Waste Form Colloid-Associated Concentration Limits: Abstraction and Summary Rev 00, ICN 00

ANL-WIS-MD-000012

April 2000

1. PURPOSE 
The purpose of this Analysis/Model Report (AMR) is to present and describe the abstraction of 
process models and data developed primarily by Argonne National Laboratory (ANL) and Los 
Alamos National Laboratory (LANL) for the Total System Performance Assessment-Site 
Recommendation (TSPA-SR) calculations to be performed with the TSPA code, GoldSim. The 
process models describe the observed types, formation, and stability of radionuclide-bearing 
colloids that result from degradation and decomposition of the waste forms contained in failed 
waste packages. In addition, contaminant-colloid attachment/detachment mechanisms and 
transport characteristics anticipated in the repository are discussed for colloids formed from 
degradation of the waste forms, colloids formed from man-made materials introduced into the 
repository, and naturally occurring colloids. The abstraction of the process models is intended to 
capture the most important characteristics of radionuclide-colloid behavior for use in predicting 
the potential impact on repository performance of colloid-facilitated radionuclide transport. 
It is anticipated that quantities of colloids will be mobilized as a result of alteration of both the 
high-level radioactive waste (HLW) and spent nuclear fuel (SNF) waste forms. The abundance 
of colloids within the breached waste package will depend on the extent of waste form alteration 
and the alteration products formed. Colloid abundance and stability also depend on many 
environmental factors, including the ionic strength, pH, cation concentrations, colloid content of 
groundwater entering the waste package from the drift, presence of fulvic and humic acids, and 
microbe fragments. Suspended colloids may subsequently flocculate and settle by gravity, be 
chemically or mechanically filtered, or dissolve. If the environmental factors change, colloids 
may be peptized; colloid-sized particles may precipitate; or other natural processes may occur. 
In addition, colloids may sorb readily at the interfaces between air and water in rocks and 
engineered barriers and, depending upon the degree of saturation of the porous medium as well 
as its configuration, may be retarded, even immobilized or transported. These issues are relevant 
to colloid transport, which is of primary concern within the engineered barrier system (EBS) in 
the drift and in the near- and far-fields. They are relevant as well to transport, which also may 
occur to some extent within the waste package. 
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. This abstraction analysis is restricted to this inpackage 
portion of the repository system. 
Colloidal systems are defined here as those in which the component phases are dispersed in an 
aqueous medium (groundwater, leachate) ranging in size from approximately 1 nm to 1 m. 
Much has been learned about environmental behavior of colloids over the past few decades, but 
due largely to their small size and unique surface-dominated chemical and physical 
characteristics, they have proven to be difficult to study. Therefore, the approach taken in this 
analysis was to bound the probable behavior of colloid-associated radionuclides. The lack of 
certain data can be mitigated as well by comparing these bounding predictions with relevant 
laboratory and field studies and with analog sites in which environmental conditions are 
analogous in critical respects to conditions used in experiments and anticipated in the repository. 
These approaches have been used to the extent possible to reduce the uncertainties associated

ANL-WIS-MD-000012 REV00 13 April 2000 
with predictions of colloid-associated radionuclide concentrations, stability, and transport 
characteristics. 
In accordance with AP-2.13Q, Technical Product Development Planning, a development plan, 
Waste Form Colloid-Associated Concentration Limits -Abstraction and Summary, was 
developed, issued, and utilized in the preparation of this AMR (CRWMS M&O 2000c). 
2. QUALITY ASSURANCE 
The Quality Assurance (QA) program applies to this AMR. All types of WPs were classified 
(per QAP-2-3) as Quality Level-1. CRWMS M&O (1999a, p. 7) in Classification of the MGR 
Uncanistered Spent Nuclear Fuel Disposal Container System is cited as an example of a WP 
type. The Performance Assessment Operations responsible manager has evaluated this technical 
document activity in accordance with QAP-2-0, Conduct of Activities. The activity evaluation, 
Conduct of Performance Assessment, determined that this activity is subject to the Quality 
Assurance Requirements and Description (DOE 2000) requirements. 
3. COMPUTER SOFTWARE AND MODEL USAGE 
No computer software was used for calculations and/or modeling in support of this AMR. No 
codes or routines were developed. 
Reference to parameters used in calculations by GoldSim is made in this AMR. The parameters 
are provided for use by the TSPA-SR team in GoldSim, but no GoldSim calculations were 
performed for this AMR. 
Graphs and figures were prepared with Microsoft Excel 97, Microsoft PowerPoint 97, and 
Adobe Illustrator 8.0.1. Excel was used for tabular presentation of data only in Attachment XVI, 
Table XVI-5. 
4. INPUTS 
The rationale for the approach used in the waste form colloid source term abstraction for TSPASR 
is based largely on the observation and characterization of radionuclide-bearing colloids 
formed during waste form degradation tests conducted at ANL. The general approach used in 
the analysis was extended and developed from that used in the TSPA Viability Assessment 
(VA). Additional data and improved process models from ongoing and recently completed 
laboratory and field testing programs at ANL and LANL are incorporated into this AMR. The 
data and models were developed according to AP-3.10Q and the respective Quality Assurance 
Plans (QAPs) for Yucca Mountain Project activities performed at ANL and LANL. Results of 
these activities enabled reduction in uncertainties through the refinement of parameter ranges and 
a better understanding of fundamental mechanisms of colloid formation, stability, and behavior 
and the association of radionuclides with colloids.

ANL-WIS-MD-000012 REV00 14 April 2000 
4.1 DATA AND PARAMETERS 
Technical product inputs and sources are listed in Table 1 and on the Document Input Reference 
System as stipulated by AP-3.10Q. Primary technical inputs have been taken from CRWMS 
M&O (2000a). 
Table 1. Summary of Parameters Used in Abstraction 
Data Tracking Number 
(DTN) 
Parameter 
Name 
Parameter Description Unitsa Basis for 
Value/ 
Source 
Value and 
Distribution 
Waste Form Colloids 
LL991109851021.095 CRNcoll,wf,irrev,max Highest observed or expected 
concentration of Pu associated with both 
types of waste-form colloids 
mol/L Figures 7 and 
13 (Pu only) 
6E-8 mol/Lb 
MO0003SPALOW12.00 
1 
CRNcoll,wf,irrev,min Lowest observed or expected 
concentration of Pu associated with both 
types of waste-form colloids 
mol/L Figures 7 and 
13 (Pu only) 
1E-11 mol/Lb 
LL991109851021.095 Mcoll,wf,both,max Highest observed or expected mass of 
both types of waste-form colloids per unit 
volume or mass of water 
mg/L Figure 10 5 mg/Lc 
N/A, intermediate 
parameter resulting from 
development of TSPA in 
this AMR 
CRNcoll,wf,irrev Concentration of irreversibly attached (or 
engulfed as a solid inclusion) 
radionuclide element RN associated with 
waste-form colloids 
mol/L Determined in 
PA 
calculation; 
Flow-charts 
1a, 1b 
N/A 
N/A, intermediate 
parameter resulting from 
development of TSPA in 
this AMR 
CRNcoll,wf,rev Concentration of reversibly attached 
radionuclide element RN associated with 
waste-form colloids 
mol/L Determined in 
PA 
calculation; 
Flow-chart 1d 
N/A 
N/A, intermediate 
parameter resulting from 
development of TSPA in 
this AMR 
Mcoll,wf,both Mass of waste-form colloids with 
reversibly sorbed or irreversibly attached 
(or engulfed) radionuclide element RN 
per unit volume or mass of water 
mol/L Determined in 
PA 
calculation; 
Flow-chart 1c 
N/A 
MO0003SPAION02.003 Ilo-thresh,coll,wf Ionic strength below which waste-form 
colloids are stable 
mol/L Figures 7 and 
13; Figures 5 
and 12 
0.01 mol/L 
MO0003SPAION02.003 Ihi-thresh,coll,wf Ionic strength above which waste-form 
colloids are unstable 
mol/L Figures 7 and 
13; Figures 5 
and 12 
0.05 mol/L 
MO0004SPAKDS42.005 Kd,RNcoll,wf,rev Distribution coefficient for reversible 
sorption of radionuclide element RN onto 
waste-form colloids 
mL/g 
colloid 
Data 
developed 
from LANL 
data; 
CRWMS 
M&O (2000c) 
Geometric 
mean values: 
Pu = 1e+4 
Am = 1e+5 
log-normal 
distributions; 
geometric 
std. dev. = 10 
N/A, intermediate 
function resulting from 
development of TSPA in 
this AMR 
STcoll,wf,is Function relating stability of waste-form 
colloids to ionic strength, based on 
results of HLW glass degradation 
experiments (based on Pu concentration) 
mol/L Figures 7 and 
13 
[Pu colloid] = 
-1.25E-6  I 
+7.25E-8

ANL-WIS-MD-000012 REV00 15 April 2000 
Data Tracking Number 
(DTN) 
Parameter 
Name 
Parameter Description Unitsa Basis for 
Value/ 
Source 
Value and 
Distribution 
N/A, intermediate 
functions resulting from 
development of TSPA in 
this AMR 
STcoll,wf,pH Function relating stability of waste-form 
colloids to ionic strength and pH, over 
limited pH range, based on results of 
montmorillonite stability experiments 
mol/L Figures 5 and 
12 
I = pH/200 
Iron-(hydr)oxide Colloids 
N/A, intermediate 
parameter resulting from 
development of TSPA in 
this AMR 
CRNcoll,FeOx Concentration of reversibly attached 
radionuclide element RN associated with 
iron-(hydr)oxide colloids 
mol/L Determined in 
PA 
calculation; 
Flow-chart 2b 
N/A 
N/A, intermediate 
parameter resulting from 
development of TSPA in 
this AMR 
Mcoll,FeOx Mass of iron-(hydr)oxide colloids per unit 
volume or mass of water 
mg/L Determined in 
PA 
calculation; 
Flow-chart 2a 
N/A 
MO0003SPAHIG12.002 Mcoll,FeOx,max Highest observed or expected mass of 
iron-(hydr)oxide colloids per unit volume 
or mass of water 
mg/L Reasonable 
conservative 
estimate 
1 mg/L 
MO0003SPAHIG12.002 Mcoll,FeOx,min Lowest observed or expected mass of 
iron-(hydr)oxide colloids per unit volume 
or mass of water 
mg/L Reasonable 
conservative 
estimate 
1E-3 mg/L 
MO0003SPAION02.003 Ilo-thresh,coll,FeOx Ionic strength below which iron- 
(hydr)oxide colloids are stable 
mol/L Figures 6 and 
11 
0.01 mol/L 
MO0003SPAION02.003 Ihi-thresh,coll,FeOx Ionic strength above which iron- 
(hydr)oxide colloids are unstable 
mol/L Figures 6 and 
11 
0.05 mol/L 
MO0004PAKDS42.005 Kd,RNcoll,FeOx Distribution coefficient for reversible 
sorption of radionuclide element RN onto 
iron-(hydr)oxide colloids 
mL/g 
colloid 
Data 
developed 
from LANL 
data; 
CRWMS 
M&O (2000c) 
Geometric 
mean values: 
Pu, Am = 
1e+4 
log-normal 
distributions; 
geometric 
std. dev. = 10 
N/A, intermediate 
function resulting from 
development of TSPA in 
this AMR 
STcoll,FeOx,pHlo Function relating stability of iron- 
(hydr)oxide colloids to pH and ionic 
strength at relatively low pH values 
mol/L Figure 6 and 
11 
I = -0.02  pH 
+0.17 
N/A, intermediate 
function resulting from 
development of TSPA in 
this AMR 
STcoll,FeOx,pHhi Function relating stability of iron- 
(hydr)oxide colloids to pH and ionic 
strength at relatively low pH values 
mol/L Figure 6 and 
11 
I = +0.02  
pH - 0.17 
N/A, intermediate 
function resulting from 
development of TSPA in 
this AMR 
STcoll,FeOx,pHint Function relating stability of iron- 
(hydr)oxide colloids to pH and ionic 
strength at intermediate pH values 
Boolean Figure 6 and 
11 
Boolean 
(pH = 8 AND 
pH = 9) 
Groundwater Colloids 
N/A, intermediate 
parameter resulting from 
development of TSPA in 
this AMR 
CRNcoll,gw Concentration of reversibly attached 
radionuclide element RN associated with 
groundwater colloids 
mol/L Determined in 
PA 
calculation; 
Flow-chart 3b 
N/A

ANL-WIS-MD-000012 REV00 16 April 2000 
Data Tracking Number 
(DTN) 
Parameter 
Name 
Parameter Description Unitsa Basis for 
Value/ 
Source 
Value and 
Distribution 
N/A, intermediate 
parameter resulting from 
development of TSPA in 
this AMR 
Mcoll,gw Mass of groundwater colloids per unit 
volume or mass of water 
mg/L Determined in 
PA 
calculation; 
Flow-chart 3a 
N/A 
MO0003SPAHLO12.004 Mcoll,gw,max Highest observed or expected mass of 
groundwater colloids per unit volume or 
mass of water 
mg/L Figures 9 and 
14 
3E-2 mg/L 
MO0003SPAHLO12.004 Mcoll,gw,min Lowest observed or expected mass of 
groundwater colloids per unit volume or 
mass of water 
mg/L Figures 9 and 
14 
3E-6 mg/L 
MO0004SPAKDS42.005 Kd,RN,coll,gw Distribution coefficient for reversible 
sorption of radionuclide element RN onto 
groundwater colloids 
mL/g 
colloid 
Data 
developed 
from LANL 
data; 
CRWMS 
M&O (2000c) 
Geometric 
mean value: 
Pu, Am = 
1e+3 
log-normal 
distribution; 
geometric 
std. dev. = 1 
order-ofmagnitude 
MO0003SPAION02.003 Ilo-thresh,coll,gw Ionic strength below which groundwater 
colloids are stable 
mol/L Figures 9 and 
14 
0.01 mol/L 
MO0003SPAION02.003 Ihi-thresh,coll,gw Ionic strength above which groundwater 
colloids are unstable 
mol/L Figures 9 and 
14 
0.05 mol/L 
N/A, intermediate 
function resulting from 
development of TSPA in 
this AMR 
STcoll,gw Function relating stability of groundwater 
colloids to ionic strength at intermediate 
ionic strengths 
mg/L Figures 9 and 
14 
Function in 
Section 
6.3.3.3, 
Figure 15g 
Other output 
N/A, intermediate 
parameter resulting from 
development of TSPA in 
this AMR 
CRNcoll Concentration of mobile colloidal 
radionuclide element RN per unit volume 
or mass of water 
mol/L Determined in 
PA 
calculation; 
Figure 15j 
N/A 
N/A, intermediate 
parameter resulting from 
development of TSPA in 
this AMR 
CRNdiss Concentration of dissolved radionuclide 
element RN per unit volume or mass of 
water 
mol/L Input 
parameter 
and 
determined in 
PA 
calculation 
N/A 
N/A, intermediate 
parameter resulting from 
development of TSPA in 
this AMR 
Mcoll Total mass of all mobile colloids per unit 
volume or mass of water 
mg/L Determined in 
PA 
calculation; 
Figure 15j 
N/A 
Input parameters 
N/A, parameter provided 
by GoldSim calculation 
I Ionic strength mol/L Input 
parameter 
N/A 
N/A, parameter provided 
by GoldSim calculation 
pH pH pH units Input 
parameter 
N/A

ANL-WIS-MD-000012 REV00 17 April 2000 
Data Tracking Number 
(DTN) 
Parameter 
Name 
Parameter Description Unitsa Basis for 
Value/ 
Source 
Value and 
Distribution 
N/A, parameter provided 
by GoldSim calculation 
CRNdiss Concentration of dissolved radionuclide 
element RN per unit volume or mass of 
water 
mol/L Input 
parameter 
and 
determined in 
PA 
calculation 
N/A 
LA0003NL831352.002 
LA0003NL831352.003 
[referred to in 
AMR] 
Distribution coefficients and forward and 
reverse adsorption/desorption rates for 
element RN onto waste-form colloids 
source data 
Various LANL 
laboratory 
data 
Various 
NOTE: a mol/L is actually moles per liter of dispersion and is not molarity in the strict sense 
b uncertainty is linked to uncertainty for ionic strength, which is an input parameter; no additional uncertainty is 
added 
c mass is correlated with irreversible radionuclide concentration 
4.2 CRITERIA 
The NRC Total System Performance Assessment and Integration (TSPAI) IRSR (NRC 1998a) 
establishes generic technical acceptance criteria considered by the NRC staff to be essential to a 
defensible, transparent, and comprehensive assessment methodology for the repository system. 
These regulatory acceptance criteria address five fundamental elements of the DOE TSPA model 
for the Yucca Mountain site, namely: 
1. Data and model justification (focusing on sufficiency of data to support the conceptual 
basis of the process model and abstractions) 
2. Data uncertainty and verification (focusing on technical basis for bounding assumptions 
and statistical representations of uncertainties and parameter variabilities) 
3. Model uncertainty (focusing on alternative conceptual models consistent with available 
site data) 
4. Model verification (focusing on testing of model abstractions using detailed process-level 
models and empirical observations) 
5. Integration (focusing on appropriate and consistent coupling of model abstractions). 
Relevant to the topic of this AMR, elements (1) through (4) of the acceptance criteria are 
addressed herein and/or in the supporting Calculation document(s). Element (5) of the NRC 
acceptance criteria, which strictly applies to the completed synthesis of process-level models and 
abstractions, will be addressed separately in the TSPA-SR. As well, this AMR addresses 
relevant KTIs raised in the NFE IRSR Revision 2 (NRC 1999a) and CLST IRSRs Revisions 1 
(NRC 1998b) and 2 (NRC 1999b).

ANL-WIS-MD-000012 REV00 18 April 2000 
4.3 CODES AND STANDARDS 
This AMR was prepared to comply with the above NRC TSPAI acceptance criteria, as well as 
the DOE interim guidance (Dyer 1999) which requires the use of specified subparts/sections of 
the proposed NRC high-level waste rule, 10 CFR Part 63 (64 FR 8640). Subparts of this 
proposed rule that are particularly applicable to data include Subpart B, Section 15 (Site 
Characterization) and Subpart E, Section 114 (Requirements for Performance Assessment). 
Subparts applicable to models are outlined in Subpart E, Sections 114 (Requirements for 
Performance Assessment) and 115 (Characteristics of the Reference Biosphere and Critical 
Group). 
5. ASSUMPTIONS 
Assumptions made as a part of the analysis and model development are presented in Table 2 and 
incorporated into the discussion in Section 6.3, Abstraction and Implementation in GoldSim. 
Table 2. Assumptions Used in the Analysis and Model Development 
Category Assumption Rationale 
Section 
Used 
The colloids produced from corrosion of HLW glass and 
SNF are predominantly smectite; a good approximation 
for abstraction is that all waste form colloids are 
smectite, or are irreversibly attached to smectite. 
ANL experimental data indicate 
that preponderance of HLW glassderived 
colloids is smectite, and 
that degradation of SNF produces 
smectite as well (CRWMS M&O 
2000a). 
6.3 
Metal (hydr)oxide colloids, large particles, and scale 
may form from degradation of the waste package and 
its internal components; a good approximation for 
abstraction is that all metal (hydr)oxides are iron- 
(hydr)oxides. 
The preponderance of metal in the 
waste package steels is iron. 6.3 
Smectite colloids may form from spallation of alteration 
phases from HLW as well as precipitation from solution. 
Based on observations and 
inferences in connection with ANL 
experimental program (CRWMS 
M&O 2000a) 
6.3 
SNF produces colloids less readily under experimental 
conditions approximating repository conditions for 
reasons related to the mechanisms of alteration of the 
SNF. 
Based on observations and 
inferences in connection with ANL 
experimental program (CRWMS 
M&O 2000a) 
6.3 
DOE-SNF and commercial SNF (CSNF) are considered 
together as generic SNF and are not treated 
separately in the abstraction 
The preponderance of SNF is 
CSNF. Metallic fuels may 
produce fine particulates as they 
degrade, but it is currently 
considered that DOE-SNF will not 
adversely affect repository 
performance (see Section 6.1.1.1) 
6.3 
Colloid 
Formation 
The waste form degrades according to in-package 
geochemistry process models and abstractions and are 
represented by a single GoldSim mixing cell 
TSPA methodology (in progress) 6.3 
Colloid 
Stability and 
Concentration 
(continued on 
next page) 
The stability of smectite colloids will depend to varying 
degrees on their concentration and the composition, 
ionic strength, and pH of the solution 
Observed in experiments (e.g., 
Tombacz et al. 1990) 6.3

ANL-WIS-MD-000012 REV00 19 April 2000 
Category Assumption Rationale 
Section 
Used 
The stability of smectite is less sensitive to pH (is 
increased) at higher values of pH, more sensitive to pH 
(is decreased) at lower values 
Observed in experiments 
(Tombacz et al. 1990) 
6.3 
Am is assumed to behave as does Pu in terms of 
mobilization from waste degradation and formation of 
irreversible association with smectite colloids and clay 
layers on HLW 
Reasonable assumption based on 
chemical characteristics of Pu and 
Am 
6.3 
The properties of colloidal smectite and iron- 
(hydr)oxides are well known and applicable to 
radionuclide-bearing colloids 
Smectite and iron-(hydr)oxide 
colloid behavior (Tombacz et al. 
1990; Liang and Morgan 1990) 
6.3 
Iron-(hydr)oxide colloid stability and concentration 
depend upon both the pH and ionic strength that may 
be encountered in the repository 
Generally acknowledged 
relationship (Liang and Morgan 
1990) 
6.3 
Waste form colloids produced at 90oC are stable at 
25oC 
Stability is considered to increase 
as T drops 
6.3 
Some radionuclides released from degraded fuel occur 
as aqueous (dissolved) species and are available to 
sorb reversibly to smectite and iron-(hydr)oxide colloids, 
forming pseudocolloids 
Experimental results and solubility 
determinations indicate that 
radionuclides may dissolve, and 
colloids are available for sorption 
of radionuclides (CRWMS M&O 
2000a) 
6.3 
Smectite colloids contain entrained radionuclide-bearing 
phases; further, all discrete radionuclide-bearing colloidsized 
phases (besides smectite) are entrained in 
smectite colloids 
ANL experimental data indicate 
that smectite colloids contain 
entrained or embedded RNbearing 
phases; conservative 
assumption made because 
reversibility of attachment was not 
analyzed (CRWMS M&O 2000a). 
6.3 
Smectite colloids with or without entrained radionuclidebearing 
phases may have adsorbed dissolved 
radionuclides. 
Smectite colloids are considered 
to have the capability of sorbing 
dissolved radionuclides to form 
pseudocolloids (e.g., Kim 1994). 
6.3 
Pu, Am, Th, Pa, Cs, and Sr are assumed to be the most 
significant radionuclides available for colloid 
association; data indicate that smectite colloids may 
contain Pu, Am, Th, U, Cm, Np, and rare earth 
elements (REEs), and that Pu generally behaves 
similarly to Th, Am and REEs 
These are RNs of interest for 
several TSPA investigators. 
Smectite colloid RN compositions 
observed in ANL experiments 
(CRWMS M&O 2000a). 
6.3 
Radionuclide association with colloids is irreversible 
within the waste package environment, except for 
dissolved radionuclides available for pseudocolloid 
formation. 
A conservative assumption; 
transport times for colloids leaving 
the waste package are considered 
to be short in comparison with 
sorption-desorption times (DTNs: 
LA0003NL831352.002 and 
LA0003NL831352.003). 
6.3 
Radionuclide 
Attachment to 
Colloids 
Adsorption/desorption of radionuclides on colloids to 
form pseudocolloids is defined according to 
experimentally determined rates of adsorption and 
desorption of Pu and Am on several types of colloids, 
and consequently to an effective Kd incorporating both 
the forward and reverse attachment rates. 
Based on results of experiments 
performed at LANL (DTNs: 
LA0003NL831352.002 and 
LA0003NL831352.003) 
6.3

ANL-WIS-MD-000012 REV00 20 April 2000 
Category Assumption Rationale 
Section 
Used 
Desorption of Pu and Am from pseudocolloids is 
assumed to be slow relative to transport rates within the 
waste package, i.e., pseudocolloid sorptive attachment 
is effectively irreversible within the waste package 
Estimates or over-estimates the 
consequences and is bounding 6.3 
Physical 
Filtration of 
Colloids 
Physical filtration of colloids may occur within the waste 
package but is not considered in the abstraction. 
All stable colloids are considered 
to leave a failed waste package 
through the failure opening, which 
over-estimates the consequences 
and is considered bounding 
6.3 
Colloid 
Sorption at 
the Air-Water 
Interface 
Colloid sorption at the air-water interface may occur 
within the waste package, but is not considered in the 
abstraction 
This over-estimates the 
consequences and is considered 
bounding, as colloid sorption to 
stationary air-water interfaces 
would retard colloid transport 
6.3 
Gravitational 
Settling of 
Colloids 
Gravitational settling of colloids may occur within the 
waste package, but is not considered in the abstraction. 
This over-estimates the 
consequences and is considered 
bounding, as all stable colloids are 
assumed not to settle but to leave 
a failed waste package through 
the failure opening 
6.3 
Colloid 
Diffusion 
Colloid diffusion may occur within the waste package, 
but is not considered in the abstraction. 
This over-estimates the 
consequences and is considered 
bounding, as all stable colloids are 
assumed not to settle but to leave 
a failed waste package through 
the failure opening 
6.3 
Microbes and 
Colloidal 
Organic 
Components 
Processes involving interactions between contaminants 
and microbes and organic components (such as humic 
and fulvic acids) are not considered in this abstraction 
This is considered to overestimate 
the consequences and is 
considered bounding, as 
biological influences have been 
known to encourage colloid 
agglomeration 
6.3 
Colloid 
Transport 
All stable colloids are assumed to leave a failed waste 
package through the failure opening 
This over-estimates the 
consequences and is considered 
bounding; processes such as 
sorption to in-package 
components, gravitational settling, 
and filtration would tend to retain 
colloids in the waste package. 
6.3

ANL-WIS-MD-000012 REV00 21 April 2000 
6. ANALYSIS/MODEL 
Colloids may affect repository performance if they are generated in significant quantities; are 
stable within the waste package, Engineered Barrier System (EBS), and unsaturated and 
saturated zones (UZ and SZ, respectively); carry a significant radionuclide load; and transport 
readily. Processes relevant to the waste package environment are discussed in this section in the 
context of conceptual models and process models and in the approach to abstraction used in this 
effort. 
6.1 CONCEPTUAL MODELS AND BACKGROUND 
Features, events, and processes (FEPs) relevant to colloid processes in the waste package are 
listed and summarized in Table 3. The complete FEPs are presented in Attachments I through 
XV. 
Table 3. Features, Events, and Processes (FEPs) Included in this AMR 
FEP YMP No. FEP Title FEP Summary FEP Location in AMR 
2.1.09.14.00 Colloid Formation in 
Waste and EBS 
Colloids may form as true colloids, 
colloids formed by co-precipitation with 
waste alteration phases, pseudocolloids, 
and microbial colloids. 
Sections 6.1.1, 6.2.1, 
6.3 
2.1.09.15.00 
Formation of True Colloids 
in Waste and EBS 
True colloids (radionuclide intrinsic 
colloids) are colloid-sized assemblages 
of hydrolyzed, polymerized dissolved 
radionuclides. 
Sections 6.1.1, 6.2.1, 
6.3 
2.1.09.16.00 
Formation of Pseudocolloids 
(Natural) in Waste 
and EBS 
Pseudocolloids are colloid-sized 
particles with attached radionuclides. 
Natural pseudocolloids may include 
mineral fragments in groundwater, 
microbes and fragments, and organic 
acids. 
Sections 6.1.3, 6.2.3, 
6.3 
2.1.09.16.01 Colloid Phases Produced 
by Co-precipitation (in 
Waste and EBS) 
Colloidal phases containing embedded 
radionuclides and radionuclide phases 
occur at the waste form. 
Sections 6.1.3, 6.2.3, 
6.3 
2.1.09.17.00 Formation of Pseudocolloids 
(Dependent on 
Corrosion Products) (in 
Waste and EBS) 
Pseudocolloids are colloid-sized 
particles with attached radionuclides. 
Colloid-sized particles to which 
radionuclide may attach may form from 
the corrosion of materials introduced 
into the repository during construction 
and waste emplacement. 
Sections 6.1.3, 6.2.3, 
6.3 
2.1.09.18.00 Microbial Colloid Transport 
in the Waste and EBS 
Colloid-sized microbes and microbe 
fragments may form pseudocolloids. 
Microbes can also affect waste package 
corrosion rates, possibly affecting 
corrosion colloid production rates, as 
well as increasing colloid size, resulting 
in increased gravitational settling and 
filtration. 
Sections 6.1.8, 6.2.8 
2.1.09.19.00 Colloid Transport and 
Sorption 
Interactions between radionuclidebearing 
colloids and the waste form and 
EBS materials may result in retardation 
from sorption. 
Sections 6.1.9, 6.2.9, 
6.3 
2.1.09.20.00 Colloid Filtration Filtration processes may affect transport 
of radionuclide-bearing colloids in the 
Sections 6.1.4, 6.2.4

ANL-WIS-MD-000012 REV00 22 April 2000 
FEP YMP No. FEP Title FEP Summary FEP Location in AMR 
waste and EBS. 
2.1.09.21.00 Suspensions of Particles 
Larger Than Colloids 
Groundwater flow through the waste 
could remove radionuclide-bearing 
particles larger than colloids by a rinse 
mechanism. 
Not analyzed in AMR 
2.1.10.01.00 Biological Activity in 
Waste and EBS 
Biological activity in the waste and EBS 
may affect radionuclide transport 
through, for example, formation of 
colloids and biofilms. 
Sections 6.1.8, 6.2.8 
2.1.13.03.00 Mutation Radiation fields could cause mutation of 
microorganisms, leading to unexpected 
chemical reactions and impacts. 
Not analyzed in AMR 
2.1.09.22.00 
(proposed) 
Colloid Sorption at the Air- 
Water Interface 
Colloids may sorb irreversibly at the 
gas-water interface under partially 
saturated conditions. 
Sections 6.1.5, 6.2.5 
2.1.09.23.00 
(proposed) 
Colloid Stability and Concentration 
Dependence on 
Aqueous Chemistry 
Colloid stability is determined in part by 
ionic strength, pH, and cation 
concentrations of the aqueous medium. 
Sections 6.1.2, 6.2.2, 
6.3 
2.1.09.24.00 
(proposed) 
Colloid Diffusion Colloids may diffuse into intercrystalline 
porosity and be physically retarded. 
Sections 6.1.7, 6.2.7 
2.1.09.25.00 
(proposed) 
Colloid Gravitational Settling 
Larger colloids may experience 
gravitational settling, thereby inhibiting 
transport. 
Sections 6.1.6, 6.2.6 
6.1.1 Colloid Formation 
Several types of colloids are expected to be generated within the Yucca Mountain repository 
system. Colloids will be present at the waste form and in the drift and the unsaturated and 
saturated zones. At the waste form, colloids are produced during the degradation of the waste 
form and waste package. In the repository, additional colloids may be produced from the 
degradation of steel and concrete components of the EBS. Several types of colloids (mineral and 
organic colloids) are present in the unsaturated zone, as well as the saturated zone, and may be 
introduced into the repository. Mineral colloids include clays, iron-(hydr)oxides (goethite, 
hematite), and silica; organic colloids include microbes and fragments as well as humic and 
fulvic acids. Formation and introduction of colloids to the waste form/waste package 
environment are discussed below in the context of the conceptual models considered in the 
abstraction. 
6.1.1.1 Colloids Derived from Waste Form Degradation 
Several types of radionuclide-bearing colloids may result from the degradation of waste forms. 
Formation of radionuclide-bearing colloids is discussed in this section; colloid stability and the 
nature of the attachment of radionuclides to colloids is discussed in Sections 6.1.2 and 6.1.3, 
respectively. 
Real colloids are formed from polymerization of hydrolyzed radionuclide ions (Figure 1) (Kim 
1994). For real colloids to form it is necessary that the fluid be saturated with respect to a solid 
phase, sometimes amorphous; thus, the formation of real colloids is solubility limited. No real 
colloids were detected in the experimental investigations at ANL (CRWMS M&O 2000a), and 
so they are not included here.

ANL-WIS-MD-000012 REV00 23 April 2000 
Pseudocolloids may be formed by the incorporation, adsorption, and ion exchange of dissolved 
radionuclides onto existing colloids, either formed from waste form degradation or introduced 
with groundwater (Figure 1) (Kim 1994). Thus, radionuclides, which might otherwise sorb to 
minerals in the rock or other immobile substrates, instead attach to mobile colloidal particles and, 
added to the dissolved radionuclides, increase the mobile radionuclide source term. Colloids 
may facilitate migration within the repository system and potentially the far-field. 
Waste form colloids, or primary colloids, may provide the most significant contribution to 
colloid-facilitated radionuclide transport. These colloids form from nucleation of colloids from 
waste form dissolution and spallation of colloid-sized waste form alteration products that contain 
radionuclides and have the potential for increasing mobile concentration to levels higher than 
achievable with real colloids or pseudocolloids (Figures 1 and 2) (CRWMS M&O 2000a). 
O-Pu-O-Pu-O 
O-Pu-O-Pu-O 
O 
O
O 
O
O O 
groundwater 
colloid Rn+ 
waste form 
alteration phases 
Rn+ 
(a) (b) (c) 
NOTE: Several types of radionuclide-bearing colloids are depicted: (a) pseudo-colloids, (b) real colloids, and (c) 
waste form colloids. The radionuclide (Rn+) associated with the pseudo and waste form colloids can be an 
ionic species, a real colloid, or a discrete radionuclide-bearing phase. The size range for the colloids is 
approximately 1 to 1000 nm. (CRWMS M&0 2000a, Fig. 1) 
Figure 1. Several Types of Radionuclide-Bearing Colloids 
Three broad categories of waste form are to be placed in the repository. The most abundant 
waste type is commercial spent nuclear fuel (CSNF) from commercially owned and operated 
electric power reactors. DOE spent nuclear fuel (DOE-SNF) is a diverse collection of waste 
from reactors at U.S. Department of Energy (DOE) nuclear complex sites. Defense high-level 
radioactive waste (DHLW) is a borosilicate glass-based waste form containing radionuclides. In 
the following descriptions, CSNF and DOE-SNF are treated together. 
Corrosion of DHLW-Testing on corrosion of DHLW has gone on for over ten years in the case 
of static tests in which glass samples are immersed in fluid and for over four years in the case of 
drip tests in which fluid is dripped at specified rates onto the glass sample. In the static tests, 
surface area (SA) of the sample was varied relative to the volume (V) of fluid used; this 
proportion is referred to as SA/V. SA was determined using geometric estimates of crushed 
materials, not by a gas adsorption method. Glass waste forms tested include Savannah River 
Laboratory (SRL) and West Valley Demonstration Project (WVDP) glasses. Fluids used were 
equilibrated J-13 water (EJ-13) and deionized water. Methods used to characterize colloids 
produced during the tests include transmission electron microscopy (TEM) and filtration 
(analysis of filtrates with inductively coupled plasma-mass spectrometry or alpha spectroscopy; 
CRWMS M&O 2000a). Descriptions of the tests may be found in CRWMS M&0 (2000a) and 
Ebert (1995).

ANL-WIS-MD-000012 REV00 24 April 2000 
NOTE: Schematic of colloid formation from waste form corrosion whereby several processes are represented: (1) 
spallation of radionuclide (RN) 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 RN 
species or RN-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 micron or solution 
chemistry destabilizes the colloids, deposition, or coagulation and gravitational settling of RN-bearing 
alteration phases occur (modified Figure 2 from CRWMS M&0 2000a). 
Figure 2. Schematic of Colloid Formation from Waste Form Corrosion 
Colloids produced from both types of glass waste testing were primarily smectite clay containing 
discrete radionuclide-bearing phases which are incorporated (entrained) in the clay (CRWMS 
M&0 2000a). Iron silicate colloids were also observed. The entrained phases were identified as 
primarily brockite (thorium calcium orthophosphate) and an amorphous thorium-titanium-iron 
silicate, similar to thorutite. The phases also contained other actinides and REEs. Uranium was 
detected within the clays and iron silicates in some samples. 
Currently, no data are available on the chemical and physical properties of the entrained phases, 
so that their solubility and disassociation/association properties with the clay cannot be derived 
and used in the model. The colloidal properties of the entrained radionuclide-bearing phases, 
therefore, are assumed to be governed by the properties of the smectite, and the stability (and 
therefore mobility) of smectite colloids will control the mobility of the entrained phases. 
(Testing is currently underway at LANL to determine the attachment/detachment properties of 
the entrained radionuclide-bearing phases in smectite colloids produced from the ANL glass 
tests.) Accordingly, the radionuclides are assumed to be irreversibly attached to the smectite 
colloids.

ANL-WIS-MD-000012 REV00 25 April 2000 
As the glass corrodes, the clay alteration product forms, grows, and evolves as a layer on the 
surface of the glass. The clay layer grows as the glass beneath the clay alters and as the clayforming 
elements saturate in the solution and precipitate on the clay surface. Through these 
mechanisms colloid-sized clay that does not attach to the clay layer also precipitates from the 
solution (CRWMS M&O 2000a). 
It was observed in the static tests that colloids developed and increased in concentration 
(population) with time, up to a point where the colloid concentration reached a maximum and 
then the dispersion became unstable. This was attributed by the investigators primarily to the 
concomitant increase in ionic strength to a threshold above which the colloids flocculated and 
settled by gravity (CRWMS M&O 2000a). 
It was noticed that as the drip tests proceeded, whereas relatively soluble elements such as boron 
and neptunium increased smoothly with time, after about eight years the concentration in the 
leachate of the less soluble elements Pu and Am increased sharply (Figure 3). This was 
interpreted as evidence that colloid-sized fragments of the clay layer spalled from the clay layer 
on the surface of the corroding glass and were released into the fluid. This behavior was not 
observed in the static tests, in which the glass sample was not subjected to dripping water influx 
and remained submerged in the leachate. 
10-4 
10-3 
10-2 
10-1 
100 
101
0 2 4 6 8 10 12 
Np 
Pu 
Am 
B 
Time (years) 
Normalized Release (g/m2) 
NOTE: The cumulative normalized release of the elements B, Np, Pu, and Am from one of the tests in the N2 Series 
[Fortner-1999]. (Figure 10 from CRWMS M&0 2000a) 
Figure 3. The Cumulative Normalized Release of the Elements B, Np, Pu, and Am

ANL-WIS-MD-000012 REV00 26 April 2000 
The experimental data on colloid generation resulting from HLW glass corrosion were used to 
develop a conceptual model for the rate and evolution of radionuclide release, described here, 
and an expression for use in predicting the rate of radionuclide release, described in Section 
6.4.2, Alternative Models. The amount of boron released was used as a marker for the amount of 
glass corrosion as B was observed to increase smoothly in the leachate with glass corrosion, and 
it did not become incorporated in alteration phases. The normalized (to SA) mass loss of Pu, 
NL(Pu), from the glass was compared to the normalized mass loss of B, NL(B), as a function of 
reaction time; the ratio of the two indicates the quantity of Pu released relative to the quantity of 
glass corroded. In some of the tests, NL(Pu) was consistently lower than NL(B) by at least one 
order of magnitude. Marked decreases from this ratio can be attributed in specific tests to an 
increase in the glass corrosion rate due to secondary phase formation, as well as destabilization 
of the colloids (CRWMS M&0 2000a). NL(Pu)/NL(B) can be less than 10-3 under solution 
conditions that inhibit colloid stability (see Section 6.4.2.1.2). 
During the early stages of waste glass corrosion, the glass dissolves and releases components at a 
rate proportional to the rate of corrosion of the waste form. As this occurs, clay colloids nucleate 
in the solution and clay alteration layers form on the surface of the glass. These clay phases, 
which contain entrained radionuclide-bearing phases, also sorb ionic radionuclide species and/or 
colloidal radionuclide-bearing phases. If the ionic strength of the solution is low, the colloids are 
generally stable (see Section 6.1.2) and the rate of formation of the radionuclide-bearing colloids 
is proportional to the amount of altered glass (CRWMS M&0 2000a). 
As the glass alters, the rate of radionuclide-bearing colloids production decreases; this is 
attributed by the investigators to colloid deposition (by sorption) on the fixed clay alteration 
layer. Once the cumulative normalized B release is greater than approximately 1-3 g m-2, the 
radionuclide release is controlled by spallation of the clay layer. Figure 4 illustrates the observed 
increases in colloid production due to the onset of spallation. 
Thus, radionuclide release may be viewed as 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 (CRWMS M&0 2000a). A preliminary quantitative version of this model is presented 
in Section 6.4.2, Alternative Models. The observed concentrations of HLW-derived colloids is 
discussed in Section 6.1.2, Colloid Stability and Concentration. 
Corrosion of Commercial SNF-Testing on the corrosion of SNF has been conducted for over 
four years. Two types of spent fuel from pressurized water reactors with different burnups were 
used. EJ-13 water was introduced in two ways, at the bottom of the test vessel to maintain 100% 
relative humidity and onto the top of the spent fuel fragments. Generated colloids were 
characterized by using the same analytical techniques as were used in the glass testing. 
Descriptions of the SNF tests may be found in CRWMS M&0 (2000a) and Finn et al. (1994).

ANL-WIS-MD-000012 REV00 27 April 2000 
(a) 
(b) 
(c) 
NOTE: Plutonium release rate (calculated for each sampling interval) versus cumulative boron 
release: (a) N2-9, (b) N2-10, (c) N2-12. The total plutonium release is used, including 
vessel acid strip. (Figure 29 from CRWMS M&0 2000a) 
Figure 4. Plutonium Release Rate versus Cumulative Boron Release

ANL-WIS-MD-000012 REV00 28 April 2000 
(d) 
(e) 
(f) 
NOTE: Plutonium release rate (calculated for each sampling interval) versus cumulative boron 
release: (d) N3-9, (e) N3-10, (f) N3-12. The total plutonium release is used, including 
vessel acid strip. (Figure 29 from CRWMS M&0 2000a) 
Figure 4. Plutonium Release Rate versus Cumulative Boron Release (continued) 
In comparison to what was observed in the glass waste form tests, very little radionuclidebearing 
colloidal material was detected in the fluid samples. (It is possible that this was partly a 
result of the test vessel configuration [CRWMS M&O 2000a].) In one sample, colloids

ANL-WIS-MD-000012 REV00 29 April 2000 
characterized were primarily a hydrous uranium oxide (schoepite) and a partially crystalline 
uranium silicate (soddyite). The colloids contained the rare earth elements (REEs) La, Ce, Pr, 
Nd, Sm, and Eu. Plutonium was not detected in these colloids; however, alteration on the 
corroded surface of the spent fuel contained significant concentrations of Pu, along with Zr, Ru, 
and REEs. 
In another sample, smectite clay colloids were formed that were similar structurally and 
compositionally to the smectites produced in the glass tests, although they contained Mn, Ni, Fe, 
and Cr. 
Elevated Pu concentrations in surficial corrosion mentioned above were found in several samples 
of both types of SNF. The Pu-bearing materials were overlain by a mat of cryptocrystalline Ubearing 
phase occurring as agglomerates of colloid-sized particles. The Pu layers were about 
100 nm thick and may be colloidal aggregates within the alteration layer or may have 
accumulated as a result of filtering by the outer uranium silicate-aggregate mat (CRWMS M&0 
2000a). It is the opinion of the investigators that the Pu layers were formed during dissolution of 
the fuel (CRWMS M&0 2000a), based largely on the fact that the Pu layer is sandwiched 
between the uranyl alteration phases and the fuel (CRWMS M&0 2000a). 
Throughout the period of SNF testing, it was observed that most of the Pu was retained in the 
surficial corrosion concentration phase and little was released into the fluid. This may be due to 
the fact that the mat of uranium-bearing agglomerates, in effect, maintains the structural integrity 
of the alteration layer and prevents the Pu concentrations from spalling(CRWMS M&0 2000a). 
Based on experimental results, it is apparent that the Pu concentration phase is not released in a 
form in which it is irreversibly attached to colloids, and it is assumed in the abstraction that it 
will not be released in such form. (However, if these Pu regions became friable, e.g., due to 
humidity transients, Pu colloids might be released; another conservative approach would allow 
for the possibility of spallation of the Pu-concentration regions, although this process is not 
supported by the experimental data.) Further, filtered material with measurable associated Pu 
does not appear in TEM images to contain entrained radionuclide-bearing bases in the same 
manner as HLW glass corrosion colloids. 
The experimental data on colloid generation resulting from SNF corrosion were used to develop 
a conceptual model for the rate and evolution of radionuclide release, described here, and an 
expression for use in predicting the rate of radionuclide release, described in Section 6.4.2. The 
amount of technetium released was used as a marker for the amount of SNF corrosion, as Tc was 
observed to increase smoothly in the leachate with glass corrosion, and it did not become 
incorporated in alteration phases. The normalized mass loss of Pu, NL(Pu), from the glass was 
compared to the normalized mass loss of Tc, NL(Tc); the ratio of the two indicates the quantity 
of Pu released relative to the quantity of SNF corroded. 
Dissolved Pu concentrations in the fluid were determined to be between 10-10 and 10-8, less than 
to approximately the same as, the solubility of Pu in J-13 water determined by Efurd et al. 
(1998). Further, no correlation between the rate of Pu-bearing colloid formation and rate of SNF 
corrosion was observed. Consequently the Pu release as a function of Tc release was bounded 
by assuming that NL(Pu)/NL(Tc) is equal to 10-5 (see Section 6.4.2.1.2).

ANL-WIS-MD-000012 REV00 30 April 2000 
In the conceptual model, most of the Pu and Am remain at the surface of the corroded fuel 
beneath and are protected by the mat of agglomerated uranium silicates. The Pu and Am 
concentrations associated with pseudocolloids are assumed to be similar to the Pu solubility as 
described above. 
Corrosion of DOE-SNF-DOE-SNF consists of more than 250 distinct types of fuel. One type, 
an irradiated uranium metal fuel from the N-Reactor at Hanford, is being characterized as part of 
an ongoing program at ANL. In addition, drip tests and flowthrough tests are being performed 
on unirradiated uranium metal fuel. Preliminary results suggest that the metal degrades 
relatively rapidly, and there is evidence that colloid-size particulate material may form as a result 
of fuel degradation. It should be emphasized that the test data are preliminary and represent the 
initial results of an ongoing program. Clearly, however, the possible effects on repository 
performance of rapidly-formed colloid-size uranium particles must be considered. 
A potentially mitigating factor is the small fraction of the DOE-SNF inventory represented by NReactor 
fuel. Out of almost 4000 co-disposal packages, only 160 are planned to contain NReactor 
fuel, or four percent of the co-disposal packages. On this basis, and because the data are 
preliminary and the ANL test program is ongoing, it is currently assumed that corroded Nreactor 
fuel will have little or no effect on repository performance. However, the assumption 
may be modified to reflect new information obtained during a verification process. It is 
proposed that verification be performed in three parts: (1) examine current data from the 
ongoing testing program at ANL; (2) perform sensitivity analyses using GoldSim; and (3) track 
the test program and its findings as the program proceeds. 
6.1.1.2 Iron-(Hydr)oxide Colloids from Internal Waste Package Corrosion 
As it is likely that some corrosion of the metallic components occurs and oxidizing conditions 
exist and that most of the metal is iron contained in several types of steel, it is assumed in the 
abstraction that iron will corrode and form iron-(hydr)oxides. It is further assumed that the iron- 
(hydr)oxides will occur in three forms: immobile scale, large particles that will settle out, and 
colloid-sized particles. 
The proportions of these iron-(hydr)oxide occurrences may be represented by ranges or 
estimated from calculations of quantities of corrosion products based on known masses of 
metallic components. The iron-(hydr)oxide phases may be expected to provide abundant 
sorption sites for dissolved (aqueous) radionuclides in amounts determined by the appropriate 
effective Kds. Stable iron-(hydr)oxide colloid suspensions may serve to increase the mobility of 
sorbed radionuclides (see Section 6.1.3). Their stability is controlled by fluid conditions 
including ionic strength and pH (see Section 6.1.2). 
6.1.1.3 Natural Colloids from Unsaturated Zone Groundwater 
Three types of groundwater colloids may be present at Yucca Mountain: 
1. Mineral fragments are generally hydrophobic (see Section 6.1.2), hard particles that are 
kinetically stabilized or destabilized by electrostatic forces and may consist of crystalline 
or amorphous solids (Kim 1994). Steric coatings may enhance the stability of mineral

ANL-WIS-MD-000012 REV00 31 April 2000 
fragments by preventing close contact. Mineral fragments may act as substrates for 
sorption of radionuclides (they may also consist of precipitated or coprecipitated actinide 
solids). 
2. Humic substances are generally hydrophilic particles that are stabilized by solvation 
forces. They can be powerful substrates for uptake of metal cations and are relatively 
small (less than 100,000 atomic mass units). 
3. Microbes are relatively large colloidal particles that are stabilized by hydrophilic coatings 
on their surfaces, which behave as steric stabilizing compounds. They may act as 
substrates for extracellular actinide sorption, or they may actively bioaccumulate 
radionuclides intracellularly. 
Clays, silica, hematite, and goethite colloids occur in groundwater in the vicinity of Yucca 
Mountain, and it is assumed that small quantities of these colloids will enter a failed waste 
package and be available to interact with released radionuclides. The presence and potential 
influence of natural colloids on formation of radionuclide-bearing colloids in the waste package 
are considered by assuming that groundwater colloids are smectite and iron (hydr)oxide. 
An assessment of the concentration of humic substances in groundwaters collected near the 
Yucca Mountain was conducted by Minai et al. (1992). In that study, humic substances were 
extracted from several thousand gallons of J-13 well water through a series of steps including 
acidification, purifying with ion exchange columns, and freeze drying. The humic substances 
that were collected characterized with NMR and other spectroscopic techniques. Experiments 
were conducted to characterize site binding densities and complexation with Am. Results and 
supporting calculations indicated that the presence of humic substances in J-13 well water could 
affect oxidation speciation by reducing some radionuclide species. Considering the presence of 
calcium in J-13 well water, however, the authors estimated that the effective complexation 
capacity of humic substances is about 2.3 x 10-10 eq/L (equivalents per liter) at pH 6.9 and 2.7 x 
10-11 eq/L at pH 8.2 (Minai et al. 1992, p. 199). Considering the very low complexation capacity 
of humic substances in this system, they are not considered as groundwater colloids. 
Because of the relatively large sizes of microbes, they are very susceptible to filtration in 
geologic media. Consequently, microbe-facilitated contaminant transport is not considered in 
this abstraction. Refer to section 6.1.8 for more discussion on microbes. 
Consequently, mineral colloids are considered in this abstraction and will, henceforth, be 
referred to generically as groundwater colloids. 
6.1.2 Colloid Stability and Concentration 
In order for radionuclide-bearing colloids to affect repository performance, the colloidal 
dispersion must be stable for the time frame of transport and must carry significant amounts of 
radionuclides. Transport times can range from days/months/years for transport out of a failed 
waste package under a large seep, to hundreds of thousands of years for retarded transport to the 
accessible environment. Thus, some relatively unstable colloids generated at the waste form 
may persist long enough to be transported out of the waste package, increasing the radionuclide 
release from the waste package, but not traveling a significant distance away from the repository.

ANL-WIS-MD-000012 REV00 32 April 2000 
The more stable colloids, however, may remain suspended for years and travel a much greater 
distance. The clay colloids observed in the HLW and SNF corrosion tests are reasonably 
considered to have similar stability characteristics as natural clay colloids because the waste 
form-derived clays exhibit similar crystallographic and chemical properties as natural clays. The 
iron-(hydr)oxide colloids assumed to form from internal corrosion of the waste package are 
reasonably considered similar to natural iron-(hydr)oxide colloids for the same reasons. 
Following is a discussion of parameters affecting colloid stability. 
Background-The stability and the coagulation, or agglomeration, of colloids is controlled by 
electrostatic and chemical processes at the colloid surfaces. In the 1940s and 1950s research on 
particle coagulation centered around the electrostatic properties of the particles in which the 
DLVO theory (independently developed by Derjaguin and Landau, and Verwey and Overbeek) 
of the particle electrical double layer was used to explain particle stability in terms of surface 
charges and potentials, the ionic strength of solutions, and the van der Waals forces between 
particles (Liang and Morgan 1990, p. 33). It was subsequently realized that the double-layer 
theory by itself was not complete, and that the origin of surface charges and the role of specific 
chemical interactions in the stabilization and destabilization of colloidal particles should be 
considered. At low surface potential and low surface charge density of particles, experiments to 
verify DLVO theory agreed with theoretical prediction. However, at high surface potential and 
charge density, due to the large repulsive forces of the particles, the theory predicted orders of 
magnitude higher stability ratios than were observed (Liang and Morgan 1990, p. 34). 
Beginning in the 1970s, surface complexation adsorption models were developed (Langmuir 
1997b, p. 771); these models were used to analyze changes in particles surface potential and 
charge brought about by complexation reactions (described by mass law equations) at the 
particle surface involving specific adsorbed chemical species (Dzombak and Morel 1987, p. 430; 
Liang and Morgan 1990, p. 34). 
Hydrophobic colloidal particles are kinetically stabilized and destabilized by electrostatic forces. 
Mineral fragments, which are a hydrophobic colloid type, are affected by ionic strength in this 
way. Hydrophilic colloidal particles are stabilized by solvation forces, which are largely 
independent of the ionic strength of the dispersant. This type of colloidal particle is essentially a 
dissolved macromolecule. Humic materials are an example of the traditional hydrophilic colloid 
type. 
DLVO Theory and Surface Complexation-The surfaces of dispersed, stable colloids have an 
electric charge sufficient to maintain the particles separation by mutual repulsion. The surface 
charge of a particle can have two general causes: (1) specific characteristics of the crystal 
structure (see below), and preferential adsorption of specific ions onto the surface (van Olphen 
1977, p. 18). The surface charges are compensated by the attraction of counter-ions, which have 
the opposite charge to the surface charges and to the vicinity of the particle surface. The charges 
at the particle surface and the oppositely charged, closely associated counter-ions make up the 
electric double layer (EDL) (van Olphen 1977, p. 18). The counter-ions are electrostatically 
attracted to the oppositely charged surface. However, they tend in water to diffuse away from 
the surface along a decreasing concentration gradient. As a result, counter-ions develop an 
atmosphere around the colloid that decreases in concentration away from the particle (van 
Olphen 1977, p. 30).

ANL-WIS-MD-000012 REV00 33 April 2000 
There are three causes of surface charge related to crystal structure (Langmuir 1997a, pp. 346-7): 
(1) isomorphous substitution in the crystal lattice (e.g., Al3+ for Si4+ in the tetrahedral layer and 
Mg2+ for Al3+ in the octahedral layer resulting in an excess of O2- bonds, the chief cause of 
negative surface charge in smectites); (2) lattice imperfections or defects (e.g., deficits in 
octahedral Al3+ or interlayer K+ causing a net negative surface charge, also important for 
smectites); and (3) broken or unsatisfied bonds (e.g., at crystal plate corners and edges usually 
resulting in a net negative surface charge due to exposed O2- and OH-, important for smectites). 
As clay particles decrease in size, the importance of broken bonds increases (Langmuir 1997a, p. 
346). 
The surface charge of oxides, hydroxides, phosphates, and carbonates originates chiefly from 
ionization of surface groups or surface chemical reactions (Langmuir 1997a, p. 349). Surface 
charge among this group is strongly dependent upon pH, being positive at lower pH and negative 
at higher pH. For example, in oxides and hydroxides, this is due largely to the adsorption of H+ 
at lower pH and OH- at higher pH. The pH at which the net surface charge is neutral is called the 
zero point of charge (ZPC); if the change in charge is solely due to adsorption of H+ or OH- ions, 
the pH at neutrality is called the point of zero net proton charge (PZNPC) (Langmuir 1997a, p. 
350). 
Far from the ZPC or PZNPC, the surface-charge density is relatively high; the particles are 
mutually repelled, and the suspension is stable. At or near the ZPC or PZNPC the repulsion 
force decreases to near zero, and the particles are more likely to agglomerate (stick) when 
particle collisions occur. 
Stability of Smectite Colloids as a Function of pH-The ZPC of smectite is approximately pH 
2. Since a pH this low is not anticipated in the repository, it is a reasonable assumption that 
smectite colloids will remain in a stable pH range much of the time under anticipated repository 
conditions (unless ionic strength exceeds a certain threshold). However, it should be recognized 
that with decreasing pH, the charge density of smectite particle surfaces will decrease as more 
and more H+ sorb to the surface and offset negative charges, generally decreasing stability (Buck 
and Bates 1999; Tombacz et al. 1990). This decrease in stability has been taken into account in 
Section 6.2.2 and 6.3. 
The HLW tests conducted at ANL resulted in measured pH ranging between approximately 9 
and 11.5 (Buck and Bates 1999), which is part of the range at which smectite colloids exhibit the 
highest surface charge and, hence, are most stable. Tombacz et al. (1990) investigated the 
stability of montmorillonite (a smectite clay mineral) suspensions as a function of pH and ionic 
strength in an 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 5). 
Stability of Iron-(Hydr)oxide Colloids as a Function of pH-The ZPC of iron-(hydr)oxide 
colloids is around pH 8.5, and at this pH they will be unstable and tend to agglomerate. At 
higher or lower pH, however, iron-(hydr)oxide colloids may be stable, depending upon ionic 
strength (see Sections 6.2.2 and 6.3.6). Liang and Morgan (1990) demonstrated that for a given 
ionic strength iron-(hydr)oxide stability increases as pH both increases and decreases away from 
the ZPC (Figure 6). In general the higher the ionic strength, the wider the pH range about the 
ZPC that iron-(hydr)oxide is unstable. For example, at an ionic strength of 0.05 M, iron

ANL-WIS-MD-000012 REV00 34 April 2000 
(hydr)oxide is unstable between approximately pH 6 and 11. This relationship is incorporated in 
the abstraction and is discussed in Section 6.2.1.3. 
NOTE: Data points represent combinations of ionic strength and pH at which montmorillonite stability is 
significantly decreased. Data from Tombacz et al. (1990). 
Figure 5. Experimental Determination of Montmorillonite Stability as a Function of pH and 
Ionic Strength 
Ionic Strength (NaCl) 
2 6 8 10 
zpc 
pH 
0.01 
0.02 
4 
0.03 
0.04 
0.05 
MONTMORILLONITE 
UNSTABLE 
MONTMORILLONITE 
STABLE

ANL-WIS-MD-000012 REV00 35 April 2000 
NOTE: The pH of zero proton condition is indicated. Dashed lines are drawn through the 
experimental points as a guide. The solid line represents a DLVO model calculation for 
ionic strength 0.05 M. Note that Wexp 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; a value of zero indicates rapid agglomeration. At ionic 
strengths of about 0.5 m or more, the stability artio is very low between pH values of 
about 6 and 11. A relatively low log Wexp of approximately 0.2 was selected for 
abstraction of pH and ionic strength data for use in Figure 11 (Figure 1 from Liang and 
Morgan 1990). 
Figure 6. Experimentally Derived Stability Ratio, Wexp, of a Hematite Suspension Plotted as a Function 
of pH for Differing Ionic Strengths 
Ionic Strength-In nearly pure water, dissolved ions are likely to contact only other water 
molecules. As the concentrations of the ions increase, the ions are likely to come in contact with 
one another. The interactions between and among the ions is proportional to their charge, as 
reflected in the definition of ionic strength: 
I =  S (mi zi
2) (Eq. 1) 
where I (and m) is in molal units and zi is the charge of ion i (e.g., Langmuir 1997a). 
According to DLVO theory, colloids become unstable as ionic strength in a solution increases 
due to compression of the diffuse counter-ion atmosphere and decrease in surface potential, i.e., 
decrease in repulsive forces. Further, the degree of compression of the EDL is dependent on the 
concentration and valence of ions of opposite sign from that of the surface charge (van Olphen

ANL-WIS-MD-000012 REV00 36 April 2000 
1977, p. 34). The effect of valence is captured by the Schulze-Hardy rule, which incorporates 
the principle that for a given counter-ion concentration, the higher the valence of the counterions, 
the more unstable the colloids and the greater the tendency to flocculate (van Olphen 1977, 
p. 24). 
In his treatise on clay colloidal dispersions, van Olphen (1977 p. 24) summarized empirically 
determined critical coagulation concentration (c.c.c.) values as follows: 
Counterion Valence c.c.c. (M) 
1 25,000 to 150,000 
2 500 to 2000 
3 10 to 100 
At pH values below the pHZPC for a mineral, the negatively charged counterions are important; 
above the pHZPC, positively charged counterions are important. 
For comparison, concentrations of counterions in J-13 well water are as follows (CRWMS M&O 
2000d): 
Ion Valence Concentration (M) 
Na +1 1990 
K +1 129 
Ca +2 324 
Mg +2 82.7 
Al +3 0.0255 
Cl -1 201 
S (sulfate) -2 192 
Note that none of the concentrations of individual ions in J-13 well water exceed van Olphens 
empirical c.c.c. values, although Ca2+ is close. 
Stability and Concentration of Waste Form Colloids-Colloids were formed in the course of 
static HLW tests on several of the glasses for different SA/V. 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 very low quantities 
(Figure 7) (CRWMS M&0 2000a). As this threshold was approached, it was also observed for 
one glass that the colloid size increased significantly due to aggregation of the colloids. The 
threshold at which flocculation occurred was approximately I = 0.05m. During the drip tests, 
ionic strength remained below this value, and colloids were stable throughout the tests. 
Stability and Concentration of Groundwater Colloids-Colloids are thought to exist naturally 
in groundwater in all 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 iron-(hydr)oxides and clay colloids is a potentially important transport

ANL-WIS-MD-000012 REV00 37 April 2000 
mechanism, and is thought to be generally more prevalent than transport by true colloids 
(McCarthy and Zachara 1989). 
McCarthy and Degueldre (1993) conducted a significant effort to characterize the colloids of 
groundwaters from around the world from crystalline and sedimentary rocks in saturated and 
unsaturated hydrologic regimes. They showed that the colloid concentration and stability is 
dependent, among other things, upon pH, redox potential, ionic strength (including relative 
concentrations of the major cations Na, K, Ca, and Mg), counterion valence, and organic carbon; 
these factors were studied in the investigation. The results indicate that, in general, colloids 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 8). This provides a 
guideline for assessing colloid stability and allows for very rough estimation of concentration. 
Whereas the DLVO calculations provide useful inferences on mechanisms of colloid 
destabilization, these empirical observations are more useful in developing approaches to PA 
calculations. 
NOTE: Plutonium-bearing colloids as a function of ionic strength for corrosion tests on glass samples SRL 202A 
and SRL 131A at SA/V or 2,000 and 20,000/m (at 90C) (Figure 8 in CRWMS M&0 2000a). 
Figure 7. Plutonium-Bearing Colloids as a Function of Ionic Strength in Corrosion Tests 
0E+0 
2E-8 
4E-8 
6E-8
0.000 0.020 0.040 0.060 0.080 0.100 
Relative Ionic Strength (mol/kg) 
202A_2K 
202A_20K 
131A_2K 
131A_20K 
[Pucolloids] (M)

ANL-WIS-MD-000012 REV00 38 April 2000 
NOTE: Here, concentrations of colloids are compared on the basis of alkali and alkaline-earth element 
concentration for colloid size > 100 nm. (Figure 1 from CRWMS M&0 2000b) 
Figure 8. Colloid Concentrations Versus Alkali and Alkaline-Earth Concentration for Groundwaters 
from Around the World 
In Chapter 4 of Total System Performance AssessmentViability Assessment (TSPA-VA) 
Analyses Technical Basis Document (CRWMS M&O 1998), data on colloid concentration and 
ionic strength for these groundwaters were plotted in an attempt to provide a predictive tool for 
estimation of colloid concentration, given the ionic strength. The plot illustrating the 
relationship is presented in Figure 9. Note that the uncertainty indicated by the error bars is quite 
high, approximately several orders of magnitude in both directions about the data points. 
Further, the shape of the curve is dominated by one groundwater with an ionic strength of 
approximately 0.07m. What the plot shows is that at higher ionic strengths there is a tendency 
toward lower colloid concentrations. 
Colloid concentration is difficult to correlate precisely with ionic strength as there are other 
factors that can significantly affect concentration, such as pH. As described above, for example, 
at pH 8.5, hematite colloids would have a zero net surface charge (since their ZPC is 
approximately 8.5) resulting in elimination of mutual repulsive forces and the tendency to 
flocculate, even if the ionic strength were very low.

ANL-WIS-MD-000012 REV00 39 April 2000 
DTN: MO0003SPAHLO12.004 
NOTE: Colloid concentration plotted as a function of ionic strength for the same groundwaters as those in 
Figure 8. Error bars denote uncertainty in concentration values. Note that this plot may illustrate 
that concentrations may be lower at higher ionic strengths. Ionic strength units: m (mol/kg). 
Figure 9. Colloid Concentration Versus Ionic Strength for Groundwaters from Around the World 
Further, site-specific studies of groundwater colloids (e.g., Kingston and Whitbeck 1991) suggest 
that when groundwater conditions favor colloid stability, there may be a very wide range of 
colloid concentrations observed for a narrow range of groundwater chemistry. 
6.1.3 Radionuclide Attachment to Colloids 
A brief description of radionuclide attachment to colloids is presented in this section. A more 
comprehensive treatment may be found in Attachment XVI. 
Non-radioactive colloids are only important to repository performance insofar as radionuclides 
attach to them. Attachment as used here is a general term that implies no particular mechanism. 
Attachment may occur by a number of mechanisms and may be reversible or irreversible in the 
time frame of transport. Some of the mechanisms of attachment include coprecipitation (e.g., 
radionuclide phases entrained in smectite colloids), surface sorption (e.g., Pu sorbed onto 
smectite or iron-(hydr)oxide colloids), and ion exchange. The rate of attachment, detachment, 
and transport determine the effective reversibility of the attachment. When the attachment is fast 
and reversible, transport on colloids will be controlled by local equilibrium, and simple models, 
e.g., linear isotherm models, using effective distribution coefficients reasonably describe the 
radionuclide behavior. When attachment is irreversible or detachment is slow relative to 
transport times, then local equilibrium models may fail, resulting in over- or under-predictions of 
transport.

ANL-WIS-MD-000012 REV00 40 April 2000 
Reversible Attachment-Perhaps the most common approach used in assessment of 
contaminant-rock interactions in the subsurface is the linear isotherm, or Kd, approach, based on 
results of batch sorption experiments. The linear isotherm model relationship is defined as 
follows (Langmuir 1997a): 
S = Kd C (Eq. 2) 
where, 
Kd distribution coefficient (mass-based) 
S mass of a solute adsorbed on a unit mass of solid 
C concentration of the adsorbing solute 
The amount of solute adsorbed on a solid can also be defined on the basis of area as follows: 
S = Ka C 
where, 
Ka distribution coefficient (area-based) 
S amount of a solute adsorbed on a unit area of solid 
C concentration of the adsorbing solute 
For this relationship to be applicable, i.e., for the assumption of constant Kd to be valid, three 
critical assumptions must be met. First, the water-contaminant-rock system must be in 
thermodynamic equilibrium. In other words, 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 interaction of some radionuclides, particularly some actinides, with 
mineral surfaces, critical assumptions of the linear isotherm model are not met, and special 
treatment must be made. Those approaches are described below. 
Reversible radionuclide uptake chiefly occurs by ion exchange in clay minerals and adsorption in 
most other minerals. Because of the assumption of reversibility, coprecipitation of radionuclides 
along with other non-radionuclide ions is excluded. A large amount 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 strontium (Sr), cesium (Cs), thorium (Th), uranium (U), protactinium (Pa), neptunium 
(Np), plutonium (Pu), and americium (Am) on colloids is described, along with caveats. 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 embedded in uranium-bearing 
colloids. 
To a large extent, the effectiveness of colloids at facilitating contaminant transport is due to their 
very large mobile surface area available for sorption. Depending on the size distribution of

ANL-WIS-MD-000012 REV00 41 April 2000 
colloids in the dispersion, the impact of choosing a mass-based Kd, or a surface-area-based Ka, 
may be significant. The greatest deviation exists in situations in which an inordinately large 
number of very small colloids exist, which have a high surface-area-to-mass ratio. Based on 
experimental measurements and observations of colloids in Yucca Mountain Project (YMP) 
groundwaters, this situation does not exist at YMP, and the use of a mass-based Kd is 
satisfactory. That is fortuitous, since nearly all available sorption information is mass-based, and 
conversion to surface-area-based information is not always straightforward. 
To develop colloid Kd information, information provided by an AMR (CRWMS M&O 2000b) 
and published literature were considered. Several significant compendia of sorption data have 
been assembled in the past decade. One useful example is the compendium developed by the 
National Cooperative for the Disposal of Radioactive Waste (NAGRA, Switzerland; Stenhouse, 
1995). The objective of that work was 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, all with reducing groundwaters (or porewater, in the case of 
bentonite). Ionic strengths are similar to J-13 water or greater. Despite differences in those 
conditions, the information compiled in the NAGRA document is useful for YMP, in that it 
includes consideration of phenomena affecting sorption. 
A second compendium that is particularly useful is a recent one developed for the U.S. 
Environmental Protection Agency (EPA 1999). In that work, information on radionuclides 
including Sr, Cs, Th, U, and Pu is compiled (no information for Pa, Np, or Am). 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 YMP. Unfortunately, the group of radionuclides covered does not provide an 
analog element for trivalent or pentavalent elements, such as Am(III), Pa(V), or Np(V). 
Several sources of YMP-specific sorption data exist from LANL research. Some specific 
experiments were conducted to evaluate sorption of Pu(IV) and Pu(V) on colloidal dispersions of 
hematite, goethite, montmorillonite, and silica. Information on Am(III) was also collected 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. In Attachment XVI, Table XVI-5, available sorption data collected at 1 day and longer 
periods is compiled. To reduce these data, sorption measurements from tests conducted at 1, 2, 
and 4 days, or 1, 2, 4, and 10 days were averaged. The rationale is that sorption is typically quite 
fast, requiring less than one day to approach an asymptote. That hypothesis is supported by the 
LANL data, in that after one day, oscillation in values were observed, but Kd values did not 
increase appreciably. The sorption averages were used in a second averaging operation, by 
including the results of the desorption experiments. In general, desorption and the averaged 
sorption Kds are very similar or within about a factor of 3 or 4, except in a few cases. It is clear, 
however, that desorption is significantly slower. A third averaging was conducted to consider 
the results of experiments conducted in synthetic J13 water and actual J13 water. The reason for 
the different results in the two waters is not clear. For plutonium, since both Pu(IV) and Pu(V) 
may be present in Yucca Mountain, results from the goethite experiments were averaged. 
Goethite was selected rather than the more sorptive hematite, because goethite, as a less 
crystalline material, should more closely resemble corrosion-product colloids produced from 
waste degradation. The developed values are listed in Attachment XVI, Table XVI-6.

ANL-WIS-MD-000012 REV00 42 April 2000 
A second source of sorption data is compiled in the AMR focusing on transport properties in the 
unsaturated and saturated zones (CRWMS M&O 2000b), which also considers previously done 
Yucca-Mountain reviews by Meijer (1992) and others. Values developed for devitrified tuff and 
iron-(hydr)oxides in the saturated zone are listed in Attachment XVI, Table XVI-6. These 
ranges are generally comparably to values extracted from the NAGRA report (Table XVI-3). 
For sorption on colloids, it is likely to be conservative to use the large Kd values in the ranges. 
Table 4 summarizes recommended values, which have been extracted from the LANL data 
discussed above. Values for Sr, Cs, Pa, U, and Np were taken from the AMR discussing 
transport properties in the saturated and unsaturated zones (CRWMS M&O 2000b). The 
maximum Kd values from the ranges for devitrified tuff and iron-(hydr)oxides were used as an 
analog for colloids consisting of smectite and corrosion-product colloids, respectively. Values 
for Am were also used for Th, because of similar tendencies for sorption, not as an oxidation 
species analog. For Pu and Am, values were developed from Stenhouse (1995). Given the nonexact 
nature of developing these Kd values, a plus-or-minus one-order-of-magnitude uncertainty 
band can be assigned to each value.

ANL-WIS-MD-000012 REV00 43 April 2000 
Table 4. Sorption Data for Colloid Transport 
Radionuclide Sorbate 
and Oxidation State(s) 
at YMP 
Recommended Kd (mL/g) 
Iron-(hydr)oxide Colloids 
Smectite Colloids 
Source 
Sr(II) 200 
30 
Maximum values developed in CRWMS M&O 
(2000b, Table 16), for devitrified tuff and iron- 
(hydr)oxides, respectively, for saturated zone units. 
Cs(I) 1000 
500 
As above 
Th(IV) (use same values as developed 
for Am) 
As above 
Pa(V) 100 
1000 
As above 
U(VI) 5
1000 
As above 
Np(V) 10 
100 
As above 
Pu(IV, V, VI) 70,000 a 
30,000 a 
Average of results of each data set (~1 day tests) 
for Pu(IV) and Pu(V) sorption on goethite and Camontmorillonite 
(DTN: LA0003NL831352.002) 
Am(III) 600,000 a 
40,000 b 
Average of results of each data set (~1 day tests) 
for Am sorption on hematite and Camontmorillonite 
(DTN: LA0003NL831352.003) 
NOTES: a A value of 10,000 mL/g was provided to PA, based on preliminary data 
b A value of 100,000 mL/g was provided to PA, based on preliminary data 
Irreversible Attachment-The assumptions required for the linear sorption model described 
above are not met in HLW-glass degradation experiments, which show that plutonium is 
irreversibly attached to smectite colloids generated during the experiments (CRWMS M&O 
2000a). Second, evidence on plutonium sorption experiments (Stenhouse 1995) with colloidal 
hematite and goethite show that the rate of desorption (backward rate) of Pu is significantly 
slower than the rate of sorption (forward rate). More importantly, over a significant time period 
(up to 150 days in some experiments), the extents of desorption is somewhat less than the extent 
of sorption. Special considerations must be made for these situations. 
For the HLW-glass experiments, the conceptual model and abstraction developed below 
accounts for smectite colloids with engulfed plutonium and other radionuclides by treating 
them as a separate colloid subtype and assuming that the engulfed radionuclides are an intrinsic 
part of the colloid, not in equilibrium with the aqueous system. For the non-ideal sorption 
behavior of plutonium, a larger Kd was selected based on the desorption experiments, pending 
further results from LANL. 
6.1.4 Physical Filtration of Colloids 
Colloid filtration as discussed here refers to the physical removal of colloids from a flow system 
by pore clogging, sieving, and straining. Filtration of colloids generally means the retention of

ANL-WIS-MD-000012 REV00 44 April 2000 
colloids moving with the suspending fluid in pores, channels, and fracture apertures that are too 
small or dry to allow passage of the colloids. Two types of physical filtration are recognized, 
conventional straining and film straining. Colloids are filtered by conventional straining if they 
are larger than a pore throat diameter or fracture aperture. Where water saturation is low, 
colloids may be filtered by film-straining if their size is greater than the thickness of the adsorbed 
water film coating the grains of the rock. The rate of colloid transport through thin water films 
depends upon the colloid size relative to the film thickness. 
Filtration of colloids could conceivably occur within the waste package. Colloids may form 
within the waste form (e.g., corroded waste fuel pellets) and at its outer surfaces (e.g., corroded 
HLW glass). They could be filtered within fractures in fuel pellets or trapped at the boundaries 
of disaggregating grains. Colloids forming within fuel rods with breached cladding could be 
filtered at perforations in the cladding; colloids formed and spalled from the HLW glass could be 
filtered at perforations in the stainless steel HLW container. Colloids reaching the interior of the 
waste package (after escaping from fuel rod cladding and HLW containers) could be filtered at 
perforations in the skin of the waste package. 
Existing colloid filtration models are empirically derived and to a considerable extent are 
specific to the system experimentally characterized. There have been no comprehensive studies 
of colloid filtration within spent fuel and HLW glass waste package environments. Therefore, 
meaningful analysis of colloid filtration within the waste package is currently not feasible. 
The conservative assumption is therefore made that all colloids formed within the waste package 
(the calculated colloid source term) are assumed to exit the waste package and enter the drift; 
thus, consideration of colloid filtration processes is excluded from the TSPA-SR analysis. 
6.1.5 Colloid Sorption at the Air-Water Interface 
Both hydrophilic and hydrophobic colloids (defined in Section 6.1.1.3) may be sorbed 
irreversibly at the gas-water interface under partially saturated conditions (Wan and Wilson 
1994). Models of colloid transport in partially-saturated media have been developed in recent 
years, but there have been no experimental studies of transient flow in partially saturated porous 
media for model comparison and calibration. 
As summarized, the concentration of colloids sorbed at the gas-water interface is a function of 
the following conditions: 
 The interface surface area available for colloid uptake, which is a function of the 
total gas saturation 
 The affinity of colloids for the gas-water interface (hydrophobic colloids have 
higher affinities than hydrophilic colloids) 
 The electrostatic charge on the colloidless negatively charged colloids exhibit a 
stronger affinity 
 The salinity of the aqueous phase, with higher salinity promoting sorption.

ANL-WIS-MD-000012 REV00 45 April 2000 
Empirical evidence suggests that the sorption of colloids at the gas-water interface is irreversible 
and the affinity may be stronger than to the rock matrix (Wan and Wilson 1994). 
Partially saturated conditions may be described or classified by considering degrees of 
saturation. At low water saturations the surface area of the gas-water interface approximates that 
of the rock matrix. Overall, colloid migration is retarded, although colloids may still move 
through the adsorbed water films. At intermediate water saturations there is still an 
interconnected gas phase, although gas flux may be lower. The interface may act as a static 
sorbing surface, but 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, transporting sorbed colloids. However, a proportion of the 
bubbles may become trapped in the rock and will effectively immobilize sorbed colloids. 
Colloid migration rates depend more strongly on colloid size as lower saturation states are 
considered (CRWMS M&O 2000b; McGraw 1996). To examine the influence of colloid size on 
transport, McGraw (1996) investigated transport of monodisperse colloids (five different sizes, 
between 20 nm and 1900 nm) under both saturated and unsaturated conditions in a quartz sand. 
The results indicated that under saturated conditions the time required for breakthrough of each 
colloid at C/Co = 50% was the same as that of the breakthrough of a non-reactive tracer, 
indicating no relationship of colloid size to migration under saturated conditions. However, the 
times required for breakthroughs of the colloids under highly unsaturated conditions exhibited a 
strong relationship between the colloid breakthrough and the colloid size with fairly complete 
breakthrough of the 20-nm colloid and little or no breakthrough of the 1900-nm colloid. 
In another set of experiments, McGraw (1996) compared four sets of hydrophobic and 
hydrophilic colloids (functionalized latex microspheres). Her results indicated that transport of 
hydrophobic colloids depends on colloid size, water film thickness, and colloid charge density. 
In contrast, hydrophilic colloids were not affected by these variables and were rapidly 
transported through the system even under very low moisture contents. McGraw (1996) 
concluded that for hydrophobic colloids, the cumulative mass of colloids recovered relative to 
that input into the column was logarithmically dependent upon the ratio of the water film 
thickness to colloid diameter. In contrast, for hydrophilic colloids, the cumulative mass of 
colloids recovered relative to that input into the column was linearly dependent upon the ratio of 
the water film thickness to colloid diameter, similar but more pronounced than the effect with the 
non-reactive tracer. McGraws findings suggest that the unsaturated zone is not necessarily an 
effective barrier to colloid migration even for relatively large colloids, although larger colloids 
will tend to be retarded more than smaller ones. 
Although the potential effects of degree of saturation on colloid transport are varied and 
complex, on balance it appears that colloids would be at least somewhat retarded under lowsaturation 
conditions. Therefore, neglect of colloid sorption onto the air-water interface in the 
TSPA calculations is believed to be conservative. 
6.1.6 Gravitational Settling of Colloids 
Over the relatively short transport distances within the waste package, larger colloidal particles 
may experience gravitational settling, thereby inhibiting transport. However, GoldSim does not

ANL-WIS-MD-000012 REV00 46 April 2000 
model settling; because of this and the fact that there are no conceptual models and data for 
gravitational settling within the waste package, this process is not incorporated into the 
abstraction. Instead, the conservative assumption is made that gravitational settling does not 
occur, but that all colloid-associated radionuclides leave a breached waste package. 
6.1.7 Colloid Diffusion 
Colloidal particles, together with any associated actinides, that are sufficiently small may diffuse 
into intercrystalline porosity and will be physically retarded. 
Physical retardation of dissolved actinides is best evaluated using Fick's Law, in which the mass 
diffusion constant (D; also referred to as diffusivity or the free-water diffusion constant) is a 
critical parameter. At a fundamental level, diffusivity is related to the physical characteristics of 
the diffusing molecule and the molecules of the host medium and the intermolecular forces 
acting between them. For example, the diffusivity of a gas molecule in another gas (the host gas) 
at low density is inversely proportional to the square root of the reduced mass of the diffusing 
and host gas molecules and the collision diameter. Whereas a rigorous theory of solute diffusion 
in liquids is apparently not available, order of magnitude estimates may be made on the basis of 
hydrodynamical theory. With that theory, the diffusivity of a solute in a liquid is inversely 
proportional to the radius of the diffusing particles (Bird et al. 1960, p. 513). Rates of diffusion 
of colloidal particles can be estimated by scaling those experimentally determined free water 
diffusion constants for dissolved actinides to dissolved colloidal materials on the basis of size 
(Stokes-Einstein relationship) as follows: 
Dcoll = Dion 
. . .
. . .
rion 
rcoll 
where, 
Dcoll diffusion constant for a colloidal actinide of radius rcoll 
Dion diffusion constant for a dissolved actinide of radius rion 
rcoll radius of the colloidal actinide 
rion radius of the dissolved actinide 
For example, given an ionic radius and a colloidal particle radius of 1  and 10 nm, respectively, 
the free-water diffusion constant for the colloidal particle would be that of the dissolved actinide 
reduced by a factor of 100. That approach is consistent with discussions in Hiemenz (1986, p. 
81). 
6.1.8 Microbes and Organic Components 
The colloid source term from the repository may be sensitive to the presence of colloid-sized 
microbes, which could change aqueous chemical conditions, facilitate contaminant transport, 
affect the amount of colloidal material such as hematite, goethite, and silica that will be mobile, 
influence corrosion of the waste package, and effect agglomeration of colloids. The potential

ANL-WIS-MD-000012 REV00 47 April 2000 
effects of microbes in the repository on the colloidal source term concentration are described 
below. 
Changes in near field chemistry in micro-environments in and around waste packages that can 
cause localized pH and Eh conditions to vary considerably from the measured or predicted bulk 
chemistry. For example, acidophilic microorganisms have been found in alkaline mine spoils 
actively creating acidic micro-environments. These localized changes in pH affect colloid 
stability. In the case of hematite, colloids are more stable at lower pH values (Langmuir 1997a). 
Microbial influenced corrosion (MIC) can accelerate radionuclide release into the surrounding 
environment; microbial oxidation of metallic iron (Fe) can produce Fe oxide colloids and 
aggregates. Conversely, microorganisms can decrease the concentration of stable colloids by 
aggregating colloidal material that they use as a food source (electron acceptors) and can result 
in a decrease in colloid concentrations of up to 91% per ml (Hersman 1995). 
It is unlikely that a significant concentration of radionuclide-associated microbes associated with 
radionuclides will be transported within the repository system due to the limited energy sources 
present near the repository. Although microbial transport in the SZ has been demonstrated in the 
literature, it is only significant for the YMP if microbial transport occurred from the repository 
through the UZ (e.g., Harvey and Garabedian 1991). The transport of microbes within the waste 
package, EBS, UZ, and SZ is not currently modeled in TSPA. 
Considerable uncertainty surrounds the rate that microbes would affect the system around the 
repository. Thus, given our current knowledge, it is difficult to estimate their effects on the 
colloid source term, in particular the effect of microbes on colloid concentration and 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 conservative. 
6.1.9 Colloid Transport 
This section summarizes and integrates the preceding subsections of Section 6.1. 
On the basis of the experiments conducted at ANL, it is anticipated that radionuclide-bearing 
colloids (primarily containing plutonium and americium) will form in the waste package from 
the degradation of the waste form. It is further anticipated that the majority of the colloids will 
be composed of clays with entrained radionuclide-bearing phases, with lesser amounts of 
pseudocolloids with reversibly attached radionuclides. The stability of the clay colloids will 
depend on ionic strength and to a certain extent on pH. In addition there are expected to be 
dissolved fractions of plutonium, americium, and other low-solubility, strongly sorbing 
radionuclides. 
It is likely that some corrosion of the interior of the waste package and certain metal components 
will occur (although this process is not explicitly modeled in TSPA), producing primarily iron- 
(hydr)oxides as scale, large particles, and colloid-sized particles. Further, portions of the 
dissolved radionuclide fractions will likely sorb, probably reversibly, to these iron corrosion 
products; a fraction of the sorbed radionuclides would be onto the colloids, forming 
pseudocolloids.

ANL-WIS-MD-000012 REV00 48 April 2000 
The stability of the pseudo-colloids from metals corrosion will depend on such environmental 
factors as the concentration of colloids in the fluid and the fluids ionic strength and pH. In 
general, colloids tend to agglomerate and settle out due to gravity (not explicitly analyzed) when 
ionic strength exceeds a certain value or the concentration of colloids exceeds a certain value that 
depends on the characteristics of the system. Further, iron-(hydr)oxide colloids would likely 
agglomerate at a fluid pH at around 8. 
Because of the short transport distances within the waste package, it is assumed that all stable 
radionuclide-bearing colloids, including pseudocolloids, retain their radionuclides and exit the 
waste package. 
6.2 PROCESS MODELS AND DATA 
6.2.1 Colloid Formation 
The conceptual models for the formation of colloids from HLW and SNF described in Section 
6.1.1 are based primarily on testing performed at ANL over the past several years. The relatively 
long duration of the tests enabled a large amount of data to be collected which allowed useful 
and reasonable generalizations to be drawn. 
6.2.1.1 Colloids Formed from HLW 
Colloids formed from HLW in the ANL tests were predominantly smectite containing discrete 
radionuclide-bearing phases entrained within the clay. For modeling purposes it is assumed that 
colloids derived from HLW are all radionuclide-bearing smectite. In addition, the radionuclides 
are assumed to be irreversibly attached to the smectite colloids, pending forthcoming data from 
LANL on radionuclide attachment and detachment to/from waste form colloids. 
Based on ANLs experimental results (Figure 10) the following approximate relationship may be 
used to relate the Pu concentration to colloid concentration: 
1 ppm colloids = 1 x 10-8 M Pucoll 
Pucoll is the concentration of Pu associated with colloids. This relationship is specific to the ANL 
experiments, but it provides a useful guideline for estimating colloid concentration from colloidassociated 
Pu concentration as defined by ionic strength and pH (see Section 6.2.2, Colloid 
Stability and Concentration). Since Pu and Am were observed to behave similarly, the colloidassociated 
Am concentration was calculated by multiplying the ratio of Am to Pu in the 
inventory by the Pu concentration.

ANL-WIS-MD-000012 REV00 49 April 2000 
1 10 -8 
2 10 -8 
3 10 -8 
4 10 -8 
5 10 -8 
6 10 -8 
0.0 
1.0 
2.0 
3.0 
4.0 
5.0 
0 50 100 150 200 250 300 350 
Pu-colloids (filt) Total Colloids (DLS) 
[Pu 
colloid 
] (M) 
[Colloid] 
total 
(ppm) 
Time (days) 
NOTE: Concentration of Pu colloids (from filtration data, Ebert 1995) and total colloid concentration (from dynamic 
light scattering [DLS] measurements) as a function of test duration for the SRL 131A glass at 2,000/m (T = 
90C). The triangles, representing concentration of plutonium in filtered colloids, correspond to the left axis. 
The circles, represengint the total colloid mass, correspond to the right axis. Note that, with the scales used 
in the graph, the triangles and circles coincide (Figure 24 from CRWMS M&0 2000b). 
Figure 10. Concentration of Pu Colloids as a Function of HLW Corrosion Test Duration 
6.2.1.2 Colloids Formed from SNF 
Results from the ANL tests on corrosion of SNF indicate that very little colloidal material was 
detected. Colloids formed were smectites, some with apparently adsorbed Pu and uranium 
silicates. Further, for all leachate analyses but one, dissolved Pu was below the limit of 
detection; the one value obtained for dissolved Pu was 4 x 10-10 M. For modeling purposes it is 
assumed that colloid concentrations, and consequently, reversibly sorbed colloidal Pu and Am 
concentrations, are estimated based on consideration of ionic strength and pH. Further, the 
extent to which SNF corroded in GoldSim calculations contributes to the calculated I and pH 
will determine SNF contribution to the colloid source term. 
6.2.1.3 Iron-(Hydr)oxide Colloids 
As described in Section 6.1.1.2, it is assumed for modeling purposes that iron will corrode and 
form iron-(hydr)oxides. It is further assumed that the iron-(hydr)oxides will occur as immobile 
scale, large particles that will settle out, and colloid-sized particles. 
Because of the uncertainty in estimating the mass and surface area of iron-(hydr)oxide scale and 
the large particles that may be produced from corrosion of waste package components, they were

ANL-WIS-MD-000012 REV00 50 April 2000 
ignored for modeling purposes. This approach is conservative because it also ignores potential 
sorption sites on scale and large particles for dissolved radionuclides, which would tend to 
reduce the mobility of the dissolved radionuclides. 
Concentrations of colloidal iron-(hydr)oxides are estimated by assuming that the maximum mass 
of colloids will occur for a given set of aqueous chemical conditions. This appears to be 
reasonable, as there is abundant source material for corrosion colloid formation. The stability 
and concentration of the colloids are determined by the fluid conditions including ionic strength 
and pH (see Section 6.1.2). 
Dissolved (aqueous) radionuclides are assumed to sorb onto the iron-(hydr)oxide colloids (as 
well as onto waste form colloids) in amounts determined by the concentrations of both the 
colloids and dissolved radionuclides and by the respective Kds for specific radionuclides onto the 
iron-(hydr)oxide colloids (see Section 6.1.3). 
NOTE: Schematic representation (used in abstraction) of iron-(hydr)oxide colloid stability as a function of ionic 
strength and pH. At and near the ZPC colloids are unstable, even at low ionic strengths. At higher ionic 
strengths the pH range at which colloids are unstable is greater. Above ionic strength 0.05 colloids are 
assumed to be unstable for all pH (Abstracted from Liang and Morgan 1990, Figure 1). 
Figure 11. Schematic Representation of Iron-(Hydr)oxide Colloid Stability as a Function of pH and Ionic 
Strength 
6.2.2 Colloid Stability and Concentration 
The stability of smectite colloids is determined primarily by ionic strength, but also to an extent 
by pH, as described in Section 6.1.2 and shown in Figure 5 and schematically for abstraction 
purposes) in Figure 12. The stability of iron-(hydr)oxide colloids is determined by both ionic 
strength and pH as described in Section 6.1.2 and shown schematically in Figure 11. 
6 8 9 11 
zpc 
pH 
0.01 
0.05 
Ionic Strength 
iron oxyhydroxide colloid 
zone of instability 
I = -0.02 pH + 0.17 
I = 0.02 pH - 0.17

ANL-WIS-MD-000012 REV00 51 April 2000 
NOTE: Schematic representation (used in abstraction) of smectite stability as a function of pH and ionic strength, 
abstracted from data presented in Figure 5. 
Figure 12. Schematic Representation of Smectite Stability as a Function of pH and Ionic Strength 
The concentration of total colloid-associated radionuclides in the pH range 6 to 12 is calculated 
according to ionic strength as shown schematically in Figure 13. The derived relationship shown 
in Figure 13, which is based on bounding of the data shown in Figure 7, is: 
[Pu colloid] = -1.25E-6  I +7.25E-8 (Eq. 3) 
valid for ionic strengths greater than 0.01 M and less than 0.05 M. Concentration at ionic 
strengths greater than 0.05 M is assumed to be 1 x 10-11 M/kg, approximately three orders of 
magnitude lower, based on experimental observations (CRWMS M&0 2000a). At ionic 
strengths less than 0.01 M, concentration is assumed to be constant at 6 x 10-8 M, based on 
experimental observations (CRWMS M&0 2000a; see Figure 13).

ANL-WIS-MD-000012 REV00 52 April 2000 
NOTE: Schematic relationship (used in abstraction) between RN-bearing colloid concentration and ionic strength. 
The function represents the bound of the HLW corrosion experimental data (see Figure 7); the maximum 
value represented, 6 x 10-8, is the maximum concentration of colloids observed in the HLW experiments. 
Figure 13. Schematic Relationship Between Radionuclide-Bearing Colloid Concentration and Ionic 
Strength 
6.2.3 Treatment of Groundwater Colloids 
Figure 9 illustrates the effect of ionic strength on stability, in terms of number population, of 
colloids from sites around the world. To be useful for calculations, these data must be converted 
to mass or surface area per unit volume. In general, the number of colloids increases by a factor 
of 1000 as colloid diameter decreases by a factor of 10 (assuming spherical colloids). Expressed 
mathematically, population = diameter-3. In natural systems, the power law exponent may be 
between about 2 and 5. By taking the first derivative of the cumulative distribution, combining it 
with masses of individual colloids assuming spherical geometry, and then integrating the 
function between a lower and upper size limit, the total colloidal mass is obtained. Figure 9 was 
constructed using populations of colloids greater than 100 nm (CRWMS M&O 1998). In Figure 
14, colloid masses were calculated using number populations from Figure 9 extrapolated to 
include colloid sizes between 10 nm and 1 micrometer. A colloid mineral density of 2.5 g/cm3 
was assumed.

ANL-WIS-MD-000012 REV00 53 April 2000 
NOTE: Schematic relationship (used in abstraction) between groundwater colloid concentration and ionic strength. 
Colloid masses were calculated using number populations from Figure 9 extrapolated to include colloid sizes 
between 10 nm and 1 micrometer. A colloid mineral density of 2.5 g/cm3 was assumed. 
Figure 14. Schematic Relationship between Groundwater Colloid Concentration and Ionic Strength 
In Figure 14, the upper and lower bounding lines capture the highest and lowest concentrations 
of colloids shown in Figure 9. Between ionic strengths of 0.01 and 0.05 M, number and mass 
concentrations decrease as groundwater colloids are destabilized. The decrease is a bounding 
approximation that is considered to be linear in log-log space, consequently derived as follows: 
function(Scoll,gw): 10^[A  log(I) + B] 
where, 
(Eq. 4) 
(Eq. 5) 
The mineralogy of YMP groundwater colloids is assumed to be clays. Consequently, the effect 
of pH on groundwater colloid stability can be simulated using the same approach described 
above for waste-form colloids (Figures 5 and 12). Radionuclide sorption on groundwater 
colloids could be diminished due to sorption-site saturation by mono- and di-valent cations 
( ) ( ) [ ] ( ) ( ) [ ] 
( ) ( ) gw coll, thresh, h gw coll, thresh, - lo 
gw coll, thresh, hi max gw, coll, gw coll, thresh, - lo min gw, coll, 
I log I log 
I log M log I log M log 
B 
- 
- 
- 
 -  = 
i 
( ) ( ) 
( ) ( ) gw coll, thresh, hi gw coll, thresh, lo 
min gw, coll, max gw, coll, 
I log I log 
M log M log 
A 
- - -
- =

ANL-WIS-MD-000012 REV00 54 April 2000 
during transport. Conversely, radionuclide sorption could be enhanced by the presence of thin 
coatings of organic materials or iron-(hydr)oxides. Because the surface chemistry of 
groundwater colloids is not known, it is assumed that sorption is similar to clay minerals, and the 
same Kd values used for waste-form colloids are used for groundwater colloids. 
6.2.4 Radionuclide Attachment to Colloids 
According to the conceptual models, a large fraction of the radionuclide-bearing colloids will be 
smectite clay with entrained radionuclide-bearing phases; the radionuclides are effectively 
irreversibly attached. Clay and iron-(hydr)oxide pseudocolloids will be available for reversibly 
sorbing aqueous radionuclides or radionuclide-bearing colloids. 
Effective kds for combined adsorption and desorption of Pu and Am on waste-form, iron- 
(hydr)oxides, and groundwater colloids are used to determine effective reversibility of the 
radionuclides on colloids according to the relationship: 
RNadsorbed/desorbed = RNdissolved x kd,RN x Mcolloid (Eq. 6) 
6.2.5 Physical Filtration of Colloids 
Physical filtration of colloids in the waste package is not analyzed in TSPA. 
6.2.6 Colloid Sorption at the Air-Water Interface 
Colloid sorption at the air-water interface is not analyzed in TSPA. 
6.2.7 Gravitational Settling of Colloids 
Gravitational settling of colloids is not analyzed in TSPA. 
6.2.8 Colloid Diffusion 
The coefficient of diffusivity of colloids relative to dissolved solutes in water is assumed to be 
0.01. 
6.2.9 Microbes and Organic Components 
Microbes and organic components are not analyzed in TSPA. 
6.2.10 Colloid Transport 
For modeling purposes, calculated colloid-associated radionuclide concentrations all leave the 
waste package. Further, desorption is assumed to be slow, relative to the transport time out of 
the waste package. 
6.3 ABSTRACTION AND IMPLEMENTATION IN GOLDSIM 
This section describes the method devised to incorporate the various process models described 
above into an abstraction of those process models, a simplified model intended to retain the

ANL-WIS-MD-000012 REV00 55 April 2000 
important principles and features of the process models. The following narrative describes the 
abstraction; Figure 15 shows the logic and flow of the abstraction approach. 
6.3.1 Inputs to GoldSim 
The abstraction uses a set of input parameters for the TSPA code GoldSim; these parameters are 
listed and described in Table 1 in Section 4.1. Input parameters have been taken from the ANL 
colloid-related AMR (CRWMS M&0 2000a); others have been extracted from available project 
documents and the scientific literature. 
6.3.2 Calculated GoldSim Parameters Needed for Abstraction 
The abstraction requires three parameters to be calculated by GoldSim as output from in-package 
chemistry calculations: CRN,diss, I, and pH (see Table 1). 
6.3.3 Calculation of Colloid and Colloid-Associated Radionuclide Concentrations 
The concentrations of colloids and colloid-associated radionuclides were calculated on the basis 
of abstractions of the experimental observations and chemical principles described above. A 
stepwise approach was taken in the abstraction, and each step is set out below with a flow 
diagram and logic statements that explicitly describe the abstraction step.

ANL-WIS-MD-000012 REV00 56 April 2000 
6.3.3.1 Waste-Form (Smectite) Colloids 
The mass concentration of radionuclides irreversibly attached to waste form colloids, produced 
from corrosion of HLW glass, as a function of ionic strength is calculated in Step 1a (Figure 
15a). 
Input I 
IF I = Ihi-thresh,coll,wf 
THEN CRNcoll,wf.irrev = CRNcoll,wf,irrev,min 
ELSE 
IF (I > Ilo-thresh,coll,wf) AND (I < Ihi-thresh,coll,wf) 
THEN CRNcoll,wf,irrev = function(STcoll,wf,is) 
ELSE CRNcoll,wf,irrev = CRNcoll,wf,irrev,.max 
function(STcoll,wf,is): -1.25E-6  I +7.25E-8 
Figure 15a. Flow Chart and Logic Statements: Effect of Ionic Strength on the Concentration of Wasteform 
Colloids Generated during HLW Glass Degradation Based on Experiments Conducted 
at ANL (see Figure 13) 
The stability of irreversibly attached waste form colloids from HLW glass is determined in Step 
1b on the basis of the fluid ionic strength and pH (Figure 15b).

ANL-WIS-MD-000012 REV00 57 April 2000 
IF (I = Ihi-thresh,coll,wf) OR (I = function(STcoll,wf,pH)) 
THEN CRNcoll,wf,irrev = CRNcoll,wf,irrev,min 
function(STcoll,wf,pH): pH/200 
Figure 15b. Flow Chart and Logic Statements: Effect of pH and Ionic Strength on Waste-form Colloid 
Stability Based on Stability Behavior of Montmorillonite Colloids (see Figure 12) 
The mobile mass of waste form colloids from HLW glass is calculated in Step 1c (Figure 15c). 
Mcoll,wf,both = Mcoll,wf,both,max  CRNcoll,wf,irrev  CRNcoll,wf,irrev,max 
Figure 15c. Flow Chart and Logic Statements: Determination of Mobile Mass of Waste-form Colloids

ANL-WIS-MD-000012 REV00 58 April 2000 
The concentration of radionuclides reversibly sorbed on waste form colloids from HLW glass is 
calculated in Step 1d (Figure 15d), based on the mass of waste form colloids, kds describing the 
distribution of radionuclides between the fluid and montmorillonite clay, and the dissolved 
concentration of radionuclides as calculated by GoldSim. 
CRNcoll,wf,rev = CRNdiss  Kd,RN,wf  Mcoll,wf,both 
Figure 15d. Flow Chart and Logic Statements: Sorption of Radionuclides on Waste-form Colloids

ANL-WIS-MD-000012 REV00 59 April 2000 
6.3.3.2 Corrosion (Iron-[Hydr]oxide) Colloids 
The stability of iron-(hydr)oxide colloids assumed to form as a result of the corrosion of waste 
package components is determined in Step 2a on the basis of the fluid ionic strength and pH 
(Figure 15e). 
IF 
BEGIN 
(I > Ihi-thresh,coll,FeOx) 
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,max 
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 15e. Flow Chart and Logic Statements: Effect of Ionic Strength and pH on Stability of Iron- 
(Hydr)oxide Colloids (see Figure 11)

ANL-WIS-MD-000012 REV00 60 April 2000 
The concentration of radionuclides reversibly sorbed on iron-(hydr)oxide colloids is calculated in 
Step 2b (Figure 15f), based on the mass of iron-(hydr)oxide colloids, kds describing the 
distribution of radionuclides between the fluid and iron-(hydr)oxides and the dissolved 
concentration of radionuclides as calculated by GoldSim. 
CRNcoll,FeOx = CRNdiss  Kd,RN,FeOx  Mcoll,FeOx 
Figure 15f. Flow Chart and Logic Statements: Sorption of Radionuclide RN on Iron-(Hydr)oxide 
Colloids.

ANL-WIS-MD-000012 REV00 61 April 2000 
6.3.3.3 Groundwater (Smectite) Colloids 
The mobile mass of groundwater colloids is calculated on the basis of fluid ionic strength in Step 
3a (Figure 15g). 
IF I = Ihi-thresh,coll,gw 
THEN Mcoll,gw = Mcoll,gw,min 
ELSE IF (I > Ilo-thresh,coll,gw) AND (I < Ihi-thresh,coll,gw) 
THEN Mcoll,gw = function(Scoll,gw) 
ELSE Mcoll,gw = Mcoll,gw,max 
function(Scoll,gw): 10^[A  log(I) + B] 
where: 
Figure 15g. Flow Chart and Logic Statements: Effect of Ionic Strength on Mobile Mass of Groundwater 
Colloids. 
( ) ( ) [ ] ( ) ( ) [ ] 
( ) ( ) gw coll, thresh, h gw coll, thresh, - lo 
gw coll, thresh, hi max gw, coll, gw coll, thresh, - lo min gw, coll, 
I log I log 
I log M log I log M log 
B 
- 
- 
- 
 -  = 
i 
( ) ( ) 
( ) ( ) gw coll, thresh, hi gw coll, thresh, lo 
min gw, coll, max gw, coll, 
I log I log 
M log M log 
A 
- - -
- = 
(see Figure 14)

ANL-WIS-MD-000012 REV00 62 April 2000 
The stability of groundwater colloids is determined in Step 3b on the basis of the fluid ionic 
strength and pH (Figure 15h). 
IF (I = Ihi-thresh,coll,wf) OR (I = function(STcoll,wf,pH)) 
THEN CRNcoll,wf,irrev = CRNcoll,wf,irrev,min 
function(STcoll,wf,pH): pH/200 
Figure 15h. Flow Chart and Logic Statements: Effect of pH and Ionic Strength on Groundwater Colloids 
Stability Based on Stability Behavior of Montmorillonite Colloids 
The concentration of radionuclides reversibly sorbed on groundwater colloids is calculated in 
Step 3c (Figure 15i), based on the mass of groundwater colloids, kds describing the distribution 
of radionuclides between the fluid and montmorillonite, and the dissolved concentration of 
radionuclides as calculated by GoldSim. 
CRNcoll,gw = CRNdiss  Kd,RN,gw  Mcoll,gw 
Figure 15i. Flow Chart and Logic Statements: Radionuclide Sorption on Groundwater Colloids 
(see Figure 12)

ANL-WIS-MD-000012 REV00 63 April 2000 
6.3.3.4 Colloid-Associated Radionuclide Source Term in Waste Package 
The colloid-associated radionuclide source term is calculated in Step 4 by summing the 
contributions to the source term from Steps 1d, 2b, and 3c (Figure 15j). 
CRNcoll = CRNcoll,wf,irrev + CRNcol,wf,rev + CRNcoll,FeOx + CRNcoll,gw 
Mcoll = Mcoll,wf,both + Mcoll,FeOx + Mcoll,gw 
Figure 15j. Flow Chart and Logic Statements: Calculation of Colloid-associated Radionuclide Source 
Term from Summation of Calculated Colloid and Radionuclide Masses 
6.4 ASSESSMENT OF ABSTRACTION 
6.4.1 Validity of Abstraction and Comparison to Data 
Generally the model is considered valid for its intended purpose based on several considerations. 
The assumptions used are considered conservative and/or bounding, depending upon the 
assumption and the basis for the assumption. The exception is the assumption, described in 
Section 6.1.1.1, regarding the potential effects of colloids from corroded DOE-SNF on repository 
performance. This assumption is subject to update and revision, and a process for verification of 
the assumption is presented in Section 6.1.1.1. 
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. Because of the minute dimensions and 
unique properties of colloids, their study presents formidable difficulties; the body of work used 
for the abstraction developed in this AMR represents a significant fraction of the state-of-the-art 
colloid research performed to date. 
Certain processes were not incorporated into the abstraction, including filtration of colloids in the 
waste package, sorption to the air-water interfaces in unsaturated environments, and gravitational 
settling. It was considered realistic to conservative to do so, as it was assumed that all of the

ANL-WIS-MD-000012 REV00 64 April 2000 
colloids generated within a breached waste package left the waste package and entered the invert 
in the drift. It may be argued that gravitational settling would facilitate the exit of colloids from 
a breach at the bottom of the waste package; the net effect, however, is the same. 
Attachment mechanisms of radionuclides to colloids was also based on laboratory tests and field 
observations. The assumption that all radionuclides associated with colloids produced from 
corroded HLW glass were irreversibly attached is conservative and is based on numerous direct 
observations. An unknown quantity of radionuclides may be reversibly attached to the colloids, 
but measurements were not undertaken to determine this quantity, and so all were assumed to be 
irreversibly attached. The assumption that radionuclide attachment to colloids produced from 
degradation of CSNF is reversible is also based on numerous direct observations, in which no 
embedded phases of the sort observed in the HLW glass-derived colloids were detected. The 
data support the assumption; however, fewer data were obtained from the CSNF testing than 
from the HLW glass testing. 
Test runs of the abstraction using GoldSim are planned in order to help assess code verification, 
model sensitivity, and confidence in model results. The test run results may be presented in a 
subsequent revision to this AMR or other vehicle. 
6.4.2 Alternative Models 
The following sections describe a possible approach to estimating the rate of release of Pu from 
degraded waste, based on the rate of waste degradation. This modeling approach has not been 
developed to the same extent as the abstraction described in Section 6.3. It is preliminary and is 
provided here for information and as a basis for potential discussion and planning. 
6.4.2.1 Rate-of-Colloid-Generation Model 
6.4.2.1.1 Conceptual Model 
The conceptual models for the formation of colloids from HLW and SNF described in Section 
6.1.1 are based on testing performed at ANL and are relevant to the rate-of-colloid-generation 
model. Section 6.1.1 describes the observed relationships between B release and extent of HLW 
corrosion, and between Tc release and SNF corrosion. 
6.4.2.1.2 Process Model 
The process models presented here were developed by ANL investigators and are based on 
observed relationships mentioned in Section 6.4.2.1.1 above and described in Section 6.1.1. 
Colloids Formed from HLW-The normalized mass loss of Pu, NL(Pu), was compared to the 
normalized mass loss of boron, NL(B), to indicate the amount of Pu released relative to the 
amount of B released, which is assumed to represent the amount of HLW glass corroded. As 
indicated in Figure 16 the NL(Pu) is 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 conditions which inhibit stable colloids.

ANL-WIS-MD-000012 REV00 65 April 2000 
10-8 
10-7 
10-6 
10-5 
10-4 
10-3 
10-2 
10-1 
100 
0 200 400 600 800 1000 
202A/2K 
202A/20K 
131A/2K 
131A/20K 
NL(Pu)/NL(B) 
time (days) 
NOTE: Ratio of normalized mass loss of plutonium to normalized mass loss of boron, NL(Pu)/NL(B), as a function of 
test duration (in days) for static corrosion tests on the SRL 202A and SRL 131A glasses at 2,000 and 
20,000/m (T = 90C) [Ebert 1995] (Figure 27 from CRWMS M&0 2000a). 
Figure 16. Ratio of Normalized Mass Loss of Plutonium to Normalized Mass Loss of Boron, 
NL(Pu)/NL(B) 
Under low ionic strength conditions, the colloids are stable in the fluid and the rate of formation 
of Pu-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&0 2000a). 
The rate of colloid formation can be described by the equation: 
dmPu -coll 
dt 
= a Malt ( )-b (Eq. 7) 
The constants, a and b, derived for the various glasses may be a function of the SA/V, leachate 
composition (I, etc.), colloid composition, or other parameters. Table 5 summarizes the ranges 
and bounding values for a and b.

ANL-WIS-MD-000012 REV00 66 April 2000 
Table 5. Ranges and bounds for the constants, a and b, in Equation 7 
Static Tests Drip Tests 
Constant Low High Low High 
Bound 
a 9 x 10-11 3 x 10-4 2 x 10-6 5 x 10-6 10-4 
b 2.5 11 0.3 1.7 1 
At the high ionic strengths encountered in certain static tests, there was no correlation observed 
between Pu and B production rates since the quantity of colloids was very low. In the drip tests, 
ionic strength remained low because EJ-13 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. The first step may be represented by the 
following equation (CRWMS M&O 2000a): 
Malt = A NRi ( t) - NRPu ( t) ( ) 
0
t.
d t (Eq. 8) 
where Malt is the mass of altered waste form containing Pu, A is the surface area (geometric), t is 
the elapsed time, and NRi( t) is the normalized release rate of marker element i (B for glass). 
(Note the term NRPu, which is the normalized release rate of Pu; this term is considered 
negligible by the investigators over the time scale of the experiments since NRB is several orders 
of magnitude larger but could become significant over long periods.) 
Release of colloids due to spallation is given by the derived relationship: 
dmPucoll 
dt 
= .Malt (Eq. 9) 
where mPu-coll is the mass release of Pu suspended 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 2000a). An 
empirical bound for . is 10-4 d-1. 
These relationships have been incorporated into the abstraction in Section 6.4.2.1.2; t is 
represented in the abstraction by the time steps at which GoldSim performs iterative calculations. 
Colloids Formed from SNF-The normalized mass loss of Pu, NL(Pu), was compared to the 
normalized mass loss of technetium, NL(Tc), to indicate the amount of Pu released relative to the 
amount of Tc released, which is assumed to represent the amount of SNF corroded. It was noted 
that the reaction of SNF under unsaturated drip test conditions does not exhibit a correlation

ANL-WIS-MD-000012 REV00 67 April 2000 
between the rate of Pu-bearing colloid formation and the corrosion of spent fuel (see Figure 17). 
The scattered data indicating Pu release rate versus Tc release rate (which are proportional to the 
NL(Pu) and NL(Tc), respectively) were bounded by assuming conservatively that 
NL(Pu)/NL(Tc) is equal to 10-5 d-1. 
10-11 
10-10 
10-9 
10-8 
10-7 
10-6 
0.000 0.005 0.010 0.015 0.020 0.025 0.030 0.035 
ATM-103 
ATM-106 
10e-5*Cum Rel Tc 
Pu Release Rate (day -1) 
Cumulative Fractional Release of Tc 
NOTE: The fractional release rate of Pu as a function of cumulative fractional Tc release for the ATM-103 and ATM- 
106 high-drip-rate tests [SN 1644, p. 37]. The solid line is the conservative bound based upon a constant of 
10-5 d-1 (Figure 30 from CRWMS M&0 2000a). 
Figure 17. Fractional Release Rate of Pu as a Function of Cumulative Fractional Tc Release for the 
ATM-103 and ATM-106 High-drip-rate Tests 
This relationship has been incorporated into the abstraction in Section 6.4.2.1.2. 
6.4.2.1.3 Abstraction 
This section describes the method devised to incorporate the process models described above 
into an abstraction of those process models, a simplified model intended to retain the important 
principles, and features of the process models. The following narrative describes the abstraction; 
Figure 18 shows the logic and flow of the abstraction approach. 
Inputs to GoldSim-The abstraction uses a set of input parameters for the TSPA code GoldSim; 
these parameters are listed and described in Table 6 below. Input parameters have been taken 
from the ANL and LANL AMRs (CRWMS M&0 2000a; CRWMS M&0 2000b); certain others 
have been extracted from available project documents and the scientific literature. 
Calculation of Quantities of Radionuclides Mobilized at Each Time Step-The masses of 
radionuclides mobilized by degradation of the waste at each time step is calculated directly by

ANL-WIS-MD-000012 REV00 68 April 2000 
GoldSim. Alternatively, the masses of the radionuclides may be determined by the product of 
mass of waste degraded and radionuclide inventory. 
Calculation of Quantity and Rate of Production of Irreversibly Attached Radionuclides- 
After the quantities of total mobilized radionuclides are calculated, they must be distributed 
between radionuclides irreversibly attached to colloids, dissolved (aqueous) radionuclides, and 
radionuclides precipitated in immobile secondary phases. Since the experimental evidence 
indicates that most waste form colloids are radionuclides and/or radionuclide-bearing phases 
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 = RNmob - RNdiss - RNsecondaryphases (Eq. 10) 
The irreversibly attached radionuclides are in turn distributed between the clay layers on altered 
HLW and clay colloids produced from both HLW and SNF; the aqueous radionuclides are 
available for reversible sorption onto colloids. 
It was discussed in Sections 6.1.1 and 6.4.2.1.2 that boron was used as a marker to indicate the 
amount of HLW degraded at any point in time. If the cumulative release of boron, Bcum, is less 
than 1 g m-2, then the following relationship (called Rate Model 1 in Figure 18) 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. 11) 
where Bcum represents the mass of HLW degraded and a and b are empirically-derived constants 
(see Section 6.4.2.1.2).

ANL-WIS-MD-000012 REV00 69 April 2000 
Figure 18. Flow Chart Illustrating Alternative Abstraction Logic 
DETERMINATION OF 
COLLOID STABILITY 
COLLOID 
SOURCE 
TERM 
N I <= 0.01 
AND 
8 <= pH <= 9 
Y 
0.01 <= I <= 0.05 
AND 
[ I > (-0.02*pH + 0.17) 
OR 
I > (0.02*pH - 0.17) ] 
CRNcoll,FeOx = 
CRNcoll,FeOx x 10-3 
Y 
N 
CRNcoll,tot = 
CRNcoll,smec-irrev + 
CRNcoll,smec-revers + 
CRNcoll,FeOx 
6 
7 
a 
DETERMINATION OF 
SMECTITE 
COLLOID STABILITY 
5 
CRNcoll,smec-irrev = 
CRNcoll,smec-irrev 
pH >= 6 
Y 
N
I >= 
.004 + 2.714 x 10-2 ln pH 
Y 
CRNcoll,smec-irrev = 
1 x 10-11 
N CRNcoll,smec-irrev = 
CRNcoll,smec-irrev 
CRNcoll,FeOx = 
CRNcoll,FeOx x 10-3 
CALCULATION OF CRNcoll,smec-irrev 
(RN IRREVERSIBLY ATTACHED 
TO SMECTITE COLLOIDS) 
If cumul 
Bnorm = 1 
g/m2 
If cumul 
1 < Bnorm < 3 
g/m2 
If cumul 
Bnorm = 3 
g/m2 
Y
Y
Y 
N
N 
Colloid generation 
Rate Model 1 
(smectite colloids) 
Colloid generation 
Rate Model 1 or 2 
(0.5 probability) 
Colloid generation 
Rate Model 2 
(spalled smectite layer) 
CALCULATION OF RNentrained 
AND RNdissolved 
(RNcoll, irrev 
+ 
RNsmec layer, irrev) 
= 
RNmob - RNdissolved 
- RNsecondaryphases 
(RNs entrained in 
smectite 
colloids and 
smectite layer) 
RNdissolved 
RNsmec = Tc x 10-5 
HLW: 
SNF: 
1 2 
CALCULATION OF 
CRNcoll 
REVERSIBLY 
ATTACHED 
TO SMECTITE AND 
IRON OXIDE 
CALCULATION OF 
CRNFeOx 
PARTITIONED 
BETWEEN IRON 
OXIDE COLLOIDS 
AND IMMOBILE 
IRON OXIDE 
a 
CRNcoll,smec-revers = 
0.9 RNdissolved 
x kad, RN-smec 
CRNFeOx = 
0.1 RNdissolved 
x kad, RN-FeOx 
CRNcoll,FeOx = 
0.001 CRN,FeOx 
CRNimmob,FeOx = 
0.999 RNFeOx 
4 3

ANL-WIS-MD-000012 REV00 70 April 2000 
Table 6. Summary of Parameters for Prospective Alternative Abstraction 
Parameter Description Value /Range/ 
Distribution 
Source 
WFdeg Fraction of waste degraded during 
time step 
-------- Would be generated by 
GoldSim 
WFSA Surface area of waste form during time 
step 
-------- Would be generated by 
GoldSim 
Binv, Tcinv, 
Puinv, Aminv 
Masses of B, Tc, and RNs in 
unaltered waste form during time step 
-------- Would be generated by 
GoldSim 
Bmob, Tcmob, 
Pumob, Ammob 
Masses of B and RNs mobilized in 
WFdeg during time step 
-------- Would be generated by 
GoldSim 
Bcum Cumulative mass of B mobilized in 
WFdeg for all previous time steps 
-------- Would be generated by 
GoldSim 
a Empirically derived constant used in 
expression for precipitated smectite 
colloids 
10-4 (bound) CRWMS M&O (2000a) 
b Empirically derived constant used in 
expression for precipitated smectite 
colloids 
1 (bound) CRWMS M&O (2000a) 
. Empirically derived constant used in 
expression for spalled smectite 
colloids 
10-4 (bound) CRWMS M&O (2000a) 
I Ionic strength of fluid during time step -------- Would be generated by 
GoldSim 
Ithresh Ionic strength at and above which all 
colloids are unstable 
0.05 or 
(narrow) range 
CRWMS M&O (2000a) 
pH pH of fluid during time step -------- Would be generated by 
GoldSim 
Kad, Pu-smec 
Kad, Am-smec 
Kad, Pu-FeOx 
Kad, Am-FeOx 
Effective attachment-detachment Kds 
for Pu and Am on smectite and iron- 
(hydr)oxide colloids based on 
measured rates of adsorption and 
desorption 
TBD 
DTN pending 
FRACcoll,min Fraction of colloids remaining under 
conditions in which colloids 
agglomerate 
0.001 Bounding from ANL results 
(CRWMS M&O 2000a) 
FRACsmec Fraction of smectite colloids versus 
(smectite + iron-[hydr]oxide) available 
for pseudocolloids 
0.9 or range To be derived 
FRACFeOx Fraction of iron-(hydr)oxides (colloids 
plus immoble) versus (iron-[hydr]oxide 
+ smectite colloids) available for 
pseudocolloids 
0.1 or range To be derived 
FRACFeOx,coll Fraction of iron-(hydr)oxide colloids 
versus total iron-(hydr)oxides available 
for pseudocolloids 
0.001 or range To be derived 
If Bcum is greater than 3 g m-2, then Rate Model 2 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. 12) 
where . is an empirically-derived constant. 
If Bcum is between 1 and 3, then either Rate Model 1 or 2 is chosen randomly (0.5 probability).

ANL-WIS-MD-000012 REV00 71 April 2000 
In the case of SNF, the quantity of radionuclides irreversibly attached to colloids was bounded 
according to the expression: 
Mass of irreversibly attached RN = (Tcmob / Tcinv) * 10-5 (Eq. 13) 
at each time step. 
Calculation of Quantity and Rate of Production of Reversibly Attached Radionuclides- 
Quantities of radionuclides reversibly attached to colloids are calculated by taking the mass of 
dissolved (aqueous) RN resulting from the waste degradation at each time step. This approach is 
similar to that described in Section 6.3. 
Quantities of radionuclides reversibly attached to colloids are calculated by taking the mass of 
dissolved (aqueous) Pu and Am resulting from the waste degradation at each time step. 
Dissolved Pu and Am are assumed to sorb onto smectite and iron-(hydr)oxide colloids as 
determined by the effective adsorption-desorption kd (kad) according to the following 
relationships: 
RNrev,smec = 0.9 * RNdiss * Kad,RNsmec (Eq. 14) 
RNrev,FeOx = 0.1 * RNdiss * Kad,RNFeOx (Eq. 15) 
The values 0.9 and 0.1 are placeholders and are represented by a range. 
Iron-(hydr)oxides are assumed to be 0.001 colloids and 0.999 immobile corrosion products 
(placeholders) and are represented by a range. 
Determination of Stability of Smectite and Iron-(Hydr)oxide Colloids-This approach is 
identical to that described in Section 6.3.6. Figure 12 illustrates for purposes of the abstraction 
where smectite colloids are stable in I-pH space and was derived from experimental data (Figure 
5). The upper limit of ionic strength is assumed to be 0.05 for the pH range 6 to 12. Below pH 
6, stability is pH-dependent; at lower pH, smectite becomes unstable at lower ionic strength. 
Smectite stability is calculated as follows (see also Figure 13): 
1. IF pH >= 6 AND I >= 0.05 THEN RNsmec,coll = 10-11 M 
2. IF pH >= 6 AND 0.01 < I < 0.05 THEN RNsmec,coll = 4.327 x 10-11 x I-1.654 
3. IF pH >= 6 AND I <= 0.01 THEN RNsmec,coll = 6 x 10-8 M 
4. IF pH < 6 AND I >= (0.004 + 2.714 x 10-2 ln pH) THEN RNsmec,coll = 10-11 M 
5. IF pH < 6 AND I < (0.004 + 2.714 x 10-2 ln pH) THEN RNsmec,coll = 
4.327 x 10-11 x I-1.654

ANL-WIS-MD-000012 REV00 72 April 2000 
Figure 6 illustrates for purposes of the abstraction where iron-(hydr)oxide colloids are stable in IpH 
space and was derived from a plot of stability ratio versus pH from Liang and Morgan 
(1990). Iron-(hydr)oxide stability is calculated as follows (see also Figure 13): 
1. IF I >= 0.05 M, THEN RNFeOx,coll = RNFeOx,coll x 10-3 
2. IF (I <= 0.01) AND (8 <= pH <= 9), THEN RNFeOx,coll = RNFeOx,coll x 10-3 
OTHERWISE RNFeOx,coll = RNFeOx,coll 
3. IF (0.01 < I < 0.05) AND I > (-0.02 * pH + 0.17) OR I > (0.02 * pH  0.17) THEN 
RNFeOx,coll = 0.01 RNFeOx,coll 
OTHERWISE RNFeOx,coll = RNFeOx,coll 
Calculation of Colloid Source Term-The colloid source term is the sum of radionuclides 
irreversibly attached to smectite colloids, radionuclides reversibly attached to smectite colloids, 
and radionuclides reversibly attached to oxyhydroxide colloids: 
CRNcoll,sum = CRNcoll,irrev + CRNcoll,smec-revers + CRNcoll,FeOx (Eq. 16) 
Assessment-This alternative abstraction takes into account many of the results from the ANL 
waste form corrosion experimental program as well as models derived from them. It 
incorporates an analytical approach that would couple the waste form degradation calculations 
by GoldSim more closely to the process model derived from the experiments. 
While it takes advantage of the most complete information available on waste form corrosion, at 
the same time it is tailored very closely to the experimental conditions and configurations. There 
are aspects of the experiments which 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 HLW, which is 
likely due to the impact of the drips falling on the sample in the HLW drip tests (recall that no 
spallation was observed in the static tests). If the corroded contents of a failed waste canister 
assumes 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. 
Implementation of this abstraction requires parameters that, it is believed, are not available under 
the current TSPA approach, such as the surface area of the waste form at each time step, and the 
total masses and chemistries of secondary phases at each time step. In lieu of these parameters, 
assumptions would have to be made that may be difficult to defend and uncertainties 
incorporated to which the results may be overly sensitive. Should in-package chemistry 
calculations be developed along these lines, the applicability of this approach could improve.

ANL-WIS-MD-000012 REV00 73 April 2000 
6.4.3 Issues and Comments from the NRC IRSRs 
Table 7. NRC IRSR Issues 
Source Issue Resolution 
NRC (1998b, 
p. 49) 
DOE should specifically determine colloid 
contribution to actinide release from four types 
of colloid formation processes: condensation, 
detachment of secondary phases, sorption onto 
existing colloids, detachment of WF fragments. 
This abstraction addresses the lack of 
condensation (true) colloids which may be 
expected in a failed waste canister, as well as 
sorption onto existing colloids. Detachment of 
secondary phases and waste form fragments is 
addressed in the alternative abstraction, which is 
impractical at this stage of TSPA development. 
NRC (1998b, 
p. 50) 
DOE should demonstrate the validity of its 
empirical correlation for colloid stability versus 
pH and ionic strength over a wide range of 
evolving environmental conditions on the bases 
of mechanistic interpretations. 
This has been accomplished in the abstraction by 
taking existing data on smectite concentration and 
smectite and iron-(hydr)oxide stability as a 
function of ionic strength and pH, bounding the 
data, and developing simplified relationships 
useful in the abstraction. 
NRC (1998b, 
p. 50) 
The influence of corrosion products generated 
by steel components on the formation of 
colloids needs to be examined by DOE. 
This has been incorporated as an assumption. 
Quantitative estimates of corrosion products and 
colloids are not currently available. 
NRC (1998b, 
p. 53) 
DOE should develop appropriate scenarios and 
estimations of colloid transport through 
package perforations. 
In this abstraction all colloids are assumed to exit 
the failed canister by whatever perforation is 
present. 
NRC (1998b, 
p. 54) 
DOE should consider higher RN releases from 
glass due to colloids and pulse-release of 
colloidal RNs from hydrated glass. 
Detachment of secondary phases and waste form 
fragments is addressed in the alternative 
abstraction, which is impractical at this stage of 
TSPA development. 
NRC (1998b, 
p. 56) 
DOE should account for large excursions in 
glass corrosion behavior due to spallation of 
colloids. 
Detachment of secondary phases and waste form 
fragments is addressed in the alternative 
abstraction, which is impractical at this stage of 
TSPA development. 
NRC (1999b, 
p. 23) 
DOE has identified and considered likely 
processes for SNF degradation [including] 
dissolution of the irradiated UO2 matrix, 
withformation of secondary minerals and 
colloids. 
Colloid formation from degradation of SNF was 
considered by evaluating the test data from 
Argonne National Laboratory and incorporating 
the results into the abstraction. 
NRC (1999b, 
p. 25) 
DOE has identified and considered likely 
processes for degradation of HLW glass 
[including] dissolution of the primary phase 
[and] formation of secondary minerals and 
colloids. 
Colloid formation from degradation of HLW glass 
was considered by evaluating the test data from 
Argonne National Laboratory and incorporating 
the results into the abstraction. 
NRC (1999b, 
p. 46) 
resolution [of Subissue 3]...is achieved 
through examination of [among other things] 
the consistency of the information with 
evidence generated by studies in other 
countries. 
A comprehensive review of the programs of other 
countries was not performed for this AMR. 
NRC (1999b, 
p. 55) 
it has been recommended that DOE 
determine colloid contribution to actinide 
release [from SNF]in PA calculations. 
Colloid contribution to actinide release was 
considered by evaluating the test data from 
Argonne National Laboratory and incorporating 
the results in the abstraction. 
NRC (1999b, 
p. 59) 
DOE has not conducted any relevant scaledown 
test to evaluate the location, morphology, 
and characteristics of corrosion penetrations. 
Corrosion penetrations were not addressed in the 
abstraction.

ANL-WIS-MD-000012 REV00 74 April 2000 
Source Issue Resolution 
NRC (1999b, 
p. 60) 
DOE should provide confirmation 
thatradionuclides are not released [from HLW 
glass] at rates greater than SNF. Otherwise, 
DOE should consider the effects of [release 
from the hydrated layer] andradionuclide 
release from HLW glass in PA 
Release of colloid-associated radionuclides from 
HLW glass was incorporated into the abstraction, 
based on test results from Argonne National 
Laboratory. 
NRC (1999b, 
p. 62) 
It is important that the long-term radionuclide 
release rate in the corrosion models used by 
DOE include the influence of stage III behavior 
[in which reprecipitation of secondary phases 
occurs, as well as cracking and exfoliation of 
the altered surface layer]. 
Release of colloid-associated radionuclides from 
HLW glass due to such processes was 
incorporated into the abstraction, based on test 
results from Argonne National Laboratory. 
NRC (1999b, 
p. 63) 
Recent researchha[s] shown formation of 
colloids in the alteration phases[and] 
transport of 100 percent of the actinides as 
colloids to the environment. 
Transport of actinides has been treated in the 
abstraction by considering colloids with 
irreversibly attached actinides and dissolved 
actinides available for reversible attachment.

ANL-WIS-MD-000012 REV00 75 April 2000 
7. CONCLUSIONS 
The following conclusions may be drawn regarding this development of an abstraction for the 
calculation of the waste form colloid source term in TSPA: 
 A useful model has been developed for GoldSim to calculate colloid concentration as a 
function of ionic strength, assuming that colloids are stable, as well as for determining 
the stability of smectite and iron-(hydr)oxide colloids as a function of both ionic 
strength and pH. 
 The abstraction employs bounding relationships that are closely tied to the colloid 
generation and characterization experimental programs conducted at ANL and LANL. 
 The abstraction is conservative, relies on bounding estimates, and incorporates as much 
realism as is considered defensible. 
 The abstraction addresses many of the comments and concerns presented in the NRC 
IRSRs. 
 Regarding spallation of altered waste form, this observation in the experiments may be 
due in part to the vessel configuration and may or may not apply to the interior of a 
failed and corroded waste canister. The alternative abstraction does begin to address 
spallation-related colloid release by incorporating the colloid rate generation developed 
by ANL investigators. 
There are several significant sources of uncertainty attached to this abstraction. The potential 
formation of colloids from degradation of N-Reactor fuel, and its potential contribution to 
repository performance, must be investigated. At this time the data are preliminary; however, 
the program is ongoing, and more data are anticipated. For now, it is assumed in the abstraction 
that, due to the small quantity of N-reactor fuel, any colloids generated from degradation of the 
fuel will have little or no effect on repository performance. A plan for verifying this assumption 
and concomitantly reducing the associated uncertainty, is described in Section 6.1.1.1. If this 
assumption, after examination, proves untrue, use of the assumption could result in 
underestimation of the contribution of colloids derived from degradation of N-reactor fuel to 
repository performance. 
Another, but less significant, uncertainty is the assumed nature of radionuclide attachment to 
colloids produced as a result of CSNF degradation. The assumption, that radionuclides may only 
be reversibly attached to smectite colloids produced from degradation of CSNF, is based on 
observations, i.e., that no embedded radionuclide phases were seen to occur in the few clay 
colloids that were produced during the degradation testing. This is described in Section 6.4.1. If 
any of the colloids contain embedded (irreversibly attached) radionuclides, the consequences for 
waste package releases would be minimal since it is assumed that all colloid-associated 
radionuclides leave a breached waste package. However, this could result in underestimation of 
mobile coloids in the drift and unsaturated and saturated zones and, possibly, underestimation of 
their impact on repository performance.

ANL-WIS-MD-000012 REV00 76 April 2000 
The abstraction is considered valid and usable in TSPA calculations for any time after the 
temperature in the repository has decreased to below boiling after the thermal pulse. Many of 
the waste degradation tests were performed at 90oC but mostly sampled at near room 
temperature. Therefore the test results may be applied to this post-thermal period. The range of 
ionic strength and pH for which colloid masses and stability are calculated in the abstraction are 
within the ranges anticipated from in-package chemistry calculations and abstraction. 
In general, the bounding relationships employed in the abstraction incorporate uncertainty 
present in the data used. Additional uncertainty may result from GoldSim and the ways in which 
GoldSim performs calculations. For example, the choices of distributions and the method of 
sampling a particular distribution may result in uncertainties in determination of colloid 
concentrations, ionic strength, pH, and radionuclide concentrations. 
Data used directly as input were taken from Liang and Morgan (1990) and Tombacz et al. (1990) 
and are currently TBV, awaiting approval as accepted data. The data form a part of the basis for 
this abstraction, and accordingly this approval is necessary for the abstraction in its current form. 
The data are from frequently cited peer-reviewed journal articles in highly respected journals, so 
approval is anticipated. 
This document may be affected by technical product input information that requires 
confirmation. Any changes to the document that may occur as a result of completing the 
confirmation activities will be reflected in subsequent revisions. The status of the input 
information quality may be confirmed by review of the Document Input Reference System 
database.

ANL-WIS-MD-000012 REV00 77 April 2000 
8. INPUTS AND REFERENCES 
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Technology, 25, 175-185. Washington, D.C.: American Chemical Society. TIC: 245733. 
Harvey, R.W.; George, L.H.; Smith, R.L.; and LeBlanc, D.R. 1989. "Transport of Microspheres 
and Indigenous Bacteria Through a Sandy Aquifer: Results of Natural- and Forced-Gradient 
Tracer Experiments." Environmental Science and Technolology, 23, (1), 51-56. Washington, 
D.C.: American Chemical Society. TIC: 224869. 
Hersman, L. 1995. Microbial Effects on Colloidal Agglomeration. LA-12972-MS. Los Alamos, 
New Mexico: Los Alamos National Laboratory. ACC: MOL.19971210.0253. 
Hiemenz, P.C. 1986. Principles of Colloid and Surface Chemistry. 2nd edition, Revised and 
Expanded. New York, New York: Marcel Dekker. TIC: 246392. 
Kim, J.I. 1994. "Actinide Colloids in Natural Aquifer Systems." MRS Bulletin, 19, (12), 47-53. 
Pittsburgh, Pennsylvania: Materials Research Society. TIC: 246128. 
Kim, S. and Corapcioglu, M.Y. 1996. "Kinetic Approach to Modeling Mobile Bacteria- 
Facilitated Groundwater Contaminant Transport." Water Resources Research, 32, (2), 321-331. 
Washington, D.C.: American Geophysical Union. TIC: 247566.

ANL-WIS-MD-000012 REV00 79 April 2000 
Kingston, W.L. and Whitbeck, M. 1991. Characterization of Colloids Found in Various 
Groundwater Environments in Central and Southern Nevada. DOE/NV/10384-36. Las Vegas, 
Nevada: U.S. Department of Energy. ACC: NNA.19930607.0073. 
Langmuir, D. 1997a. Aqueous Environmental Geochemistry. Upper Saddle River, New Jersey: 
Prentice Hall. TIC: 237107. 
Langmuir, D. 1997b. "The Use of Laboratory Adsorption Data and Models to Predict 
Radionuclide Releases from a Geological Repository: A Brief History." Scientific Basis for 
Nuclear Waste Management XX, Symposium held December 2-6, 1996, Boston, Massachusetts. 
Gray, W.J. and Triay, I.R., eds. 465, 769-780. Pittsburgh, Pennsylvania: Materials Research 
Society. TIC: 238884. 
Liang, L. and Morgan, J.J. 1990. "Chemical Aspects of Iron Oxide Coagulation in Water: 
Laboratory Studies and Implications for Natural Systems." Aquatic Sciences, 32, (1), 32-55. 
Basel, Switzerland: Birkhauser Verlag. TIC: 246125. 
McCarthy, J.F. and Degueldre, C. 1993. "Sampling and Characterization of Colloids and 
Particles in Groundwater for Studying Their Role in Contaminant Transport." Chapter 6 of 
Environmental Particles. Buffle, J. and van Leeuwen, H.P., eds. Volume 2. Boca Raton, Florida: 
Lewis Publishers. TIC: 245905. 
McCarthy, J.F. and Zachara, J.M. 1989. "Subsurface Transport of Contaminants." Environmental 
Science & Technology, 23, (5), 496-502. Easton, Pennsylvania: American Chemical Society. 
TIC: 224876. 
McGraw, M.A. 1996. The Effect of Colloid Size, Colloid Hydrophobicity, and Volumetric Water 
Content on the Transport of Colloids Through Porous Media. Ph.D. dissertation. Berkeley, 
California: University of California. TIC: 245722. 
Meijer, A. 1992. "A Strategy for the Derivation and Use of Sorption Coefficients in Performance 
Assessment Calculations for the Yucca Mountain Site." Proceedings of the DOE/Yucca 
Mountain Site Characterization Project Radionuclide Adsorption Workshop at Los Alamos 
National Laboratory, September 11-12, 1990. Canepa, J.A., ed. LA-12325-C. Pages 9-40. Los 
Alamos, New Mexico: Los Alamos National Laboratory. ACC: NNA.19920421.0117. 
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. Oldenbourg Verlag. 
TIC: 238763. 
Stenhouse, M.J. 1995. Sorption Databases for Crystalline, Marl and Bentonite. NAGRA 
Technical Report 93-06. Wettingen, Switzerland: National Cooperative for the Disposal of 
Radioactive Waste. On Order Library Tracking Number-1527.

ANL-WIS-MD-000012 REV00 80 April 2000 
Tombacz, 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 Science. TIC: 246046. 
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. 
Wan, J. and Tokunaga, T.K. 1997. "Film Straining on Colloids in Unsaturated Porous Media: 
Conceptual Model and Experimental Testing." Environmental Science and Technology, 31, (8), 
2413-2420. Washington, D.C.: American Chemical Society. TIC: 234804. 
Wan, J. and Wilson, J.L. 1994. "Colloid Transport in Unsaturated Porous Media." Water 
Resources Research, 30, (4), 857-864. Washington, D.C.: American Geophysical Union. TIC: 
222359. 
Weiss, T.H.; Mills, A.L.; Hornberger, G.M.; and Herman, J.S. 1995. "Effect of Bacterial Cell 
Shape on Transport of Bacteria in Porous Media." Environmental Science Technolology, 29, (7), 
1737-1740. Washington, D.C.: American Chemical Society. TIC: 247490. 
Yates, M.V. and Yates, S.R. 1988. "Modeling Microbial Fate in the Subsurface Environment." 
Critical Reviews in Environmental Control, 17, (4), 307-345. Boca Raton, Florida: CRC Press. 
On Order Library Tracking Number-1530 
8.2 CODES, STANDARDS, REGULATIONS, AND PROCEDURES 
10 CFR 960. 1988. Energy: General Guidelines for the Recommendation of Sites for Nuclear 
Waste Repositories. Readily Available 
64 FR 8640. Disposal of High-Level Radioactive Wastes in a Proposed Geologic Repository at 
Yucca Mountain, Nevada. Proposed rule 10 CFR 63. Readily Available 
AP-3.10Q, Rev. 02, ICN 0. Analyses and Models. Washington D.C., Washington D.C.: Office of 
Civilian Radioactive Waste Management. ACC: MOL.20000217.0246. 
DOE (U.S. Department of Energy) 2000. Quality Assurance Requirements and Description. 
DOE/RW-0333P, Rev. 9. Washington, D.C.: U.S. Department of Energy, Office of Civilian 
Radioactive Waste Management. ACC: MOL.19991028.0012. 
NLP-2-0, Rev. 5. Determination of Importance Evaluations. Las Vegas, Nevada: CRWMS 
M&O. ACC: MOL.19981116.0120. 
NRC (U.S. Nuclear Regulatory Commission) 1998a. Issue Resolution Status Report Key 
Technical Issue: Total System Performance Assessment and Integration. Rev. 1. Washington, 
D.C.: U.S. Nuclear Regulatory Commission. ACC: MOL.19990105.0083.

ANL-WIS-MD-000012 REV00 81 April 2000 
NRC (U.S. Nuclear Regulatory Commission) 1998b. Issue Resolution Status Report Key 
Technical Issue: Container Life and Source Term. Rev. 1. Washington, D.C.: U.S. Nuclear 
Regulatory Commission. ACC: MOL.19990105.0081 
NRC (U.S. Nuclear Regulatory Commission) 1999a. Issue Resolution Status Report Key 
Technical Issue: Evolution of the Near-Field Environment. Rev. 2. Washington, D.C.: U.S. 
Nuclear Regulatory Commission. ACC: MOL.19990810.0640. 
NRC (U.S. Nuclear Regulatory Commission) 1999b. Issue Resolution Status Report Key 
Technical Issue: Container Life and Source Term. Rev. 2. Washington, D.C.: U.S. Nuclear 
Regulatory Commission. TIC: 245538. 
QAP-2-3, Rev. 10. Classification of Permanent Items. Las Vegas, Nevada: CRWMS M&O. 
ACC: MOL.19990316.0006. 
QAP-2-0, Rev. 5, ICN 1. Conduct of Activities. Las Vegas, Nevada: CRWMS M&O. ACC: 
MOL.19991109.0221. 
8.3 SOURCE DATA 
LL991109851021.095. Colloid Size and Concentration Investigations in Scientific Notebook SN 
1381. Submittal date: 01/10/2000. 
MO0003SPALOW12.001. Lowest Observed Or Expected Concentration Of Radionuclide 
Element Rn Associated With Waste-Form Colloids. Submittal date: 03/02/2000. 
MO0003SPAHIG12.002. Highest and Lowest Observed or Expected Masses of Iron- 
(hydr)Oxide Colloids Per Unit Volume or Mass of Water. Submittal date: 03/02/2000. 
MO0003SPAION02.003. Values Of Ionic Strength That Define The Stability Limits Of Iron- 
(Hydr)Oxide Colloids. Submittal date: 03/03/2000. 
MO0003SPAHLO12.004. Highest and Lowest Observed or Expected Groundwater Colloid 
Masses Per Unit Volume or Mass of Water; Values of Ionic Strength Above Which Groundwater 
Colloid Dispersions Are Unstable Concentration and Stability of Groundwater Colloids as 
Defined by Ionic Strength of Groundwater. Submittal date: 03/16/2000. 
MO0004SPAKDS42.005. Kds for Pu and Am on Waste Form, Iron (hydr)Oxide, and 
Groundwater Colloids. Submittal date: 04/10/2000. 
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. 
LA0003NL831352.003. The Kd Values of 243Am on Colloids of Hematite, Montmorillonite and 
Silica in Natural and Synthetic Groundwater. Submittal date: 03/29/2000.

ANL-WIS-MD-000012 REV00 I-1 April 2000 
ATTACHMENT I - 
YMP NO. 2.1.09.14.00 
NEA NO. 2.1.09AOCOLLOID FORMATION IN WASTE AND EBS 
I.1 YMP PRIMARY FEP DESCRIPTION 
Colloids in the waste and EBS may affect radionuclide transport. Several types of colloids may 
occur naturally and may form during the evolution of the system by a variety of mechanisms. 
This FEP is broad in scope, general in its descriptions, and groups the several types of colloids 
into a single category. Discussions of colloids for the Yucca Mountain repository are presented 
elsewhere for true colloids (YMP No. 2.1.09.15.00), colloids formed by co-precipitation of waste 
alteration phases (YMP No. 2.1.09.16.01), natural pseudo-colloids (YMP No. 2.1.09.16.00), 
pseudo-colloids formed from corrosion products (YMP No. 2.1.09.17.00), and microbial colloids 
(YMP No. 2.1.09.18.00). 
I.2 SCREENING DECISION 
Include 
I.3 SCREENING ARGUMENT 
Colloids are particles and large molecules ranging from approximately 1 nanometer to 1 
micrometer in size. Several types of colloids may form in the waste form/waste package and 
EBS environment. Radionuclides may become associated with colloids in several possible ways: 
(1) radionuclides may form true colloids through hydrolysis and polymerization; (2) dissolved 
(aqueous) radionuclides may sorb to existing colloids (e.g., clay, iron-(hydr)oxides, silica), 
forming pseudocolloids; (3) radionuclide-bearing phases may coprecipitate with waste form 
alteration phases, such as clay, forming multiphase colloids in which the radionuclide-bearing 
phases are entrained in (embedded in, or occluded by) the alteration-phase colloid; and (4) 
colloid-sized microbe fragments or large organic molecules may associate with radionuclides. 
Colloids thus formed in the waste and EBS may influence radionuclide transport. Depending on 
their size and on their surface and chemical characteristics, colloids may travel at velocities 
greater than the average groundwater velocity or may travel more slowly due to retardation 
processes. Analyses and field studies show that certain radionuclides attached to colloids may be 
transported long distances relative to the aqueous species. 
YMP No. 2.1.09.14.00 (this FEP) groups the several modes of colloid formation as a set of 
processes. Specific colloid processes relevant to the Yucca Mountain repository are discussed in 
separate FEPs: True colloids (YMP No. 2.1.09.15.00), colloids from co-precipitation of waste 
alteration phases (YMP No. 2.1.09.16.01), natural-colloid-based pseudocolloids (YMP No. 
2.1.09.16.00), pseudocolloids formed from corrosion of introduced materials in the EBS and the 
WP (YMP No. 2.1.09.17.00) and microbial colloids (YMP No. 2.1.09.18.00). Retardation is 
discussed in YMP No. 2.1.09.19.00 and filtration in YMP No. 2.1.09.20.00.

ANL-WIS-MD-000012 REV00 I-2 April 2000 
DOE-SNF and DHLW are being evaluated for the purpose of improving our understanding of 
colloid formation and stability and colloids effect on radionuclide transport. Issues recently 
resolved or under investigation include: 
 Mechanisms of waste-derived colloid formation (ANL) 
 Stability and concentration of waste-derived and intrinsic colloids (ANL, LANL) 
 Stability and concentration of pseudo-colloids (LANL) 
 Attachment/detachment rates and effective Kds of pseudo-colloids (LANL) 
 Fraction of effectively irreversibly attached colloids and pseudo-colloids (LANL). 
A model abstraction that defines the waste form colloid source term for TSPA calculations has 
been developed from the data and models that result from these investigations. 
I.4 TSPA DISPOSITION 
A waste-form colloid source term will be abstracted from the data and models produced from 
several investigations (see section 3 above, Screening Argument). 
I.5 BASIS FOR SCREENING DECISION 
Plutonium and several other radionuclides of concern have low solubilities under the 
geochemical conditions anticipated in the repository, and as dissolved metal ions, they tend to 
sorb readily to the mineral phases present in the host rock. However, there have been several 
cases at the Nevada Test Site and DOE weapons complex sites where radionuclides were 
observed to have traveled distances greater than expected or predicted from models. Upon 
investigation it was determined that the radionuclides formed or were associated with colloidal 
particles. 
As described in section 3 above, Screening Argument, radionuclides may become attached to 
colloid-sized mineral fragments, waste form alteration phases, or organically derived colloids 
and as a result may travel greater distances than would be predicted for the aqueous radionuclide. 
True, or primary, colloids (formed from hydrolysis or polymerization of dissolved radionuclides) 
and colloid particles formed by coprecipitation of phases that include a radioactive component 
may result from corrosion of SNF and HLW glass within the waste package. Radioactive 
pseudocolloids may form where dissolved or colloidal radionuclides sorb onto colloid-sized clay, 
silica, or iron-(hydr)oxides. These minerals occur naturally in the groundwater; they may also 
form from the degradation of introduced repository materials such as steel and concrete and be 
introduced into a failed waste package. They may also be formed within the waste package from 
degradation of the waste form (e.g., clays) and the waste package itself (e.g., iron-(hydr)oxides). 
Microbes and microbe fragments may occur in groundwater as colloid-sized particles to which 
radionuclides can sorb. They may affect the repository environment in several ways and are a 
potential cause of localized changes in groundwater chemistry in and around the waste package, 
oxidation of metallic iron, and aggregation of colloidal materials used as a food source (resulting 
in decreased concentration). Colloid-sized organic (e.g., humic) materials may sorb 
radionuclides as well.

ANL-WIS-MD-000012 REV00 I-3 April 2000 
The occurrence and stability of colloids, including radioactive colloids, depends on such 
environmental factors as the concentration of colloids in the fluid, the fluids ionic strength, and 
the fluids pH. In general, colloids tend to agglomerate (clump together to form larger particles) 
and settle out due to gravity when ionic strength exceeds a certain value or the concentration of 
colloids exceeds a certain value that depends on the characteristics of the system. Agitation and 
high groundwater velocities are additional (but low probability) processes that may facilitate 
colloid formation and promote stability. These are not analyzed in TSPA.

ANL-WIS-MD-000012 REV00 II-1 April 2000 
ATTACHMENT II - 
YMP NO. 2.1.09.15.00 
NEA NO. 2.1.09CFORMATION OF TRUE COLLOIDS 
IN WASTE AND EBS 
II.1 YMP PRIMARY FEP DESCRIPTION 
True colloids are colloidal-sized assemblages (between approximately 1 nanometer and 1 
micrometer in diameter) of hydrolyzed and polymerized dissolved radionuclides. They may form 
in the waste package and EBS during waste-form degradation and radionuclide transport. True 
colloids are also called radionuclide intrinsic colloids (or actinide intrinsic colloids, for those 
including actinide elements). 
II.2 SCREENING DECISION 
Exclude (Note: this FEP addresses waste forms only.) 
II.3 SCREENING ARGUMENT 
The formation of true colloids involves the release of radionuclide ions from the waste form into 
solution and the subsequent hydrolysis and polymerization of the radionuclide ions to form 
colloid-sized species. However, based on the results of investigations of SNF and DHLW, in the 
course of which no true colloids were observed, it is considered unlikely that true colloids will 
form during the corrosion of waste forms. 
DOE-SNF and DHLW are being evaluated with regard to their effect or contribution to the 
technical issues discussed in this FEP. 
II.4 TSPA DISPOSITION 
Not applicable. 
II.5 BASIS FOR SCREENING DECISION 
For the past several years tests have been conducted at ANL to characterize the modes of 
degradation of SNF and DHLW under anticipated repository groundwater and geochemical 
conditions. As a part of these tests, colloids produced as a result of degradation have been 
sampled and characterized. The results of the colloid characterization may be summarized as 
follows: (1) true (intrinsic) colloids have not been observed from corrosion tests in glass or spent 
fuel, and (2) the colloids which form from waste-form corrosion are associated with waste-form 
corrosion phases (largely clays). 
These results are consistent with the argument that true colloids are unlikely to form under 
anticipated repository conditions. The formation of true colloids requires supersaturation or 
near-supersaturation conditions with respect to the radionuclide. The test results indicate that the 
concentration of dissolved plutonium, for example, was below its concentration limit. Since (1)

ANL-WIS-MD-000012 REV00 II-2 April 2000 
the test conditions were intended to approximate anticipated repository conditions, (2) the test 
conditions were not conducive to true colloid formation, and (3) true colloids were not observed, 
it was concluded that true colloids are not likely to form as a result of waste degradation. It is, 
therefore, recommended that this FEP be excluded.

ANL-WIS-MD-000012 REV00 III-1 April 2000 
ATTACHMENT III - 
YMP NO. 2.1.09.16.00 
NEA NO. 2.1.09DFORMATION OF PSEUDO-COLLOIDS (NATURAL) 
IN WASTE AND EBS 
III.1 YMP PRIMARY FEP DESCRIPTION 
Pseudocolloids are colloidal-sized particles (between approximately 1 nanometer and 1 
micrometer in diameter), usually (but not necessarily) nonradioactive, that have radionuclides 
sorbed to them. Natural pseudocolloid particles may include mineral fragments formed from the 
host rock and present in the groundwater, microbes and microbe fragments, and humic and fulvic 
acids. This FEP addresses radionuclide-bearing colloids formed from host-rock materials and 
interactions of the waste and EBS with the host rock environment. Pseudocolloids formed from 
corrosion of the waste form and EBS are discussed in YMP No. 2.1.09.17.00. The occurrence, 
stability, and transport of bacteria and organic colloids is addressed in YMP No. 2.1.09.18.00. 
III.2 SCREENING DECISION 
Include 
III.3 SCREENING ARGUMENT 
Natural colloids (clay, silica, and iron-(hydr)oxides) may be transported in groundwater into the 
repository from the vadose zone above it or may be formed from the erosion of natural backfill 
and invert materials (e.g., crushed tuff). Pseudo-colloids 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 EBS may influence radionuclide transport. 
Analyses and field studies show that radionuclides attached to colloids may be transported long 
distances relative to the aqueous radionuclide. 
DOE-SNF and DHLW are being evaluated with regard to their effect or contribution to the 
technical issues discussed in this FEP. 
III.4 TSPA DISPOSITION 
This will be incorporated into the waste form colloid-associated radionuclide concentration 
limits. 
III.5 BASIS FOR SCREENING DECISION 
Plutonium and several other radionuclides of concern have low solubilities under the 
geochemical conditions anticipated in the repository, and as dissolved metal ions they tend to 
sorb readily to the mineral phases present in the host rock. However, there have been several 
cases at the Nevada Test Site and DOE weapons complex sites where radionuclides were 
observed to have traveled distances greater than expected or predicted from models. Upon

ANL-WIS-MD-000012 REV00 III-2 April 2000 
investigation it was determined that the radionuclides formed or were associated with colloidal 
particles. 
As described in section 3, Screening Argument, radionuclides may become attached to colloidsized 
mineral fragments (clay, silica, iron-[hydr]oxides), forming radioactive pseudocolloids, and 
as a result may travel greater distances than would be predicted for the aqueous radionuclide. 
Radionuclides tend to sorb reversibly to the mineral colloids, i.e., after they sorb, they eventually 
desorb. As observed in tests conducted at LANL, the radionuclides tend to desorb more slowly 
than they sorb. Sorption and desorption rates of plutonium and americium on several types of 
mineral colloids have been and are currently being determined. The results of these 
determinations will indicate whether desorption should be considered fast or slow within 
transport times of interest in different parts of the repository system and, therefore, whether the 
sorption should be considered effectively reversible or effectively irreversible in the TSPA 
analysis. 
Microbes and microbe fragments may occur in groundwater as colloid-sized particles to which 
radionuclides can sorb or be actively bioaccumulated across the cell membrane. They may effect 
localized changes in groundwater chemistry in and around the waste package, oxidation of 
metallic iron, and aggregation of colloidal materials used a food source (resulting in decreased 
concentration). Colloid-sized organic (e.g., humic) materials may sorb radionuclides as well. 
The occurrence and stability of pseudo-colloids, as any colloids, depend on such environmental 
factors as the concentration of colloids in the fluid, the fluids ionic strength, and in some cases 
the fluids pH. In general, colloids tend to agglomerate (clump together to form larger particles) 
and settle out due to gravity when ionic strength exceeds a certain value or the concentration of 
colloids exceeds a certain value that depends on the characteristics of the system. Agitation and 
high groundwater velocities are additional (but low probability) processes that may facilitate 
colloid formation and promote stability. These are not analyzed in TSPA.

ANL-WIS-MD-000012 REV00 IV-1 April 2000 
ATTACHMENT IV - 
YMP NO. 2.1.09.16.01 
NEA NO. 2.1.09A-COLLOID PHASES PRODUCED BY CO-PRECIPITATION 
(IN WASTE AND EBS) 
(SECONDARY TO YMP 2.1.09.16.00 [NEA 2.1.09D]) 
IV.1 YMP PRIMARY FEP DESCRIPTION 
Colloidal phases of minerals containing radionuclides are formed by co-precipitation that occurs 
at the waste form. (YMP) 
IV.2 SCREENING DECISION 
Include (Note: only production of colloid phases by co-precipitation in the WP is considered 
here) 
IV.3 SCREENING ARGUMENT 
The most important processes contributing to the formation of waste form colloids, as observed 
in investigations of the degradation of SNF and DHLW, are: (1) co-precipitation of radionuclidebearing 
phases with colloid-sized clay phases formed homogeneously in the solution contacting 
the waste, and (2) under conditions in which water drops onto the waste, spallation of colloidsized 
fragments of co-precipitated clay (primarily) and radionuclide-bearing phases from the 
alteration layers formed on the waste form surfaces. Through these mechanisms, radionuclidebearing 
phases can become entrained in alteration phases, primarily clays. 
Colloids thus formed from the waste may have a large influence on radionuclide transport 
because of the apparently stable attachment of the radionuclides. Depending on their size and on 
their surface and chemical characteristics, colloids may travel at velocities greater than the 
average groundwater velocity or may travel more slowly due to retardation processes. Analyses 
and field studies show that radionuclides attached to colloids may be transported long distances 
relative to the aqueous radionuclide. 
DOE-SNF and DHLW are being evaluated with regard to their effect or contribution to the 
technical issues discussed in this FEP. 
IV.4 TSPA DISPOSITION 
This will be incorporated into the waste form colloid-associated radionuclide concentration 
limits abstraction. 
IV.5 BASIS FOR SCREENING DECISION 
Plutonium and several other radionuclides of concern have low solubilities under the 
geochemical conditions anticipated in the repository, and as dissolved metal ions they tend to 
sorb readily to mineral phases present in the host rock. Further, there have been several cases at

ANL-WIS-MD-000012 REV00 IV-2 April 2000 
the Nevada Test Site and DOE weapons complex sites where radionuclides were observed to 
have traveled distances greater than expected or predicted from models. Upon investigation it 
was determined that the radionuclides formed or were associated with colloidal particles. 
As described in section 3, Screening Argument, radionuclides may become entrained in 
colloid-sized fragments of waste form alteration phases, clay colloids in particular, and as a 
result may travel greater distances than would be predicted for the aqueous radionuclide. 
It is considered that radionuclide-bearing phases coprecipitating with clay from alteration of the 
waste form will form colloids that may be stable under anticipated repository conditions. The 
occurrence and stability of colloids, including coprecipitated radioactive colloids, depend on 
such environmental factors as the concentration of colloids in the fluid, the fluids ionic strength, 
and the fluids pH (although clay colloid stability is relatively pH-independent). In general, 
colloids tend to agglomerate (clump together to form larger particles) and settle out due to 
gravity when ionic strength exceeds a certain value or the concentration of colloids exceeds a 
certain value that depends on the characteristics of the system. Agitation and high groundwater 
velocities are additional (but low probability) processes that may facilitate colloid formation and 
promote stability (these are not analyzed in TSPA). 
DOE-SNF and DHLW are being evaluated with regard to their effect or contribution to the 
technical issues discussed in this FEP.

ANL-WIS-MD-000012 REV00 V-1 April 2000 
ATTACHMENT V - 
YMP NO. 2.1.09.17.00 
NEA NO. 2.1.09E-FORMATION OF PSEUDO-COLLOIDS 
(DEPENDENT ON CORROSION PRODUCTS) (IN WASTE AND EBS) 
V.1 YMP PRIMARY FEP DESCRIPTION 
Pseudocolloids are colloid-sized particles (between approximately 1 nanometer and 1 
micrometer in diameter), usually (but not necessarily) nonradioactive, that have radionuclides 
sorbed to them. Possible pseudocolloid particle types formed from corrosion of materials 
introduced into the repository during construction and waste emplacement include iron- 
(hydr)oxides from corrosion of waste package and EBS components; silica from degradation of 
cementitious materials; and construction materials including diesel fuel, rockbolt resin, extruded 
polystyrene, bentonite clay, and cable lubricant. This FEP addresses pseudocolloids formed from 
corrosion of the metals in the waste package and EBS. Radionuclide-bearing colloids formed 
from natural materials are discussed in YMP No. 2.1.09.16.00. The formation, stability, and 
transport of bacteria and organic colloids are addressed in YMP No. 2.1.09.18.00. 
V.2 SCREENING DECISION 
Include 
V.3 SCREENING ARGUMENT 
Artificial materials (e.g., cement grout, carbon steel, stainless steel, aluminum, titanium-7, and 
Alloy-22) will be introduced during construction of the repository. The corrosion of these 
materials may produce significant quantities of colloids, primarily metal oxyhydroxides. 
Radionuclides will tend to sorb onto the colloids, forming pseudo-colloids as well as onto larger 
particles and scale. In addition, the degradation of waste glass and spent nuclear fuel will likely 
produce clays (with and possibly without entrained radionuclide-bearing phases), another 
substrate for pseudocolloids in the drift. 
Pseudocolloids thus formed in the waste and EBS may influence radionuclide transport. 
Analyses and field studies show that radionuclides attached to colloids may be transported long 
distances relative to the aqueous radionuclide. 
DOE-SNF and DHLW are being evaluated with regard to their effect or contribution to the 
technical issues discussed in this FEP. 
V.4 TSPA DISPOSITION 
This will be incorporated into the waste form colloid source term model abstraction.

ANL-WIS-MD-000012 REV00 V-2 April 2000 
V.5 BASIS FOR SCREENING DECISION 
Plutonium and several other radionuclides of concern have low solubilities under the 
geochemical conditions anticipated in the repository, and as dissolved metal ions they tend to 
sorb readily to the mineral phases present in the host rock. However, there have been several 
cases at the Nevada Test Site and DOE weapons complex sites where radionuclides were 
observed to have traveled distances greater than expected or predicted from models. Upon 
investigation it was determined that the radionuclides formed or were associated with colloidal 
particles. 
As described in section 3 above, Screening Argument, radionuclides may become attached to 
colloid-sized particles resulting from the corrosion of metals in the waste package and EBS, 
forming radioactive pseudocolloids, and as a result may travel greater distances than would be 
predicted for the aqueous radionuclide. 
Radionuclides tend to sorb reversibly to the metal oxyhydroxide colloids, i.e., after they sorb 
they eventually desorb if the dissolved radionuclide concentration decreases. As observed in 
tests conducted at LANL, the radionuclides tend to desorb more slowly than they sorb. Sorption 
and desorption rates of plutonium and americium on several types of mineral colloids have been 
and are being determined. The results of these determinations will indicate whether desorption 
should be considered fast or slow within transport times of interest in different parts of the 
repository system and, therefore, whether the sorption should be considered effectively 
reversible or effectively irreversible in the TSPA analysis. 
The occurrence and stability of pseudocolloids from metals corrosion depend on such 
environmental factors as the concentration of colloids in the fluid, the fluids ionic strength, and 
in some cases the fluids pH. In general, colloids tend to agglomerate (clump together to form 
larger particles) and settle out due to gravity when ionic strength exceeds a certain value or the 
concentration of colloids exceeds a certain value that depends on the characteristics of the 
system. Agitation and high groundwater velocities are additional (but low probability) processes 
that may facilitate colloid formation and promote stability. These are not analyzed in TSPA.

ANL-WIS-MD-000012 REV00 VI-1 April 2000 
ATTACHMENT VI - 
YMP NO. 2.1.09.18.00 
NEA NO. 3.2.04AAMICROBIAL COLLOID TRANSPORT 
IN THE WASTE AND EBS 
VI.1 YMP PRIMARY FEP DESCRIPTION 
This FEP addresses the formation and transport of microbial colloids in the waste and EBS. 
Pseudocolloids formed from corrosion and degradation of the metals in the waste package and 
EBS are discussed in YMP No. 2.1.09.17.00. Pseudocolloids formed from host-rock materials 
and interactions of the waste and EBS with the host rock environment are discussed in YMP No. 
2.1.09.16.00. 
VI.2 SCREENING DECISION 
Exclude 
VI.3 SCREENING ARGUMENT 
Microbes can affect the amount of mobile colloidal material such as clay, hematite, goethite, and 
silica by influencing the rate of waste package corrosion. Given the present state of knowledge, 
estimates of the effects of microbes on corrosion processes are highly uncertain. However, it is 
unlikely that sufficient quantities of microbes will be available to accelerate corrosion rates 
significantly. Also, microbial action tends to increase colloid size, which would result in 
increased gravitational settling and filtration. Therefore, exclusion of microbial effects from 
TSPA may be considered conservative. 
VI.4 TSPA DISPOSITION 
N/A 
VI.5 BASIS FOR SCREENING DECISION 
The colloid source term from the repository may be sensitive to the presence of microbes, which 
could affect the amount of colloidal material such as clay, hematite, goethite, and silica that will 
be mobile. The effects of microbes in the repository on the colloidal source term concentration 
can include: 
 Changes in near-field in micro-environments in and around waste packages that can cause 
localized pH and Eh conditions to vary considerably from the measured or predicted bulk 
chemistry. For example, acidophilic microorganisms in alkaline mine spoils may actively 
create acidic micro-environments. These localized changes in pH affect colloid stability. For 
example, hematite colloids are more stable at lower (acidic) pH values (Langmuir 1997a). 
 Microbial oxidation of metallic iron (Fe) can produce Fe oxide colloids and aggregates.

ANL-WIS-MD-000012 REV00 VI-2 April 2000 
 Microbially influenced corrosion can accelerate radionuclide release into the surrounding 
environment. 
 Microorganisms can decrease the concentration of stable colloids by aggregating colloidal 
material that they use as a food source and can result in a decrease in colloid concentrations 
of up to 91% under certain conditions (Hersman 1995). 
The transport of microbes, which are colloidal in size, through the vadose zone is not considered 
in the current unsaturated zone (UZ) or saturated zone (SZ) colloid model. Microbes could be 
included in the model if data were available on concentration, size, shape, sorption properties, 
the effect of microbial activity on mobility, biofilm formation, rock-microbe interaction, plume 
characteristics, nutrient availability, longevity, etc. Nevertheless, it is unlikely that a significant 
concentration of microbes associated with radionuclides will be transported through the UZ due 
to the availability of energy sources near the repository. Although microbial transport in the SZ 
has been demonstrated in the literature, it is only significant when microbial transport occurs 
from the repository through the UZ (Fontes et al. 1991; Harvey et al. 1989; Kim and 
Corapcioglu 1996; Weiss et al. 1995; Yates and Yates 1988). 
Considerable uncertainty surrounds the rate that microbes would affect the system around the 
repository. Thus, given our current knowledge, it is difficult to approximate a colloid source 
term from the repository and even more difficult to determine the effect of microbes on colloid 
concentration and stability. 
However, the effect of overriding interest here is the fact that microbial action will tend to 
increase the colloid size, which will increase gravitational settling and filtration. Conservatism 
thus indicates that microbial transport may be excluded from the TSPA analyses. 
DOE-SNF and DHLW were evaluated with regard to their effect or contribution to the technical 
issue(s) discussed in this FEP.

ANL-WIS-MD-000012 REV00 VII-1 April 2000 
ATTACHMENT VII - 
YMP NO. 2.1.09.19.00 
NEA NO. 3.2.04ZCOLLOID TRANSPORT AND SORPTION 
VII.1 YMP PRIMARY FEP DESCRIPTION 
Interactions between radionuclide-bearing colloids and the waste and EBS may result in 
retardation of the colloids during transport by sorption mechanisms. 
VII.2 SCREENING DECISION 
Exclude in waste form and waste package. 
VII.3 SCREENING ARGUMENT 
It is assumed in TSPA that all radionuclide-bearing colloids generated from waste form 
degradation within a failed waste package will leave the waste package and enter the drift and 
EBS, and that a relatively small fraction of dissolved radionuclides may be retarded under certain 
chemical conditions. Currently, it is planned that retardation of colloids within the EBS and drift 
will not be analyzed in TSPA, although retardation of colloids in the unsaturated zone (UZ) and 
saturated zone (SZ) may be considered in UZ and SZ transport analyses. 
DOE-SNF and DHLW were evaluated with regard to their effect or contribution to the technical 
issue(s) discussed in this FEP. There was not any unique or significant effect not already 
accounted for by CSNF. 
VII.4 TSPA DISPOSITION 
Waste form/waste package: will not be analyzed. 
VII.5 BASIS FOR SCREENING DECISION 
On the basis of experiments conducted at ANL, it is anticipated that radionuclide-bearing 
colloids (primarily containing plutonium and americium) will form in the waste package from 
the degradation of the waste form. It is further anticipated that the majority of the colloids will 
be composed of coprecipitated clays and radionuclide-bearing phases (see YMP No. 
2.1.09.16.01). In addition there are expected to be dissolved fractions of plutonium, americium, 
and other relatively insoluble and strongly sorbing radionuclides, at concentration levels below 
their solubility limits. 
It is likely that some corrosion of the interior of the waste package and certain metal components 
will occur, producing primarily iron-(hydr)oxides as scale, large particles, and colloid-sized 
particles. Further, portions of the dissolved radionuclide fractions will likely sorb, probably 
reversibly, to these iron corrosion products; a fraction of the sorbed radionuclides would be onto 
the colloids, forming pseudo-colloids.

ANL-WIS-MD-000012 REV00 VII-2 April 2000 
The stability of the pseudo-colloids from metals corrosion will depend on such environmental 
factors as the concentration of colloids in the fluid and the fluids ionic strength and pH. In 
general, colloids tend to agglomerate (clump together to form larger particles) and settle out due 
to gravity when ionic strength exceeds a certain value or the concentration of colloids exceeds a 
certain value that depends on the characteristics of the system. Further, iron-(hydr)oxide colloids 
would likely agglomerate at a fluid pH of around 8.5.

ANL-WIS-MD-000012 REV00 VIII-1 April 2000 
ATTACHMENT VIII - 
YMP NO. 2.1.09.20.00 
NEA NO. 3.2.04YCOLLOID FILTRATION 
VIII.1 YMP PRIMARY FEP DESCRIPTION 
Filtration processes may affect transport of radionuclide-bearing colloids in the waste and 
engineered barrier system (EBS). 
VIII.2 SCREENING DECISION 
Exclude 
VIII.3 SCREENING ARGUMENT 
Filtration processes may affect transport of radionuclide-bearing colloids in the waste and EBS. 
Physical colloid filtration within the waste package will not be analyzed in the TSPA, and it is 
anticipated that it will not be analyzed in the EBS. Instead the conservative assumption is made 
that all colloids produced within the waste package (the calculated colloid source term) exit the 
waste package through a perforation in the waste package and enter the drift/EBS. 
DOE-SNF and DHLW were evaluated with regard to their effect or contribution to the technical 
issue(s) discussed in this FEP. 
VIII.4 TSPA DISPOSITION 
N/A 
VIII.5 BASIS FOR SCREENING DECISION 
Colloid filtration as discussed herein refers to the physical removal of colloids from a flow 
system by pore clogging, sieving, and straining. Filtration of colloids generally means the 
retention of colloids moving with the suspending fluid in pores, channels, and fracture apertures 
that are too small or dry to allow passage of the colloids. Two types of physical filtration are 
recognized (e.g., Wan and Tokunaga 1997), conventional straining and film straining. Colloids 
are filtered by conventional straining if they are larger than a pore throat diameter or fracture 
aperture. Where water saturation is low, colloids may be filtered by film-straining if their size is 
greater than the thickness of the adsorbed water film coating the grains of the rock. The rate of 
colloid transport through thin water films depends upon the colloid size relative to the film 
thickness. 
These processes could conceivably occur within the waste package, the EBS, and the host rock. 
Colloid filtration within the waste package is not analyzed in TSPA, primarily because the 
requisite research has not been performed and because a conservative assumption can be made 
which incorporates its potential effects. Chemical and electrostatic processes affecting colloid

ANL-WIS-MD-000012 REV00 VIII-2 April 2000 
flow (e.g., sorption), included as processes of filtration in much of the literature, are treated 
separately in YMP No. 2.1.09.19.00. 
Within the waste package, colloids may form within the waste form (e.g., corroded waste fuel 
pellets) and at its outer surfaces (e.g., corroded HLW glass). They could be filtered within 
fractures in fuel pellets or trapped at the boundaries of disaggregating grains. Colloids forming 
within fuel rods whose cladding has been breached could be filtered at perforations in the 
cladding; colloids formed and spalled from the HLW glass could be filtered at perforations in the 
stainless steel HLW container. Colloids reaching the interior of the waste package (after 
escaping from fuel rod cladding and HLW containers) could be filtered at perforations in the skin 
of the waste package. 
Existing colloid filtration models are empirically derived and to a considerable extent are 
specific to the system experimentally characterized. There have been no comprehensive studies 
of colloid filtration within the spent fuel and HLW glass waste package environments. 
Therefore, meaningful analysis of colloid filtration within the waste package is currently not 
feasible. 
Therefore, a conservative assumption is made: all colloids formed within the waste package (the 
calculated colloid source term) are assumed to exit the waste package and enter the drift. This is 
the justification for excluding colloid filtration processes from the TSPA-SR analysis. 
DOE-SNF and DHLW are being evaluated with regard to their effect or contribution to the 
technical issues discussed in this FEP.

ANL-WIS-MD-000012 REV00 IX-1 April 2000 
ATTACHMENT IX - 
YMP NO. 2.1.09.21.00 
NEA NO. 3.2.08C-SUSPENSIONS OF PARTICLES LARGER THAN COLLOIDS 
IX.1 YMP PRIMARY FEP DESCRIPTION 
Groundwater flow through the waste could remove radionuclide-bearing particles by a rinse 
mechanism. Particles of radionuclide-bearing material larger than colloids could then be 
transported in water flowing through the waste and EBS by suspension. 
IX.2 SCREENING DECISION 
Exclude. 
IX.3 SCREENING ARGUMENT 
Although it is conceivable that some particles larger than colloids could get rinsed out of the 
waste package, depending on the flow-through scenario under consideration, particles leaving 
the waste package is a process considered to have low consequence. The following discussion 
focuses on the possible fate of particles downstream of the waste package. 
The particles may flow through the EBS, or may be trapped. Beyond that, suspension of 
particles in groundwater flowing downward through discrete fractures in the UZ is possible. It is 
unlikely, however, that larger-than-colloid particles will have access to a sufficient number of 
vertical or subvertical fractures whose apertures permit their passage to be transported in 
significant quantities. The relatively small quantity of particles that may make it through the UZ 
would encounter low groundwater velocities in the SZ, which would likely result in the particles 
settling and becoming immobile. 
IX.4 TSPA DISPOSITION 
Suspension of particles larger than colloids is not analyzed in the TSPA. 
IX.5 BASIS FOR SCREENING DECISION 
Suspensions of particles having sizes above the colloid range are by definition not stable. 
(Colloids by definition may form stable suspensions under certain conditions; their properties 
derive from their high surface area and surface charge relative to their mass.) Unlike stable 
suspended colloids, they are subject to gravitational settling. They do not have the requisite 
surface properties to effect mutual repulsion and stability in suspension. 
Small particles could be entrained by high groundwater flow velocities. It is conceivable that 
vertical or subvertical fractures in the UZ could, under relatively high-flow conditions, pass 
larger-than-colloid particles. It is unlikely, however, that significant quantities of radionuclidebearing 
particles would be transported to the SZ in this fashion.

ANL-WIS-MD-000012 REV00 IX-2 April 2000 
This would require numerous continuous vertical fractures through the geologic sequence 
between the floor of the repository and the SZ, which is highly unlikely. During fracture 
transport, particles would be physically removed along the length of the fracture in surface 
irregularities, dead ends, and possibly by water films of insufficient thickness to transmit them. 
In the event that particles did reach the SZ, it is unlikely that sufficiently high velocities will be 
encountered in the SZ to enable further transport. The particles would instead become immobile 
due to gravitational settling.

ANL-WIS-MD-000012 REV00 X-1 April 2000 
ATTACHMENT X - 
YMP NO. 2.1.10.01.00 
NEA NO. 2.1.10D-BIOLOGICAL ACTIVITY IN WASTE AND EBS 
X.1 YMP PRIMARY FEP DESCRIPTION 
Biological activity in the waste and EBS may affect disposal-system performance by altering 
degradation processes such as corrosion of the waste packages and waste form (including 
cladding) by affecting radionuclide transport through the formation of colloids and biofilms and 
by generating gases. 
Identified as biological activity by originator. 
X.2 SCREENING DECISION 
Exclude 
X.3 SCREENING ARGUMENT 
Microbes can affect the mobility of colloidal material as well as influence the rate of waste 
package corrosion. Given the present state of knowledge, estimates of the effects of microbes on 
corrosion processes are highly uncertain. However, it is unlikely that sufficient quantities of 
microbes will be available to generate the biological activity needed to affect colloid mobility or 
accelerate corrosion rates significantly. Exclusion of effects of biological activity in the waste 
and EBS from TSPA may be considered inconsequential. 
X.4 TSPA DISPOSITION 
Not considered in TSPA. 
X.5 BASIS FOR SCREENING DECISION 
The colloid source term from the repository may be sensitive to the presence of colloid-sized 
microbes, which could change chemistry, affect the amount of colloidal material such as clay, 
hematite, goethite, and silica that will be mobile, influence corrosion of the waste package, and 
effect agglomeration of colloids. The potential effects of biological activity from microbes in the 
repository on the colloidal source term concentration are described below. 
Changes in near field chemistry in micro-environments in and around waste packages that can 
cause localized pH and Eh conditions to vary considerably from the measured or predicted bulk 
chemistry. For example, acidophilic microorganisms in alkaline mine spoils may actively create 
acidic micro-environments. These localized changes in pH affect colloid stability. In the case of 
hematite, colloids are more stable at lower pH values (Langmuir 1997a). 
MIC can accelerate radionuclide release into the surrounding environment; microbial oxidation 
of metallic iron (Fe) can produce Fe oxide colloids and aggregates. Conversely, microorganisms

ANL-WIS-MD-000012 REV00 X-2 April 2000 
can decrease the concentration of stable colloids by aggregating colloidal material that they use 
as a food source and can result in a decrease in colloid concentrations of up to 91% per ml 
(Hersman 1995). 
The transport of microbes within the waste package and EBS is not currently modeled in TSPA. 
This is due to the lack of data on concentration, size, shape, sorption properties, and the effects 
of microbial activity on mobility. Nevertheless, it is unlikely that a significant concentration of 
microbes associated with radionuclides will be transported within the repository system due to 
the lack of energy sources near the repository. Although microbial transport in the SZ has been 
demonstrated in the literature, it is only significant when microbial transport occurs from the 
repository through the UZ (Fontes et al. 1991; Harvey et al. 1989; Kim and Corapcioglu 1996; 
Weiss et al. 1995; Yates and Yates 1988). 
Considerable uncertainty surrounds the rate that microbes would affect the system around the 
repository. Thus, given our current knowledge, it is difficult to estimate their effects on the 
colloid source term, in particular the effect of microbes on colloid concentration and stability. 
However, it is understood that microbial action will tend to increase the colloid size, which 
would increase gravitational settling and filtration. Thus, excluding the effects of biological 
activity from microbes in the colloid source term and transport analysis is conservative. 
DOE-SNF and DHLW were evaluated with regard to their effect or contribution to the technical 
issue(s) discussed in this FEP.

ANL-WIS-MD-000012 REV00 XI-1 April 2000 
ATTACHMENT XI - 
YMP NO. 2.1.13.03.00 
NEA NO. 2.1.13AMUTATION 
XI.1 YMP PRIMARY FEP DESCRIPTION 
Radiation fields could cause mutation of microorganisms, leading to unexpected chemical 
reactions and impacts. 
XI.2 SCREENING DECISION 
Exclude 
XI.3 SCREENING ARGUMENT 
Microbes can affect the mobility of colloidal material as well as influence the rate of waste 
package corrosion. Given present knowledge, estimates of the effects of microbes on corrosion 
processes are highly uncertain; the potential effects of mutated microbes are more uncertain. No 
analyses or experimental research have been performed to investigate this problem specifically. 
However, general principles of population genetics indicate that most mutations are either 
neutral or deleterious to the fitness of an organism and, in the absence of strong natural selection, 
are unlikely to produce any definite change in the phenotypes of the organisms. Thus, exclusion 
of effects of mutated microbes from TSPA is probably conservative. 
XI.4 TSPA DISPOSITION 
Not considered. 
XI.5 BASIS FOR SCREENING DECISION 
The colloid-associated radionuclide concentration from the repository may be sensitive to the 
presence of microbes, which could affect the amount of mobile colloidal material. The effects of 
microbes in the repository on the colloid-associated radionuclide concentration can include: 
(1) changes in near field chemistry in micro-environments in and around waste packages, (2) 
microbial oxidation of metallic iron (Fe) to produce Fe oxide colloids and aggregates, (3) MIC, 
accelerating radionuclide release into the surrounding environment, and (4) decrease in the 
concentration of stable colloids by aggregating colloidal material used as a food source. 
Considerable uncertainty surrounds the degree to which small, isolated populations of microbes 
affect chemical systems around the repository. In the absence of much organic material, 
microbial effects are believed to be small in comparison to changes induced in chemistry by 
global variables such as heat and radiation. 
Regarding the process of mutation: Populations of microorganisms not destroyed by heat and 
radiation in the post-closure repository could in time evolve phenotypic traits very different from 
traits of the pre-closure populations; and it is possible that these altered phenotypes could be

ANL-WIS-MD-000012 REV00 XI-2 April 2000 
associated with unusual manifestations of effects (items 14 noted above). The process of 
mutation, however, is not sufficient to drive evolution in a definite direction; most mutations are 
either neutral or deleterious with regard to an organisms phenotype. Thus, mutational processes 
alone are highly unlikely to produce the aforementioned effects. The process dominating 
directional evolution of isolated populations of prokaryotes is natural selection which acts over 
successive generations to maintain the fitness of an organism for life in its environment; but the 
manner and direction in which selective forces move an organisms phenotype is not 
predictable except under the most closely controlled conditions (artificial selection). 
In any case, if mutated microbes were to occur, it is unlikely that sufficient quantities will be 
available to accelerate corrosion rates significantly or have unpredictable effects. Also, as 
mentioned above, microbial action tends to increase colloid size, which would result in 
decreased colloid stability. Therefore, the potential effects of mutated microbes are excluded 
from TSPA.

ANL-WIS-MD-000012 REV00 XII-1 April 2000 
ATTACHMENT XII - 
YMP NO. 2.1.09.22.00 
COLLOID SORPTION AT THE AIR-WATER INTERFACE 
XII.1 YMP PRIMARY FEP DESCRIPTION 
Both hydrophilic and hydrophobic colloids may be sorbed irreversibly at the gas-water interface 
under partially saturated conditions. 
XII.2 SCREENING DECISION 
Exclude 
XII.3 SCREENING ARGUMENT 
Models of colloid transport in partially-saturated media have been developed in recent years, but 
there have been no experimental studies of transient flow in partially saturated porous media for 
model comparison and calibration. Although the potential effects of degree of saturation on 
colloid transport are varied and complex, it appears that colloids would be at least somewhat 
retarded under low-saturation conditions. Therefore, neglect of colloid sorption onto the airwater 
interface in the TSPA calculations is believed to be conservative. 
XII.4 TSPA DISPOSITION 
Not included in TSPA. 
XII.5 BASIS FOR SCREENING DECISION 
The concentration of colloids sorbed at the gas-water interface is a function of the following 
conditions: 
 The interface surface area available for colloid uptake, which is a function of the 
total gas saturation 
 The affinity of colloids for the gas-water interface (hydrophobic colloids have higher 
affinities than hydrophilic colloids) 
 The electrostatic charge on the colloidless negatively charged colloids exhibit a 
stronger affinity 
 The salinity of the aqueous phase, with higher salinity promoting sorption. 
Empirical evidence suggests that the sorption of colloids at the gas-water interface is irreversible 
and the affinity may be stronger than to the rock matrix (Wan and Wilson 1994). 
Partially saturated conditions may be described or classified by considering degrees of 
saturation. At low water saturations the surface area of the gas-water interface approximates that

ANL-WIS-MD-000012 REV00 XII-2 April 2000 
of the rock matrix. Overall, colloid migration is retarded, although colloids may still move 
through the adsorbed water films. At intermediate water saturations there is still an 
interconnected gas phase, although gas flux may be lower. The interface may act as a static 
sorbing surface, but 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, transporting sorbed colloids. However, a proportion of the 
bubbles may become trapped in the rock and will effectively immobilize sorbed colloids. 
McGraw (1996) concluded that for hydrophobic colloids, the cumulative mass of colloids 
recovered relative to the mass input into the column was logarithmically dependent upon the 
ratio of the water film thickness to colloid diameter. In contrast, for hydrophilic colloids, the 
cumulative mass of colloids recovered relative to that input into the column was linearly 
dependent upon the ratio of the water film thickness to colloid diameter, similar but more 
pronounced than the effect with the non-reactive tracer. McGraws findings suggest that the 
unsaturated zone is not necessarily an effective barrier to colloid migration even for relatively 
large colloids, although larger colloids will tend to be retarded more than smaller ones.

ANL-WIS-MD-000012 REV00 XIII-1 April 2000 
ATTACHMENT XIII - 
YMP NO. 2.1.09.23.00 
COLLOID STABILITY AND CONCENTRATION DEPENDENCE ON AQUEOUS 
CHEMISTRY 
XIII.1 YMP PRIMARY FEP DESCRIPTION 
In order for radionuclide-bearing colloids to affect repository performance, the colloidal 
dispersion must be stable for the time frame of transport and must carry significant amounts of 
radionuclides. The stability of smectite colloids is determined primarily by ionic strength but 
also to an extent by pH. The stability of iron-(hydr)oxide colloids is determined by both ionic 
strength and pH as well as the concentrations of polyvalent cations. 
XIII.2 SCREENING DECISION 
Include. 
XIII.3 SCREENING ARGUMENT 
In order for radionuclide-bearing colloids to affect repository performance, the colloidal 
dispersion must be stable for the time frame of transport and must carry significant amounts of 
radionuclides. Transport times can range from days/months/years for transport out of a failed 
waste package under a large seep to hundreds of thousands of years for retarded transport to the 
accessible environment. Thus, some relatively unstable colloids generated at the waste form 
may persist long enough to be transported out of the waste package, increasing the radionuclide 
release from the waste package, but not traveling a significant distance away from the repository. 
The more stable colloids, however, may remain suspended for years and travel a much greater 
distance. 
XIII.4 TSPA DISPOSITION 
Considered in waste form colloid source term abstraction. 
XIII.5 BASIS FOR SCREENING DECISION 
The ZPC of smectite is approximately pH 2. Since a pH this low is not anticipated in the 
repository, it is a reasonable assumption that smectite colloids will remain in a stable pH range 
much of the time under anticipated repository conditions (unless ionic strength exceeds a certain 
threshold). However, with decreasing pH, the charge density of smectite particle surfaces will 
decrease as more and more H+ sorb to the surface and offset negative charges, generally 
decreasing stability (Buck and Bates 1999; Tombacz et al. 1990). This decrease in stability 
should be taken into account. 
The HLW tests conducted at ANL resulted in measured pH ranging between approximately 9 
and 11.5 (Buck and Bates 1999), which is part of the range at which smectite colloids exhibit the 
highest surface charge and hence are most stable.

ANL-WIS-MD-000012 REV00 XIII-2 April 2000 
The ZPC of iron-(hydr)oxide colloids is around pH 8.5, and at this pH they will be unstable and 
tend to agglomerate. At higher or lower pH, however, iron-(hydr)oxide colloids may be stable, 
depending upon ionic strength. Liang and Morgan (1990) demonstrated that for a given ionic 
strength iron-(hydr)oxide stability increases as pH both increases and decreases away from the 
ZPC. In general the higher the ionic strength, the wider the pH range about the ZPC that iron- 
(hydr)oxide is unstable. For example, at an ionic strength of 0.05 M, iron-(hydr)oxide is 
unstable between approximately pH 6 and 11 
Colloids were formed in the course of certain HLW glass corrosion tests on several glass 
samples. 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 very low quantities. The threshold at which flocculation occurred 
was approximately I = 0.05m. 
Colloid concentration and stability is also dependent upon the concentrations of the major 
cations Na, K, Ca, and Mg. In an investigation of natural groundwaters, it appeared that in 
general colloids 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.

ANL-WIS-MD-000012 REV00 XIV-1 April 2000 
ATTACHMENT XIV - 
YMP NO. 2.1.09.24.00 
COLLOID DIFFUSION 
XIV.1 YMP PRIMARY FEP DESCRIPTION 
The diffusivity of a solute in a liquid is inversely proportional to the radius of the diffusing 
particles. Colloidal particles, together with any associated actinides, that are sufficiently small 
may diffuse into intercrystalline porosity and will be physically retarded. 
XIV.2 SCREENING DECISION 
Include. 
XIV.3 SCREENING ARGUMENT 
Physical retardation of dissolved actinides should be evaluated; it is best evaluated using Fick's 
Law, in which the mass diffusion constant (D; also referred to as diffusivity or the free-water 
diffusion constant) is a critical parameter. 
XIV.4 TSPA DISPOSITION 
TSPA uses a diffusion coefficient of 10-2, i.e., the free-water diffusion constant for a colloidal 
particle is that of a dissolved actinide reduced by a factor of 100. 
XIV.5 BASIS FOR SCREENING DECISION 
At a fundamental level, diffusivity is related to the physical characteristics of the diffusing 
molecule and the molecules of the host medium and the intermolecular forces acting between 
them. For example, the diffusivity of a gas molecule in another gas (the host gas) at low density 
is inversely proportional to the square root of the reduced mass of the diffusing and host gas 
molecules and the collision diameter. Whereas a rigorous theory of solute diffusion in liquids is 
apparently not available, order of magnitude estimates may be made on the basis of 
hydrodynamical theory. With that theory, the diffusivity of a solute in a liquid is inversely 
proportional to the radius of the diffusing particles (Bird et al. 1960, p. 513). Rates of diffusion 
of colloidal particles can be estimated by scaling those experimentally determined free water 
diffusion constants for dissolved actinides to dissolved colloidal materials on the basis of size 
(Stokes-Einstein relationship) as follows: 
Dcoll = Dion 
. . .
. . .
rion 
rcoll 
where: 
Dcoll diffusion constant for a colloidal actinide of radius rcoll 
Dion diffusion constant for a dissolved actinide of radius rion

ANL-WIS-MD-000012 REV00 XIV-2 April 2000 
rcoll radius of the colloidal actinide 
rion radius of the dissolved actinide 
For example, given an ionic radius and a colloidal particle radius of 1  and 10 nm, respectively, 
the free-water diffusion constant for the colloidal particle would be that of the dissolved actinide 
reduced by a factor of 100. That approach is consistent with discussions in Hiemenz (1986, p. 
81).

ANL-WIS-MD-000012 REV00 XV-1 April 2000 
ATTACHMENT XV - 
YMP NO. 2.1.09.24.00 
COLLOID GRAVITATIONAL SETTLING 
XV.1 YMP PRIMARY FEP DESCRIPTION 
Over the relatively short transport distances within the waste package, larger colloidal particles 
may experience gravitational settling, thereby inhibiting transport. 
XV.2 SCREENING DECISION 
Exclude. 
XV.3 SCREENING ARGUMENT 
The conservative assumption is made that gravitational settling of radionuclide-bearing colloids 
does not occur, but instead all of the colloids (if stable) leave a breached waste package. 
XV.4 TSPA DISPOSITION 
Not included in TSPA. 
XV.5 BASIS FOR SCREENING DECISION 
GoldSim does not model settling; because of this and the fact that there are no conceptual models 
and data for gravitational settling within the waste package, this process is not incorporated into 
the abstraction. Instead, the conservative assumption is made that gravitational settling does not 
occur, but instead all of the colloids (if stable) leave a breached waste package.

ANL-WIS-MD-000012 REV00 XVI-1 April 2000 
ATTACHMENT XVI  
RADIONUCLIDE UPTAKE MECHANISMS 
XVI.1 RADIONUCLIDE UPTAKE MECHANISMS 
Perhaps the most common approach used in assessment of contaminant-rock interactions in the 
subsurface is the linear isotherm, or Kd, approach, based on results of batch sorption 
experiments. The linear isotherm model relationship is defined as follows: 
S = Kd  C (Eq. XVI-1) 
where Kd is the distribution coefficient (mass or surface-area based), S is the mass of a solute 
sorbed on a unit mass (or surface area) of solid, and C is the concentration of the sorbing solute. 
For this relationship to be applicable (i.e., for the Kd to be constant), three critical assumptions 
must be met. First, the water-contaminant-rock system must be in thermodynamic equilibrium. 
In other words, sorption must be completely reversible. Second, contaminant uptake must scale 
linearly with contaminant concentration (i.e., Kd must be a constant value). 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 interaction of some radionuclides, particularly some actinides, 
with mineral surfaces, critical assumptions of the linear isotherm model are not met, and special 
treatment must be made. Those approaches are described in the next two sections. 
XVI.1.1 REVERSIBLE UPTAKE 
Because of the assumption of reversibility, coprecipitation of radionuclides along with other nonradionuclide 
ions is excluded. A large amount 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 strontium 
(Sr), cesium (Cs), thorium (Th), protactinium (Pa), uranium (U), neptunium (Np), plutonium 
(Pu), and americium (Am) on colloids is described, along with caveats. 
To a large extent, the effectiveness of colloids at facilitating contaminant transport is due to their 
very large mobile surface area available for sorption. Depending on the size distribution of 
colloids in the groundwater, the impact of choosing a mass-based Kd, or a surface-area-based Ka, 
may be significant. The greatest variability exists in situations in which an inordinately large 
number of very small colloids exist, which have a high surface-area-to-mass ratio. Based on 
experimental measurements and observations of colloid characteristics in Yucca Mountain 
groundwaters, this situation does not exist at Yucca Mountain, and the use of a mass-based Kd is 
satisfactory. 
For Yucca Mountain colloid transport calculations, Kd values are in units of volume of sorbate 
per mass solid sorbent (e.g., mL/g). To calculate the concentration of a radionuclide associated 
with colloids in a colloidal dispersion, the following expression is used (RN = radionuclide): 
CRNcoll = CRNdiss  Kd,RN  Mcoll  10-6 (Eq. XVI-2) 
where:

ANL-WIS-MD-000012 REV00 XVI-2 April 2000 
CRncoll = concentration of radionuclide RN associated with the colloidal dispersion (e.g., 
moles RN/L dispersion) 
CRndiss = concentration of dissolved RN (moles RN/L) 
Kd,RN = distribution coefficient (Ml/g) 
Mcoll = mass of colloids in dispersion (mg/L) 
To develop colloid Kd information, information provided by project-related data (CRWMS 
M&O 2000b) and published literature was considered. Several significant compendia of 
sorption data have been assembled in the past decade. One useful example is the compendium 
developed by the National Cooperative for the Disposal of Radioactive Waste (NAGRA, 
Switzerland; Stenhouse, 1995). The objective of that work was 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, all with reducing groundwaters (or 
porewater, in the case of bentonite). Ionic strengths are similar to J-13 well water, or greater. 
Despite differences in those conditions, the information compiled in the NAGRA document is 
useful for YMP, in that it includes consideration of phenomena affecting sorption. A second 
compendium that is particularly useful is a recent one developed for the U.S. EPA (EPA, 1999). 
In that work, information on radionuclides including Sr, Cs, Th, U, and Pu is compiled (no 
information for Pa, Np, or Am). 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 YMP. Unfortunately, 
the group of radionuclides covered does not provide an analog element for trivalent or 
pentavalent elements, such as Am(III), Pa(V), or Np(V). 
By considering the effect of aqueous chemical conditions on sorption, the selection of relevant 
Kd values is simplified. The behavior of the mineral surface is primarily controlled by pH and 
ionic strength (particularly concentrations of sodium, calcium, and potassium). The behavior of 
the sorbate is primarily controlled by its oxidation state, pH, and the partial pressure of carbon 
dioxide (Pco2). In Table XVI-1, the effect of key phenomena affecting sorption are summarized 
for each radionuclide, for clay and metal-(hydr)oxide minerals. For minerals in which the 
sorption mechanism is primarily by ion exchange (e.g., clay minerals), ionic strength impacts 
sorption of Sr and Cs on clay minerals (especially at their edges), because cations compete with 
radionuclides for exchange sites. Ionic strength is less important for actinides, in that some 
anions and cations contributing to ionic strength affect sorption by forming complexes with the 
actinides. Under the oxidizing redox condition at Yucca Mountain, radionuclides occur in 
oxidized forms, which, in general results in decreased sorption. Plutonium will disproportionate 
and be simultaneously present in multiple valence states (IV, V, and VI at YMP). The actinide 
elements discussed exhibit strong hydrolysis behavior and a tendency to form strong carbonato 
complexes. The actinides tend to be completely hydrolyzed 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 hydroxo complexes. Consequently, sorption is greatest 
in the near neutral pH range. The Pco2 of the system has minor or negligible effect on sorption 
of Sr and Cs, but a strong effect on sorption of the actinides. Carbonate is a strong complexant, 
and at alkaline pH values, the actinides may form mono-, di-, and tri-carbonato complexes,

ANL-WIS-MD-000012 REV00 XVI-3 April 2000 
depending on the actinide (and valence). The presence of these complexes tends to decrease 
sorption. 
XVI.1.1.1 NAGRA Results 
In the NAGRA study, the mineralogic compositions of three relevant rock and engineered barrier 
types were compiled, and used to weight relevant sorption values from their compendium. The 
rock or barriers selected were crystalline rock, marl, and bentonite. The chemical compositions 
of relevant groundwaters or porewaters, all at least somewhat reducing, were also compiled. For 
Yucca Mountain colloid calculations, the sorption values for crystalline rock were selected for 
two reasons. First, the mineral phase dominating sorption is similar in mineralogy to Yucca 
Mountain waste-form and groundwater colloids. A significant observation made in the NAGRA 
report, is that sorption by a rock is typically dominated by one mineral constituent, usually a 
phyllosilicate mineral. Second, the groundwater is closest to Yucca Mountain groundwater in 
terms of pH, ionic strength, and redox conditions (Table XVI-2). NAGRA proposed realistic 
values and conservative values. The conservative values for the actinides were developed by 
NAGRA 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 would be more 
relevant to Yucca Mountain. However, to compensate for the use of a mass-based Kd approach 
on colloids, the realistic values are useful. In Table XVI-3, the rock composite Kds and 
individual Kds for smectite and oxides/quartz are compiled. The latter are probably more 
suitable to consider for Yucca Mountain colloid Kds. 
XVI.1.1.2 EPA Results 
In the EPA (1999) study, published literature was reviewed to determine the predominant 
aqueous chemical conditions affecting sorption in soils for each contaminant. For Sr, soil clay 
content and groundwater pH were used as criteria to recommend Kds. For Cs, clay content was 
also used, but pH is less important because Cs exists as an aquo ion and does not complex, and it 
sorbs by ion exchange, a process that is not affected by pH-dependent surface charge. For Th, 
pH and sorbate concentration are important. At concentrations close to its solubility limit, Kds 
tend to increase. For U, pH was crucial. For Pu carbonate complexation has a strong influence 
on sorption, and so carbonate concentration, as well as clay content, was used as criteria to select 
a relevant Kd. In Table XVI-4, Kd values are compiled based on aqueous chemical conditions 
closest to those expected in ambient groundwater at Yucca Mountain. Unfortunately, the Kd 
values are expressed in a minimum to maximum format, with no indication of distribution. 
Some ranges, for example U, span almost six orders of magnitude. to narrow the range, the 
geometric means were calculated in this AMR, and reported in Table XVI-5. It would be useful 
to examine the data in the EPA (1999) report, and extract a subset that is particularly applicable 
to Yucca Mountain, and decrease the ranges of Kd values. 
XVI.1.1.3 LANL Results 
Several sources of YMP-specific sorption data exist from LANL research. Stenhouse (1995) 
provides 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

ANL-WIS-MD-000012 REV00 XVI-4 April 2000 
measured after 150 days, because desorption tends to be much slower in these mineral-sorbate 
systems. In Table XVI-5, available sorption data collected at 1 day and longer periods is 
compiled. To reduce these data, sorption measurements from tests conducted at 1, 2, and 4 days, 
or 1, 2, 4, and 10 days were averaged. The rationale is that sorption is typically quite fast, 
requiring less than one day to approach an asymptote. That hypothesis is supported by the 
LANL data, in that after one day, oscillation in values were observed, but Kd values did not 
increase appreciably. The sorption averages were used in a second averaging operation, by 
including the results of the desorption experiments. In general, desorption and the averaged 
sorption Kds are very similar or within about a factor of 3 or 4, except in a few cases. It is clear, 
however, that desorption is significantly slower. A third averaging was conducted to consider 
the results of experiments conducted in synthetic J13 water and actual J13 water. The reason for 
the different results in the two waters is not clear. For plutonium, since both Pu(IV) and Pu(V) 
may be present in Yucca Mountain, results from the goethite experiments were averaged. 
Goethite was selected rather than the more sorptive hematite, because goethite, as a less 
crystalline material, should more closely resemble corrosion-product colloids produced from 
waste degradation. The developed values are listed in Table XVI-6. 
A second source of sorption data is compiled in the AMR focusing on transport properties in the 
unsaturated and saturated zones (CRWMS M&O 2000b), which also considers previously done 
Yucca-Mountain reviews by Meijer (1992) and others. Values developed for devitrified tuff and 
iron-(hydr)oxides in the saturated zone are listed in Table XVI-6. These ranges are generally 
comparably to values extracted from the NAGRA report (Table XVI-2). For sorption on 
colloids, it is likely to be conservative to use the large Kd values in the ranges. 
XVI.1.1.4 Recommendations 
Table XVI-7 summarizes recommended values, which have been extracted from the LANL data 
discussed above. Values for Sr, Cs, Pa, U, and Np were taken from the AMR discussing 
transport properties in the saturated and unsaturated zones (CRWMS M&O 2000b). The 
maximum Kd values from the ranges for devitrified tuff and iron-(hydr)oxides were used as an 
analog for colloids consisting of smectite and corrosion-product colloids, respectively. Values 
for Am were also used for Th, because of similar tendencies for sorption, not as an oxidation 
species analog. For Pu and Am, values were developed from Stenhouse (1995). Given the nonexact 
nature of developing these Kd values, a plus-or-minus one-order-of-magnitude uncertainty 
band can be assigned to each value. 
XVI.1.2 Irreversible Uptake 
The assumptions required for the linear sorption model described above are not met in two 
instances. First, evidence from HLW-glass degradation experiments shows that plutonium is 
irreversibly attached to smectite colloids generated during the experiments (CRWMS M&O 
2000a). Second, evidence on plutonium sorption experiments with colloidal hematite, goethite, 
montmorillonite, and silica, show that the rate of desorption (backward rate) of Pu is 
significantly less than the rate of sorption (forward rate). Moreover, over a significant time 
period (up to 150 days in some experiments), the extents of desorption is significantly less than 
the extent of sorption. Special considerations must be made for these situations.

ANL-WIS-MD-000012 REV00 XVI-5 April 2000 
For the HLW-glass experiments, the conceptual model and abstraction developed below 
accounts for smectite colloids with engulfed plutonium and other radionuclides, by treating 
them as a separate colloid subtype, and by assuming that the engulfed radionuclides are an 
intrinsic part of the colloid, not in equilibrium with the aqueous system. For the non-ideal 
sorption behavior of plutonium, a larger Kd was selected based on the desorption experiments, 
pending further results from LANL.

ANL-WIS-MD-000012 REV00 XVI-6 April 2000 
Table XVI-1. Sorption Behavior of Key Radionuclides as a Function of Aqueous Chemical Conditions 
Radionuclide 
Sorbate and 
Oxidation 
State(s) at YMP 
Sorption Mechanisms Redox Effects Sorbate Speciation Ionic Strength Effects pH 
Effects 
pCO2 
Effects 
Sr(II) Ion exchange in clay 
minerals; 
Surface complexation 
on minerals with pHdependent 
surface 
charge 
Not important Sr2+; 
in carbonate systems, 
may form SrCO3
0 
Sorption by ion 
exchange decreases 
with increasing IS 
because of surface 
competition of ions such 
as Na+, Ca2+ 
Change in pH from 6 to 
9 results in slight 
increase in sorption 
Sorption increases 
with increasing 
[CO3
2-], due to 
adsorption of 
carbonato 
complexes 
Cs(I) Ion exchange in clay 
minerals 
Not important Cs+ Sorption by ion 
exchange decreases 
with increasing IS 
because of surface 
competition of ions such 
as Na+, Ca2+, K+ 
Change in pH from 6 to 
9 results in slight 
increase in sorption 
Not important; Cs2+ 
does not tend to 
complex with CO3
2- 
Th(IV) Surface complexation Not important Sparsely soluble; 
Very strong hydrolysis; 
forms strong hydroxy 
carbonate complexes; 
Th(OH)3CO3
- 
Sorption decreases with 
increasing ionic strength 
because of surface 
competition of ions such 
as Na+, Ca2+ 
Sorption decreases 
with increasing 
[CO3
2-] 
Pa(V) Not ion exchange Pa(IV) sorption 
greater than Pa(V) 
sorption 
Strong hydrolysis 
behavior; 
PaO(OH)2+, PaO(OH)2
+ 
Not important 
U(VI) U(IV) sorption 
greater than U(VI) 
sorption 
UO2(CO3)2
2-, UO2(CO3)3
4-, 
UO2CO3
0, UO2(OH)3
-, 
UO2(OH)2
0 
pH dependence of 
sorption similar to 
Am(III) 
Sorption decreases 
with increasing 
[CO3
2-] 
Np(V) Strong hydrolysis

ANL-WIS-MD-000012 REV00 XVI-7 April 2000 
Radionuclide 
Sorbate and 
Oxidation 
State(s) at YMP 
Sorption Mechanisms Redox Effects Sorbate Speciation Ionic Strength Effects pH 
Effects 
pCO2 
Effects 
Pu(IV, V, VI) At lower pH, cationic 
species sorb by cation 
exchange; at higher pH, 
anionic carbonato 
species sorb by 
exchange with surface 
hydroxyl or carbonate 
groups; neutral species 
sorb by surface 
complexation 
Sorption of reduced 
species greater than 
oxidized species 
Strong hydrolysis; 
forms strong carbonato 
complexes; Pu(IV) may 
form polynuclear 
complexes; 
Pu(OH)2(CO3)2
2-, 
Pu(OH)4
0 
pH dependence of 
sorption similar to 
Am(III) 
Am(III) at lower pH cations sorb 
by cation exchange; 
Anionic carbonato 
complexes sorb through 
exchange of carbonate 
with surface hydroxyl 
groups on oxides 
Not important Strong hydrolysis; 
forms strong carbonato 
complexes; 
Am(OH)(CO3)0, 
Am(OH)(CO3)2
2-, 
Am(CO3)+, Am(CO3)2
- 
Sorption highest at pH 
7-8 due to neutral 
complexes, e.g., 
Am(OH)3
0; 
Kd decreases above 
pH 9 because of 
anionic species 
Anionic species (dicarbonato) 
at higher 
pH (8.6) results in 
decreased Kd

ANL-WIS-MD-000012 REV00 XVI-8 April 2000 
Table XVI-2. YMP Groundwater Compositions, Compared to NAGRA Groundwater (and Porewater) 
Compositions 
Parameters Crystalline Rock Valanginian Marl Bentonite YMP* 
Ions ( M) West-area 
Water 
East-area 
Water 
NaCl Water NaHCO3 
Water 
Porewater J-13 
Na+ 14000 8000 490000 20000 80000 1990 
K+ 200 100 700 51 170 129 
Ca2+ 400 300 4500 57 11 324 
Cl- 4000 700 470000 300 300 201 
HCO3
- 5000 5000 4000 16000 52000 2,090 (TC) 
SO4
2- 3000 1000 10.7 100 4500 192 
Eh (mV) -180 -30 -125 -125 -400 na 
pH 7.7 7.7 7.5 8.3 9 na 
ionic strength 
(M) 
0.02 0.01 0.2 0.02 0.08 na 
Table XVI-3. Summary of Sorption Data Developed by NAGRA for Bentonite, Crystalline Rock, and 
Marl 
Radionuclide 
Sorbate and 
Oxidation 
State(s) at YMP 
NAGRA Database NAGRA Rock Kd (mL/g) 
(specific reference) 
NAGRA Mineral Kd (mL/g) 
Sr(II) 461 entries; 
39 references 
1,000 
(p. Sr-11) 
smectite = 20; 
oxides, quartz = 0 
Cs(I) 516 entries; 
41 references 
1,000 
(p. Cs-11) 
smectite = 2,000; 
oxides, quartz = 40 
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 analog 
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

ANL-WIS-MD-000012 REV00 XVI-9 April 2000 
Table XVI-4. Summary of Sorption Data Developed by EPA (1999) 
Radionuclide 
Sorbate and 
Oxidation 
State(s) at YMP 
Database 
Category selected: 
Smectite colloid Kd; 
Iron-(hydr)oxide colloid Kd 
Range: Minimum to Maximum 
(geom. mean): 
Smectite Kd (mL/g); 
Iron-(hydr)oxide Kd (mL/g) 
Sr(II) 63 entries 
(166 in Appendix) 
high clay (20-60 wt%); pH 8-10; 
low clay (<4 wt%); pH 8-10 
300 to 1,700 (714); 
3 to 120 (19) 
Cs(I) 177 Kd values high clay (20-60 wt%) 
low clay (<4 wt%) 
80 to 26,700 (1,462); 
30 to 9,000 (520) 
Th(IV) 17 entries pH 8-10; [Th] <10-9 M 20 to 2,000 (200) 
Pa(V) not discussed, 
no good analog 
U(VI) 20 references pH 8 0.4 to 250,000 (316) 
Np(V) not discussed, 
no good analog 
Pu(IV, V, VI) high clay (51-70 wt%); 3-4 meq/L CO3
2-; 
low clay (<4 wt%); 3-4 meq/L CO3
2- 
1,860 to 2,550 (2,178); 
80 to 470 (194) 
Am(III) not discussed, 
no good analog

Table XVI-5. YMP-specific Data for Sorption of Plutonium and Americium on Colloidal Hematite, Goethite, Ca-montmorillonite, and Silica: (a) Pu(IV), (b) Pu(V), (c) 
Am(III) 
sorbate = 239Pu(IV) 
Kd (mL/g * 1000) mean 
source mineral [colloid] (mg/L) sorption directiontime (d) J13 actual J13 simulant 
A1 hematite 1 forward 1 270 18000 
A1 hematite 1 forward 2 570 1900 
A1 hematite 1 forward 4 1800 790 
A1 hematite 1 forward arithmetic mean 880 6897 X Kd,forward / Kd,backward 0.7 0.5 
A5 hematite 1 backward 150 1200 13000 X Kd,backward / Kd,forward 1.4 1.9 
arithmetic mean 1040 9948 5494 
Kd (mL/g * 1000) mean 
source mineral [colloid] (mg/L) sorption directiontime (d) J13 actual J13 simulant 
A2 goethite 1 forward 1 1.9 12 
A2 goethite 1 forward 2 2.3 7.1 
A2 goethite 1 forward 4 1.5 18 
A2 goethite 1 forward arithmetic mean 1.9 12 X Kd,forward / Kd,backward 0.9 0.3 
A5 goethite 1 backward 150 2.1 49 X Kd,backward / Kd,forward 1.1 4.0 
arithmetic mean 2.0 31 16 
Kd (mL/g * 1000) mean 
source mineral [colloid] (mg/L) sorption directiontime (d) J13 actual J13 simulant 
A3 Ca-montmorillonite 1 forward 1 5.3 180 
A3 Ca-montmorillonite 1 forward 2 3.7 240 
A3 Ca-montmorillonite 1 forward 4 3.5 140 
A3 Ca-montmorillonite 1 forward 10 3.3 190 
A3 Ca-montmorillonite 1 forward arithmetic mean 4.0 188 X Kd,forward / Kd,backward 2.8 64.7 
A5 Ca-montmorillonite 1 backward 150 1.4 2.9 X Kd,backward / Kd,forward 0.4 0.0 
arithmetic mean 2.7 95 49 
Kd (mL/g * 1000) mean 
source mineral [colloid] (mg/L) sorption directiontime (d) J13 actual J13 simulant 
A4 silica-PST-1 1 forward 1 3.4 7.7 
A4 silica-PST-1 1 forward 2 3.5 10 
A4 silica-PST-1 1 forward 4 4.0 16 
A4 silica-PST-1 1 forward 10 4.9 30 
A4 silica-PST-1 1 forward arithmetic mean 4.0 16 X Kd,forward / Kd,backward 1.4 2.8 
A5 silica-PST-1 1 backward 150 2.8 5.7 X Kd,backward / Kd,forward 0.7 0.4 
arithmetic mean 3.4 11 7.1 
J13 actual J13 simulant 
ANL-WIS-MD-000012 REV00 XVI-10 April 2000

Table XVI-5. YMP-specific Data for Sorption of Plutonium and Americium on Colloidal Hematite, Goethite, Ca-montmorillonite, and Silica: (a) Pu(IV), (b) Pu(V), (c) 
Am(III) 
sorbate = 239Pu(V) 
Kd (mL/g * 1000) mean 
source mineral [colloid] (mg/L) sorption directiontime (d) J13 actual J13 simulant 
A1 hematite 1 forward 1 10 2100 
A1 hematite 1 forward 2 19 1300 
A1 hematite 1 forward 4 200 430 
A1 hematite 1 forward arithmetic mean 76 1277 X Kd,forward / Kd,backward 0.2 0.7 
A5 hematite 1 backward 150 430 1800 X Kd,backward / Kd,forward 5.6 1.4 
B1 hematite 200 forward 1 23 150 
B1 hematite 200 forward 2 75 170 
B1 hematite 200 forward 4 97 700 
B1 hematite 200 forward 10 110 120 
B1 hematite 200 forward arithmetic mean 94 330 X 
arithmetic mean 200 1136 668 
Kd (mL/g * 1000) mean 
source mineral [colloid] (mg/L) sorption directiontime (d) J13 actual J13 simulant 
A2 goethite 1 forward 1 3.5 77 
A2 goethite 1 forward 2 6.0 760 
A2 goethite 1 forward 4 8.7 140 
A2 goethite 1 forward arithmetic mean 6.1 326 X Kd,forward / Kd,backward 0.3 3.0 
A5 goethite 1 backward 150 18 110 X Kd,backward / Kd,forward 3.0 0.3 
arithmetic mean 12 218 115 
Kd (mL/g * 1000) mean 
source mineral [colloid] (mg/L) sorption directiontime (d) J13 actual J13 simulant 
A3 Ca-montmorillonite 1 forward 1 4.2 8.3 
A3 Ca-montmorillonite 1 forward 2 5.1 13 
A3 Ca-montmorillonite 1 forward 4 6.4 26 
A3 Ca-montmorillonite 1 forward 10 10 47 
A3 Ca-montmorillonite 1 forward arithmetic mean 6.4 24 X Kd,forward / Kd,backward 0.4 3.4 
A5 Ca-montmorillonite 1 backward 150 17 7.0 X Kd,backward / Kd,forward 2.7 0.3 
B1 Ca-montmorillonite 200 forward 1 0.8 1.1 
B1 Ca-montmorillonite 200 forward 2 1.0 1.2 
B1 Ca-montmorillonite 200 forward 4 4.0 1.7 
B1 Ca-montmorillonite 200 forward 10 5.8 6.5 
B1 Ca-montmorillonite 200 forward arithmetic mean 2.9 2.6 X 
arithmetic mean 8.8 11 10 
ANL-WIS-MD-000012 REV00 XVI-11 April 2000

Table XVI-5. YMP-specific Data for Sorption of Plutonium and Americium on Colloidal Hematite, Goethite, Ca-montmorillonite, and Silica: (a) Pu(IV), (b) Pu(V), (c) 
Am(III) 
Kd (mL/g * 1000) mean 
source mineral [colloid] (mg/L) sorption directiontime (d) J13 actual J13 simulant 
A4 silica-PST-1 1 forward 1 0.6 0.4 
A4 silica-PST-1 1 forward 2 0.8 0.9 
A4 silica-PST-1 1 forward 4 1.5 1.8 
A4 silica-PST-1 1 forward 10 2.2 2.0 
A4 silica-PST-1 1 forward arithmetic mean 1.3 1.3 X Kd,forward / Kd,backward 0.3 0.1 
A5 silica-PST-1 1 backward 150 4.1 24 X Kd,backward / Kd,forward 3.2 18.9 
B1 silica-PST-1 200 forward 1 5.5 10 
B1 silica-PST-1 200 forward 2 5.8 10 
B1 silica-PST-1 200 forward 4 6.7 12 
B1 silica-PST-1 200 forward 10 8.1 15 
B1 silica-PST-1 200 forward arithmetic mean 6.5 12 X 
arithmetic mean 4.0 12 8 
sorbate = 241Am 
Kd (mL/g * 1000) mean 
source mineral [colloid] (mg/L) sorption directiontime (d) J13 actual J13 simulant 
B2 hematite 200 forward 1 270 200 
B2 hematite 200 forward 2 200 510 
B2 hematite 200 forward 4 320 6200 
B2 hematite 200 forward 10 79 120 
B2 hematite 200 forward arithmetic mean 217 1758 X Kd,forward / Kd,backward 4.3 9.3 
B3 hematite 200 backward 42 51 190 X Kd,backward / Kd,forward 0.2 0.1 
arithmetic mean 134 974 554 
Kd (mL/g * 1000) mean 
source mineral [colloid] (mg/L) sorption directiontime (d) J13 actual J13 simulant 
B2 Ca-montmorillonite 200 forward 1 24 46 
B2 Ca-montmorillonite 200 forward 2 65 110 
B2 Ca-montmorillonite 200 forward 4 100 100 
B2 Ca-montmorillonite 200 forward 10 14 23 
B2 Ca-montmorillonite 200 forward arithmetic mean 51 70 X Kd,forward / Kd,backward 3.6 7.2 
B3 Ca-montmorillonite 200 backward 42 14 10 X Kd,backward / Kd,forward 0.3 0.1 
arithmetic mean 32 40 36 
Kd (mL/g * 1000) mean 
source mineral [colloid] (mg/L) sorption directiontime (d) J13 actual J13 simulant 
B2 Silica-PST-1 200 forward 1 7.4 5.0 
B2 Silica-PST-1 200 forward 2 21 32 
B2 Silica-PST-1 200 forward 4 4.6 20 
B2 Silica-PST-1 200 forward 10 12 5.6 
B2 Silica-PST-1 200 forward arithmetic mean 11 16 X Kd,forward / Kd,backward 0.9 1.6 
B3 Silica-PST-1 200 backward 42 12 10 X Kd,backward / Kd,forward 1.1 0.6 
arithmetic mean 12 13 12 
ANL-WIS-MD-000012 REV00 XVI-12 April 2000

ANL-WIS-MD-000012 REV00 XVI-13 April 2000 
Table XVI-6. Summary of Sorption Data 
Radionuclide 
Sorbate and 
Oxidation 
State(s) at 
YMP 
Kds from CRWMS M&O 
(2000b) 
(Table 16  Kds for SZ units): 
Devitrified tuff (min  max) 
Iron-(Hydr)oxides (min  max) 
CRWMS M&O (2000b)  
Minimum Kd approach 
(Section 6.4) 
NAGRA Data Set 
(Stenhouse 1995) 
Sr(II) 10  200 
0  30 
>100 not discussed 
Cs(I) 20  1000 
0  500 
>100 not discussed 
Th(IV) (see Am) ? not discussed 
Pa(V) 0  100 
500  1000 
like Np(V) not discussed 
U(VI) 0  5 
100  1000 
quartz  0.1 to 0.07 from pH 
7 to 8.5 (Table 4) 
not discussed 
Np(V) 0  10 
0  100 
montmorillonite = 30; 
hematite = 1,800 (Figure 6) 
not discussed 
Pu(IV, V, VI) 5  150 
1000  5000 
hematite > montmorillonite 
> quartz; 
>10,000 (Table 1) 
iron-(hydr)oxides = 7 x 104; 
Ca-montmorillonite = 3 x 104 
Am(III) 100  2000 
1000  5000 
>100 hematite = 6 x 105; 
Ca-montmorillonite = 4 x 104

ANL-WIS-MD-000012 REV00 XVI-14 April 2000 
Table XVI-7. Recommendations for PA 
Radionuclide Sorbate and 
Oxidation State(s) at YMP 
Recommended Kd (mL/g) 
Iron-(hydr)oxide Colloids 
Smectite Colloids 
Source 
Sr(II) 200 
30 
Maximum values developed in Concas UZ transport 
AMR, for devitrified tuff and iron-(hydr)oxides, 
respectively, for saturated zone units. 
Cs(I) 1000 
500 
as above 
Th(IV) (use same values as developed 
for Am) 
as above 
Pa(V) 100 
1000 
as above 
U(VI) 5
1000 
as above 
Np(V) 10 
100 
as above 
Pu(IV, V, VI) 70,000 
30,000 
Values from NAGRA data compendium (Stenhouse 
1995) 
Am(III) 600,000 
40,000 
Values from NAGRA data compendium (Stenhouse 
1995)