from Table 2, p. 870). 
DSNF constitutes <3% of the total repository waste inventory. The N-Reactor fuel comprises 
approximately 85% of the DSNF inventory (DOE 2002 [DIRS 158405]) and is the surrogate for 
DSNF in the TSPA-LA model. Uranium metal is unstable in water and will rapidly oxidize 
depending on the temperature. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 7-10 August 2005 
ConclusionThe corroborating data match qualitatively with commercial and DOE-owned 
spent nuclear fuel -specific test results and peer-reviewed literature on field observations near 
uranium mines and deposits and are in agreement with model predictions. 
7.1.4 Validation of Corrosion Product and Groundwater Colloid Concentration Model 
Results 
Question G: Are the ranges of corrosion product and groundwater colloid 
concentrations reasonable (VA 1 and/or VA 3)? 
Honeyman and Ranville (2002 [DIRS 161657]) point out that although most systems of interest 
for radioactive waste disposal exhibit relatively low colloid concentrations (less than 10 mg/L), 
concentrations exceeding 100 mg/L are not uncommonly reported. Based on a review by 
Triay et al. (1995 [DIRS 100789]), high reported colloid concentrations in groundwaters may be 
the result of colloid sampling techniques. Hence, colloid concentrations between 50 mg/L and 
200 mg/L have been assigned only a sampling probability of 2% when the ionic strength of the 
solution is less than 0.05 M within the waste form and engineered barrier system components of 
the TSPA-LA model calculations. 
Iron oxyhydroxide colloids are expected to form as a result of internal and external corrosion of 
the waste package and associated engineered barrier system components (Section 6.3.1.3). There 
are few, if any, peer-reviewed publications that present information on iron oxideoxyhydroxide 
colloid concentrations as a function of aqueous chemistry. For the colloid model, a reasonable 
estimate was made based on expert judgment, corroborated by laboratory experiments 
investigating the corrosion of small-scale waste package replicas, and by maximum colloid 
concentrations observed in groundwater. 
The range of values for maximum iron oxideoxyhydroxide colloids concentration (0.05 mg/L to 
50 mg/L) was chosen so that it was substantially greater than iron oxyhydroxide colloid levels 
found in groundwater, but high enough to reflect the masses of corroded ferrous metal 
anticipated in the repository. Additionally, the value acknowledges the high end of the range of 
colloids in general reported in groundwaters. 
Experiments examining the corrosion of scaled waste packages suggest that abundant colloids 
will be produced (as much as 20 mg/L in the first four weeks of the experiment) 
(DTN: MO0312UCC034JC.008 [DIRS 166367]). Further, it has been demonstrated in every 
case in which a hole has been placed in a canister to simulate a breach that the hole, after passing 
quantities of corrosion particles, quickly plugs with corrosion product and effectively seals 
against further outflows. 
ConclusionThe stipulation that continuous production and outflow of iron oxide 
oxyhydroxide colloids in the waste package would continue unimpeded is, therefore, considered 
conservative and the concentrations recommended in the model area also conservative. The 
ranges of corrosion product and groundwater colloid concentrations in the corroborating data 
match qualitatively and are bounded by model predictions. Therefore, the model is validated 
through corroboration (VA 1 and VA 3). 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 7-11 August 2005 
7.1.5 Confidence Building for Distribution Coefficient Values 
Question H: Are the distribution coefficients for radionuclides onto colloids (Kd 
values) consistent with peer-reviewed literature and analytical modeling results 
(VA N/A)? 
The Kd value ranges chosen for reversibly attached radionuclides are based on laboratory results 
and peer-reviewed literature, which are described in detail in Section 6.3.3. Validation of Kd 
values for several radionuclides adopted in the abstraction, and particularly validation of the use 
of published literature, was accomplished by considering alternative calculations for determining 
effective Kd values. Beckman et al. (1988 [DIRS 144956]) reported sorption studies for cesium 
on zeolites and silicate phases (Figure 7-2 and Table 7-5). These are similar to those reported in 
Section 6. Table 6-6 lists the radionuclides to which Kd values have been assigned in this report. 
Tables 6-5, 7-6, and 7-7 list colloid types and Kd values based on the results and literature review 
(Section 6.3.3.1). Table 7-8 lists Kd values that are the results of modeling described by 
Honeyman and Ranville (2002 [DIRS 161657], Figure 8-6). As described in Section 6.3.3.1, 
Yucca Mountain vicinity-specific colloid concentrations and groundwater chemistry by Kingston 
and Whitbeck (1991 [DIRS 113930]) were used as input to a mechanistic sorption model and 
surface complexation calculations. The Kd ranges vary in part as a function of colloid 
concentration, which ranged from 0.1 mg/L to 100 mg/L. 
Source: Adapted from data reported by Beckman et al. 1988 [DIRS 144956]. 
NOTE: The distribution of Kd values for cesium on smectite is bound within this data obtained for related silicate 
mineral phases. 
Figure 7-2. Cumulative Distribution Function Showing the Distribution of Cesium Sorption Coefficients 
Obtained on Clay, Zeolite, and Glass 
Table 7-5. Cesium Sorption with 95% Lower- and Upper-Confidence Bounds (ml/g) 
Type/ ID Lower Estimate Upper 
D G1-1883 1.25E+02 1.85E+02 2.75E+02 
D G1-1882 6.87E+02 1.20E+03 2.10E+03 
D G1-2363 2.66E+02 3.95E+02 5.86E+02 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 7-12 August 2005 
Table 7-5. Cesium Sorption with 95% Lower- and Upper-Confidence Bounds (ml/g) (Continued) 
Type/ ID Lower Estimate Upper 
D JA-32 8.40E+01 1.01E+02 1.23E+02 
D JA-37 5.24E+02 6.39E+02 7.80E+02 
D YM-22 2.12E+02 3.14E+02 4.66E+02 
D YM-45 2.75E+02 4.80E+02 8.39E+02 
D YM-54 1.16E+02 1.61E+02 2.22E+02 
D JA-18 1.06E+04 1.30E+04 1.60E+04 
D YM-49 2.21E+04 3.28E+04 4.86E+04 
D G1-2289 1.98E+04 2.94E+04 4.36E+04 
D YM-38 9.64E+03 1.33E+04 1.84E+04 
D YM-48 8.58E+03 1.50E+04 2.62E+04 
D G1-3658 3.34E+04 4.95E+04 7.35E+04 
Source : Beckman et al. 1988 [DIRS 144956]. 
Sorption of cesium onto montmorillonite was investigated by Serne and Relyea (1982 
[DIRS 147681] over a range of pH values and ionic strengths (see Figure 7-3). The results 
demonstrate the effect of salt concentration on the sorption capacity of clays owing to ion 
exchange reactions and virtually no pH dependence on sorption. Clay formed during DHLW 
glass reaction would also result in a large quantity of sodium, which would block sites in the 
smectite clays and reduce overall sorption of cesium. 
Source: Serne and Relyea (1982 [DIRS 147681]) p. 45, Figure 6. 
Figure 7-3. Schematic Plot Adapted from Data Reported for Cesium Sorption on Montmorillonite with 
pH and NaCl Concentration. 
Two compendia of sorption data are available. The National Cooperative for the Disposal of 
Radioactive Waste (NAGRA, Switzerland) (Stenhouse 1995 [DIRS 147477]) developed one to 
compile a set of Kd values useful for evaluation of waste disposal in hypothetical sites in 
Switzerland. The sorbents considered were crystalline rock, marl, and bentonite with reducing 
groundwaters (or pore water, in the case of bentonite). Ionic strengths are similar to (or greater 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 7-13 August 2005 
than) J-13 well water. The information compiled by Stenhouse (1995 [DIRS 147477]) needs to 
be applied carefully to the YMP because Kd values are only applicable to a single set of 
environmental conditions. 
Stenhouse (1995 [DIRS 147477]) has proposed realistic values and conservative values to 
account for uncertainty in oxidation states (composite Kd values were reduced by an order of 
magnitude). With respect to the oxidation state, the conservative values may be more relevant to 
Yucca Mountain. However, the realistic values are useful to compensate for the use of a massbased 
Kd approach on colloids. In Table 7-6, the rock composite Kd values and individual Kd 
values for smectite and oxides or quartz are compiled. The individual mineral values are 
probably more suitable to consider for Yucca Mountain colloid Kd values. 
Table 7-6. Summary of Sorption Data Developed by National Cooperative for the Disposal of 
Radioactive Waste for Bentonite, Crystalline Rock, and Marl 
Radionuclide and 
Oxidation State(s) Database 
NAGRA Rock Kd 
(mL/g) (page) NAGRA Mineral Kd (mL/g) 
Th(IV) 79 entries; 
22 references 
5,000 
(p. Th-7) 
Smectite = 2,000; 
Oxides, Quartz = 1,000 
Pa(V) 21 entries; 
12 references 
1,000 
(p. Pa-3) 
NAGRA uses Zr(IV) as analogue for Pa(IV); 
use Np(V) for Pa(V) for YMP 
U(VI) 382 entries; 
32 references 
5,000 
(p. U-7) 
Smectite = 100; 
Oxides, Quartz = 20 
Np(V) 389 entries; 
53 references 
5,000 
(p. Np-8) 
Smectite = 100; 
Oxides, Quartz = 20 
Pu(IV, V, VI) 298 entries; 
25 references 
5,000 
(p. Pu-10) 
Smectite = 1,000; 
Oxides, Quartz = 200 
Am(III) 198 entries; 
34 references 
5,000 
(p. Am-9) 
Smectite = 10,000; 
Oxides, Quartz = 1,000 
Cesium 516 entries; 
41 references 
1,000 
(p. Cs-11) 
Smectite = 2,000; 
Iron oxide = 40 
Source: Stenhouse 1995 [DIRS 147477]. 
NOTE: NAGRA = National Cooperative for the Disposal of Radioactive Waste. 
A second useful compendium was developed for the U.S. Environmental Protection Agency 
(EPA 1999 [DIRS 147475]), which compiles information on radionuclides including thorium, 
uranium, and plutonium, with information on neptunium and americium reported in its third 
volume (EPA 2004 [DIRS 172215]. The assembled data are interpreted to predict ranges of Kd 
values for soils in shallow subsurface environments. Redox conditions for that system are 
oxidizing, which makes it useful for the redox-sensitive radionuclides at the YMP. 
The EPA (1999 [DIRS 147475]; 2004 [DIRS 172215]) studies reviewed published literature to 
determine the predominant aqueous chemical conditions affecting sorption in soils for thorium, 
uranium, plutonium, americium, strontium, and cesium. For thorium, pH and sorbate 
concentration are important. At concentrations close to its solubility limit, Kds tend to increase. 
For uranium and neptunium, the pH was important. For plutonium, carbonate complexation has 
a strong influence on sorption, so carbonate concentration, as well as clay content, must be 
considered as criteria to select a relevant Kd. In Table 7-7, Kd values have been expressed in a 
minimum to maximum format, with no indication of distribution. Some ranges, for example 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 7-14 August 2005 
uranium, span almost six orders of magnitude. To narrow the range, the geometric means were 
calculated below and reported in column 4 of Table 7-7. 
Table 7-7. Summary of Sorption Data Developed by U.S. Environmental Protection Agency 
Radionuclide 
and Oxidation 
State(s) Database 
Soil and pH Categories 
20 wt % to 60 wt % clay (high clay) 
<4 wt % (low clay) 
Kd (mL/g) 
Range of Geometric Means 
Th(IV) 17 entries pH 8 to 10; [Th] <10-9 mol/L 
(pH and Th concentration important) 
20 to 2,000 (200) 
U(VI) 20 references pH 8 
(pH important) 
0.4 to 250,000 (316) 
Pu(IV, V, VI) 2 references high clay (51 wt % to 70 wt %); 3-4 
meq/L CO3
2-; 
low clay (<4 wt %); 3 to 4 meq/L CO3
2- 
1,860 to 2,550 (2,178); 
80 to 470 (194) 
Np(V) clay 
(pH important) 
0.4 to 2,575 (55) 
Am(III) 11 observations Clay 
(pH important) 
25 to 400,000 (8,400) 
Cesium 11 references high clay (20 wt % to 60 wt %) 
low clay (<4 wt %) 
80 to 26,700 (1,500); 
10 to 3,500 (200) 
Source: EPA 1999 [DIRS 147475]; EPA 2004 [DIRS 172215] 
Honeyman and Ranville (2002 [DIRS 161657]) reviewed aspects of colloid-facilitated 
radionuclide transport, including reduction of effective retardation of radionuclides by colloids 
and the effectiveness of colloids with respect to radionuclide mobility in the presence of 
stationary phases available for radionuclide sorption. The colloid-stationary phase system is 
discussed further in Section 6.3.3.2 in the context of irreversible sorption. Contardi et al. (2001 
[DIRS 162732]) developed models using Yucca Mountain groundwater chemistry and colloid 
size and concentration data from Kingston and Whitbeck (1991 [DIRS 113930]). Kd values were 
calculated with the use of a diffuse layer model and based on three principles: (1) the system is 
symmetrical (i.e., the colloids and stationary phases are similar with respect to mineralogy); 
(2) the colloids are stable; and (3) the system is in sorptive equilibrium. The ratio of stationary 
phase to colloid surface areas ranged up to roughly two orders of magnitude (3.2 to 63). Colloid 
concentrations of up to about 100 mg/L were included. In the system analyzed, approximate 
Kd value ranges (mL/g) for five radionuclides were given (Table 7-8). 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 7-15 August 2005 
Table 7-8. Modeled Kd Values for Plutonium, Americium, Thorium, Neptunium, and Uranium Sorption 
onto Yucca Mountain-Vicinity Colloids 
Radionuclide Sorbate and 
Oxidation State(s) at YMP Kd Values, mL/g 
U(VI) 1  100 to 6  102 
Np(V) 1  101 to 1  102 
Pu(V) 1  103 to 1  104 
Th(IV) 2  103 to 9  104 
Am(III) 1  104 to 1  107 
Source: Honeyman and Ranville 2002 [DIRS 161657] (extracted 
from Figure 7-5 in reference). 
ConclusionThe distribution coefficients for radionuclides onto colloids (Kd values) in 
Tables 7-5 through 7-8 are bounded by the model values in Table 6-6. The data in the model has 
been shown to be consistent with peer-reviewed literature and analytical modeling results. 
Corroborating data match qualitatively and are bounded by model predictions. Therefore, the 
model is validated through corroboration (VA 1 and VA 3). 
7.1.6 Post Model Development Validation Conclusion 
Except for question C, the preceding sections have met the validation criteria set forth in the 
technical work plan (BSC 2005 [DIRS 173246], Table 2-3) that the corroborating data must 
match qualitatively or must be bound by model predictions. 
7.2 PROCESSES NOT INCORPORATED INTO THE ABSTRACTION 
Sorption to the airwater interfaces in unsaturated environments was not included. Attachment of 
colloids to airwater interfaces depends on the hydrophobicity of the colloidal particles, which is 
generally low in smectites and hematite. Because partitioning to immobile air-water interfaces 
will tend to retard colloid transport (see Section 6.5.3), exclusion of this process in modeling 
colloid transport in the EBS is conservative. 
Neither the effects of filtration nor gravitational settling processes on colloid concentrations as 
they are transported through the EBS are included. This exclusion will result in use of 
conservatively bounding colloid concentrations for radionuclide transport in the EBS. 
The role of minor phases in the colloid-assisted radionuclide transport is not included in the 
model. In particular, the small amount of manganese present in waste package materials could 
lead to sorption of plutonium at lower pH than for iron oxides. Sorption of Pu(IV) on to 
manganese oxide minerals has been shown to be fast and reversible (Duff et al. 1998 
[DIRS 170925]). Sorption on manganese phases increased at pH 3 and reached a maximum 
value of 100% at pH 8 and then decreased over the pH range from 8 to 10. The plutonium 
sorption appeared to occur via a reduction mechanism on the mineral surface and was consistent 
with the formation of an inner-sphere complex (Duff et al. 1998 [DIRS 170925]). No evidence of 
PuO2 precipitation was detected. Although manganese minerals are excluded, by applying a 
conservative pH independent Kd distribution, the potential effect of manganese oxide minerals is 
bounded. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 7-16 August 2005 
Sorption of metals to organic matter can be greater than sorption to oxide phases (Ramsey 1988 
[DIRS 144603]). Organic films on colloids will modify uptake of plutonium and other actinides. 
Organics may result from thermal decomposition of grease and paint. Microbial activity was 
discussed in the report but is not addressed within the model analysis. Microbes can bind metals, 
secrete metal-solubilizing compounds, and transform metal ions to altered speciation states. 
Owing to a lack of suitable data, predicting the fate of radionuclide transport processes 
associated with biocolloids is limited. However, the quantities of organic matter and microbes 
either emplaced or generated (through microbial growth processes) in the waste emplacement 
drifts are sufficiently small that they are unlikely to be of significance. 
The potential effect of radiolysis has been ignored and there is little data on the subject. 
Radiolysis can result in more-soluble states for the actinides, which might otherwise form 
colloids. Other effects include degradation of organic materials. 
Sorption coefficients are dependent on pH with a sharp sorption edge observed near pH values 8 
and 4 for plutonium and americium sorption onto hematite, respectively. However, J-13 well 
water was used in most sorption tests upon which the model was based. 
ConclusionProcesses not included in the model would reduce the impact of colloids in TSPALA; 
hence, the model bounds processes that were not incorporated into the abstraction. 
7.3 VALIDATION ACTIVITIES EXTENDING BEYOND CURRENT 
DOCUMENTATION OF THE MODEL 
The disposition of the various validation criteria described in Section 7.1 that the level of 
confidence required for validation of the various components considered in this model analysis 
has been achieved. No further validation activities are needed to complete the model validation 
for its intended use. 

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MDL-EBS-PA-000004 REV 02 8-1 August 2005 
8. CONCLUSIONS 
8.1 MODEL SUMMARY AND IMPLEMENTATION 
The colloids process models predict the potential concentrations of these colloids in the waste 
package from degradation of the waste forms and the corrosion of waste package materials and 
the concentrations of colloids present in natural waters in the vicinity of Yucca Mountain 
that could potentially come in contact with the repository waste. In addition, stability and 
sorptiondesorption mechanisms and transport characteristics of colloids anticipated in the 
repository are addressed, based on the range in geochemical conditions anticipated in the 
repository environment within the regulatory compliance period. 
Based on corrosion tests on waste glass, the major waste-form colloid type anticipated is 
smectite clay. This is supported by waste glass tests and validated by similar borosilicate glass 
corrosion studies by others (Mnard et al. 1998 [DIRS 171053]). Based on a few observations of 
radionuclide-bearing particles being associated with clay-like phases, the model has assumed that 
plutonium and americium are irreversibly attached (embedded) particles, such that the smectite 
(montmorillonite) clay dictates the radionuclide release behavior. Wilson (1990 [DIRS 100793]) 
and McNamara et al. (2005 [DIRS 174068]) observed plutonium-bearing colloids in spent 
nuclear fuel immersion tests. In the presence of even defected Zircaloy materials, plutonium 
concentrations observed by Wilson (1990 [DIRS 100793]) were so low that plutonium colloids 
may not have been present. Supporting data on spent nuclear fuel intermittent drip tests by 
Mertz et al. (2003 [DIRS 162032]) suggests that plutonium concentrations are extremely low. 
Although the exact reasons for these reductions in plutonium concentrations remain uncertain, it 
is likely that in a waste package environment, plutonium colloids from spent nuclear fuel will be 
low. 
The TSPA-LA colloid model has been constructed in such a way that a large fraction of 
dissolved plutonium and americium sorbed onto iron oxyhydroxide colloids is irreversibly 
sorbed (i.e., becomes permanently attached) and the remainder is reversibly sorbed according to 
an equilibrium Kd value model with a small fraction occurring as dissolved species. Further, a 
parameter (the target-flux out ratio, FRN) defines the ratio of plutonium and americium mass 
sorbed to colloids (irreversibly and reversibly) to the total plutonium and americium mass 
leaving the waste package (sorbed to colloids plus dissolved) for advective conditions. This ratio 
ranges between 0.9 and 0.99. The experimentally based forward rate constant (k) defines the 
irreversible sorption rate of plutonium and americium onto iron oxyhydroxide colloids and 
stationary corrosion products for zero or near zero advective flux conditions or where colloids 
are unstable. 
In the TSPA-LA model, natural seepage water/groundwater colloids are modeled to come in 
contact with and enter breached waste packages and will subsequently sorb radionuclides. The 
resulting pseudocolloid (radionuclide complexes) will subsequently be subjected to transport 
from the engineered barrier system to the downstream unsaturated and saturated zones. 
Furthermore, in contrast to the modeling effort for the TSPA-SR, wherein colloid concentration 
parameters were based upon data from groundwaters around the world (Degueldre et al. 2000 
[DIRS 153651]; DTN: SNT05080598002.001 [DIRS 162744]), more recent data from 
groundwater samples collected from wells within the Yucca Mountain area and Idaho National 

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MDL-EBS-PA-000004 REV 02 8-2 August 2005 
Laboratory are used to develop the parameters used for the TSPA-LA. These data suggested 
higher seepage watergroundwater colloids concentrations (perhaps as high as one to two orders 
of magnitude greater than those used previously in the TSPA-SR) and are appropriate for the 
TSPA-LA model calculations. 
Recent experimental work on the corrosion of miniature steel waste packages have provided 
needed data on the probable concentrations of iron oxideoxyhydroxide colloids that might be 
expected within the engineered barrier system upon contact of seepage water and subsequent 
corrosion of the waste packages. These data indicate that values as high as 50 mg/L should be 
sampled in the TSPA-LA model calculations. These high values represent a reasonable 
maximum for iron oxideoxyhydroxide colloid releases from a waste package following a 
potential breach (Section 6.3.2.2). 
There are several sources of uncertainty attached to this abstraction. The reduction of the waste 
package colloid environment to two major types of particles was necessary given the complexity 
of the issue and modeling constraints. The surface electrical heterogeneity in the corroded waste 
package environment may not be represented fully by adopting measurements of zeta potentials 
from the two types of ideal phases (Elimelech et al. (2000 [DIRS 171066]). For instance, the 
presence of manganese in the corroding metal components may have a strong influence on 
plutonium chemistry (Duff et al. 1999 [DIRS 170925]). The role of spent nuclear fuel in primary 
colloid generation has not been dealt with directly in this model. For instance, in vapor corrosion 
tests on CSNF (which were not used in the development of this model), rinds were observed 
containing fine-grained plutonium-bearing particles and uranium oxyhydroxides. Possible 
spallation of such surface corrosion layers could result in the formation of plutonium-bearing 
waste form colloids. 
The data that is used to provide information on the formation of primary colloids from waste 
forms are not directly comparable because of differences in test designs. These differences 
include, water contact mode, waste-form surface area, the total amount of corrosion and the 
corrosion rate, and incorporation or absence of metallic filters. Particles of U(VI) phases and 
uraninite were observed on the Zircaloy holder at 0.8 years in an ATM106 HDR test 
(Buck et al. 1997 [DIRS 112904]; Finn et al. 1998 [DIRS 100392]). In UO2 drip tests containing 
metallic grids, uranophane (uranium silicate) particles were observed in the leachates 
(Wronkiewicz et al. 1996 [DIRS 102047]) and were detected on filter residues on immersion 
tests on spent nuclear fuel by Wilson (1990 [DIRS 100793]). Such U(VI) silicate particles may 
remain stable suspensions indefinitely under Yucca Mountain groundwater conditions, while 
uranium oxide hydrates would be predicted to be thermodynamically unstable in this 
environment. These types of particles were not considered in the model; however, the acicular 
(needle-like) shape might result in filtration of these types of potential colloids. 
These uncertainties are thought to have been bounded through the application of conservative 
values for colloid stability ranges (high ionic strength cutoff), colloid concentrations, and 
sorption coefficients (large Kd ranges) for this model. The abstraction is considered valid and 
usable in TSPA-LA calculations for any time after the temperature in the repository has 
decreased to below boiling after the thermal pulse. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 8-3 August 2005 
8.2 MODEL OUTPUTS 
Developed data for use directly by the TSPA-LA calculations or other submodel components in 
the TSPA-LA model calculations include intermediate parameters (see Table 6-12 for parameter 
listing and associated DTNs) and model output parameters (Table 8-1). 
The parameters calculated as intermediate results within the TSPA-LA model calculations are 
listed in Table 6-12. 
Table 8-1. Summary of Model Output Parameters 
Parameter 
Name Parameter Description Units Basis for Value/ Source 
Value / Range / 
Distribution 
CRncoll,wf,embed 
CRncoll,wf,rev 
CRncoll,FeOx,irrev 
CRncoll,FeOx,rev 
CRncoll,gw 
Concentration of mobile 
colloidal radionuclide element 
radionuclides per unit volume 
or mass of water 
mol/L Determined in performance 
assessment calculation; 
Figure 6-30 
Calculated at each 
time step; uncertainty 
propagated through 
the TSPA-LA model 
calculations 
Additionally, parameters developed in this analysis for use elsewhere in the TSPA-LA model are 
listed in Table 6-12. These include waste form colloid mass concentrations (derived from 
DHLW glass), corrosion product colloid mass concentrations, Kd values for selected 
radionuclides onto smectite and iron oxideoxyhydroxide colloids, specific surface area on iron 
oxideoxyhydroxide colloids, and ionic strength thresholds for iron oxideoxyhydroxide colloid 
stability (Ilo-thresh,coll,FeOx, Ihi-thresh,coll,FeOx). Developed DTNs include: SN0306T0504103.005 
(groundwater colloid concentration parameters), SN0306T0504103.006 (sorption partition 
coefficients, Kd values), SN0309T0504103.010 (parameters related to corrosion products and 
iron oxideoxyhydroxide colloids), and SN0311T0504103.011 (waste form colloid parameters) 
(Table 6-12). 
8.3 YUCCA MOUNTAIN REVIEW PLAN ACCEPTANCE CRITERIA 
Yucca Mountain Review Plan, Final Report (NRC 2003 [DIRS 163274]) contains Acceptance 
Criteria intended to establish the basis for the review of the material contained in the License 
Application. As this report serves, in part, as the basis for the License Application, it is important 
to show how the information contained herein addresses each of the applicable Acceptance 
Criteria. 
The Acceptance Criteria applicable to this report are identified in Technical Work Plan for 
Postclosure Waste Form Modeling (BSC 2005 [DIRS 173246], Table 3-1). For each applicable 
criterion, the criterion is quoted in italics, followed by pointers to where within the model report 
the information addressing the criterion can be found. In some cases, the criterion is only 
partially addressed in this report. A demonstration of full compliance requires a review of 
multiple reports. 

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MDL-EBS-PA-000004 REV 02 8-4 August 2005 
Radionuclide Release Rates and Solubility Limits (NRC 2003 [DIRS 163274], 
Section 2.2.1.3.4.3) 
Acceptance Criterion 1System Description and Model Integration Are Adequate 
(1) Total system performance assessment adequately incorporates important 
design features, physical phenomena, and couplings, and uses consistent and 
appropriate assumptions throughout the radionuclide release rates and solubility 
limits abstraction process. 
The colloid model developed in this model report is consistent with the design features relevant 
to models from which input is taken and models to which output is provided. The values (or 
ranges of values) of parameters used in the colloid model are either developed within the model 
or obtained from project or literature sources. Implementation of the colloid model requires 
definition of variables for solution pH, and ionic strength that will be obtained from other 
models. 
(2) The abstraction of radionuclide release rates and solubility limits uses 
assumptions, technical bases, data, and models that are appropriate and 
consistent with other related U.S. Department of Energy abstractions. For 
example, the assumptions used for this model abstraction are consistent with the 
abstractions of Degradation of Engineered Barriers (Section 2.2.1.3.1); 
Mechanical Disruption of Waste Packages (Section 2.2.1.3.2); Quantity and 
Chemistry of Water Contacting Waste Packages and Waste Forms (Section 
2.2.1.3.3); Climate and Infiltration (Section 2.2.1.3.5); and Flow Paths in the 
Unsaturated Zone (Section 2.2.1.3.6). The descriptions and technical bases 
provide transparent and traceable support for the abstraction of radionuclide 
release rates and solubility limits; 
Coordination in modeling of colloids within the repository and natural system is discussed in 
Section 6.5.3. 
(3) The abstraction of radionuclide release rates and solubility limits provides 
sufficient, consistent design information on waste packages and engineered 
barrier systems. For example, inventory calculations and selected radionuclides 
are based on the detailed information provided on the distribution (both spatially 
and by compositional phase) of the radionuclide inventory, within the various 
types of high-level radioactive waste; 
Design information on the in-package materials and waste forms was used in determining which 
colloid types to model. 
(4) The U.S. Department of Energy reasonably accounts for the range of 
environmental conditions expected inside breached waste packages and in the 
engineered barrier environment surrounding the waste package. For example, 
the U.S. Department of Energy should provide a description and sufficient 
technical bases for its abstraction of changes in hydrologic properties in the near 
field, caused by coupled thermal-hydrologic-mechanical-chemical processes; 

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The colloid models account for the range of environmental conditions (pH, temperature, and 
ionic strength) expected, as described in Section 6.5. 
(5) ... 
Not applicable (applies to thermal-hydrologic processes). 
(6)  Technical bases for inclusion of any thermal-hydrologic-mechanicalchemical 
couplings and features, events, and processes in the radionuclide 
release rates and solubility limits model abstraction are adequate. For example, 
technical bases may include activities, such as independent modeling, laboratory 
or field data, or sensitivity studies; 
The technical bases for inclusion of features, events, and processes in the colloid model are 
discussed in Section 6. 
(7) ... 
Not applicable (applies to criticality). 
(8) Guidance in NUREG-1297 and NUREG-1298 or other acceptable approaches 
for peer reviews and data qualification is followed. 
Section 4.1 and Attachment II address data inputs to the model and qualification of data. 
Acceptance Criterion 2Data Are Sufficient for Model Justification 
(1) Geological, hydrological, and geochemical values used in the license 
application are adequately justified. Adequate description of how the data were 
used, interpreted, and appropriately synthesized into the parameters is provided; 
The inputs are discussed in Section 4.1. Section 6 discusses how these inputs were developed 
into parameters. 
(2) Sufficient data have been collected on the characteristics of the natural system 
and engineered materials to establish initial and boundary conditions for 
conceptual models and simulations of thermal-hydrologic-chemical coupled 
processes. For example, sufficient data should be provided on design features, 
such as the type, quantity, and reactivity of materials, that may affect 
radionuclide release for this abstraction; 
Section 6 describes the data on the characteristics of the natural system and engineered materials 
that are important to the formation and stability of colloid suspensions, and the attachment of 
radionuclides to colloids. 
(3) ... 
Not applicable (applies to solubility rather than colloids). 

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(4) The corrosion and radionuclide release testing program for high-level 
radioactive waste forms intended for disposal provides consistent, sufficient, and 
suitable data for the in-package and in-drift chemistry used in the abstraction of 
radionuclide release rates and solubility limits. For expected environmental 
conditions, the U.S. Department of Energy provides sufficient justification for the 
use of test results, not specifically collected from the Yucca Mountain site, for 
engineered barrier components, such as high-level radioactive waste forms, drip 
shield, and backfill. 
The corrosion and radionuclide release-testing program for high-level radioactive waste forms 
has provided chemistry and colloid data as well. These data, augmented, by literature data, form 
the basis for the colloid model as discussed in Section 6. 
Acceptance Criterion 3Data Uncertainty Is Characterized and Propagated Through the 
Model Abstraction 
(1) Models use parameter values, assumed ranges, probability distributions, and 
bounding assumptions that are technically defensible, reasonably account for 
uncertainties and variabilities, and do not result in an under-representation of the 
risk estimate; 
Parameter values, assumed ranges, probability distributions, and bounding assumptions are 
discussed in Sections 5 and 6. 
(2) Parameter values, assumed ranges, probability distributions, and bounding 
assumptions used in the abstractions of radionuclide release rates and solubility 
limits in the total system performance assessment are technically defensible and 
reasonable based on data from the Yucca Mountain region, laboratory tests, and 
natural analogs. For example, parameter values, assumed ranges, probability 
distributions, and bounding assumptions adequately reflect the range of 
environmental conditions expected inside breached waste packages; 
Parameter values and uncertainty are discussed in Sections 4.1 and 6.6.2 and throughout the 
document. 
(3) ... 
Not applicable (applies to thermal-hydrologic-chemical coupled processes). 
(4) Uncertainty is adequately represented in parameter development for 
conceptual models, process models, and alternative conceptual models 
considered in developing the abstraction of radionuclide release rates and 
solubility limits, either through sensitivity analyses or use of bounding analyses. 
Uncertainty is addressed in Section 6.6.2 and throughout the document. 
(5-6) ... 

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Not applicable (applies to water flow and criticality). 
(7) The U.S. Department of Energy uses as appropriate range of time-history of 
temperature, humidity, and dripping to constrain the probability for microbial 
effects, such as production of organic by-products that act as complexing ligands 
for actinides and microbially enhanced dissolution of the high-level radioactive 
waste glass form. 
The effects of microbes are discussed in Section 6.3.4. 
(8) The U.S. Department of Energy adequately considers the uncertainties, in the 
characteristics of the natural system and engineered materials, such as the type, 
quantity, and reactivity of material, in establishing initial and boundary 
conditions for conceptual models and simulations of thermal-hydrologic-chemical 
coupled processes that affect radionuclide release. 
Uncertainty is addressed in Section 6.6.2 and throughout the document. 
(9) ... 
Not applicable (applies only when insufficient data exists). 
Acceptance Criterion 4 - Model Uncertainty Is Characterized and Propagated Through the 
Model Abstraction 
(1) Alternative modeling approaches of features, events, and processes are 
considered and are consistent with available data and current scientific 
understanding, and the results and limitations are appropriately considered in the 
abstraction. 
Alternative modeling approaches are discussed in Section 6.4. 
(2) In considering alternative conceptual models for radionuclide release rates 
and solubility limits, the U.S. Department of Energy uses appropriate models, 
tests, and analyses that are sensitive to the processes modeled for both natural 
and engineering systems. Conceptual model uncertainties are adequately defined 
and documented, and effects on conclusions regarding performance are properly 
assessed. For example, in modeling flow and radionuclide release from the drifts, 
the U.S. Department of Energy represents significant discrete features, such as 
fault zones, separately, or demonstrates that their inclusion in the equivalent 
continuum model produces a conservative effect on calculated performance. 
Alternative modeling approaches are discussed in Section 6.4. 
(3) Consideration of conceptual model uncertainty is consistent with available 
site characterization data, laboratory experiments, field measurements, natural 
analog information and process-level modeling studies; and the treatment of 

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conceptual model uncertainty does not result in an under-representation of the 
risk estimate. 
Alternative modeling approaches are discussed in Section 6.4. 
(4) ... 
Not applicable (applies to thermal-hydrologic-chemical coupled processes). 
Acceptance Criterion 5 Model Abstraction Output Is Supported by Objective 
Comparisons. 
(1) The models implemented in this total system performance assessment 
abstraction provide results consistent with output from detailed process-level 
models and/or empirical observations (laboratory and field testing and/or natural 
analogs). 
As discussed in Section 7, the colloid model is consistent with empirical observations. 
(2) ... 
Not applicable (applies to thermal hydrologic models). 
(3) ... 
Not applicable (applies to thermal-hydrologic-chemical coupled processes). 
(4) ... 
Not applicable (applies to the performance confirmation program). 

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Thompson, J.L. 1999. Migration of Plutonium in Ground Water at the Nevada Test 
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109521 Kim, J.I. 1994. Actinide Colloids in Natural Aquifer Systems. MRS Bulletin, 19, (12), 
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113930 Kingston, W.L. and Whitbeck, M. 1991. Characterization of Colloids Found in Various 
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Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 9-7 August 2005 
106449 Laaksoharju, M.; Degueldre, C.; and Skarman, C. 1995. Studies of Colloids and Their 
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100051 Langmuir, D. 1997. Aqueous Environmental Geochemistry. Upper Saddle River, New 
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168991 Leenaers, A.; Sannen, L.; Van den Berghe, S.; and Verwerft, M. 2003. Oxidation of 
Spent UO2 Fuel Stored in Moist Environment. Journal of Nuclear Materials, 317, ([2- 
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109524 Liang, L. and Morgan, J.J. 1990. Chemical Aspects of Iron Oxide Coagulation in 
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52, (1), 32-55. Basel, Switzerland: Birkhauser Verlag. TIC: 246125. 
101876 Lide, D.R., ed. 1995. CRC Handbook of Chemistry and Physics. 76th Edition. Boca 
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161667 Litaor, M.I. and Ibrahim, S.A. 1996. Plutonium Association with Selected Solid 
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172864 Liu, C.; Cox, D.S.; Barrand, R.D.; and Hunt, C.E.L. 1992. Particle Size Distributions 
of U3O8 Produced by Oxidation in Air at 300-900C. Proceeding of the 13th Annual 
Conference, Saint John, NB, June 7-10, 1992. 1, 1-22. [Toronto, Ontario, Canada]: 
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166315 Lu, N.; Conca, J.; Parker, G.R.; Leonard, P.A.; Moore, B.; Strietelmeier, B.; and Triay, 
I.R. 2000. Adsorption of Actinides onto Colloids as a Function of Time, Temperature, 
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5121. Los Alamos, New Mexico: Los Alamos National Laboratory. 
ACC: MOL.20031204.0108. 
170883 Lu, N.; Reimus, P.W.; Parker, G.R.; Conca, J.L.; and Triay, I.R. 2003. Sorption 
Kinetics and Impact of Temperature, Ionic Strength and Colloid Concentration on the 
Adsorption of Plutonium-239 by Inorganic Colloids. Radiochimica Acta, 91, 713-720. 
Mnchen, Germany: Oldenbourg Wissenschaftsverlag. TIC: 256462. 
100946 Lu, N.; Triay, I.R.; Cotter, C.R.; Kitten, H.D.; and Bentley, J. 1998. Reversibility of 
Sorption of Plutonium-239 onto Colloids of Hematite, Goethite, Smectite, and Silica. 
LA-UR-98-3057. Los Alamos, New Mexico: Los Alamos National Laboratory. 
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170884 Luckscheiter, B. and Nesovic, M. 2002. Sorption Behaviour of Am on Precorroded 
HLW Glass in Water and Brines. Radiochimica Acta, 90, 537-541. Mnchen, 
Germany: Oldenbourg Wissenchaftsverlag. TIC: 256461. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 9-8 August 2005 
108215 McCarthy, J.F. and Degueldre, C. 1993. Sampling and Characterization of Colloids 
and Particles in Groundwater for Studying Their Role in Contaminant Transport. 
Environmental Particles. Buffle, J. and van Leeuwen, H.P., eds. Environmental 
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100778 McCarthy, J.F. and Zachara, J.M. 1989. Subsurface Transport of Contaminants. 
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Pennsylvania: American Chemical Society. TIC: 224876. 
113270 McEachern, R.J. and Taylor, P. 1998. A Review of the Oxidation of Uranium Dioxide 
at Temperatures Below 400C. Journal of Nuclear Material, 254, 87-121. Amsterdam, 
The Netherlands: Elsevier. TIC: 246427. 
174068 McNamara, B.; Hanson, B.; Buck, E.; Soderquist, C. 2005. Corrosion of Commercial 
Spent Nuclear Fuel. 2. Radiochemical Analyses of Metastudtite and Leachates. 
Radiochimica Acta, 93, (2005), 169-175. Mnchen, Germany: Oldenbourg 
Wissenschaftsverlag. TIC: 257131. 
171053 Mnard, O.; Advocat, T.; Ambrosi, J.P.; and Michard, A. 1998. Behavior of Actinides 
(Th, U, Np and Pu) and Rare Earths (La, Ce and Nd) During Aqueous Leaching of a 
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13, 105-126. New York, New York: Pergamon. TIC: 256470. 
162161 Mertz, C.; Fortner, J.; Goldberg, M.; and Shelton-Davis, C. 2000. Colloid Generation 
from Metallic Uranium Fuel. Argonne, Illinois: Argonne National Laboratory. 
TIC: 254296. 
162032 Mertz, C.J.; Finch, R.J.; Fortner, J.A.; Jerden, J.L., Jr.; Yifen, T.; Cunnane, J.C.; and 
Finn, P.A. 2003. Characterization of Colloids Generated from Commercial Spent 
Nuclear Fuel Corrosion. Activity Number: PAWTP30A. Argonne, Illinois: Argonne 
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100801 Minai, Y.; Choppin, G.R.; and Sisson, D.H. 1992. Humic Material in Well Water from 
the Nevada Test Site. Radiochimica Acta, 56, 195-199. Munchen, Germany: R. 
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113272 Murakami, T.; Ohnuki, T.; Isobe, H.; and Sato, T. 1997. Mobility of Uranium during 
Weathering. American Mineralogist, 82, 888-899. Washington, D.C.: Mineralogical 
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163274 NRC (U.S. Nuclear Regulatory Commission) 2003. Yucca Mountain Review Plan, 
Final Report. NUREG-1804, Rev. 2. Washington, D.C.: U.S. Nuclear Regulatory 
Commission, Office of Nuclear Material Safety and Safeguards. TIC: 254568. 
121867 Nuttall, H.E. 1986. Population Balance Model for Colloid Transport NNWSI 
Report/Milestone R318. LA-UR 86-1914. Los Alamos, New Mexico: Los Alamos 
National Laboratory. ACC: NNA.19891102.0024. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 9-9 August 2005 
174071 Painter, S.; Cvetkovic, V.; Pickett, D.; Turner, D.R. 2002. Significance of Kinetics for 
Sorption of Inorganic Colloids: Modelling and Experiment Interpretation Issues. 
Environmental Science & Technology, 36, (2002), 5369-5375. Washington, D.C.: 
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174361 Parks, G.A. 1964. The Isoelectric Points of Solid Oxides, Solid Hydroxides, and 
Aqueous Hydroxo Complex Systems. Chemical Reviews, [65], ([2]), 177-198. 
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110223 Pearcy, E.C.; Prikryl, J.D.; and Leslie, B.W. 1995. Uranium Transport Through 
Fractured Silicic Tuff and Relative Retention in Areas with Distinct Fracture 
Characteristics. Applied Geochemistry, 10, 685-704. Oxford, United 
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174351 Pirlet, V. 2001. Overview of Actinides (Np, Pu, Am) and Tc Release from Waste 
Glasses: Influence of Solution Composition. Journal of Nuclear Materials, 298, 47- 
54. Elsevier Science. TIC: 257455 
171519 Poulton, S.W.; Krom, M.D.; and Raiswell, R. 2004. A Revised Scheme for the 
Reactivity of Iron (Oxyhydr)Oxide Minerals Toward Dissolved Sulfide. Geochimica 
et Cosmochimica Acta, 68, (18), 3703-3715. New York, New York: Pergamon. 
TIC: 256498. 
171518 Roden, E.E. and Zachara, J.M. 1996. Microbial Reduction of Crystalline Iron(III) 
Oxides: Influence of Oxide Surface Area and Potential for Cell Growth. 
Environmental Science & Technology, 30, (5), 1618-1628. Washington, 
D.C.: American Chemical Society. TIC: 256499. 
144603 Ramsay, J.D.F.; Avery, R.G.; and Russell, P.J. 1988. "Physical Characteristics and 
Sorption Behaviour of Colloids Generated from Cementitious Systems." Radiochimica 
Acta, 44/45, 119-124. Mnchen, Germany: R. Oldenbourg Verlag. TIC: 226318. 
170890 Romanovski, V.V.; Brachmann, A.; Palmer, C.E.; Shaw, H.F.; Bourcier, W.L.; and 
Jardine, L.J. 1999. Characterization of Pu Colloidal and Aqueous Species in Yucca 
Mountain Groundwater Surrogate. UCRL-JC-134460. Livermore, 
California: Lawrence Livermore National Laboratory. ACC: MOL.20040930.0291. 
168432 Runde, W.; Conradson, S.D.; Efurd, D.W.; Lu, N.P.; VanPelt, C.E.; and Tait, C.D. 
2002. Solubility and Sorption of Redox-Sensitive Radionuclides (Np, Pu) in J-13 
Water from the Yucca Mountain Site: Comparison between Experiment and Theory. 
Applied Geochemistry, 17, (6), 837-853. New York, New York: Pergamon. 
TIC: 254046. 
170923 Santschi, P.H.; Roberts, K.A.; and Guo, L. 2002. Organic Nature of Colloidal 
Actinides Transported in Surface Water Environments. Environmental Science & 
Technology, 36, (17), 3711-3719. Washington, D.C.: American Chemical Society. 
TIC: 256463. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 9-10 August 2005 
174507 Schindler, M.; Mutter, A.; Hawthorne, F.C.; and Putnis, A. 2004. Prediction of Crystal 
Morphology of Complex Uranyl-Sheet Minerals. I. Theory. The Canadian 
Mineralogist, 42, (6), 1629-1649. Ottawa, Ontario, Canada: Mineralogical Association 
of Canada. TIC: 257517. 
147681 Serne, R.J. and Relyea, J.F. 1982. The Status of Radionuclide Sorption-Desorption 
Studies Performed by the WRIT Program. PNL-3997. Richland, Washington: Pacific 
Norwest Laboratory. ACC: NNA.19900416.0121. 
170928 Sondi, I.; Bican, J.; and Pravdic, V. 1996. Electrokinetics of Pure Clay Minerals 
Revisited. Journal of Colloid and Interface Science, 178, 514-522. New York, New 
York: Academic Press. TIC: 256468. 
174080 Spino, J.; Vennix, K.; Coquerelle, M. 1996. Detailed characterisation of the rim 
microstructure in PWR fuels in the burn-up range 40-67 GWD/tM. Journal of Nuclear 
Materials, 231, (1996), 179-190. Elsevier Science B.V. TIC: 257447. 
147477 Stenhouse, M.J. 1995. Sorption Databases for Crystalline, Marl and Bentonite for 
Performance Assessment. NAGRA Technical Report 93-06. Wettingen, 
Switzerland: National Cooperative for the Disposal of Radioactive Waste. 
TIC: 247885. 
170891 Taunton, A.E.; Welch, S.A.; and Banfield, J.F. 2000. Microbial Controls on Phosphate 
and Lanthanide Distributions During Granite Weathering and Soil Formation. 
Chemical Geology, 169, 371-382. New York, New York: Elsevier. TIC: 256459. 
125815 Taylor, P.; Wood, D.D.; and Owen, D.G. 1995. Microstructures of Corrosion Films on 
UO2 Fuel Oxidized in AirSteam Mixtures at 225C. Journal of Nuclear Materials, 
223, ([3]), 316-320. New York, New York: Elsevier. TIC: 256013. 
163048 Thomas, E. 2003. Transmittal of Unsaturated Testing of Bare Spent UO2 Fuel 
Fragments: Data Report, Argonne National Laboratory. Interoffice memorandum 
from E. Thomas (BSC) to J.C. Cunnane, July 2, 2003, 0702037939, with attachment. 
ACC: MOL.20030702.0116; MOL.20030311.0097. 
112690 Tombcz, E.; Abraham, I.; Gilde, M.; and Szanto, F. 1990. The pH-Dependent 
Colloidal Stability of Aqueous Montmorillonite Suspensions. Colloids and Surfaces, 
49, 71-80. Amsterdam, The Netherlands: Elsevier. TIC: 246046. 
174073 Tombcz, E; Szekeres, M. 2004. Colloidal Behavior of Aqueous Montmorillonite 
Suspensions: the Specific Role of pH in the Presence of Indifferent Electrolytes. 
Applied Clay Science, 27, (2004), 75-94. New York, New York: Elsevier. 
TIC: 257450. 
171770 Torstenfelt, B.; Andersson, K.; and Allard, B. 1982. Sorption of Strontium and Cesium 
on Rocks and Minerals. Chemical Geology, 36, 128-137. Amsterdam, The 
Netherlands: Elsevier. TIC: 256583. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 9-11 August 2005 
172072 Traexler, K.A.; Utsunomiya, S.; Kersting, A.B.; and Ewing, R.C. 2004. Colloid 
Transport of Radionuclides: Yucca Mountain Performance Assessment. Scientific 
Basis for Nuclear Waste Management XXVII, Symposium held April 12-16, 2004, San 
Francisco, CA, USA. 807, 1-6. Warrendale, Pennsylvania: Materials Research 
Society. TIC: 256677. 
100789 Triay, I.; Simmons, A.; Levy, S.; Nelson, S.; Nuttall, H.; Robinson, B.; Steinkampf, W.; 
and Viani, B. 1995. Colloid-Facilitated Radionuclide Transport at Yucca Mountain. 
LA-12779-MS. Los Alamos, New Mexico: Los Alamos National Laboratory. 
ACC: NNA.19930628.0067. 
170927 van Middlesworth, P.E. and Wood, S.A. 1998. The Aqueous Geochemistry of the Rare 
Earth Elements and Yttrium. Part 7. REE, Th and U contents in Thermal Springs 
Associated with the Idaho Batholith. Applied Geochemistry, 13, (7), 861-884. New 
York, New York: Pergamon. TIC: 256470. 
114428 van Olphen, H. 1977. An Introduction to Clay Colloid Chemistry for Clay 
Technologists, Geologists, and Soil Scientists. 2nd Edition. New York, New 
York: John Wiley & Sons. TIC: 208918. 
174503 Vernaz, E. and Godon, N. 1992. Leaching of Actinides from Nuclear Waste Glass: 
French Experience. Scientific Basis for Nuclear Waste Management XV, Symposium 
held November 4-7, 1991, Strasbourg, France. Sombret, C.G., ed. 257, 37-48. 
Pittsburgh, Pennsylvania: Materials Research Society. TIC: 204618. 
108285 Wan, J. and Tokunaga, T.K. 1997. Film Straining on Colloids in Unsaturated Porous 
Media: Conceptual Model and Experimental Testing. Environmental Science & 
Technology, 31, (8), 2413-2420. Washington, D.C.: American Chemical Society. 
TIC: 234804. 
161732 Wilkinson, K.J.; Ngre, J.-C.; and Buffle, J. 1997. Coagulation of Colloidal Material 
in Surface Waters: The Role of Natural Organic Matter. Journal of Contaminant 
Hydrology, 26, (1-4), 229-243. New York, New York: Elsevier. TIC: 253708. 
102147 Wilson, C.N. 1985. Results from NNWSI Series 1 Spent Fuel Leach Tests. HEDL-TME 
84-30. Richland, Washington: Hanford Engineering Development Laboratory. 
ACC: HQS.19880517.2579. 
171996 Wilson, C.N. 1986. Test Plan for Series 3 NNWSI Spent Fuel Leaching/Dissolution 
Tests. HEDL-7577. Washington, D.C.: U.S. Department of Energy. 
ACC: NNA.19900216.0069. 
102150 Wilson, C.N. 1987. Results from Cycles 1 and 2 of NNWSI Series 2 Spent Fuel 
Dissolution Tests. HEDL-TME 85-22. Richland, Washington: Hanford Engineering 
Development Laboratory. ACC: HQS.19880517.2581. 
121808 Wilson, C.N. 1990. Indications for the Formation of Pu, Am, and Cm Colloids in 
Semi-Static Spent Fuel Dissolution Tests. 1990 Spent Fuel Workshop, Gull Harbour, 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 9-12 August 2005 
Manitoba, Canada, Spetember 3-5, 1990. Richland, Washington: Pacific Northwest 
Laboratory. ACC: NNA.19920302.0064. 
100793 Wilson, C.N. 1990. Results from NNWSI Series 3 Spent Fuel Dissolution Tests. 
PNL-7170. Richland, Washington: Pacific Northwest Laboratory. 
ACC: NNA.19900329.0142. 
174895 Wittman, R.S.; Buck, E.C.; and Hanson, B.D. 2005. Data Analysis of Plutonium 
Sorption on Colloids in a Minimal Kinetics Model. PNNL-15285. Richland, 
Washington: Pacific Northwest National Laboratory. ACC: MOL.20050811.0087. 
102047 Wronkiewicz, D.J.; Bates, J.K.; Wolf, S.F.; and Buck, E.C. 1996. Ten-Year Results 
from Unsaturated Drip Tests with UO2 at 90C: Implications for the Corrosion of 
Spent Nuclear Fuel. Journal of Nuclear Materials, 238, (1), 78-95. Amsterdam, The 
Netherlands: North-Holland. TIC: 243361. 
171238 Zarrabi, K.; McMillan, S.; Elkonz, S.; and Cizdziel, J. 2003. Corrosion and Mass 
Transport Processes in Carbon Steel Miniature Waste Packages. Document TR-03- 
003, Rev. 0. Task 34. Las Vegas, Nevada: University of Nevada, Las Vegas. 
ACC: MOL.20040202.0079. 
170892 Zhao, P. and Steward, S.A. 1997. Literature Review of Intrinsic Actinide Colloids 
Related to Spent Fuel Waste Package Release Rates. UCRL-ID-126039. Livermore, 
California: Lawrence Livermore National Laboratory. ACC: MOL.19971212.0585. 
9.2 CODES, STANDARDS, AND REGULATIONS 
173273 10 CFR 63. 2005 Energy: Disposal of High-Level Radioactive Wastes in a Geologic 
Repository at Yucca Mountain, Nevada. ACC: MOL.20050405.0118. 
AP-2.22Q, Rev. 1, ICN 1. Classification Analyses and Maintenance of the Q-List. 
Washington, D.C.: U.S. Department of Energy, Office of Civilian Radioactive Waste 
Management. ACC: DOC.20040714.0002. 
LP-2.29Q-BSC, Rev. 0, ICN 1. Planning for Science Activities. Washington, D.C.: U.S. 
Department of Energy, Office of Civilian Radioactive Waste Management. 
ACC: DOC.20050718.0002. 
LP-3.15Q-BSC, Rev. 0, ICN 2. Managing Technical Product Inputs. Washington, 
D.C.: U.S. Department of Energy, Office of Civilian Radioactive Waste Management. 
ACC: DOC.20050801.0010. 
LP-SIII.2Q-BSC, Rev. 0. Qualification of Unqualified Data. Washington, D.C.: U.S. 
Department of Energy, Office of Civilian Radioactive Waste Management. 
ACC: DOC.20050119.0002. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 9-13 August 2005 
LP-SIII.10Q-BSC, Rev. 0, ICN 1. Models. Washington, D.C.: U.S. Department of 
Energy, Office of Civilian Radioactive Waste Management. 
ACC: DOC.20050623.0001. 
105725 ASTM C 1174-97. 1998. Standard Practice for Prediction of the Long-Term Behavior 
of Materials, Including Waste Forms, Used in Engineered Barrier Systems (EBS) for 
Geological Disposal of High-Level Radioactive Waste. West Conshohocken, 
Pennsylvania: American Society for Testing and Materials. TIC: 246015. 
LP-SI.11Q-BSC, Rev. 0, ICN 1. Software Management. Washington, D.C.: U.S. 
Department of Energy, Office of Civilian Radioactive Waste Management. 
ACC: DOC.20041005.0008. 
9.3 SOURCE DATA, LISTED BY DATA TRACKING NUMBER 
149232 LA0002SK831352.001. Total Colloidal Particles Concentration and Size Distribution 
in Groundwaters from the Nye County Early Warning Drilling Program. Submittal 
date: 02/24/2000. 
149194 LA0002SK831352.002. Total Colloidal Particles Concentration and Size Distribution 
in Groundwaters Around Yucca Mountain. Submittal date: 02/25/2000. 
161771 LA0002SK831352.003. Colloid Size Distribution and Total Colloid Concentration in 
INEEL Groundwater Samples. Submittal date: 02/25/2000. 
161579 LA0002SK831352.004. Colloid Size Distribution and Total Colloid Concentration in 
Groundwater's from the Nuclear Test Site. Submittal date: 02/25/2000. 
148751 LA0003AM831341.001. Probability Distributions for Sorption Coefficients (Kds). 
Submittal date: 03/29/2000. 
149172 LA0003NL831352.001. Experimental Data on Sorption and Desorption Amounts for 
Plutonium Onto Clay Colloids. Submittal date: 03/16/2000. 
148526 LA0003NL831352.002. The KD Values of 239Pu on Colloids of Hematite, Ca- 
Montmorillonite and Silica in Natural and Synthetic Groundwater. Submittal 
date: 03/29/2000. 
150272 LA0004NL831352.001. Pu(IV) and Pu(V) Sorption on Hematite and Goethite Colloids 
with Natural and Synthetic J-13 Water (1997-98 Data). Submittal date: 04/04/2000. 
149623 LA0005NL831352.001. The Kd Values of 243Am on Colloids of Hematite, 
Montmorillonite and Silica in Natural and Synthetic Groundwater. Submittal 
date: 05/03/2000. 
161580 LA0211SK831352.001. Colloid Concentration and Size Distribution from Crater Flats 
Wells. Submittal date: 12/04/2002. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 9-14 August 2005 
161581 LA0211SK831352.002. Colloid Concentration and Size Distribution from Nye County 
Early Warning Drilling Program, Alluvial Testing Complex (ATC) Wells. Submittal 
date: 12/04/2002. 
161582 LA0211SK831352.003. Colloid Concentration and Size Distribution from Nellis/Desert 
Research Institute (DRI) Wells. Submittal date: 12/04/2002. 
161458 LA0211SK831352.004. Colloid Concentration and Size Distribution from Nye County 
Early Warning Drilling Program, Fall 2000 Field Sampling. Submittal 
date: 12/04/2002. 
163196 LA0304PR831232.001. Calculation of Ionic Strengths Based on Cation Concentrations 
in Various Groundwater Samples in Which Colloid Concentrations Were Measured. 
Submittal date: 05/13/2003. 
163197 LA0304PR831232.002. Calculation of Ionic Strengths Based on Non-Q Cation 
Concentrations in Various Groundwater Samples in Which Colloid Concentrations 
Were Measured. Submittal date: 05/13/2003. 
170621 LA0407AM831341.002. Batch Sorption Coefficient Data for Cesium on Yucca 
Mountain Tuffs in Representative Water Compositions. Submittal date: 07/12/2004. 
142466 LA9910SK831341.003. Sorption Plutonium (V) of Ferrihydrite, Goethite, and 
Hematite. Submittal date: 12/07/1999. 
144991 LA9910SK831341.005. Total Colloidal Particles Concentration and Size Distribution 
in NTS-ER-20-5-1, NTS-ER-20-5-3, and J-13 Groundwater. Submittal 
date: 12/07/1999. 
152621 LL000905312241.018. Data Associated with the Detection and Measurement of 
Colloids Recorded in Scientific Notebook 1644. Submittal date: 09/29/2000 
143308 LL000123351021.117. ANL 94/34 - The Effects of the Glass Surface Area/Solution 
Volume Ratio on Glass Corrosion: A Critical Review. Submittal date: 01/28/2000. 
162871 MO0302UCC034JC.003. Graphical X-Ray Diffractometer Data and Mineral Analysis 
of Filtered Solids from Effluent Solution During Miniature Waste Package Corrosion. 
Submittal date: 02/10/2003. 
163910 MO0306ANLSF001.459. Colloids Generated from Irradiated N Reactor Fuel, Data 
Report. Submittal date: 06/04/2003. 
166367 MO0312UCC034JC.008. Mass Transport of Solids in Effluent Solution During 
Miniature Waste Package Corrosion. Submittal date: 12/05/2003. 
171277 MO0407ANLGNN02.608. Normalized Mass Losses of Radionuclides in N2 Tests. 
Submittal date: 07/27/2004. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 9-15 August 2005 
170760 MO0407SEPFEPLA.000. LA FEP List. Submittal date: 07/20/2004. 
162744 SNT05080598002.001. Engineered Barrier System (EBS) Transport Sorption and 
Colloid Parameters for the TSPA-VA (Total System Performance Assessment-Viability 
Assessment) Rip Calculations. Submittal date: 11/06/1998. 
9.4 SOFTWARE CODES 
161572 Software Code: GoldSim. V7.50.100. PC. 10344-7.50.100-00. 
9.5 OUTPUT DATA, LISTED BY DATA TRACKING NUMBER 
SN0306T0504103.005. Revised Groundwater Colloid Mass Concentration Parameters 
for TSPA (Total System Performance Assessment). Submittal date: 06/30/2003. 
SN0306T0504103.006. Revised Sorption Partition Coefficients (Kd Values) for 
Selected Radionuclides Modeled in the TSPA (Total System Performance Assessment). 
Submittal date: 06/30/2003. 
SN0309T0504103.010. Updated Iron Oxyhydroxide Colloid Concentration 
Parameters, Specific Surface Area (SA) of Iron Oxyhydroxide Colloids, Target Flux 
Out Ratio (FRN) and Forward Rate Constant (K) for Pu & Am Sorption to Iron 
Oxyhydroxide Colloids & Stationary Corrosion. Submittal date: 09/18/2003. 
SN0311T0504103.011. Updated Concentration of Embedded Radionuclides (Pu and 
Am) in Waste Form Colloids and Mass Concentrations of Waste Form Colloids. 
Submittal date: 09/18/2003. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 9-16 August 2005 
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Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 I-1 August 2005 
APPENDIX I 
CALCULATION OF GROUNDWATER COLLOID PARAMETERS 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 I-2 August 2005 
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Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 I-3 August 2005 
I. CALCULATION OF GROUNDWATER COLLOID PARAMETERS 
The groundwater colloid parameters for use in the TSPA-LA model were based on a data set 
including 79 groundwater samples collected in the Yucca Mountain vicinity and 11 samples 
collected from the Idaho National Laboratory (Table I-1). Inclusion of the Idaho groundwater 
colloid data (DTN: LA0002SK831352.003 [DIRS 161771]) was deemed appropriate for 
inclusion in the data analysis because the areas climate in (arid) and bedrock geology (volcanic 
fractured basalts) are similar. 
Colloid populations (particles/mL) in the groundwater samples were converted to mass 
concentrations (mg/L) (Figure 6-12, Section 6.3.2.4). 
I.1 CONVERSION OF THE COLLOID POPULATIONS TO MASS 
CONCENTRATIONS 
In order to derive particle size and concentrations from light-scattering instruments, particles 
must be modeled as spherical with uniform refractive index. The slightly higher refractive 
indicies of the iron oxide minerals compared to silicate minerals would have resulted in an 
overestimation of colloid particle populations in the groundwater, if these were present. 
Mass concentrations were obtained by applying a mineral density of 2.5 g/cm3 
(2.5  10-18 mg/nm3). Silicate minerals have particle densities ranging from 2 g/cm3 to 3 g/cm3 
for montmorillonite to 2.67 g/cm3 for quartz (Lide 1995 [DIRS 101876], pp. 4-132 through 
4-138). The iron oxide minerals were not included in the mass determination. The densities of 
hematite, goethite, and lepidocrockite are 5.26, 3.3 to 4.3, and 4.05 to 4.31, respectively 
(Lide 1995 [DIRS 101876], pp. 4-132 through 4-138). This could potentially result in an 
underestimation of particle mass concentrations. 
The volume of a sphere (V) can be calculated from the radius (R) by the following relationship: 
V = 4/3 p R3 (Eq. I-1) 
The colloid population data for each groundwater sample consisted of population data recorded 
at 10-nm increments from the particle-size diameter range (50 nm to 200 nm) as reported in the 
DTNs. The midpoint of the particle diameter size class ranges that are reported in the 
groundwater datasheets was used to establish the radius for the particle-volume calculations 
using Equation I-1. For example, those particles with diameter ranging from 50 nm to 60 nm 
were assigned a size of 55 nm. Thus, the radius (R), for this particle-diameter size population 
was 27.5 nm. 
After calculating an estimate of volume for each particle diameter size class interval, the mass 
concentration (mg/L) for each particle-size interval was calculated by multiplying particle 
volume (nm3) by particle density (mg/nm3); this product was then multiplied by the total number 
of particles in the size class interval to derive an estimate of the total colloid mass in each 
particle-size class interval. The total colloid mass in the water sample (Column 4 in Table I-1) 
was then estimated by summing the mass concentrations in each particle-size class interval. For 
example, the total colloid concentration for the groundwater sample extracted from Well 
NC-EWDP-01s was calculated as outlined in Table I-2. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 I-4 August 2005 
Table I-1. Groundwater Samples Used to Develop Cumulative Distribution Function Developed to 
Establish Colloid Concentration Sampling Frequency in GoldSim Calculations for TSPA-LA 
for Solutions with Ionic Strength less than 0.05 M 
Groundwater 
Sample Well ID Colloid Population 
(pt/mL) 
Colloid 
Concentration 
(mg/L) 
Data Source DTN: LA0002SK831352.001 [DIRS 149232], Alluvium 
1 NC-EWDP-01s, Depth 170 1.07E+07 3.02E-02 
2 NC-EWDP-01s, Depth 250 2.08E+07 4.62E-02 
3 NC-EWDP-03s, Depth 449 3.56E+08 3.48E-01 
4 NC-EWDP-03s, Depth 390 1.78E+09 1.90E+00 
5 NC-EWDP-09SX, Depth 310 3.85E+07 4.49E-02 
6 NC-EWDP-09SX, Depth 270 7.24E+07 8.74E-02 
7 NC-EWDP-09SX, Depth 150 9.73E+07 1.62E-01 
8 NC-EWDP-09SX, Depth 112 7.14E+07 8.89E-02 
9 Nc airport 7.95E+07 1.41E-01 
Data Source DTN : LA0002SK831352.002 [DIRS 149194], Volcanics 
10 UE-25 J-13 1.19E+06* 1.35E-03 
11 UE-25 J-13 1.046E+06* 1.75E-03 
12 USW SD-6 5.07E+08* 7.24E-01 
13 USW SD-6 1.05E+08* 1.43E-01 
14 UE-25 WT #17 2.11E+09* 3.49E+00 
15 UE-25 WT #17 1.475E+09* 2.32E+00 
16 UE-25 WT #17 2.633E+07* 2.69E-02 
17 UE-25 WT #17 8.236E+07* 1.32E-01 
18 UE-25 WT #3 7.78E+06* 2.02E-02 
19 UE-18 WW #8 2.93E+06* 2.79E-03 
20 UE-18 WW #8 3.31E+06* 3.70E-03 
21 U-20 WW #20 5.16E+07* 5.83E-02 
22 U-20 WW #20 5.72E+07* 7.71E-02 
23 UE-29 a #1 7.45E+06* 1.44E-02 
24 UE-29 a #2 3.66E+07* 6.73E-02 
25 UE-25 c #1 4.63E+06* 9.68E-03 
26 UE-25 c #1 2.82E+06* 4.49E-03 
Data Source DTN : LA9910SK831341.005 [DIRS 144991], Volcanics 
27 NTS-ER-20-5-3 1.43E+10 1.02E+01 
28 NTS-ER-20-5-1 1.25E+10 6.74E+00 
29 UE-25 J-13 1.54E+06 3.72E-03 
Data Source DTN: LA0211SK831352.004 [DIRS 161458], Nye Co. Phase 2  Alluvium 
30 NC=EWDP-15P 1.34E+08 1.73E-01 
31 NC-EWDP-7S 3.02E+06 9.01E-03 
32 NC-EWDP-4PB 1.41E+09 2.03E+00 
33 NC-EWDP-4PA 2.34E+08 3.05E-01 
34 NC-EWDP-5SB 9.38E+08 2.69E+00 
35 NC-EWDP-12PA 7.52E+07 1.42E-01 
36 NC-EWDP-12PB 8.54E+07 1.53E-01 
37 NC-EWDP-12PC 2.11E+07 6.29E-02 
38 NC-EWDP-19D 8.27E+07 7.73E-02 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 I-5 August 2005 
Table I-1. Groundwater Samples Used to Develop Cumulative Distribution Function Developed to 
Establish Colloid Concentration Sampling Frequency in GoldSim Calculations for TSPA-LA 
for Solutions with Ionic Strength Less Than 0.05 M (Continued) 
Groundwater 
Sample Well ID Colloid Population 
(pt/mL) 
Colloid 
Concentration 
(mg/L) 
Data Source DTN: LA0211SK831352.002 [DIRS 161581], ATC Wells 
39 ATC-IM1-A 1.01E+10 7.39E+00 
40 ATC-IM1-B 9.51E+09 7.20E+00 
41 NC-EWDP-19IM-2-A 1.71E+10 1.35E+01 
42 NC-EWDP-19IM-2-B 1.76E+10 1.46E+01 
43 NC-EWDP-19IM1-Z1 6.40E+09 5.61E+00 
44 NC-EWDP-19IM1-Z2 3.23E+09 2.70E+00 
45 NC-EWDP-19IM1-Z3 7.73E+09 5.67E+00 
46 NC-EWDP-19IM1-Z4 8.52E+09 6.40E+00 
47 NC-EWDP-19IM1-Z5 2.33E+10 1.93E+01 
48 NC-EWDP-19D-Z5 1.27E+10 1.08E+01 
49 ATC-19D1-011602-1350C 1.01E+10 6.57E+00 
50 ATC-19D1-011702-1215C 9.28E+09 5.64E+00 
51 ATC-19D1-011602-011702 7.33E+10 4.84E+01 
52 ATC-19D1-012802-1500 5.19E+10 3.15E+01 
53 NC-EWDP-19D, Z1 5.13E+09 4.12E+00 
54 NC-EWDP-19D-Z1 8.27E+07 7.74E-02 
55 NC-EWDP-19D, Z2 8.73E+07 9.75E-02 
56 NC-EWDP-19D, Z2 2.29E+08 2.47E-01 
57 NC-EWDP-19D, Z2 2.24E+08 2.24E-01 
58 NC-EWDP-19D 9.34E+08 7.47E-01 
59 NC-EWDP-19D 2.77E+08 2.52E-01 
60 NC-EWDP-19D 3.51E+09 1.87E+00 
Data Source DTN: LA0211SK831352.001 [DIRS 161580], USGS wells at Crater Flat 
61 USGS-VH-1 1.33E+07 2.45E-02 
62 USGS-VH-2 1.18E+06 5.03E-03 
63 USGS-VH-1-A 2.17E+07 2.07E-02 
64 USGS-VH-1-B 1.09E+07 1.34E-02 
Data Source DTN: LA0211SK831352.003 [DIRS 161582] 
65 DRI-EH-2-091200-0800 2.17E+06 4.36E-03 
66 DRI-BLM-091200-0915 3.24E+07 3.93E-02 
67 DRI-TTR3B-091200-1430 4.25E+06 6.77E-03 
68 DRI-Cedar Pass-091200-1325 1.01E+06 2.67E-03 
69 DRI-EH-7-091200-1040 8.58E+05 1.89E-03 
70 DRI-Sandia 7-091300-1030 1.05E+07 1.70E-02 
71 DRI-Sandia 6-091300-0930 6.29E+07 8.29E-02 
72 DRI-Tolicha Peak-091400-0745 6.98E+06 7.62E-03 
73 DRI-Airforce 3A-091300-1500 1.34E+09 1.58E+00 
74 DRI-Roller Coaster-091300-0930 1.12E+06 4.48E-03 
Data Source DTN: LA0002SK831352.003 [DIRS 161771], INEEL Wells 
75 USGS 117-1 4.23E+06 8.97E-03 
76 M7S-1 1.30E+06 2.35E-03 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 I-6 August 2005 
Table I-1. Groundwater Samples Used to Develop Cumulative Distribution Function Developed to 
Establish Colloid Concentration Sampling Frequency in GoldSim Calculations for TSPA-LA 
for Solutions with Ionic Strength Less Than 0.05 M (Continued) 
Groundwater 
Sample Well ID Colloid Population 
(pt/mL) 
Colloid 
Concentration 
(mg/L) 
77 USGS 120-1 3.76E+06 5.80E-03 
78 USGS 119-1 8.13E+06 1.48E-02 
79 M3S-1 1.24E+06 1.97E-03 
80 M1S-1 3.28E+06 5.42E-03 
81 USGS-87-1 2.63E+06 4.63E-03 
82 M10S-1 9.37E+07 1.71E-01 
83 BLR-99-1 7.73E+07 1.38E-01 
84 M14S-1 4.28E+06 8.21E-03 
85 USGS 92-1 8.00E+10 1.72E+02 
Data Source DTN: LA0002SK831352.004 [DIRS 161579], NTS Wells 
86 NTS-ER-20-5-1 1.25E+10 6.74E+00 
87 NTS-ER-20-5-3 1.43E+10 1.02E+01 
88 NTS-U20n, ~2511 ft 2.57E+08 3.20E-01 
89 NTS-U20n, lower zone 5.19E+09 5.47E+00 
90 NTS-U19q 2.72E+10 3.51E+01 
Table I-2. Example Calculation for Total Colloid Mass in Well NC-EWDP-01s, Depth 170 
Particle Diameter 
Size Range (nm) 
Particle Diameter 
(nm) 
Colloid Number 
(pt/mL) 
Calculated Colloid Mass 
(mg/L) 
50 to 60 55 9.27E+05 2.0189E-04 
60 to 70 65 1.01E+06 3.6308E-04 
70 to 80 75 8.64E+05 4.7713E-04 
80 to 90 85 8.01E+05 6.4392E-04 
90 to 100 95 8.51E+05 9.5508E-04 
100 to 110 105 1.14E+06 1.7275E-03 
110 to 120 115 8.01E+05 1.5947E-03 
120 to 130 125 7.26E+05 1.8561E-03 
130 to 140 135 5.26E+05 1.6941E-03 
140 to 150 145 6.26E+05 2.4981E-03 
150 to 160 155 5.76E+05 2.8077E-03 
160 to 170 165 4.51E+05 2.6520E-03 
170 to 180 175 4.13E+05 2.8974E-03 
180 to 190 185 3.13E+05 2.5942E-03 
190 to 200 195 3.63E+05 3.5233E-03 
200 200 3.51E+05 3.6757E-03 
Totals 1.07E+07 3.02E-02 
Source: DTN: LA0002SK831352.001 [DIRS 149232]. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 I-7 August 2005 
I.2 ESTABLISHMENT OF GROUNDWATER COLLOID PARAMETERS FOR USE 
IN THE TSPA-LA MODEL 
Mass concentrations (mg/L) of the groundwater in Table I-1 were pooled and a discrete 
cumulative distribution function was established to evaluate the uncertainty in colloid 
concentration distribution (Figure 6-12, Section 6.3.2.4). Filtered water samples were not 
considered in the development of the cumulative distribution function. Since the colloid 
particles were measured by dynamic light-scattering techniques, any potential error in 
measurement of colloid populations for the various particle size classes due to filter ripening 
(i.e., clogging of filter pores through time during the filtration process) was avoided by using 
only nonfiltered sample data. 
The data shown in Figure 6-12, Section 6.3.2.4 reflect observed variability in groundwater 
colloid concentrations. The goal of the uncertainty distribution for groundwater colloid 
concentrations is to numerically capture our knowledge about uncertainty on a large scale. 
Based on the groundwater colloid concentration collected at the two sites (Yucca Mountain and 
the Idaho National Laboratory) and the tremendous uncertainty associated with the collection of 
groundwater colloids, a reasonable representation of the uncertainty is shown in the cumulative 
distribution function. Major factors that contribute to uncertainty in the concentrations of 
colloids in the groundwater samples include: (1) collection techniques, (2) differences in 
pumping rates at each well, and (3) unknown factors including the types of additives introduced 
in the wells during the drilling process. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 I-8 August 2005 
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Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 II-1 August 2005 
APPENDIX II 
QUALIFICATION OF UNQUALIFIED DATA USED AS DIRECT INPUTS 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
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Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 II-3 August 2005 
II. QUALIFICATION OF UNQUALIFIED DATA USED AS DIRECT INPUTS 
Data from references used as direct input to this model must meet the definition of qualified in 
accordance with LP-3.15Q, Managing Technical Product Inputs. Input data that do not meet the 
definition of qualified data require qualification in accordance with the requirements in 
LP-SIII.10Q-BSC, Models, Section 5.2.1(k), and LP-SIII.2Q-BSC, Qualification of Unqualified 
Data. 
Data obtained from outside sources that are not established facts must be demonstrated to be 
suitable for the specific application. When appropriately justified, these data are considered 
qualified for use within the technical product. The extent to which the data demonstrate 
properties of interest shall be addressed. Two or more of the following factors have been used 
when presenting the case that data are suitable for intended use: 
 Reliability of data source 
 Qualifications of personnel or organizations generating the data 
 Prior uses of the data 
 Availability of corroborating data. 
Data from the sources Listed in Table II-1a are qualified for intended use in accordance with 
LP-SIII.10Q-BSC, Section 5.2.1(k). Data listed in Table II-1b are qualified for intended use in 
accordance with LP-SIII.2Q-BSC. 
Table II-1a. Data Qualified for Direct Input using LP-SIII.10Q-BSC, Section 5.2.1(k) 
DIRS Qualification Title Section Brief Input Description 
109524 Chemical Aspects of Iron Oxide Coagulation in Water: 
Laboratory Studies and Implications for Natural 
Systems. (1990) Liang, L. and Morgan, J. J., Aquatic 
Sciences, 52, (1), 32-55. 
II.1 Hematite colloid 
characteristics 
112690 The pH-Dependent Colloidal Stability of Aqueous 
Montmorillonite Suspensions, (1990) Tombcz et al., 
Colloids and Surfaces, 49, 71-80. 
II.1 Smectite clay colloid 
characteristics 
171518 Microbial Reduction of Crystalline Iron(III) Oxides: 
Influence of Oxide Surface Area and Potential for Cell 
Growth, (1996) Roden, E. E. and Zachara, J. M., 
Environ. Sci. Technol. 30 (5), 1618-1628. 
II.2 Iron Oxide and 
Oxyhdroxide Surface 
Areas 
171519 A Revised Scheme for the Reactivity of Iron 
(Oxyhydr)Oxide Minerals Toward Dissolved Sulfide, 
(2004) Poulton et al., Geochimica Cosmochimica Acta, 
68, 3,703-3,715. 
II.2 Iron Oxide and 
Oxyhdroxide Surface 
Areas 
171770 Sorption of Strontium and Cesium on Rocks and 
Minerals (1982) Torstenfelt et al., Chemical Geology, 
36, 128-137. 
II.2 Cesium Sorption on Iron 
Oxides 
172864 Particle Size Distributions of U3O8 Produced by 
Oxidation in Air at 300-900C. Proceeding of the 13th 
Annual Conference, Saint John, NB, June 7-10, 1992. 
1, 1-22. Liu, C.; Cox, D.S.; Barrand, R.D.; and Hunt, 
C.E.L. 
II.2 Particle-Size 
Distribution 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 II-4 August 2005 
Table II-1b. Data Qualified for Direct Input using LP-SIII.2Q-BSC 
DIRS Qualification Title Section Brief Input 
Description 
166315 Adsorption of Actinides onto Colloids as a Function of Time, 
Temperature, Ionic Strength, and Colloid Concentration, 
Waste Form Colloids Report for Yucca Mountain Program 
(Colloid Data Summary from 1999 to 2000 Research). 
(2000) Lu, N.; Conca, J.; Parker, G.R.; Leonard, P.A.; 
Moore, B.; Strietelmeier, B.; and Triay, I.R. LA-UR-00-5121. 
Los Alamos, New Mexico: Los Alamos National 
Laboratory. ACC: MOL.20031204.0108. 
II.3 Iron oxide surface area 
(53.5 m2/g) and iron 
oxide colloid 
concentration 
(200 mg/L). Americium 
and plutonium sorption 
on smectite clay and 
hematite colloids 
100793 Results from NNWSI Series 3 Spent Fuel Dissolution Tests. 
PNL-7170. Richland, Washington: Pacific Northwest 
Laboratory. 1990. Wilson, C.N. 
ACC: NNA.19900329.0142. 
II.4 Spent nuclear fuel 
colloids 
Data Qualification Plans 
Adsorption of Actinides onto Colloids as a Function of Time, 
Temperature, Ionic Strength, and Colloid Concentration, 
Waste Form Colloids Report for Yucca Mountain Program 
(Colloid Data Summary from 1999 to 2000 Research). 
(2000). Lu, N.; Conca, J.; Parker, G.R.; Leonard, P.A.; 
Moore, B.; Strietelmeier, B.; and Triay, I.R. LA-UR-00- 
5121. Los Alamos, New Mexico: Los Alamos National 
Laboratory. ACC: MOL.20031204.0108. 
III 
Results from NNWSI Series 3 Spent Fuel Dissolution Tests. 
PNL-7170. Richland, Washington: Pacific Northwest 
Laboratory. 1990. Wilson, C.N. 
ACC: NNA.19900329.0142 
III 
Arguments and justification in support of the qualification of the data used in the model are 
presented in the sections as listed in Table II-1. 
II.1 QUALIFICATION OF DATA FROM LIANG AND MORGAN (1990 
[DIRS 109524]) AND TOMBACZ ET AL. (1990 [DIRS 112690]) 
Liang, L. and Morgan, J.J. 1990. Chemical Aspects of Iron Oxide Coagulation in Water: 
Laboratory Studies and Implications for Natural Systems. Aquatic Sciences, 52, (1), 32-55. Basel, 
Switzerland: Birkhauser Verlag. TIC: 246125. 
The data contained within this report regard hematite colloid characteristics, which are directly 
applicable to the colloid model and thus represent a property of interest. 
Tombcz, E.; Abraham, I.; Gilde, M.; and Szanto, F. 1990. The pH-Dependent Colloidal 
Stability of Aqueous Montmorillonite Suspensions. Colloids and Surfaces, 49, 71-80. 
Amsterdam, The Netherlands: Elsevier. TIC: 246046. 
The data contained within this report regard colloid characteristics of smectite clay, which are 
directly applicable to the colloid model and, thus, represent a property of interest. 
The following factors from LP-SIII.10Q-BSC, Models, Section 5.2.1(k), have been used to 
demonstrate that the data incorporated into the model are suitable for their intended use: 
 Reliability of the data source 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 II-5 August 2005 
 Qualifications of personnel or organizations generating the data 
 Prior uses of the data. 
These two references provide fundamental data on the colloid behavior of iron oxide and 
smectite clay colloids. Liang and Morgan (1990 [DIRS 109524]) document experiments to 
investigate the chemical characteristics of water that result in the coagulation of hematite. Liang 
and Morgan (1990 [DIRS 109524], Figure 1) show the dependence of hematite colloid 
suspension stability on pH and ionic strength. Tombcz et al. (1990 [DIRS 112690], p. 77, 
Figure 3) established the stability of montmorillonite clay colloid suspensions with pH and ionic 
strength. These figures were transformed to produce a set of relationships that could be used in 
the TSPA-LA model to determine whether or not smectite clay and iron oxideoxyhydroxide 
colloid suspensions are stable under specific pH and ionic strength conditions. These data were 
required to enable implementation of the colloid stability portion of the model. The qualification 
of the data, based on the three factors listed above (reliability of source, qualifications of 
personnel, and prior uses) is described below. 
Reliability of the data source. The paper by Liang and Morgan (1990 [DIRS 109524]) was 
published in Aquatic Sciences, a peer-reviewed journal currently published by the Swiss Federal 
Institute for Environmental Science and Technology (EAWAG). In 1990, the year of publication 
of the subject reference, Aquatic Sciences was published by the Swiss Academy of Sciences. 
The journal is a credible and reliable data source of high quality. The paper by Tombcz et al. 
(1990 [DIRS 112690]) was published in Colloids and Surfaces, an important journal in the 
colloidal sciences. 
Qualification of personnel or organizations generating the data. The experimental work 
described in Liang and Morgan (1990 [DIRS 109524]) was performed at the W. M. Keck 
Laboratories, Environmental Engineering Science, California Institute of Technology, Pasadena, 
California. James J. Morgan holds a Ph.D. from Harvard University (1964) and has been at 
Cal Tech since 1965. He is currently the Marvin L. Goldberger Professor of Environmental 
Engineering Science at the California Institute of Technology. He has been the Editor of 
Environmental Science and Technology and on the Editorial Board of Limnology and 
Oceanography, Chairman of Water Resources Research Committee, National Research Council. 
Morgan cowrote, with Werner Stumm, the seminal textbook Aquatic Chemistry, first published 
in 1981. He also was coeditor, with C. P. Huang, of the book Aquatic Chemistry: Interfacial and 
Interspecies Processes, published in 1995. Liyuan Liang holds a Ph.D. in Environmental 
Engineering Science from California Institute of Technology (1988). Since 1990, she has been a 
Staff Scientist at Oak Ridge National Laboratory (ORNL). Her work is also published in 
Environmental Science and Technology, Mineralogical Magazine, Water Resources Research, 
and Geochimica and Cosmochimica Acta. The credentials of Liang and Morgan are outstanding, 
and it is concluded that the data selected from their report (1990 [DIRS 109524]) may be used 
with the highest confidence. 
Dr. Etelka Tombcz, C.Sc. is from the Department of Colloid Chemistry, Univeristy of Szeged, 
Hungary where she is an associate professor. She has over 30 journal and book chapter 
publications dealing with colloid chemistry. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 II-6 August 2005 
Prior uses of the data. The work described in Liang and Morgan (1990 [DIRS 109524]) was 
first noted in the widely used textbook by Donald Langmuir, Aqueous Environmental Chemistry, 
published in 1997. In Chapter 12, Iron Geochemistry, Langmuir discusses suspended colloidal 
iron oxyhydroxides, factors influencing coagulation, and coagulation rates. Liang and Morgans 
work (1990 [DIRS 109524]) has been cited at least 88 times, including: 
 Geckeis H, Grambow B, Loida A, (1998) Formation and stability of colloids under 
simulated near field conditions,  Radiochim. Acta, 82: 123-128 
 Zhang JW, Buffle J, (1996) Multi-method determination of the fractal dimension of 
hematite aggregates, Colloid Surfaces A107: 175-187 
 Ryan JN, Elimelech M (1996) Colloid mobilization and transport in groundwater, 
Colloid Surfaces A 107: 1-56 
 Schafer T, Artinger R, Dardenne K, (2003) Colloid-borne Americium migration in 
gorleben groundwater: Significance of iron secondary phase transformation, Environ. 
Sci. Technol. 37: 1528-1534. 
The following works have cited the paper by Tombcz et al. (1990 [DIRS 112690]): 
 Joe P, Antal K, (1998) Transport processes in clay and humate treated clay layers: 
an electrochemical and radioabsorption survey  Colloids and Surfaces A141: 
365-377 
 Lagaly G, Ziesmer S, (2003) Colloid chemistry of clay minerals: the coagulation of 
montmorillonite dispersions, Adv. Coll. Surface Inter. Sci. 100: 105-128 
Waite TD, Cleaver JK, Beattie JK, (2001) Aggregation kinetics and fractal structure 
of gamma-alumina assemblages, J. Coll. Inter. Sci. 241: 333-339 
 Cool P, Clearfield A, Crooks RM, (1999) The development of microporous pillared 
layered materials for volatile organic compound adsorption and N-2/O-2 separation, 
Adv. Environ. Res. 3 (2): U4-151 
 Buck EC, Bates JK, (1999) Microanalysis of colloids and suspended particles from 
nuclear waste glass alteration, Appl. Geochem. 14: 635-653. 
It is concluded that the suitability of the data from reports by Liang and Morgan (1990 
[DIRS 109524]) and Tombcz et al. (1990 [DIRS 112690]) for its intended use in this report has 
been amply demonstrated in the discussions presented above. Further, on the basis of the 
evidence presented relevant to the three qualification criteria used, the data are qualified. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 II-7 August 2005 
II.2 QUALIFICATION OF DATA FROM RODEN AND ZACHARA (1996 
[DIRS 171518]), POULTON ET AL. (2004 [DIRS 171519]), AND TORSTENFELT 
ET AL. 1982 [171770] 
Roden, E.E. and Zachara, J.M. 1996. Microbial Reduction of Crystalline Iron(III) Oxides: 
Influence of Oxide Surface Area and Potential for Cell Growth. Environmental Science & 
Technology, 30, (5), 1618-1628. Washington, D.C.: American Chemical Society. TIC: 256499. 
The data contained within this report regard the surface area of iron oxides, which are directly 
applicable to the colloid model and, thus, represent a property of interest. 
Poulton, S.W.; Krom, M.D.; and Raiswell, R. 2004. A Revised Scheme for the Reactivity of Iron 
(Oxyhydr)Oxide Minerals Toward Dissolved Sulfide. Geochimica et Cosmochimica Acta, 68, 
(18), 3703-3715. New York, New York: Pergamon. TIC: 256498. 
The data contained within this report regard the surface area of iron oxides, which are directly 
applicable to the colloid model and thus represent a property of interest. 
Torstenfelt, B.; Andersson, K.; and Allard, B. 1982. Sorption of Strontium and Cesium on Rocks 
and Minerals. Chemical Geology, 36, 128-137. Amsterdam, The Netherlands: Elsevier. 
TIC: 256583. 
The data contained within this report regard cesium sorption onto hematite, which are directly 
applicable to the colloid model and thus represent a property of interest. 
The following factors from LP-SIII.10Q-BSC, Section 5.2.1(k), have been used to demonstrate 
that the data incorporated into the model are suitable for their intended use: 
 Reliability of the data source 
 Qualifications of personnel or organizations generating the data. 
Roden and Zachara (1996 [DIRS 171518]) reported surface areas from various iron oxide 
preparations that were determined by multipoint BET analysis with N2 as the adsorbate. They 
were unable to determine the surface area of an amorphous Fe(III) gel by BET analysis, so they 
adopted a value of 600 m2/g based on reviews of the literature (Table II-2). The different 
iron(III) oxides were also examined by transmission electron microscopy to determine the typical 
size range of individual particles and to assess the degree of particle aggregation. 
Table II-2. Data Modified from Roden and Zachara (1996 [DIRS 171518], p. 1,621, Table 1) 
Oxide Morphology 
Typical Particle 
Size Range (nm) 
Surface 
Area (m2/g) 
Amorphous Fe(III) Isometric 2 to 6 600* 
2-line ferrihydrite Aggregated Not determined 240.8 
Goethite Lath 15 to 30 153.3 
Goethite Lath 50 to 70 55.2 
Goethite Lath 100 to 200 31 
Hematite Spheroid 100 to 200 5.4 
NOTE: *calculated value only. 

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MDL-EBS-PA-000004 REV 02 II-8 August 2005 
Data reported by Poulton et al. (2004 [DIRS 171519]) was obtained with multipoint BET 
methods using a Beckman/Coulter SA 1300 analyzer with a Beckman/Coulter SA-PREP 
outgasser unit (Table II-3). 
Table II-3. Iron Oxide Surface Areas 
Mineral SA 
Lepidrocrocite 61.4 m2/g 
Goethite 36.5 m2/g 
Hematite 2.5 m2/g 
Magnetite 2.7 m2/g 
Source: Poulton et al. 2004 [DIRS 171519], 
Section 3.1, p. 3,705. 
NOTE: No surface area measurement was 
obtained for HFO material. 
Data by Torstenfelt et al. (1982) includes Kd values (reported in m3/kg) for cesium sorption onto 
hematite, magnetite, calcite, apatite, orthoclase, serpentine, biotite, muscovite, hornblende, 
biotite, gneiss, diabase, and granite. Data for hematite was used in the model (Tables II-4 
and II-5). 
Table II-4. Effect of pH on Cesium Sorption (Kd) 
Mineral 
pH Hematite (mL/g) Magnetite (mL/g) Biotite (mL/g) 
Muscovite 
(mL/g) 
5 2 3 900 700 
7 4 5 790 810 
9 8 9 700 840 
Source: Torstenfelt et al. 1982 [DIRS 171770], Table III, p. 130. 
NOTE: The original data were reported in m3/kg and has been converted to mL/g. 
Table II-5. Effect of Contact Time on Sorption (Kd) 
Contact Time 
Rock/Mineral 
Nuclide 
(10-7M) 1 day (mL/g) 1 week (mL/g) 3 months (mL/g) 6 months (mL/g) 
Hematite Cesium 6 6 NA 6 
Magnetite Cesium 8 7 NA 6 
Biotite Cesium 439 624 668 617 
Muscovite Cesium 627 765 NA 527 
Source: Torstenfelt et al. 1982 [DIRS 171770], Table IV, p.133. 
NOTES: Contact time: 1 week, particle size 45 m to 63 m. NA = not available. 
The original data were reported in m3/kg and has been converted to mL/g. 
Reliability of the data source. The paper by Roden and Zachara (1996 [DIRS 171518]) was 
published in Environmental Science & Technology, a well respected journal in environmental 
chemistry. Data from Roden and Zachara (1996 [DIRS 171518]) has been used by Poulton and 
many others. This paper has been cited 153 times. 

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MDL-EBS-PA-000004 REV 02 II-9 August 2005 
Poulton and coworkers have previously reported surface area determinations of iron oxides using 
BET (Poulton et al. 2004 [DIRS 171519], 825 to 834). The journal, Geochimica Cosmochimica 
Acta, is one of the most highly respected publications in the geosciences. 
Data from Torstenfelt et al. (1982 [DIRS 171770]) was used as input in Table D.3.0 Kd Data Set 
for Soils and Pure Mineral Phases, Appendix D, Partition Coefficients for Cesium in the 
Environmental Protection Agency, Volume II on sorption coefficients. Publications from the 
group of Torstenfelt and Allard are widely cited in the scientific literature. These were reported 
in the peer-reviewed journal Chemical Geology and were part of a Ph.D. dissertation by 
Dr. Torstenfelt awarded by Chalmers University in Sweden. 
Qualifications of personnel or organizations generating the data. Dr. John Zachara is Chief 
scientist and one of four associate scientific directors of the William Wiley Environmental 
Molecular Sciences Laboratory at Battelle, Pacific Northwest National Laboratory. Dr. Zachara 
received his Ph.D. in soil chemistry from Washington State University in 1986. His research 
focuses on adsorption reactions between organic, metal, and radionuclide contaminants and solid 
matter in the subsurface. His research has ranged from fundamental surface chemical studies to 
site evaluations of solute mobilization and transport in the field. He has authored or coauthored 
over 80 publications on these subjects. His current research is focused on the geochemical 
behavior of metals and radionuclides complexed by organic ligands and on the influence of 
subsurface microbial processes on mineral surface chemistry and reactivity in groundwater. The 
coauthor on this paper is Eric Roden, an associate professor in Biological Sciences at the 
University of Alabama. He earned a B.S., Lebanon Valley College, Biology, 1983 and a Ph.D. 
at the University of Maryland College Park, Marine-Estuarine Environmental Science, 1990. He 
previously held a postdoctoral position at Pacific Northwest National Laboratory and has more 
than 20 peer-reviewed journal publications. Zachara and Roden have been funded through the 
Office of Biological and Environmental Research (OBER) of the Office of Science (SC), U.S. 
Department of Energy (DOE), Natural and Accelerated Bioremediation Research (NABIR) 
program for several years. The results of DOE SC investigations are generally published in peerreviewed 
scientific literature. The experimental investigations of iron oxide behavior were 
performed under the auspices of the DOE NABIR program and the for DOE Subsurface Science 
Program. It is concluded, therefore, that the quality of the PNNL laboratory investigations 
deems the resulting data reliable. The credentials of John Zachara and Eric Roden are 
outstanding, and it is concluded that the data selected from Zachara and Roden (1996 
[DIRS 171518) may be used with the highest confidence. 
Additional surface area data has been used from a paper by Poulton et al. (2004 [DIRS 171519]). 
Simon Poulton has a B.Sc. in Geology (1992) from the University of Portsmouth, United 
Kingdom (UK), an M.Sc. in Geochemistry (1994) from the University of Leeds, UK, and a Ph.D. 
in Geochemistry (1998) from the University of Leeds, UK. The coauthor on the paper, Rod 
Raiswell, is a Professor of Geochemistry at Leeds University, UK. Dr. Poulton was a Lecturer at 
the School of Earth Sciences, University of Leeds, UK and is now an Assistant Professor at the 
Danish Centre for Earth System Science, University of Southern Denmark. The Earth Sciences 
Department at Leeds University in the UK is an internationally renowned research department 
and is where Poulton and coworkers performed this study. The Department received the top 
research grade (grade 5) in all Higher Education Funding Council for England (HEFCE) 
Research Assessments made to date. This grading is used to rate scientific research departments 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 II-10 August 2005 
within the UK. The credentials of Simon Poulton are acceptable and it is concluded that this data 
may be used with confidence. 
Dr. Brje Torstenfelt and Dr. Bert Allard have been studying the mechanisms of radionuclide 
attenuation in a repository environment for over 20 years. Work from this group has been 
reported in numerous technical reports sponsored by the Swedish Repository Program and 
presented at the Scientific Basis for Nuclear Waste Management sponsored by the Materials 
Research Society and in peer-reviewed publications. Dr. Torstenfelt has a Ph.D. from Chalmers 
University in Gteborg, Sweden in radionuclide sorption on geological materials. Chalmers 
University has produced many reports on nuclear waste management over the past few decades. 
It is concluded that the suitability of the data from Roden and Zachara (1996 [DIRS 171518]), 
Poulton et al. (2004 [DIRS 171519]), and Torstenfelt et al. (1982 [DIRS 171770]) for its 
intended use in this report has been amply demonstrated in the discussions presented above. 
Further, on the basis of the evidence presented relevant to the two qualification criteria used, the 
data are qualified. 
II.3 QUALIFICATION OF DATA FROM LU ET AL. 2000 [DIRS 166315] 
The data contained within this report regard the surface area of iron oxides and colloid 
concentration, which are directly applicable to the colloid model and thus represent a property of 
interest. 
II.3.1 Data set(s) for qualification 
The purpose of this Data Qualification Report is to qualify the surface area and colloid 
concentration data from Adsorption of Actinides onto Colloids as a Function of Time, 
Temperature, Ionic Strength, and Colloid Concentration, Waste Form Colloids Report for Yucca 
Mountain Program (Colloid Data Summary from 1999 to 2000 Research) (Lu et al. 2000 
[DIRS 166315]) for use as direct input to this document. 
II.3.2 Method(s) of qualification selected and rationale 
Data on surface areas for iron oxides and mass of colloids (Lu et al. 2000 [DIRS 166315], 
Table 1) have been qualified in accordance with LP-SIII.2Q-BSC, Qualification of Unqualified 
Data. The qualification process used for these data is necessary because the documentation of 
proof of proper data acquisition is unavailable for review (LP-SIII.2Q-BSC, Attachment 3, 
Section 5.b). This qualification process is designed to provide the desired level of confidence for 
the intended use in this report. This report documents the data qualification task conducted in 
accordance with the approved Data Qualification Plan (Appendix III) for the technical 
assessment qualification method. 
II.3.3 Evaluation criteria 
The data will be considered qualified if attributes 1 through 5 from LP-SIII.2Q-BSC, 
Attachment 4 are adequately addressed when compared to the generally accepted scientific and 
engineering practices at the time the data was generated. 

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MDL-EBS-PA-000004 REV 02 II-11 August 2005 
The attributes are as follows: 
1) Qualifications of personnel or organizations generating the data are comparable to 
qualification requirements of personnel generating similar data under the approved 
10 CFR 60, Subpart G, program. 
2) The technical adequacy of equipment and procedures used to collect and analyze the 
data. 
3) The extent to which the data demonstrate the properties of interest (e.g., physical, 
chemical, geologic, mechanical). 
4) The environmental conditions under which the data were obtained if germane to the 
quality of data. 
5) The quality and reliability of the measurement control program under which the data 
were generated. 
II.3.4 Evaluation of the technical correctness of the data 
Determination that the employed methodology is acceptable (Attributes 2 through 4). 
The measurements of the surface area of iron oxides were performed using ethylene glycol 
monoethyl ether sorption. This method is typically used for soil surface area determinations, 
particularly for clays. Its application to iron oxides (e.g., hematite) is reasonable and the values 
obtained are comparable to the results of Roden and Zachara (1996 [DIRS 171518]) and 
Poulton et al. (2004 [DIRS 171519]) on similar iron oxide particles. The data demonstrate the 
properties of interest. 
The concentration of colloids was obtained through standard weighing and dilution. 8 g of 
ground Fe2O3 powder were added to 500 mL of J-13 well water or SYN.J-13 well water and 
shaken vigorously. After leaving to stand for 1hr, the 2/3 of the upper portion of the suspension 
was diluted with 500 mL of J-13 well water or SYN.J-13 well water. The mass of colloids was 
determined by weight difference of the colloid suspension before and after drying. The J-13 
solution had been passed through a 50-nm membrane filter under ~2 atmospheres and the 
SYN.J-13 solution was prepared by 0.028 g Na2CO3 and 1.92 g NaHCO3 to 10 L of nanopure 
deionized water (Lu et al. 2000 [DIRS 166315]). These procedures are technically adequate and 
provide the property of interest under appropriate environmental conditions. 
Determination that confidence in the data acquisition or developmental results is 
warranted (Attributes 1 and 5). 
Los Alamos National Laboratory (LANL) is, and has been for years, one of the few resources in 
the U.S. with the ability to perform complex laboratory investigations with radioactive materials. 
The results of the investigations are generally published as LANL reports and eventually are 
published in peer-reviewed literature. The experimental investigations were performed under the 
auspices of the Yucca Mountain Project. Ningping Lu has been senior author or a coauthor on 
11 LANL reports related to radioactive elements and isotopes and their behavior in the 
environment. She has published in peer-reviewed journals, including AIP Journal, Applied 
Radiation and Isotopes, Applied Geochemistry, Radiation Physics and Chemistry, and 
Radiochimica Acta. 

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MDL-EBS-PA-000004 REV 02 II-12 August 2005 
II.3.5 Evaluation results/conclusion/limitations 
The data are qualified through attributes 1 through 5 from LP-SIII.2Q-BSC, Attachment 4, which 
are adequately addressed when compared to the generally accepted scientific and engineering 
practices at the time the data was generated. The input data being considered here is qualified 
only for the purposes of this report. 
II.4 QUALIFICATION OF DATA FROM WILSON 1990 [DIRS 100793] 
II.4.1 Data set(s) for qualification 
The purpose of this data qualification report is to qualify the 239Pu + 240Pu series 3 data in 
Tables A.2 through A.7 of Results from NNWSI Series 3 Spent Fuel Dissolution Tests, PNL-7170 
(Wilson 1990 [DIRS 100793]) for use as a direct input to this report. 
II.4.2 Method(s) of qualification selected and rationale 
The 239+240Pu Series 3 data, Tables A.2 through A.7 (Wilson 1990 [DIRS 100793]), are qualified 
in accordance with LP-SIII.2Q-BSC, Qualification of Unqualified Data. The qualification 
processes used for these data are necessary because the documentation of proof of proper data 
acquisition is unavailable for review (LP-SIII.2Q-BSC, Attachment 3, Section 5.b). This 
qualification process is designed to provide the desired level of confidence for the data in its 
intended use for this modeling report. This report documents the data qualification task 
conducted in accordance with the approved Data Qualification Plan (Appendix III) for the 
technical assessment qualification method. 
II.4.3 Evaluation criteria 
The data will be considered qualified if attributes 1 through 5 and 7 from LP-SIII.2Q-BSC, 
Attachment 4 are adequately addressed when compared to the generally accepted scientific and 
engineering practices at the time the data was generated. The attributes are as follows: 
1) Qualifications of personnel or organizations generating the data are comparable to 
qualification requirements of personnel generating similar data under the approved 
10 CFR 60, Subpart G, program. 
2) The technical adequacy of equipment and procedures used to collect and analyze the 
data. 
3) The extent to which the data demonstrate the properties of interest (e.g., physical, 
chemical, geologic, mechanical). 
4) The environmental conditions under which the data were obtained if germane to the 
quality of data. 
5) The quality and reliability of the measurement control program under which the data 
were generated. 
7) Prior uses of the data and associated verification processes. 

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MDL-EBS-PA-000004 REV 02 II-13 August 2005 
II.4.4 Evaluation of the technical correctness of the data 
Determination that the employed methodology is acceptable (attributes 2 through 4) 
The spent nuclear fuel samples used in the results reported in Series 3 data, Tables A.2 through 
A.7 (Wilson 1990 [DIRS 100793]) were from the H. B. Robinson and Turkey Point reactors and 
had been characterized by Materials Characterization Center at PNL as approved testing 
materials. The 239+240Pu content was verified by using both a computer code and radiochemical 
analyses. For the H. B. Robinson fuel the reported calculated 239+240Pu content was 7.44  102 
Ci / g fuel and the measured value was 7.16  102 Ci/g fuel (Wilson 1990 [DIRS 100793]). 
An important aspect of the spent nuclear fuel sample preparation was the removal of fine 
particulates from sectioning and handling in the hot cells before testing. This type of 
contamination would have biased the results if not removed. The testing set-up and design is 
appropriate for spent nuclear fuel leaching experiments. 
The series 3 tests were conducted in sealed stainless steel vessels similar to other waste form 
tests (e.g. Bates et al. (1995 [DIRS 170880]) on DHLW glass and Finn et al. (1994 
[DIRS 100746]) on spent nuclear fuel). Temperature of tests was controlled using an oil bath 
immersion technique. Periodic solution samples were extracted using a preleached 20-ml plastic 
syringe with 6-in-long stainless steel needles. Any attachment of plutonium to the walls of the 
syringe may have reduced the plutonium concentration. For the high temperature tests, samples 
were extracted at temperature. Before drawing the sample, a small amount of air was bubbled 
into the leachate. This may have resulted in the agitation of some settle plutonium particles; 
however, it was used to maintain consistency in the sampling procedure. Solutions were placed 
in preleached vials. The sample extraction methods are similar to those reported in other waste 
form tests (see Bates et al. 1995 [DIRS 170880]; Vernaz and Godon 1992 [DIRS 174503]). 
Hence, the waste form testing methodology is an acceptable method. 
The data in Series 3 data, Tables A.2 through A.7 (Wilson 1990 [DIRS 100793]) is plutonium 
concentration with filtered fraction. The plutonium analysis was performed with alpha 
spectroscopy and the filtration with a series of Amicon Corporation type CF25 Centriflo 
membrane filter cones (this product was discontinued in March 2000). Centriflo was a 
trademark of the Amicon Corporation, Lexington, MA, now owned by Millipore Corporation, 
Billerica, MA. 
After around 1 hr, the sample was separated into aliquots for the unfiltered, 0.4-m filtered, and 
18- filtered fractions. Filtration is commonly employed for examining colloidal populations 
(van Middlesworth and Wood 1998 [DIRS 170927]; Vernaz and Godon 1992 [DIRS 174503]). It 
is an imperfect method but is the only method available for examining colloid sizes of a 
particular species in a complex system, such as a waste for corrosion tests. However, as the 
range of filter sizes is always smaller than the theoretical colloidal size range, the range reported 
by Wilson (1990 [DIRS 100793]) is likely to be biased lower than the actual range. Hence, the 
use of filtration is an acceptable method for isolating colloid concentrations from corrosion tests. 
Alpha spectroscopy was used to determine the 239+240Pu content for each aliquot. The detection 
limits for 239+240Pu by alpha spectroscopy were listed as 0.003 ppb for a spent nuclear 
fuel-derived sample. Use of alpha spectroscopy for analyzing spent nuclear fuel solutions has 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 II-14 August 2005 
been reported by Finn et al. (1994 [DIRS 100746]) and McNamara et al. (2005 [DIRS 174068]). 
Results by Finn et al. (1994 [DIRS 100746]) demonstrate that similar results can be obtained for 
both alpha spectroscopy and mass spectrometry when analyzing spent nuclear fuel leachates. 
Hence, use of alpha spectroscopy for analyzing plutonium colloid concentrations in spent nuclear 
fuel tests is an acceptable method of determining these properties of interest. 
Determination that confidence in the data acquisition or developmental results is 
warranted (attributes 1, 5, and 7) 
Pacific Northwest National Laboratory (PNNL) is, and has been for years, one of the few 
resources in the U.S. with the ability to perform complex laboratory investigations with spent 
nuclear fuel. The results of the investigations are generally published as PNNL reports and 
eventually are published in peer-reviewed literature. The experimental investigations of spent 
nuclear fuel corrosion were performed under the auspices of the Yucca Mountain Project; the 
experimental conditions chosen were based on the environmental conditions anticipated in the 
repository. The experiments represent probably the only laboratory work performed that 
specifically addresses the formation of plutonium colloids from spent nuclear fuel produced 
under immersion conditions. It is concluded, therefore, that the quality of the PNNL 
investigations and the Project-specific design of the experiments deem the resulting data reliable. 
The experimental work described by Wilson (1990 [DIRS 100793]) was performed at PNNL. 
Dr. Chuck Wilson was a senior author or a coauthor on ~40 PNNL reports, task plans, and peerreviewed 
publications related to spent nuclear fuel corrosion. Dr. Wilson has a Doctorate from 
the University of Washington (1974) and B.S. from the University of Washington (1969) in 
Ceramic Engineering (Cum Laude). Dr. Wilson has worked on the Basalt Waste Isolation Pilot 
Program and was principal investigator for spent nuclear fuel radionuclide release in support of 
the Yucca Mountain repository in 1984. He has coauthored papers with other well-respected 
scientists within the Yucca Mountain Program, including: Drs. Henry Shaw, Carol Bruton, Walt 
Gray, Robert Einziger, and Virginia Oversby. 
As described in the source document (Wilson 1990 [DIRS 100793], Acknowledgements), the 
Westinghouse Hanford Company (WHC) and Pacific Northwest National Laboratory (PNNL) 
developed these data for the Yucca Mountain Project (YMP). The work was conducted under 
Contract No. W-7405-ENG-48 and the YMPapproved WHC/PNNL quality assurance 
programs, which met the contractual quality assurance requirements. The laboratory activities 
were completed under the Westinghouse Hanford Company QA program, from Feb. 1986 to 
May 1987, and much of the data evaluation and the preparation by Wilson (1990 
[DIRS 100793]) were performed under the PNNL QA program. This work was transferred from 
WHC to PNNL in June 1987 as part of the Hanford site contract consolidation effort. The WHC 
and PNNL QA programs were reviewed and approved by LLNL, which managed the waste 
package task for YMP and treated WHC and PNNL as suppliers. Technical implementing 
documents, including test plans, test procedures, and technical drawings were reviewed and 
approved by LLNL and WHC prior to their implementation, and determined to be technically 
adequate for the collection and analyses of the data. Test Plan for Series 3 NNWSI Spent Fuel 
Leaching/Dissolution Tests (Wilson 1986 [DIRS 171996]) also included an appendix that 
identified all test instructions, technical procedures, and analytical procedures, including solution 
and solids analyses, and burnup/radiochemical analyses. The initial evaluation of the overall 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 II-15 August 2005 
data quality has determined that the controls and methods used to plan, collect, and analyze the 
data in question were adequate when compared to generally accepted scientific and engineering 
practices at the time the data were generated. 
The spent nuclear fuel release data reported by Wilson (1990 [DIRS 100793]) has been used in 
following project products: 
 ANL-DSD-MD-000001 Rev. 0, Probability Analysis of Corrosion Rates for Waste 
Package Materials 
 ANL-EBS-MD-000015 Rev. 02, CSNF Waste Form Degradation: Summary 
Abstraction 
 ANL-EBS-MD-000019 Rev. 0, ICN 1 Secondary Uranium-Phase Paragenesis and 
Incorporation of Radionuclides into Secondary Phases 
 ANL-WIS-MD-000010 Rev. 02, Dissolved Concentration Limits of Radioactive 
Elements 
 ANL-WIS-MD-000021 Rev. 0 ICN 0, Clad Degradation - Summary and 
Abstraction for LA 
 CAL-WIS-MD-000012 Rev. 0 ICN 0, Dissolved Concentration Limits for Certain 
Radioactive Elements 
 TDR-CRW-GS-000001 Rev. 02 ICN 01, Yucca Mountain Site Description  
Section 5. 
II.4.5 Rationale for Abandoning Qualification Method 
Direct comparison of the series 1, 2, and 3 tests conducted by Wilson (1985 [102147]; 1987 
[DIRS 102150]; 1990 [DIRS 100793]) was considered in the data qualification plan in REV01 of 
the document because the same group at PNL used the same materials in all test series. 
However, there were differences that make direct useful comparison with respect to determining 
plutonium colloid concentrations. The series 1 and 2 tests were run in unsealed vessels. Series 1 
was conducted in deionized water and series 2 and 3 in J-13 well water. The series 3 tests were 
conducted in sealed stainless steel vessels. Sequential filtration was only carried out on the 
series 2 and 3 tests (Wilson 1987 [DIRS 102150]; 1990 [DIRS 100793]). Hence, comparison 
with Series 1 tests could not be made with respect to plutonium colloid concentrations. 
Although, Wilson (1987 [DIRS 102150]) reports the filtered solution analyses for the series 2 
test, the original samples used in the tests were not cleaned after sectioning and cutting in the hot 
cells. This led to fines and particulates in the solutions. Hence, these results were not 
representative of colloid formation following weathering of spent nuclear fuel and could not be 
compared to the series 3 tests, where precleaning was performed. Therefore, a technical 
assessment was made of the data in the series 3 tests related to plutonium colloid concentrations. 

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MDL-EBS-PA-000004 REV 02 II-16 August 2005 
II.4.6 Evaluation Results/Conclusion/Limitations 
The data reported in the series 3 data, Tables A.2 through A.7 (Wilson 1990 [DIRS 100793]) are 
qualified through attributes 1 through 5 and 7 from LP-SIII.2Q-BSC, Attachment 4, which are 
adequately addressed when compared to the generally accepted scientific and engineering 
practices at the time the data was generated. The input data being considered here is qualified 
only for the purposes of this report. 
II.5 QUALIFICATION OF DATA FROM LIU ET AL. (1992 [DIRS 172864]) 
172864 Liu, C.; Cox, D.S.; Barrand, R.D.; and Hunt, C.E.L. 1992. 
Particle Size Distributions of U3O8 Produced by Oxidation in Air 
at 300-900C. Proceeding of the 13th Annual Conference, Saint 
John, NB, June 7-10, 1992. 1, 1-22. [Toronto, Ontario, Canada]: 
Canadian Nuclear Society. TIC: 256950. 
The data contained within this paper regard particle size characteristics following high 
temperature oxidation, which are directly applicable to the colloid model and thus represent a 
property of interest. 
The following factors from LP-SIII.10Q-BSC, Section 5.2.1(k), have been used to demonstrate 
that the data incorporated into the model are suitable for their intended use: 
 Qualifications of personnel or organizations generating the data. 
Data were obtained at the Atomic Energy Canada Laboratory, Whiteshell, Canada. This 
organization has performed many studies on spent nuclear fuel behavior over many years. Hunt, 
Cox, and Barrand have published several papers in the peer-reviewed scientific publications 
including; Journal of Nuclear Materials and Nuclear Engineering and Design on uranium oxide 
oxidation and fission product release over 10 yrs to 20 yrs. 

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APPENDIX III 
DATA QUALIFICATION PLANS 

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INTENTIONALLY LEFT BLANK 

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III. DATA QUALIFICATION 
PLANS 

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Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 IV-1 August 2005 
APPENDIX IV 
CR 5286 

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INTENTIONALLY LEFT BLANK 

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IV. CR 5286 
CR5286 states that the high sorption coefficients (Kd) for thorium, protactinium, and cesium 
result in noncredible nonphysical sorption of these radionuclides onto colloids. In one case, a 
TSPA run indicated that there would be 10 g of thorium sorbed onto 1g of FeOOH. 
RESPONSE 
The situation of sorbing 10 g of thorium onto 1 g of iron oxide may be physically impossible. 
The range of surface areas for hematite, goethite, and ferrihydrite in Table 6-8 of REV01 of this 
document is from 1.8 to 720 m2/g. Based on a density of 3.93 g/cm3 and spherical geometry, a 
particle with a surface area of 720 m2/g would be about 2 nm in diameter, which is within 
accepted colloidal size range. 
Figure A-1. Plot Showing Particle Size with Maximum Mass of Thorium Sorbed. 
The maximum sorbed quantity of an element is provided through the density of available 
sorption sites (Davis and Kent (1990 [DIRS 143280], Table 1, p. 182). The values for 
ferrihydrite are between 2.6 and 16.8 /nm2. In Figure A-1, this information is presented 
graphically. A particle with a surface area of 720 m2/g and a site density of 16.8 /nm2 could 
sorb: 
720 (m2/g)  1.68  1019 (sites/m2) 232 (g/mole) / NA = 4.66 g Th 
where: 
NA = 6.022  1023 sites/mole 
To sorb close to 10 g of thorium the particle would need to have a surface area of 1,545 m2/g. 
and would be ~1 nm in diameter. This is beyond the colloid size range and predicted surface 
areas for iron oxides. 

Waste Form and In-Drift Colloids-Associated Radionuclide Concentrations: Abstraction and Summary 
MDL-EBS-PA-000004 REV 02 IV-4 August 2005 
In conclusion, CR5286 is correct that the amount of sorption of thorium, protactinium, and 
cesium should be limited according to the available sorption sites on the colloids. The base-case 
model, which has no limit, is conservative.